vvEPA
United States
Environmental Protection
Agency
Office of Research and
Development
Washington DC 20460
EPA/600/R-94/120
August 1994
Evaluation of
Technologies for In-Situ
Cleanup of DNAPL
Contaminated Sites
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EPA/600/R-94/120
August 1994
EVALUATION OF TECHNOLOGIES FOR IN-SITU CLEANUP
OF DNAPL CONTAMINATED SITES
by
Dennis G. Grubb and Nicholas Sitar
Department of Civil Engineering
University of California
Berkeley, California 94720
Cooperative Agreement No. CR-81 8956
Project Officer
S.G. Schmelling
Processes and Systems Research Division
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, IL 60604-3590
ROBERT S. KERR ENVIRONMENTAL RESEARCH LABORATORY
OFFICE OF RESEARCH AND DEVELOPMENT
U.S. ENVIRONMENTAL PROTECTION AGENCY
ADA, OK 74820
Printed on Recycled Paper
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NOTICE
The information in this document has been funded wholly or in part by the United States
Environmental Protection Agency (EPA) under Cooperative Agreement No. CR-818956 to the University
of California, Berkeley. It has been subjected to the Agency's peer and administrative review, and it has
been approved for publication as an EPA document. Mention of any trade names or commercial products
does not constitute endorsement or recommendation for use.
All research projects making conclusions or recommendations based on environmentally related
measurements and funded by the Environmental Protection Agency are required to participate in the
Agency Quality Assurance Program. This project did not involve environmentally related measurements
and did not involve a Quality Assurance Project Plan.
11
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FOREWORD
EPA is charged by Congress to protect the Nation's land, air and water systems. Under a mandate
of national environmental laws focused on air and water quality, solid waste management and the control of
toxic substances, pesticides, noise and radiation, the Agency strives to formulate and implement actions
which lead to a compatible balance between human activities and the ability of natural systems to support
and nurture life.
The Robert S. Kerr Environmental Research Laboratory is the Agency's center of expertise for
investigation of the soil and subsurface environment. Personnel at the Laboratory are responsible for
management of research programs to: (a) determine the fate, transport, and transformation rates of pollutants
in the soil, the unsaturated zone, and the saturated zones of the subsurface environment; (b) define the
processes to be used in characterizing the soil and subsurface environment as a receptor of pollutants; (c)
develop techniques for predicting the effects of pollutants on the ground water, soil, and indigenous organisms;
and (d) define and demonstrate the applicability and limitations of using natural processes, indigenous to the
soil and subsurface environment, for the protection of this resource.
This report provides a review and technical evaluation of in-situ technologies for remediation of DNAPL
contamination occurring below the ground-water table. Various in-situ technologies are reviewed and are
evaluated on the basis of their theoretical background, field implementation, level of demonstration and
performance, waste, technical and site applicability/limitations, and cost and availability.
4
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ABSTRACT
Ground water contamination by non-aqueous phase liquids poses one of the greatest remedial
challenges in the field of environmental engineering. Denser-than-water non-aqueous phase liquids
(DNAPLs) are especially problematic due to their low water solubility, high density, and capillary forces
arising from interfacial tension between the DNAPLs and water. As a result, conventional pump-and-treat
technologies have met poor success in remediation of DNAPL contaminated aquifers. In fact, in certain
situations, conventional pump-and-treat methods may actually extend existing contamination into previously
uncontaminated areas. The problems associated with current pump-and-treat remedial approaches have
served as the impetus to develop alternative technologies to accelerate in-situ DNAPL contamination
remediation. This report provides a review and technical evaluation of in-situ technologies for remediation
of DNAPL contamination occurring below the ground-water table. Various in-situ technologies are reviewed
and are evaluated on the basis of their theoretical background, field implementation, level of demonstration
and performance, waste, technical and site applicability/limitations, and cost and availability.
IV
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CONTENTS
SECTION PAGE
Notice ii
Foreword iii
Abstract iv
Figures vii
Tables xii
Acknowledgement xiii
1.0 Introduction 1
1.1 Overview and Organization 1
1.2 Methodology 1
1.3 Limitations of the Report 2
2.0 DNAPL Fate and Transport Processes 3
2.1 Physics of Multiphase Flow 3
2.1.1 Darcy's law and capillary forces 3
2.1.2 Saturation and relative permeability 9
2.1.3 Other factors contributing to DNAPL mobility 11
2.2 Physical and Chemical Properties of NAPLs 12
2.2.1 Aqueous solubility 13
2.2.2 Density 13
2.2.3 Interfacial tension 13
2.2.4 Wettability and spreading 14
2.2.5 Viscosity 16
2.2.6 Vapor pressure and Henry's law constant 16
2.2.7 Octanol-water partitioning coefficient (K0J 17
2.2.8 Boiling point 17
2.2.9 Dielectric constant 17
2.3 Physical Properties of Subsurface Systems 17
2.3.1 Porosity 18
2.3.2 Permeability 18
2.3.3 Clay-pore fluid interactions 19
2.3.4 Organic matter 19
2.4 Multicomponent-Multiphase Equilibria 20
2.4.1 Multicomponent NAPLs . . . 21
2.4.2 Surfactants, cosolvents and multicomponent NAPLs 22
2.5 Unsaturated and Saturated Zone Transport Mechanisms 23
2.5.1 Unsaturated zone transport 23
2.5.2 Saturated zone transport 24
2.6 Estimation of the Extent of Site Contamination and Site Characterization. . .26
2.6.1 Ground-water samples 26
2.6.2 Soil gas samples 28
2.6.3 Well product thickness 33
2.6.4 Soil samples 34
2.7 Challenges Facing In-Situ Technologies 34
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3.0 Technology Descriptions 37
3.1 Technology Evaluation Format 37
3.2 Biological Processes 39
3.2.1 Aerobic biodegradation 42
3.2.2 Anaerobic biodegradation 46
3.3 Electrolytic Processes 50
3.3.1 Electro-osmosis (EO) 51
3.3.2 Electroacoustic soil decontamination (ESD) 55
3.4 Containment and Ground Modification 58
3.4.1 Isolation and containment 58
3.4.2 Stabilization/solidification 65
3.4.3 Permeable treatment walls 69
3.5 Soil Washing Processes 72
3.5.1 Alkali soil washing 73
3.5.2 Cosolvent soil washing 79
3.5.3 Surfactant soil washing 87
3.5.4 Water flooding and ground-water extraction 94
3.6 Air Stripping 99
3.6.1 Air sparging and vacuum extraction 99
3.6.2 Vacuum vaporizer wells (UVB) 106
3.7 Thermal Processes 114
3.7.1 Contained recovery of oily wastes (CROW®) 114
3.7.2 Steam enhanced extraction (SEE) 121
3.7.3 Radio frequency heating 127
3.7.4 Vitrification -131
4.0 In-Situ Technology Comparisions 137
4.1 Introduction 137
4.2 Explanation of terms 137
4.3 Promising technologies 138
References 141
Copyright Permissions 167
VI
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FIGURES
Number
2.1.1 Schematic of a multiphase system consisting of DNAPL, and air, aqueous, and
solid phases 6
2.1.2 Mobilization of residual oil ganglia relating NAPL saturation ratio, S /S01*
(initial/final NAPL S%), and the capillary number, Nc. Nc* and Nc denote initial
mobilization and 100% removal, respectively 6
2.1.3 Random vertical migration of PCE in saturated bead pack (dia. 0.49-0.70 mm) 8
2.1.4 NAPL-water relative permeability curves in a porous medium 10
2.1.5 Infiltration of PCE (top) and DCM (bottom) in static glass trough experiments.
Volume of DNAPL released: 10 liters. Elapsed time and heights of unsaturated
(hj, saturated (hs) and capillary fringe (hc) shown 12
2.2.1 Relationship between contact angle (<)>) and wettability 14
2.5.1 Schematic of the distribution of subsurface contamination emanating from residual
DNAPL source in the vadose zone 25
2.5.2 Schematic of the distribution of subsurface contamination emanating from residual
DNAPL sources in the vadose and water saturated zones, and DNAPL pools 25
2.5.3 Schematic of fractured bedrock contamination resulting from mobile and pooled
DNAPL 26
2.6.1 Schematic of a simplified flow geometry of ground water sweeping past a NAPL lens .... 28
2.6.2 Computed average TCE concentrations of ground water (a) and of soil gas (b)
sweeping past a NAPL lens using a simplified geometry (Figure 2.6.1) 29
2.6.3 Predicted and observed evolutions of: (top) aqueous hydrocarbon concentrations
in equilibrium with bicomponent NAPL; and (bottom) mole fractions of the
bicomponent NAPL 30
2.6.4 Evolution of toluene and o-xylene soil gas concentrations in a homogeneous
sand pack 32
2.6.5 Soil gas composition as a function of time during soil venting at a gasoline
contaminated site 32
2.6.6 Bypassing air flow mechanism and its effect on the composition profile of an
evaporating bicomponent NAPL pool trapped within low permeability zone 33
3.2.1 Relative rates of reduction and oxidation as a function of halogenation 40
vii
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3.2.1.1 Methanotrophic conversion of methane to methanol by Methane Monooxygenase
(MMO) and the formation of TCE-epoxide as the initial step of TCE oxidation 42
3.2.1.2 Schematics illustrating oxygen and nutrient delivery using spargers (a) and an
infiltration gallery (b) 44
3.2.2.1 Pathways for anaerobic biotransformation of chlorinated aliphatics including
abiotic (a) transformations 47
3.3.1.1 Ion distribution adjacent to clay particle surface 52
3.3.1.2 Schematic of electro-osmotic flow resulting from an applied electric field in a
charged porous medium 52
3.3.1.3 Schematic of an in-situ electro-osmotic extraction system 54
3.3.1.4 Schematic layout of electrode arrays for the in-situ electrokinetic application at
Baton Rouge field test site 54
3.3.2.1 Conceptual layout of the electroacoustical soil decontamination process 56
3.4.1.1 Relationship between the permeability and bentonite content of SB backfill materials 60
3.4.1.2 Primary and secondary overlapping patterns for in-situ soil mixing processes 60
3.4.1.3 Schematic configuration of a coupled impervious barrier and hydraulic gradient
control system. Ground-water flow across barrier is maintained into contaminated
groundwater region 61
3.4.1.4 Schematic of SB slurry wall installation process 61
3.4.1.5 Schematic showing different grouting techniques 62
3.4.1.6 Vertical section taken through a composite geomembrane-SB slurry wall impervious
barrier system, Liguria, Italy 64
3.4.1.7 Vertical section taken through utility corridor in which jet grouted impervious
barrier was constructed to join SB slurry walls 64
3.4.2.1 Schematic of crane mounted in-situ shallow soil mixing (SSM) process 67
3.4.2.2 Schematic of drilling pattern for in-situ deep soil mixing (DSM) process 67
3.4.2.3 Schematics of various final soil treatment patterns of SSM and DSM in-situ
stabilization/solidification processes 67
3.4.3.1 Conceptual plan views for possible configurations of in-situ permeable treatment
walls 70
3.5.1.1 Schematic of alkali recovery process 74
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3.5.1.2 pH comparison of commonly available alkali chemicals 75
3.5.1.3 Comparison of experimental and theoretical alkali breakthrough times for NaOH,
Na4SiO4, and Na2CO3 as a function of pH 76
3.5.1.4 IFF values of Wilmington Ranger zone crudes with alkalis at 52°C 77
3.5.1.5 IFT values of Dome Lloydminster "A" pool crude as a function of alkali and
surfactant addition 78
3.5.2.1 Schematic of the fluid-fluid displacement process 80
3.5.2.2 Column effluent histories of miscible displacements 80
3.5.2.3 Viscosity enhancement of water and NAPL by isopropanol (IPA) in the H2O-IPA-
Naphtha ternary liquid system at 20°C 81
3.5.2.4 Equilibrium phase diagrams for the IPA-Soltrol-Brine (2% CaCl2) and TBA(tert-butyl
alcohol)-Soltrol-Brine (2% CaCl2) systems showing binodal curves and inclination
of tie lines 81
3.5.2.5 Idealized fluid-fluid displacement using a cosolvent (IPA) slug 82
3.5.2.6 Effect of mobility ratio on displacement fronts and injected pore volumes until
breakthrough using quarter of five-spot method 83
3.5.2.7 Area contacted by fluid drive after breakthrough, quarter of five spot method 84
3.5.2.8 Effect of mobility ratio on fluid recovery from segmented-stratified porous media
model 84
3.5.2.9 Comparison of effluent histories for (a) vertical and (b) horizontal H2(
miscible displacements in soil cores. IPA->TCE mobility ratios stable for both
displacements whereas H2O—>IPA are not. The IPA-»TCE interface in (b) is unstable
due to gravity effects, while the H2O->IPA interfaces in both (a) and (b) are unstable .... 86
3.5.3.1 Physical property changes of aqueous solutions of sodium lauryl sulfate in vicinity
of critical micelle concentration 88
3.5.3.2 Relationships between salt concentration, oil chain length, surfactant concentration
on (a) interfacial tension, and (b) surfactant partitioning and micelle formation in
petroleum sulfonate systems 88
3.5.3.3 Schematic illusffating the l-»m-»u-* transition and the factors influencing its
determination in surfactant/oil/brine/alcohol systems 89
3.5.3.4 Schematic illustrating fluid bank formation as a function of saturationand distance
in a surfactant/polymer flood 90
3.5.3.5 Schematic of dual drain line system for the 1988 field test using water and combined
alkali/surfactant flooding of heavy oils 92
IX
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3.5.3.6 Schematic of field test using water and surfactant flooding for enhanced PCE recovery
(Borden, Canada) 92
3.5.4.1 Schematic illustrating the upconing phenomena of a dense fluid phase to pumping
stress in the overlying fluid phase 96
3.5.4.2 Schematic of dual drain line system for pumping of both light and dense fluid phase
to enhance the recovery of the underlying, denser phase 96
3.6.1.1 Schematic of air sparging/vacuum extraction system 100
3.6.1.2 Schematic of (a) typical air sparging system configuration and (b) the effect of
subsurface heterogeneities on gas channeling 101
3.6.1.3 Possible air sparging well configurations 103
3.6.1.4 Effect of gas injection pressure on air sparging system 103
3.6.2.1 Streamlines for longitudinal vertical recirculation patterns for several ground-water
flow velocities: (a) 0 m/d; (b) 0.3 m/d; (c) 1.0 m/d 107
3.6.2.2 Schematic of three-dimensional capture zone for anisotropic soil conditions using
(a) single and (b) dual UVBs. Effect of recirculation cell on incoming flow is
indicated by the depressed areas 108
3.6.2.3 Schematic of vacuum vaporizer well (UVB) configured with (a) separation plate
and vacuum extraction; (b) no separating plate and vacuum extraction; and, (c)
separation plate and closed air recirculation 109
3.6.2.4 Field data obtained from Mannheim-Kaefertal site (Germany). Measured hydraulic
heads (a) indicate vertical flow patterns in aquifer. Downflow in well occurs until
6/13/89, upflow thereafter. Corresponding PCE concentrations in ground water
monitoring locations in the lower UVB (b), upper UVB (c), and in a downgradient
well (d) which is screened in the upper portion of the aquifer illustrate the
importance of upflow in the UVB well on PCE recovery Ill
3.7.1.1 Influence of temperature on fluid viscosity (a) and density (b) for several DNAPLs 115
3.7.1.2 Conceptual schematic of the CROW® process 116
3.7.1.3 Temperature dependence of DNAPL recovery using hot water and surfactant solutions
in one-dimensional column tests by CROW® process 116
3.7.1.4 DNAPL removal and corresponding temperature isotherms using hot (a) water and
(b) surfactant solutions in 3-D tests by CROW® process 119
3.7.1.5 Temperature profiles at well location BP24 during CROW® pilot test (MN) 119
3.7.1.6 NAPL saturation profiles in soil samples (CT1,CT2) taken in vicinity of injection
well (IW1) after CROW® pilot test 120
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3.7.2.1 Temperature distribution near steam condensation front 122
3.7.2.2 Effect of soil heterogeneity on steam front advancement 122
3.7.2.3 Total hydrocarbon concentration measured in column effluent reported per liter of
displaced fluid. Concentration spike indicates presence of NAPL "bank" 124
3.7.2.4 Temperature profile and calculated separate phase o-xylene saturation before
(residual water flood saturation) and after steam flood commencement (t=5000s).
O-xylene "bank" situated ahead of steam front 124
3.7.2.5 Schematic of in-situ steam enhanced extraction process 125
3.7.3.1 Loss tangent and dielectric constant of tar sand samples (Vernal, UT) 129
3.7.3.2 Schematic of radio-frequency soil heating process showing electromagnetic electrode
array and vacuum hood 129
3,7.4.1 Schematic illustrating the in-situ vitrification (ISV) process 133
3.7.4.2 Chemical processes and reactions occurring within and near the soil melt zone 133
3.7.4.3 Effect of soil moisture on cost of in-situ vitrification 136
XI
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TABLES
Number Page
2.1 Most prevalent DNAPLs at U.S. Superfund Sites 4
2.2 Typical soil retention values for organic liquids in different soil types 11
2.3 Experimentally measured contact angles of DNAPLs in different soil types 15
2.4 Typical equilibrium concentrations of pure and gasoline-derived BTX compounds 27
2.5 BTX concentrations in water and soil from same borehole at a gasoline contaminated
site 34
3.2.1 Microbial utilization of organic compounds as a function of biological process type
and environmental conditions 39
3.2.2 Comparison of substrate utilization rates by mixed cultures using different electron
acceptors 41
3.3.1.1 Direct and coupled flow phenomena occurring in the subsurface 50
3.6.1.1 Summary of data published on air sparging sites 105
3.7.4.1 Typical organic destruction/removal efficiencies by ISV 135
4.1.1 In-situ technology comparisons 139
xn
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ACKNOWLEDGMENT
The authors would like to thank Dr Stephen G. Schmelling, the Robert S. Kerr Environmental
Research Laboratory Project Officer, for his assistance and input throughout this project. The assistance
of all individuals that responded to the "In-Situ DNAPL Remediation Technology Questionnaires" and
provided information via discussion and correspondence is much appreciated.
Thanks also to Professors Lisa Alvarez-Cohen, Clayton J. Radke, and Kent S. Udell of the
University of California, Berkeley, for their assistance in reviewing the manuscripts of sections 3.2,3.5, and
3.7, respectively. Ms. Selina Tarn and Mr. Jared Dunn provided invaluable assistance during the literature
review and data base management phases of the project. Ms. Elizabeth Turner assisted in word
processing and figure preparation.
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SECTION 1.0
INTRODUCTION
1.1 OVERVIEW AND ORGANIZATION
Ground-water contamination by immiscible hydrocarbons, often referred to as non-aqueous phase
liquids (NAPLs, or LNAPLs and DNAPLs denoting those lighter or denser than water, respectively), poses
one of the greatest remedial challenges in the field of environmental engineering. DNAPLs are especially
problematic due to their low water solubility, high density, and capillary forces arising from interfacial tension
between the DNAPLs and water. As a result, conventional pump-and-treat technologies have met poor
success in remediation of DNAPL contaminated aquifers [Wilson, 1992].
The objective of this report is to provide a comprehensive assessment of the current state-of-the-art
of in-situ treatment technologies as they pertain to the treatment, mobilization, and recovery of DNAPLs
from the subsurface. Focus is placed on attempting to identify in-situ technologies capable of addressing
the remediation of DNAPLs situated below the water table; secondary importance is placed on
contaminants dissolved in the aqueous phase. Several of the evaluated technologies were not originally
developed for remediation of contaminated sites, much less DNAPLs. As a result, some of the
technologies have not yet been demonstrated on DNAPLs, and owing to their developmental stage, have
not been demonstrated in the field and below the water table. However, their applicability to remediation
of DNAPLs is nonetheless considered in order to not rule them out prematurely. On the other hand, while
the technology required to implement certain remedial approaches may be currently available, the expertise
required for successful full-scale field application may be lacking. Also, some of the technologies have
been fully demonstrated only in non-environmental applications and are just being adapted for
environmental applications.
Aside from technology evaluation and selection, there are several factors controlling remedial
options. Containment, recovery, and remediation options are usually dictated by site considerations,
regulations, cost, extent of contamination, and presence of other waste types. The problem of mixed
inorganic and organic wastes is a complicated one that has not been fully addressed here because this
project was limited to DNAPLs. Nonetheless, site heterogeneity and regulatory approval are seen to be
the most critical factors controlling remedial options. Extensive site heterogeneity can render all
technologies ineffective, some more than others, and most technologies will require regulatory approval to
implement. Albeit mentioned, regulatory acceptance and related issues are beyond the scope of this
document.
In order to identify the various physical and technical barriers limiting DNAPL treatment,
mobilization and recovery, the mechanisms responsible for DNAPL fate and transport are outlined in section
2.0. In-situ technologies are arranged alphabetically by major process type in Section 3.0 and are
evaluated for their applicability to cleanup of DNAPL contaminated sites. The following aspects of each
relevant in-situ technology were evaluated: theoretical background, field implementation, level of
demonstration and performance, applicability/limitations, and cost and availability. Finally, the in-situ
technologies are compared and contrasted on several different levels in section 4.0.
1.2 METHODOLOGY
This study was conducted between December, 1991, and May, 1993; and the major effort in this
study was the review and compilation of information on in-situ DNAPL treatment technologies: no actual
experiments were conducted. Approximately 400 references were compiled during this study. Much of this
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information was collected from journal articles, conference proceedings, vendor and manufacturer fact
sheets and literature, and federal, state, and local agency reports and publications. The authors also
attended a number of conferences to obtain information as current as possible.
To supplement these sources of information, an "In-Situ DNAPL Remediation Technology
Description Questionnaire," was developed in cooperation with EPA personnel at the Robert S. Kerr
Environmental Research Laboratory. The questionnaires were sent to academic, federal and state, and
institute and industry professionals working in the area of DNAPL cleanup. The first mailing of these
questionnaires occurred in February, 1992. Positive responses were followed up with letters and personal
contacts to the extent necessary for an adequate technology evaluation. As the project progressed, the
correspondence was expanded as additional responses were solicited and received.
The technology descriptions of the relevant in-situ technologies were then prepared. The following
aspects of each relevant in-situ technology were evaluated: theoretical background, field implementation,
level of demonstration and performance, applicability/limitations, and cost and availability. Several
technologies have been demonstrated at different stages of development or have been demonstrated
numerous times. In such cases, an effort was made to evaluate the most current information and to select
representative applications illustrating the more interesting or impressive capabilities of the technology.
However, an exhaustive compilation of relevant case studies (as in the case of slurry wall construction) was
beyond the scope of this effort. Thus, the technology descriptions are intended to provide a basic technical
assessment of the technology and to identify its problem areas using basic principles.
1.3 LIMITATIONS OF THE REPORT
Due to the limited time frame of this project (18 months), the technology descriptions included
within this report cannot be considered exhaustive nor are they intended to be. Several additional factors
contribute to this fact: poor literature reporting; gaps due to unavailability of information; nature of
proprietary research and/or confidential information; and stage of development of technology. Furthermore,
the past performance of certain in-situ technologies which were originally designed for other applications
and/or targeted waste groups is not directly transferable. Consequently, the anticipated performance of
these technologies can be difficult to interpret within the context of DNAPL cleanup.
While this report can help identify potentially applicable in-situ technologies for cleanup of DNAPL
contaminated sites, it is not intended to be the sole basis for selecting a technology for a particular DNAPL
at a given site. Consequently, this report should serve as a complement to, not a substitute for,
engineering judgement, analysis, and design. Potential in-situ technologies must be further evaluated by
contacting technology developers (vendors, contractors, etc.,) and by performing bench-, and/or pilot-scale
treatability tests as necessary under site-specific conditions. This is especially true for undemonstrated
technologies and for technologies whose success depends heavily on the characteristics of the waste
matrix.
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SECTION 2.0
DNAPL FATE AND TRANSPORT PROCESSES
A dense non-aqueous phase liquid (DNAPL) is a sparingly soluble hydrocarbon having a specific
gravity greater than that of water at a typical soil temperature, usually less than 20-25°C. The distribution
of a DNAPL within the subsurface is the net result of coupled chemical and physical interactions between
the DNAPL, pore water, pore gases, and porous media. The chemical properties of the NAPL and
chemical equilibria relationships determine the partitioning of the compounds among the various phases,
while the physical properties of the pore fluids and porous media determine the mobility of each fluid phase.
A summary of the most important physical and chemical properties of DNAPLs found most commonly at
Superfund sites is presented in Table 2.1.
As shown in Figure 2.1.1, four distinct phases, each consisting of many chemical species, can be
present in the subsurface: the gas phase (in the vadose zone); the solid phase (rock, soil grains, soil
organic matter); the aqueous (polar) phase; and the DNAPL (non-polar phase). The fluid phases may be
mobile or immobile, and interphase partitioning is determined by such factors as the aqueous solubility limit,
Henry's constant, octanol-water partitioning coefficient, and sorption coefficients.
It is impossible within the scope of this report to describe on a site specific basis every aspect of
DNAPL transport in the unsaturated and saturated zones and the available modeling techniques. The aim
here is to conceptually describe the basic physical and chemical processes to aid in the discussion of
problems facing in-situ technologies (Section 2.7), and to aid in the evaluation of these technologies
(Sections 3,4).
2.1 PHYSICS OF MULTIPHASE FLOW
While the actual DNAPL flow and distribution will be complex owing to soil heterogeneities, two
major generalizations about the migration of DNAPL can be made. In order for the DNAPL to migrate as
a separate phase in any direction, both the capillary pressure resisting DNAPL flow and the DNAPL
retention capacity of the soil must be exceeded. For generality, portions of this discussion will be cast in
terms of "NAPL" because the relationships apply equally to both LNAPLs and DNAPLs.
2.1.1 Darcy's Law and Capillary Forces
Gas or liquid phase flow is governed by Darcy's Law for multiphase flow which incorporates both
capillary pressures and fluid properties and is written for each fluid phase (H2O, NAPL, etc.) as [Freeze
and Cherry, 1979; Muskat, 1982]:
-™ri
'M/
P,SS
(1)
Here, the subscript i denotes the fluid phase i; v, is the interstitial velocity of fluid phase i; x denotes
position; the fluid properties of viscosity and density are denoted by n, and pr respectively; parameters of
the porous media are denoted by <)>, k, and kn which are the porosity, permeability, and the relative
permeability of fluid i to the porous media, respectively; P, indicates the capillary pressure of fluid phase
i; g denotes the acceleration due to gravity; and 8 is the inclination of the porous media from the horizontal.
Furthermore, in water saturated porous media, the capillary pressure, Pc, between the NAPL (n) and the
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Solid
Water
DNAPL
Figure 2.1.1 Schematic of a multiphase system consisting of DNAPL and air, water, and solid phases.
[USEPA, 1992a].
0*0
V i
-------
aqueous phase (w) can be expressed as [Muskat, 1982; Villaume, 1985; Hunt et al., 1988b]:
2a.IM. coscj)
P = P - P = (2)
c n w
where anw is the NAPL-H20 interfacial tension; (|> is the contact angle (usually assumed to be zero when
medium is water wet); and r, is the mean pore throat radius.
Capillary pressures arise from the density differences in liquid properties, viscous forces caused
by hydraulic gradients, and due to interfacial tension between the two liquid phases. The Bond and
Capillary numbers are the ratios of buoyancy and viscous forces to interfacial forces in the vertical and
horizontal directions, respectively, and are used to estimate the mobility of NAPLs [Wilson and Conrad,
1984]:
Bond number (NB):
4 (5)
Capillary number (Nc):
PM'V>V h ' > 4 (6)
where Lv and Lh are the minimum lengths of ganglia for the NAPL mobilization to occur in the vertical and
horizontal directions, respectively. When Equations 5 and 6 are not satisfied, i.e., Lv and Lh are less than
the limiting lengths the NAPL ganglia become immobile or trapped. From equations 5 and 6, L^ and Lh
are seen to be a function of liquid properties, pore throat size and pressure gradients. It can also be
-------
inferred from these ratios that pore and macroscopic level soil heterogeneity (i.e., the actual distribution of
the pore throat diameters) can greatly affect the in-situ flow path of the advancing NAPL. An example of
an actual DNAPL distribution resulting from pore level soil heterogeneity is provided in Figure 2.1.3, while
the effects of macroscopic (field scale) heterogeneities are detailed later in Section 2.5.
Figure 2. 1.3
Random vertical migration of PCE in saturated bead pack (dia. 0.49 to 0.70 mm.)
[Schwille, 1988].
During infiltration at the ground surface, the limiting value of NB in the unsaturated and saturated zones
is frequently exceeded in most soils, with the possible exception of silts and clays, and downward migration
of the NAPL occurs. However, once emplaced, it is very difficult to mobilize the NAPL ganglia horizontally,
because the requisite hydraulic gradients and flow velocities are large and unattainable [Wilson and
Conrad, 1984; Sitar et al., 1987; Hunt et al., 1988b]. For example, in laboratory experiments, emplaced
TCE and PCE ganglia could not be mobilized by horizontal flow velocities of up to 14 m/day [Schwille,
1988].
The ability of an injected fluid to displace the resident pore fluids from the porous media can be
estimated using the mobility ratio, M. The mobility ratio, which neglects gravity and interfacial forces
[Buckley-Leverett assumption], is often used evaluate the potential success of a proposed fluid-fluid
displacement. Using water flooding as an example, the mobility ratio is defined as the ratio of the velocity
of the displacing fluid (water, w) to that of the resident pore fluids (i.e., NAPL or n):
M = _ =
(7)
-------
A favorable displacement is usually indicated by M < 1 [Buckley and Leverett, 1942; Muskat, 1982]. This
relationship is often referred to in soil washing applications, and its importance is discussed in section 3.5.
The Gravity Number NG, which is the ratio of gravity forces to viscous forces (NB/NC), provides an
indication of the potential for gravity under-ride or over-ride during fluid displacement processes. NG also
provides a measure of the slope of the advancing displacing fluid saturation front in a homogeneous porous
medium.
2.1.2 Saturation and Relative Permeability
The influence of fluid saturations on rate of migration of the advancing NAPL in both the
unsaturated and saturated zones is accounted for by the relative permeability (Equation 1). Also, as the
continuous NAPL advances in the vadose and water saturated zones, the soil retention capacity diminishes
the volume of mobile NAPL, causing stranding and isolation of discontinuous ganglia within in the soil
interstices.
Fluid saturation is defined as the ratio of the volume of the fluid in the pore space to the total pore
volume. The saturations of primary interest are the irreducible water saturation (Sow) and the residual
NAPL saturation (Sor or Sr). These fluid saturations correspond to the capillary pressure at which the
capillary pressure increases rapidly with negligible decreases in saturation [Corey, 1986]. Water flooding
of sandstone cores has shown that Sor can be on the order of 27% to 43% for low Bond and Capillary
numbers [Chatzis and Morrow, 1981]. The residual saturation definition is a matter of practical utility
because the residual NAPL saturation is affected by numerous factors. (1) pore size distribution; (2) pore
aspect ratio; (3) wettability; (4) soil texture: (5) clay fraction; (6) and the ambient Bond and Capillary
numbers [Wilson and Conrad, 1984; Sitar et al, 1987; Wilson et al., 1990]. The latter of these factors has
been observed to be the most important.
For example, laboratory studies employing glass beads showed that the typical range of Sor values
for NAPLs was on the order of 16% to 18% at Nc values less than 7x10"6, which is the range of typical
ground-water conditions [Morrow and Songkran, 1981; Morrow and Chatzis, 1982; Chatzis et al., 1983].
However, in experiments in which the hydraulic gradient and thus the viscous forces were increased,
complete NAPL displacement was observed at Nc > 1x10"3 [Wilson and^Conrad, 1984]. This implies that
any initial residual NAPL saturation (Sor*) can be reduced to zero (Sor**=0) by elevating NB and Nc, as
shown in Figure 2.1.2. However, unless the chemical and physical properties of the NAPL and porous
media are changed in some way, complete displacement of NAPL requires pressure increases beyond
those realistically attainable under normal groundwater conditions [Sitar et al., 1987].
Residual NAPL saturation can be alternatively expressed as soil retention capacity or a retention
factor, R, having the units of liters of NAPL per m3 soil (l/m3) [Schwille, 1967; de Pastrovich et al., 1979]:
R = S *//*1000 (8)
Table 2.2 presents experimentally determined values of residual saturations and retention factors of several
hydrocarbons in a variety of soil types under saturated and unsaturated conditions [Mercer and Cohen,
1990]. The specific retention of NAPLs in unsaturated and saturated sandy soils has been shown to be
on the order of 3 to 30 l/m3 and 5 to 50 l/m3, respectively [Schwille, 1984; Wilson and Conrad, 1984;
Mercer and Cohen, 1990]. In water-wet systems, vadose zone residual saturations are lower than residual
saturations in the saturated zone for a variety of reasons: continuity of films during drainage; presence of
non-wetting gas phase; larger buoyancy forces between NAPL and the gas phase compared to NAPL and
water; lower interfacial tension between the NAPL and the gas phase [Wilson et al, 1990]. Retention is
also affected by soil gradation (soil pore distribution). For example, under dry conditions, unsaturated fine
-------
grained soils (~S=55%) have been shown to retain more gasoline (multicomponent LNAPL) than coarse
sands (~S=14%) [Pfannkuch, 1983; Hoag and Marley, 1986].
The dependence of the relative permeability of water (kTO) and NAPL (krn) on saturation in a binary
fluid system is shown in Figure 2.1.4. The relative permeability can exhibit hysteresis and is a function of
NAPL properties, soil type, fluid saturation and thus, capillary pressure. The kTO and krn equal zero at Sow
and Sor, respectively. At any value above the irreducible water and residual NAPL saturations, the fluids
are considered to flow simultaneously, albeit, not necessarily in the same pores [USEPA, 1992a]. Since
the NAPL is the non-wetting fluid, it is likely to be flowing in the larger pore channels, and this partially
explains why small increases in NAPL saturation in initially water-wet soils can result in very large
decreases in the relative permeability of water [Schwille, 1988].
Ideally, relative permeability curves should be determined either experimentally or empirically fitted
to existing data, or converted from existing capillary pressure-saturation curves. This requires great care
because measurement methods and other experimental considerations such as testing techniques (steady
vs. unsteady), saturation determination (in-situ vs. ex-situ), viscous fingering, capillary end effects,
hysteresis, and scaling effects may affect two- and three-phase relative permeability estimations [Saraf and
McCafferty, 1982; Honarpour et al., 1986]. As a result, the determination of site- and NAPL-specific relative
permeability curves or capillary pressure-saturation curves is expensive, and often difficult, particularly in
the case of three-phase relative permeabilities. While NAPL-water relative permeability data is certainly
sufficient for estimating immiscible fluid flow in the saturated zone, three-phase (NAPL-H2O-air) relative
permeabilities are required for vadose zone transport analysis, and for analysis of situations when the gas
phase is introduced into the saturated zone, such as in air sparging and steam injection.
For these and other reasons, theoretical models have been developed to estimate three-phase
relative permeability and capillary pressures from two-phase data [Stone, 1973; Parker et al., 1987;
Delshad and Pope; 1989]. The two-phase relative permeabilities are often taken from Corey (1954) and
the relative permeability to gas is only taken to be a function of total liquid saturation. Oil-water and gas-
NAPL capillary pressures are assumed to be solely functions of water and gas saturations, respectively
[Leverett, 1941; Parker et al., 1987].
100V,
noo%
Figure 2.1.4 NAPL-water relative permeability curves in a porous medium [adapted from Schwille,
1988].
10
-------
TABLE 2.2 TYPICAL SOIL RETENTION VALUES FOR ORGANIC LIQUIDS IN DIFFERENT SOIL
TYPES [adapted from Mercer and Cohen, 1990]
NAPL
DNAPL
DNAPL
Tetrachloroethene
Benzene
Benzyl alcohol
p-Cymene
o-Xylene
Trichloroethene
Tnchloroethene
Trichloroethene
Tnchloroethene
1,1,1-Tnchloroethane
Tetrachloroethene
System
vad.
sat
vad.
sat
sat.
sat
sat.
vad.
vad.
vad.
vad.
sat.
sat.
Soil
sandy soils
sandy soils
fracture with 0.2 mm aperture
sand (92% sand, 5% silt, 3% clay)
sand (92% sand, 5% silt, 3% clay)
sand (92% sand, 5% silt, 3% clay)
sand (92% sand, 5% silt, 3% clay)
medium sand
fine sand
fine sand
loamy sand
coarse Ottawa sand
coarse Ottawa sand
Residual Saturation (Sr)
Retention factor, R (1/ m3)
Sr>0.01-0.10(2)
R > 3-30 (1)
Sr> 0.02-01 5 (2)
R > 5-50 (1)
R = 0.5 1 m-2 '2>
Sr = 0 24 (3)
Sr = 0.26 (3)
Sr = 0 16 (3)
Sr = 0.19(3)
Sr = 0 20 <4)
Sr = 0.19(4)
Sr = 0 15-020 (4)
Sr = 0 08 (5)
Sr = 0.15-040(6)
Sr = 0.15-025(6)
References:
1. Feenstraand Cherry (1988); 2' Schwille (1988); 3, Lenhard and Parker (1987); 4' Lin etal. (1982), 5 Gary et al (1989), 6' Anderson
(1988)
Delshad and Pope (1989) compared predicted versus experimental relative permeabilities using
three sets of data and seven three-phase relative permeability models. All seven methods provided
reasonably good fits to the experimental data and were seen to be dependent on the range of relative
permeability modeled. Certain methods allow for more parameter adjustment than others, but this
advantage can be offset by additional data requirements or assumptions. Despite these efforts, relative
permeability data of DNAPLs of environmental concern remain sparse [Mercer and Cohen, 1990].
While experimental data can be reasonably approximated by the relative permeability models,
recent visual experiments have shown that all aspects of DNAPL migration in a three-phase system cannot
be fully captured using current analyses [Wilson et al., 1990]. For example, under imbibing conditions in
a three-phase system, the sudden appearance of several discontinuous interpore DNAPL ganglia in the
downgradient direction could not be explained in relation to either stationary or slowly migrating continuous
DNAPL. Closer inspection revealed that the formation of new DNAPL ganglia directly resulted from DNAPL
film flow occurring at the air-water interface. Also, DNAPL film flow has been attributed to the propensity
of the DNAPL to spread (see section 2.2.4) [Wilson et al., 1990]. Hence, DNAPLs may be mobile at low
saturations, and phase continuity is not an essential prerequisite for appreciable DNAPL migration.
2.1.3 Other Factors Contributing to DNAPL Mobility
Factors other than capillary forces and retention capacity which contribute to the mobility of the
DNAPL include: volume of contaminant release; area of infiltration of contaminant; and time period over
11
-------
which the release occurred [de Pastrovich et al., 1979; Feenstra and Coburn, 1986; Pantazidou, 1991].
All of these factors relate to the soil volume contacted and thus contaminated by the DNAPL, the
preponderance of a distinct separate phase, and the likely modes of transport.
DNAPL infiltration in the vadose zone can attain significant depths rapidly. In laboratory
experiments using relatively homogeneous sands (K~1-2x10~4 m/s, n=50%), PCE traversed 100 cm of
vadose zone at residual water saturation, -40 cm of capillary fringe, and approximately 30 cm of the
saturated zone in 4 hours, see Figure 2.1.5. In clean, water-saturated sands (k~10~6 cm2), the vertical
migration of pure TCE and PCE has been observed to be 1 to 4 cm/s under a small DNAPL driving
gradient [Schwille, 1988; Kueperand Frind, 1991 a]. In other laboratory experiments using saturated sands,
pure PCE moved around soil heterogeneities and penetrated up to depths of 35 cm in approximately 5.25
minutes [Kueperand Frind, 1991a,b]. Infield experiments using the same DNAPL volume, the penetration
depth was greater for the slower release and smaller application area [Poulsen and Kueper, 1992].
2.2
PHYSICAL AND CHEMICAL PROPERTIES OF NAPLS
Physical and chemical properties of hydrocarbons and their partitioning are influenced by the
ambient pressure and temperature, and the type and quantity of other species in the system. For a single-
component NAPL, the aqueous solubility and the NAPL properties at 20-25°C are appropriate for analysis
and modeling purposes. In contrast, a multicomponent NAPL acquires properties reflecting the aggregated
contribution of all hydrocarbons comprising it, and the former assumption is either completely inappropriate
or only provides a first-order approximation. Hence, such values should be used judiciously, as discussed
later.
r
Y:
CH2C12
hu = 100 cm
1h 20 mm
h. = 70 cm
Figure 2.1.5 Infiltration of PCE (top) and DCM (bottom) in static glass trough experiments. Volume of
DNAPL released: 10 liters. Elapsed time and heights of unsaturated (hu), saturated (hs)
and capillary fringe (hc) shown [Schwille, 1988].
12
-------
The physical properties of a multicomponent NAPL mass located in the subsurface also evolve over
time. The more soluble, volatile, and biodegradable components of the NAPL mass can be more rapidly
depleted with time, leaving the more viscous, sorptive, and less volatile components behind [Geller, 1990;
Hunt et al., 1988a; Mercer and Cohen, 1990; Sitar et al., 1992]. As a result, the characteristic properties
such as viscosity, density, and interfacial tension of the NAPL mass are likely to change.
2.2.1 Aqueous Solubility
The aqueous solubility limit, Ciwso|, refers to the maximum dissolved concentration of that
compound in pure water. Sparingly soluble hydrocarbons have aqueous solubilities on the order of less
than 10 mg/l. When such hydrocarbons are present at quantities exceeding the solubility limit, a second
liquid phase forms consisting of nearly pure hydrocarbon with trace quantities of water. This phase is
commonly referred to as the "NAPL" or as the "immiscible" or "separate" phase. The Ciwso| is an
expression of the chemical equilibrium in the water-NAPL binary liquid system, and its exact value may be
influenced by hydrocarbon molecular structure, degree of halogenation, polarity, pH, temperature, and
pressure.
Rather than existing as a single continuous phase in porous media, the NAPL is more likely to
exist predominantly as discontinuous droplets, or "ganglia," which are very difficult to locate in the
subsurface. Once emplaced, single- or multicomponent NAPL ganglia slowly dissolve into the adjacent
pore water, serving as long-term sources of contamination. The concentration of each hydrocarbon in the
aqueous phase then depends on its solubility limit, its mole fraction in the NAPL, and on mass transfer
limitations [Sitar et al., 1987, 1992].
2.2.2
Fluid density, defined as the mass per unit volume, is a useful parameter for estimating the
potential for the downward migration of an NAPL in the subsurface. An analogous term, specific gravity
(SG) is often used to describe fluid density. The NAPL specific gravity (SG|) is the ratio of NAPL density
to that of water. An NAPL with an SG less than unity is termed an "LNAPL" (lighter-than-water NAPL) or
a "floater" because the LNAPL normally resides at or above the water table in the subsurface. However,
LNAPL ganglia can be also trapped below a fluctuating water table, as discussed later.
Generally, the greater the hydrocarbon molecular weight and degree of halogenation (CI', Br"
substitutions, etc.), the denser the NAPL will be, as can be seen in Table 2.1. The downward vertical
mobility of DNAPL increases with increasing molecular weight and density. Density differences as small
as ~1% influence fluid movement in the subsurface, and most DNAPLs possess densities 10-50% greater
than water [Mackay et al., 1985]. Density is often a strong function of temperature, and a DNAPL may be
effectively changed to an LNAPL by increasing temperature.
2.2.3 Interfacial Tension
Interfacial tension develops at the phase (solid, polar liquid, non-polar liquid, gas) boundaries and
refers to the surface energy that develops at the physical interfaces between immiscible phases, such as
the air-water interface or between polar and non-polar liquids (e.g., water-DNAPL). Interfacial tension has
the units of force per unit length and it is a measure of the deformability of the interfacial contact. In
general, the water-NAPL interfacial tension increases with the degree of halogenation (see Table 2.1); it
decreases with increasing temperature and it is affected by pH, and gases and surfactants [Mercer and
Cohen, 1990].
13
-------
Interfacial tension is a controlling factor in the prediction of ganglia mobilization under a variety of
conditions. The inference made from Equation (2) is that increasing interfacial tension results in a greater
degree of emplacement and lesser DNAPL mobility. Ganglia mobilization can be facilitated by reducing
or eliminating the interfacial tension through addition of surfactants and hydrophilic cosolvents, and by
increasing the temperature.
