1>EPA
United States
Environmental Protection
Agency
Office of Research and
Development
Washington DC 20460
EPA/625/6-89/025b
July 1990
Assessing the
Geochemical Fate of
Deep-Well-lnjected
Hazardous Waste:
Summaries of Recent
Research
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EPA/625/6-89/025b
July 1990
Assessing the Geochemical Fate of
Deep-Weil-Injected Hazardous Wastes;
Summaries of Recent Research
U.S.Environmental Protection Agency
Office of Research and Development
Center for Environmental Research Information
Cincinnati, OH 45268
Robert S. Kerr Environmental Research Laboratory
Ada, Oklahoma 74820
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Notice
This report has been reviewed by the U.S. Environmental Protection Agency and approved for
publication. Mention of trade names or commercial products does not constitute endorsement
or recommendation for use.
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Contents
Chapter page
1 EXECUTIVE SUMMARY 1
1.1 Overview 1
1.2 Processes Affecting the Geochemical Fate of Deep-Well-lnjected Wastes 2
1.3 Major Environmental Factors Affecting Deep-Well-lnjection Geochemical Processes 4
1.4 Geochemical Characteristics and Fate of Hazardous Waste 5
1.5 Methods and Models for Predicting the Geochemical Fate of Deep-Well-lnjected Wastes 7
1.6 Laboratory Procedures and Protocols 12
1.7 Field Case Studies '. 13
1.8 Further Research Needs 15
References 16
2 RESEARCH SUMMARY NO. 1: STATE-OF-THE-ART REPORT:
INJECTION OF HAZARDOUS WASTES INTO DEEP WELLS 22
2.1 Overview 22
2.2 Processes Affecting the Geochemical Fate of Deep-Well-lnjected Wastes 22
2.3 Major Environmental Factors Affecting Deep-Well-lnjection Geochemical Processes 27
2.4 Geochemical Characteristics and Fate of Hazardous Waste 28
2.5 Methods and Models for Predicting the Geochemical Fate of Deep-Well-Injected Wastes 32
2.6 Laboratory Procedures and Protocols 34
2.7 Field Case Studies 34
2.8 Further Research Needs 35
References 36
3 RESEARCH SUMMARY NO. 2: THE CHEMISTRY OF WASTE FLUID
DISPOSAL IN DEEP INJECTION WELLS 40
3.1 Overview 40
3.2 Processes Affecting the Geochemical Fate of Deep-Well-lnjected Wastes 40
3.3 Major Environmental Factors Affecting Deep-Well-lnjection Geochemical Processes 41
3.4 Geochemical Characteristics and Fate of Hazardous Waste 42
3.5 Methods and Models for Predicting the Geochemical Fate of Deep-Well-lnjected Wastes 42
3.6 Laboratory Procedures and Protocols 44
3.7 Field Case Studies ; 44
3.8 Further Research Needs 44
References 44
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Contents (Continued)
Chapter
Page
4 RESEARCH SUMMARY NO. 3: LABORATORY PROTOCOL FOR DETERMINING FATE
OF WASTE DISPOSED IN DEEP WELLS 46
4.1 Overview ,.. 46
4.2 Processes Affecting the Geochemical Fate of Deep-Well-lnjected Wastes 47
4.3 Major Environmental Factors Affecting Deep-Well-Injection Geochemical Processes 48
4.4 Geochemical Characteristics and Fate of Hazardous Waste 48
4.5 Methods and Models for Predicting the Geochemical Fate of Deep-Well-lnjected Wastes 49
4.6 Laboratory Procedures and Protocols 50
4.7 Field Case Studies • • 53
4.8 Further Research Needs 53
References • 53
5 RESEARCH SUMMARY NO. 4: GEOCHEMICAL INTERACTIONS OF HAZARDOUS WASTES
WITH GEOLOGICAL FORMATIONS IN DEEP-WELL SYSTEMS 55
5.1 Overview 55
5.2 Processes Affecting the Geochemical Fate of Deep-Well-Injected Wastes 55
5.3 Major Environmental Factors Affecting Deep-Well-Injection Geochemical Processes 57
5.4 Geochemical Characteristics and Fate of Hazardous Waste 58
5.5 Methods and Models for Predicting the Geochemical Fate of Deep-Well-lnjected Wastes 58
5.6 Laboratory Procedures and Protocols 60
5.7 Field Case Studies 61
5.8 Further Research Needs '. 61
References • 61
6 RESEARCH SUMMARY NO. 5: CURRENT GEOCHEMICAL MODELS TO PREDICTTHE
FATE OF HAZARDOUS WASTES IN THE INJECTION ZONES OF DEEP DISPOSAL WELLS 64
6.1 Overview 64
6.2 Processes Affecting the Geochemical Fate of Deep-Well-lnjected Wastes 65
6.3 Major Environmental Factors Affecting Deep-Well-Injection Geochemical Processes 68
6.4 Geochemical Characteristics and Fate of Hazardous Waste 70
6.5 Methods and Models for Predicting the Geochemical Fate of Deep-Well-lnjected Wastes ..70
6.6 Laboratory Procedures and Protocols , 86
6.7 Reid Case Studies 87
6.8 Further Research Needs : 87
References • 87
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CHAPTER ONE
EXECUTIVE SUMMARY
1.1 Overview
This report compiles and summarizes the results of recent
research funded by the U.S. Environmental Protection
Agency on topics related to geochemical-fate assessment
of deep-well-injected hazardous wastes. Its purpose is
twofold:
1. To make the results of this research available to a wider
audience of scientists and professionals who are involved
in various aspects of regulating and implementing the deep-
well injection of hazardous wastes under federal and state
Underground Injection Control programs.