2.2.4 Wettabilitv and Spreading
Wettability refers to the preferential spreading of one fluid over the solid surfaces in a multiphase
system. In porous media, the wetting fluid has the tendency to spread and occupy the smaller pore spaces
and channels, and flow of the non-wetting fluids is generally limited to the larger pore flow channels
[Schwille, 1988; Mercer and Cohen, 1990]. Wettability is a function of the intertacial tension and it is
normally indexed to the contact angle between two immiscible fluid phases at the solid surface as shown
in Figure 2.2.1. The contact angle is determined from Young's Equation [Adamson, 1982]:
cos<|) = —— — (9)
nw
where d> is the contact angle, and on<: and a.... are the NAPL-solid and water-solid interfacial tensions,
n rt
respectively. Porous media systems are usually described as: water wet if <|> < 70 , NAPL-wet if <|> > 110 ,
and neutral if = 70°-110° [Anderson, 1986a]. Others use <|> < 90° and <|> > 90° to define water-wet and
NAPL-wet systems, respectively [Wardlaw, 1982; Villaume, 1985]. In most natural porous media systems,
preferential wettability decreases in the order of H2O, NAPL and air, unless the medium has been
previously contacted by the NAPL. Wettability is affected by mineralogy, presence of organic matter,
presence of surfactants, NAPL composition, pore water chemistry, and saturation history [Mercer and
Cohen, 1990]. Some crystalline compounds such as dolomite, graphite, limestone, sulfides, sulfur, talc,
and talc-like silicates, may be preferentially NAPL-wet [Craig, 1971; Anderson, 1986a]. Table 2.3 presents
the contact angles of several DNAPL's in natural soils [Arthur D. Little, Inc., 1981]. Contact angles studies
have shown that NAPL wettability increases with time [Craig, 1971]. Hysteresis between the contact angles
of advancing NAPL in initially water-wet medium and of receding NAPL from NAPL-contaminated medium
is a well known phenomenon [Villaume, 1985; Morrow, 1990]. Differences on the order of 5 to 10 degrees
are common. An extensive review of the various factors influencing wettability, its measurement, and its
effect on capillary pressures, relative permeability, residual saturation and NAPL recovery is presented in
Anderson (1986a,b,c; and 1987a,b,c).
NAPL-WET
WATER-WET
Hz° NAPL
V/////////////SOLID //////7/
Figure 2.2.1 Relationship between contact angle (<|>) and wettability [Mercer and Cohen, 1990].
14
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TABLE 2.3 EXPERIMENTALLY MEASURED CONTACT ANGLES OF DNAPLS IN DIFFERENT SOIL
TYPES [adapted from Arthur D. Little, Inc., 1981].
DNAPL
Tetrachloroethene
Tetrachloroethene
1 ,2,4-Trichlorobenzene
1 ,2,4-Trichlorobenzene
Hexachlorobutadiene
Hexachlorocyclopentadiene
2,6-Dichlorotoluene
4-Chlorobenzotri fluoride
Carbon tetrachloride
Chlorobenzene
Chloroform
S-Area DNAPL
S-Area DNAPL
S-Area DNAPL
S-Area DNAPL
S-Area DNAPL
S-Area DNAPL
Tetrachloroethene
Tetrachloroethene
Tetrachloroethene
S-Area DNAPL
S-Area DNAPL
S-Area DNAPL
S-Area DNAPL
S-Area DNAPL
S-Area DNAPL
S-Area DNAPL
S-Area DNAPL
S-Area DNAPL
S-Area DNAPL with solvents
Substrate
clay
clay
clay
clay
clay
clay
clay
clay
clay
clay
clay
clay
clay
clay
fine sand and silt
clayey till (30-40% clay)
Ottawa fine to coarse sand
Ottawa fine to coarse sand
Lockport Dolomite
Lockport Dolomite
Lockport Dolomite
Lockport Dolomite
NAPL-contammated fine sand
soils with vegetative matter
paper
wood
cotton cloth
stainless steel
clay
clay
Medium
APL
air
APL
air
water
water
water
water
water
water
water
APL
water
air
water
water
water
water
water
air
water
air
APL
water
water
water
water
water
water (SA)
water
f (°)
23-48
153-168
28-38
153
32-48
32-41
30-38
30-52
27-31
27-34
29-31
21-54
20-37
170-171
30-40
20-37
33-50
33-45
16-21
171
16-19
164-169
45-105
50-122
31
34-37
31-33
131-154
25-54
15-45
Adsorbed S-Area (New York, USA) chemicals were detected on some of the clay samples APL refers to aqueous phase liquids
(water containing dissolved chemicals) SA refers to surface-active agents (Tide® and Alconox®) which were added to the water.
S-Area DNAPL is comprised primarily of tetrachlorobenzene, trichlorobenzenes, tetrachloroethene, hexachlorocyclopentadiene, and
octachlorocyclopentene.
Usually, the spreading of NAPLs is not prevalent in the saturated zone because the soil is typically
water-wet. However, since transport in the subsurface occurs in both the vadose and saturated zones and
the vadose zone is a three-phase system (NAPL-water-gas), NAPL spreading and migration in the vadose
zone will almost certainly contribute to the degree of contamination occurring in the saturated zone. The
spreading behavior of a NAPL in the water-NAPL-gas system can be estimated from its spreading
coefficient, I. [Adamson, 1982; Wilson et al., 1990]:
15
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where aaw, oow and oao are the air-water, NAPL-water and NAPL-air interfacial tensions, respectively.
Negative spreading coefficients indicate the tendency of the NAPL to "bead" at the air-water interface. On
the other hand, positive Z indicates the potential for the formation of an organic liquid film at the air-water
interface [Wilson et al., 1990]. Laboratory studies using porous media have indicated that these films,
because of their continuity, allow the DNAPL to migrate even though DNAPL ganglia may appear
discontinuous and isolated [Wilson et al., 1990]. Hence DNAPL films can potentially lead to the net
migration of the DNAPL in the downgradient direction. Other implications of film flow have yet to be
assessed, such as the effect of films on mass transfer processes in the subsurface.
2.2.5 Viscosity
The absolute, or dynamic viscosity of a fluid is defined as its resistance to flow. In general,
absolute viscosity increases with increasing molecular complexity, molecular size and polarity, but
decreases with increasing temperature and degree of halogenation. The kinematic viscosity is one of the
better indicators of NAPL mobility since it incorporates both viscous and density effects. The infiltration
rates of NAPLs with low kinematic viscosities are expected to exceed those with high kinematic viscosities
[Schwille, 1984; Pantazidou, 1991; Pantazidou and Sitar, 1992]. For example, it has been shown that
certain LNAPLs and DNAPLs comprised of polynuclear aromatics are 2-10 times less mobile than water,
while DNAPLs comprised of aliphatics are 1.5-3.0 times more mobile than water in porous media [Schwille,
1981, 1988]. However, Figure 2.1.5 shows that in identical DNAPL releases of PCE (v = 0.54 mm2/s) and
DCM (dichloromethane, v = 0.32 mm2/s), the PCE not only migrated faster, but it also remained confined
to a zone with a smaller radius; and thus, the quantity of PCE retained by the soil was smaller. Schwille
attributed the slower penetration and spreading of DCM to possible vaporization of DCM owing to its high
vapor pressure; but more plausible explanations include that for the same soil, the Bond Number, NB, of
PCE is greater than that of DCM, and that some heterogeneity may have existed in the DCM column.
2.2.6 Vapor Pressure and Henry's Law Constant
The vapor pressure is the pressure exerted by the vapor when it is in equilibrium with its pure solid
or liquid phase at a specified temperature (usually 20°C). The vapor pressure of a hydrocarbon represents
an "air solubility limit" expressed as pressure, not concentration, and is therefore analogous to the aqueous
solubility limit. Vapor pressure, and thus volatility, generally increase with increasing hydrocarbon
aliphaticity and degree of halogenation. The Henry's Law Constant (KH) is an air/water partitioning constant
which is defined as the ratio of the hydrocarbon vapor pressure (atm) to its molar aqueous solubility limit
(mole/m3) at a reference temperature of 20°C or 25°C.
Hydrocarbons are usually classified as "volatile" if their vapor pressures at 20°C are greater than
1 mm Hg (1.31x10~3 atm), and as "semi-volatile" if their vapor pressures are between 10~10 mm Hg to 1
mm Hg (1.31x10~13to 1.31x10"3 atm) [USEPA, 1992a]. Hydrocarbons with vapor pressures greater than
0.5 mm Hg can be expected to significantly volatilize from leaking underground storage tanks (USTs)
[Bennedsen et al., 1985], while hydrocarbons with vapor pressures less than 10 7 mm Hg are not expected
to significantly volatilize [Dragun, 1988].
In subsurface applications, these rule-of-thumb conventions based on vapor pressure are useful
indicators in cases in which pure products are in contact with the gaseous phase in the vadose zone. For
NAPL ganglia emplaced below the water table or for a dissolved hydrocarbon plume, the hydrocarbon
volatility is more dependent on its aqueous concentration because the NAPL is no longer in direct contact
16
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with the vadose zone. In this situation, it is much more realistic to evaluate hydrocarbon volatility on the
basis of Henry's constant, KH. Hydrocarbons having KH greater than 10"5 atm m3/mole can be successfully
air stripped from ground water [Nirmalakhandan et al., 1987; Speece et al., 1987], or in-situ air-sparged
from porous media [Brown et al., 1991]
2.2.7 Octanol-Water Partitioning Coefficient (Kow)
The octanol-water partitioning coefficient (Kow) is used to estimate the hydrophobicity and sorptive
tendencies of hydrocarbons. Octanol is used to simulate an immiscible organic phase such as soil organic
matter (section 2.3.4). Most often, Kow is used to correlate hydrocarbon sorption in aquifer media. The
Kow is the ratio of the hydrocarbon concentration in the octanol (C10H22) phase to the aqueous phase:
C = (K )~]C- (H)
iw \ ow' 10
where CIW (g hydrocarbon/I H2O) and CIO (g hydrocarbon/I C10H22) are the hydrocarbon concentrations in
the water and octanol phases, respectively For convenience, Kow is often reported in logarithmic form
(logKow) because representative values of Kow for the class of immiscible hydrocarbons which are of
environmental concern span several orders of magnitude (Table 2.1). Negative logKow values indicate
hydrophilicity, that is, preference for the aqueous phase. Conversely, positive logKow value indicate
hydrophobicity and the hydrocarbon's preference to form separate phases, sorb strongly to solids, or
potentially volatilize.
2.2.8 Boiling Point
The boiling point (b.p.) is the temperature at which the vapor pressure of the liquid equals the
pressure of gases above the liquid, causing bubbles of vapor to form throughout the liquid. This
temperature varies with pressure, and the normal boiling point is given at a reference pressure of 1 atm.
The boiling point provides a measure of the volatility of the fluid; low boiling point (b.p.<100°C) and high
boiling point (b.p.>100°C) liquids are classified according to the normal boiling point of water (100°C). This
convention is used as a benchmark for comparisons, such as required energy input and mass transfer
between the contaminant and water, when considering stripping and thermal process applications.
2.2.9 Dielectric Constant
The dielectric constant (e) is the ratio of the permitivity of a medium to that of a vacuum, and is a
reflection of the ability of the medium to interact electrostatically and conduct an electrical current. A fluid
with a low dielectric constant (most sparingly soluble hydrocarbons) does not respond well to an applied
electric field: it acts as an insulator. Behavior of this type is important when considering the interactions
of clayey soils and pore fluids which contain hydrocarbons Because hydrocarbons do not align themselves
in the electrostatic field (diffuse double layer) generated by the surface charge of the clay particles, the
diffuse double layer effectively shrinks allowing interparticle interactions to occur. These particle
interactions lead to clay flocculation and increase hydraulic conductivity. Certain technologies (see sections
3.3, 3.7.3) make use of the differences between dielectric constants of NAPLs (e<10) and water (e~80) to
aid in detection and/or cleanup.
2.3 PHYSICAL PROPERTIES OF SUBSURFACE SYSTEMS
Darcy's Law for multiphase flow and the capillary pressure expression (Equations 1 and 2) indicate
which parameters govern the mobility of the DNAPL. These parameters are: porosity, permeability, relative
permeability, and mean pore throat diameter.
17
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2.3.1 Porosity
The porosity, <|>, of the aquifer media is defined as the dimensionless ratio of the pore volume to
the total bulk volume of aquifer media. Fluid flow, however, may only occur in a portion of the total pore
space, since preferential flow may occur through macropores, cracks and other features, and flow will not
occur in dead end pores [Anderson et al., 1985a,b; Bowders, 1985; Connor et al., 1989]. Hence, in the
case of fissured clays, for example, the effective porosity may be considerably smaller than the total
porosity.
Even in soils without macroscopic discontinuities, the gradation and spatial distribution of pore and
pore-throat sizes directly affect the distribution of capillary pressures and, thus, DNAPL flow and
emplacement [Wilson and Conrad, 1984; Sitar et al., 1987]. Emplacement is thought to occur by two
mechanisms: snap-off and by-passing [Mohanty et al., 1980, Chatzis et al., 1983]. Both mechanisms are
observed in varying proportions depending on the characteristics of the porous media. Snap-off, that is,
the formation of discrete intrapore ganglia, is favored when the aspect ratio of the pore body to the pore
throat is large. Low aspect ratios and heterogeneities usually promote by-passing, that is, the formation
of large clusters or interpore ganglia [Chatzis et al., 1983; Wilson and Conrad, 1984]. Since the
measurement of the actual pore and pore throat size distribution is difficult at best, empirical relationships
have been developed to obtain an estimate of the average pore throat size from the mean soil grain
diameter [Villaume, 1985].
2.3.2 Permeability
The permeability, k, of natural soils, sediments, and rocks spans approximately 13 orders of
magnitude [Freeze and Cherry, 1979]. The factors that affect the permeability of the soil include the soil
porosity and gradation, the spatial distribution of the soil grains and pore throat sizes, and the scales over
which they vary. Permeability is related to soil gradation through the square of the mean soil grain
diameter and a constant of proportionality which incorporates such factors as particle shape, packing, and
the shape of the gradation curve [Hubbert, 1940]. Macroscopically, the changes in grain size and gradation
are manifested as stratigraphy and soil heterogeneity, two of the most important factors affecting the
mobility and the distribution of the DNAPL.
Of particular importance is the presence of clay minerals, their activity, and the proportion of the
clay fraction in the media. Permeability decreases with increasing clay fraction, and permeabilities of pure
and compacted clays can range from 10~11 to 10"15 cm2 [Freeze and Cherry, 1979]. Therefore, NAPL
migration is not likely to occur in aquitards, and liners have been commonly thought to serve as adequate
barriers against DNAPL migration. However, hydrocarbons can traverse clayey aquitards via diffusion, and
NAPLs can migrate through new or pre-existing fissures [Mitchell and Madsen, 1987; Feenstra and Cherry,
1988]. These modes of transport are most important in the vicinity of "pools" and "lenses" of DNAPL.
Pools of DNAPL tend to accumulate in the depressions along relatively impermeable layers such
as clayey aquitards. Large diffusive fluxes resulting from simple Fickian diffusion of concentrated and
mobile contaminants across a competent, impermeable clay soil (k<5x10~13 cm2) have been observed
[Johnson et al., 1989]. Thus, diffusion into clayey aquitards and more permeable fine grained soil layers
should be expected.
At locations where the DNAPL pool intersects pre-existing cracks and discontinuities in clayey
aquitards or fissured clays (or fractured rock), penetration of DNAPL into these macroscopic features has
been documented. The ability of a DNAPL to penetrate into fractures depends on the DNAPL driving
gradient, the DNAPL-water interfacial tension, and aperture of the fissure [Kueper and McWhorter, 1991;
Middleton et al., 1992]. Modeling studies indicate that a pool thicknesses on the order of 0.1 to 1.0 meters
18
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may be sufficient to permit the flow of PCE into fissures with apertures of 10 urn to 50 u.m, if the top of the
pool exists under imbibition conditions [Kueper and McWhorter, 1991].
2.3.3 Clay-Pore Fluid Interactions '
Due to the surface charge and large specific surface of clay particles, clayey soils are susceptible
to changes in pore water chemistry. Studies have revealed that concentrated organic compounds can
influence the ability of clayey soils to act as barriers against organic chemical transport by causing hydraulic
conductivity changes of up to several orders of magnitude [Acar et al., 1985a,b, Anderson et al., 1985a,b;
Bowders, 1985; Fernandez and Quigley, 1985; Mitchell and Madsen, 1987]. Good summaries of the
performance of clay soils [Daniel et al., 1985; Mitchell and Madsen, 1987] and clay liners [Mitchell and
Jaber, 1990] with respect to contaminant and clay type, testing conditions, and testing apparatus are
available.
Clay-organic compound interactions such as sorption, intercalation, and cation exchange may result
in clay fabric changes, cracking, and shrinking [Mitchell and Madsen, 1987]. For example, clay swelling
in pore throats of granular (sandy) soils can also lead to DNAPL trapping by the bypassing mechanism
(sections 2.1.2, 2.3.1). Sorption of polar organic compounds (alcohols, ketones) and ionizable
hydrocarbons such as phenols [Lee et al., 1990], quinoline [Zachara et al., 1987, 1988], and nitrogen
heterocyclic compounds [Zachara et al., 1987] has been observed to be a function of porewater pH
conditions and their acidity constants (pKa's). Hydrocarbon sorption to chemically altered clay soils has
also been observed [Estes et al., 1988]. However, most dense sparingly soluble hydrocarbons are both
nonionizable and often non-polar, so their sorption onto clays is small compared to their partitioning into
soil organic matter. Of primary concern is desiccation of clays that generally results from differences in
dielectric constants (e) between pore fluids composed of sparingly soluble hydrocarbons (e<10) and water
(e=80.4). As a result, flow of non-polar hydrocarbons has been observed in interconnected cracks and
macropores [Anderson et al., 1985a,b; Bowders, 1985].
More importantly, the migration of hydrocarbons through clayey soils may be enhanced in the
presence of polar and hydrophilic organic compounds. Experiments in which a water saturated clay was
sequentially permeated with water, ethanol, and benzene (in that order), revealed that the hydraulic
conductivity of the clay could be increased by approximately four orders of magnitude [Fernandez and
Quigley, 1985], Since water and ethanol are completely miscible, ethanol displaced most (75% wt.) of the
water from the clay minipores and macropores. Benzene, which is completely miscible with ethanol, was
then used to displace the ethanol. Approximately half (50% wt.) of the ethanol was displaced, and
approximately 30% (wt.) benzene composed the total pore fluid In separate experiments in which no
ethanol was used, benzene displacement of water resulted in only 8% (wt.) residual benzene in the total
pore fluid. Hence, the net effect of ethanol was to quadruple the residual saturation of benzene in the clay.
Similar trends were observed in sequential permeation experiments using ethanol and xylene and
cyclohexane [Fernandez and Quigley, 1985]. Since the dielectric constants of carbon tetrachloride (e=2.2)
and TCE (e=3.4) are comparable to that of benzene (e=2.2), analogous behavior appears likely, especially,
if density effects are also taken into consideration.
2.3.4 Organic Matter
The portion of a sediment or soil that consists of organic detritus (waste material such as dead
microorganisms, slimes, plant fibres, etc.) is referred to as the natural organic matter. The fraction of
organic matter (foc) is important with respect to hydrocarbon partitioning and sorption within the soil matrix.
Karickhoff (1984) provides a detailed review of sorption in aquatic systems, and this work contains many
useful references.
19
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The sorption isotherms of hydrophobic organic compounds to natural sediments have been shown
to be linear, if their dissolved concentration is less than 10 5 M, or less than half of their solubility limit,
whichever is lower [Karickhoff, et al., 1979]. Experimental observations encompassing a wide variety of
soils indicate that this partitioning is directly related to the foc, and the octanol-water partitioning coefficient,
Kow [Karickhoff et al., 1979, Chiou et al., 1983], For hydrocarbons having solubilities between 0.5-1800
mg/l and for soil particles < SOujm, sorption was observed to be reversible, independent of grain size, and
there was no competitive adsorption between multiple contaminants. An empirical relationship developed
by Karickhoff et al. [1979] is stated as:
where Kd is the distribution coefficient of the compound between the aqueous and soil phases and has the
units of (g hydrocarbon/g soil)/(g hydrocarbon/I H2O), and
-------
The usage of a, and p as the polar (water) and nonpolar (NAPL) phases in equation (13) is
deliberate. This usage is taken from the solvophobic theory which recognizes that NAPL-water equilibrium
is a subset of the larger, more general class of nonpolar-polar equilibrium interactions which exhibit similar
behavior [Smanoglu and Abdulnur, 1965; Sinanoglu, 1968]. "Nonpolar-polar" equilibrium terminology better
accommodates systems in which compounds such as acids, alcohols, caustics, ketones, and amines are
the polar solvents instead of water. This terminology also seems better suited for concentrated systems
in which the composition and properties of the polar and non-polar phases are significantly changing with
time, as would be expected in in-situ acid, caustic, surfactant, and hydrophilic cosolvent soil washing
applications. For other less concentrated ground-water systems, it is convenient to refer to the polar phase
in terms of the principal solvent, water.
2.4.1 Mujticomponent NAPLs
In the discussion of multicomponent NAPLs we assume that the presence of other dissolved
chemical species (metals, cosolvents, surfactants) in the system will, in general, not affect partitioning of
the individual compounds. This assumption allows the multicomponent NAPL to be treated as an ideal
liquid phase. Similarly, the aqueous phase, is also considered as an ideal liquid phase in that all solutes
are considered to be infinitely dilute; and therefore, they do not interact. Several of the constants
mentioned previously in this chapter including the aqueous solubility (C,WSO|), octanol-water partitioning
coefficient (Kow), and the sorption coefficient (Kd) are also based on the same assumptions.
Cancelling the reference fugacities in Equation (13) from each liquid phase in the water-NAPL
binary liquid system yields the expression:
Y' Y- (14)
in i in
or,
(
Y,,,
(15)
where the subscripts w and n denote the water and NAPL respectively, and the mole fractions in each
phase sum to unity. For fluid phase equilibria of nonelectrolytes, Roault's Law is used for activity coefficient
normalization. This convention states that as x -> 1, y -> 1 [Prausnitz et al., 1986]. As the
multicomponent NAPL is considered as an ideal liquid phase, a further simplification is made that all activity
coefficients in the NAPL equal unity (y|n =1). In order to obtain the aqueous solubility limit of a
hydrocarbon, an excess quantity of single-component NAPL is contacted with water. Under these
conditions, the right hand side of equation (14) becomes unity, and upon rearrangement,
where x(W so, is the aqueous solubility limit of the hydrocarbon expressed as a mole fraction. The aqueous
solubility limits of some of the compounds appear in Table 2.1. The infinite dilution assumption for the
aqueous phase is reasonable when y|W > 1000 [Fu and Luthy, 1986a], and this condition is satisfied in the
absence of solubility enhancers. If ylw < 1000, the mole fraction solubility calculation must account for the
appreciable solubility of the compound and that the mole fraction does not approach unity [Fu and Luthy,
1986a]. Substituting ylw = (1 / x|W S0|) back into equation (15) provides
21
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or, in terms of aqueous concentrations (mass hydrocarbon/vol. H2O),
Ciw = (Ciw,sol)xin (18)
Equation (18) states that the aqueous concentration of a hydrocarbon is directly proportional to its mole
fraction in the NAPL and its solubility limit. Thus, Equation (18) implies that dissolved concentrations
occurring below their respective solubility limits do not preclude the existence of multicomponent NAPL
pools [Sitar et al., 1992]. Comparison of equations (15) and (18) reveals that CIWSO| is none other than
the ratio of activity coefficients, or similarly, the partitioning coefficient, between the'water and NAPL. The
octanol-water partitioning coefficient (Kow) and the soil organic matter sorption (Kd) coefficients can be
similarly obtained. These partitioning coefficients can be respectively expressed on a mass basis by
equating fugacity between the phases, or:
CIW = (CiWt5ol)xin = (KowylCio = (K(,ylCis = ... d9)
where CIO and CIS are the hydrocarbon concentrations in the octanol and solid (i.e., soil organic matter)
phases, respectively. Recall that Kd can be expressed as equation (12) [Karickhoff et al., 1979; Karickhoff
1981, 1984]. These three partitioning coefficients can be treated as constants, as long as the solutes in
the aqueous phase are infinitely dilute. This may not be the case when surfactants and hydrophilic
cosolvents are present in the ground-water environment.
On an equilibrium basis, and considering only one immiscible hydrocarbon, the extent to which the
NAPL will be present is limited by the mass of available hydrocarbon [Mackay 1979; Mackay and Paterson,
1981; Mackay and Shiu, 1992]. For example, a NAPL will not form in a well mixed, homogeneous, non-
reactive system (i.e., air, water, soil, sediment, etc.) until each preexisting phase is saturated with respect
to the given compound. In sub-saturated systems, the hydrocarbon will equilibrate amongst the preexisting
phases according to fugacity. Upon saturation, further addition of hydrocarbon results in the formation of
a NAPL into which all subsequent hydrocarbon accumulates. Now, if other immiscible hydrocarbons are
then added to the system, they will partition among all phases (including the NAPL) according to fugacity.
The original hydrocarbon will re-equilibrate among the phases according to changes in its NAPL mole
fraction. Hence, if reasonable estimates of porosity, soil organic matter, and water saturation are made,
soil concentration data can be used to obtain a first-order estimate of the amount of NAPL (and its
composition) in the soil sample [Sitar et al., 1992] In reality, subsurface systems are heterogeneous,
reactive, and mass transfer limited, so even modest addition of a sparingly soluble organic compound may
result in NAPL formation.
2.4.2 Surfactants, Cosolvents and Multicomponent NAPLs
The occurrence of hydrocarbons above their solubility limits in ground water has been observed
in the presence of surfactants (natural or synthetic) and hydrophilic organic compounds (cosolvents).
Surfactants are surface active compounds which accumulate at interphase contacts and increase NAPL
solubility primarily by formation of micelles, composite NAPL-surfactant psuedophases consisting of a NAPL
interior and surfactant exterior which enable enhanced water solubility of the NAPL [Adamson, 1982;
Fountain et al., 1990]. Natural organic matter such as humic and fulvic acids [Chiou et al., 1986, 1987;
Abdul et al., 1990a,b] and dissolved organic matter [Kan and Tomson, 1986] may facilitate transport of
22
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hydrocarbons in a micellar and surfactant-like fashion. In contrast, cosolvents do not form micelles per se.
Because of their mutual solubility with both water and immiscible hydrocarbons, cosolvents partition into
both water and NAPL thereby altering phase properties (i.e., polarity, dielectric constant) such that they
become more similar. The net effect is that the electrostatic associations between water and sparingly
soluble hydrocarbons are increased, and this is manifested by increased solubility.
In terms of equilibria, surfactants and cosolvents are seen to alter the activity coefficients ratios
between phases. Depending on their concentrations, surfactants [Kile and Chiou, 1989; Fountain et al.,
1991] and cosolvents [S0renson and Arlt, 1980; Groves, 1988] have been shown to increase the aqueous
solubility of hydrocarbons by several orders of magnitude. Surfactants and cosolvents have been shown
to strongly desorb hydrocarbons from soil organic matter and to enhance hydrocarbon partitioning into the
aqueous (polar) phase [Fu and Luthy, 1986b; Woodburn et al, 1986; Walters and Guiseppi-Elie, 1988,
Zachara et al., 1988; Abdul et al., 1990a,b; Rao et al., 1990]. More specific information on surfactants and
cosolvents is presented in Section 4.0.
Whereas the activity coefficients in the NAPL and aqueous phase were formerly treated as
constants, the activity coefficients in surfactant and cosolvent systems are seen to vary with chemical
species and total system composition. Hence, the limits of the activity coefficient must be determined from
partitioning data, phase diagrams, or suitable equilibria models which can account for nonideal behavior.
The solubility enhancement of surfactants is normally measured experimentally and presented in
log-linear format. There are currently several approaches to modeling the effects of cosolvents in
multiphase multicomponent systems, including: log linear [Fu and Luthy, 1986a,b], excess free energy,
molecular surface area (MSA), and group contribution models [Fu and Luthy, 1986a; Prausnitzetal., 1986].
The merits and disadvantages of these approaches are discussed elsewhere [Fu and Luthy, 1986a]. The
log linear and excess free energy approaches are extremely accurate, but their infinite dilution activity
coefficients must be calculated by group contribution models or equivalent methods when experimental data
are lacking. The MSA approach receives limited use because few parameter data are available.
Experimental results are not needed for group contribution models such as NRTL, UNIQUAC and
UNIFAC, which can accurately approximate (within a factor of 2) compound solubilities in multicomponent
multiphase systems [Prausnitz et al., 1986]. These models employ interaction parameters based on
molecular structure and functional groups, and large amounts of data are available on group interaction
parameters [Gmehling et ai., 1982]. UNIFAC has enjoyed widespread use, but information is not always
available for interaction parameters which are specific to certain compounds of environmental concern.
2.5 UNSATURATED AND SATURATED ZONE TRANSPORT MECHANISMS
2.5.1 Unsaturated Zone Transport
Although the emphasis in this report is on the saturated zone, it is necessary to understand
transport through the vadose zone because most contaminant releases typically occur at the ground
surface or in the vadose zone. The conceptual framework presented here is summarized from several
sources [Schwille, 1967, 1981, 1984, 1988; Feenstra and Cherry, 1988, 1990; USEPA, 1992a].
Figure 2.5.1 shows a release of a fixed volume of DNAPL at the ground surface which becomes
immobilized in the vadose zone by capillary forces, adsorption, and by forming annular rings and surface
films before encountering the saturated zone. The immobilized, or residual, DNAPL is shown by the cross-
hatched region. Infiltrating rainwater passing through the residually contaminated soil zone may leach and
transport hydrocarbons to the water table [USEPA, 1992a]. Consequently, a dissolved plume is created
within the aquifer.
23
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Even if there is no rainwater infiltration through the residually contaminated zone, considerable
contamination of the aquifer may still develop [Hinchee and Reisinger, 1987; Schwille, 1988]. Vapor
transport of volatile organic compounds, VOCs, acts independently of the leaching mechanism. The
stippled areas within the unsaturated and saturated zones in Figure 2.5.1 indicate the regions affected by
hydrocarbons emanating from the soil zone contaminated by residual DNAPL.
Denser-than-air vapors emanating from the contaminated soil zone have been observed to migrate
downward to spread along the water table [Marrin and Thompson, 1987; Schwille, 1988], and this has been
the subject of several recent studies [Hunt et al., 1988a; Falta et al., 1989; Sleep and Sykes, 1989;
McClellan and Gillham, 1990: Mendoza and Frind, 1990a,b; Gierke et al., 1990, 1992]. At the water table,
vapors equilibrate with the aqueous phase according to Henry's law, and a dissolved plume develops. The
lateral spreading and diffusion of vapors at the water table can be significant as vapors may migrate below
buildings, parking lots, and other structures [Hinchee and Reisinger, 1987]. Ground-water contamination
occurring upgradient of the DNAPL source also has been observed as a result of vapor transport in the
vadose zone [Marrin and Thompson, 1987].
2.5.2 Saturated Zone Transport
Figure 2.5.2 shows a conceptual view of a release of a sufficient quantity of DNAPL to overcome
the capillary forces and the retention capacities of the vadose zone, capillary fringe and saturated zone.
As before, the cross-hatched area in Figure 2.5.2 shows the soil regions which are contaminated by the
residual DNAPL. While the vapor transport within the vadose zone is almost identical between Figures
2.5.1 and 2.5.2, the dissolved plume within the saturated zone is noticeably larger because the residual
DNAPL and DNAPL pools (layers or lenses) are in direct contact with ground water
Figure 2.5.2 also shows DNAPL pools which can form in the depressions of low permeability strata
such as silty or clayey lenses, aquitards, and bedrock. DNAPL pools can form when mobile DNAPL
encounters water-wet strata with very small pore throats that result in prohibitively large DNAPL entry
pressures. DNAPL accumulation up to saturations of 70-80% of the pore space may occur at the strata
interface. Because pooled DNAPL occurs in excess of its residual saturation, it should be considered
mobile because it may penetrate into preexisting fissures in the underlying clayey strata [Kueper and
McWhorter, 1991]. Preexisting fissures in naturally occurring clays are known to exist at substantial depths
below the water table [D'Astous et al., 1989; Sabourin, 1989]. DNAPL pools may also drain through newly
created fissures in a clayey strata resulting from clay desiccation, as already mentioned. Upgradient
DNAPL migration along horizontal strata is possible also, as shown in Figure 2.5.2. Migration is obviously
enhanced when the underlying strata is inclined.
DNAPL will also penetrate into bedrock fractures as shown in Figure 2.5 3, and the resulting
downward vertical migration of DNAPL occurring within the fractures may be extensive owing to the low
retention capacities of fractured bedrock systems. For example, based on laboratory experiments
employing planar fractures with a frequency of 5 fractures/meter and 0.2 mm apertures, Schwille (1988)
estimated DNAPL retention capacities on the order of 0.25 I hydrocarbon/m3 for weakly fractured rock
systems of moderate hydraulic conductivity. This value is an order of magnitude smaller than that for
unsaturated and saturated soils. Hence, once the DNAPL enters a fractured bedrock system, it can
contaminate a much larger region, given volumetric considerations.
While the influence of pronounced soil heterogeneities such as clay aquitards and bedrock on
DNAPL migration can be dramatic, it is important to note that even subtle hydraulic conductivity changes
in clean sands, on the order of a factor of 2, may be sufficient to cause preferential flow of DNAPL [Kueper
and Frind, 1991 a]. Site heterogeneities on that order is quite common, thus complicating the NAPL flow
and often making even the identification of residually contaminated soil zones and DNAPL pools difficult.
24
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ONAPL
AIR OR WATER-
FILLED PORE SPACE
TOP OF
CAPILLARY FRINGE
i±~ WATER TABLE
DISSOLVED
CHEMICAL
PLUME
GROUNDWATER
"FLOW
LOWER
PERMEABILfTY
STRATA
Figure 2.5.1
Schematic of the distribution of subsurface contamination emanating from residual DNAPL
source in the vadose zone [Feenstra and Cherry, 1990].
DNAPL
AIR OR WATER-
FILLED PORE SPACE
DISSOLVED
CHEMICAL
PLUME
TOP OF
CAPILLARY FRINGE
WATER TABLE
\
LOWER
PERMEABILITY
STRATA
DNAPL
VNAiER-FILLED
~ r?c SPACE
Figure 2.5.2 Schematic of the distribution of subsurface contamination emanating from residual DNAPL
sources in the vadose and water saturated zones, and DNAPL pools [Feenstra and Cherry,
1990],
25
-------
DNAPLHr
GROUNDWATER
FLOW
RESIDUAL DNAPL
DNAPL LAYERS
TOP OF
CAPILLARY FRIf
12. WATER TABLE
OVERBURDEN
FRACTURED
POROUS ROCK
DISSOLVED
CHEMICAL
IN FRACTURES
DNAPL
GROUN
FLOW
Figure 2.5.3 Schematic of fractured bedrock contamination resulting from mobile and pooled DNAPL
[Feenstra and Cherry, 1990].
2.6
ESTIMATION OF THE EXTENT OF SITE CONTAMINATION AND SITE CHARACTERIZATION
Injection, extraction, observation wells and other invasive monitoring, sampling and remedial
structures locally disrupt the stratigraphy and therefore introduce bias. Since testing procedures such as
well evacuation prior to ground-water sampling perturb resident pore fluids, further bias is introduced.
Sampling data themselves can often be misleading relative to the nature and extent of contamination,
principally in the delineation of DNAPL in the subsurface. Frequently the importance of multicomponent-
multiphase equilibria and interphase transport phenomena has been ignored or underestimated. This
section addresses some of these issues and the implicit difficulties regarding data interpretation and
estimation of contamination. Since we are ultimately interested in being able to evaluate the effectiveness
of the different technologies, these issues must be acknowledged in the technology assessment process.
2.6.1 Ground-water Samples
Ground-water samples are used to delineate the extent of contamination in the saturated zone and
to assess the success of remedial applications. Since much emphasis is placed on the "actual" values,
it is important to consider the various interpretations that can be made based on the data. For example,
both multiphase multicomponent equilibria and mass transfer limitations can have dramatic effects on
observed ground-water concentrations. To illustrate this point, the comparison between the partitioning
behavior of single- and multicomponent NAPLs is presented. Gasoline (multicomponent LNAPL) and its
components are arbitrarily chosen because gasoline spills are reasonably well studied and are helpful in
illuminating the limitations of subsurface sampling techniques and data.
Benzene, toluene, ethylbenzene and xylenes (BTEX) are the compounds of primary concern in
fresh and weathered gasoline. First, using Equation (18) and assuming that the NAPL is an ideal liquid
26
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phase, the aqueous phase equilibrium concentrations corresponding to a single-component NAPL
comprised of either pure benzene, toluene, or xylene, are 1800 mg/l (Cbwsol), 540 mg/l (Ctwso,), 185mg/l
(CXWSO|), respectively. Recall that the hydrocarbon mole fraction in a single-component NAPL is unity.
Now, if a bicomponent NAPL consists of a 50/50 mixture of benzene and toluene, the hydrocarbon
mole fraction in the NAPL is equal to 0.5. Thus, in comparison to the single-component NAPL, the
equilibrium concentration of benzene (Cbw) and toluene (C,J are effectively halved to 900 mg/l and 270
mg/l, respectively. Table 2.4 shows that for a multicomponent NAPL such as gasoline, the typical
equilibrium concentrations of benzene, toluene, and xylene amount to an approximate combined total of
47 mg/l. The total dissolved concentration (Cgws0|) of all gasoline-derived hydrocarbons is on the order of
100 mg/l. Hence, while BTX compounds are only minor constituents (21%) of gasoline, their combined
solubility accounts for almost 50% of the total dissolved hydrocarbons [Sitar et al., 1992]. The significance
of the low equilibrium concentrations of the individual components in the presence of a mulitcomponent
NAPL is that they are far below their respective solubility limits, and a convincing argument for the
existence of a NAPL cannot be made even though the multiphase equilibria is taken into account. On the
other hand. BTEX compounds may be detected in ground water above their solubility limits if gasoline
additives such as methanol, ethanol, and methyl tertiary-butyl ether (MTBE) are present [Mihelcic, 1990;
Barker et al., 1992; Kan et al., 1992], Since ground-water data is usually only obtained for regulated
compounds unless otherwise specified, it would be easy to overlook the presence of hydrophilic compounds
which enhance the solubilities of the regulated compounds.
TABLE 2.4 TYPICAL EQUILIBRIUM CONCENTRATIONS OF PURE AND GASOLINE-DERIVED BTX
COMPOUNDS [Sitar et al., 1992]
NAPL
BenzenG
Toluene
Xylene
% in Gasoline
23
83
103
Aqueous Solubility (g/m3)
Pure Phase
1800
540
185
Gasoline-Derived
11
24
12
Sub-solubility concentrations can also be caused by other phenomena. Consider the simplified flow
geometry depicted in Figure 2.6.1, in which ground water sweeps past a NAPL lens. As shown in Figure
2.6.2a for a single-component NAPL (i.e., TCE), the percent solubility of TCE in the aqueous phase is
dependent on such factors as the flow velocity and mixing, as indicated by the value of transverse
dispersivity (a). The response of a bi- or multicomponent NAPL lens is more complex [Geller, 1990].
For example, it is possible for the dissolved hydrocarbon concentrations and the mole fractions in
the bicomponent NAPL lens to temporally evolve as flowing ground water depletes the more soluble
component (benzene) from the NAPL lens more rapidly, as shown in Figure 2.6.3 The preferential removal
of soluble components from a NAPL mass via another flowing phase (aqueous or gas) is referred to as
selective dissolution or fractionation. As benzene is depleted from the leading edge of the NAPL lens, the
toluene mole fraction at the leading edge soon exceeds that of the downstream edge. From the
perspective of the aqueous phase, the equilibrium concentration of toluene at the leading edge of the NAPL
lens is greater than at the downstream end Consequently, toluene may dissolve into the aqueous phase
at the leading edge only to repartition back into the NAPL lens at the downstream end [Geller, 1990;
Adenekan, 1992]. in this way, both temporal and spatial concentration gradients within each liquid phase
may develop. In the case of multicomponent NAPLs such as gasoline, this process is very complex.
Other temporal and spatial gradients may develop as a result of the actual flow geometry (see
section 2.6.2 for discussion). Because of selective dissolution, the mole fraction ratios at the NAPL lens
27
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T
U
Y
V//////////////A
NAPL
Figure 2.6.1 Schematic of a simplified flow geometry of ground water sweeping past a NAPL lens [Sitar
et al., 1992].
interface may be different than in the NAPL interior Due to mass transfer limitations occurring within the
multicomponent NAPL, it therefore takes longer for aqueous concentrations to reach equilibrium with a
multi-component NAPL than with a single-component NAPL [Geller, 1990; Adenekan, 1992]. Hence, mass
transfer limitations will contribute to aqueous hydrocarbon concentrations appearing below their solubility
limits [Hunt et al., 1988b; Miller et al., 1990; Powers et al., 1991; Brusseau, 1992; Sitar et al., 1992]. Sub-
solubility limit aqueous NAPL concentrations may result from mixing and dilution within the porous media
and at the wellhead [Hunt et al., 1988b; Feenstra and Cherry 1988; Feenstra, 1990; Chiang et al., 1992],
and misleading interpretations of ground-water data may result if these factors are not given due
consideration.
2.6.2 Soil Gas Samples
Since the gaseous diffusion of volatile compounds is about four orders of magnitude greater than
in the liquid phase, soil gas monitoring and vapor extraction are useful in delineating vadose zone
contamination, and possible saturated zone contamination. For this reason, soil gas sampling of volatile
hydrocarbons has enjoyed widespread use.
Hydrocarbon vapors have been observed in excess of 100 meters from the NAPL source under
quiescent conditions [Marrin and Thompson, 1987; Hunt et al., 1988a]. Migration of this magnitude
indicates potential success for soil vapor extraction and for the identification of potential volatile DNAPL
"hotspots." However, while soil gas measurements can be used to delineate the contaminated region in
the unsaturated zone, they yield little information on the actual distribution of DNAPL within the saturated
zone [Marrin 1988], as illustrated in Figures 2.5.2 and 2.5.3. Also, nonvolatile hydrocarbons which
comprise DNAPLs will not be detected. Soil heterogeneity, density and quantity will limit the DNAPL
exposure, access and partitioning to the unsaturated zone. Furthermore, in terms of multiphase equilibria
alone, the problems associated with soil gas sampling are essentially identical to those of ground-water
sampling [Hunt et al., 1988a; Sitar et al., 1992] When coupled with mass transfer limitations, interpretation
of subsurface sampling data becomes very complex.
For example, as a result of continuous soil gas sampling or extraction, the tailing-off of hydrocarbon
concentrations is frequently observed. While this is commonly attributed to successful compound removal
from a NAPL lens, numerical simulations have shown that hydrocarbon concentrations in soil gas are a
28
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a = transverse dispersivity (m)
10" 10 -a 10" 1
hLOW VELOCITY (m/d)
CD
m -
CJ 1
Of
L_l
CL
10" 10" 10"
FLOW VELOCITY (m/d.
1 o
Figure 2.6.2 Computed average TCE concentrations ot ground water (a) and soil gas (b) sweeping past
a NAPL lens using a simplified geometry (Figure 2 6.1) [Sitar et al, 1992],
29
-------
1000
BOO
Z BOO
O
F
Z 400
LJ
O
O
200
1. I I 1 I I 1 T 1 1 1 1 1 1 1 1 1 1 1 1 1—
"*-..,. Computed Benzene
ODOOO Observed Benzene
Computed Toluene
ooooo Observed Toluene
0 2 4 6 8 10 12 14 IB 18 20 22
DISPLACED PORE VOLUMES
e a to 12 14 ie IB
DISPLACED PORE VOLUMES
20 22
Figure 2.6.3 Predicted and observed evolutions of: (top) aqueous hydrocarbon concentrations in
equilibrium with bicomponent NAPL; and (bottom) mole fractions of the bicomponent NAPL
[Sitaret al., 1992].