2. To provide an overall assessment of the state of the art of
predicting the geochemical fate of deep-well-injected was-
tes and to identify possible future directions for research in
this area.
Chapters Two through Six summarize individual research
reports. These five reports include two literature surveys
(ArdenStrykerand A. Gene Collins, State-of-the-Art Report:
Injection of Hazardous Wastes into Deep Wells; John A.
Apps, Current Geochemical Models to Predict the Fate of
Hazardous Wastes in the Injection Zones of Deep Disposal
Wells); two reports focusing on laboratory procedures for
predicting the geochemical fate of injected hazardous waste
(J. Apps, L. Tsao, and O. Weres, The Chemistry of Waste
Fluid Disposal in Deep Injection Wells; A. Gene Collins and
M. E. Crocker, Laboratory Protocol for Determining Fate of
Waste Disposed in Deep Wells); and one report comparing
empirical data with the predictions from solution
geochemistry models (W. R. Roy, S. C. Mravik, I. G. Krapac,
D. R. Dickerson, and R. A. Griffin, Geochemical Interactions
of Hazardous Wastes with Geological Formations in Deep-
Well Systems).
This Executive Summary synthesizes the current under-
standing of the geochemistry of deep-well hazardous waste
injection. It is drawn largely from information presented in
the research summaries in this document, supplemented by
additional literature review. These additional areas of study
include the effects of organic matteron geochemical proces-
ses and biodegradation of hazardous organics in the deep-
well environment Since many of the conclusions of the
additional literature are based on scientific data developed
under near-surface environmental conditions, the
similarities and differences between near-surface and deep-
surface geochemical environments are noted where
relevant.
The research summarized in this document represents an
end to a 10-year hiatus in the study of the geochemical fate
of deep-well-injected industrial wastes. The last period of
active research on this subject took place from the late
1960s to the late 1970s. Most waste-reservoir interaction
studies were published between 1972 and 1978, and vir-
tually all reports of field studies of the geochemical fate of
injected wastes appeared between 1971 and 1978. (The
most recent citation, Vecchioli et al., 1984, reports no data
after 1977.) Most of the post-1978 literature that is cited in
Chapters Two and Six (the reports containing the most
comprehensive literature reviews) does not relate directly to
the geochemistry of deep-well waste injection, although it
may provide insights into deep-well geochemical proces-
ses.
Thus, very little research specifically addresses the
geochemical fate of deep-well-injected hazardous wastes,
particularly in the context of the current federal and state
regulatory environment for deep-well injection. A broad
range of scientific literature is available on the geochemical
fate of hazardous wastes in soil and near-surface
groundwater systems. However, most of this literature is
based on laboratory and/or field studies that do not simulate
deep-well environmental conditions, and transferring results
to estimate deep-well geochemical fate must be done
cautiously.
The following uniform format is used for presenting material
in this chapter and each of the five reports:
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• Section 1 (Overview) presents the title and authors of
the reports, where it can be obtained, a brief description
of its contents, and a summary of its major conclusions.
• Section 2 (Processes Affecting Geochemical Fate)
summarizes information on basic processes that may
affect geochemical reactions between injected waste
and fluids and solids in the injection zone. Specific
chemical processes are classified into three categories:
— Partition processes affect the form or state of a
compound but not its chemical structure or toxicity.
These processes include: acid-base reactions, ad-
sorption-desorption, immiscible phase separation
and precipitation- dissolution.
— Transformation processes afterthe chemical structure
of a substance. These processes include bfodegrada-
tfon, complexatton, hydrolysis, neutralization, oxidation-
reduction, polymerization, and thermal degradation.
Catalysis is included in this category. [
— Transport processes cany wastes through the subsur-
face environment. Only those transport processes that
significantly affect geochemistry (hydro-dynamic disper-
sion, osmotic potential, and particle migration) are covered.
• Sections (Major Environmental Factors Affecting
Geochemical Processes) contains any information in
the report related to the significance and effect of
environmental factors such as pH, Eh, salinity,
reservoir-matrix minerals, temperature, and pressure
on geochemical processes. It includes any
information on the actual environmental conditions
that exist in deep-well injection zones.
• Section 4 (Geochemical Characteristics of
Hazardous Waste) summarizes information on the
chemical characteristics of specific organic and
inorganic substances (both hazardous and
nonhazardous) that may be injected into deep-well
formations.
• Section 5 (Methods and Models for Predicting
Geochemical Fate) summarizes information on
basicapproachesto geochemical modeling and specific
methods or models for predicting adsorption, aqueous
and solution geochemistry, biodegradation, hydrolysis,
and transport.
• Section 6 (Laboratory Procedures and Protocols)
summarizes any laboratory procedures described in
the report for obtaining empirical data on
waste-reservoir geochemical interactions.
• Section 7 (Field Case Studies) summarizes any
information in the report on field observations of
geochemical interactions between injected wastes and
the injection zone.
• Section 8 (Further Research Needs) lists any
recommendations for further research.
The Executive Summary provides an overview of the
state of the art in geochemical fate assessment for
deep-well injection of hazardous wastes. This chapter
follows the same format used in the research-summary
chapters to identify strengths and weaknesses in current
knowledge. The conclusions in this chapter are drawn
from a synthesis of information in the research summaries
and additional review of relevant scientific literature.
The uniform format of each report is designed to facilitate
locating information on specific topics. For example, any
information on the processes of adsorption will be included
under Partition Processes in Section 2 of each chapter.
Similarly, any information on aqueous- and solution-
gochemistry models will be found in Section 5 of each
chapter.
Most of the reports summarized in this document contain
some discussion of EPA's 1988 Final Underground Injection
Control regulations concerning injection of hazardous was-
tes (53 Federal Register 28118-28157). Discussions
specific to regulatory issues have not been included in the
Research Summaries.
1.2 Processes Affecting the Geochemical
Fate of Deep-Well-lnjected Wastes
Environmental conditions in the deep-well environment (see
also Section 1.3) restrict the number of basic chemical
processes that may immobilize or transform hazardous
wastes. For example, absence of sunlight and air-water
interfaces means that photolysis and volatilization do not
occur. The significance of geochemical processes that may
affect deep-well-injected waste are briefly discussed below.
1.2.1 Partition Processes
Acid-base equilibria are fundamental to the aqueous
geochemistry of injected waste and the injection zone. The
high salinities of injection zones make predicting such reac-
tions more difficult than predicting those occurring in the fresh
or moderately saline waters typically found in near-surface
environments (excluding marine environments).
Adsorption-desorption is likely to be a significant process
affecting the mobility of heavy metals and organic wastes.
The basic mechanisms for these processes are still not well
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understood. Such deep-well environmental conditions as
increased temperature, pressure, and high salinities make
the study and prediction of adsorption more difficult com-
pared to those for near-surface environments. Organic-mat-
ter content is a major factor affecting adsorption in the near
surface, but its significance in the deep-well environment
has received little attention.
Precipitation-dissolution reactions are particularly impor-
tant because incompatibilities between injected wastes and
reservoir fluids commonly result in precipitation reactions
that can plug or at least reduce the permeability of the
injection zone. Both precipitation and dissolution reactions
affect the dissolved species that are available for acid-base
reactions.
Immiscible - Phase Separation is not a major process in
the deep-well environment because deep-well injection is
generally limited to waste streams that are soluble in water.
Blowout, caused when gaseous carbon dioxide forms as a
result of the injection of concentrated acids into carbonate
formations, is an example of a process unique to the deep-
well environment.
\
1.2.2 Transformation Processes
Neutralization of highly acidic wastes in injection zones with
carbonate lithology is one of thef ew geochemical processes
that can be predicted with any confidence. Neutralization of
acidic and alkaline wastes in injection zones with other
lithologies will also occur to varying degrees.
Complexation is likely to be an important process affecting
the mobility of heavy metals in the deep-well environment.
High salinities in the deep-well environment make prediction
of complexation reactions more difficult. Humic substances
can be major factors in complexation reactions, but their
significance to such reactions in deep-well environments
has received little attention.
Hydrolysis may be a significant process for a limited num-
ber of organic compounds in the deep-well environment
(see Table 1-1). EPA's 10,000-year- no-migration standard
establishes a time frame in which half-lives on the order of
thousands of days may be adequate to allow hydrolysis to
be a significant transformation process. The half-lives
reported in Table 1 -1 are based on rate constants measured
at surface conditions and do not necessarily reflect what
would occur under the temperatures, pressures, and
salinities typical of the deep-well environment. The higher
temperatures and pressures that exist may increase rates
of hydrolysis; however, the high salinities may affect rate
constants in unpredictable ways.
Oxidation-reduction (redox) reactions involving inorganic
constituents in the deep-well environment will affect the
mobility of heavy metals and precipitation reactions and will
strongly influence the type of microbiological activity. Many
Table 1-1 Listed Hazardous Organic Wastes for Which
Hydrolysis May Be a Significant Transform-
ation Process in the Deep-well Environment
Group Compound
Half-life3
Pesticides
AWrin 750
Dieldrin 3330
DDT* _
Endosulfan/Endosulfan sulfateb 21
Heptachtor 1
Halogenated Aliphatic Hydrocarbons
Chloroethane (ethyl chloride)6 38
1,2-Dichloropropaneb
1,3-Dfchloropropeneb _
Hexachk>rocyclopentadieneb 14
Bromomethane (methyl bromide)13 20
Bromodichioromethane 5,000
Methyl chloride
Halogenated Ethers
bis(Chloromethyl)etherb < 1
2-Chloroethyl vinyl ether 1,800
bis(2-Chtoroethoxy) methaneb —
Monocyclic Aromatics
Pentachlorophenol 200
Phthalate Esters
Dimethyl phthalate 1,200
Diethyl phthalate 3700
Di-n-butyl phthalate 7,'eoo
Di-n-octyl phthalate 4^00
Polycyclic Aromatic Hydrocarbons —
Nitrosamines and Misc. Compounds —
aHalf-life measured in days at pH 7 and ambient
temperature.
Hydrolysis identified as a significant process by Callahan et
al. (1979).
Sources: Callahan et al. (1979), Mills et al. (1985);
Schwarzenbach and Giger (1985); Ellington et al. (1988).
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organic compounds are degraded by biologically mediated
redox reactions (see Bfodegradation, below). Aerobic con-
ditions may exist near the injection well when injected
wastes contain dissolved oxygen. However, the deep-well
environment is typically mildly to strongly reducing
(anaerobic), and any oxygen in injected wastes is likely to
be depleted rapidly.
Catalysis may increase the rate of other transformation
reactions (such as hydrolysis, redox reactions, and
polymerization). Catalytic reactions underthe temperatures
and pressures typical of the deep-well environment have
received little attention in the literature.