30
-------
function of the bulk phase sweep velocity [Hunt et al., 1988b; Sitar et al., 1992] and that NAPL removal is
limited by gas phase molecular diffusivity. Because liquid phase molecular diffusivities are about four
orders of magnitude less than gas phase diffusivities, gas phase removal is more efficient, as suggested
by Figure 2.6.2b.
Additionally, the tailing-off of concentrations of volatile compounds is often attributed to the
fractionation of the multicomponent NAPL mass, as shown in Figure 2.6.4. The gas phase concentrations
imply that the mole fractions of each compound in the NAPL mass are changing with time. Yet, field data
obtained by Johnson et al. (1990) reveal that the relative vapor concentrations of BTEX compounds
emanating from gasoline (multicomponent LNAPL) in the subsurface remain essentially constant over time,
see Figure 2.6.5. In this case, the soil gas concentrations suggest that the mole fractions in the gasoline
are constant and that fractionation may not be occurring.
The difference between the observed concentrations shown in Figure 2.6 4 and 2 6.5 can be
explained in terms of air flow geometry, i.e., flow-through and/or bypass drying mechanisms. The flow-
through drying mechanism is observed to occur when the emplaced NAPL is situated within a
homogeneous soil, or within a less permeable soil zone in which the permeability ratio between the
adjacent soil layers is less than 10:1 [Ho and Udell, 1992]. Flow-through air flow implies that the air is
flowing directly through the DNAPL contaminated soil zone, and compound removal is vapor solubility
controlled [Ho and Udell, 1992]. Fractionation of the leading edge (and to a lesser degree the periphery
edges) of the NAPL contaminated soil zone occurs as the more volatile components are preferentially
vaporized. Depending on the length of the contaminated soil zone in the direction of air flow, the relative
soil gas concentrations may not evolve (Figure 2.6.4), but may remain constant for some time (Figure
2.6.5), especially if contamination is extensive. Since DNAPL mass fractionation occurs in the direction of
air flow, the effluent soil gas concentrations will sequentially evolve as each component is removed, as
shown in Figure 2.6.4.
The bypass drying mechanism is observed to occur when the emplaced NAPL is situated within
a less permeable soil zone in which the permeability ratio between the adjacent soil layers is greater than
100:1 [Ho and Udell, 1992]. Bypassing air flow occurs through the more permeable zone, and the removal
of the compound is controlled by vapor diffusion within the less permeable soil zone, as shown in Figure
2.6.6. Under these conditions the liquid is stagnant, and although equilibrium is attained at the vapor-liquid
interface, fractionation of the DNAPL mass will not be manifested until the mass is nearly diminished or
unless liquid-phase mass transfer limitations exist.
Three distinct stages mark the evolution of the soil gas concentrations under bypassing conditions:
(1) an early, rapid decrease in the more volatile components arising from fractionation at the DNAPL
surface; (2) a period of quasi-steady state compound removal at near constant relative gas concentrations
as the DNAPL surface recedes further into the impermeable media; and (3), a long-term gradual decrease
in relative gas concentrations as the more volatile species are sequentially removed from the diminishing
DNAPL mass. These mechanisms have been observed on the laboratory scale [Ho and Udell, 1992].
Thus, the relative soil gas concentrations emanating from a multicomponent DNAPL mass may be
constant under two of the aforementioned conditions: (a) during flow-through vapor flow in an extensive
(long) DNAPL mass; and (b) during condition (2) of bypassing air flow. If the flow-through mechanism is
operating, the relative soil gas ratios will approximate the relative soil gas solubility limit ratios, whereas the
relative soil gas concentrations should be close to the initial compound mass ratios in the DNAPL (from
soil samples), if the bypassing mechanism is operating.
31
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uo 5
0.02
6/15/91
«
j*
U
0.015
001
0.005
Id/xyl horn dan
toluene
-91% of toluene
v removed at 100 min.
\ \
o-xylene *
B B -91% of o-xylene
ffiE removed al 2(X) mm.
50
100 150
Time (min)
POO
Figure 2.6.4 Evolution of toluene and o-xylene soil gas concentrations in a homogeneous sand pack
[Ho and Udell, 1992].
Vapor
Cone. 60-
(mg/1)
Isopentane-Benzem
Benzene-Toluene
Toluene-Xylene
>Xylene
0
40 W) K(J
Time (d)
Figure 2.6.5 Soil gas composition as a function of time during soil venting at a gasoline contaminated
site [Johnson et al., 1990].
32
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Permeability
Zone
Low
Permeability
Zone
Airflow
bypasses
contaminant
Interface
V
.'•:.'• oj
V ' "O
('.'•'•'•'••.'•'•
Sc.
;N
ft
ffff^'- J. •' '. •' '. '
. •'. •' • •' • -' '• .'
mole fraction, x
.v.'/.^>-- binary
•' ''':•'•' '•' contaiiiinaiii
..•' . • . pool
Figure 2.6.6 Bypassing air flow mechanism and its effect on the composition profile of an evaporating
bicomponent NAPL pool trapped within low permeability zone [Ho and Udell, 1992].
2.6.3 Well Product Thickness
Evidence of floating LNAPL product in monitoring wells has been used to detect subsurface
contamination and to estimate the quantity of recoverable LNAPL [Zilliox and Muntzer, 1975; de Pastrovich
etal., 1979; Hall et al., 1984; Abdul etal., 1989; Farret al., 1990; Kemblowski and Chiang; 1990; Lenhard
and Parker, 1990; Mercer and Cohen, 1990]. While these efforts have been made for petroleum spills
(LNAPLs), they are of interest to us from the perspective that specific hydrocarbons (see Table 2.1) are
often components of multicomponent LNAPLs. Based on the arguments of soil heterogeneity, soil retention
capacity and saturation-capillary pressure relationships alone, it should be evident that the absence of
floating LNAPL in the monitoring wells does not preclude the existence of NAPLs in the subsurface. While
the presence of floating LNAPL in a monitoring well confirms its existence in the subsurface, it indicates
little else. For example, LNAPL can become trapped by capillary forces as a result of water table
fluctuations. Once immobilized, the LNAPL may no longer be in communication with a pre-existing well,
nor will it be easy to locate the trapped LNAPL during subsequent site characterization Therefore, the use
of monitoring wells for detection of NAPLs (consider immobile ganglia) and for NAPL recovery estimations
is questionable [Abdul et al., 1989; Lenhard and Parker, 1990].
Similar difficulties are encountered in trying to assess the presence of DNAPL. Most importantly,
however, boreholes and completed wells will act as downward conduits for preferential flow, therefore,
drilling into suspected DNAPL zones generally is not recommended. In cases when DNAPL is
encountered, the height of the DNAPL can be estimated only if the exact bottom of the DNAPL pool is
known and if the capillary rise at the DNAPL-water interface is included in the analysis. However, current
correlation models do not incorporate the Bond and Capillary number constraints which lead to
discontinuous NAPL emplacement under dynamic conditions. Thus, volume estimates of in-situ DNAPL
developed using well correlation methods may be highly misleading.
33
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2.6.4 Soil Samples
Because of the complexities introduced by multiphase-rnulticomponent equilibrium and mass
transfer limitations, accurate estimates of the extent of subsurface contamination are difficult to obtain using
both ground-water and soil gas samples, as already discussed. The remaining alternative is soil sampling
which has the advantage of giving very specific data at discrete locations. However, results of soil sample
analyses also require careful interpretation and an understanding of the principles already discussed.
The data of Dresen et al. (1986) illustrate that ground-water and soil samples taken from the same
borehole can lead to conflicting conclusions regarding the presence of NAPL [Sitar et al., 1992]. These
data are presented in Table 2.5. While the measured ground-water concentration for each BTX compound
is well below its individual solubility limit, the concentrations are consistent with saturation concentrations
of BTX in contact with separate phase gasoline, as indicated in Section 2.6.1. The reported ground-water
concentrations can-be used to estimate the amount of adsorbed hydrocarbon in equilibrium with the ground
water using Equations (12) and (19), and assuming a conservative value of foc=0.01 and Kow= 135, 490,
and 1300 for benzene, toluene, and xylene, respectively. The computed soil concentrations assuming
sorption alone are 22, 76, and 109 mg/kg for benzene, toluene, and xylene, respectively, which are at least
an order of magnitude below the measured soil concentrations. Thus, both the ground-water concentration
and soil concentration data point to the presence of free product gasoline.
TABLE 2.5 BTX CONCENTRATIONS IN WATER AND SOIL FROM SAME BOREHOLE AT A
GASOLINE CONTAMINATED SITE [Dresen et al., 1986]
NAPL
Benzene
Toluene
Xylene
Total Hydrocarbons
Measured Concentrations
in water (g/m3)
27
26
14
NA
in soil (mg/kg)
270
1100
1100
9400
The important point is that once the separate phase is present, its distribution is highly variable,
governed by the heterogeneity of the subsurface environment, and the soil concentration data will appear
highly variable and inconsistent. Moreover, current soil concentration reporting practices often provide no
information on the soil porosity, water content, density or other soil parameters [Sitar et al., 1992]. Hence,
volume or mass estimates of each phase (i.e., soil, water, air) are precluded, as are accurate estimates
of the hydrocarbon distribution and the volume of NAPL [Mackay, 1979, Mackay and Paterson, 1981;
Mackay and Shiu, 1992].
2.7
CHALLENGES FACING IN-SITU TECHNOLOGIES
A variety of the phenomena and technical issues associated with the fate and transport of DNAPLs
have been addressed in the previous sections. The migration of DNAPLs was described in terms of their
physical and chemical properties and in terms of the porous media characteristics under natural subsurface
conditions. In this respect, it is important to distinguish between three major zones of contamination: the
source zone, which contains very high concentrations of potentially mobile DNAPL; the residual zone,
through which the mobile DNAPL has already traversed, leaving behind considerable contamination; and
the dissolved zone (or plume), which emanates from the source and residual zones carrying dissolved
hydrocarbons usually at or below their respective solubility limits. In addition, in all zones, hydrocarbons
are likely to have partitioned into the solid phase (i.e., organic matter). The major concern with DNAPLs,
34
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in terms of remediation, revolves around the fact that DNAPL sources and residual zones are often quite
deep, making access and detection extremely problematic.
Soil heterogeneity is an important factor affecting DNAPL fate and transport. The site stratigraphy
affects the distribution of the DNAPL in the subsurface and the contaminant distribution then plays a critical
role in the selection of the overall approach for site remediation. Ultimately, the success of any passive
or active in-situ technology is largely associated with its susceptibility to soil heterogeneities and its ability
to favorably alter the DNAPL properties to facilitate recovery or remediation.
Soil heterogeneity and manipulation of subsurface conditions are not the only challenges facing
in-situ DNAPL cleanup technologies. Successful technologies also have to be able to adapt to other site
specific conditions such as depth to the water table, depth of the contaminated zone, volume of
contaminated soil, site access, and man-made structures and other obstructions.
Finally, current remedial goals often require that a baseline aqueous contaminant concentration
(maximum contaminant level, MCL) be attained, or that in excess of 99% of the DNAPL be treated or
recovered. This standard by itself poses a significant technical challenge to many technologies even under
the most favorable of conditions. Thus, all of these issues and challenges have to be kept in mind when
considering the potential viability of the remedial technologies which are discussed in the following sections
of this report.
-------
SECTION 3.0
TECHNOLOGY DESCRIPTIONS
3.1 TECHNOLOGY EVALUATION FORMAT
The technology descriptions contain a synthesis of the relevant information about each technology.
Focus is placed on attempting to identify in-situ technologies capable of addressing the remediation of
DNAPLs situated below the water table; secondary importance is placed on contaminants dissolved in the
aqueous phase. Several of the evaluated technologies were not originally developed for remediation of
contaminated sites, much less DNAPLs. As a result, some of the technologies have not yet been
demonstrated on DNAPLs, and owing to their developmental stage, have not been demonstrated in the
field and below the water table. Some in-situ technologies which have potential applicability to remediation
of DNAPLs occurring below the water table have been demonstrated in the vadose zone only; but beyond
this, the evaluation of technologies used to clean up contamination in the vadose zone has been omitted
from this report. Also, several in-situ technologies have been fully demonstrated only in non-environmental
applications, and are currently being adapted for environmental applications. In all cases, the applicability
to remediation of DNAPLs occurring below the water table is nonetheless considered in order to not rule
them out prematurely.
DNAPL treatability data were specifically sought for this report, but were often difficult to obtain for
the reasons indicated above. In cases where information on DNAPLs is sparse or not available,
performance data relating to treatment/recovery of LNAPLs and metals are provided for illustrative
purposes, when applicable. Technologies must therefore be evaluated within the context of the specific
application, even though an attempt is made to anticipate the theoretical and practical effectiveness of
these technologies to the DNAPL case. With certain technologies such as air sparging or in-situ
vitrification, it was very difficult to separate the theoretical background from the field implementation.
However, every effort was made to maintain the division.
Effort was made to evaluate the most current information and to select representative applications
illustrating the more interesting or impressive capabilities of each technology. However, an exhaustive
compilation of relevant case studies (as in the case of slurry wall construction) was beyond the scope of
this effort. Thus, the technology descriptions are intended to provide a basic technical assessment of the
technology and to identify its problem areas using basic principles.
Technologies are grouped by major process type (i.e., biological, soil washing, thermal) and are
arranged alphabetically. An attempt was also made to keep multifaceted technologies separate. For
instance, air sparging may enhance in-situ biological degradation of organic compounds. To avoid
repetition, this fact is only briefly stated in the air sparging section; and air sparging is reported as an
oxygen delivery method in the in-situ aerobic biodegradation section.
The Technology Descriptions, which appear later in this section, are organized into the following
subsections:
Theoretical Background-
The theory of each technology is evaluated in terms of the approach, reaction
types, dominant phenomena and important considerations or operating parameters.
Theory is distinguished from field implementation in order to assess the adequacy of the
theory separately from its application in the field. If the theory is poorly understood,
37
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empirical relationships are discussed.
Field Implementation--
A conceptual description or the layout of the technology is presented and
evaluated. Details regarding the development, construction, geometry, operating
parameters and process rates are provided under this heading, when available.
Level of Demonstration and Performance--
Information on lab-, pilot-, and field-scale performance is provided to demonstrate
capabilities of the technology. Information at the most advanced stage of development is
presented to the extent possible. Frequency of implementation is stated under this
heading and performance assessed versus predictions where possible.
Applicability/Limitations-
Information regarding targeted contaminants, soil matrix limitations, site
considerations, health and safety issues, and inefficiencies are addressed under this
heading.
Cost and Availability-
The promise and commercial availability of the technology is evaluated. Although
necessary technology hardware may be available, the requisite expertise may be lacking.
Patent and license information is provided where possible. Total, operating, maintenance,
partial, or relative costs are presented where available. Untreated residuals requiring
further treatment are usually identified.
References-
The most relevant citations selected and used in each technology description are
provided at the end of the report.
Although not formally addressed here, regulatory acceptance and/or approval will be required for
most of these technologies and, therefore, plays a major role in the viability of actual technology
implementation. Regulatory issues require serious consideration and the involvement of state and federal
entities should be sought at the earliest possible date to explore remedial alternatives.
Finally, the significance of site heterogeneity cannot be stressed enough. The effectiveness of all
of the reviewed technologies can be affected by the presence of subsurface heterogeneities. However,
some technologies are still more susceptible than others; therefore, this issue is an important criterion in
determining the potential effectiveness of a given technology.
38
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3.2 BIOLOGICAL PROCESSES
Introduction--
In-situ biodegradation is a process in which aqueous phase organic compounds are completely or
partially metabolized by microorganisms situated in the subsurface. Bacteria are largely responsible for
the biological transformations which occur in porous media and are generally considered as a stationary
phase, either through attachment to solid surfaces or via agglomeration [Criddle et al., 1991]. These
organisms convert natural and xenobiotic organic compounds into energy and end products, and utilize a
portion of the organic material for cell synthesis [Lee et al., 1988; McCarty, 1988, 1991; Sims et al., 1992].
In this section, some of the general features of biodegradation are presented and the in-situ aerobic and
anaerobic biodegradation processes are specifically evaluated in sections 3.2.1 and 3.2.2, respectively.
Metabolic processes of aerobic and anaerobic microbial consortia are distinguished by the nature
of carbon substrate utilization, and three metabolic processes are recognized: primary metabolism,
secondary metabolism, and cometabolism. The metabolic utilization of a compound depends on such
factors as its molecular structure, concentration, environmental conditions, bioavailability of nutrients,
presence of competing or inhibitory substrates, the nature of the microbial consortia and the enzymes and
cofactors they possess, and toxicity effects.
Primary metabolism of an organic compound occurs when it yields sufficient energy for both cell
maintenance and growth, and it is present at concentrations large enough to sustain the microbial
population [McCarty, 1988, 1991]. Petroleum hydrocarbons are generally good examples of primary
substrates, while compounds such as ammonia can serve as a primary energy source but not a carbon
source. Examples of halogenated primary substrates and the conditions under which they are utilized are
presented in Table 3.2.1. Many stoichiometric relationships describing the oxidation and reduction of
organic compounds by microbes have been enumerated [McCarty, 1975; Criddle et al., 1991]. From the
stoichiometric relations, nutrient (electron acceptor, primary substrate, nitrogen, phosphorus, etc.,) demands
can be estimated and Monod kinetics can be used to relate the growth and decay of the microbial consortia
to the degradation reactions [Monod 1942; McCarty, 1971]. Also, provisions can be made to incorporate
sorption and biofilm effects [Criddle et al., 1991; Semprini and McCarty, 1992].
Secondary metabolism describes the utilization of trace organic compounds which, by virtue of their
low concentrations, cannot sustain microbial growth [McCarty, 1988]. Cometabolism occurs when
nonspecific microbial enzymes or cofactors fortuitously biotransform organic compounds that provide
insignificant energy and organic carbon for growth [McCarty. 1988]. Cometabolism has been identified as
one of the major mechanisms in the transformation of chlorinated hydrocarbons and pesticides [Horvath,
1972].
TABLE 3.2.1 MICROBIAL UTILIZATION OF ORGANIC COMPOUNDS AS A FUNCTION OF
BIOLOGICAL PROCESS TYPE AND ENVIRONMENTAL CONDITIONS [McCarty, 1988]
Primary Subsl idles
Co-nietaholism
(secondary substiates)
Aerobic and Anaeiobic
Aerobic Primarily
Oxidations
Reductions
Glucose, acetone, isopropanol, acetate, benzoate, phenol
Alkanes, benzene, toluene, xylene, vinyl chloride.
1 .2 dichloroethane, chlorobenzene
Trichloiethylene, dichloroethylene. dichloioethane,
vinyl chloride, chloroform
1 .1 .1 -Tnchloioethane. inchloiethylene. letiachloioethylene,
dichloioethylene. dichloioethane,
caibon tetrachlonde. chloioform, DDT. hndane, PCBs
39
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Organic substrates may be oxidized under aerobic conditions and reduced under anaerobic
conditions. With increasing degree of halogenation, the carbon atoms within the organic compound
become more oxidized, making reductive metabolism more likely than oxidative [Vogel et al., 1987]. Hence,
for aliphatic compounds, rates of oxidation generally increase with decreasing degree of halogenation;
whereas rates of reduction generally increase with increasing degree of halogenation [Vogel et al., 1987].
The influence of the degree of halogenation on the transformation (reaction) rate is summarized in Figure
3.2.1. Table 3.2.2 presents several compound transformation rates by mixed cultures under a variety of
environmental conditions. Carbon dioxide or methane, water and inorganic salts (chlorine, bromine, etc.)
are produced by the complete mineralization of halogenated hydrocarbons and other compounds by
microbial consortia.
If naturally occurring microorganisms can be identified in the subsurface environment and are
capable of degrading the targeted compounds, they are usually enriched or biostimulated with amendments
which may include electron donors (primary substrates), electron acceptors (i.e., oxygen, nitrates, etc.), and
nutrients such as nitrogen, phosphorus and other trace metals. Many case studies exist where indigenous
microorganisms have been successfully biostimulated. Introduction of exogenous cultures into the
subsurface is often considered on occasions when indigenous consortia are incapable of degrading the
targeted compound, or are non-existent. However, to date, in-situ aerobic biodegradation resulting from
introduction of exogenous cultures into the subsurface has not been convincingly demonstrated [Lee et al.,
1988; Alvarez-Cohen, 1993a]. Introduction of genetically engineered microorganisms into the subsurface
is presently prohibited by law in the U.S. without prior approval [Thomas and Ward, 1989].
Elements common to successful applications of in-situ aerobic biodegradation include: adequate
aquifer permeability, usually K>10~4cm/s [Thomas and Ward, 1989]; prior removal of free product [Thomas
and Ward, 1989; Alvarez-Cohen, 1993a]; a suitable microbial population [McCarty 1991; Alvarez-Cohen,
1993a]; sufficient hydrodynamic control for plume containment and delivery of required amendments [Sims
I
1
g
o
-------
TABLE 3.2.2 COMPARISON OF SUBSTRATE UTILIZATION RATES BY MIXED CULTURES USING
DIFFERENT ELECTRON ACCEPTORS [Griddle et al., 1991]
Substrates
Primary
acetate
Secondary
chlorobenzene
o-dichloro benzene
p-dichlorobenzene
1,2,4- trichlorobenzene
ethylbenzene
styrene
naphthalene
bromoform
chloroform
carbon tetrachloride
1,1,1,- tnchloroethane
tetrachloroethene
1,2-dibromomethane
dibromochloropropane
hexachloroe thane
k' (L / mg day)
Aerobic
(02)
38
2.5
100
11.0
50
350
500
40.0
Anaerobicb
Damnification
(NO.,')
1.4
0.23
036
0002
00019
0056
Sulfate Reduction
(SO/')
1 0
071
0.2
0005
00076
023
038
Methanogenesis
(CO,,)
0.63
2.0
0 21
063
096
0094
2 1
24
061
References. a Bouwer and McCarty (1985), b Bouwer and Wright (1988)
et al., 1992; Alvarez-Cohen, 1993a]; and a complete monitoring system [Sims et al., 1992; Alvarez-Cohen,
1993a]. In common engineering practice, successful in-situ biodegradation is often explained in terms of:
changes in dissolved electron acceptor concentrations (oxygen, nitrate, etc.), reduction in dissolved
compound concentrations, increased carbon dioxide (or methane) concentrations, increased biomass in-
situ, and the ability of indigenous microorganisms to biologically transform the targeted compounds in
laboratory microcosms. Evidence of this kind is putative because evaluation of successful in-situ
biodegradation is complicated by: uncontrollability of the field sites [Madsen, 1991]; aquifer heterogeneities
[Madsen, 1991; McCarty, 1991; Alvarez-Cohen, 1993a]; and, a wide range of competing contaminant fates
[Madsen, 1991, Alvarez-Cohen, 1993a]. For example, most abiotic transformations are slow, but are still
significant within the time scales commonly associated with groundwater movement [Vogel et al., 1987].
Actual proof and sufficient confirmation of halogenated hydrocarbon destruction by microbial
degradation has been provided in only a few studies which rely on several convergent lines of evidence
[Madsen 1991; Alvarez-Cohen, 1993a]. Madsen [1991] summarizes sources of evidence which indicate
successful in-situ biodegradation: production of specific intermediate metabolic compounds; changes in
organic compound, stereoisomer, isotope or electron acceptor/tracer ratios after onset of in-situ
biodegradation; amendment utilization and aqueous phase concentration responses coincident with pulsed
41
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injection; and increased presence of microbial predators. Other factors may exist which are compound and
site specific.
Because of compound toxicity, in-situ microorganisms cannot degrade the pure phase organic
liquids. However, in-situ biological degradation should be considered when these compounds exist in
residual or trace amounts in the saturated zone, and for any dissolved plumes emanating from DNAPL
source areas. Considerable benefit can be derived from "ground-water polishing" when in-situ
biodegradation is used in conjunction with other technologies capable of addressing the removal of the
separate phase.
3.2.1 Aerobic Biodegradation
Theoretical Background-
Aerobic biodegradation is a process by which aqueous phase organic compounds are completely
or partially metabolized by oxygen utilizing microorganisms. Using oxygen as the terminal electron
acceptor, microorganisms convert natural and xenobiotic organic compounds into end products, energy,
and utilize a portion of the organic material for cell synthesis [Lee et al., 1988; McCarty, 1988, 1991; Sims
et al., 1992]. For low molecular weight chlorinated aliphatics, the metabolic intermediates and end products
exuded by aerobes are not usually recalcitrant and/or toxic. However, degradation pathways for more
complex aliphatics and aromatics may involve recalcitrant and/or toxic intermediates and end products.
Monooxygenases mediate halogenated hydrocarbon oxidation in three generalized ways in aerobic
systems: oc-hydroxylation and halosyl oxidation of halide substituted alkanes, which results in the formation
of easily biodegradable alcohols and organic acids; and epoxidation of ethene bonds which produces
unstable epoxides. Epoxidation is recognized as being the first step in the overall mineralization of several
halogenated hydrocarbons in microbial systems [Stirling and Dalton, 1979; Patel et al , 1982; Janssen et
al., 1987]. Methanotrophs were among the first bacteria recognized to utilize the epoxidation mechanism
during the cometabolism of NAPLs [Leadbetter and Foster, 1959]. The epoxidation mechanism is shown
in Figure 3.2.1.1 for the degradation of TCE by methane monooxygenase (MMO) as originally proposed
by Hou (1984) and as modified by Henry and Grbic-Galic (1986). Epoxidation is afforded by the unusually
broad substrate specificity of MMO [Stirling and Dalton, 1979]. Epoxidation pathways are recognized for
several nonspecific oxygenases such as methane, propane, toluene, and ammonia monooxygenases and
toluene dioxygenase [McCarty, 1992].
CELL CONSTITUENTS
MMO
1
HCHO— ^— -HCOOH-™^-^ CO2 + H2O
4 _ - ,__3-— —
NADH NAD * XH NAD NADH NAD NADH
TCE —-? "x^ TCE EP°xide
NADH NAD
Figure 3.2.1.1 Methanotrophic utilization of methane by Methane Monooxygenase (MMO) and the
formation of TCE-epoxide as the initial step of TCE oxidation [Hou, 1984; Henry and Grbic-
Galic, 1986].
42
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MMO catalyzes the oxidation of methane (CH4) to methanol (CH3OH); and energy is liberated, as
indicated in Figure 3.2.1.1. During methane oxidation, energy is expended as TCE is transformed to TCE-
epoxide. TCE-epoxide can then undergo hydrolysis (abiotic oxidation) to intermediates which can be easily
degraded to carbon dioxide, chloride, and water [Little et al., 1988]. Subsequent oxidation of methanol,
formaldehyde (HCHO), and formate (HCOOH) catalyzed by other oxygenases yields additional energy.
While the use of methanol, formaldehyde, and formate as alternate substrates appears attractive from the
perspective that reduced competition between methane and the targeted compounds for the MMO will
result in higher compound removal [Semprini et al., 1991], field [Semprini et al., 1990,1991] and laboratory
[Alvarez-Cohen and McCarty, 1991; Henry and Grbic-Galic, 1991] studies have indicated that MMO enzyme
production, and subsequently halogenated hydrocarbon epoxidation, is curtailed when methane is absent.
However, MMO can remain active in the cell for extended periods of time.
Finally, some bacteria can oxidize simple chlorinated aliphatics such as vinyl chloride [Hartmans
et al., 1985], 1,2 dichloroethane [Stucki et al., 1983], and methylene chloride [LaPat-Polasko et al., 1984]
as sole carbon sources for energy. Aromatic degrading bacteria using phenol and toluene as the primary
substrates have been shown to degrade di- and tri-halogenated ethenes [Nelson et al., 1987].
Field Implementation--
Most in-situ aerobic biodegradation applications are variations on the approach patented by
Raymond (1974). Wellhead injection and infiltration galleries are two common configurations used for in-
situ aerobic biodegradation, as shown in Figure 3.2.1.2. In-situ biostimulation is usually achieved using
strategies analogous to hydraulic gradient control and pump-and-treat methods except that the injected
fluids are amended. Pulsed (cyclic) injection of amendments has been employed to prevent biofouling in
the vicinity of the injection well [Semprini et a!., 1990]. Conventional injection and extraction well
construction equipment can be used. Well placement strategy will depend on the nature and extent of
contamination, soil heterogeneities, and anticipated subsurface flow behavior. In-situ biostimulation may
initiate changes in fluid flow as a result of: changes in aqueous chemistry, pH, porosity, and fluid viscosity;
corrosion of the support media, and surface property alterations [Griddle et al., 1991].
Because the subsurface is usually anaerobic and the oxygen demand for sustained biological
degradation can be appreciable, oxygen, the primary electron acceptor, must be supplied to accelerate
aerobic processes. Primary substrate, oxygen, and nutrient demands can be ascertained from treatability
studies using aseptically obtained aquifer samples. Delivery of primary substrates, nitrogen, and
phosphorus amendments is facilitated by aqueous phase injection due to their high water solubilities.
However, actual delivery of nutrients to the contaminated soil zone depends on such factors as
heterogeneity, hydraulic conductivity, etc.
Oxygen can be supplied in several ways which may require judicious selection given site
conditions. The dissolved oxygen (DO) content of injected or recycled water can be increased to saturation
(approximately 8 to 12 mg/l) prior to injection [Lee et al., 1988; Semprini et al., 1990; Semprini and McCarty
1992]. Air spargers can locally increase the DO of the saturated zone to approximately 8 to 12 mg/l if air
is injected [Lee et al., 1988], or up to 40 mg/l if pure oxygen is used [USEPA, 1988; Thomas and Ward,
1989]. Spargers, as described in Section 3.6.2, may introduce such problems as precipitation of hardness
ions and pore clogging under initially reduced subsurface conditions, changes in the local hydraulic
gradients due to mounding, and uncontrolled migration of dissolved contaminants and DNAPL away from
the treatment zone.
Two patents have been issued on hydrogen peroxide assisted in-situ aerobic biodegradation
[Raymond et al., 1986; Lawes and Litchfield, 1988]. Hydrogen peroxide (H2O2) solutions can be directly
injected into the saturated zone to provide a source of oxygen. H2O2 solutions can be stabilized by use
of peroxidases, oxidases, or phosphates so that excessive decomposition by iron or soil catalyzed reactions
43
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(a)
To Sewer or
Recirculate
Water Supply
Injection Well
Clay
(b)
Air Compressor or
Hydrogen Peroxide
Tank
Nutrient Addition
Infiltration Gallery
,Trapped Hydrocarbons
V
Monitoring Well
Water Table
Recovery Well
Figure 3.2.1.2 Schematics illustrating oxygen and nutrient delivery using spargers (a) and an infiltration
gallery (b) [Thomas and Ward, 1989].
44
-------
does not occur [Raymond et al., 1986; Lee et al., 1988; USEPA, 1988]. Also, some organic compounds
may be oxidized by H2O2 and the mobility of inorganic metals such as lead and antimony may be increased
via reactions with H2O2 [Alvarez-Cohen, 1993a]. Gradually increasing H2O2 concentrations from 50 mg/l
to as high as 1000 mg/l fosters microbial consortia acclimation to H2O2 and mitigates microbial toxicity
[Thomas and Ward, 1989]. However, Lee et al. (1988) report that degassing of molecular oxygen may
occur at H202 concentrations > 100 mg/l.
Laboratory studies have indicated that colloidal gas aphrons (CGAs) may be an alternate source
of oxygen delivery into the subsurface [Lofti and Michelsen, 1991; Michelsen et al., 1984, 1991]. CGAs,
are a colloidal microdispersion of oxygen (65 vol%) in a thin soapy film (surfactant matrix) that have bubble
dimensions of 25 to 50 jim in diameter. When injected into the saturated zone, the microbubbles eventually
become immobilized and adhere to the solid phase. Oxygen transfer efficiencies of 5.4 to 59% have been
achieved in two-dimensional tank studies using sandy soils. However, the viscous fluid (16 cp) may
channel when injected or cause ground-water diversion, thus reducing contact efficiency. About 12% of
the oxygen transferred is required to biodegrade the surfactant.
Level of Demonstration and Performance-
Numerous sites containing DNAPLs are reported to have been remediated by in-situ aerobic
biodegradation, and many are underway [USEPA, 1992b]. An extensive listing of ongoing bioremediation
activities is available [USEPA, 1992c]. Approximately 20% of the entries contained within the SITE
Demonstration Program Technologies in the USEPA's ATTIC database include bioremediation applications
[USEPA, 1992b]. As indicated earlier, actual proof and sufficient confirmation of targeted compound
destruction by microbial degradation has only been provided in a few studies [Madsen 1991; Alvarez-
Cohen, 1993a].
In 1986-1988, Moffet Field Naval Air Station in Mountain View, California, was the site of several
detailed in-situ aerobic biodegradation studies which illustrated the successful biotransformation of
chlorinated aliphatics [Roberts et al., 1990; Semprini et al., 1990; 1991; Semprini and McCarty, 1991,1992].
The studies were conducted in a shallow confined sandy and gravelly aquifer which was not strongly
anaerobic. Bromide tracer studies were conducted to define the flow characteristics and capture efficiency,
and the sorption and retardation of each compound was evaluated prior to biostimulation. Mass balances
performed on trichloroethene (TCE), cis-dichloroethene (cis-DCE), trans-dichloroethene (trans-DCE) and
vinyl chloride (VC) revealed that abiotic transformations were negligible. Laboratory [Henry and Grbic-Galic,
1986] and field experiments [Semprini et al., 1990] demonstrated that the indigenous methanotrophic
consortia could be biostimulated. Methane and oxygen were pulsed at various intervals, whereas small
concentrations of the dissolved hydrocarbons were continuously injected.
In-situ aerobic biodegradation was demonstrated by several corroborating facts. Decreased and
increased concentrations of the different compounds were observed to be coincident with the onset and
cessation of methane utilization, respectively. The elimination of electron donors/acceptors, and
appearance of a specific biotransformation product {trans-dichloroethene oxide (epoxide)} were a further
confirmation of actual in-situ biodegradation [Semprini et al., 1990]. More importantly, quantitative
knowledge of the contaminant releases and extensive instrumentation and monitoring afforded accurate
mass balances that are atypical of most in-situ biodegradation applications.
Madsen et al., (1991) successfully demonstrated in-situ aerobic biodegradation of coal tar
constituents (naphthalene and phenanthrene) in a shallow confined aquifer setting. Increased populations
of microbes, particularly protozoans (predators), were detected within the plume area. The degradation
activities and microbial population comparisons between contaminated and pristine soil samples taken from
the site served as indirect and qualitative evidence of in-situ biodegradation.
45
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Applicability/Limitations-
In-situ aerobic biodegradation applies only to the remediation of the aqueous phase. Regions
containing the separate phase cannot be treated because the large compound concentrations result in
microbial toxicity; therefore, major accumulations of free product must be removed by other means [Thomas
and Ward, 1989; Alvarez-Cohen, 1993a]. Many DNAPLs were previously thought to be non-biodegradable,
but new laboratory studies continue to demonstrate the biodegradability of DNAPLs. While site data may
suggest in-situ biodegradation, it may be difficult to prove because of site conditions and competing
mechanisms [Madsen, 1991; Alvarez-Cohen, 1993a].
Site characterization is required and hydraulic gradient control is necessary to effectively deliver
nutrients. Hydraulic conductivities should be above 10"4 cm/s [Thomas and Ward, 1989]. Soil
heterogeneities will greatly affect the ability to implement in-situ biodegradation. At very heterogeneous
sites, in-situ biodegradation may be completely ruled out because of the inability to effectively delivery the
nutrients to the contaminated areas [Alvarez-Cohen, 1993b].
Treatability studies are required to assess the viability of biostimulation, nutrient demands, ability
of culture to degrade DNAPL, and other factors such as pH, redox potential, moisture conditions, DNAPL
toxicity, and temperature effects. Microorganisms may convert DNAPLs into more recalcitrant, toxic, or
inhibitory intermediate products. In the particular case of petroleum hydrocarbons, in-situ bioremediation
can accelerate the time scale of pump-and-treat and natural attenuation of NAPLs from decades and
centuries to months and years. However, degradation rates generally decrease as concentrations decrease
and total cleanup may not be attainable [Ellison, 1992].
Cost and Availability-
In-situ aerobic biodegradation has been implemented on the full-scale numerous times, but mostly
for petroleum hydrocarbons (usually non-halogenated). The technology hardware is generally available
for full-scale applications. Although pulsed injection can be accommodated by two wells, more
sophisticated (automated) injection systems may require site specific design, fabrication, or assembly.
Aerobic biodegradation is a good candidate for dissolved plume management in the saturated zone in
granular (aquifer) media It is not well suited for low permeability and fractured media, or in areas where
DNAPL is present,
In-situ aerobic bioremediation often costs less than other remedial technologies [Sims et al., 1992].
The usual cost range is $15-60/yd3 [Ellison, 1992], Ex-situ hardware will include such items as nutrient
feedstocks, air strippers, and granular activated carbon.
3.2.2 Anaerobic Biodegradation
Theoretical Background-
Anaerobic biodegradation involves complete or partial metabolization of aqueous phase organic
compounds by non-oxygen utilizing microorganisms. Using such compounds as nitrates, sulfates, carbon
dioxide, or possibly ferric iron and other metal oxides as terminal electron acceptors [Ball et al., 1992],
microorganisms convert natural and xenobiotic organic compounds into end products, energy, and utilize
a portion of the organic material for cell synthesis [Lee et al., 1988; McCarty, 1988, 1991; Sims et al.,
1992]. In certain cases, polyhalogenated aliphatic organic compounds can also serve as electron acceptors
[Vogeletal., 1987],
Reduction of halogenated hydrocarbons occurs through two generalized dehalogenation
mechanisms under anaerobic conditions: hydrogenolysis and dihalo-elimination. Hydrogenolysis entails
the substitution of halogen atoms by hydrogen atoms. Dihalo-elimination results in the replacement of two
adjacent halogen atoms by an ethene bond. Figure 3.2.2.1 illustrates that the hydrogenolysis mechanism
46
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Ci
CI-C
I
Ci
CT
CI
I
-CI
carbon dioxide
a Cl cDCEa C( (DCE H
Figure 3.2.2.1 Pathways for anaerobic biotransformation of chlorinated aliphatics including abiotic (a)
transformations [Vogel et al., 1987].
is the predominant reductive pathway for aliphatic hydrocarbons. These reductions may be mediated by
a variety of enzymes and cofactors. Under severely reducing conditions, such as those typified by
methanogenesis, the organic compound can actually serve as the electron acceptor [Alvarez-Cohen,
1993bj.
Reductive dehalogenation also occurs in aromatic hydrocarbons. Descriptions of sequential
dehalogenation of aromatic compounds and aromatic ring cleavage mechanisms are beyond the scope of
this review, but are available [USEPA, 1986a].
Field Implementation--
Most in-situ biodegradation applications are variations on the approach patented by Raymond
(1974). In anaerobic systems, wellhead injection is the usual configuration, and the strategy employed is
analogous to aerobic processes (section 3.2.1) with the exception of electron acceptor delivery Unlike
oxygen, anaerobic electron acceptors (nitrate, sulfate, etc.) are extremely water soluble. The subsurface
is usually anaerobic which makes this process advantageous. Delivery of primary substrates electron
acceptors, nitrogen and phosphorus amendments is easily facilitated by aqueous phase injection owing to
47
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their high water solubilities.
Level of Demonstration and Performance-
Anaerobic processes are naturally occurring and have been observed to occur in-situ. Field studies
have been conducted, but no full-scale engineered applications are known to exist. In 1980, St. Louis Park
(Minnesota) was the site of a study which evaluated the in-situ anaerobic degradation of creosote
constituents [Ehrlich et al., 1982]. The constituents of interest were the phenolics (2-17%). The 20 m thick
aquifer consisted of three units: an upper drift of lake deposits and till; a middle drift consisting of glacial
sands; and a lower unit consisting of till and deeply weathered bedrock. The presence of methane and
methanogenic consortia only within the plume area and their absence elsewhere was evidence of in-situ
biodegradation. Disappearance of phenolics relative to other less biodegradable creosote constituents,
such as naphthalene, suggested that dilution was not a source of attenuation. Supplemental evidence was
provided by laboratory sorption studies using field samples which indicated low phenolic sorption. Methane
production in laboratory microcosms, inoculated with indigenous bacteria from the contaminated soil zone,
lent further credibility to the qualitative conclusion that phenolics were being biodegraded in-situ under
anaerobic conditions.
In 1988-1989, Moffet Field Naval Air Station in Mountain View, California, was the site of a detailed
in-situ anaerobic biodegradation study which illustrated the successful biotransformation of chlorinated
aliphatics [Semprini et al., 1992]. Experiments were conducted in a shallow confined sandy and gravelly
aquifer which was not strongly anaerobic. Bromide tracer studies were conducted to define the flow
characteristics and capture efficiency, and the sorption and retardation of each organic compound was
evaluated prior to biostimulation. Mass balances performed on carbon tetrachloride (CT) revealed that
abiotic transformations were negligible. Two potential electron acceptors were naturally occurring: nitrate
(25 mg/l; as nitrate) and sulfate (700 mg/l; as sulfate). Other contaminants present in the groundwater
included 50 u,g/l trichloroethane (TCA), 6 jag/l Freon-113 and 3 |ag/l Freon 11. Field experiments [Semprini
et al., 1992] demonstrated that the indigenous consortia could be biostimulated using acetate as a primary
substrate without any other amendments. To avoid biofouling near the injector, acetate (320 mg/l) and
nitrate (25 mg/l) were pulsed at various intervals, whereas CT (40 u.g/1) was continuously injected.
The field response indicated that the main denitrifying population was not responsible for the
transformation of CT. While most of the acetate (80-90%) and nitrate were consumed within the first meter
of transport, the most rapid rates of CT transformation occurred further downstream. Potential inhibition
of CT transformation by high nitrate concentrations may have caused this trend. It was therefore
hypothesized that a secondary microbial consortia that utilized the remaining acetate and decay products
of the denitrifiers was responsible for the CT transformation [Semprini et al., 1992]. This hypothesis
appears to be corroborated by the results of the transient experiments in which no direct evidence was
found for the stimulation of sulfate-reducing or methanogenic bacteria when nitrate was completely
removed.
In-situ anaerobic biodegradation was demonstrated by several corroborating facts. Decreased
organic compound concentrations were observed to be coincident with the onset of acetate utilization,
whereas increased organic compound concentrations coincided with the cessation of acetate utilization.
Appearance of chloroform (CF), an intermediate biotransformation product which is not produced abiotically,
was further confirmation of actual in-situ biodegradation [Semprini et al., 1992], Transformation of CT was
on the order of 70 to 97% within the test zone, and CF accounted for 30 to 40 % of the CT transformed.
The other compounds, TCA, Freon 113, and Freon 11 were also transformed, but to a lesser extent.
Quantitative knowledge of the compound releases, extensive instrumentation and monitoring afforded
accurate mass balances that are atypical of most in-situ biodegradation applications.
48
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Applicability/Limitations--
In-situ anaerobic biodegradation applies only to the remediation organic compounds in the aqueous
phase. Via the aqueous phase, sorbed and residual organics can be rapidly biotransformed. The separate
phase cannot be directly treated because of compound toxicity, and major accumulations of free product
must be removed [Thomas and Ward, 1989; Alvarez-Cohen, 1993a]. Many halogenated organic
compounds were previously thought to be non-biodegradable, but new laboratory studies continue to
demonstrate the biodegradability of these compounds by mixed or pure cultures and to elucidate the
metabolic pathways. While site data may suggest in-situ biodegradation, it may be difficult to prove
because of site conditions and competing mechanisms [Madsen, 1991; Alvarez-Cohen, 1993a].
Treatability studies are required to assess the viability of biostimulation, nutrient demands, ability
of culture to degrade a targeted compound, and other factors such as pH, redox potential, moisture
conditions, compound toxicity, and temperature effects. Site characterization is required and hydraulic
gradient control is necessary to effectively deliver nutrients. Hydraulic conductivities should be above 10"4
cm/s [Thomas and Ward, 1989]. Soil heterogeneities will greatly affect the ability to implement in-situ
biodegradation, and even under the most favorable conditions, total cleanup may not be attainable [Ellison,
1992].