Polymerization reactions of some monocyclic aromatic
compounds (catalyzed by clays) may serve to enhance
adsorption of these compounds. Such reactions under
deep-well environmental conditions have not been studied.
Thermal-degradation reactions under deep-well environ-
mental conditions have not received much study. In general,
however, temperatures and pressures typical of deep-well
injection zones are probably too low for initiating high-
temperature reactions.
Blodegradatlon of at least some components of deep-well-
injected wastes has been observed in all cases whfere
mixtures of injected wastes and formation waters have been
used to make direct observations of microbiological and/or
geochemical effects (see Section 1.7). Degradation of or-
ganic wastes by denitrifying, sulfate-reducing, and
methanogenic bacteria have been either observed directly
or inferred from geochemical evidence. The conditions of
the deep-well environment are well within the range of
conditions to which anaerobic bacteria are adapted.
Denitrifying and methanogenic bacteria have been ob-
served to degrade a number of hazardous halogenated
aliphatic and monocyclic aromatic hydrocarbon compounds
in near-surface environments. Sulfate-reducing bacteria
appear to be more abundant and adapted to a wider range
of environmental conditions in the deep-well environment
than denitrifying or methanogenic bacteria, but are less
capable of degrading hazardous organic compounds; data
are, however, very limited on this subject. The ecologies of
denitrifying, sulfate-reducing, and methanogenic bacteria in
both near-surface and deep-well environments are not well-
understood, and this area of research has received little
attention in the context of deep-well injection of wastes.!
1.3 Major Environmental Factors Affecting
Deep-Well-lnjection Geochemical Processes
Every injection well has a unique set of enviromental factors
that determine the chemical reactions that may occur when
waste is injected. Compared to those of near-surface en-
vironments, the parameters that define the deep-well en-
vironment are much narrower in range.
1.3 Typical Range of Environmental Factors
pH and Eh (Redox Potential). The pH of most deep-well
injection zones ranges from 5.0 to 8.5 and the Eh ranges
from 0 to -400 mV. The pH range is well suited for neutraliza-
tion reactions (although the lithology of the injection zone is
an equally important factor affecting n eutralization capacity).
As noted in the previous section, the low Eh values of typical
injection zones indicate that aneraboic, biologically
mediated reducing reactions will predominate.
Salinity and Water Chemistry. The salinity of most deep-
well injection zones ranges from 20,000 to 100,000 rng/L,
although values as high as 350,000 mg/L are possible.
Sodium chloride, a strong electrolyte, is a major constituent
of water in deep-well injection zones. High salinities and
electrolytic characteristics of waters in injection zones
complicate the modeling and prediction of the
aqueous geochemistry of injected-waste and formation-
water mixtures (see Section 1.5).
Reservoir Matrix. Sedimentary roc*; forms the solid matrix
of most deep-well injection zones. Many chemical reactions
occurring when hazardous wastes are injected are largely
determined by the physical and chemical properties of that
rock. Approximately two-thirds of active waste-injection
wells use sands and sandstones for the injection zone, and
about one-third use carbonate rock (limestone and
dolomite). Particle-size distribution influences the surface
area available for waste-solid interactions. Mineralogy also
strongly influences the types of chemical reactions that will
occur at the waste-solid interface.
Silicate clays (particularly the smectite group, vermiculite,
and illite) and hydrous-oxide clays are the most reactive
minerals in deep-well formations because of their high
surface area and cation-exchange capacity. Clays are par-
ticularly important in adsorption reactions and may catalyze
other geochemical reactions. Organic matter composed of
stable humic substances is even more reactive than clay
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minerals in near-surface environments, but this material has
not received much attention in the context of geochemical-
fate assessment in the deep-well environment. Organic
matter in sedimentary rocks (bitumin and kerogen) is
derived largely from humic substances that originated at
near-surface conditions, but it is possible that burial may
alter the reactivity of these materials.
Temperature and Pressure. The temperatures of most
deep-well injection zones range from 40° to 75°C with
extremes from 20° to 150°C. Pressures most likely range
from 50 to 150 bars. The rates of most acid-base and
dissolution reactions increase with temperature, but the
exact effect of the interactions among competing reactions
is difficult to predict. Increased temperature usually
decreases the rate of adsorption because these reactions
are primarily exothermic. The separate or combined effects
of temperature and pressure may result in different reactions
from those occurring among wastes and the injection zone
fluids and solids under near-surface conditions. Conse-
quently, laboratory compatibility tests and evaluations of
individual geochemical processes must simulate actual
temperatures and pressures in the injection zone.
1.3.2 Influence of Environmental Factors on Waste-
Reservoir Compatibility
Injected wastes tend to be less saline (on the order of 10,000
mg/Ltotal dissolved solids), and theirdissolved constituents
usually have attained equilibrium at lowertemperatures and
pressures than those found in the injection zone. When
such wastes are injected without considering possible in-
compatibilities, adverse chemical and physical reactions
can occur near the injection site. Major operational
problems that can occur in this situation include: (1) well
plugging, (2) well-casing/confining-layerfailure, and (3) well
blowout. These problems are discussed below.
Well Plugging. Table 1-2 lists 10 possible causes of
reduced permeability in the injection zone or plugging of well
screens and possible remedial actions. The most common
causes of plugging are (1) swelling and migration of water-
sensitive clays (such as montmorillonite) when lower-
salinity wastes replace the formation fluids, (2) precipitation
reactions, and (3) biological clogging of well screens.
Well-Casing and/or Confining-Layer Failure. Corrosion
of well casing and packing can threaten the integrity of a well
if proper materials have not been used in construction (see
U.S. EPA, 1989). Dissolution of the confining-layer forma-
tion by highly acidic or alkaline wastes may also allow
upward migration of wastes. Chemically active injected
fluids may reduce surface energy, surface cohesion, and
breaking strength of a confining formation, and stresses
from increased injection pressure may fracture rock to
create channels in a confining formation (Swolf, 1972).
Carbonate-confining layers are most susceptible to breach-
ing from dissolution by acidic wastes. Lower density and
viscosity of injected wastes compared with those of the
injection-zone fluids tend to increase the potential for waste-
confining-layer interactions and upward migration.
Well Blowout. Injection of hot, concentrated hydrochloric
acid into a carbonate zone can result in phase separation of
carbon-dioxide gas. In this situation, formation pressures
increase to the point where waste and reservoir fluids are
forced up the injection well to the surface. This problem can
be controlled readily by keeping temperatures and con-
centrations of the acidic waste within certain recommended
limits: (1) less than 6% HCI (Pangiotopoulos and Reid, 1986)
and (2) less than 88T (Kamath and Salazar, 1986).
1.4 Geochemical Characteristics and Fate of
Hazardous Waste
1.4.1 Sources and Composition of Deep-Well-lnjected
Wastes
An estimated 11.5 billion gallons of hazardous wastes were
injected in 1983, of which about half (50.9%) came from the
Table 1-3 Estimated Volume of Deep-Well- Injected
Wastes by Industrial Category, 1983
Industrial
Category
Volume
(MGY)
Percent
of Total
Organic chemical
5,868
Petroleum refining and
petrochemical products 2,888
Misc. chemical products 687
Agricultural chemical
products 525
Inorganic chemical
products 254
Commercial disposal 475
Metals and minerals 672
Aerospace and related
industry 169
Total 11,539
50.9
25.0
6.0
4.6
2.2
4.1
5.8
1.5
100.0
Source: U.S. EPA (1985).
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Table 1-2 Causes of Well Plugging and Possible Remedial Actions
Cause
Possible Action
Failure to remove paniculate solids
and/or colloids before injection.
Bacterial growth on well screen and
formation.
Emulsiftcatfon of two fluid phases.
Precipitates resulting from mixing
of injection and reservoir fluids.
Expansion and dispersion of water-
sensitive clays (particularly
montmorillonite).
Migration of fines (very small
particles) released by dissolution.
Repreclp'rtatfon of dissolved
material (iron or calcium sulfate).
Change in wettabifity or reduction
in pore dimensions by adsorption
(organfcs w'rth large molecular
weight).
Flow of unconsolidated sands into
well-bore.
Scaling on injection equipment by
precipitation from injection fluid.
Filtration before injection.
Treatment with bactericides.
Do not exceed solubility limits of
organic wastes in water.
Pretreatment; buffer of non-
reactive water.
Avoid injection of tow-salinity
solutions in water-sensitive
formations. Use clay stabilizers.
Preinjection neutralization to avoid
dissolution.
Pretreatment.
Difficult to remedy.
Gravel pack well screen. Inject a slug
of brine after every period of interrupted
flow.
Pretreatment; flushing with solutions
to remove accumulated scale.
Source: Adapted from Barnes (1972), Donaldson and Johanson (1973),
Hower et al. (1972), Davis and Funk (1972), and Veley (1969).
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production of organic chemicals and one-quarter (25.0%)
came from the petroleum-refining and petrochemical-
products industry (see Table 1 -3). The remaining 24% came
from six other industrial categories: miscellaneous
chemical production (6.0%), agricultural chemical products
(4.6%), inorganic chemical products (2.2%), commercial
disposal (4.1%), metals and minerals (5.8%), and
aerospace (1.5%).
Although no systematic data exist on the exact chemical
composition of deep-well-injected wastes, EPA has
gathered data for 108 wells (55% of total active wells)
operating in 1983. Table 1-4 summarizes the total quantity
of undiluted waste in six major categories, provides a break-
down of average concentrations of individual constituents
for which data were available, and indicates the number of
wells involved. A little more than half the undiluted-waste
volume was composed of nonhazardous inorganics
(52.0%). Acids were the next most important constituent by
volume (20.3%), followed by organics (17.4%). Heavy me-
tals and other hazardous inorganics made up less than 1%
of the total volume in the 108 wells. About a third of the wells
injected acidic wastes and about two-thirds injected organic
wastes. Although the percentage of heavy metals by volume
was tow, almost one-fifth of the wells injected wastes con-
taining heavy metals.
1.4.2 Geochemlcal Fate of Deep-Well-Injected Wastes
Several dozen inorganic elements and compounds are
classified as hazardous or exhibit toxic characteristics at
relatively low concentrations, and hundreds of organic
compounds are classified as hazardous. Much of the data
available on the geochemical properties of individual
complexity of natural environments, where competing
processes and chemical reactions may lead to outcomes
different from those indicated by laboratory studies. Injected
hazardous wastes tend to be complex mixtures of
hazardous and nonhazadous organic and inorganic
constituents. The number of possible chemical interactions
increases factorially as the number of compounds in the
waste stream increases, confounding geochemical fate
predictions for specific substances. Variations in
particle-size distribution and mineral composition of the
injection zone (which is difficult to characterize from a few
boreholes) further complicate predictions of geochemical
reactions between a waste and the injection zone.