Depending on the environmental conditions and the exact compound and its metabolic pathway,
the intermediate or end products exuded by the anaerobic consortia may be recalcitrant, toxic, or
undesirable. For example, vinyl chloride is a byproduct which poses a greater human health hazard than
the parent compound [Vogel et al., 1987], while formation of chloroform is undesirable from a water quality
standpoint [Semprini et al., 1992]. This issue may preclude anaerobic processes from being implemented
at certain sites, and it is one of the reasons why much attention has been devoted to aerobic processes.
Another aspect to consider is the condition of the aquifer after remediation: it will be anaerobic and possibly
very reduced and characterized by relatively high concentrations of Fe, Mn, H2S, and CH4.
Cost and Availability-
In-situ anaerobic biodegradation is naturally occurring. Field studies have been conducted, but no
full-scale applications exist. Much remains unknown about the anaerobic mineralization of halogenated
hydrocarbons in-situ. At this time, full-scale application of anaerobic biodegradation is generally
discouraged because of the formation of potentially toxic end products. However, the hardware is available
for full-scale applications.
The cost of anaerobic biodegradation is comparable to that of aerobic processes ($15-60/yd3
[Ellison, 1992]). Ex-situ hardware may include such items as nutrient feedstocks, air strippers, and granular
activated carbon.
49
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3.3
ELECTROLYTIC PROCESSES
Introduction-
In-situ electrolytic processes use applied electric fields to enhance organic contaminant removal.
The effectiveness of these processes in soils is controlled by coupled flow phenomena [Mitchell, 1976,
1991; van Olphen, 1977; Mitchell and Yeung, 1990; Yeung 1990]. In most cases, the flow results from the
presence of fluid, heat, electrical, and chemical flow potentials; or any of these potentials may be created
even though only one driving force is applied [Mitchell, 1991]. These relationships are shown in Table
3.3.1. The electrolytic processes reviewed in this chapter include electro-osmosis (section 3.3.1), and
electroacoustic soil decontamination (section 3.3.2) which employs both electrical and acoustical fields to
enhance contaminant treatment.
TABLE 3.3.1 DIRECT AND COUPLED FLOW PHENOMENA OCCURRING IN THE SUBSURFACE
[Mitchell, 1991]
Flow J
Fluid
Heat
Current
Ion
Gradient X
Hydraulic Head
Hydraulic
conduction:
Darcy's law
Isothermal heat
transfer
Streaming current
Streaming current
Temperature
Thermo-osmosis
Thermal
conduction.
Fourier's law
Thermo-electricity.
Seebeck effect
Thermal diffusion
of electrolyte
Soret effect
Electrical
Electro-osmosis
Peltier effect
Electrical conduction-
Ohm's law
Electrophoresis
Chemical
Chemical-osmosis
Dufour effect
Diffusion and
membrane potentials
Diffusion
Pick's law
Electro-osmotically and chemico-osmotically driven fluid flows dominate transport in saturated, fine
grained soils having hydraulic conductivities less than approximately 10"9 m/s [Mitchell, 1991] because the
electrical conductivity of a soil is independent of soil particle size and pore size, whereas the hydraulic
conductivity is related to particle size [Casagrande, 1952; Mitchell, 1976, 1991; Shapiro and Probstein,
1993]. Also, in clays a portion of the relatively large surplus of cations, which are required to balance the
net negative charge of the clay particle surfaces, can be mobilized under applied electrical gradients to
produce hydraulic flow that would not otherwise be possible by hydraulic means alone [Mitchell, 1976].
In general, electrolytic methods attempt to mobilize ionic species, namely dissolved heavy metals,
radionuclides, and charged organic compounds towards their respective electrodes by applying potential
(electrical and acoustical) fields. Recovery of contaminants including neutral compounds transported by
advection occurs at the electrodes. Dissolved contaminants are pumped to the surface for above-ground
treatment. However, DNAPLs will not be appreciably mobilized by electrolytic methods because of their
uncharged, non-polar attributes.
Subsurface structures and objects [Lageman et al., 1989] and electrode corrosion [Segall and
Bruell, 1992; Shapiro and Probstein, 1993] are among the factors that will interfere with the electrolytic
process. If not properly applied, electrolytic processes may also cause cracking and fabric changes in the
porous media [Mitchell, 1991].
50
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3.3.1 Electro-Osmosis (EO)
Theoretical Background-
Electro-osmosis uses current flow and electric potential gradients to enhance organic contaminant
removal. The primary mechanisms are: ionic migration of charged species (cations, a'nions) resulting from
the applied electrical potential [Casagrande, 1952]; and advection of neutrally charged species in the
direction of the bulk diffusive flow of the major ions, usually cations [Mitchell, 1991]. Other phenomena that
may contribute to overall contaminant removal when an electric field is applied to a wet soil mass include:
development of osmotic and pH gradients, desiccation of soils due to heat generation at the electrodes,
precipitation, electrolysis, hydrolysis, oxidation, reduction, adsorption, and soil fabric changes [Mitchell and
Yeung, 1990].
The electrostatic distribution of ions in the vicinity of a negatively charged clay particle surface is
schematically shown in Figure 3.3.1.1. When an external electric field is applied, both cations and anions
"drag" water with them toward the cathode and anode, respectively. However, the net bulk fluid flow is in
the direction of the cathode owing to the abundance of cations, as shown in Figure 3.3.1.2.
Several factors will affect contaminant removal. While the electrical and hydraulic gradients may
be held constant during electro-osmosis, chemical gradients evolve from migration of cations and anions
[Mitchell and Yeung, 1990]. Chemical gradients acting counter-current to the bulk flow direction will retard
contaminant removal. Depending on the applied electrical potential and current flow, joule resistance
heating of the porous media may also result [Shapiro and Probstein, 1993; Smith and Hinchee, 1993].
Additionally, hydrolysis of water at the electrodes causes the pH to rise at the cathode and decrease at the
anode [Mitchell and Yeung, 1990]. In the initial stages of the process, the pH at the anode and cathode
can approach 2 and 12, respectively. This may result in the propagation of an acid front towards the
cathode [Shapiro et al. 1989a,b; Named et al., 1991; USEPA, 1992d]. The rate of advance of this front
will be affected by the buffer capacity of the soil [Acar, 1992] and the pH variations may initiate ionization
or valence changes of organic compounds which can affect their removal [Shapiro et al., 1989b] Alkaline
conditions occurring near the cathode can desorb organics, pesticides, and heavy metals from the solid
surfaces, thereby enhancing their removal [Segall et al., 1980].
Field Implementation-
A schematic of an in-situ environmental application of electroosmosis is shown in Figure 3.3.1.3.
Ions migrate to the respective electrodes in response to the applied electrical field. Recovery of
contaminants occurs at the electrodes, and the dissolved contaminants are pumped to the surface for
above-ground treatment. Electrode placement strategy depends on the nature and extent of contamination,
soil heterogeneities, flow resistance, and response of the electrical potential field to subsurface features.
The site must be thoroughly characterized and bench studies are required. The strategy of electroosmosis
is in many ways analogous to pump-and-treat, but using current, voltage, and electrical gradients as major
design variables.
Electrodes can be placed in conventional injection and extraction wells [Banerjee, 1993].
Electrodes are usually fabricated of a metal such as iron [Mitchell, 1991; Segall and Bruell, 1992; Banerjee,
1993] or graphite [Segall and Bruell, 1992]. Graphite electrodes have non-wetting surfaces which offer
greater corrosion resistance, reduce hydrogen gas formation, and can be constructed to facilitate fluid
injection, extraction, or recycling [Segall and Bruell, 1992]. To keep the clay saturated in order to mitigate
consolidation effects, water or purge solutions containing salts, surfactants, or chelating agents are usually
injected at the anode [Acar, 1992].
Dewatering applications can have electrode spacings on the order of 30 ft. In contaminant recovery
applications, electrode spacings are usually on the order of 3 to 15 feet depending on the soil type and
51
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applied current. Applied DC potentials on the order of 25-500 volts are typical.
Level of Demonstration and Performance--
Geotechnical applications of electro-osmosis for purposes of clay dewatering and enhancing rates
of clay consolidation, and for general site and soil improvement are well known and date back to the 1930's
[Casagrande, 1952; Mitchell, 1991; Civil Engineering, 1992]. Environmental applications of electro-osmosis
are relatively new and have mostly addressed recovery and treatment of heavy metals and radionuclides
[Lageman, 1989]. Muralidhara et al. (1990) provide a brief summary of several electro-osmotic
applications. Only laboratory studies have addressed the removal of dissolved organics, including TCE.
Q)
O
CO
t
O
©
0
- ©
© -
©
© ^
© ©0
©
©"
Distance
Figure 3.3.1.1 Ion distribution adjacent to clay particle surface [Mitchell, 1991]
ELECTRIC FIELD;
ANODE
CATHODE
WATER Q
VELOCITY
PROFILE
Figure 3.3.1.2 Schematic of electro-osmotic flow resulting from an applied electric field in a charged
porous medium [Shapiro et al., 1989b].
52
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In 1986, electro-osmosis was implemented at a former chrome-plating facility in Corvallis, Oregon
[Civil Engineering, 1992; Banerjee, 1993]. The subsurface soils consisted mostly of clays and clayey silts,
and the water table was located at a depth of 10 ft. Seven electrodes (iron reinforcing bars) were installed
in monitoring wells to a depth of 20-22 ft. The seven spot well configuration was used with the anode
located in the center of the test cell. The spacing between the central well and perimeter wells was 5 ft.
The applied current was on the order of 5-10 amps. Hexavalent chrome and other metals were recovered
in the central well. In this test, ground-water concentrations of chrome were reduced from 1,000 mg/l to
35 mg/l [Civil Engineering, 1992].
Most environmental applications of electro-osmosis have been conducted abroad, principally by
Geokinetics, Inc., in Europe. Two field tests and one commercial application have been completed
[Lageman et al., 1989; Civil Engineering, 1992]. One field experiment [Lageman et al., 1989] was
conducted in sandy, clayey soils located near the ground surface (1 m depth) having soil concentrations
of zinc of 7,101 ppm (max.) and 2,410 ppm (ave.). After approximately 8 weeks (including some downtime)
using an energy supply of approximately 160 kW/ton, the zinc concentrations were reduced to 5,300 ppm
(max.) and 1,620 ppm (ave.). Another field experiment was conducted in sediments (peat and fines)
located in a drainage ditch which had soil concentrations of copper and lead as high as 5,000 and 10,000
ppm, respectively [Lageman et al., 1989]. In the commercial application, approximately 340 tons of arsenic
contaminated clayey soils were treated [Lageman et al., 1989]. Arsenic concentrations were reduced from
110 ppm to 30 ppm in approximately 10 weeks [Lageman et al., 1989; Civil Engineering, 1992].
Electro-osmosis is currently being evaluated as part of the USEPA SITE Program for the removal
of tetraethyl lead from clayey soils at a former refinery [Acar, 1992; USEPA, 1992d]. Soil concentrations
of lead as high as 100,000 ppm have been measured. For perspective, the drinking water standard is 5
ppm, and the SITE cleanup goal is approximately 500 ppm [Acar, 1992]. The 10 x 25 ft test cell is shown
in Figure 3.3.1.4. At this time, soil is being treated to a depth of 3 ft. A central anode array consisting of
7 anodes at 2.5 ft spacings is surrounded by 24 cathodes. Two rows of 11 cathodes are aligned parallel
to the anode array. As of fall 1992, initial results indicated that electro-osmosis was unsuccessful and the
test program was temporarily suspended [Civil Engineering, 1992]. Enhancement schemes are currently
being investigated.
Soluble (and polar) organic compounds such as acetic acid and phenol have been successfully
removed from clay soils in laboratory samples [Shapiro et al., 1989b]. More than 94% of dissolved acetic
acid (0.5 mole/l) and phenol (450 ppm) were recovered from kaolinitic soils [Shapiro and Probstein, 1993].
Ongoing bench studies are examining the viability of electro-osmotically enhanced removal of BTEX, TCE
and other non-polar organic compounds [Bruell et al., 1992; Acar, 1992; Marks et al., 1992; Segall and
Bruell, 1992]. In one kaolinite soil column study, 25 wt% of TCE (150 ppm) was removed within 5 days
[Bruell et al., 1992],
Applicability/Limitations-
Electro-osmosis pertains mostly to the removal of ionic species, namely dissolved heavy metals,
radionuclides, and charged (valence bearing) organic compounds whose sign (if any) will depend on the
ambient ground-water pH and the compound's pKa. Therefore, it may have applicability to such
compounds as pentachlorophenol. The ability of electro-osmosis to recover dissolved organic compounds
is dependent on their sorption properties and solubility [Bruell et al., 1992]. It remains to be seen if
DNAPLs can be mobilized by electro-osmosis [Mitchell, 1991]. Electro-osmosis is best applied in fine
grained soils having active and plastic clays and low void ratios [Mitchell, 1991]. In soils with high
concentrations of electrolytes, the soil zeta potential may drop to zero thereby ceasing electro-osmotic flow;
or, if the zeta potential is negative, double layer charge reversal may result in electro-osmotic flow to the
cathode [Casagrande, 1952]. Subsurface structures like utilities and other objects such as drums and scrap
metals will interfere with the electro-osmotic process [Lageman et al., 1989].
53
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- Generate'
or main
PURIFICATION PURIFICATION
_— Circulation system
Boundary of electrokmelic treatnieni
Figure 3.3.1.3 Schematic of an in-situ electro-osmotic extraction system [Lageman, 1989].
Anode array
Air compressor slab
Figure 3.3.1.4 Schematic layout of electrode arrays for the in-situ electrokinetic application at Baton
Rouge field test site [Acar, 1992].
54
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Desaturation, thermal drying and pore chemistry changes may result in cracking of the clay soils
[Mitchell, 1991]. Corrosion of the electrodes may affect the efficiency of electro-osmosis [Segall and Bruell,
1992; Shapiro and Probstein, 1993]. Gases may be evolved at the electrodes depending on electrode
composition, power input, and the prevailing electrochemical reactions taking place [Segall and Bruell,
1992]. Electrode corrosion can be mitigated by electrode type and composition (graphite vs. metal), and
by cycling or injecting of fluids [Lageman, 1989; Mitchell, 1991; Acar, 1992; Segall and Bruell, 1992].
Cost and Availability-
This technology has been demonstrated on the full-scale in both geotechnical and environmental
applications. Electro-osmosis equipment for geotechnical applications is readily available, and the design
criteria are well established [Casagrande, 1952]. Geokinetics holds the European patent for in-situ electro-
osmosis for electrokinetic remediation. Probstein et al. (1991) have been awarded U.S Patent (5,074,986)
for "Electro-osmosis Techniques for Removing Hazardous Materials from Soil." Another U.S. patent
(5,137,608) entitled "Electrochemical Decontamination of Soils and Slurries" has been awarded to Louisiana
State University [Acar, 1993].
However, no known full-scale applications of electro-osmosis pertaining specifically to DNAPLs in
the saturated zone are known to exist. Electro-osmosis is generally not a good candidate for DNAPL
cleanup because DNAPLs are generally nonpolar, making them generally unsusceptible to electrical fields.
Furthermore, electro-osmosis is best applied in saturated fine (clay) soils where major quantities of DNAPL
are usually not found.
Electro-osmosis costs are dependent on initial contaminant concentration, energy supply and time
duration. For low energy delivery over long periods (months), the total treatment can be as low as $50/ton
which can increase up to $400/ton for short (weeks) energy intensive applications [Lageman, 1989]. Of
the total cost, the electric power costs are typically in the range of $2-20/ton of remediated fine grained
soils [Marks et al., 1992; Shapiro and Probstein, 1993].
3.3.2 Electroacoustic Soil Decontamination (ESP)
Theoretical Background-
Electroacoustic Soil Decontamination (ESD) employs both electrical and acoustical (pressure)
gradients to enhance organic contaminant removal. The electrically-induced phenomena and removal
mechanisms are identical to electro-osmosis (see section 3.3.1). It may be convenient to think of this
technology as a hybrid of electro-osmosis and radio frequency heating (section 3.7.3), except that the lower
applied frequencies (usually 100-1000 Hz) do not result in appreciable soil heating; but they do enhance
fluid flow through the porous media. The primary removal mechanisms and phenomena derived from
acoustical fields which are believed to contribute to overall contaminant removal include: orthokinetic forces,
Bernoulli's forces, rectified diffusion, "rectified" Stake's forces, decreased apparent viscosity and radiation
pressure [Muralidhara et al., 1990]. Since the elements of electroosmosis have been discussed previously,
attention will be given here to the prevailing acoustical phenomena.
Acoustic fields generate fluctuating (sinusoidal) pressure waves which vary as a function of time
and position [Muralidhara et al., 1990]. Buried acoustical sources will produce mainly compression and
shear waves and negligible surface waves [Muralidhara et al., 1990]. Wave intensity decreases with the
inverse of distance squared, while soil attenuation of waves increases with the square of frequency.
Travelling waves impart mechanical energy to particles in the form of velocity. Particle velocities are seen
to be a function of applied frequency, and are related to the acoustic pressure through the acoustical
impedance of the porous medium [Muralidhara et al., 1990], Which of the previously mentioned removal
mechanisms and phenomena will prevail in-situ is in large part governed by the physical and chemical
properties of the porous medium. Since the process is neither fully understood [USEPA, 1992d] nor
55
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accurately predictable (a priori) [Muralidhara et al., 1990], the approach is empirically based.
The following descriptions of the contributing phenomena have been summarized from Hinchee
et al. (1989) and Muralidhara et al. (1990). Orthokinetic and Bernoulli's forces refer to the forces that cause
small and large particles to agglomerate, respectively. Cavitation of pore fluids and gas bubble generation
within soil particle capillaries with the resultant expulsion of trapped pore fluids from minipores and
macropores aid in dewatering, this is referred to as rectified diffusion. Decreased apparent viscosities are
thought to arise from the high strain rates and localized heating. The nonlinear spatial variation in fluid
viscosity, which aids in transport to the source, is referred to as the rectified Stoke's force. Radiation
pressure, a static pressure, is a second order effect which adds to the normal pressure differential.
Of the acoustically induced phenomena, particle rearrangement and viscosity reduction are believed
to be among the major factors contributing to the contaminant removal. Ex-situ applications of electrical
and acoustic fields to dewatering of sludges and soils has indicated that the dewatering effects are
synergistic [Hinchee et al., 1989]. It has been postulated [Hinchee et al., 1989; Muralidhara et al., 1990]
that the rearrangement of particles creates new flow channels which enhance fluid flow and, therefore,
electro-osmosis.
Field Implementation-
A conceptual layout for a vadose zone application of ESD is shown in Figure 3.3.2.1. Since this
technology is still under development, details on the construction and configuration of the requisite acoustic
sources are not available. However, an acoustical electrode or array is likely to be similar in certain
respects to the radio frequency devices (section 3.7.3). The electro-osmotic elements and their
implementation have been previously described (section 3.3.1), and the strategy for laying out the
electrodes for ESD is essentially analogous to that of electro-osmosis.
Water
+
Contaminants
Ground
Surface
Optional
Anolite Treatment
- Contaminants
Water (Optional)
Cathode J Acoustic
(-) Water Source
Velocity
Profile
Anode
Figure 3.3.2.1 Conceptual layout of the electroacoustical soil decontamination process [USEPA, 1992d].
56
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Level of Demonstration and Performance--
To date, ESD has not been implemented in the field. Bench scale results indicate that metals (zinc,
cadmium) were removed more effectively than decane (a non-polar organic compound) [Hinchee et al.,
1989; Muralidhara et al., 1990]. The soils tested were classified as clay loam, silty and sandy clays, and
clay. The clays were slightly acidic (pH= 5.5) and had organic matter contents of 1.85 wt%. Samples had
a diameter of 3.5 in and length dimensions of 2 to 8 in depending on the test. On a dry soil basis, samples
were uniformly spiked with either 2,000 ppm zinc, 11,000 ppm zinc, 1,000 ppm cadmium, or 8 wt% decane.
In the zinc removal experiments, the soil moisture contents were on the order of 42 wt%. Constant
currents of 50 amps were applied, and the voltage drop across the 4.5 in long samples increased from 0.3
to 20 V/in during the test. Electrical energy consumption was 1.423 watts, and no acoustics were applied.
In 100 hours of operation, more than 90 wt% of the zinc was removed from three-quarters of the sample.
Precipitation of zinc hydroxide was detected near the cathode. Zinc removal efficiencies were found to
increase with longer test durations and elevated power levels. Similar results were obtained for cadmium
removal.
The decane spiked soil samples were 2.5 in long and had approximate weight proportions as
follows: solids, 52.7%; water, 39.3%; decane, 8%. Four tests were completed with average applied
currents of 0.11 amps, with voltage drops ranging from 25 to 45 V/in. Of these four samples, acoustic
energy was applied to only one sample at 400 Hz and 0.697 watts. The test duration was 2 hours.
Decane removal was estimated to be between 10-25 wt%. However, the positive effect of the acoustical
field on decane removal could not be confirmed owing to data discrepancies between the two analytical
laboratories [Muralidhara et al., 1990]. Deoiling of petroleum sludges and removal of jet fuel from sandy
soils are reported to have been successful with electroacoustics [Hinchee et al., 1989]. However, the exact
contribution of acoustics above and beyond that of simple electro-osmosis was not reported.
Applicability/Limitations-
Essentially the same as electro-osmosis.
Cost and Availability-
This technology has been demonstrated on the lab scale only. No field- or full-scale applications
of any kind have been attempted. The technology is currently under development by Battelle Memorial
Institute as part of the USEPA SITE Emerging Technology Program [USEPA, 1992d]. The process has
been patented by Battelle Memorial Institute [USEPA, 1992d].
Electroacoustics does not appear to be a good candidate for DNAPL cleanup because DNAPLs
are usually nonpolar making them relatively unsusceptible to electrical fields. Furthermore, electroacoustics
is best applied in saturated (fine) soils where major quantities of DNAPL are usually not found.
The cost is similar to electro-osmosis, plus acoustical hardware and incremental energy cost. ESD
should be comparable to RF heating coupled with soil venting (i.e., $40-100/ton) [Muralidhara et al., 1990].
57
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3.4 CONTAINMENT AND GROUND MODIFICATION
Introduction--
Containment systems and ground modification methods are used to contain and immobilize
dissolved contaminants and, in certain cases, DNAPLs. Containment systems are usually placed on the
periphery of the contaminated area or along specified boundaries so that the encompassed area becomes
effectively isolated from its surroundings, thereby preventing further spreading. Impermeable barriers and
ground-water injection/extraction systems are examples of containment systems. The ground modification
methods are usually confined to DNAPL source areas, and aim to immobilize or neutralize the
contaminants. Stabilization/solidification (S/S), vitrification, permeable treatment walls, and variations of
these methods are examples of ground modification. Containment and ground modification can be either
passive or active, the distinction being made on the required energy expenditure after installation.
Impermeable barriers constructed of soil-bentonite (SB) slurry walls, composite geomembrane-slurry
walls, grout curtains, and sealable joint (sheetpile) cut-off walls are the focus of section 3.4.1. The primary
issues related to these systems are cost, durability, compatibility, and constructibility [Elsevier Science,
1989; ASCE, 1990; ASTM, 1990a,b]. These are passive systems which rely on low hydraulic conductivity
to inhibit contaminant migration [Mitchell and van Court, 1992]. Active ground-water extraction and
recharge systems are also briefly evaluated in section 3.4.1.
Stabilization/solidification (S/S) by in-situ soil mixing is addressed in section 3.4.2. Immobilization
of contaminants is achieved by neutralization, precipitation, sorption, and physical encapsulation of the
contaminants within a solidified soil matrix. In the broadest sense, many treatment technologies can be
considered as stabilization technologies [USEPA, 1986b]. In-situ vitrification (ISV) is a good example of
a technology which achieves both stabilization and solidification, but has been placed in section 3.7
because solidification and stabilization are attained by heat application. The major issues surrounding in-
situ S/S are chemical compatibility, and the durability and leachability of the treated soil mass.
In-situ permeable treatment walls (section 3.4.3) are granular backfill walls which provide treatment
of dissolved contaminants but no containment or immobilization. The composition of the porous backfill
(additives, surface coatings, etc.,) can promote favorable conditions for in-situ biodegradation, precipitation,
and chemical oxidation or reduction. The major issues regarding in-situ permeable treatment walls pertain
to changes in ground-water flow direction, clogging, long-term performance, and incomplete treatment of
wastes.
Isolation and containment systems and in-situ S/S are commercially available and have been
successfully demonstrated on the full scale. In-situ permeable treatment walls are being tested on the pilot
scale. All of these technologies except hydraulic controls constitute permanent structures.
3.4.1 Isolation and Containment
Theoretical Background-
Passive systems such as impermeable barriers and active systems such as hydraulic controls are
commonly used to attain dissolved plume or contaminant source area isolation and containment. Grout
curtains, geomembranes, in-situ soil mixed zones, and clay slurry walls, caps, and liners are impermeable
barriers which rely on low hydraulic conductivity to inhibit contaminant migration [Mitchell and van Court,
1992]. Ground-water extraction and recharge via arrays of wells and trenches facilitate hydraulic gradient
control, create capture zones, and permit redirection of local ground-water flow [Mercer and Cohen, 1990;
Mitchell and van Court, 1992]. The focus here is on slurry walls, grouting, sealable joint cut-off walls, and
hydraulic controls. In-situ soil mixing is described in the next section (3.4.2), owing to its similarity to in-situ
stabilization/solidification.
58
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Slurry walls can be constructed of clay (usually bentonite), cement, pozzolans (polymers, resins,
asphaltic emulsions), and native soils [McFarlane and Holtz, 1992]. The lower cost and hydraulic
conductivity, and greater plasticity and chemical resistance of soil-bentonite (SB) slurry walls usually make
them preferred over cement-bentonite (CB) walls [D'Appolonia, 1980; Ryan, 1980, 1987]. Plastic concrete
and concrete cut-off walls can also be used [Evans et al., 1987; ASTM, 1992]. During the construction of
a slurry wall, a filter cake develops on the walls of the trench. The filter cake also serves to reduce the
hydraulic conductivity. Experience has shown that the lowest filter cake permeabilities are attained when
the slurry mixture contains approximately 5-7 wt% bentonite, producing slurry viscosities greater than 40
sec-Marsh [D'Appolonia, 1980]. The hydraulic conductivity of the SB backfill decreases with increasing
fraction of fines appearing in the backfill material and the final percentage of bentonite in the SB wall, as
shown in Figure 3.4.1.1. Since the initial composition of the backfill becomes relatively unimportant at
bentonite contents greater than 2 wt%, this minimum limit is often recommended or used [D'Appolonia,
1980; Millet et al., 1992; Mitchell and van Court 1992]. In order to decrease the hydraulic conductivity,
composite slurry walls have been designed incorporating inclusions such as concrete panels, sheetpiles,
and, most recently, geomembranes into the traditional SB cut-off walls [Ryan, 1987; Hayward Baker, 1991;
ASTM, 1992]. Design criteria for the use of geosynthetics [Koerner, 1990], and clay cap and liner systems
[ASCE, 1990; Goldman et al., 1990] are readily available.
Penetration and jet grouting are the two main grouting techniques for construction of grout curtains
[Mitchell, 1981; Ryan, 1987; Mitchell and van Court, 1992]. Penetration (pressure) grouting entails the
pressure injection of paniculate or fluid slurries to fill interparticle voids and fissures. Jet grouting uses high
pressure nozzle injection to destroy the soil fabric thereby mixing the soil and slurry in-situ. Grout design
parameters are usually water/solids ratio, viscosity, bleed, and permeability [Weaver et al., 1992].
Penetration grouting hole spacings are usually on the order of 1.3 to 2.5 m. Jet-grouted soil column
diameter can be on the order of 0.3 to 1.5 m [Gazaway and Jasperse, 1992; Mitchell and van Court, 1992],
and the spacing of the columns is somewhat less to provide for overlap of adjacent soil columns.
Overlapping patterns using at least two to three grout injection rows are often used to ensure continuity of
the grout barrier as shown in Figure 3.4.1.2 [Gazaway and Jasperse, 1992; Mitchell and van Court, 1992],
but in some instances this convention has been waived [Weaver et al., 1992]. Hydraulic conductivities of
grouted soil samples can be on the order of 10"5-10"8 cm/s depending on the grouting method and media
type [Gazaway and Jasperse, 1992; Weaver et al., 1992].
Sealable joint sheetpile cut-off wall systems have been recently proposed [Starr et al., 1991; Starr
and Cherry, 1992]. To reduce leakage through sheetpile joints, the joints are sealed using bentonite slurry
or proprietary polymers which lower the overall hydraulic conductivity of the wall. However, because
sheetpiles are installed by driving, this technology may cause cracking and damage to aquitards such as
stiff clay soils.
Active ground-water controls are often used in conjunction with impermeable barriers to minimize
the potential for advective transport of contaminants across the impermeable barrier, as shown in Figure
3.4.1.3. Judicious well pumping may be used to alter the flow direction of plumes. Depending on the
quantity of DNAPL and the overall remedial strategy, more aggressive ground-water pumping approaches
may be practiced (see section 3.5.4). However, the cost of pumping and water treatment in absence of
cut-offs can be very high.
Field Implementation-
The configuration of the containment barrier is dictated by such factors as the overall remediation
strategy, cost, site limitations, liability reduction, preservation of uncontaminated or drinking water supplies,
and the ability to contain the contaminant by other means. Two approaches are commonly used: (1) the
DNAPL source areas and dissolved plumes are completely encompassed with a continuous barrier, or (2)
the dissolved plume migration is locally arrested or stalled using discrete barrier sections. Vertical barriers
59
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10'
ICP
o
O I0"6
tr
UJ
CL
WELL GRADED
COARSE GRADATIONS
(30-70% + 20SIEVE)
W/IO TO 25% NP FINES
POORLY GRADED
SILTY SAND W/
30 TO 50V. NP FINES
CLAYEY SILTY SAND
W/ 30 TQ50% FINES
J I
_L
01234
% BENTONITE BY DRY WEIGHT Of SB BACKFILL
Figure 3.4.1.1 Relationship between the permeability and bentonite content of SB backfill materials
[D'Appolonia, 1980].
Drilling Pattern
SECONDARY
COMPLETED OVERLAPPING
AND COMPLfc i h TREATMENT
Figure 3.4.1.2 Primary and secondary overlapping patterns for in-situ soil mixing processes [Geo-Con,
Inc., 1990].
60
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SI URR1 WALL
GRAVEL-FILLED TRENCH
-DRAIN PIPE
CONTAMINATED GROUNDWATER
Figure 3.4.1.3 Schematic configuration of a coupled impervious barrier and hydraulic gradient control
system. Groundwater flow across barrier is maintained into contaminated groundwater
region [Ryan, 1987].
n
^ ^ ^-^r^
| [\V/ BACKHOt "^^
Y KEYS TRENCH ^\
/"
\^^
~\ CLAY LAtEP J/
)
BACKF'L^
PuACE Z
HERE — -
LEVEL
"BENTONITE SLURRY \
BACKFILL
SLOUGHS
FORWARD
Figure 3.4.1.4 Schematic of SB slurry wall installation process [Ryan, 1980].
are usually grouted into underlying aquitards or bedrock. Detailed descriptions of slurry wall installation
and grouting technology are available in a number of publications [D'Appolonia, 1980, Mitchell, 1981; Ryan
1987; ASCE, 1992; ASTM, 1992].
Figure 3.4.1.4 illustrates slurry wall installation using the trench method. Trenches are typically 2-3
ft wide, and trenches more than 100 ft deep and 1,000 ft long have remained stable and open for several
weeks between excavation and backfilling [D'Appolonia, 1980]. Construction to depths of 400 ft have been
reported [Ressi di Cervia, 1992], and materials such as weathered shales and conglomerates, sands,
gravels, clay, and tills have been used successfully as backfill materials in geotechnical practice
[D'Appolonia, 1980]. Well mixed SB backfills having a slump of 2-6 in and a minimum viscosity of 40 sec-
Marsh are considered as ideal for placement [D'Appolonia, 1980; Ryan, 1987]. To ensure that the SB
backfill efficiently displaces the slurry from the trench, its unit weight should be at least 15 Ib/ft2 greater than
that of the slurry [D'Appolonia, 1980; Ryan, 1980, 1987].
61
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1
I
Penetration (intrusion)
Penetration (permeation)
\
Grouted column
Ro:a!e and t>
Grout iet
n.iplacement
(cc -i(i iction grouting)
Jet grouting
(displacement, replactv ent)
Figure 3.4.1.5 Schematic showing different grouting techniques [Hausmann, 1990].
Figure 3.4.1.5 depicts the penetration and jet grouting installation methods. Penetration grouting
can be accomplished within the entire soil horizon by packing off lower sections after grouting. A jet grouted
soil column is created as the rotating nozzle injector is removed from the hole. Jet grouting can also be
facilitated in panel sections using directional injection [Mitchell and van Court, 1992]. Available grouting
materials include particulate grouts (clay, lime, fly ash, microfine cement) and chemical grouts (bitumen,
phenolic and acrylic resins, silicates) [Hausmann, 1990; Mitchell and van Court, 1992]. Grout selection is
affected by such factors as the hydraulic conductivity of the porous media, cost, permanency, and
compatibility [Hausmann, 1990]. Penetration grouting is quite generic and flexible and it can handle a
variety of conditions. Typical parameters for jet grouting include: nozzle diameters of 1.8-2.2 mm diameter;
injection pressures up to 6,000 psi; and injector rotation rates and lift rates of 1-2 rpm and 1-1.5 fl/min,
respectively [Gazaway and Jasperse, 1992]
62
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Level of Demonstration and Performance--
Slurry walls have a long history of use. One recent application features the installation of an
approximately 3 km long composite high density polyethylene (HOPE) SB slurry wall at an industrial facility
in Liguria, Italy [Manassero and Viola, 1992]. The contaminants were unspecified. The subsurface soils
at this site were characterized by 3 to 15 m high permeability sands and gravels underlain by marl. Depth
to ground-water at the site was 1 to 8 m. The cross section through the composite wall is shown in Figure
3.4.1.6. The construction sequence entailed the initial installation of a 1.2 m thick provisional slurry wall
to bedrock using the clamshell method. After the wall had set, a combination of bedrock drilling and
clamshell excavation facilitated removal of the middle 0.5 m of the provisional slurry wall to an approximate
depth of 2 m into bedrock. The excavated portions of the wall were kept open using a tremied CB mixture
which also served as the final slurry wall. Prior to setting of the cement, 2.5 m wide geomembrane panels
having tongue-and-groove connections with hydrolite seals were lowered using guides into the center of
the final slurry wall. In-situ tests revealed that the hydraulic conductivities of the keyed areas and
composite wall were on the order of 10"6 and 10~7 cm/s, respectively.
Grouting technology has also been used to handle complex subsurface conditions. Jet grouting
was recently implemented at a chemical plant in Michigan with unspecified subsurface contamination
[Gazaway and Jasperse, 1992]. The project entailed the construction of a jet-grouted cut-off wall through
a utility corridor to join two existing slurry walls as shown in Figure 3.4.1.7. Numerous utilities ranging from
diameters of 2 to 48 in appeared within the utility corridor up to depths of 17 ft. The jet-grouted wall with
a design permeability of 1x10"6 cm/s was installed to a depth of 24 ft and was keyed into both existing
slurry walls. Field tests indicated that soil columns with diameters of up to 4-5 ft (1.2-1.5 m) were possible,
and a conservative spacing of 2 ft was adopted. Operating parameters included: injection pressures up
to 6,000 psi and injector rotation rates and lift rates of 1 rpm and 1.3 ft/min, respectively. The rotation and
lift rates were decreased near the larger utilities to ensure good mixing and void filling. In some cases the
injection pressures were reduced as well. The 5,700 ft2 wall was completed within 3.5 weeks with no
detectable damage to the utilities and no stoppage in utility service.
Applicability/Limitations-
Slurry walls [ASTM, 1992] have been installed at sites with difficult and complex conditions [ASCE,
1992; ASTM, 1992; Davidson et al., 1992]. Impermeable barriers are generally used to isolate and contain
dissolved phase contaminants. The actual remediation of pure and dissolved phase DNAPL requires other
techniques. Hydraulic controls, on the other hand, may be utilized for containment and partial recovery of
DNAPL.
The primary issues related to these systems are cost, durability, compatibility, and constructibility
[Elsevier Science, 1989; ASCE, 1990; ASTM, 1990a,b]. Chemical compatibility and permanency of slurry
walls and grouted barriers are major issues. Clay barriers are susceptible to cracking and leaking when
exposed to concentrated solutions of acids, caustics, and non-polar organic compounds [Fernandez and
Quigley, 1985; Mitchell and Madsen, 1987; Madsen and Mitchell, 1989; Quigley and Fernandez, 1992].
In general, clay barriers are not affected by dilute organic solutions, and the effects of inorganic solutions
are consistent with their effects on clay particle double layers, surface and edge charges, and pH [Madsen
and Mitchell, 1989]. While CB walls are more chemically resistant, they are prone to cracking because of
their brittle nature. For these reasons, ductile CB walls have received recent attention [Evans et al., 1987;
Millet et al., 1992], Some of the same general trends are observed for grouts, but the actual response is
contaminant and grout specific due to the many grout mixes and grout pozzolans available [Spooner et al.,
1982, 1984; May et al, 1986; Bodosci et al., 1987]. Sulphate attack is one of the most common problems
related to grouts [Jefferis, 1992].
Penetration grouting is best suited for sealing fissures in bedrock [Ryan, 1987]. However, May et
al. (1986) indicate that the main disadvantage of penetration grouting is that the grout cannot be controlled
63
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Final slurry wall
;:i]jl~~ Altered marl
EL caprock
i
Figure 3.4.1.6 Vertical section taken through a composite geomembrane-SB slurry wall impervious barrier
system, Liguria, Italy. [Manassero and Viola, 1992].
Drainage Swale
Soil/Bentonite
Slurry Wall
USoil/Bentonite ., ~,
Slurry Wall_/A/" t
7 , 0 / A*~ Jet Grout
Cutoff Wall
Figure 3.4.1.7 Vertical section taken through utility corridor in which jet grouted impervious barrier was
constructed to join SB slurry walls [Gazaway and Jasperse, 1992].
64
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to ensure a complete seal. Grout curtain continuity is difficult to verify, and this is among the reasons why
slurry wall installation is favored [Ryan, 1987]. Similarly, poor slurry wall construction will result in the
entrapment of pockets of slurry and other wall heterogeneities which may lead to increases in permeability.
Composite slurry walls using synthetic panels as inclusions offer challenges regarding wall alignment.
Placement of slurry walls is limited to areas without ground structures, buried utilities, other subsurface
obstructions, and above bedrock. In such situations, the installation of jet-grouted impermeable barriers
in utility corridors has been successful [Gazaway and Jasperse, 1992].
Cost and Availability--
Slurry wall technology has been commercially available for some time, and hundreds of slurry walls
have been installed to depths [D'Appolonia, 1980; Ryan, 1987]. Depths of more than 30 m are common.
The installation is typically provided by a specialty contractor. While grouting technology dates back
centuries [Bowen, 1981] and there are numerous contractors providing this service, the applications in site
remediation are relatively new. Slurry walls and grout curtains can be designed by qualified professionals.
If properly installed, impermeable barriers are usually good candidates for isolating the DNAPL
source and for preventing the migration of dissolved contaminants. Grout curtains and slurry walls can be
used heterogeneous soils; however, their effectiveness in fractured bedrock is limited.
Isolation and containment systems usually cost less than other remedial alternatives [Ryan, 1987;
Ellison, 1992]. Backhoe slurry wall construction ranges from $7-13/ft2 and in-situ grouting (using drill rig)
ranges from $60-100/yd3 [Ellison, 1992].
3.4.2 Stabilization and Solidification
Theoretical Background-
Stabilization and solidification (S/S) methods are intended to immobilize dissolved contaminants
and, in certain cases, DNAPLs. Stabilization refers to techniques that reduce contaminant hazard potential
by converting the contaminants to less soluble, mobile, or toxic forms, which does not necessarily imply
a change in physical nature and handling characteristics. Solidification refers to techniques that
encapsulate the contaminant in a monolithic solid of high structural integrity [Conner, 1992]. These two
ends may be achieved by a variety of means, and often occur simultaneously. S/S has been applied to
a wide variety of wastes, including sludges, waste pits, liquids, lagoon sediments, and contaminated soils
on an ex-situ and in-situ basis. Good reviews of S/S processes and technology are available [Tittlebaum
et al., 1985; Cullinane et al., 1986; Wiles, 1987; USEPA, 1989a,b; Jones, 1990; Conner, 1992]; and,
therefore, only a brief overview is given here.
S/S processes are distinguished by reagent type into inorganic and organic processes [Conner,
1992]. Organic systems (urea formaldehyde, polyethylene, bitumen, asphalt emulsions) attain S/S by
thermoplastic encapsulation and by polymerization [USEPA, 1989b,c], and these processes pertain mostly
to radioactive wastes [USEPA, 1989b, Conner, 1992]. Inorganic systems utilize a combination of cementing
agents (cement, lime) and bulking agents to encapsulate and/or mechanically bind the waste material within
a solid matrix to restrict contaminant migration [USEPA, 1989b; Conner, 1992]. Bulking agents are
admixtures that contribute to the solids content and viscosity of the waste and prevent suspended waste
components from settling out before solidification occurs [Conner, 1992]. Bulking agents can be inert or
pozzolanic. Pozzolanic S/S admixtures (kiln dust, fly ash, organophilic clays, proprietary admixtures) are
aluminosilicate materials that do not possess cementing behavior in themselves, but form cementitious
materials when combined with lime, cement and water [USEPA, 1989b; Jones 1990; Conner, 1992]. Data
suggest that the use of silicates in conjunction with lime, cement and other setting agents can stabilize a
broader range of materials (including oily sludges and soils) than cement based mixtures [USEPA, 1988].
65
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In addition to physical isolation of organics in an impermeable matrix, other chemical fixation
mechanisms such as neutralization, precipitation, partitioning and sorption are thought to be active [Martin
et al., 1990]. Since cementitious reactions occur at a pH greater than 10, hydrolysis, oxidation, reduction
and other compound reactions involving organics may contribute to their immobilization and/or ultimate fate
[Conner, 1992].
The exact nature of the reactions occurring in cement-pozzolan-soil-waste systems is not clearly
understood [USEPA, 1989b, Jones, 1990, Conner, 1992]. Pozzolan chemistry is not well known, but is
thought to be analogous to that of portland cement [Jones, 1990]. Waste components such as polar
organics, salts, and certain heavy metals often interfere with cement and pozzolan chemistry and may
retard or altogether stop setting and hardening of the paste [Wiles, 1987; Jones, 1990]. Also, admixtures
may be used to retard the setting process, but they ultimately produce treated soil of greater integrity and
strength [Jones, 1990; Conner, 1992]. Non-polar organic wastes of low volatility should not hinder strength
development in cement or pozzolan systems [Jones, 1990]. Because the reaction chemistry and the
corresponding immobilization mechanisms are poorly understood, treatability studies are required. The S/S
processes are therefore empirical.
The strength and durability of the concrete products are directly related to the number of voids in
the final product [Jones, 1990]. The integrity of the monolithic mass is assessed using a number of testing
criteria such as [USEPA, 1989b, Jones, 1990]; index property (suspended solids, pumpability); density;
permeability; strength; durability (wet/dry and free/thaw); and contaminant leaching (EP TOX, TCLP
analysis) [USEPA, 1989b].
Field Implementation-
Once the S/S reagent blend is selected, it can be administered in-situ using a variety of
conventional construction equipment and techniques [USEPA, 1989b,c]. In order to minimize contaminant
volatilization and to ensure more complete mixing [USEPA, 1989c], in-situ soil mixing using rotating auger
heads is preferred. For instance, Geo-Con, Inc., has developed shallow (SSM) and deep (DSM) soil mixing
technologies which are capable of achieving S/S of subsurface soils to depths of 10 m and 30 m,
respectively [Broomhead and Jasperse, 1992]. Furthermore, these processes are generic construction
technologies, and any combination of admixtures into the soil for purposes of S/S, soil strength
improvement, and reinforced wall and impermeable barrier construction [Jasperse and Ryan, 1992].
The SSM process is depicted in Figure 3.4.2.1. The process uses a crane-mounted multi-blade
auger having a diameter of 1 to 3.7 m for mixing of soft soils and sludges. Soil admixtures are fed from
the ex-situ mixing plant via hoses and a hollow kelly bar to the SSM auger. The soil admixtures enter the
soil via three ports located at the bottom of the mixing auger. Primary and secondary overlapping bore
patterns are used for area treatment, as shown in Figure 3.4.1.2.