1.5 Methods and Models for Predicting the
Geochemical Fate of Deep-Weil-lnjected
Wastes '
1.5.1 Specific Methods and Models
Aqueous-Geochemistry Computer Codes. Four general
types of computer codes are used to model aqueous
geochemistry:
1. Thermodynamic codes process empirical data so that
thermodynamic data at a standard reference state (25°C
and 1 bar) can be obtained for individual species. They are
also used to recalculate thermodynamic properties of the
species of interest at the temperatures, pressures, and ionic
concentrations being simulated.
2. Distribution-of-species codes, also called equilibrium
codes, solve a simultaneous set of equations that describe
equilibrium reactions and mass balances of the dissolved
elements. The output is the theoretical distribution of the
aqueous species for the dissolved elements.
3. Reaction-progress codes, also called mass-transfer
codes, calculate both the equilibrium distribution of
aqueous species (as in distribution-of-species codes) and
the new composition of the solution, as selected minerals
and compounds are precipitated or dissolved.
4.Transport codes model chemical transport by combining
aqueous-geochemistry codes with physical- transport
codes. Two major approaches have been used: integrated
codes simultaneously solve all mass, momentum, and ener-
gy-transfer equations, including those in which chemical
reactions participate, for each time step in the evolution of
the system; two-step models first solve mass momentum
and energy balances for each time step and then re-equi-
librate the chemistry using a distribution-of- species code.
Dozens of codes have been developed to model aqueous
geochemistry, but most have been developed for near-sur-
face environments. Only a few are suitable for modeling the
high salinities, temperatures, and pressures that exist in the
deep-well environment. Thermodynamic, distribution-of-
species, and reaction-progress codes that may be useful for
modeling deep-well injection are listed and described in
Table 1-5. Table 1-6 describes integrated and two-step
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Table 1-4 Waste Characteristics of 108 Hazardous-Waste Wells Active In 1983 In the United
Waste Type/
Components
Acids
Hydrochloric acid
Sulfuric acid
Nitric acid
Formic acid
Unspecified acid
Heavy Metals
Chromium
Nickel
Unspecified metals
Metal hydroxides
Hazardous Inorganics
Selenium
Cyanide
Organlcs
Total organic carbon (TOC)
Phenol
Oil
Organic acids
Organic cyanide
Isopropyl alcohol
Formaldehyde
Acetophenone
Urea-N"
Chlorinated organfcs
Formic acid
Organic peroxides
Pentachtorophenol
Acetone
Nit rile
Methacrytonitrile
Ethylene chloride
Carbon tetrachtoride
Nonnazardous
Inorganics
Other
Total
Gallons8
44,140,900 (20.3)b
'
1,517,600(0.7)
89,600 (< 0.1)
39,674,500(17.4)
'
118,679,700 (52.0)
22,964,600 (9.9)
228,021,800°
States
Avg. Concentration
(mg/L) No. of Wells
78,573
43,000
75,000
75,000
44,900
1.4
600
5,500
1,000
0.3
391
11,413
805
3,062
10,000
400
1,775
15,000
650
1,250
35,000
75,000
4,950
7.6
650
700
22
264
970
37
15
6
2
2
12
19 (17.6)
11
5
2
1
4 (3.7)
2
2
71 (65.7)
24
22
6
3
3
3
2
2
2
2
2
2
2
2
1
1
1
1
50
'33 (30.5)
108
aGaltons of undiluted wastes.
bNumber in parentheses is percent of total.
°ExcIudes overlaps between organics and acids.
Source: U.S. EPA (1985).
-------
Table 1-5 Aqueous- and Solution-Geochemistry Models of Potential Value for Modeling Deep-Well-Injection
Name/Developer(s)
Description/Comments
Thermodynamic Codes
SUPCRT
Shock and Helgeson, 1988,1989
Tangerand Helgeson, 1988
PHAS20
Haas, 1974
Haas and Fisher, 1976
Can be used to calculate dissolution-reaction con-
stants at any specified temperature between 0° and
800"C and 1-5,500 bars.
Developed by USGS forthermodynamic calculations.
Distribution-of-Species Codes
SOLMNEQ
Kharaka and Barnes, 1973
Handles temperature of 0°-350°C, pressure from
1 -1,000 bars, and salinities up to about 29,000
mg/L It includes organic complexes and ion-exchange
equilibria. The model has been used by Ehrlich
et al.(1979) and Roy et al. (1988) to simulate injected
waste-reservoir interactions.
Reaction-Progress Codes
EQ3/6
Wolery and Walters, 1975
Wolery, 1979
Wolery, 1983
Jackson and Wolery, 1985
Wolery, 1986
PHREEQE
Parkhurst et al., 1980
Plummeretal., 1983
Plummer and Parkhurst, 1985
Plummeretal., 1983).
PHREEQEP
Crowe and Longstaffe, 1987
ECESa
Scrivneretal., 1986
Handles temperatures of 0°-350°C and pressures from
1 -500 bars. Earlier version handles salinities up to
about 0.5 molal (~29,000 mg/L); latest version con-
tains Pitzer interaction electrolyte model. Has been
used to model geochemical evolution of Gulf Coast
brines (Apps et al., 1988) and to simulate evolution of
ground waters in basalt (Solomon, 1986). Most
thoroughly documented of available models.
Temperature range 20°-150°C, pressure range
50-300 bar, salinity range 10,000-350,000 TDS.
Has successfully modeled the evolution of ground
water with the mineralogy of a limestone and dolomite
aquifer in Florida (Plummer et al., 1983).
Incorporates Pitzer interaction electrolyte
model into PHREEQE up to 150°C.