The DSM process is depicted in Figure 3.4.2.2. This process uses a crane mounted rig consisting
of an interlocking series of four 36-in diameter auger flights. Soil admixtures from a mixing plant are
introduced into the soil via the hollow auger flights. Overlapping patterns are used to ensure complete
mixing and continuity of the treated area, as shown in Figure 3.4.2.1 and Figure 3.4.2.3.
The SSM and DSM processes also incorporate a vacuum hood at the ground surface in order to
capture any emissions that may be liberated during in-situ mixing [Jasperse and Ryan, 1992]. A
combination of horizontal and vertical leads, and template and guide systems are employed to ensure
proper alignment of soil mixed zones [Geo-Con Inc., 1989, 1990].
66
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BULK
STORAGE
TANKS
1REATMENT
TRANSFER
TANK
ACTIVATED
CARBON
DUST TREATMENT EXHAUST
COLLECTOR TANKS FAN
Figure 3.4.2.1 Schematic of crane mounted shallow soil mixing (SSM) process [Geo-Con, Inc., 1990].
9'-0"
Figure 3.4.2.2 Schematic of drilling pattern for deep soil mixing (DSM) process [Geo-Con, Inc., 1989].
WALL TYPE
GRID TYPE
BLOCK TYPE
AREA TYPE
Figure 3.4.2.3 Schematics of various final soil treatment patterns of SSM and DSM in-situ
stabilization/solidification processes [Geo-Con, Inc., 1989].
61
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Level of Demonstration and Performance--
Soil mixing technology was developed in the US in the 1950s, but to date it has been used mostly
in Japan for a variety of construction related purposes [Ryan, 1987; Broomhead and Jasperse, 1992]. The
SSM and DSM soil mixing technologies have been used more than a dozen times for
stabilization/solidification purposes [Jasperse and Ryan, 1992]. Environmental applications include waste
lagoons, sludges, pits and subsurface soils [Jasperse and Ryan, 1992]. In at least one case, DSM has
been applied to soils contaminated with DNAPLs (PCBs).
In 1988, DSM utilizing S/S admixtures supplied by International Waste Technologies was
demonstrated as part of the USEPA SITE Program at a site in Hialeah, Florida [USEPA, 1990a]. The soil
profile at the site consisted of sands, silty sands and karstic lime rock. The water table was situated at a
depth of 2 m. The average hydraulic conductivity of the soils was approximately 1.8x10"2 cm/s. Initial PCB
concentrations varied between non-detectable to approximately 1,000 ppm. Soil was contaminated to a
depth of 15 m, and in the affected areas PCB concentrations ranged from 200-600 ppm.
Bench scale studies indicated that a proprietary pozzolanic additive containing a treated clay
adsorbent was effective in chemically binding the PCBs and other organics [Jasperse and Ryan, 1992].
Two 10x20 ft test cells were treated to depths of 14 and 18 ft using a 3-ft diameter SSM auger. The auger
was rotated at 15 rpm, and the reagent addition rate was 0.18 Ib reagent/ 20lb of dry soil. Thirty-six soil
columns were used in an overlapping pattern to completely treat each test cell. The average properties
of the treated soil included: hydraulic conductivity 4x10~7 cm/s; unconfined compressive strength of 410 psi;
and a volume increase of 8.5%. The treated soil performed satisfactorily on the wet/dry test but not in the
freeze/thaw test.
TCLP analyses were performed on both untreated and treated soil samples. Untreated soil
samples containing PCBs below 60 ppm had no PCBs detected in the leachate by the TCLP, whereas
untreated soil samples having above 300 ppm PCBs had PCBs detected in their leachates. Between these
limits, the TCLP results were varied. The maximum PCB concentration in the treated soil samples was 170
ppm, with most being below 100 ppm [Jasperse and Ryan, 1992]. Blending of high- and low-contaminated
soils by the SSM auger is likely to have contributed to lower contaminant levels in the treated soil [USEPA,
1989a]. Since no PCBs were detected in the leachate, it appeared that they were effectively immobilized
according to TCLP analyses [USEPA, 1989a]. However, because of detection limit problems (as indicated
above), effective treatment could not be absolutely confirmed [USEPA, 1989a; Jasperse and Ryan, 1992].
Subsequent to the test, complete treatment of the site began in late 1990 and was completed in 1991
[Jasperse and Ryan, 1992].
Applicability/Limitations-
"S/S is one of the most important Best Demonstrated Available Technologies (BDATs) for both
'listed' and 'characteristic' wastes and will continue to be in the future" [Conner, 1992]. Effective treatment
of non-polar organic compounds (which include DNAPLs) has been demonstrated under certain conditions
[USEPA, 1989b]. However, many organics have been claimed to be effectively treated by S/S processes,
but little data is available for confirmation [Conner, 1992]. Treatability studies are required to assess
contaminant effects on the physical properties of the treated soil mass.
S/S is a good candidate for sites extensively contaminated with DNAPLs since the DNAPL will
become mixed throughout the treated soil column. However, to date, DSM has not been specifically used
for DNAPLs and it is not clear whether DNAPL migration from the treatment zone can be prevented.
Metals and organics can be treated simultaneously. Dissolved phase plumes and DNAPL in fractured rock
are better addressed by other techniques.
68
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Although PCBs constituted only ~1 wt% of the soil to be treated in the USEPA SITE demonstration,
the IWT/Geo-Con system is thought to be capable of treating wastes containing up to 25 wt% organics
[USEPA, 1990a,b]. The in-situ mixing process is intrusive, and surface and subsurface obstructions such
as boulders and concrete blocks must be removed. Drums, trash, and other refuse may be penetrated and
incorporated into the treated soil mass [USEPA, 1988]. Although S/S is a BOAT, durability, leachability and
longevity are matters which still require attention on a site specific basis.
Cost and Availability--
Cements, pozzolans and other S/S admixtures are commercially available, as are in-situ mixing
technologies such as the Detoxifier [USEPA, 1988] and SSM and DSM technologies [Jasperse and Ryan,
1992]. In-situ soil mixing technologies can also be used to deliver steam to the subsurface to drive off
volatiles during S/S [USEPA, 1988]. Typical costs for the SSM and DSM technologies are $20-50/m3 and
$100-200/m3, respectively, excluding reagent cost [Jasperse and Ryan, 1992],
3.4.3 Permeable Treatment Walls
Theoretical Background-
In-situ treatment walls are permeable and reactive structures installed using conventional slurry wall
construction technology The walls are constructed of granular materials to permit ground-water flow
through the structure under ambient ground-water gradients. Treatment is achieved by using a combination
of reactive granular backfill and a variety of additives or surface coatings such as [Gillham and Burris,
1992]: nutrients and bacteria for in-situ biodegradation; redox controls and/or metal catalysts to aid in
metals precipitation and chemical dehalogenation [Blowes and Ptacek, 1992; Xu and Schwartz, 1992;
O'Hannesin and Gillham, 1993]; organic carbon for enhanced denitrification; and selective sorbents to
increase the retardation capacity of the in-situ wall [Burris and Antworth, 1992; Chapman, 1992]. The
dissolved phase contaminants are exposed to the reactive amendments and/or microbial consortia in the
permeable treatment wall. Factors such as rates of reaction and the maintenance of favorable conditions
will affect the wall thickness and its longevity.
The remedial strategy requires that the treatment wall either span the entire width of the
contaminant plume or the plume be directed through the treatment wall. Figure 3.4.3.1 schematically
illustrates two possible wall configurations; an independently acting wall system (Fig. 3.4.3.1 a), and a
system employing impermeable wing walls which aid in the channeling of contaminated ground water to
the reactive sections (Fig. 3.4.3.1b).
Field Implementation-
Installation of in-situ treatment walls will depend on the desired wall configuration and its
composition. For example, continuous treatment wall sections such as that shown in Figure 3.4.3.1 a, can
be constructed in a manner analogous to slurry cut-off walls using biodegradable polymer slurries rather
than bentonite [Gillham and Burris, 1992]. For the geometry shown in Figure 3.4.3.1b, large diameter
borings filled with soil amendments are used in conjunction with cut-off walls. Using conventional
construction technology, installation depths greater than 100 ft are possible [O'Hannesin and Gillham,
1992]. In addition, in-situ treatment walls with hydraulic feed controls can be installed to deliver gaseous
or liquid phase amendments [Gillham and Burris, 1992],
Level of Demonstration and Performance-
In most cases, current development of in-situ treatment walls is at the conceptual and laboratory
scales [Blowes and Ptacek, 1992; Burris and Antworth, 1992: Xu and Schwartz, 1992]. At the laboratory
scale, many chlorinated aliphatics have been transformed under abiotic (iron reducing) conditions and
appear to be first order in nature [Gillham et al., 1993]. Zinc and iron appear to transform halogenated
aliphatics faster than other metals, and transformation rates were pH dependent [Gillham et al., 1993;
69
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Gillham and O'Hannesin, 1993].
No full-scale applications are known to exist [Gillham and Burris, 1992]. However, in 1991, one
pilot test was completed at the Canadian Forces Base (CFB) Borden in which dissolved TCE and PCE
were treated [Gillham and O'Hannesin, 1992; Gillham et al., 1993]. At the site, the depth to ground water
was approximately 1 m and the ground-water velocity was approximately 9 cm/d. The centroid of the
dissolved plume was 3 m below the water table and had a thickness and width of 1 m and 2 m,
respectively. The maximum dissolved concentrations of PCE and TCE were 43 and 250 ppm, respectively.
A sheetpile system permitted local dewatering and excavation of aquifer soils and their replacement
by amended soils. The in-situ treatment wall consisting of a mixture of iron grindings (22 wt%) and sand
(78 wt%) was constructed 5 m downgradient from the DNAPL source. The wall was constructed in the
depth interval of 3.8-6.0 m, and was 1.6 m thick and 5.5 m long. Rows of multilevel sampling wells were
placed in the aquifer approximately 0.5 m from the upgradient and downgradient faces of the wall, and
within the wall at distances of 0.5 and 1.0 m from the upgradient face. The 348 sampling points were
monitored for 500 days. Most transformation occurred in the first half of the wall. The results indicated
that the plume did not bypass the wall; and reductions of 91% and 95% in PCE and TCE concentration,
respectively, were achieved. Mass balances confirmed that chlorine concentrations appearing
downgradient of the wall were consistent with degradation of the two compounds. Trace amounts of
dichloroethene were detected which may have potentially resulted from abiotic or biotic transformations.
(b)
Figure 3.4.3.1 Conceptual plan views for possible configurations of in-situ permeable treatment walls
[Envirometal Technologies, Inc., 1992]
Applicability/Limitations-
Laboratory column studies have shown that in-situ soil amendments (treatment zones) can
effectively treat both inorganic and organic dissolved phase compounds [Blowes and Racek, 1992; Burris
and Antworth, 1992; Xu and Schwartz, 1992; Gillham et al., 1993]. Since most DNAPL is immobile, once
emplaced, the intended use of this technology is the management and treatment of the dissolved phase
70
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in aquifers; extension to fractured media is unlikely.
The active amendments must be reactive, non-toxic, and must be both soluble enough to supply
ample reagent mass for reactions and stationary enough to persist for long periods of time. Excessive
biological growth and precipitation may compromise the long-term performance of the permeable wall. Use
of inorganic catalysts that promote organic compound reduction may also foster anaerobic biodegradation
which may potentially occur within the wall or downgradient. If the treated compounds are not completely
dehalogenated, formation of toxic products such as vinyl chloride is conceivable. Since abiotic reactions
are aspecific, formation of toxic intermediate degradation products is also conceivable.
The placement of permeable treatment walls and impermeable wing walls is limited to areas without
ground structures, buried utilities, and other subsurface obstructions and heterogeneities. By design, in-situ
permeable treatment walls are intended to be passive; and no ex-situ treatment is needed. Thus, the
treatment costs associated with pump-and-treat can be eliminated. Because the system is passive, site
heterogeneity greatly affects site selection to ensure the plume flows through the wall. Seasonal effects
may cause the plume migration direction to change, making additional hydraulic controls necessary; i.e.,
hydraulic gradient controls such as slurry and sheetpile cut-off walls, and/or judicious well pumping.
Cost and Availability-
Degradation of dissolved contaminants has been shown on the pilot scale only. The technology
and elements required to construct and implement permeable treatment walls are readily available. The
Waterloo Centre for Groundwater Research (Canada) has patents pending on in-situ treatment wall
technology [O'Hannesin and Gillham, 1993].
Permeable treatment walls have potential promise for dissolved plume treatment in aquifers
providing the plume continues to flow through the wall. Since DNAPLs are generally immobile under
ambient groundwater conditions, it is expected that they will remain unaffected. In-situ permeable treatment
walls are thought to be less expensive than pump-and-treat [O'Hannesin and Gillham, 1993].
71
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3.5 SOIL WASHING PROCESSES
Introduction--
In-situ soil washing (or fluid flushing/flooding) relies on fluid-fluid displacement processes to
enhance contaminant removal. Fluids (alkalis, cosolvents, surfactants, water, etc.) can be injected into the
porous media in order to mobilize the resident pore fluids, water and DNAPL, by a combination of physical
forces which can be aided by favorably altering chemical partitioning so that bulk fluid properties change,
i.e., solubility enhancement and interfacial tension reduction. The exact nature of the displacement and
the prevailing physical and chemical behavior occurring in these systems depends on the liquid properties
and environmental conditions. In-situ soil washing using alkalis, cosolvents, surfactants, and water are
evaluated in sections 3.5.1 through 3.5.4, respectively.
The effectiveness of the displacement process is controlled by phase equilibria and the
hydrodynamics of frontal propagation in porous media. The mechanics of miscible and immiscible fluid
front propagation in homogeneous, isotropic porous media are well established [Buckley and Leverett,
1942; Miller, 1975; McWhorter and Sunada, 1990]. Factors affecting the orientation and shape of the
advancing saturation front include: matrix heterogeneities, fluid properties, geometry of the aquifer and
injection strategy, initial moisture and boundary conditions, and injection rates.
As a liquid progresses through the porous media, two types of fluid displacement may occur.
Miscible displacement characterizes the removal of resident pore fluids by a mutually soluble displacing
liquid (i.e., cosolvent) in which the displacing-resident fluid interfacial region is a continuous liquid phase
that is free of interfacial tension. Immiscible displacement occurs when the two fluids are mutually insoluble
and capillary forces arising from interfacial tension exist between the two fluids. Both displacement
processes consist of two recovery phases: primary and secondary. In a successful primary recovery phase,
a very large and concentrated "bank" of resident pore fluid is removed just prior to breakthrough of the
injected liquid. Secondary recovery occurs as a result of increased hydrocarbon solubilization or leaching
after the breakthrough of the injected liquid.
In saturated homogeneous isotropic porous media, the stability of the propagating front is related
to M and NG (mobility ratio and gravity number, see section 2.1) These ratios have been shown to be very
important in terms of viscous fingering, gravity override and effective sweep-out [Saffman and Taylor, 1958;
Basel and Udell, 1989]. If viscous forces dominate in saturated porous media, a propagating front is stable
with respect to gravity, and propagation can be reasonably predicted and controlled [Buckley and Leverett,
1942; Morrow et al., 1985; Udell and Stewart, 1989].
The mobility ratio, which neglects gravity and interfacial forces (Buckley-Leverett assumption), is
often used to evaluate the potential success of a proposed displacement. A value of M < 1 usually
indicates a favorable displacement [Buckley and Leverett, 1942; Muskat, 1982]. System miscibility may
not seriously affect primary recovery if M is low enough. For example, Everett et al. (1950) found that for
very low mobility ratios, miscible and immiscible displacement of oil from unconsolidated clean sands using
water at flow velocities on the order of 20 m/d could lead to primary recoveries as high as 100% and 91%,
respectively. Since ground-water flow velocities of 20 m/d are unrealistic, alternative methods have been
sought to favorably affect recovery at ambient ground-water velocities (1-5 m/d).
Alkalis enhance the removal of NAPL, heavy oils, creosotes, etc., that contain organic acids such
as carboxylic acids, phenolics, and asphaltenes by saponifying the organic acids from the NAPL which
results in natural surfactant production and, thus, interfacial tension reduction. Alkalis also disrupt
adsorption, precipitation and ion exchange processes between the pore fluids and the porous media which
reduces surfactant losses.
72
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Cosolvents rely on viscosity and density differences between the injected and resident fluids and
on solubility enhancement to overcome the capillary forces and to mobilize the DNAPL. For enhanced
primary recovery, a concentrated cosolvent slug is injected to locally create a displacing front that is
essentially one fluid phase composed of water-cosolvent-DNAPL which is free of surface tension effects.
Surfactants rely on interfacial tension reduction and enhanced hydrocarbon solubilization to improve
DNAPL recovery. The surfactant can be injected as a concentrated slug followed by a polymer for mobility
control to maximize primary recovery. If this is undesirable, a secondary recovery approach emphasizing
continuous injection of surfactants at low concentrations can result in significant compound recovery.
Waterflooding and ground-water extraction usually do not lead to appreciable recovery of DNAPL
and are therefore primarily used for-containment purposes. However, at extensively contaminated sites,
certain pumping strategies can enhance DNAPL mobilization.
All of these processes have been utilized separately and in combination for enhanced oil recovery
in the petroleum industry with varied success. The technologies are available, but are highly compound-
and site-specific. If administered properly, they can lead to significant DNAPL recovery. At this time,
environmental applications are being explored.
3.5.1 Alkali Soil Washing
Theoretical Background-
In-situ alkali soil washing involves the injection of alkaline agents such as sodium carbonate
(Na2CO3), sodium hydroxide (NaOH), and sodium orthosilicate (Na4SiO4) to enhance contaminant removal.
Alkalis enhance NAPL recovery by a number of mechanisms: formation of natural surfactants via NAPL-
alkali reactions; the porewater salinity is altered to enhance surfactant formation; precipitation of calcium
and magnesium hardness which enhances interfacial activity; reduction of surfactant adsorption to aquifer
solids; NAPL wettability changes; emulsification and coacervation of NAPL into a middle-phase emulsion;
enhanced NAPL ganglia mobilization as a result of interfacial tension reduction; coalescence of individual
NAPL ganglia into a NAPL bank; and displacement of the NAPL bank to a recovery well by viscous forces.
Alkalis have been frequently used in enhanced oil recovery methods either alone [Breit et al., 1981;
Janssen-van Rosmalen and Hesselink, 1981; de Zabala et al., 1982; Mayer et al., 1983], or in conjunction
with cosolvents and surfactants (see sections 3.5.2,3.5.3) [Reed and Healy, 1977; Krumrine et al., 1982a,b;
Clark et al., 1988; Manji and Stasiuk, 1988; Peru and Lorenz, 1990; Surkalo, 1990].
The effect of alkalis on the subsurface transport of petroleum NAPLs is complex. Johnson [1976]
and de Zabala et al. [1982] have enumerated the sometimes contradictory mechanisms taking place during
alkaline flooding of petroleum reservoirs: emulsification and entrapment; emulsification and entrainment;
emulsification with coalescence; wettability reversal (NAPL-wet to water-wet, or water-wet to NAPL-wet);
wettability gradients; oil-phase swelling; disruption of rigid films; and low interfacial tensions. The presence
of acidic components in petroleum NAPLs appears to be one unifying factor which is common to the
observed phenomena [de Zabala et al., 1982]. However, no correlation between the acid number of the
NAPL (oil) and its recovery has been established [Ehrlich and Wygal, 1977; Janssen-van Rosmalen and
Hesselink, 1981; Mayer et al., 1983]. The acidic components such as carboxylic acids, carboxyphenols,
phenolics, porphyrins, and asphaltene fractions of multicomponent petroleum NAPLs can form hydrolyzed
surfactant products when saponified in-situ by alkalis [de Zabala et al., 1982; Mayer et al.,1983]; and the
hydrolyzed surfactants, which are negatively charged, are presumed to be responsible for enhanced NAPL
recovery, not the alkalis themselves [de Zabala et al., 1982].
Alkali and surfactant soil washing differ in the nature of the passage of the surfactant through the
system. In surfactant flooding, the surfactant can either be continuously supplied or pulsed. In alkali soil
73
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washing, in-situ surfactant production is coincident with the hydroxide (OH") front created by the
concentrated alkali; and it continues until the surfactant precursors are completely depleted from the NAPL
mass [de Zabala et al., 1982]. Since the amount of surfactant precursors is finite, the surfactant effluent
history resulting from this in-situ process is analogous to a single pulsed injection of surfactant. Hence,
if the NAPL has not been mobilized by the time its acidic components have been saponified and the alkali
front has passed it by, the NAPL will probably remain emplaced, since no other surfactant will be supplied
to reduce its interfacial tension.
As the alkaline front propagates through porous media, the aqueous phase pH is elevated and
diffusion of acidic NAPL components into the aqueous phase is initiated. The frontal advance of alkali is
retarded by saponification of acidic NAPL components at the NAPL-water interface and, more importantly,
by reversible and irreversible reactions involving dissolved cations and the porous media [Ehrlich and
Wygal, 1977; de Zabala et al., 1982; Jensen and Radke, 1988]. Interactions with the porous media include
bulk mineral dissolution and reversible sodium-hydrogen ion exchange at mineral surfaces [Jensen and
Radke, 1988]. The saponification and reversible ion exchange reaction equilibria are depicted in Figure
3.5.1.1. The frontal propagation is chromatographic, and its rate of advance has been shown to be
dominated by reversible ion exchange reactions [de Zabala et al., 1982; de Zabala et al., 1986; Jensen et
al., 1986; Jensen and Radke, 1988]. While there is a tendency to use optimal pH-interfacial tension
relationships as the sole basis of design, the beneficial use of alkalis at the optimal pH may be annulled
because ion exchange interactions will involve alkali and thus alter pH [de Zabala et al., 1982].
High pH alkalis such as NaOH are obviously desirable because: the reversible ion exchange
reactions are overpowered allowing the OH" front to propagate faster; the surfactant precursor solubilities
concomitantly increase with increasing pH; and surfactant adsorption is mitigated by anion exclusion at high
pHs [de Zabala et al., 1982]. The pH behavior of several alkalis is shown in Figure 3.5.1.2. At high pHs,
solubilization of carbonaceous materials up to 0.5 wt% has been observed after alkali breakthrough [de
Zabala et al., 1982]. However, considerable alkali consumption and accelerated rock dissolution make
NaOH undesirable [Burk, 1987; Jensen and Radke, 1988; Peru and Lorenz, 1990], Therefore, alkali buffers
such as Na2C03 and Na4Si04 are desirable because they effectively solubilize surfactant precursors and
achieve NAPL mobilization at lower pHs, can be used at lower concentrations, and buffer against the
reversible sodium-hydrogen ion exchange reactions which consume OH and cause its frontal attenuation
[Jensen and Radke, 1988]. Buffered systems are usually characterized by an intermediate pH region which
occurs from ion exchange and alkaline buffer respeciation [Jensen and Radke, 1988]. The disparity of
breakthrough times for alkaline buffers vs. a non-buffered alkali (NaOH) is shown in Figure 3.5.1.3.
Figure 3.5.1.1 Schematic of alkali recovery process [de Zabala et al., 1982].
74
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140
130
120
100
0 I
05 I 0
ALKALI CONC (wt %)
50
100
Figure 3.5.1.2 pH comparison of commonly available alkali chemicals [Mayer et al., 1983].
Alkalis can significantly reduce interfacial tension, producing values on the order of 0.01 mN/m, as
shown in Figure 3.5.1.4. Figure 3.5.1.5 illustrates that interfacial tension reduction can be synergistic in
alkali-surfactant floods [Campbell, 1981; Manji and Stasiuk, 1988]. Co-injection of surfactants may also
circumvent NAPL recovery limitations imposed by the finite amount of in-situ saponification that can be
realized. Mobility control agents can be used with alkalis to improve sweep efficiency of alkaline flooding
[Burk, 1987]. Sodium chloride (NaCI) is often pre-injected or co-injected with alkalis to achieve the optimal
salinity for interfacial activity [de Zabala et al., 1982; Mayer et al., 1983].
Field Implementation-
The petroleum engineering literature is replete with field applications of alkaline flooding. Mayer
et al. [1983] provides a good summary of alkaline flooding projects. Details on the necessary equipment
are also available [Clark et al., 1988; Manji and Stasiuk, 1988; Sale et al., 1989].
Maximum NAPL recovery occurs when a concentrated and viscous middle-phase microemulsion
and NAPL bank are created near the advancing front of the alkali. This can be accomplished using a
concentrated alkali slug, usually followed by a mobility control agent [Mayer et al., 1983; Burk, 1987], or
via continuous injection [Mayer et al., 1983; de Zabala et al., 1982]. Alkaline injection can be continuous
for several pore volumes until breakthrough of the alkali and OH" front and/or the saponification capacity
of the residual NAPL is exhausted (i.e., in-situ surfactant production ceases). Tertiary injection of mobility
control agents may then commence if they were not co-injected with the alkaline agent.
Conventional injection and extraction well construction equipment can be used for in-situ alkaline
soil washing [Sale et al., 1989]. For extremely corrosive injectates, stainless steel well construction may
be required. Both horizontal and vertical well configurations have been successfully used (see below).
Well placement strategy depends on the nature and extent of contamination, soil heterogeneities, and
anticipated subsurface flow behavior once washing commences.
75
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50
30
20-
I 0
T 1
nmQ)c=OI8meq/IOOg
10
12
pH
80
>
.c
G
-c" 60
cr
n
o
jc
o
Qj
o 4 0
CJ
E
i—
2 0
„ No2CO
i?
pH
13
Figure 3.5.1.3 Comparison of experimental and theoretical alkali breakthrough times for NaOH, Na4SiO4,
and Na2CO3 as a function of pH [Burk, 1987; Jensen and Radke, 1988].
76
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K> ' 10 > 10 '
CONCENTRATION — WtV S,,C
Figure 3.5.1.4 IFT values of Wilmington Ranger zone crudes with alkalis at 52°C [Burk, 1987].
Level of Demonstration and Performance-
Enhanced recovery of petroleum hydrocarbons from oil reservoirs using alkalis either alone or in
combination with surfactants, cosolvents, and mobility control polymers has been found to significantly more
effective than conventional water-flooding [Mayer et al., 1983; Manji and Stasiuk, 1988; Clark et al., 1988].
Alkaline agents were employed with surfactants to enhance recovery of dense oils (SG=1.04, u=54 cp) from
a shallow alluvial aquifer at a former wood-treating site [Sale et al., 1989]. The description of the field test
is provided in section 3.5.3.
Applicability/Limitations-
Alkalis can be incorporated into almost any soil washing treatment of DNAPLs, providing a
compound can be selected such that the phase behavior and the resultant changes in bulk liquid properties
are favorable. Since most multicomponent DNAPLs are not likely to contain acidic components, in-situ
saponification is precluded, and surfactant must be supplied in these instances. The favorable influence
of alkalis will still be realized with respect to such factors as optimal salinity, hardness precipitation,
surfactant adsorption mitigation, and interfacial tension reduction. However, water-DNAPL interfacial
tension reductions below 5 dynes/cm have resulted in rapid downward vertical migration of DNAPLs in
laboratory studies [Fountain et al., 1991].
Engineered alkaline flooding in petroleum reservoirs has an entirely different focus than that of
environmental applications. In petroleum engineering, the scale of application, profit motive, and reservoir
conditions such as high NAPL viscosity and saturations, salinity, and geologic confinement favor
approaches geared toward high primary recovery; that is, the creation of concentrated and viscous middle-
phase microemulsions and banks having low interfacial tensions which promote good sweep efficiency.
In general, this approach is not well suited for environmental applications because of the relatively small
scale of contamination, environmental sensitivity, and liability. However, high primary recovery applications
can be considered as appropriate at sites having large scale contamination, as was the case at the former
wood treating facility in Laramie, Wyoming [Sale et al., 1989].
77
-------
o
UJ
1
001 01 1
CONCENTRATION (WT %)
O
uJ
001
0001 001 01
CONCENTRATION (WT %)
Figure 3.5.1.5 IFT values of Dome Lloydminster "A" pool crude as a function of alkali and surfactant
addition [Manji and Stasiuk, 1988].
Alkalis themselves often reduce the viscosity of water, which often promotes unfavorable mobility
ratios. If a more displacement-like operation is preferred over solubilization, any combination of compounds
such as polymers (density enhancement) and viscosifiers may be added to the alkaline agents to ensure
more favorable mobility ratios [Burk, 1987].
Compatibility issues also arise and a careful study of the interactions of the alkali (including co-
injected surfactants, cosolvents and brine) with the field soils and pore fluids is essential or pore clogging,
excessive alkali consumption, or insufficient pHs may lead to poor NAPL recovery. Alkali consumption and
rock dissolution may be excessive for long applications [Breit et al., 1981]. Clayey soils may cause
dispersion of the alkali and OH" fronts which will delay and affect NAPL recovery, as well a potentially
increasing alkali consumption and surfactant adsorption [Jensen et al., 1986].
78
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The condition of the aquifer upon competletion of akali soils washing is likely to be reduced with
a relatively high pH. Therefore, it may be desirable to oxygenate or neutralize the groundwater.
With the exception of borings, in-situ alkali soil washing is not likely to be intrusive; therefore, few
limitations exist with respect to interference from ground structures, overhead or buried utilities, and other
subsurface obstructions. Site grading is not a problem. The above ground hardware can be trailer-
mounted and constructed of readily available materials and standard unit operations equipment.
Cost and Availability-
The hardware required for in-situ alkali soil washing is readily available, but their are few experts
on this technology within the environmental community. Many full-scale applications have been completed
in the petroleum industry. A field scale application using compatible surfactants and alkalis has been
successfully completed.
Alkalis appear to be good candidates for aquifer remediation of DNAPLs, probably serving best as
complements to surfactants because most DNAPLs lack acidic components required for saponification
(natural surfactant production). Alkalis will have their biggest impact on DNAPL source areas.
Cost information from petroleum applications is not directly applicable. No information is available
at this time for environmental applications. Major issues requiring consideration are surfactant costs,
surfactant recycling, tankage requirements, and effect of field-scale soil heterogeneity on displacement
efficiency. Since the emulsions created can be very stable and the quantity of extracted fluids can be large,
management of produced fluids must be carefully considered.
3.5.2 Cosolvent Soil Washing
Theoretical Background-
In-situ cosolvent soil washing uses hydrophilic organic compounds (i.e., alcohols, ethers, ketones)
to enhance contaminant removal. The primary mechanisms are: displacement of the contactable NAPL
by a propagating cosolvent front; and solubility enhancement and interfacial tension reduction of NAPLs
which assists in their recovery. Although the exact displacement application differs, this process is in many
ways identical to enhanced oil recovery methods utilizing steam, surfactants, hot-water, or caustic floods
[Shah, 1981; Janssen-van Rosmalen and Hesselink, 1981].
Figure 3.5.2.1 shows the relationships between the resident fluid, r, and the displacing fluid, d, in
a column of porous media inclined at an angle 9 from the horizontal. The interfacial region possesses a
mixing length e, and is inclined at an angle a to the direction of flow. The importance of the mobility ratio,
M, on advancing cosolvent front stability is clearly shown in Figure 3.5.2.2 as pure isopropanol (IPA)
miscibly displaces naphtha (M=0.271) [Gatlin, 1959]. Conversely, the effluent history for an unstable
miscible displacement of IPA by naphtha (M=3.69) suggests that viscous fingering of naphtha into IPA
precludes effective sweep-out of IPA.
While miscible displacement is preferred so that interfacial forces are eliminated, pore level mixing
within interfacial regions will almost certainly lead to emulsification. This generally does not inhibit removal
because cosolvents may introduce a high degree of non-linearity in fluid phase properties such as viscosity,
density, and interfacial tension [Gatlin, 1959], The degree of non-linearity is ternary specific. For example,
Figure 3.5.2.3 shows the viscosities of equilibrated liquid pairs for the water-IPA-naphtha ternary as a
function of IPA content in each phase Gatlin (1959) found that while the miscible displacement of naphtha
by pure IPA yielded M=0.271, a water-lPA slug can be chosen with M=0.147 to induce a potentially
equivalent immiscible displacement.
79
-------
a = inclination angle
0 = dip angle
Figure 3.5.2.1 Schematic of the fluid-fluid displacement process [adapted from Boyd and Farley, 1992].
Figure 3.5.2.2
Column effluent histories of miscible displacements [adapted from Gatlin, 1959].
To estimate or optimize M for displacement, the selection of an appropriate cosolvent requires
specific knowledge of the phase behavior of water-cosolvent-NAPL ternary systems. The differences in
phase behavior can be pronounced for small differences in cosolvent structure. Figure 3.5.2.4, for example
shows the ternary phase diagrams for the IPA-water-Soltrol and tert-butyl alcohol (TBA)-water-Soltrol
ternary phase diagrams [Taber et al., 1960]. The miscible (single) liquid phase and binary (immiscible)
liquid phase regions are situated above and below the binodal curve, respectively. The small area under
the binodal curve indicates good cosolvent solvation ability. The end of the tie lines on the binodal curve
show the compositions of equilibrated liquid pairs. The location of the plait point and the tie lines illustrates
that for the entire range of possible compositions, IPA remains hydrophilic, while TEA becomes more
hydrophobic with increasing TBA content. Given the commensurate changes in bulk liquid properties that
result from partitioning, accumulation of TBA (u=2.9cp) in the Soltrol (u=1.5cp) phase may produce less
favorable mobility ratios than would be possible with IPA (|i=1 .9cp). Among the other factors that must be
considered when selecting a cosolvent include: low human health hazard; biodegradation potential;
80
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IO ?O SO «3 SO *O T
PER CENT IfW IN EACH
Figure 3.5.2.3 Viscosity enhancement of water and NAPL by isopropanol (IPA) in the H2O-IPA-Naphtha
ternary liquid system at 20°C [Gatlin, 1959].
Figure 3.5.2.4 Equilibrium phase diagrams for the IPA-Soltrol-Brine (2% CaCI2) and TBA(tert-butyl
alcohol)-Soltrol-Brine (2% CaCI2) systems showing binodal curves and inclination of tie
lines [Taberet al., 1960].
81
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T
e
s
3
WATER
ISOPROFYL
ALCOHOL
OIL
WATER
j Miscibl« ".nterfac*" Irmuscit*) .nf«rfoce
Figure 3.5.2.5 Idealized fluid-fluid displacement using a cosolvent (IPA) slug [adapted from Gatlin, 1959].
cosolvent losses to the DNAPL phase (excessive partitioning into DNAPL phase), strong ability to desorb
DNAPLs from soil organic matter; and low sorption to solids.
Field Implementation-
The displacement process entails the injection of the cosolvent as a finite, concentrated "slug" into
the porous medium which is then followed by a driving fluid. Injection of a dilute cosolvent may result in
solubility enhancement of the hydrocarbons, but not physical displacement of the NAPL. A slug is preferred
over continuous cosolvent injection for economic reasons. The driving fluid, water, pushes the cosolvent
slug through the porous medium and the cosolvent, in turn, miscibly or immiscibly displaces the resident
DNAPL and water in a plug flow manner. This process is schematically shown in Figure 3.5.2.5 for the
water-IPA-naphtha system.
The cosolvent slug is bounded by the water/cosolvent interface on the upgradient side and by the
cosolvent/resident fluid interface on the downgradient side. Stability at both interfaces must be considered
to avoid excessive viscous fingering and/or gravity override/underride which deteriorate slug integrity. The
size of the slug is then related to interfacial mixing at its upgradient and downgradient boundaries [Sievert
et al., 1958; Blackwell et al., 1959; Habermann, 1960] If the resident fluids are locally contacted by
essentially pure cosolvent, a local condition may be created which plots very close to or within the miscible
phase region of the ternary phase diagram. Under these conditions capillary forces arising from interfacial
tension can be mitigated or altogether eliminated.
While the effects of cosolvents on dissolved plume behavior have been studied [Barker et al.,
1992], no known field applications of in-situ cosolvent soil washing have been reported in the environmental
literature. Site selection is currently under way for a pilot study of in-situ cosolvent soil washing of
trichloroethylene (TCE) using ethanol which will be conducted in conjunction with the Robert S. Kerr
Environmental Research Laboratory (RSKERL) in Ada, Oklahoma [Wood, 1992].
In 1959, it was reported that 39 miscible displacement projects (some of which included cosolvents)
were being performed in the US for enhanced oil recovery from deep reservoirs [Habermann, 1960].
Displacements performed under these conditions are not directly transferable to the environmental industry
owing to differences in hydrogeologic context and purpose. Although the field implementation cannot be
adequately evaluated, certain aspects of its implementation are analogous to in-situ alkali and surfactant
soil washing (sections 3.5.1 and 3.5.3) and the CROW® process (section 3.7.1).
Level of Demonstration and Performance-
Previous work in the area of chemical displacements, floods, and soil washing was principally
82
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conducted using one-dimensional cores and LNAPLs, in which the LNAPL was displaced using a stable
flood configuration with respect to either density or viscosity, or both [Everett et al., 1950; Offeringa and
van der Poel, 1954; Sievert et al., 1958; Gatlin, 1959; Gatlin and Slobod, 1960; Morrow et al., 1985]. Gatlin
(1959) performed numerous alcohol floods of LNAPLs using methanol, IPA and tert-butyl alcohol (TBA) in
various combinations. Floods were conducted in 1.059-inch ID, 100 ft long, galvanized steel pipes packed
with Ottawa sand to porosities on the order of 35% and permeabilities of about 4 darcys. Cores were
oriented vertically, and frontal advances were on the order of 6 ft/hr. Figure 3.5.2.2 shows the removal of
naphtha by IPA for a stable slug size of 13.5 % pore volume. Gatlin (1959) examined the LNAPL removal
for a variety of slug sizes and injection strategies and found that comparable LNAPL removal and smaller
IPA requirements were possible when methanol was injected prior to the IPA slug, because methanol
preferentially displaced the residual water.
Horizontally oriented, two-dimensional, miscible displacement experiments have been conducted
by Habermann (1960). Radial displacement experiments using sands having permeabilities between 4.5
and 20 darcys were conducted for a variety of mobility ratios. The effect of adverse mobility ratios on
displacement efficiency is illustrated in Figure 3.5.2.6. For M<1, uniform propagation of the slug front
occurs until the influence of the extraction point on the flow field causes rapid breakthrough of the slug.
D
M « 71 5
• PRODUCING VrfLL
X 'N.'ECTION WfLi ,
Pv - POPE VOLUWF INjrCTFD
BT , BRLAKTHROJGH
Figure 3.5.2.6 Effect of mobility ratio on displacement fronts and injected pore volumes until breakthrough
using quarter of five-spot method [Habermann, 1960].
83
-------
I'X)
'O ?lt>
(T
Q
CD
Q
LLJ
O
O
O
tr
<
10 2O 3<-
PORE VOLUME INJECTED
Figure 3.5.2.7 Area contacted by fluid drive after breakthrough, quarter of five spot method [Habermann,
1960].
100
Sand contained 12
segments with 6 strata
4 6 8 I.O t 2
Solvent Injected pore volumes
14
Layer
PERMEABILITY PATTERN FOR SEGMENTED-STRATIFIED MODEL
Segment
H
I
1 9.5 33.0 10.2 9.5 41.5 56.0 16.2 41.5 33.0 9.5 33.0 41.i
2 56.0 16.2 41.5 56.0 16.2 33.0 6.5 16.2 9.5 56.0 6.5 16.2
3 6.5 41.5 33.0 6.5 33.0 41.5 9.5 6.5 56.0 33.0 41.5 9.5
4 41.5 6.J 56.0 33.0 6.5 9.5 41.5 56.0 6.5 16.2 56.0 33.0
5 16.2 56.0 9.5 16.2 56.0 6.5 33.0 9.5 16.2 41.5 9.5 6.5
6 33.0 9.5 6.5 41.5 9.5 16.2 56.0 33.0 41.5 45 14.2 56.0
Figure 3.5.2.8 Effect of mobility ratio on fluid recovery from segmented-stratified porous media model
[Blackwell et al., 1959]
84
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At breakthrough, primary recovery accounts for a sweep efficiency of approximately 70% on an area basis,
and appreciable secondary recovery can lead to total recoveries in excess of 90%, as shown in Figure
3.5.2.7.
Blackwell et al. (1959) investigated miscible displacement in lucite models containing
heterogeneously packed sands. In one series of two-stratum experiments, a 3/8x6x72 in Lucite model was
packed with two 3-in wide sand strata having permeabilities of 190 and 43 darcys, respectively. For an
M=1, 4.6 and 75, oil recovery from the model after 2.5 pore volumes of cosolvent injection amounted to
88, 72, and 60%, respectively. Cross flow of cosolvent into the more permeable stratum and bypassing
of the oil in the tight sand were observed. A second series of experiments was conducted in a 1/4x3x72
in Lucite model in which the sand was packed in a segmented-stratified configuration. Flow velocities of
41 to 72 ft/day were used. Sand "segments" (1/2 in wide and 6 in long) having permeabilities ranging from
6.5 to 56 darcys were arranged such that adjacent sand segment permeabilities were different, as shown
in Figure 3.5.2.8. Oil recovery is also shown for three mobility ratios, illustrating that oil recovery can be
appreciable in heterogeneous soils under controlled hydraulic conditions.
Boyd and Farley (1992) studied the 1-D displacement of TCE by IPA and examined flow stabilities
and TCE recovery in upflow, downflow and horizontal configurations using flow rates of 18 ft/day.
Displacement (H2O^IPA-4TCE) was conducted in 2.5 cm diameter, 80 cm long, glass chromatography
columns packed with glass beads. As shown by the effluent concentrations presented in Figure 3.5.2.9a,
TCE recovery in the downflow displacement occurred within two pore volumes, since viscous and gravity
forces at the IPA/TCE interface complemented one another. However, IPA recovery was much slower
since both M and buoyancy forces at the water/I PA interface were unfavorable in the direction of flow. The
effluent history for the horizontal displacement shown in Figure 3.5.2.9b, illustrates that displacement
efficiency is strongly influenced by the orientation of viscous and density forces.
Results of analogous experiments employing soils with a clay content of 16 wt% and using flow
rates of 9.5 ft/day were consistent with the glass bead experiments except at small slug sizes. Poorer TCE
removal at small slug sizes was attributed to possible TCE adsorption on clay surfaces [Boyd and Farley,
1992]. Considering that hydrophobic organic compound sorption to clay surfaces is low [Karickhoff, 1981],
a more likely explanation is that pore clogging caused by migration of fines led to TCE entrapment.
Sequential permeation of hydrocarbon contaminated water-wet clays by ethanol and water (in that order)
resulted in the successful leaching of benzene by ethanol, and subsequently a two order of magnitude
reduction in hydraulic conductivity after benzene removal [Fernandez and Quigley, 1985]. Similar results
were obtained for other cosolvent-hydrocarbon pairs
Wood et al. (1992) conducted elution experiments of 2.3+0.3 ppm poly chlorinated biphenyl (PCB)
contaminated soils using ethanol-water solutions in 2.54 cm diameter columns of 5 cm length The soils
possessed 2 g/kg organic matter content. PCB displacement efficiencies of 85.1, 96.1, and 98.3% were
achieved for binary ethanol-water mixtures containing 47 5, 57, and 76% ethanol, respectively.
Applicability/Limitations-
Theoretically, this technology can be applied to almost any DNAPL providing a cosolvent can be
selected such that the phase behavior and the resultant changes in bulk liquid properties are favorable.
Solubility enhancement and desorption of hydrophobic hydrocarbons from soils by cosolvents are widely
acknowledged [S0renson and Arlt, 1980; Fu and Luthy, 1986a,b; Prausnitz et al., 1986; Woodburn et al.,
1986; Groves, 1988; Zachara et al., 1988; Rao et al., 1990; Broholm et al., 1992; Kan et al., 1992; Wood
et al., 1992].
Interfacial instability (M>1) caused by large viscosity differences (5-200 cp) of reservoir oils was
among the primary reasons for deterioration of miscible slugs observed in petroleum recovery. Any
85
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combination of compounds such as alkaline agents, surfactants, polymers (density enhancement) and
viscosifiers may be added to the injectates to ensure a more favorable mobility ratio between the cosolvent
and the DNAPL or to enhance the solubility of DNAPL components. Finally, displacement efficiency will
be greatly affected by soil heterogeneities and cosolvent buoyancy.
Experiment I
JEL
10
—r~
10
—I—
12
—T~
14
—r~
16
18
Cumulative Injection (PV)
Enpenment I 100% IPA Injection Ftee ptoduct 'ecovaiy, IPA tlooc,
follow-up waterflood were conducted in the downtlow direction
-i 1 1 1 r
3456
Cumulative Injection (PV)
Experiment IV. 100% IPA Injection. Free product re:
follow-up waterflood were conducted in the horizontal a
Figure 3.5.2.9 Comparison of effluent histories for (a) vertical and (b) horizontal H2O->IPA^TCE miscible
displacements in soil cores. IPA^TCE mobility ratios stable for both displacements
whereas H2O-^IPA are not. The IPA^TCE interface in (b) is unstable due to gravity effect,
while the H2O-»IPA interfaces in both (a) and (b) are unstable [Boyd and Farley, 1992].
86
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With the exception of borings, in-situ cosolvent soil washing is not likely to be intrusive; therefore,
there are few limitations due to interference from ground structures, overhead or buried utilities, and other
subsurface obstructions. Site grading is not a problem. The above ground hardware can be trailer-
mounted and constructed of readily available materials and standard unit operations equipment.
To prevent chemical interactions between well construction materials and pore fluids, stainless steel
well construction may be required. Chemical interactions occurring between concentrated organic pore
fluids and clayey soils may result in clay desiccation and cracking [Fernandez and Quigley, 1985; Mitchell
and Madsen, 1987]. Co-metabolic assimilation of the remaining contamination by indigenous
microorganisms may potentially occur after the main DNAPL displacement.
Cost and Availability-
Cosolvents have been employed on the full-scale in petroleum engineering applications. At present
there is limited experience with this technology within the environmental community. Environmental
applications are on the lab scale at this time, and a pilot study has been proposed.
Cosolvents appear to hold promise for cleanup of NAPL source areas, but many issues are
unresolved at this time. However, because of density considerations, it is likely that cosolvents will
probably be more effective for LNAPL cleanup rather than DNAPLs.
No cost information is available at this time. Major issues requiring consideration are cosolvent
costs, cosolvent recycle, and effect of field scale soil heterogeneity on displacement efficiency.
3.5.3 Surfactant Soil Washing
Theoretical Background-
In-situ surfactant soil washing employs cationic, anionic or non-ionic surface active compounds to
enhance contaminant removal. The primary mechanisms are: micellular solubilization of sparingly soluble
hydrocarbons into the aqueous phase; coacervation of NAPL into a middle-phase emulsion; enhanced
NAPL ganglia mobilization as a result of interfacial tension reduction; coalescence of individual NAPL
ganglia into a NAPL bank; and the displacement of the NAPL bank to a recovery well by viscous forces.
Although the exact displacement application differs, this process is in many ways identical to enhanced oil
recovery methods utilizing steam, cosolvents, hot-water, or caustic floods [Shah, 1981; Janssen-van
Rosmalen and Hesselink, 1981]. In fact, cosolvent and alkalis have often been used to complement
surfactants [Reed and Healy, 19" " Shah, 1981; Clark et al., 1988; Manji and Stasiuk, 1988]
Two regions of low interfacial tensions may occur in surfactant systems: at low surfactant
concentrations (0.1-0.2 wt% and often less) which corresponds to a two-phase system [Rosen, 1978; Shah,
1981; Vignon and Rubin, 1989]; and at high surfactant concentrations (2-10 wt% and often up to 30-40
wt%), which corresponds to a three-phase system containing a middle-phase microemulsion that is in
equilibrium with the two bulk phases [Reed and Healy, 1977; Shah, 1981; Chan and Shah, 1981; Radke,
1993].
Micellization describes the process in which surfactant monomers form spheroid or lamellar
structures possessing organic psuedophase interiors. At low surfactant concentrations, low interfacial
tensions and pronounced solubility enhancement normally coincide with the onset of micellization in the
aqueous phase, as shown in Figure 3.5.3.1 [Rosen, 1978; Shah, 1981; Chan and Shah, 1981]; under these
conditions, the surfactant is described as being at its "apparent" critical micelle concentration. This will
differ from the critical micelle concentration measured in pure water. Factors such as the aqueous phase
salinity, hydrocarbon chain length (degree of hydrophobicity), and surfactant type (hydrophile-lipophile
balance, HLB) and concentration affect the exact value of the apparent critical micelle concentration and
87
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Critical
Concentration
Detergency
/mass ol dispersed and dissolved contaminants
*
volume of aqueous solution of surfactant
Solubitization
/ _ mass of dissolved contaminant _ \
* volume ol aqueous solution of surfactant '
\
'
Surface Tension
0 01 0.2 03040506070809
Sodium Lauryl Solfale (%)
Figure 3.5.3.1 Physical property changes of aqueous solutions of sodium lauryl sulfate in vicinity of critical
micelle concentration [Rosen. 1978]
PETROLEUM SULFONATES
(a)
OIL
BRINE
s, r
-!r-j,+
20
NoCI CONC "X,
2 5
( | | V | (
o u CT~D
sc
3-0
(b)
Figure 3.5.3.2 Relationships between salt concentration, oil chain length, surfactant concentration on (a)
interfacial tension, and (b) surfactant partitioning and micelle formation in petroleum
sulfonate systems [Chan and Shah, 1981],
88
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m
The transition I •
Parameter Increasing
- u occurs by:
1 Increasing Salinity
2 Decreasing oil chain length
3 Increasing alcohol concentration (C C Cr )
4 b ' b
4 Decreasing temperature
5 Increasing total surfactant concentration
6 Increasing brine/oil ratio
7 Increasing surfactant solution/o'1 i.n"
8 Increasing molecular weight of sutladant
Figure 3.5.3.3 Schematic illustrating the l^m^u^ transition and the factors influencing its determination
in surfactant/oil/brine/alcohol systems [Shah, 1981].
extent of surfactant partitioning [Chan and Shah, 1981]. For example, Figure 3.5.3.2 illustrates the
generalized effect of salinity on surfactant partitioning and phase behavior. At low salinity, the surfactant
and micelles prevail in the aqueous phase, but as the salinity increases, "salting out" of the surfactant
results in its partitioning into the NAPL.
Concentrated (>10 wt%) surfactant systems are characterized by complex phase behavior and
middle-phase microemulsions whose formation is precluded at lower surfactant concentrations simply from
mass considerations. Since water is more dense than oil, an emulsion resulting from surfactant
accumulation in the denser aqueous phase is referred to as a lower (I) phase emulsion, using the
petroleum engineering convention [Reed and Healy, 1977]. To avoid confusion arising from density
considerations, it can also be referred to as Winsor type I emulsion [Winsor, 1954]. A middle (m) phase
or Winsor type III microemulsion forms when the surfactant is concentrated at the water-NAPL interfacial
region. Surfactant accumulation in the NAPL phase is referred to as an upper (u) phase or Winsor type
II emulsion. These relationships are shown in Figure 3.5.3.3.
Much like Figure 3.5.3.2, Figure 3.5.3.3 shows that a continuum of phase behavior, and l-m-u
transitions can be achieved by changing any of the following variables: aqueous salinity, surfactant
concentration, hydrocarbon chain length, molecular weight of surfactant, cosolvent structure and chain
length, surfactant/oil ratios, and surfactant/brine ratios [Reed and Healy, 1977; Salager et al., 1979; Shah,
1981; Graciaa et al., 1982]. Middle-phase or Winsor type III microemulsions are the most favorable
emulsions for the displacement process because this emulsion: has ultra-low interfacial tensions (~0.01-
0.001 dyne/cm) making displacement at realistic hydraulic gradients possible; contains large quantities of
NAPL which enhances primary recovery; and is viscous, thus promoting residual NAPL mobilization and
formation of a NAPL bank ahead of the microemulsion which enhances primary recovery.
Field Implementation--
Two strategies are considered, depending on the surfactant concentration used and whether simple
89
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FLOW
S0!
WATER i POLYMER [SURFAC- PREFLUSH I RESIDENT
i I TANT I WATER
sof
1 ifWULTIMATE RESIDUAL^
Figure 3.5.3.4 Schematic illustrating fluid bank formation as a function of saturation and distance in a
surfactant/polymer flood [Reed and Healy, 1977],
solubility enhancement or a more displacement-like process is desired. Solubility enhancement entails
continuous surfactant injection for a specified number of pore volumes in order to leach hydrocarbons out
of the porous media. Low surfactant concentrations (<1 wt%) are usually employed for this purpose.
The displacement process is analogous to the cosolvent slug displacement process (section 3.5.1).
Liquid-liquid displacement requires the injection of highly concentrated surfactant (10-40 wt%) as a finite
slug. At the down-gradient interface of the slug, a concentrated and viscous middle-phase microemulsion
and a NAPL bank are created. Successful mobility control of the middle-phase microemulsion and NAPL
bank requires that the surfactant slug be driven by a polymer slug(s) of greater viscosity [Reed and Healy,
1977; Shah, 1981; Manji and Stasiuk, 1988]. This process is depicted in Figure 3.5.3.4.
Conventional injection and extraction well construction can be used for in-situ surfactant soil
washing. Horizontal and vertical wells have been used (see below), and well placement strategy depends
on the nature and extent of contamination, soil heterogeneities, and anticipated flow behavior once washing
commences.
Level of Demonstration and Performance-
Solubility enhancement and desorption of hydrophobic hydrocarbons from soils by surfactants are
widely acknowledged and documented [Winsor, 1954; Saito and Shinoda, 1967; Reed and Healy, 1977;
Akstinat, 1981; Shah, 1981; Ellis et al., 1985, Kile and Chiou, 1989; Edwards et al., 1990, Aronstein et al.,
1991; Rixey et al., 1991; Kan et al., 1992]. Numerous laboratory batch and column studies have
demonstrated that NAPL recovery from porous media is greatly enhanced by surfactants [Thornton, 1980;
Hesselink and Faber, 1981; Ellis et al., 1985; Vignon and Rubin, 1989; Abdul et al., 1990a,b; Abdul and
Gibson, 1991; Kan et al., 1992; Soerens et al., 1992].
Several field applications of in-situ soil surfactant washing of DNAPLs have been conducted. An
in-situ surfactant soil washing pilot study was conducted in a fire training pit at the Volk National Guard
Base (Wl). Subsurface soils had been contaminated with chlorinated organics (including DCM, chloroform,
TCA, TCE) up to 3.5 ppm, and ground water contained total organics in excess of 300 ppm [Nash, 1988].
The sandy soils had a cation exchange capacity of 0.5 meq/kg, and organic matter content ranged from
0.037 to 1.5 wt% [Nash, 1988]. Chawla et al. (1989) cite data from two reports indicating that the soil had
5 to 15 wt% fines, hydraulic conductivities of 5.2x10"4 to 1.7x10"2 cm/s, and a depth to water table of 12
ft Pits with dimensions of 1x1x1 or 2x2x1 ft were dug around each borehole (10 total) to aid in surfactant
delivery. Laboratory testing indicated that a 1.5 wt% blend (50/50) of Adsee 799 (Witco) and Hyonic NP-90
(Diamond Shamrock) was capable of 74 to 94% NAPL recoveries within 12 pore volumes [USEPA, 1991c].
90
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Based on these results, the aforementioned blend and six other synthetic and natural surfactant blends (3
each) were administered into the pits four times a day at an equivalent rate of 1.9 gal/ft2 per day for a
period of 4 to 6 days [Chawla et al., 1989; USEPA, 1991c]. This application rate corresponded to
approximately 14 pore volumes of fluid. Three test holes clogged by the third day. Chemical analysis of
soil samples taken from 2 to 4 and 12 to 14 inches below the bottom of the test holes revealed that no
statistically significant contaminant removal had occurred [USEPA, 1991 c]. Several mechanisms have been
proposed: high contaminant sorption to soil organic matter [Chawla et al., 1989]; surfactant bypassed
contaminated zone [Nash, 1986]; migration of fines caused pore clogging [USEPA, 1991c]; formation of
micelles caused pore clogging [USEPA, 1991c]; and biological activity caused pore clogging [Vignon and
Rubin, 1989].
In 1988, a field study of in-situ surfactant soil was conducted at a former wood treating facility in
Laramie, Wyoming [Sale et al., 1989], Sheetpiling was used to create a 27x27 ft test cell in alluvial soils
to a depth of 12 ft where bedrock (shale) was encountered. The cell had an estimated pore volume (PV)
of 5,000 to 5,500 gallons. Wood treating waste contaminated soils (SG=1.04, u=54 cp) saturated the lower
3 ft of the alluvium. Soil contamination within the test cell was estimated to be approximately 93,000 ppm.
A 4-inch surfactant delivery injection line and a vertically nested dual drain line were placed in parallel, and
spaced 15 ft apart. The dual drain line had a 3 ft vertical spacing and was designed to simultaneously
extract water and denser oils at different flow rates from the upper and lower drains, respectively [Sale and
Pointek, 1988; Sale et al., 1988, 1989]. The test cell is schematically depicted in Figure 3.5.3.5.
Two surfactant/alkali/polymer blends were selected. Alkalis and polymers were selected because
of wetting, surfactant sorption, pore clogging and mobility considerations. Blend 1 (10,000 gal) was used
as a prewash to increase the amount of reusable oil and Blend 2 (20,000 gal) to attain lower cleanup
levels. The blend compositions were: (Blend 1) 1.0 wt% sodium dodecyl benzene sulfonate (Polystep A-
7®), 0.72 wt% NaHCO3, 0.1 wt% Na2CO3, and 1,050 mg/l Xanthum Gum Biopolymer; and (Blend 2) 1.4
wt% ethoxylated nonphenol (Makon-10®, Stepan Chem. Co.), 0.65 wt% NaHC03, 0.825 wt% Na2CO3, and
1,050 mg/l Xanthum Gum Biopolymer. The primary water-flood (140,000 gal) and surfactant flood (30,000
gal) recovered 1,600 and 260 gal of oil, respectively. Residual oil concentrations of 15,500 and 5,100 ppm
(soil) were estimated from an analysis of soil cores taken after water flooding and surfactant flooding,
respectively. This constitutes an overall reduction of approximately 95% by weight. Water quality data
indicated that approximately 99 wt% of the surfactants were recovered.
In 1990-91, a surfactant soil washing pilot study was conducted at the Canadian Forces Base
Borden in Ontario, Canada [Fountain et al., 1990,1991; Wunderlich et al., 1992; Fountain, 1992b, in press].
The Borden sands have a hydraulic conductivity in the order of 1x10"4 m/s. The organic carbon content
and cation exchange capacity of these soils are very low. The Borden sands are extensively described
elsewhere [Sudicky et al., 1983; Sudicky, 1986], A sheetpile test cell having the dimensions of 3x3 m was
socketed into the underlying clay aquitard located at a depth of 4 meters. A secondary sheet pile isolation
system was also installed. Five surfactant and five injection wells were installed at opposite sides of the
cell, as shown in Figure 3.5.3 6. Multilevel monitoring wells each with six sampling points were placed 0.3
m from each injection and extraction well [Fountain et al., 1990]. A total of 231 liters of PCEwere released
via a pipe into the center of the test cell, and undisturbed migration of PCE was permitted for two months.
Pumping of pure PCE product which accumulated in the wells during this period recovered 48 liters of PCE.
As part of the study, PCE infiltration was studied. Excavation of near surface soils to a depth of 1 m
recovered 52 liters of PCE. The cell was subsequently backfilled with bentonite [Fountain, in press]. A
water flood was then initiated which recovered an additional 12 liters of PCE.
After extensive surfactant testing [Fountain, et al., 1991], a 2 wt% solution of a 50/50 blend of
nonylphenol ethoxylate (Alkasurf NP-10, Alkaril Co.) and a phosphate ester of a alkylphenol ethoxylate
(Rexophos 25/97, Hart Chem.) were finally selected for the pilot test. This blend lowered the water-PCE
91
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Reagent Delivery
System
.>ii!ivcryOr
-------
interfacial tension to 3.2 dynes/cm. The field test entailed the injection of approximately 14.4 pore volumes
(1 PV=2400 gals) of surfactant over a 4 month period [Wunderlich et al., 1992]. Approximately 62 liters
(27%) of PCE were recovered during the surfactant flood. The PCE concentrations in the effluent
exceeded 4,000 ppm at its maximum, and dissolved PCE concentrations on the order of 12,000 ppm PCE
were observed in the monitoring wells during flooding. Analysis of groundwater samples indicates that PCE
concentrations have been reduced below 1 ppm over most of the cell [Fountain, in press]. Soil samples
taken during flooding revealed that although some PCE had perched on fine soil lenses, this PCE was
successfully removed. Other soil samples taken after surfactant flooding revealed that less than 10 liters
of PCE remained within the test cell [Fountain, in press], or a total test recovery of 184 liters PCE (80%).
Approximately 35-50 liters of PCE remain unaccounted for, and possible explanations for it include:
volatilization of PCE; trapping of PCE near the edges of the test cell; and migration of PCE into fractures
within the aquitard caused by the sheetpile driving [Fountain, in press]. While laboratory studies had
indicated that rapid downward vertical migration of DNAPLs was possible when water-DNAPL interfacial
tensions were lowered below 5 dynes/cm [Fountain et al., 1991], evidence indicates that this was not a
problem during the pilot test [Fountain, in press].
Poly-chlorinated biphenyl (PCB) contaminated soils were the focus of a recent soil washing test
in vadose soils [Abdul et al., 1992]. The test was conducted in a sandy fill material having a hydraulic
conductivity on the order of 1x10"3 cm/s which extended to depths of 13 to 15 ft. The initial water table
was at a depth of 4 ft. Soil cores revealed PCB and carrier oil concentrations up to 6,223 and 67,000 ppm,
respectively. Estimates based on soil analyses indicated that approximately 15.3 kg of PCBs and 157.1
kg of carrier oil were present within the test cell. The 10 ft diameter test cell extended to a depth of 5 ft.
Surface application of the surfactants required construction of a small berm. A 4-inch schedule 80 PVC
extraction well was constructed in the center of the test cell with a screened interval at depths between 5.75
and 13.25 ft. Four 2-inch schedule 60 PVC monitoring wells were installed along the perimeter of the test
cell. To measure the saturation response of the vadose soils, two sets of tensiometers and moisture
measuring devices were installed within the test cell. The surfactant, a 0.75 wt% nonionic ethoxylated
alcohol (Witconol SN70, Witco), was applied to the test cell at an average daily rate of 77 gal/day for a
period of 70 days (5,375 gal total). Ground water containing leachates was extracted at an average daily
rate of 157 gal/day (10,981 gal total). During the testing period, a total of 1.6 kg PCBs (10.5 wt%) and
16.9 kg (10.7 wt%) carrier oil were recovered.
A surfactant flood is currently being conducted at an industrial facility in Corpus Christi, Texas,
where ground water has been contaminated with carbon tetrachloride [Fountain, 1992a, in press;
Wunderlich et al., 1992]. The contaminated aquifer at the site is 4 m thick and is underlain by a thick clay
deposit. The aquifer has an organic matter content of 0.025 wt%. Analysis of cores revealed carbon
tetrachloride concentrations above 1100 ppm in the upper meter of the aquifer. Since no containment
system was installed, potential downward migration of carbon tetrachloride was of obvious concern.
Therefore, Witconol 2722 (Witco), a polysorbate, which was not the best solubilizer of carbon tetrachloride,
was selected since it produced a water-carbon tetrachloride interfacial tension of about 10 dynes/cm. This
value of interfacial tension was sufficient to prevent vertical migration of the carbon tetrachloride at the site.
Thus, high solubilization was traded for added protection from potential downward migration of carbon
tetrachloride free product. No results are available at this time.
Applicability/Limitations-
Theoretically, in-situ surfactant soil washing can be applied to almost any immiscible hydrocarbon
providing a surfactant can be selected such that the phase behavior and the resultant changes in bulk liquid
properties are favorable. Water-DNAPL interfacial tension reductions below 5 dynes/cm have resulted in
rapid downward vertical migration of DNAPLs in laboratory studies [Fountain et al., 1991] DNAPL recovery
by surfactants will be greatly affected by soil heterogeneities.
93
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Surfactant flooding was originally developed in the petroleum industry. Deep reservoir conditions
such as high NAPL saturations, high NAPL viscosity, salinity, geologic confinement, scale of application,
and the attainment of ultra-low interfacial tensions favor the creation of concentrated and viscous middle-
phase microemulsions and banks to promote good sweep efficiency. An approach of this kind may be
favored at sites like the former wood treating facility in Laramie, Wyoming, at which large-scale DNAPL
contamination exists. However, for most other DNAPL-contaminated sites, micellular solubilization of
DNAPLs using lower surfactant concentrations (<1 wt%) which provide moderate interfacial tensions seems
most fitting in light of environmental sensitivity and the relatively small scale of contamination. Ongoing
research will provide further insights into the applicability of ultra-low interfacial tensions to environmental
remediation.
Surfactants themselves often reduce the viscosity of water, which often promotes unfavorable
mobility ratios. If a more displacement-like operation is preferred over solubilization, any combination of
compounds such as alkaline agents, cosolvents, polymers (density enhancement) and viscosifiers may be
added to the surfactant to ensure a more favorable mobility ratio. However, compatibility issues arise and
a careful study of the interactions of the surfactant blend (including any cosolvents and alkalis) with the field
soils and pore fluids is essential or pore clogging, surfactant precipitation and sorption may result. Anionic
and non-ionic surfactants are generally not prone to sorb to aquifer solids (organic matter interactions
excluded), whereas cationic surfactants are; and they have been intentionally used to lower aquifer
permeabilities [Brown et al., 1992; Burris and Antworth, 1992; Westall et al., 1992]. Many surfactants are
biodegradable and non-toxic, and the anaerobic degradation of surfactants has been observed to be
extensive on the time scale of months [Fountain, 1992a]. Therefore, the condition of the aquifer after
surfactant washing should be favorable for continued biodegradation of any remaining hydrocarbons.
With the exception of borings, in-situ surfactant soil washing is not likely to be intrusive; therefore,
there are few limitations due to interference from ground structures, overhead or buried utilities, and other
subsurface obstructions. Site grading is not a problem. The above ground hardware can be trailer-
mounted and constructed of readily available materials and standard unit operations equipment.
Cost and Availability-
The hardware and surfactants required for in-situ surfactant soil washing are readily available.
Expertise in this area is increasing rapidly in the environmental community. Surfactant soil washing has
been demonstrated on the full scale in petroleum applications, and environmental field applications of
surfactant soil washing have been completed. Several more field applications are planned.
Surfactants are good candidates for aquifer remediation of DNAPLs when used in conjunction with
alkalis, cosolvents and viscosifiers. Surfactants will have their largest impact on DNAPL source areas.
Application within fine-grained soils is not likely to be successful.
No information on costs is available at this time. Major issues requiring consideration are surfactant
costs, surfactant recyclability, tankage requirements, and the effect of field scale soil heterogeneity on
displacement efficiency. Since the emulsions created can be very stable and quantity of extracted fluids
can be large, management of produced fluids can be problematic.
3.5.4 Water Flooding and Ground-water Extraction
Theoretical Background-
Water flooding and/or ground-water extraction, often referred to as pump-and-treat, is a process
in which ground-water injection or reinjection and pumping is used for contaminant removal. The primary
mechanisms are: increased recovery of the DNAPL as it responds to pumping stress in the aqueous phase
94
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[Villaume, 1985; Wisniewski et al., 1985; Sale et al., 1989]; dissolution of the DNAPL components into the
aqueous phase; and containment of the dissolved plumes. The first of these recovery mechanisms has
been used alone and in conjunction with alkali and surfactant soil washing [Sale et al., 1989]. The
response of a DNAPL under pumping stress is discussed here owing to its applicability as a remedial step
at sites having widespread DNAPL contamination.
At extensively contaminated sites, that is, where sufficient separate phase is present, DNAPL flow
to recovery wells may be induced simply by gravity or by the application of ground-water (hydraulic)
gradients. However, the capillary pressure phenomena and considerations described in Sections 2, and
3.5.1 through 3.5.3 still apply. A related mechanism is the upconing of the interface between a dense fluid
phase (i.e., saltwater, DNAPL) and ground water in response to pumping of the overlying ground water.
Upconing of a dense fluid phase is initiated by total head reduction in each fluid phase in the vicinity of an
extraction well [Muskat, 1982], as illustrated in Figure 3.5.4.1. When the fluids are miscible, as in the case
of fresh and saltwater, the behavior of the interface can be described by a simple hydrostatic balance, i.e.,
the Ghyben-Herzberg approximation [Bear, 1972],
However, interfacial and viscous effects between water and DNAPLs may preclude the use of such
a simplified approach, although it is often used as a first order approximation. DNAPL recovery is seen
to be a function of the thickness of the DNAPL pool, capillary pressure, and the buoyant density and
viscosity separate phase [McWhorter et al., 1992]. Based on theoretical calculations [McWhorter and
Sunada, 1990], it is anticipated that DNAPL recovery can be maximized by utilizing small pumping rates
[McWhorter et al., 1992].
Field Implementation--
Depending on the properties of the NAPL, conventional injection and extraction well construction
equipment can be used for in-situ water flushing; and both horizontal and vertical well configurations have
been used [Villaume et al., 1983; Villaume, 1985; Sale et al., 1988]. Well placement strategy depends on
the nature and extent of contamination, soil heterogeneities, and anticipated flow behavior once water
pumping commences.
For brevity, only the horizontal dual "drain line" approach is described here [Sale et al., 1988,
1989]. This vertically nested well configuration is shown in Figure 3.5.4.2a By installing DNAPL recovery
wells as close to the bottom of a DNAPL pool as possible; i.e., near stratigraphic depressions in underlying
aquitards or bedrock, the DNAPL elevation head is maximized for recovery [Villaume, 1985]. Water
recovery wells are nested directly above the DNAPL recovery wells at an elevation which accommodates
upconing of the DNAPL interface.
Although pumping strategies are tailored to site specific needs, water recovery wells are usually
pumped continuously. Once DNAPL is detected in the water recovery well, DNAPL recovery is initiated
as illustrated in Figure 3.5.4.2b [Villaume et al , 1983; Sale et a!., 1988] DNAPL pumping is either
intermittent, or continuous, but at a lesser rate than the water recovery well. Overpumping of the DNAPL
well can decrease DNAPL recovery because DNAPL influx to the well is "pinched-off" when the cone of
depression in the DNAPL phase is large and water influx results [Sale et al., 1988; McWhorter et al., 1992;
USEPA, 1992a].
Level of Demonstration and Performance--
There are at least two documented instances in which the response of a DNAPL to pumping
stresses in overlying groun-dwater has been used advantageously as an enhanced DNAPL recovery
technique [Villaume, 1985; Sale et al., 1988].
95
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Figure 3.5.4.1 Schematic illustrating the upconing phenomena of a dense fluid phase to pumping stress
in the overlying fluid phase [Wisniewski et al., 1985].
GROUND SURFACE
GROUND WATERSUR FACE
OIL SURFACE
WATER TABLE DEPRESSION
DRAINLINE (WTDD)
OIL I." '
l .' •' OIL RECOVERY DRAINLINE (ORD)
\ y
BEDROCK
GROUND SURFACE
OIL DISTRIBUTION
THE OIL, BEING SLIGHTLY DENSER THAN
GROUNDWATER, HAS ACCUMULATED AT
THE BASE OF THE ALLUVIUM
OIL RECOVERY WITH A
SINGLE RECOVERY LINE
GROUNDWATER
SURFACE
BEDROCK
GROUND SURFACE
PUMPING ONLY THE ORD RESULTS IN
THE MORE MOBILE GROUNDWATER
TRUNCATING THE FLOW OF THE MORE
VISCOUS OIL TO THE RECOVERY LINE
OIL RECOVERY WITH A
GROUNDWATER
->
OIL SURFACE ^3H
~~~^WTDD^~—
-^- ^~^j -•
~~ ORD^~~^-
O '
SURFACEjz__
V
BEDROCK
GROUND SURFACE
GROUNDWATER
SURFACE_s_
BEDROCK
DUAL DRAINLINE TECHNIQUE
• DRAWDOWN OF THE OVERLYfNG WATLR
TABLE BY PUMPING THE WTDD
RESULTS IN MOUNDING OF THE
OIL BENEATH THE WTDD
• PUMPING FROM BOTH THE WTDD AND
ORD INDUCES OIL FLOW TO THE ORD
• SEPARATE PRODUCTION OF OIL AND
GROUNDWATER REDUCES ABOVEGROUND
SEPARATION REQUIREMENTS
• A FLOW PATH OF MAXIMUM
FORMATION PERMEABILITY TO OIL
IS ESTABLISHED AT THE BASE
OF THE ALLUVIUM
Figure 3.5.4.2 Schematic of dual drain line system for pumping of both light and dense fluid phase to
enhance the recovery of the underlying, denser phase [Sale et al., 1988].
96
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In the early 1980's, enhanced recovery of DNAPLs by water pumping was undertaken at a former
coal gasification facility in East Stroudsburg, Pennsylvania [Villaume et al., 1983; Villaume, 1985]. The
hydraulic conductivity of the near surface aquifer (clean sands and gravels) was approximately 5.3x10"3
cm/s. The aquifer, with a thickness of approximately 5 to 40 ft, is underlain by a silty sand stratum. This
stratum behaved as an aquitard because its penetration by the pure phase coal tar was prevented by large
capillary pressures. Approximate depths to ground water and natural hydraulic gradient were 10 ft and
0.015 ft/ft, respectively [Villaume et al., 1983]. In one area of the site, approximately 35,000 gal of coal tar
(SG=1.02, u.=19 cp) was pooled in a stratigraphic depression on the underlying silty sand formation. The
estimated thickness of the coal tar pool was 3-10 ft. The pure phase coal tar contamination was essentially
confined to the sand and gravel aquifer because of capillary pressures.
One 4-inch and four 6-inch PVC wells were installed within a 30-inch gravel packed borehole to
a depth of approximately 40 ft in the stratigraphic depression [Villaume et al., 1983]. Initially, only coal tar
was pumped at a slow rate and approximately 100 gal/d of coal tar was recovered, but this recovery
decreased rapidly owing to oil depletion in the vicinity of the well. At an elevation situated considerably
above the static DNAPL surface, a packer was then installed in the central 4-inch well. Water removal then
commenced at a slow rate, and extracted water was re-injected 65 ft upgradient. Using this approach,
more than 8,000 gallons of coal tar having a water content of less than 1 wt% were recovered in nine
months of operation.
In 1988, water-flooding was conducted at a former wood treating facility in Laramie, Wyoming [Sale
et al., 1988] as a precursor to in-situ surfactant/alkali soil washing (see section 3.5.3). Sheetpiling was
used to create a 27x27 ft test cell in alluvial soils to a depth of 12 ft where bedrock (shale) was
encountered. The cell had an estimated pore volume (PV) of 5,000 to 5,500 gallons. Spent wood treating
oils (SG=1.04, |o=54 cp) saturated the lower 3 ft of the alluvium. Contaminant concentrations in the soil
within the test cell were estimated to be approximately 93,000 ppm. A 4-inch injection line and vertically
nested dual drain line were placed in parallel, spaced 15 ft apart. The test cell is schematically depicted
in Figure 3.5.3.5. The water-flood (140,000 gal) recovered 1,600 gal of oil. Residual oil concentrations of
15,500 ppm were estimated from soil core analyses. This constitutes a reduction of approximately 83 wt%.
Applicability/Limitations-
Waterflooding can be applied to enhance recovery from DNAPL pools; however, depending on the
DNAPL, significant problems may be encountered as a result of chemical attack on downhole equipment
[Villaume et al., 1983; Villaume 1985]. Thus far, only relatively light DNAPLs (SG<1.1) are known to have
been recovered by this technique. Also, this treatment is suitable only as a precursor to other in-situ
cleanup measures, since the residual concentrations of DNAPL will remain significant (approx 5-20 wt%)
[Wilson and Conrad, 1984].
Simultaneous pumping of fluids from the water and DNAPL horizons can minimize ex-situ liquid
separations requirements and increase DNAPL recycle [Villaume et al, 1983]. Since a large volume of
fluids may be produced, water re-injection is often used
With the exception of borings, in-situ soil washing is not intrusive and, therefore, there are few
limitations due to interference from ground structures, overhead or buried utilities, and other subsurface
obstructions. Site grading is not a problem. The above ground hardware can be trailer-mounted and
constructed of readily available materials and standard unit operations equipment.
Cost and Availability-
There are at least two published accounts which emphasize the use of upconing and pumping
response strategies to enhance DNAPL recovery. Recovery of floating LNAPL product occurs by the same
mechanism. The hardware required for in-situ waterflooding is readily available, and the requisite expertise
97
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to design such systems exists within the environmental community.
Total operating and maintenance costs for the coal tar recovery were on the order of $1,000/month,
including repairs associated with chemical attack [Villaume et al., 1983]. Recovered coal tar (17,500 Btu/lb)
was sold as a fuel supplement. Installation costs should be comparable to pump-and-treat.
98
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3.6 AIR STRIPPING
Introduction--
In-situ air stripping processes generally rely on the air circulation through the subsurface to remove
volatile DNAPLs from the subsurface. The applications considered herein, in-situ air sparging and vacuum
extraction (section 3.6.1) and vacuum vaporizer wells (section 3.6.2), differ from conventional air stripping
and soil vapor extraction in the vadose zone in that they operate in both the saturated and unsaturated
zones.
Air sparging and vacuum extraction entail the injection of clean air directly into the saturated zone.
Stripping occurs within the porous medium and volatilized contaminants are recovered by vapor extraction
wells nested in the vadose zone. Vacuum vaporizer wells, or UVBs, create water recirculation cells within
the porous media. Stripping is performed "in-well" and contaminant laden vapors are collected at the top
of the well, while water is recycled back into the aquifer. UVBs can also simultaneously recover soil vapors
from the vadose zone.
Both processes are diffusion limited, and apply to the recovery of volatile and semi-volatile DNAPLs
only. Sparging may also result in uncontrolled migration of DNAPL out of the treatment zone. Enhanced
biostimulation may be a beneficial by-product of both processes. Both technologies are commercially
available and used.
3.6.1 Air Sparging and Vacuum Extraction
Theoretical Background--
Air sparging and vacuum extraction (ASP/VE) relies on the air stripping mechanism to remove
volatile contaminants from the saturated zone. The injection, or "sparging," of clean air into the saturated
zone is coupled with vacuum extraction to recover volatile contaminants within the vadose zone. While
analogous to in-situ air stripping and vacuum extraction, the fundamental kinetics of ASP/VE have yet to
be clearly elucidated. The ASP/VE design is empirically based [Marley et al., 1992a], and the design
strategy revolves around the limitations imposed by subsurface geology, contaminant volatility, and the
nature and areal extent of contamination.
As clean air is injected into liquid saturated, homogeneous, isotropic porous media, the region
affected by a properly pressurized air sparger is assumed to be conical in shape, having some radius of
influence, r|nf, as shown in Figure 3.6.1.1. The actual flow regime of the sparged air through the porous
media is not clearly understood at this time. One theory suggests that air flows through the porous media
as discontinuous spherical micro-bubbles, thus possessing a large surface area to volume ratio which
favors partitioning of gases across the air-liquid interface [Loden and Fan, 1992; Sellers and Schreiber,
1992]. A second theory suggests that the air flows continuously in discrete and stable channels through
pores which represent the paths of least resistance [Loden and Fan, 1992; Marley et al., 1992a]. While
micro-bubbles can be generated using an in-situ diffuser to promote micro-bubble percolation, it is more
likely that the actual flow regime is more channelized, owing to the coalescing of micro-bubbles under the
operating injection rates.
Figures 3.6.1.2a and 3.6.1.2b schematically illustrate the influence that heterogeneities can have
on the success of ASP/VE [Marley et al., 1992a; Martin et al., 1992; Loden and Fan, 1992], Depending
on the type and distribution of heterogeneities and the areal extent of subsurface contamination, different
injection strategies may be required such as those pictured in Figure 3.6.1.3. The air spargers should be
installed below the heavily contaminated soil zone, as shown in Figure 3.6 1.2, to permit the sparged air
to contact and hence vaporize aqueous and separate phase volatile NAPLs, as well as to promote their
desorption.
99
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Vertical placement of injection wells is favored in coarser soils because they possess a low air entry
pressure making a large rinf possible, all other factors being equal. While the air entry pressure governs
the lowest possible air injection pressure, the maximum injection pressure must be less than the soil
fracturing pressure to prevent fracturing and the subsequent short-circuiting of air flow [Marley et al.,
1992a]. However, a large rinf may result in by-passing of contaminated soil zones due to sparger over-
pressurization, as shown in Figure 3.6.1.4, or due to subsurface heterogeneities, Figure 3.6.1.2b.
The effective rjnf of an individual sparger (or system) can be gauged by: pressure response (>0.1
inches H20) of vadose zone to the applied vacuum; rise in water table elevation; increases in volatile
contaminant vapor concentrations in the neighboring extraction wells; and increases in the dissolved
concentration of oxygen in neighboring monitoring wells. Rinf on the order of 5 to 20 feet has been
observed in coarse soils [Marley et al., 1992a,b], while rinf in stratified environments has been observed
on the order of 40 to 60 ft. [Marley et al., 1992a,b,c; Martin et al., 1992]. Other studies report that r|nf of
50 to 150 ft is possible [Gudemann and Hiller, 1988; Brown and Fraxedas, 1991; Brown et al., 1991], and
rinf has been reported to potentially extend up to 300 ft under sealed surfaces such as geosynthetics, paved
areas, and buildings [Gudemann and Hiller, 1988].
Recovery of volatile NAPLs requires that vacuum extraction be continuous. Air sparging can be
continuous, but in normal practice it is often pulsed. The combination of sparging in the saturated zone
and reduced air pressures in the vadose zone often leads to increases in the ground water table elevation
which can be on the order of several feet.
Aerobic in-situ biodegradation of NAPLs may result as a secondary benefit of ASP/VE. In fact, air
sparing is often used as a means of oxygen delivery for in-situ aerobic processes (section 3.2.1). The
stimulation of microorganisms as a result of oxygenation is often referred to as bioventing, and its
contribution to overall removal is often reported, but is usually not accurately quantifiable. In one year-long
air sparging and vapor extraction experiment, 23% mass reduction of gasoline was attributed to in-situ
biodegradation [Johnson et al., 1992].
Sf ARGING
POINT
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INJECTION
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~1
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Figure 3.6.1.1 Schematic of air sparging/vacuum extraction system [Sellers and Schreiber, 1992].
100
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Air Injection Off-gas
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Schematic of (a) typical air sparging system configuration and (b) the effect of subsurface
heterogeneities on gas channeling [Marley et al., 1992a].
Field Implementation-
A good summary of the design of soil vapor extraction systems is presented in the Environmental
Protection Agency's "Soil Vapor Extraction Technology: Reference Handbook" [1991d]. Construction of air
sparging systems is essentially identical, with a few minor changes [Marley et al., 1992a; Martin et al.,
1992]. The injection and vacuum well risers have often been constructed of 1- to 1.5-inch diameter PVC
(schedule 40-80) or galvanized steel Stainless steel construction allows for the heating of the injected air
as well as operation in corrosive environments. Compatible well point screens have been fitted using
threaded 10-slot or 20-slot PVC screen sections.
101
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Screened intervals in injection wells are usually limited to 1 to 3 ft. Longer screen lengths are
unwarranted because the air enters the porous media at the top of the screened interval; i.e., the point of
lowest hydrostatic head. Vertical nesting of injection wells has been implemented, and header type
manifolds are used to accommodate different injection pressures [Marley et al., 1992a; Loden and Fan,
1992]. Dual injection/extraction wells are common [Kresge and Dacey, 1991; Loden and Fan, 1992].
Extraction wells are fully screened in the vadose zone to within several feet of the capillary fringe to
accommodate the water table mounding. The screened and riser zone backfill consist of coarse silica sand
and bentonite, respectively.
Level of Demonstration and Performance--
Numerous sites in North America and Europe containing dense hydrophobic organic compounds
are reported to have been remediated by ASP/VE to concentrations in the range of 10 to 1000 ppb [Brown,
1992; Loden and Fan, 1992]. Sites contaminated with PCE, TCE, TCA, 1,2-DCE, BTEX compounds, and
petroleum compounds in sandy and silty soils are among those reported to have been successfully treated
by ASP/VE [Marley et al., 1992a; Loden and Fan, 1992, Brown and Fraxedas, 1991]. A recent ASP/VE
technology review highlights 21 ASP/VE applications, nine of which include DNAPLs [Loden and Fan,
1992]. Several other case studies are also available [Marley et al., 1992a; Martin et al., 1992; Middleton
and Killer, 1990, Brown and Fraxedas, 1991; Kresge and Dacey, 1991; Brown et al., 1991]. Table 3.6.1.1
highlights several applications at DNAPL contaminated sites, three of which are described below.
In Connecticut, a 4-week pilot study was conducted at a spill site in a 2,000 sq ft test cell with a
depth of 40 ft [Marley et al., 1992a; Martin et al., 1992]. The subsurface consisted of stratified fine sands
and silts. Initially, the water table was located at a depth of 20 ft. Ground-water TCE concentrations within
the test cell ranged from 0.76 mg/l to 11 mg/l. Seven spargers were installed at depths of 27 to 40 ft.
Intermittent pairs of air spargers were pulsed at injection rates of 3 to 10 scfm and pressures of 15 to 60
psi on a 4/2 hr or 3/3 hr (on/off) cycle because of problems encountered (see next paragraph). Two
extraction wells were nested in the vadose zone. The extraction wells operated continuously at 70 scfm
(combined) at vacuums of 15 to 20 inches H2O. Four pounds of VOCs (primarily TCE) were recovered.
Two weeks after shutdown, VOC concentrations were reported to have returned to background levels.
One finding of this study was that ground-water mounding was evident approximately 60 ft outside
of the test cell. This mounding, and the consequent lateral migration of aqueous phase VOCs, resulted
from the preferential horizontal air flow caused by soil stratification. Stratification also promoted short
circuiting of air flow from spargers into monitoring wells nested at depths of 19 to 28 ft both inside and
outside of the test cell. In these monitoring wells, the VOC vapor concentrations were as high as 150 ppm.
To mitigate these effects, pulsed injection rather than continuous injection was practiced.
Full-scale ASP/VE was used at a site in Germany with a subsurface characterized as quaternary
sands and gravels to a depth of 110 ft [Gudemann and Miller, 1988]. The water table was situated at a
depth of 27 ft, and a silty sand layer was located at depths of 44 to 47 ft. The unsaturated and saturated
zones were contaminated with TCE and PCE. The soil was vented for 100 days using two extraction units
capable of 475 scfm flow. Venting alone recovered 5100 Ibs of solvents (combined TCE and PCE).
Sparging then commenced using six-injectors at depths of 37 ft with flow rates of 6 scfm. Ground-water
concentrations decreased from an initial 33 ppm to 0.027 ppm in 3 months. ASP/VE treatment removed
a total of 8900 Ibs of solvents in 8 months of application.
In the United Sates, full-scale ASP/VE was used at a site underlain by coarse sands which had
been contaminated by PCE, TCE, TCA, 1,2-DCE, and petroleum hydrocarbons (TPH) [Brown et al., 1991].
The water table was located at a depth of 11 to 14 ft., and subsurface contamination appeared to be
concentrated in the intervals of 3 to 9 ft and 15 to 18+ ft, below the ground surface. Initial readings in ten
ground-water monitoring wells included: a high of 41,000 ppb total VOCs (excluding TPH); two below
102
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Extraction
Well
Sparging
Well
Extraction
Well
SPACED CONFIGURATION
Extraction
Well
Well
NESTED WELLS
HORIZONTAL WELLS
COMBINED HORIZONTAL/VERTICAL
Figure 3.6.1.3 Possible air sparging well configurations [Loden and Fan, 1992].
PROPERLY PRESSURIZED SYSTEM
OVERLY PRESSURIZED SYSTEM
Figure 3.6.1.4 Effect of gas injection pressure on air sparging system [Loden and Fan, 1992].
103
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detection levels (BDLs); a ten well average of 5685 ppb; and, a median of 1137 ppb. The ASP/VE system
consisted of seven spargers, seven dual sparger/extraction wells, and one extraction well. The dual wells
were installed to depths of approximately 33 ft, and the seven spargers were installed to 7 to 9 ft below
the water table.