Temperature range is 0°-200°C; pressure range is
0-200 atm; and tonic strength is 0-30 molal. It incor-
porates the Pitzer interaction electrolyte model for
high salinities. It is a proprietary model licensed by
OLI Systems, Morristown, New Jersey.
Electrolyte Concentration of Equilibrium Solution
Sources: Nordstrom et al., 1979; Apps, 1988.
-------
Table 1*6. integrated Groundwater Chemical Transport Models
Developers
Description/Comments
integrated Models
Rubin and James, 1973
Valtochi, Street, and Roberts, 1981
Valtochietal., 1981
Jennings et al., 1982
Miller and Benson, 1983
Noorishad and Camahan, 1985
Camahan, 1986
Noorishad etal., 1987
Camahan, 1987
Two-Step Models
Grove and Wood, 1979
Reardon, 1981
Walsh etal., 1982
Cederberg et al., 1985
Kirkneretal., 1984,1985
Their etal., 1984
Huyakom etal., 1983
Krupka and Morrey, 1985
Narashimhan et al., 1986
Liu and Narashimhan, 1989 a,b
Simulates heterovalent ton exchange and changing concentrations
of pore fluid tons in one-dimensional flow.
Simulates multispecies heterovalent ion exchange under conditions
of varying total solution concentrations.
Multicomponent equilibrium chemistry in ground water.
CHMTRN includes dispersion/diffusion, advection, adsorption of
ions and complexes, aqueous complex formation, and dissociation
of water. THCC is a variant that simulates uranium transport with
variable temperature and oxidation potential. Latest version, called
CHMTRNS, can simulate in one dimension both homogeneous
aqueous phase and heterogeneous temperature-dependent reaction
kinetics. Has been applied to a number of simple problems involving
reversible and irreversible dissolution, and oxidation-reduction reac-
tions. Has not been tested with complex multicomponent systems.
Solved the nonreacting advective-dispersive transport equation.
Uses distribution-of-species code by Morel and Morgan (1972).
TRANQL incoporates distribution-of-species code MICROQL (Westall
et al., 1976). Modeling of ion-exchange reactions in artificial recharge
in Palo Alto Baylands project yielded same results as one-step
analysis by Valocchi, Street, and Roberts (1981).
Models multicomponent solute transport with adsorption and aqueous
complexation.
SATURN incorporates distribution-of-species code MINTEQ (Felmy
et al., 1984; Krupka and Morrey, 1985).
DYNAMIX combines the transport code TRUMP (Edwards, 1972)
with distribution-of-species code PHREEQE (Parkhurst et al., 1980).
Most recent version handles thermodynamics of hydrolysis aqueous
cpmplexation, redox reactions and precipitation-dissolution. Field-
tested by White et al. (1984). Comparison of predicted and laboratory
column uranium transport with one-step code THCC yielded similar
results.
Source: Apps (1988).
10
-------
groundwater chemical-transport models. Of those listed,
SOLMNEQ, EQ3/6, and ECES have actually been used to
model the deep-well environment. Section 1.5.2 discusses
deficiencies in available models.
Adsorption is a complex process involving one or more of
a number of bonding mechanisms (ion exchange, protona-
tion, hydrogen, Van der Waals, hydrophobic, and/or dipole
bonding). Thermodynamically and kinetically sound models
of simple systems are available (e.g., the Langmuir equation
for adsorption of gases on homogeneous surfaces and the
Stanford General Model for Adsorption [SGMA]—the triple-
layer model—for simple oxides). However, no such model
is available for modeling adsorption on more complex
minerals such as aluminosilicates, complex oxides, and
fixed-charge clay minerals, which are prevalent in the deep-
well environment. Consequently, researchers are currently
confined to using empirical Freundlich isotherms or distribu-
tion coefficients (in the unlikely event that adsorption exhibits
a linear relationship to concentration) when trying to predict
adsorption in the deep-well environment. Inherent
problems with this empirical approach are discussed in
detail in Section 6.5.2.3. If these approaches are taken,
laboratory adsorption must be measured under deep-well
temperatures and pressures.
Hydrolysis is easily predicted when the rate constants for
a compound are known. However, few if any data are
available on hydrolysis rate constants for compounds at the
salinities, temperatures, and pressures existing in the deep-
well environment.
Biodegradation. Several qualitative models for biode-
gradation in the deep-well environment have been sug-
gested. These models do not allow quantitative predictions
Figure 1-1 Proposed geochemical model of waste after
injection into subsurface.
Zones
ll.Obs.
Well
Front
(Degradation)
to be made, but they indicate what types of biodegradation
processes may occur. The conceptual geochemical model
of acidic waste injected into the subsurface proposed by
Leenheer and Malcolm (1973) involves a moving front of
microbial activity. The moving front has five zones, as
shown in Figure 1-1: (1) a dilute zone controlled by dif-
fusion, (2) a zone where substrate concentrations are high
enough to allow significant microbial activity, (3) a transition
zone, where increasing waste concentrations create un-
favorable conditions for microbial growth, (4) a neutralization
zone, where abiotic chemical reactions predominate, and
(5) a waste-storage zone, where undiluted waste no longer
reacts with the host rock.
Bouwer and McCarty (1984) have suggested a qualitative
model based on redox zones for microbial degradation of
trace organic constituents with increasing distance from the
injection point. Table 1 -7 shows the progression that would
occur as Eh declines with distance from the injection point
and lists hazardous organic compounds that would be
degraded most readily in each zone. The model implies that
most compounds not degraded in their appropriate zone will
move through the groundwater system without significant
additional degradation, except for those compounds which
are biodegraded methanogenically, for which complete
biodegradation may occur. Other factors, such as pH and
water chemistry (e.g., presence or absence of sulfates),
tend to complicate the redox-zone model.