First, a vacuum (20-30 inches H2O) was applied by the 8 extraction wells at a combined flow of
450 scfm. Sparging then commenced in the dual wells using a blower capable of delivering 270 scfm at
10 psi (specific operating parameters not provided). The remaining spargers were then engaged. After
125 days of ASP/VE treatment and one week of shutdown, VOC concentrations in the ten monitoring wells
averaged 130 ppb, had a median of 13.5 ppb, a high of 897 ppb, and two BDLs. ASP/VE removed
approximately 900 Ibs of PCE and TCE from the subsurface.
Information on the long term efficacy of ASP/VE is lacking. In one full-scale implementation of
ASP/VE at a site contaminated with up to 10,000 ppb total BTEX compounds, ground-water concentrations
were reduced to approximately 600 ppb BTEX and less than 0.5 ppb benzene and remained stable for a
period of 6 months after sparging was shut down [Marley et al., 1992a].
Applicability/Limitations-
Sites contaminated with dissolved volatile and semi-volatile hydrocarbons possessing Henry's
constants greater than 105 atm-m3/mole are good candidates for ASP/VE treatment, depending on
subsurface conditions. To obtain sufficient in-situ air flow in the saturated zone, a minimum soil hydraulic
conductivity of 0.001 cm/sec is required [Middleton and Miller, 1990]. While sites underlain by gravel, fill,
sand, and sandy and silty lenses have been treated by ASP/VE, the process is strongly controlled by
stratigraphic heterogeneities; and therefore, careful well placement and a site specific clean-up strategy is
required. In addition, to mitigate potential lateral spreading of contaminants, peripheral containment or
extraction wells may be required [Marley et al., 1992a,c; Martin et al, 1992].
Many authors state that the intended use of ASP/VE is to remediate contaminants in the aqueous
phase and sorbed on the soil [Felten et al., 1992; Loden and Fan 1992; Marley et al., 1992a,b,c; Sellers
and Schreiber, 1992; Leonard and Brown, 1992]; but considering that sparging occurs below the zone(s)
of contamination and the relatively low sorption potential of coarse soils, it is likely that NAPL lenses are
affected by spargers. Since air sparging changes the pressure regime within the vicinity of the sparger,
NAPLs may be potentially mobilized laterally beyond the treatment zone, or vertically downward below the
sparger.
The question of the presence of sparged air as discrete micro-bubbles or stabilized air channels
has very different implications in terms of mass transfer limitations and potential mobilization of NAPLs.
Sellers and Schreiber (1992) developed a simple air sparging model which estimates clean-up times and
ground-water concentrations. Sparged air is modeled as discrete micro-bubbles, and the model suggests
that contaminant removal is diffusion limited. In two simulations using published field data, the results of
the Sellers and Schreiber model compared favorably with the field observations of Marley et al. (1990,
1992c), but not those of Brown et al. (1991).
A 3-D air sparging model using Darcy's Law for multiphase flow compared well with two sets of
actual field data [Marley et al., 1992b]. This model compares predicted and measured air pressure
distributions and flow velocities in the subsurface. It is not clear whether the sparged air is treated as
discrete micro-bubbles or as stable channels. However, the actual flow regime, whether pulsed or
continuous, may have important consequences If air flow occurs as stable channels, the removal process
will be mass transfer limited. Subsurface contamination by semi- and non-volatile DNAPLs may be
potentially exacerbated by the preponderance of stable channels in the saturated zone. The spreading
behavior of certain DNAPLs may permit them to migrate as films along the air-water interfaces (see section
104
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2.2.4). Under these conditions, potential downward migration of DNAPL out of the treatment zone may
result.
With the exception of borings, ASP/VE is not intrusive; and therefore, few limitations are present
with respect to interference from ground structures, overhead or buried utilities, and other subsurface
obstructions. Site grading is not a problem. The above ground hardware can be trailer-mounted and
constructed of readily available materials and standard unit operations equipment [USEPA, 1991d]. Some
of these systems can be automated, monitored and operated from remote locations.
ASP/VE may stimulate in-situ biological degradation through oxygenation of the subsurface which
may lead to biological fouling. Precipitation of metal carbonates and oxides may clog the aquifer [Felten
et al., 1992]. Despite its empirical nature and its drawbacks, ASP/VE has been very successful in attaining
negotiated cleanup goals relatively quickly- on the order of several months in many cases.
Cost and Availability-
ASP/VE was developed specifically as a physical-chemical treatment technology in Germany in the
mid-1980's [Gudemann and Miller, 1988]. However, earlier incorporation of ASP/VE as an oxygen delivery
system as part of in-situ biological treatment in the saturated zone occurred in the late-1970's to early
1980's in the US [Marley et al., 1992a]. Regardless of its actual origin, Loden and Fan (1992), in a recent
technology review report that hundreds of sites (presumably contaminated with volatile hydrocarbons
including DNAPLs) within the US and Europe have been remediated by ASP/VE. See Table 3.6.1.1 for
data specific to dense hydrophobic organic compounds. The hardware for ASP/VE is readily available as
is the expertise.
ASP/VE is a good candidate for remediating dissolved phase plumes of volatile hydrocarbons in
aquifer media. Hot-air injection is likely to enhance stripping. The potential mobilization of the separate
phase makes its application to spreading DNAPLs questionable.
Costs are site specific, and reporting of ASP/VE costs has been poor. The pilot study which
recovered 4 Ibs of VOCs (primarily TCE) cost approximately $140,000. Using an estimated test cell volume
of 80,000 cu ft, this cost translates to approximately $50/yd3. Using soil vapor extraction alone as a bench
mark of approximately $50/yd3, full scale ASP/VE is estimated to be approximately $75-$125/yd3 [Fan,
1992]. Discharged vapors are normally treated by granular activated carbon units.
3.6.2 Vacuum Vaporizer Wells (UVB)
Theoretical Background-
Vacuum-Vaporizer-Wells (UVB, in German: Unterdruck Verdampfer Brunnen) rely on the air
stripping mechanism to recover volatile DNAPLs. In-situ air stripping is achieved in two ways: actively, by
direct "in-well" stripping of volatile NAPLs from the ground water; and, passively, by soil vapor extraction
in the vadose zone which may recover volatile compounds emanating from dissolved plumes [Herrling et
al., 1992a,b]. Since the emphasis of this document is on the saturated zone applications, soil vapor
extraction is not formally addressed here. However, a good summary of soil vapor extraction systems is
available [USEPA, 1991d],
The effectiveness of the UVB to remediate a contaminated aquifer depends on compound solubility
and volatility, and the ability of the UVB to recirculate treated ground water within the aquifer. The essential
components of the UVB design include the circulation system, sphere of influence, and capture zone of an
individual UVB or UVB field [Herrling et al., 1991, 1992a; Herrling and Stamm, 1992a]. UVB differs from
traditional ground-water wells in that the generated radial flow regime is not strictly horizontal: non-
negligible vertical components of ground-water velocity exist. In quiescent ground-water environments, the
106
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three-dimensional flow field can be represented by a simple two-dimensional (2-D) streamline analysis
[Herrling et al., 1991, 1992a].
Figure 3.6.2.1 a depicts the symmetric vertical recirculation pattern and streamlines created by a
UVB. Figures 3.6.2.1b,c are cross sections taken through the UVB normal to the flow direction which
illustrate the general effect of the overall ground water flow regime on the recirculation pattern. However,
the actual radial flow pattern generated by the UVB is asymmetric in three dimensions. Therefore, analyses
using curved separating stream surfaces and particle tracking methods are often used to delineate UVB
capture zones [Herrling and Buermann, 1990; Herrling et a!., 1991]. Anisotropic effects can be incorporated
as well. Figure 3.6.2.2 schematically shows the 3-D capture zone of the contaminated ground water. The
recycling of treated water in the upgradient direction depresses the flow lines along the path of the
contaminated water to the UVBs, thus making the capture zone within the aquifer wider at the bottom than
at the top. By overlapping capture zones, dissolved plumes of volatile DNAPLs can be effectively treated.
UVBs can also be placed in parallel, that is, a second UVB is placed immediately downstream of the first
UVB to provide additional stripping [Herrling et al., 1992b].
Field Implementation-
Three configurations of the UVB apparatus are illustrated in Figure 3.6.2.3. The well configuration
shown in Figure 3.6.2.3a has a separation plate dividing the UVB into two distinct regions: an "extraction"
region in which ground-water extraction occurs; and a "stripping" region in which air stripping, vapor
extraction, and ground water recirculation occurs [Herrling et al., 1992a]. Contaminated ground water
enters the well via the lower well screen and exits via the upper well screen after being air stripped. Clean
air, drawn from outside the well, enters the water column through an adjustable "pinhole plate" apparatus.
The pinhole plate is situated at an elevation in the water column corresponding to sub-atmospheric
pressure. The upward flow of the bubbles creates an "in well" stripping region, and due to the efficiency
of mixing, an air/water ratio of 10:1 is achieved [Herrling et al., 1991]. Contaminant laden vapors from
stripping, and soil gas vapors from the vadose zone which entered the UVB through the unsaturated portion
of the upper well screen, are exhausted to the off-gas treatment system.
(a)
(b)
(c)
Figure 3.6.2.1 Streamlines for longitudinal vertical recirculation patterns for several ground-water flow
velocities: (a) 0 m/d; (b) 0.3 m/d; (c) 1.0 m/d [Herrling et al., 1991].
107
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(a)
KH= 0.001 m/s
Kv= 00001 m/s
(b)
UVB1
= 0.001 m/s
Kvr 00001 m/s
Figure 3.6.2.2 Schematic of three-dimensional capture zone for anisotropic soil conditions using (a) single
and (b) dual UVBs. Effect of recirculation cell on incoming flow is indicated by the
depressed areas [Herrling et al., 1991].
Sub-atmospheric pressures in the stripping region are maintained by a ventilation system. The
applied vacuum concurrently induces a rise in the ground-water table elevation in the UVB, inflow of clean
air bubbles into the well via the pinhole plate, and vapor inflow from the surrounding vadose soils into the
well. Air bubbles strip compounds from the ground water. All collected vapors exit the top portion of the
well and are routed to the off-gas treatment system. Rising air bubbles induce upward convective flow
within the well which is usually sufficient to draw more contaminated ground water into the lower well
screen. In certain cases, additional ground-water pumping may be necessary. Hence, the use of partially
penetrating well screens in the lower and upper well regions and the adjustable separation plate facilitates
the ground-water circulation within the aquifer
108
-------
fresh oir activated carbon
ventilator filter
used air
cleaned air
air removed
via suction
stripping zone*
.working water level
fresh air activated carbon
ventilator scte<
circulation
borehole filling
^ Wteroravel sealing material
aquifer bottom
aquifer bottom
Figure 3.6.2.3 Schematic of vacuum vaporizer well (UVB) configured with (a) separation plate and
vacuum extraction; (b) no separating plate and vacuum extraction; and, (c) separation plate
and closed air recirculation. [Herrling et al, 1992aj.
109
-------
At many sites, the configuration without a separation plate depicted in Figure 3.6.2.3b is used
[Herrling et al., 1992a]. While it would appear that the lack of a separation plate should result in excessive
short circuiting of fluid flow within the well, density differences between the untreated and treated ground
water can result in net fluid flow in either the upward or downward direction. Ground water may either be
heated or cooled as a result of heat transfer with the fresh air: heated (less dense) and cooled (more
dense) water will exit through the upper and lower screens, respectively. In this way, water circulation
patterns are reversible, and both have been observed in the field [Herrling et al., 1992a].
Both configurations shown in Figures 3.6.2.3a and 3.6.2.3b are susceptible to clogging as a result
of iron, manganese and calcium precipitation [Herrling et al., 1992a]. Treated air recirculation and water
re-introduction below the working phreatic surface can mitigate these effects. After air is recycled a few
times and air-water equilibrium is re-established, precipitation of insoluble salts should cease or be reduced.
The incorporation of separation plates and additional water pumps, and the potential for extended ground-
water circulation within the aquifer, make the configurations shown in Figures 3.6.2.3a and 3.6.2.3c the
most preferred configurations [Herrling et al., 1992a].
UVBs have been installed to depths of 40 meters [Herrling et al., 1991, 1992a]. Multiple screened
intervals, nested ground-water sampling ports, separation plates and additional water pumps can be
installed to selectively create recirculation cells within any vertical portion of an aquifer [Herrling et al.,
1991]. Screened intervals have been in the range of 2 to 5 meters. Reported volumetric flow rates of UVB
ventilation systems are as high as 500 m3/hr [Herrling et al., 1991; 1992a]. Fresh air and soil gas vapor
inflows to UVBs have been reported as high as 180 and 320 m3/hr, respectively [Herrling et al., 1992a].
Depending on the well configuration, UVBs may stimulate in-situ biological degradation of organic
compounds through oxygenation of the subsurface [Herrling and Stamm, 1992a,b]. This observation has
led to the development of Groundwater Circulation Wells (GZB, in German: Grundwasser Zirkulations
Brunnen). GZBs facilitate continuous or pulsed introduction of aqueous phase compounds for physical or
biological treatment of ground water using the same ground-water circulation strategy as UVBs [Herrling
et al., 1992b; Herrling and Stamm, 1992a,b].
Level of Demonstration and Performance-
Numerous sites (60+) within Europe containing immiscible compounds are reported to have been
remediated to concentrations in the range of 10 to 1000 ppb [Herrling et al., 1991, 1992a,b]. Sites
contaminated with PCE, TCE, TCA, 1,2-DCE, DCM, and BTEX compounds in sandy to silty soils are
among those reported to have been successfully treated by UVB [Herrling et al., 1991, 1992a,b]. Two
European applications are summarized here.
In 1988-91, UVB was applied at a former steel processing plant in Rhine-Ruhr region of Germany
to clean up TCE contaminated ground water. The site was underlain by approximately 40 meters of
interbedded fine to medium sands and gravels with occasional silt lenses [Herrling et al., 1991]. The
ground-water table was located at a depth of 6 m, situated just below an upper layer composed of artificial
fill. Pump tests indicated hydraulic conductivity coefficients on the order of 10~3 to 5x10"4 m/s. UVB1 was
installed to a depth of 12.5 m with a screened interval of 4 to 12.5 m using the configuration without a
separation plate shown in Figure 3.6.2.3b. About 20 m from, and somewhat downgradient of, UVB1, UVB2
was installed to a depth of approximately 40 m with three screened intervals 6-8.2 m; 20-25 m; and, 35-40
m. A separation plate was installed between each screen. TCE was detected up to a maximum
concentration of 5 ppm and other organic compounds were detected at much lower concentrations. Before
pumping commenced, the concentration profile in the UVB2 sampling ports were 1.26, 1.22, 1.64 ppm at
depths of 11, 24, and 38 meters, respectively.
110
-------
• UVe,u('| < it euiu ei ,
4- B22b,uppe< pan cl ih
liO
ug/l
100
UVB,lower measurement location
UVB,upptr measurement location
r part of the aquifer
ug/l I
I 00 I
Figure 3.6.2.4 Field data obtained from Mannheim-Kaefertal site (Germany). Measured hydraulic heads
(a) indicate vertical flow patterns in aquifer. Downflow in well occurs until 6/13/89, upflow
thereafter. Corresponding PCE concentrations in ground-water monitoring locations in the
lower UVB (b), upper UVB (c), and in a downgradient well (d) which is screened in the
upper portion of the aquifer illustrate the importance of upflow in the UVB well on PCE
recovery [Herding et al., 1992a].
Ill
-------
In terms of recovery, vapor concentrations exiting UVB1 fluctuated between 120-300 mg/m3, and
average daily removal rates were approximately 2.0 kg/d. Approximately 1500 kg of VOCs were recovered
from the unsaturated and saturated zones in 15,000 hours of operation. After one month of operation,
vapor concentrations in the sampling ports of UVB2 decreased to 0.1 ppm of VOCs and remained at this
level. UVB2 commenced pumping some 800 days after UVB1. Daily vapor recovery rates at UVB2 were
initially 375 g/d and decreased to 20 g/d. In 4,000 hours of operation, LJVB2 recovered 50 kg of VOCs.
Dissolved organic compound concentrations in a monitoring well situated 60 meters upstream of
UVB1 and just beyond the estimated location of the upstream stagnation point of the recirculated water
steadily decreased from approximately 1.5 ppm to 0.3 ppm. Within the recirculation zone, concentrations
decreased from 1.5 ppm to 0.2 ppm in a well situated 10 m from UVB1. Wells located downstream from
UVB1.2 showed a decrease in concentrations of volatile NAPL from initial readings of 2-5 ppm to 0.4-0.8
ppm. However, ground-water concentrations in a monitoring well located approximately 150 meters
upgradient from UVB1 fluctuated between 0.46 and 5.08 ppm and showed an unexplained increasing trend
with time. While some dilution may have occurred and another NAPL hotspot may have been located,
recovery of dissolved contaminants, primarily TCE, was continuous and considerable.
In 1989, field experiments were conducted at a site in the Mannheim-Kaferetal area (Germany)
[Herrling et al., 1992a]. In these experiments, the effect of flow reversal in the UVB on efficacy of cleanup
was investigated. The stratified subsoils consisted of interbedded sands and gravel to a depth of
approximately 38.7 m where a clay aquitard was encountered. A discontinuous lens of clay appeared at
a depth of 16.5 to 18.3 m. One UVB was installed to a depth of 40.0 m with upper and lower screened
intervals of 8.5-14.0 m and 35.5-38.5 m, respectively. The configuration without the separating plate shown
in Figure 3.6.2.3b was used for the first six months of operation, and then was modified to include a
separation plate and a water pump. One monitoring well, situated 15 m downgradient of the UVB, was
installed only in the upper portion of the aquifer.
Chlorinated hydrocarbons (including PCE) were detected at concentrations between 0.1-0.2 ppm.
Figure 3.6.3.4a presents total hydraulic head data from the two sampling ports within the UVB.
Concentration data obtained during the period when no separation plate was used (Phase 1: 1/89-6/89)
indicate that downflow was occurring in the well and that the water flow m the surrounding aquifer was
opposite of that depicted in Figure 3.6.2.1 a After separation plate installation (Phase 2: after 6/89), the
flow direction in the well was reversed to the upflow direction making aquifer flow consistent with that
shown in Figure 3.6.2.1 a. The effect of flow reversal on the PCE concentration data taken from the UVB
sampling ports is shown in Figure 3.6.2.4b,c. For example, significant PCE concentration reductions were
observed in the downgradient well after the beginning of Phase 2.
Since treated water was cycled to the top of the aquifer and the downgradient well was only
screened in this interval, the flushing effect of the treated water on this well is manifested by PCE
reductions. The PCE reduction occurring at this location are consistent with what is expected from the
aquifer recirculation pattern shown in Figures 3.6.2.1b,c.
Applicability/Limitations-
UVBs can be applied to sites contaminated with aqueous phase volatile and semi-volatile organic
compounds having Henry's constants greater than approximately 105 atm-m3/mole. It is not known whether
the separate phase liquids will be mobilized: most likely they will not be mobilized and the solubilization
process will be diffusion limited.
Good site characterization is required to avoid cross-contamination of unconfined and confined
aquifers. Sites underlain by gravel, fill, sand, and sandy to clayey lenses have been treated by UVBs
[Herrling et al., 1991; 1992a,b]. The hydraulic conductivities of subsurface soils at treated sites have been
112
-------
in the range of 10~3 m/s to approximately 10~6 m/s [Herrling et al., 1991; 1992a]. Ambient ground-water
velocities that have been accommodated are reported as high as 1 m/d.
With the exception of the UVB itself, this process is not intrusive; therefore, there are few limitations
due to interference from ground structures, overhead or buried utilities, and other subsurface obstructions.
Site grading is not a problem.
There is no ground-water extraction to the ground surface and no overall lowering of the phreatic
surface; that is, pronounced cone of depression is not formed [Herrling et al., 1992b]. The only ex-situ
process is the treatment of the extracted gas phase. For UVB configurations which do not recirculate air
(Figures 3.6.2.3a,b), oxygenation of the ground water may potentially lead to biological fouling and
precipitation of metal carbonates and oxides in both the UVB and aquifer. Hot-air injection is likely to
enhance stripping.
Cost and Availability-
The technology is commercially available and has been implemented on the full-scale in Europe.
The UVB and GZB have been patented by I EG mbH (Reutlingen, Germany). Their US affiliate is IEG
Technologies Corp. (Charlotte, North Carolina).
UVB technology is a good candidate for remediating dissolved plumes of VOCs in aquifer media.
However, because solubilization of the separate phase is diffusion limited, application of UVBs to DNAPL
cleanup is limited.
The application in the Rhine-Ruhr area (Germany) in which approximately 1550 kg of volatile
NAPLs (primarily TCE) cost $352,000 which includes site investigations planning etc. (21.8%), monitoring
and field work (21.5%), analytical work (8.2%), borings and UVB installation (15.3%), granular activated
carbon treatment and NAPL disposal (24.1%), and energy cost (9.1%) [Herrling et al., 1991]. Average
monthly operating costs were $4,000. These costs are somewhat inflated by technology development and
specific costs associated with conducting business in Germany (insurance, patents, regulations). Because
it is composed of elements common to ASP/VE, soil vapor extraction, and pump-and-treat, UVB is likely
to cost about $50-100/yd3.
113
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3.7 THERMAL PROCESSES
Introduction-
Thermal and thermally enhanced processes rely on various means for the delivery of thermal
energy into the subsurface: the CROW® process (see below) utilizes hot water and/or low quality steam
injection; in-situ steam enhanced extraction (SEE) relies on high quality steam injection; and radio
frequency heating and in-situ vitrification facilitate heating using microwave and electrical arrays,
respectively. During these processes, steam and hot water progress through cool porous media and they
heat the interstitial fluids and porous media. These fluid-fluid displacement processes are analogous to
liquid-liquid displacement processes (see section 3.5) with the added complexity of heat transfer. The
contaminants can be recovered as vaporized gases, and as dissolved- and separate-phase liquids.
The effectiveness of the CROW® process (section 3.7.1) and steam enhanced extraction (section
3.7.2) is controlled by the thermodynamics and hydrodynamics of hot-water and steam displacement in
porous media. Thus, the thermal properties of both the porous media and the pore fluids become
important. The orientation and shape of the propagating steam fronts are governed by the matrix
heterogeneities, geometry of the aquifer, initial moisture and boundary conditions, steam quality, injection
rates, and most importantly, the ratio of buoyancy to viscous forces. In saturated homogeneous isotropic
porous media, the ratio of buoyancy to viscous forces is important in terms of gravity override and effective
sweep-out [Basel and Udell, 1989]. The same principles hold for condensation fronts propagating through
layered media, but the temperature profiles and fronts will be curved at layer interfaces owing to intrinsic
permeability differences [Udell and Stewart, 1989]. When gravity effects are negligible, the behavior of
propagating fronts can be readily predicted and controlled [Buckley and Leverett, 1942; Udell and Stewart,
1989].
Radio frequency heating (section 3.7.3) achieves subsurface heating by using an electrode array
system to transmit electromagnetic waves through the porous media. In-situ moisture is converted to a
steam front which propagates through porous media thus displacing other pore fluids, including DNAPLs,
in a manner similar to that described above.
In-situ vitrification (ISV, section 3 7.4) also employs an electrode array system, but for the purposes
of current flow. Large current flows cause electrical resistance (joule) heating of the soil to the melting
point. During this process, DNAPLS can be volatilized and pyrolized.
The CROW® process, SEE and radio frequency heating processes have their origins in the
enhanced oil recovery business. ISV was developed for the stabilization/solidification of wastes containing
radionuclides. All of these technologies have been demonstrated at the pilot scale, but only CROW® and
SEE have been successfully demonstrated in the saturated zone. A full-scale demonstration of SEE is in
progress.
3.7.1 Contained Recovery of Oily Wastes (CROW®)
Theoretical Background-
The Contained Recovery of Oily Wastes (CROW®) process uses low-quality steam and hot-water
injection to enhance contaminant removal from the subsurface. The primary mechanisms are: the flotation
of NAPL contaminants by temperature induced viscosity reduction and buoyancy; and by displacement of
dissolved contaminants and NAPL by a propagating hot-water front. Secondary mechanisms include
solubility enhancement of the targeted compounds which assists in their recovery, and enhanced in-situ
biological degradation (see Section 3.2) [Western Research Institute, 1992; Johnson and Suddeth, 1989].
This process is in many ways identical to enhanced oil recovery methods utilizing steam, solvent, surfactant
or caustic floods [Shah, 1981; Janssen-van Rosmalen and Hesselink, 1981].
114
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As low-quality steam and/or hot water enter and progress through cool porous media, they heat
the interstitial fluids and porous media. Unlike steam enhanced extraction (SEE, section 3.7.2) which relies
on steam front propagation for contaminant displacement, the low-quality steam employed by the CROW®
process results in a hot water front which immiscibly displaces NAPL contaminants. By utilizing hot water
as the displacing fluid, the viscosity, buoyancy, and capillary pressures of DNAPLs can be favorably
affected. The relative permeability of the porous media to water (kw) is also increased.
The viscosity reduction of several petroleum-derived DNAPLs is shown in Figure 3.7.1.1 a. If the
viscosity reduction is not sufficient, surfactants and viscosifiers can be added to the hot water to improve
mobility ratios. The addition of surfactants enhances DNAPL solubilization and interfacial tension reduction
[Fountain et al., 1991]. Capillary pressure-saturation relationships and parameters for Brooks-Corey or van
Genuchten equations as a function of temperature are necessary to model the displacement process.
Depending on the actual compound and applied temperature, a DNAPL may be effectively transformed to
a LNAPL, which aids in flotation and thus free product recovery. The temperature dependence of density
is shown for several petroleum-derived DNAPLs in Figure 3.7.1.1b.
Field Implementation-
A schematic of the field implementation is shown in Figure 3.7.1.2. Horizontal well configurations
are also possible [Johnson, 1992]. A specially designed injection well has the capacity to simultaneously
inject three different fluids (low-quality steam, hot, and cool water) at three separate elevations. Production
(extraction) wells recover aqueous and pure phase DNAPL which is pumped to the ground surface for
treatment and/or recycle. Hot and cool water may be re-injected after treatment.
The objective of the CROW® process is to upwardly displace, or float, DNAPLs toward the water
table and extraction wells by reducing separate phase density, viscosity, and interfacial tension. The
strategy implemented to accomplish DNAPL mobilization is summarized below.
I \ \ Coal Tar
-\\
PA
Laramie, WY
Wood Treating Wastes
\
40°C 50°C
Inverse Temperature, K'1 x 1C3
40 60 8C 100 120 140 160
T «mp«rotur«, f F
Figure 3.7.1.1 Influence of temperature on fluid viscosity (a) and density (b) for several DNAPLs [Johnson
and Suddeth, 1989].
115
-------
Injection Well
Production Well
Steam-Stripped
Water
Low-Quality
Steam
Hot-Water
Reinjection
Absorption Layer
-TT
Oil and Water
Production
Residual Oil
Saturation .
Hot-Water
Displacement
• Original Oil'.•
Accumulation '
Hot-Water
Flotation
Steam
i njection
Figure 3.7.1.2 Conceptual schematic of the CROW® process [Johnson and Suddeth, 1989].
100
£ 80
c
o
"§ 60
0)
"c 40
c
o
x 2°
6
0
6
Volume %
Che mical
.87
.95-
Q M
o No Chemical ' "
• Chemical-Added
-
*~ o
o
1 1 , i 1 1 1
0 80 100 120 140
160
180
Temperature, °F
Figure 3.7.1.3 Temperature dependence of DNAPL recovery using hot water and surfactant solutions in
one-dimensional column tests by CROW® process [Johnson and Leuschner, 1992].
116
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Low-quality steam is injected below the contaminated soil zone which results in heating of the
porous media and pore fluids. DNAPL is mobilized upward providing it becomes less dense than water
under the applied temperatures, and a favorable ratio of buoyancy to viscous forces is maintained. As
DNAPL migrates upwards, a concentrated NAPL bank (dark stippled area in Figure 3.7.1.2) may form
ahead of the propagating hot-water front. Displacement of this kind can result in a lower residual saturation
(Sor) of NAPL and an associated increase in kw in the swept-out zone. Also, enhanced DNAPL
solubilization can occur at elevated temperatures in the zone swept by hot water.
Simultaneous injection of hot water along the periphery of the contaminated zone serves the dual
purpose of lateral containment and displacement of DNAPL toward the extraction well. The lateral
displacement occurs analogously to the upward displacement, although the frontal stability condition is
somewhat different owing to the orientation of forces. DNAPL solubilization occurs at elevated levels in
the zone swept by the continual passing of hot-water.
An "absorption layer" [Johnson, 1989] or cold water cap is created above the contaminated zone
by cool-water injection. The purpose of this layer is to provide vertical containment of the rising pore fluids,
and to condense any vapors emanating from the heated contaminated soil zone situated directly below.
This may not be necessary when dealing with non-volatile DNAPLs.
Any combination of compounds such as alkaline agents, surfactants, polymers (density
enhancement) and viscosifiers may be added to the injectales to ensure a more favorable mobility ratio
between the hot water and the DNAPL or to enhance compound solubilities. After the quantity of residual
DNAPL is reduced by the hot-water front, nutrients and electron acceptors (primarily hydrogen peroxide)
may be added to the hot water to enhance biodegradation of residual DNAPL.
Level of Demonstration and Performance-
One-dimensional studies in 3.75-in. dia. and 36-inch long packed columns using former wood
treatment plant contaminated soils containing creosote, pentachlorophenol (PCP) and petroleum products,
and manufactured gas plant contaminated soils containing oily residues have shown that the residual
saturations of DNAPLs could be significantly lowered depending on the waste type, applied hot-water
temperatures and surfactant addition [Leuschner and Johnson, 1990; Johnson and Leuschner, 1992].
Two test samples of wood treating plant soil contained 2.9 wt% and 7.4 wt% hydrocarbons and
PCP concentrations of 1,500 ppm and 3,200 ppm, respectively. Using a flow rate approximately twice that
of natural ground-water velocity and hot-water temperatures of 120°F (49°C) and 140°F (60°C), 0.5 wt%
hydrocarbon concentrations and PCP concentrations below 2.5 ppm were obtained for both samples. This
constitutes hydrocarbon reductions of 84% and 94% for the two soils, respectively.
Testing on the manufacturing gas plant soils was somewhat more extensive using soils that
contained 0.13 wt% to 3 wt% organics. Initially, the extremes were tested. Using a flow rate approximately
twice that of natural ground-water velocity and ambient water temperatures of 64°F (18°C), hydrocarbon
reductions of only 15% and 21% were obtained for the two soils, respectively. Injection of hot water at
100°F (38°C), 120°F (49°C), and 140°F (60°C) showed organic concentration reductions of 23, 30, and
42 wt%, respectively for the 3 wt% soil.
One-dimensional experiments using hot-water temperatures between 155°F (68°C) and 165°F
(74°C) were then completed on samples having up to 2.8 wt% hydrocarbon. Reductions were on the order
of 55 to 63 wt% with the optimum removal occurring near 155°F (68°C), as shown by the open circles in
Figure 3.7.1.3. One sample, initially having 0.13 wt% hydrocarbon was reduced by 61 wt% compared to
the 15 wt% obtained previously at 18°C. At 155°F (68°C), NAPL reductions were improved to between
64 and 84 wt% by using 0,45 to 0.95 vol% Igepal CA-750 surfactant solutions, as shown in Figure 3.7.1.3.
117
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These results are consistent with 1-D displacements performed in reservoir sands which showed that hot
caustic floods outperformed hot-water floods by 16-18% on a pore volume basis [Janssen-van Rosmalen
and Hesselink, 1981].
Next, two three-dimensional tests using hot- and cool-water injection were also completed on the
manufactured gas plant soils in a reaction box (3x3x7 ft) fitted with 80 thermocouples and
injection/extraction ports situated on the long axis. The soils were layered to simulate site stratigraphy: 1
ft impervious base; 0.5 ft of very saturated oily sand; 0.5 ft of lightly saturated oily sand; and, 1 ft cap of
clean sand. Both tests were run at flow rate approximately twice that of natural ground-water velocity. The
cool water was injected into the clean sand and hot water was injected at 155°F (68°C) into the oily sands,
see Figure 3.7.1.4. In both tests, the temperature profiles show that a cooler adsorption layer could be
maintained over the treated soil zone.
The duration of the first test was 100 hours using hot water only. The second test used hot water
and 1.0 vol% Igepal CA-750 (surfactant) addition and lasted only 50 hours due to clogging caused by
migration of fines. Comparison of the nearly coincident 60 wt% saturation reduction and 140°F profiles in
Figure 3.7.1.4a reveals that removal efficiency exceeds that of the 1-D tests, Figure 3.7.1.3. One possible
explanation for this may be the larger sample may permit the formation of a larger oil bank [Johnson and
Leuschner, 1992].
One pilot study has been completed at a former wood treatment plant site [Fahy et al., 1992]. The
main purpose of this test was to demonstrate hydraulic control of the hot-water front as it propagated in
the subsurface. The hot-water front was successfully kept within the capture zone of the extraction well
throughout the pilot test. Creosote and pentachlorophenol (PCP) in a fuel carrier oil comprised the
subsurface contamination. The pilot test was conducted in a 23-47 ft thick aquifer consisting of uniform
silty, fine to medium gravel and sands. The water table was situated between 10 to 20 ft below the ground
surface across the site. The aquifer was underlain by a 96 ft thick till layer having a hydraulic conductivity
of approximately 10"7cm/s.
The pilot test utilized one injection well, one extraction well, four monitoring wells, and three
piezometers. The injection well was screened from the top of the till layer and extended to within 5 ft below
the ground surface. The extraction well (previously installed) was located 50 ft from the injection well. The
monitoring wells were constructed of 2-inch I.D., 0.01 inch continuous slot stainless steel screen and solid
casing risers. The screened interval extended from the top of the till layer to approximately 13 to 16 ft
below the ground surface. The monitoring wells and piezometers were each fitted with thermocouples
situated at 18, 23, 28, and 33 ft, and, 22, 32, and 37 ft below the ground surface, respectively.
The extraction well was pumped continuously for one week prior to start-up of the injection well,
and continued until the end of the test, Day 41. Injection commenced on Day 7 and continued until Day
37. The injection and extraction wells were pumped at an average rate of 4.5 and 6.5 gpm, respectively.
No surfactant addition or pH adjustment of the hot water was employed.
The initial injection temperature of the hot water was 147°F (64°C), but on Day 9 it was elevated
to 203°F (95°C) for the remainder of the test. As shown in Figure 3.7.1.5, uniform heating of the
subsurface was achieved by Day 35 in the treatment zone at monitoring well BP-24 (located near midpoint
between injection and extraction wells). The injection pressure gradually increased from 6 to 14 psig.
NAPL arrival (floating product) at the extraction well was detected on Day 21, and the hot water broke
through on Day 27. Hot-water injection totaled 193,000 gallons. Extraction totaled 390,000 gallons with
an estimated NAPL recovery of 2,000 gallons. Polynuclear aromatic hydrocarbons such as 2-
methylnaphthalene, acenaphthalene, dibenzofuran, naphthalene and phenanthrene were also present in
the effluent.
118
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Temperature Profile, °F
Temperature Profile, °F
(a) fhj
Figure 3.7.1.4 DNAPL • ~T>?"al and corresponding temperature isotherms using hj; (a) water and (b)
surfactant solutions in 3-D tests by CROW® [Johnson and I euschner, 1992]
0
TO
1 -20
c
3
O
-25
•£ -30
CD
c
O
t-
-35
-iO
Day 35
Day 25
50 mr -
-------
-10r —•— Well IW1 (Before Injection
— •*• — CT1 (After Injection 4' Away;
----- CT2 (After Injection 10' Away)
.c
Q
0)
D
-20
8- -25
-30
-35
-40
9-Foot
Zone
4 8 12
NAPL Saturation, wt %
16
Figure 3.7.1.6 NAPL saturation profiles in soil samples (CT1.CT2) taken in vicinity of injection well (IW1)
after CROW® pilot test [Fahy et al., 1992].
is currently investigating the application of CROW® to sites contaminated with much denser DNAPLs such
as TCE and PCE.
Residual saturations (Sj of DNAPL are controlled by the NB and Nc (see section 2.0), and a
reduced residual saturation of DNAPL (Sor~) on the order of 0.1-5 wt% may pers.st even after treatmen
by the CROW®proCess, as indicated by the 1-D, 3-D and pilot study findings. The CROW® process must
thereforeT be augmented with other forms of in-situ treatment. The significant Sor reductions, toxicity
reSon bettering (increased kj, compatibility of equipment, and easily facilitated oxygenatjon and
nutrient addition make CROW® an attractive precursor to in-situ biological treatment. However, to date,
in-situ biological treatment has not been used in conjunction with the CROW® process.
With the exception of borings, CROW® is not intrusive, and there are few limitations due to
interference from ground structures, overhead or buried utilities, and other subsurface obstructions. Srte
grading is not a problem. The ex-situ hardware can be trailer-mounted and constructed of read,ly available
materials and standard unit operations equipment.
Depth of application and soil type will dictate allowable steam and hot-water injection pressures^
well spacing and thus cost [USEPA, 1991 a]. Capital cost will depend on the well spacing per unit area and
depth of application basis because the boring and well construction are the major cost items.
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Cost and Availability--
One pilot study at the site of a former wood treatment plant in Minnesota has been completed
[Johnson, 1992]. Other treatability studies are being conducted. US Patent No. 4,484,460 has been
assigned to the Western Research Institute, Laramie, Wyoming, for CROW® [Johnson and Suddeth, 1989].
Design of two full-scale CROW® applications are currently underway [Johnson, 1992]. One application is
proposed as part of the USEPA SITE Demonstration Program, at the Pennsylvania Power and Light (PP&L)
Brodhead Creek site, Stroudsburg, Pennsylvania [USEPA, 1991 a]. Approximately one-half acre of
contaminated soils will be treated to a depth of 20 ft [Johnson, 1993]. A second full scale application is
planned for the former wood treatment plant site, where the pilot study was completed [USEPA, 1991 a].
Approximately 2.7 acres of contaminated soil will be treated to a depth of 20 ft [Johnson, 1993].
CROW® appears to be a viable candidate for cleanup of DNAPL contamination in aquifer media.
CROW® will have its largest impact on DNAPL source areas. However, it is not clear that the injection of
hot water and low quality steam by the CROW® process represents a distinct thermal advantage over the
high quality steam injection of SEE.
A soil treatability study requiring a minimum of 120 Ibs (two 5 gal. containers) of soil costs
approximately $20,000 [Johnson, 1993]. This estimate excludes modifications made to the hot water such
as pH adjustment or surfactant addition. The total cost of the pilot study at the site of the former wood
treatment plant in Minnesota was approximately $300,000 and the full-scale application is anticipated to
cost $2.2 million [Johnson, 1993]. The SITE program demonstration at the Stroudsburg, Pennsylvania, site
is anticipated to cost $1.2 million [Johnson, 1993]. All of these applications will be sampling and monitoring
intensive.
These estimates exclude in-situ biological treatment. In-situ biological treatment can operate using
the CROW® hardware with minor modifications and tankage conversions. Unless additional borings are
required, operating and maintenance costs for in-situ biological treatment are anticipated to be on the order
of $50-60,000/year depending on site conditions [Leuschner, 1993].
3.7.2 Steam Enhanced Extraction (SEE)
Theoretical Background-
The Steam Enhanced Extraction (SEE) process relies on several mechanisms of contaminant
removal from the subsurface. The primary mechanisms are: vaporization of low boiling point (b.p.<100°C)
contaminants at the steam condensation front; enhancement of evaporation rates of higher boiling point
(b.p.>100°C) contaminants at the condensation front and within the steam zone; displacement of dissolved
contaminants and NAPLs by a steam condensation front and by steam within the steam zone; and
desorption from solids.
As steam progresses through cool porous media, the steam condenses and transfers its latent heat
of vaporization to the interstitial fluids and porous media. Fig. 3.7.2.1 schematically shows that continuous
steam injection results in the development of three distinct thermodynamic zones: a hot isothermal (steam)
zone at the steam temperature; a relatively sharp thermal transition zone of several centimeters thickness;
and a cool isothermal zone which represents the porous media and interstitial fluids at their ambient
temperature [Udell and Stewart, 1992]. Steam temperature depends on injection pressure, which is
governed by the depth of the injection interval, while ambient temperatures may vary between sites. The
steam condenses at the interface between the steam zone and the thermal transition zone, creating a
"steam condensation front." The growth rate of the steam zone is directly related to the injected steam
enthalpy flux if the thermal transition zone does not grow in length [Hunt et al., 1988c; Stewart and Udell,
1988].
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When gravitational forces are negligible, in unsaturated porous media, for example, results of
laboratory experiments and theoretical analyses show that in a homogeneous isotropic medium, the
orientation of the propagating condensation front will remain essentially perpendicular to the direction of
steam flow [Basel and Udell, 1989]. The same is essentially true for layered soil systems having relatively
homogeneous layers of different permeabilities. For example, the temperature profiles for a propagating
steam condensation front through layered media is illustrated in Fig. 3.7.2.2. After breakthrough of all
propagating fronts, a steady state isothermal condition is achieved; that is, a steam zone is created
between the injection and extraction wells.
100
80
60
40
20
Ambient
Zone
Transition Zone
- 15
-10.
-5 0
x - Vf( (cm)
Figure 3.7.2.1 Temperature distribution near steam condensation front [adapted from Udell and Stewart,
1989]
O 3
a 7
V 9
a. :<
O i4
Cool Region in center
indicates presence of
more impermeable lens
>
TEMPERATURE
Figure 3.7.2.2 Effect of soil heterogeneity on steam front advancement [adapted from Udell and Stewart,
1989].
122
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NAPLs with low boiling points that are contacted by steam vaporize and are completely removed
from the steam zone [Udell and Stewart, 1992]. They migrate ahead of the steam zone to the leading edge
of the thermal transition zone where they recondense to a separate phase liquid. A saturated front or
"bank" of NAPL forms [Hunt et al, 1988c] and it is displaced to the extraction point. Figure 3.7.2.3 shows
the results of a one-dimensional laboratory experiment in which gasoline was displaced from a water
saturated sand column [Hunt et al, 1988c]. As can be seen, the NAPL compounds were removed from the
column just ahead of the steam front.
NAPLs with a high boiling point that are contacted by steam undergo enhanced volatilization. This
enhanced volatilization is proportional to the ratio of NAPL vapor pressures at the steam temperature
(100°C or more depending on pressure) and 20°C and the NAPL mole fraction in the remaining
multicomponent liquid phase in the pore space. NAPL vapors also migrate to, and coalesce with, the
multicomponent NAPL bank ahead of condensation front. After steam breakthrough at the extraction point
(well), contaminant vapors are recovered in the gaseous form. This recovery mechanism has been
successfully modeled using numerical simulations [Falta et al., 1992a,b], as illustrated in Figure 3.7.2.4.
Desorption of contaminants from solids is enhanced because sufficient energy is added to the
aquifer media by steam condensation to overcome the latent heat of adsorption of many organic
contaminants and inorganics [Udell and Stewart, 1992]. Hence, partitioning to the aqueous phase is made
favorable, and subsequent vaporization of NAPLs and water and/or their displacement out of the soil matrix
lead to a net reduction of sorbed contaminants.
Another significant aspect of the process is the potential for removal of interstitial water and low
boiling point fluids in dead-end or otherwise remote (uncontacted) micro- and macropores [Udell and
Stewart, 1992]. The boiling of these fluids is achieved by vacuum drying the aquifer media after the steam
front has broken through and steam injection is discontinued [Udell et al., 1991]. As the soil matrix cools
adiabatically under vacuum, it transfers its energy to the remaining pore fluids (held by capillarity) which
boil under the applied vacuum. A net vapor flux from the remote pores towards the main flow channels
in the porous media is thus realized.
Field Implementation-
A schematic of the field implementation of SEE is presented in Figure 3.7 2.5. Once the treatment
zone (vadose and/or phreatic zone) and its areal extent is defined, a system of steam injection and vapor
and liquid condensate extraction wells is installed. In experiments to date, steam injection and extraction
wells have been constructed of low carbon steel to accommodate the operating temperatures and
pressures. Borings up to 18-inches in diameter have been used to accommodate the necessary hardware
[Udell and Stewart, 1989]. Insulation is necessary in the non-screened intervals to minimize heat losses.
The extraction well must be capable of recovering both vapors and condensed liquids. Backfill for injection
and extraction wells has consisted of pea gravel and cement in the screened and non-screened intervals,
respectively, temperature monitoring wells constructed of steel pipe which house thermocouples or other
devices have been used to monitor the progression of the steam condensation front through the
subsurface.
Normal operating practice requires that the steam be slightly supersaturated to account for thermal
losses in the manifold prior to wellhead entry. As such, one hundred percent (100%) quality steam reaches
the wellhead and is injected into subsurface soils which are initially at their ambient temperatures, usually
20°C-25°C, although other temperatures are easily accommodated [Udell and Stewart, 1989]. The steam
temperature depends on the allowable injection steam pressures. Steam injection pressures must be
selected sufficiently below the fracturing pressure of the porous media which is related to depth of
application. Near surface (up to 20 ft. depth) steam injection pressures of 6 psig. have been used at one
site resulting in steam temperatures near 100°C [Udell and Stewart, 1989].
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CD
£ 10OOOO
t/J
Z
o
m
cc
<
(J
O
1 ppm) were: BTEX
compounds (19,000 ppm), 1,2-DCB (2,900 ppm), 1,1,1 -TCA (1,700 ppm), acetone (1,650 ppm), TCE (1,600
ppm) PCE (1,400 ppm), Freon 112 (480 ppm), 2-butanone (450 ppm), methylene chloride (97 ppm), 4-
methyl-2-pentanone (4.6 ppm), and cis 1,2-DCE (2.5 ppm). Soil concentrations were reduced from 2,065
124
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HOLDING STORAGE TANK
jI HYDROCARBON VAPOPj
I I & STEAM VAPOTT~"^
Figure 3.7.2.5 Schematic of in-situ steam enhanced extraction process [USEPA, 1992d].
mg/kg to 12 mg/kg total contaminant concentrations [Udell Technologies, 1991].
The well hardware for the pilot study was installed in 18-inch borings. A cylindrical treatment cell
with an injection well encircled by six injection wells at a 5-ft. radius was employed. The injection wells
were constructed of 1-inch diameter steel pipe (for steam delivery from the manifold) which was insulated
and encased within a 6-inch diameter low carbon steel casing. The screened sections for the injection and
extraction wells were constructed of 6-inch diameter low carbon schedule 40 slot wire wrap screen. The
in-situ temperatures were monitored using 3/4-inch diameter schedule 80 steel pipe which housed
thermocouples spaced at approximately 1-foot vertical intervals.
Utilizing this design, approximately 763 Ibs of contaminants were removed following 140 hours of
steam injection cycled with vacuum extraction [Udell and Stewart, 1989]. No vapors were observed
escaping from the treatment cell. Vacuum extraction alone was responsible for 29% removal, while the
total contribution resulting from steam injection was 71%. The pilot study operational parameters were:
steam injection of approximately 250 Ib/hr at injection pressures of 6 psig, vacuum rates of approximately
25 scfm, and a well spacing of 5 feet [Udell and Stewart. 1989].
Full-scale Steam Enhanced Recovery Process (SERP) is currently underway in Huntington Beach,
California. Recovery of 135,000 gallons of diesel is being attempted at depths of 40 ft [USEPA, 1991b].
Thirty-seven steam injection wells and 39 dual vacuum extraction (vapor/liquid) wells were employed to
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remediate approximately two acres of subsurface below a truck unloading facility. Vapor flow of the
extraction system was 1200 cfm. In five months of operation, 4400 gallons of free product condensate and
vapors amounting to 14,000 gallons of product have been recovered, effectively reducing the plume to 20-
30% of its original volume [van Sickle, 1992]. No off site migration of vapors or liquids was detected, and
surface activities were not interrupted. Ex-situ treatment of recovered liquids and vapors included thermal
oxidation and waste water treatment, respectively [van Sickle, 1992].
A full-scale attempt at cleanup of a gasoline contaminated aquifer is currently in operation at
Lawrence Livermore National Laboratory in Livermore, California, at depths of 137 feet [USEPA, 1991 a].
The estimated gasoline spill size was on the order of 6200 gallons. Steam was periodically injected for
approximately two months. To date, over 7400 gallons of petroleum hydrocarbons were removed [Udell,
1993]. Three weeks after shutdown of the steam injectors, contaminant removal from the aquifer remained
on the order of 50-100 gallons/day. Preliminary estimates indicate that the contaminant removal attained
by combined steam injection\vacuum extraction is approximately forty times greater than that attainable by
conventional soil vapor extraction and ground-water pumping [Udell, 1993],
Applicability/Limitations-
SEE has successfully mobilized volatile and semi-volatile NAPLs, as well as certain inorganics, from
both the unsaturated and saturated zones through steam injection coupled with vapor and condensate
extraction in a controlled manner [Hunt et al., 1988c; Udell and Stewart, 1989, 1992]. The successful
removal of halogenated semi-volatiles such as dichlorobenzene (DCB, all isomers) is possible, and dioxins
can be mobilized by the condensation front [Udell Technologies, 1991]. Treatability and pilot studies have
documented the efficacy of SEE on multicomponent mixtures of petroleum hydrocarbons and solvents [Hunt
et al., 1988c; Udell and Stewart, 1989, 1992] in a variety of saturated and unsaturated media: sand, ash,
and silty clays with gravels.
Types of media which can be treated include in situ soils and sludges both saturated and
unsaturated. Efficacy is site specific; and in general, sites dominated by silts and clays present problems,
as is the case for all in situ technologies. Hence, impermeable layers may not be remediated to the
targeted cleanup levels. However, contaminant reduction by SEE can be expected in these areas and
heterogeneous regions by the vacuum drying mechanism [Udell and Stewart, 1992]. Because SEE utilizes
both heat and mass transfer for remediation, successful treatment is less susceptible to heterogeneities
than with other in situ technologies. One-dimensional soil column treatability studies are recommended.
To date, no applications of SEE performed in fractured rock are known to exist.
It is expected that shallow applications of SEE will not result in the sterilization of aquifer media and
that microorganisms will persist in a dormant state [Alvarez-Cohen, 1993b]. Upon cooling of the porous
media, they are expected to flourish: thus, thermally enhanced in-situ biodegradation is anticipated as a
secondary benefit of SEE. However, deeper applications of SEE which require greater injection pressures
may result in the complete sterilization of aquifer media, and in-situ biodegradation is only anticipated to
occur after re-acclimation and repopulation [Alvarez-Cohen, 1992]. Therefore, no biofouling is expected.
Also, clogging of porous media is not likely to result from precipitation of inorganic compounds, as
described earlier.
With the exception of borings, SEE is not intrusive; and, there are few limitations due to
interference from ground structures, overhead or buried utilities, and other subsurface obstructions. Site
grading is not a problem. The ex-situ hardware can be trailer-mounted and constructed ot readily available
materials and standard unit operations equipment
Depth of application and soil type will dictate allowable steam injection pressures, well spacing, and
thus cost [USEPA, 1991b]. Capital cost will depend on well spacing per unit area and depth of application
126
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basis because the boring and well construction are the major cost items. For these reasons, SEE is very
attractive for use in the most contaminated hot spots for rapid pure product recovery.
Cost and Availability--
US Patent No. 5,018,576, has been awarded to the University of California for SEE [Udell et al.,
1991], and US Patent No. 5,009,266 for a similar technology is held by USPCI, a subsidiary of the Union
Pacific Corporation.
One pilot study involving dense immiscible hydrocarbons in the vadose zone has been completed
[Udell and Stewart, 1989]. Pilot-scale treatability studies require a minimum of 50 yd3 of soil [Udell
Technologies, 1991]. A demonstration of SEE is proposed as part of the SITE Demonstration Program at
McClellan AFB near Sacramento, California [USEPA, 1991 a,b]. Two full-scale SEE applications for
remediation of gasoline contaminated soils are currently under way [USEPA, 1991a,b]. No other domestic
full-scale applications are known to exist at this time
SEE can operate both above and below the water table therefore making it a good candidate for
cleanup of both LNAPLs and DNAPLs. It will have its biggest impact on NAPL source areas. Studies show
that it can enhance contaminant removal from low permeability zones It has not been applied to fractured
media at this time.
SEE is anticipated to cost about $50-125/yd3 depending on site characteristics [Udell Technologies,
1991]. Included in this estimate is the treatment of the waste streams emanating from the recovery wells:
condensible and non-condensible gases, and extraction pump liquids. Condensible gases and extraction
pump liquids are concentrated for recycling or destruction by separations equipment Non-condensible
gases are collected and treated by granular activated carbon units.
The factors cited to most influence the overall treatment cost are: areal extent of treatment, depth
of contamination, waste quantity and targeted cleanup goal, site preparation, ongoing surface activities and
waste handling [Udell Technologies, 1991].
3.7.3 Radio Frequency Heating
Theoretical Background-
Radio Frequency (RF) heating is an enhanced oil recovery process which uses electromagnetic
energy to accomplish subsurface heating, thereby enhancing contaminant removal. The primary removal
mechanisms, which depend on the actual heating strategy, are: vaporization of low boiling point
(b.p.<100°C) organic compounds and water; enhancement of evaporation rates of higher boiling point
(b.p.>100°C) organic compounds; partial or complete displacement of heated pore fluids by a propagating
steam condensation front partial or complete displacement of all contactable NAPLs by the propagating
steam front, and/or enhanced pore liquid mobilization resulting from liquid density and viscosity alterations
(increased capillary numbers). The flexibility of applied temperatures and geometries allows this technology
to potentially operate analogously to either CROW® or SEE (section 3.7.1 and 3.7.2, respectively), or it
may be used in conjunction with, or as a precursor to, soil washing processes (section 3.5)
The focus here is on the actual heating mechanisms; ohmic and dielectric heating of pore fluids
[Dev et al., 1988]. Wave frequencies in the range of 6.78 MHz to 2 45 GHz are used to achieve bulk
volumetric heating of the pore fluids and porous media [Dev et al , 1987], Ohmic heating results from ionic
or conduction current flow through the porous media. Dielectric heating refers to the mechanism by which
electromagnetic energy is converted into thermal energy. In this process, agitation and physical distortion
of the molecular structure of polar compounds (i.e., water), initiated by an applied alternating AC electric
field, result in increased kinetic activity and thus heating. Important parameters governing the success of
127
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dielectric heating are the relative dielectric constant of the porous medium and its loss-tangent, where the
loss tangent is defined as the ratio of the apparent conductivity to the frequency, relative dielectric constant,
and permitivity of a vacuum [Dev et al. 1988]. The dielectric properties of porous media decrease with
increasing applied frequencies [von Hippl, 1954].
The amount of RF power adsorbed is directly proportional to the frequency of the applied electric
field, the square of the amplitude, the relative dielectric constant, and the loss tangent [Dev et al., 1988].
Medium breakdown, corona discharge and ohmic heating result with increased power and thereby limit the
useful transmitter output [Smith and Hinchee, 1993]. The effective penetration depth, or "skin depth," of
the electromagnetic energy is defined as the distance at which the wave amplitude decreases to 37% of
its initial value [Dev et al., 1987]. The skin depth decreases with increasing loss tangent, relative dielectric
constant (er) and apparent conductivity [Morey, 1974; Johnson and Sitar, 1986]. The er for dry soils varies
between 2 and 10, and typical values for different materials are available [von Hippl, 1954; Morey, 1974:
Okrasinski et al., 1979]. Since ewater~80 is very high, RF adsorption in highly water saturated porous media
is large, making the penetration depth small. Thus, water immediately adjacent to the RF electrodes is
converted to steam which may assist in NAPL recovery, and in this way a steam front can be generated.
The impact of water on er and the loss tangent, even at low saturations, can be seen in Figure
3.7.3.1 which shows the relationship between loss tangent, relative dielectric constant and soil temperature
for a sample of Utah tar sand. Although the water content of the tar sand was not reported, ers between
4 and 8 suggest that it must be very low, considering eNAPL (see Table 2.1) is usually below 5 and ewater
is about 80. In Figure 3.7.3.1, er increases from an initial value of 6 to a peak value of 8. As water is
boiled off (T~100°C), er and the loss tangent are abruptly reduced to a final value of 4 and 0.1, respectively.
Thus, as water is boiled off near the RF electrode, the corresponding decrease in loss tangent indicates
that the skin depth effectively increases. However, the drop in loss tangent affects the efficiency of
coupling between the RF field and porous media. This represents a major challenge in RF system design.
Coupling may be maintained by changing transmission frequency and/or electrical properties of the network
[Smith and Hinchee, 1993].
Field Implementation-
A schematic of a field implementation is shown in Figure 3.7.3.2. Horizontal well configurations
are also possible [Sresty et al., 1986]. Since the generated EM waves used for soil heating can interfere
with communications and navigation equipment as well as pose a human health threat, any RF application
must be designed to effectively contain the EM radiation within the specified soil treatment zone. Triplate
line and fringing-field transmission line arrays are therefore employed [Dev et al., 1988]. The triplate line
array system, as shown in Figure 3.7.3.2, will be described here since it has been used in an enhanced
oil recovery and environmental application. The fringing-field transmission line configuration is described
elsewhere [Dev et al., 1987].
The triplate line configuration is an electromagnetic analogy to a central conductor enclosed
between two parallel plates. Tubular electrodes are arranged in three parallel rows. Whereas the
frequency of operation is governed by the dielectric properties of the porous media and treatment zone
size, the geometry and spacing between rows is governed by the thickness of the treatment zone (or
deposit), heating rate and final heating temperature. In enhanced oil recovery applications, the spacing
between rows has been taken to be less than the deposit thickness [Sresty et al., 1986]. The spacing of
electrodes in each row is generally somewhat smaller than the row spacing. Electrodes can be constructed
of thin-walled pipe, copper, steel or aluminum tubing which can be perforated to accommodate vapor flow
[Sresty et al., 1992a]. In large applications, use of low cost materials such as aluminum permits electrode
abandonment.
128
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« c«
a>
V)
O 0 _!
_)
o :
o i
O
o
IfcO 180 HO
Temperature, °C
Figure 3.7.3.1 Loss tangent and dielectnu constant of tar sand samples
restyetal., 1986].
- Transition section
- RF power fee<3 point
Vapor bamer
Concrete pac
Figure 3.7.3.2 Schematic of radio-frequency soil heating process showing electromagnetic electrode array
and vacuum hood [Dev and Downey, 1988].
As the subsurface is heated, water and contaminant vapors flow to the ground surface or the
nearest perforated electrode. For applications installed at the ground surface, Halon® tracer experiments
have confirmed that the induced vapor flow produces a draft into the soil treatment zone from the adjacent
porous media [Dev and Downey, 1988; Dev et al., 1988]. Depending on the initial water content and soil
temperature, steam and/or distilled vapor fronts can be established in-situ which aid in recovery (see
section 3.7.2 for details on steam recovery mechanisms). Vapors are recovered by a vapor collection
manifold/impermeable barrier system situated at the ground surface. Subsurface heating also promotes
gravity segregation of pore fluids because of density and viscosity alterations. In fact, gravity drainage of
viscous bitumen from tar sands is economically favorable [Sresty et al., 1986].
When the soil temperature is below 100°C, hot-water displacement of NAPLs may assist in NAPL
recovery. Any combination of compounds such as alkaline agents, surfactant, polymers (density
129
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enhancement) and viscosifiers may be introduced into the treatment zone to ensure a more favorable
mobility ratio between the hot water and the NAPL or to enhance compound solubilities.
Level of Demonstration and Performance--
Recovery of bitumen from a tar sand deposit was conducted in small-scale field experiments at the
Asphalt Ridge deposit near Vernal, Utah [Sresty et al., 1986]. On a surface outcrop, an RF electrode array
was installed vertically into the tar sand deposit. Mobilized bitumen was allowed to drain by gravity into
a collection gallery within a pre-existing mine shaft situated directly below the RF array. Approximately 33
cu yd (25 nrr) of tar sands were heated by 20 ft long electrodes arranged in a triplate geometry. The daily
power input varied from 40 to 75 kW, and soil temperatures were increased to approximately 200°C. RF
power was terminated after 20 days. Total bitumen production from this experiment was approximately 336
gals, or 35% recovery. Redistribution of bitumen as a result of gravity drainage was clearly evident, even
though bitumen viscosity at 100-150°C was on the order of 20-50 cp [Sresty et al., 1986].
The fire training pit area at the Volk Field Air National Guard Base in Wisconsin is the site of the
only known environmental application of RF heating [Dev and Downey 1988; Dev et al., 1988; Sresty et
al., 1992a]. The sandy subsurface soils have been contaminated with numerous waste oils, fuels and
solvents yielding hydrocarbon concentrations up to 4,000 ppm. Depth to the water table is approximately
12 ft. The moisture content profile for the vadose zone was not provided.
A triplate electrode array using perforated electrodes was installed within a 6 x 12 ft test cell area.
Each row consisted of 13 electrodes using a 1 ft spacing for a total row length of 12 ft. Spacing between
adjacent rows was 3 ft. The center and outer rows were installed in the vadose zone to depths of 6 and
8 ft, respectively. Thermocouples were attached to the inner sides of the electrodes to monitor soil heating.
Fluid-filled thermowells were installed to measure soil temperatures between electrode rows. At the ground
surface, a concrete frame was constructed around the test cell area and an impermeable silicon rubber
sheet was bonded to it to form the vapor barrier. Below the vapor barrier, two perforated vapor extraction
tubes were installed at the ground surface through which a low vacuum equivalent to 6 in H2O was applied.
A daily power input of 35 kW, at 6.78 MHz, was used for the first four days of the test to vaporize
the porewater. This input was reduced to 20 kW for the remainder of the test. Soil temperatures exceeded
100°C within two days of soil heating. After day 8, the soil temperatures reached 150-160°C in the center
of the test cell, and were maintained at that temperature for four days. Soil temperatures along the test
cell periphery averaged about 100°C owing to heat conduction losses to the surrounding porous media.
The test duration was 12.5 days. Since effluent concentrations were not measured and contaminant
influxes into the test were not estimated, a mass balance calculation was precluded. Contaminant removal
was based on soil samples.
Soil samples were taken after the test was shut down for 17 days, at which time the soil had cooled
to 50-60°C [Dev et al., 1988]. Soil analyses revealed that the removal rate for volatiles (b.p.<120°C) was
99.6% and 99.3% for aromatic and aliphatics, respectively. The removal rate for semi-volatiles
(120°C
-------
The soil water content greatly affects the viability of RF heating because of the large dielectric
constant of water. The initial water contents of the samples tested were on the order of 5 to 12 wt%. It
is therefore difficult to ascertain whether RF heating is a viable technology for environmental applications
in the saturated zone: and if it is, what are its possible advantages over SEE are (section 3.7.2). Radio
frequency heating equipment requires a high degree of sophistication to implement and operate [Smith and
Hinchee, 1993].
Bench and pilot treatability studies simulating in-situ heating (nitrogen/steam injection) have
indicated that the presence of clay minerals did not significantly affect contaminant removal [Sresty et al.,
1992a]. Although the RF process was not used, this result seems surprising in light of the low permeability
of clays and their high moisture contents. Furthermore, field scale stratigraphy with low permeability lenses
has been observed to not only affect steam front propagation, but also contaminant removal [Udell and
Stewart, 1989, 1992; Ho and Udell, 1992]. The affect of soil stratigraphy on the actual RF heating process
has yet to be evaluated. The high temperatures employed by RF heating may inhibit and destroy
indigenous microorganisms, and could have an adverse impact on the humic matter in soil [Smith and
Hinchee, 1993]
With the exception of borings, the RF heating process is not intrusive; therefore, there are few
limitations with respect to interference from ground structures and overhead utilities. Site grading is not
a problem. Subsurface obstructions such as buried utilities, abandoned foundations, etc., appearing within
the RF treatment zone may reduce the effectiveness of the process or potentially cause leakage of EM
energy. Leakage of EM energy is a concern because of its interference with communications and
navigation equipment and human health. The ex-situ hardware can be trailer-mounted and constructed
of readily available materials and standard unit operations equipment.
Cost and Availability-
This technology is commercially available through the Illinois Institute of Technology Research
Institute (IITRI). The process is reported to have been patented by IITRI, and is exclusively licensed to Roy
F. Weston, Inc. [Roy F. Weston, Inc., 1989]. Two field scale studies have been completed, and others are
planned [Sresty et al., 1992b]. However, no application has been completed in the saturated zone or
specifically on DNAPLs.
Radio frequency heating technology appears to hold promise for cleanup of dissolved contaminants
and DNAPLs, but many issues are unresolved at this time. Radio frequency heating is anticipated to have
its biggest impact on DNAPL source areas. However, since the objective is to heat the subsurface and
to generate a sweep front, it is not clear whether RF heating has any distinct advantage over steam
injection.
The cost of RF heating is estimated to be on the order of $40-100/ton of soil depending on soil
moisture content, and final treatment temperature [Sresty et al, 1992a]. This estimate is based on a
maximum soil moisture content of approximately 20%. Residuals produced by this process include vapors,
and steam and vapor condensates. Gas vapors can undergo cooling to condense out low boiling point
NAPLs which can be followed by carbon treatment polishing. Condensed liquids can be
reclaimed/recycled.
3.7.4 Vitrification
Theoretical Background-
In-situ vitrification (ISV) is a process that relies on joule resistance heating and consequent melting
of the contaminated zone to enhance organic contaminant removal. The primary mechanisms are:
accelerated chemical reactions in the soil surrounding the melt and the pyrolysis zone (thermal zone
131
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adjacent to the melt); recovery of organic vapors in a vacuum hood situated above the soil treatment zone;
pyrolysis of DNAPLs in the melt and pyrolysis zones; and pyrolysis of combustible vapors in the vacuum
hood [Dragun, 1991].
Joule resistance heating of the soil results from electric current flow through the porous media
[Fitzpatrick et al., 1984]. Since the soil becomes more conductive upon melting, the voltage and current
requirements must be adjusted to maintain the same power delivery. To initiate the soil melt, a conductive
mixture of graphite and glass frit is placed on the ground surface between the electrodes to serve as an
initial conductive (starter) path, see Figure 3.7.4.1 [Hansen and Fitzpatrick, 1991]. An electric potential is
applied, and the high current flowing through the graphite and glass frit causes them to melt. Heat is
transferred by conduction from the molten mass to the surrounding soils which causes boiling of pore fluids
and, ultimately, melting of the soil. Soil temperatures of 1,600 to 2,000 °C can be achieved [Hansen and
Fitzpatrick, 1991]. Upon melting, the adjacent soils become electrically conductive. In this way, the
process is continued beyond startup and the molten mass grows vertically downward at an approximate
rate of 1-2 in/hr to the desired treatment depth [Fitzpatrick et al., 1984]. The chemical reactions and
processes which are thought to occur within the soil melt and the adjacent soils are depicted in Figure
3.7.4.2.
The propagating soil-melt interface is preceded by a transition zone, dry zone, and pyrolysis zone,
respectively. The total combined thickness of these zones is approximately 9-12 inches [Hansen, 1993].
Within the transition zone (-25-100°C), enhanced vaporization of soil moisture and DNAPLs occurs. When
soil moisture and low boiling point (b.p.<100°C) DNAPLs are reached by the propagating 100°C isotherm,
they are boiled in-situ. Enhanced volatilization of high boiling point (b.p.>100°C) DNAPLs occurs in the
transition zone and dry zone (~100-400°C) until the soil temperature reaches the boiling point of the
compound. The dry zone is narrow and is reported to have temperature gradients on the order of 150-
250°C/in [Hansen and Fitzpatrick, 1991]. Since the liquid saturation of the dry zone is low and its gas
permeability is high, the dry zone is thought to serve as a conduit for water and organic vapor flow to the
ground surface. Other vapors may reach the ground surface through the perforated electrodes. Within
the oxygen-deficient pyrolysis zone and reducing environment of the high temperature melt zone, any
remaining high boiling point compounds are thermally decomposed. Convective currents within the molten
soils mass cause it to have a uniform chemical composition [USEPA, 1988; Dragun, 1991]. Because of
convective mixing, pyrolysis byproducts can also reach the ground surface. As vapor, gases and other
organic pyrolysis products escape from the treatment zone at the ground surface, they are captured in the
vacuum hood. Flammable pyrolysis products encounter oxygen in the vacuum hood and are combusted.
All emissions are treated by an off-gas treatment system.
Elevated temperatures accelerate a variety of reactions between the organic compounds, soil
moisture, and mineral surfaces of the porous media. Dragun (1991) enumerates that hydrolysis,
substitution, oxidation, reduction, and surface-catalyzed reaction rates may be increased in the affected soil
zones.
Upon cooling, the vitrified soil mass resembles obsidian. This amorphous material has a strength
5-10 times that of concrete, high leach resistance [Buelt and Westsik, undated], and its durability is similar
to that of granite [EPRI, 1988].
Field Implementation-
A schematic of a typical field implementation of ISV is shown in Figure 3.7.4.1. The objective of
the ISV process scheme is to create a molten soil mass in which heavy metals and radionuclides are
stabilized and DNAPLs are pyrolyzed. The strategy implemented to accomplish in-situ heating is
summarized below.
132
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BacWill
Figure 3.7.4.1 Schematic illustrating the in-situ vitrification (ISV) process [Smith and Hinchee, 1993].
MELT ZONE
>1700°C
PYROLYS1S ZONE
-400-C
HLAT AH-ECTED (OR DRY) ZONK
100-C ISOTHERM
IRANSniOS 7OM
AMBIENT
SOIL ZONF
Figure 3.7.4.2 Chemical processes and reactions occurring within and near the soil melt zone [Dragun,
1991].
133
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Electrodes have been constructed of molybdenum cores and perforated graphite collars [Hansen
and Fitzpatrick, 1991]. Electrodes can either be fixed, that is, installed to a specified depth, or can be
moveable, which allows them to be lowered as the soil melt progresses downwards. A 50 ft diameter
vacuum hood is placed over an electrode array to capture fugitive emissions emanating from the heated
soil zone. Under the hood, square electrode arrays having a maximum width of 35 ft are possible.
Electrode installation depths have exceeded 19 ft, and depths as great as 25 ft are considered possible
[USEPA, 1992d]. Soil has been treated to a depth of 24 ft [Smith and Hinchee, 1993]. The applied
vacuum of the off-gas treatment system is on the order of 0.5 to 1.0 in H20.
After treatment is complete, subsidence occurring at the ground surface can be back-filled. The
process can be sequentially applied and adjacent melts can be fused together [Hansen and Fitzpatrick,
1991]. Mobilization between treatment zone locations requires approximately 16 hours [Dragun, 1991].
Level of Demonstration and Performance-
Since 1980, 20 pilot-scale and 6 large-scale ISV tests processing 10-50 and 400-800 tons of
contaminated soil, respectively, have been completed using metals and simulated radionuclides [Dragun,
1991]. ISV has been tested at several private, Superfund, and U.S. Department of Energy (DOE) sites
[USEPA, 1992d]. None of these applications have included organic compounds. Testing with NAPLs has
occurred on the engineering test-scale using soil quantities of 1 ton or less [Hansen, 1993]. Table 3.7.4.1
illustrates typical removal efficiencies for porous media containing several organic compounds.
Applicability/Limitations-
The ISV process was originally developed for sites contaminated with heavy metals and
radionuclides [USEPA, 1988] where there are very few options for treatment and the treatment costs are
enormous. The ISV process can theoretically destroy DNAPLs by pyrolysis and it has been demonstrated
in small-scale tests [Hansen and Fitzpatrick, 1989, 1991]. Therefore, ISV is a potentially attractive
treatment alternative for sites containing mixed inorganic and organic wastes.
A limitation of the technology is that ground-water recharge in permeable soils with hydraulic
conductivities greater than 10"4 cm/swill stop the progress of the melt [USEPA, 1988; Hansen 1993]. This
makes the ISV process essentially applicable to treatment of vadose soils only, unless dewatering,
containment, or other hydraulic controls are engaged to minimize ground-water recharge into the treatment
zone [Hansen and Fitzpatrick, 1989]. While the process may be applicable to fine grained saturated soils
such as clays because of their low hydraulic conductivity [Hansen, 1992] and higher relative electrical
conductivity than coarse soils [McElroy, 1993], the amount of DNAPL contained within clay soils is likely
to be low compared to that of coarser media.
Subsurface obstructions and features can interfere with the operational efficiency of the ISV
process. While ISV can accommodate a very heterogenous subsurface, several rule-of-thumb limitations
apply: general metals concentration of 5-16 wt%; no continuous metal traversing a distance greater than
90% of the electrode spacing; combustible organic concentrations limits of 5-10 wt%; rubble limit of 20 wt%;
must have sufficient glass forming minerals (usually not a problem for soils); and individual void volumes
less than 150 ft3 [Hansen and Fitzpatrick, 1989, 1991; USEPA, 1992d; Smith and Hinchee, 1993]. Buried
drums, crates and cartons containing wastes may pose additional problems [Hansen and Fitzpatrick, 1991].
High concentrations of iron or other dense metals may result in its pooling near the bottom of the melt and
current short circuiting [Hansen and Fitzpatrick, 1989]. Concentrated vapor loading of pyrolyzed organics
from DNAPL pools may potentially overload the off-gas treatment system.
The ISV process produces a solidified soil mass, and the subsurface is essentially unusable once
the process is complete. However, the vitrified mass can be broken and moved. Light structures or
vegetation may be supported on the backfill materials, but there is usually little incentive to reuse land
134
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contaminated by mixed wastes containing radionuclides [Hansen, 1993]. Since vitrification results in the
collapse of the pore volume and removal of combustible organic materials, soil volume reduction on the
TABLE 3.7.4.1 TYPICAL ORGANIC DESTRUCTION/REMOVAL EFFICIENCIES BY ISV [Hansen and
Fitzpatrick, 1991].
Contaminant
PESTICIDES
4, 4 DDD/DDE/DDT
Aldrin
Chlordane
Dieldrin
Heptachlor
VOLATILES
Fuel Oil
MEK
Toluene
Trichloroethane
Xylenes
SEMI-VOLATILES
PCP
NON-VOLATILES
Glycol
PCBs
Dioxms
Furans
Concentration
(ppb)
21-240,000
113
535,000
24,000
61
230-110,000
6,000 (2)
203,000
106,000
3,533,000
>4,000,000
8,000 (3)
19,400,000
>47,000
>9,400
Percent
Destruction
99 9-99.99
>97
99.95
98-99.9
98.7
>99
>99
99.996
99995
99.998
99995
>98
99 9-99.99
99.9-99.99
99.9-99.99
Percent
Removal (1)
>999
>99.9
>99.9
>999
>99.9
>99.9
>999
>999
>99.9
>99.9
>999
>99.9
>999
>999
>99.9
Total
ORE {%)
99.9999
99.99
99 9999
99.99
9999
99999
99999
99.99999
99.99999
99 99999
99 99999
99.99
99 9999
99 9999
99.9999
(1) Percent removed from off-gas after destruction; percentages are additive for the
total ORE.
(2) 98% MEK in container, yielding 6,000 ppm in layer of container thickness
(3) 50% ethylene glycol in container, yielding 8,000 ppm in layer of container thickness
order of 25-45% is possible. In one application, 4 ft of subsidence was observed [Geosafe Corp., 1992].
The vacuum hood requires a side clearance of 15-20 ft [Hansen and Fitzpatrick, 1991],
The ex-situ hardware is trailer-mounted. The large ISV equipment can treat soil at a rate of
approximately 10,000 Ibs/hr [Hansen and Fitzpatrick, 1989]. Since typical soil applications require 0.35-0.4
kilowatt hrs/lb, 4,000 kilowatts per application are required. This is comparable to the daily energy
consumption of an average-size hotel in a major city [Hansen and Fitzpatrick, 1989]. However, because
of the efficiency of soil heating, ISV consumes less than one-third the energy of an incinerator [Hansen,
1993].
Cost and Availability-
No field- or full-scale ISV applications exist involving DNAPLs situated below the water table. The
U.S. Department of Energy has been awarded the patent (Patent No. 4,376,598) for the ISV process
135
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[Fitzpatrick et al., 1984]. DOE has licensed the technology to Battelle, which created the Geosafe
Corporation (Kirkland, Washington) and has exclusively sublicensed the ISV technology to Geosafe for
commercialization purposes [Hansen and Fitzpatrick, 1989]. Geosafe is participating in the SITE Program;
and the Parsons/ETM site in Grand Ledge, Michigan, has been selected for the demonstration [USEPA,
1992d]. The ISV process has been selected as a preferred technology at ten other sites (private,
Superfund and DOE sites) [USEPA, 1992d].
The ISV process is not a viable candidate for in-situ cleanup of DNAPLs below the water table
because the presence of water will stop the progression of the melt unless ground-water recharge is cut-off.
Other methods should therefore be sought. However, for mixed wastes containing radionuclides and
DNAPLs, few remedial alternatives exist and dewatering or impermeable barrier construction may be
warranted.
The cost of a treatability study to determine the viability of ISV is on the order of $35-40,000, and
could be more depending on unusual analytical requirements [Hansen and Fitzpatrick, 1989]. Process cost
estimates which exclude mobilization costs ($125-200,000) and sampling costs ($50-80,000) have been
provided [Fitzpatrick et al., 1984; Hansen and Fitzpatrick, 1991]. Major factors affecting the process cost
of ISV include: cost of electrical power; initial moisture content of soil and recharged water to be removed
during the ISV process; depth of treatment; and analytical requirements associated with process control
and permit compliance [Hansen and Fitzpatrick, 1991]. The dependency of process cost on moisture
content and electrical rates using 1982 dollars is shown in Figure 3.7.4.3. For a vadose soil containing
25% moisture, Figure 3.7.4.3 predicts that the maximum total cost will approach approximately $450/yd
[Fitzpatrick et al., 1984]. More current estimates place typical ISV process costs at $300-400/ton using a
non-specific moisture content basis [Hansen and Fitzpatrick, 1991].
200
u
0
468
Eloctncal Rates (
-------
SECTION 4.0
IN-SITU TECHNOLOGY COMPARISONS
4.1 INTRODUCTION
In the previous sections of this report, the factors controlling the fate and transport of DNAPLs have
been reviewed, and the various technologies with a potential for application in remediation of DNAPL
contamination have been described. The purpose of this section is to provide side-by-side comparisons
and an overview of the technologies reviewed in the report. To facilitate the overview, Table 4.1.1. lists
the major characteristic features of each technology. The entries in the table are based on the essential
features of each technology, and on judgement and interpretations made on the basis of the available
information.
4.2 EXPLANATION OF TERMS
The main factors used to rate each technology are as follows: Design Basis; Operational
Mechanism; Applicability; Scale of Demonstration; Expected Efficiency, Commercial Availability; and
Approximate Cost Range.
Design Basis-
This aspect of the technology is classified as either theoretical or empirical. The
distinction is made on the basis of how well the specific application is researched and
elucidated, and how much of the implementation is theoretically based versus common
sense and field experience. Though a technology is designated as "theoretical," this does
not imply that its implementation will be more successful than an "empirical" one, even if
designed by qualified professionals, because of site specific considerations.
Operation Mechanism-
This heading relates to the intended use and purpose of the technology and is
defined as either: treatment (i.e., degradation, destruction); recovery (i.e., enhanced
solubility, mobilization, volatilization, coupled with recovery); or containment (i.e.,
immobilization, isolation). Some technologies have multiple capabilities, but the major
emphasis is indicated.
Applicability-
Applicability refers to the type of contamination for which the technology is suitable,
and the terms used in Table 4.1.1 are self explanatory. For limitations and problems
associated with each specific technology, the reader is referred to the specific technology
descriptions (section 3.0). These issues could not be conveniently and briefly summarized
within Table 4.1.1.
Scale of Demonstration-
This indicates the most current testing or demonstration level of the technology as
it specifically pertains to environmental applications. For example, electro-osmosis has
been implemented on the full-scale in geotechnical engineering applications for several
137
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decades, but specific environmental applications are still at the pilot scale. Similarly, soil
washing (flooding) using alkalis, cosolvents and surfactants and combinations of these
compounds has been demonstrated in the petroleum engineering field, yet specific
applications to environmental problems and conditions are still at the emerging and pilot
scales.
Expected Efficiency--
Under this heading the anticipated efficiency is rated at the full scale solely within
the context of the technology's operational mechanisms and its applicability. By rating the
expected efficiency of slurry walls as "high," the proper interpretation is that for their
intended purpose, that is containment of the dissolved and separate phases, slurry walls
usually perform extremely well. As an additional example, it is not implied that air sparging
is capable of recovering separate phase DNAPLs efficiently.
Commercial Availability--
Degree of availability indicated. Many of the technologies and/or the components
comprising and utilized by the technologies were formerly established in other disciplines
and are presently available. A few technologies have completed pilot scale testing, and
are ready for scale up to full scale and/or have been previously implemented in other
disciplines (e.g., chemical flooding).
Approximate Cost Range-
Estimated costs in $/yd3 have been developed to provide a benchmark of
anticipated costs and to provide relative comparisons between technologies. Where costs
in section 3.0 were reported on a $/ton basis, a soil unit weight of 120 Ib/ft3 was used as
a conversion factor. While technologies such as slurry walls and hydraulic gradient control
can provide containment relatively inexpensively, it should be realized that the majority of
the DNAPL remains in the subsurface.
4.3 PROMISING TECHNOLOGIES
As already indicated in Section 2.7, the remediation of DNAPLs faces a number of challenges
posed by the site stratigraphy and heterogeneity, the distribution of the contamination, and the physical and
chemical properties of the DNAPL. Thus, a successful technology has to be able to overcome the
problems posed by the site complexity and be able to appropriately modify the properties of the DNAPL
to facilitate recovery, immobilization, or degradation. In addition, the methodology has to be adaptable to
different site conditions and has to be able to meet the regulatory goals. There are several ways in which
to define a "promising technology." A promising technology for the purposes of this report is defined as
a technology that is capable of effectively treating or recovering the DNAPL from the source areas, lenses
and pools, and residually contaminated zones.
Because the thermally based technologies represent perhaps the largest thermodynamic
perturbation to the subsurface system, they are among the most promising. Among thermal technologies,
steam enhanced extraction (SEE) is probably the most promising candidate. The CROW® process relies
on similar mechanisms, however, it is not clear whether the injection of hot water and low quality steam
offers an advantage over SEE. Radio frequency heating, which relies on in-situ steam generation to be
most effective, has only been tested in the vadose zone.
138
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-------
The next group of promising technologies are the soil washing technologies because they can
manipulate chemical equilibria and reduce capillary forces. A blend of alkalis, cosolvents, and surfactants
is probably the best combination for a soil washing application, and each is important for its own reasons:
alkalis can saponify certain DNAPLs and affect wettability and sorption; cosolvents provide viscous stability,
and enhance solubility and mass transfer between the aqueous phase and the DNAPL; and surfactants
have the largest impacts on solubility and interfacial tension reduction. Water flooding is best applied in
highly contaminated areas (source areas) as a precursor to these methods. Of course, the exact approach
will depend on site specific conditions.
The thermal and soil washing technologies are best suited for areas that are highly contaminated
with DNAPLs. However, even under the best conditions, these techniques by themselves still may not be
able to achieve the currently mandated regulatory cleanup standards. Thus, consideration should be given
to using these technologies in combination with the technologies suitable for long-term plume management.
In particular, the bioremediation techniques and permeable treatment walls seem to hold the best promise,
although any of the remaining technologies in Table 4.1.1 are capable of dissolved plume management,
and each has its own niche, depending on site specific considerations.
A special problem is posed by mixed wastes, heavy metals and radionuclides mixed with DNAPLs,
since recovery at the ground surface may not be desirable in many instances. In such instances,
stabilization/solidification and vitrification currently appear to be the most viable in-situ technologies.
Excluding radionuclides, in-situ S/S is the most promising candidate because of its broadly demonstrated
effectiveness, cost, and applicability to the saturated zone.
140
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FIGURES
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Fig. 3.5.2.7 Reprinted with permission of the Society of Petroleum Engineers. Habermann, B.,
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Fig. 3.5.2.8 Reprinted with permission of the Society of Petroleum Engineers. Blackwell, R., J.
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Produce Ultralow Interfacial Tension at the Oil/Brine Interface, in Surface
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Materials, San Francisco, CA, 1992.
Fig. 3.6.1.4 Reprinted with permission from Hazardous Materials Control Research Institute.
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HMCRI's National Research & Development Conference on the Control of Hazardous
Materials, San Francisco, CA, 1992.
Fig. 3.6.2.3 Reprinted from: Weyer, Udo (ed.), Subsurface contamination in immiscible fluids -
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Vermont 05036 (Telephone: 802-276-3162; telefax: 802-276-3837).
171
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Fig. 3.6.2.4 Reprinted from: Weyer, Udo (ed.), Subsurface contamination in immiscible fluids -
Proceedings of a symposium, Calgary, Alberta, 18-20 April 1990. 1993 588 pp., Hfl
195/USS115.00. Please order from: A.A. Balkema, Old Post Road, Brookfield,
Vermont 05036 (Telephone: 802-276-3162; telefax: 802-276-3837).
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Fig. 3.7.1.4 From Johnson, L., Leuschner, A., The CROW® process and bioremediation for in situ
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Groundwater, Calabrese, E., Kostecki, P., Eds., Lewis Publishers, a subsidiary of CRC
Press, Boca Raton, FL, 1992. With permission.
Fig. 3.7.1.5 Reprinted with permission from Western Research Institute. Fahy, L., L. Johnson,
D. Sola, S. Horn, and J. Christofferson, Bell pole CROW® pilot test results and
evaluation (Presented at Colorado HWMS Annual Conference), Denver, CO, October,
1992.
Fig. 3.7.1.6 Reprinted with permission from Western Research Institute. Fahy, L., L. Johnson,
D. Sola, S. Horn, and J. Christofferson, Bell pole CROW® pilot test results and
evaluation (Presented at Colorado HWMS Annual Conference), Denver, CO, October,
1992.
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by the American Geophysical Union.
Fig. 3.7.2.4 Falta, R., K. Pruess, I. Javandel, and P. Witherspoon, Numerical modeling of steam
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Fig. 3.7.3.1 Reprinted with permission of the Society of Petroleum Engineers. Sresty, G., H. Dev,
R. Snow, and J. Bridges, Recovery of bitumen from tar sand deposits with the radio
frequency process. SPE Reservoir Engineering, 1, 85-93, 1986.
Fig. 3.7.3.2 Reprinted from Zapping Hazwastes, H. Dev and D. Downey, Civil Engineering,
August, 1988 with permission of ASCE, 1994.
172
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Fig. 3.7.4.2 Reprinted with permission from Elsevier Science Publishers BV, Academic Publishing
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TABLES
Table 2.2
Table 2.4
Table 3.2.2
Table 3.3.1
Table 3.6.1.1
Reprinted with permission from Elsevier Science Publishers BV, Academic
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the subsurface: Properties, models, characterization and remediation. J. Contam.
Hydro!., 6, 107-163, 1990.
Reprinted from: Weyer, Udo (ed.), Subsurface contamination in immiscible fluids
- Proceedings of a symposium, Calgary, Alberta, 18-20 April 1990. 1993588pp.,
Hfl 195/USS115.00. Please order from: A.A. Balkema, Old Post Road,
Brookfield, Vermont 05036 (Telephone: 802-276-3162; telefax: 802-276-3837).
Criddle, C., L. Alvarez, and P. McCarty, Microbiological processes in porous
media, in Transport Processes in Porous Media, edited by J. Bear and M.
Corapcioglu, pp. 639-691, Kluwer Academic Publishers, New York, NY, 1991.
Reprinted by permission of Kluwer Academic Publishers.
Mitchell, J., Conduction phenomena: from theory to geotechnical practice,
Geotechnique, 41(3), pp. 299-340, 1991. Reprinted with permission of the
Institution of Civil Engineers and Journal, Thomas Telford Limited, Thomas
Telford House, 1 Heron Quay, London E14 4JD, United Kingdom.
Reprinted with permission from Hazardous Materials Control Research Institute.
Loden, M., and C. Fan, Air sparging technology evaluation, in Proceedings of the
HMCRI's National Research & Development Conference on the Control of
Hazardous Materials, San Francisco, CA, 1992.
• U.S. GOVERNMENT PRINTING OFFICE: I 994-550-00 1/OO 1 70
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