The most sophisticated model available for predicting
biodegradation of organic contaminants in subsurface sys-
tems is the biofilm model, originally presented by William-
son and MeCarty (1976a,b) and refined over several years
at Stanford University and the University of Illinois/Urbana
(Rittmann et al., 1980; Rittmann and McCarty; 1980a,b;
McCarty et al., 1981; Bouwer and McCarty, 1984; Chang
and Rittmann, 1987a,b). The model predicts that biofilm
development will be confined to within about a meter of the
injection zone at near-surface artificial recharge wells.
Where biological clogging is a potential problem, the biofilm
model may be of value in deep-well-injection settings. How-
ever, where injected wastes are toxic to mfcrobiota before
dilution in the injection zone (see Figure 1-1), the biofilm
model would not be applicable.
1.5.2 Deficiencies in Geochemical Models
The geochemical modeling of the fate of hazardous wastes
in saline aquifers contained in deep sedimentary formations
is in a preliminary stage. Computer codes have not been
adequately tested, so fate predictions must be corroborated
by laboratory and field studies. The major deficiencies in
currently available geochemical codes for predicting the fate
of deep-well-injected hazardous wastes include:
11
-------
Table 1-7 Redox Zones for Biodegradation of Organic Micropollutants
Increasing Distance from Injection Point
Aerobic
heterotrophto
respiration
Biolog leal Conditions
Denitrification t Sulfate
respiration
Organic Pollutants Transformed
Methanogenesis
Chlorinated
benzenes
Bhylbenzene
Styrene
Naphthalene
Carbon tetrachtoride
Bromodichloromethane
Dibromochtoromethane
Bromoform
None identified Ci and Ca
Halogenated
aliphatics
Source: Adapted from Bouwer and McCarty (1984).
• The data on thermodynamic properties of many
relevant water-miscible organic species are either
incomplete or unavailable.
• Many minerals are solid solutions (e.g., clays,
amphiboles, and plagioclase feldspars). Either
solid-solution models have not yet been developed or
appropriate algorithms have not been incorporated into
computer codes.
• Models describing the adsorption of water-miscible
organic compounds on natural materials are in
preliminary stages of development and have not been
correlated with field observations under typical
injection-zone conditions. Few computer codes contain
algorithms for calculating the distribution of species
between the adsorbed and aqueous state.
• Calcium-sodium chloride brines (which typically occur
in deep-well injection zones) require sophisticated
electrolyte models to calculate their thermodynamic
properties. Many parameters for characterizing the
partial molal properties of the dissolved constituents in
such brines have not been determined. Precise
modeling is limited to systems with relatively low
salinities, where modeling a multitude of parameters is
unnecessary, or to chemically simple systems
operating near25"C.
• Current computer codes usually calculate only the
thermodynamically most stable configuration of a
system. Modifications to these codes can simulate
nonequilibirum conditions, but the extent to which
codes can be manipulated to simulate processes that
are kinetfcally (rate) controlled is limited. The slow
reaction rates in the deep-well environment relative to
groundwater movement create particular problems for
the simulation.
Little is known about the kinetics of dissolution,
precipitation, and oxidation-reduction reactions in the
natural environment. Consequently, simulating the
kinetics of the complicated injection-zone chemistry is
very difficult.
1.6 Laboratory Procedures and Protocols
1.6.1 General
Procedures for field and laboratory characterization of in-
jected wastes, reservoir lithology, and formation water are
well established. For geochemical fate assessment, re-
searchers must pay more attention to characterizing reac-
tive minerals in the injection zone (primarily clays) and solid
and dissolved phases of organic matter, as well as to the
microbial ecology of the injection zone (as evidenced by
gaseous byproducts of microbial activity and direct obser-
vation of microbiota). Basic methods for sampling and
identifying groundwater microorganisms are reasonably
well established, but they require some refinement if they
are to be used systematically to characterize deep-well-in-
jection zones.
12
-------
1.6.2 Waste-Reservoir Interaction Tesls
Waste-reservoir-interaction tests serve at least three pur-
poses: (1) to identify possible incompatibilities between
reservoir components and wastes to be injected, (2) to
identify types of chemical interactions, and (3) to provide
empirical data for predicting the geochemical fate of injected
wastes. Specific procedures for performing interaction tests
described in the scientific literature vary considerably but
can be grouped into two types: batch tests and
flowthrough tests.
Batch tests are performed by mixing wastes and reservoir
materials in the same proportions as those expected in the
field. The materials are mixed in a series of reactors, which
may be subjected to temperatures and pressures that simu-
late the deep-well environment. The reactors are opened
in sequence at regulartime intervals and the fluids analyzed.
When waste and reservoir fluids are mixed, the presence
and type of precipitates may be the main concern; when
injection fluid is mixed with reservoir rock, adsorption or
dissolution reactions may be of primary interest, and chan-
ges in the concentration of species being adsorbed or
dissolved species may be measured. Chapter Four
describes procedures for batch tests in more detail.
Flowthrough tests, also called dynamic coreflocd tests,
are used to study interactions between fluids and solids.
The solid may be an undisturbed core or packed columns
intended to simulate subsurface conditions. In either case,
the same core is used throughout the experiment, and the
injected fluid is monitored at the outflow end at specified
intervals to observe changes in chemistry. In adsorption
experiments, equilibrium adsorption is obtained when the
outflow concentation equals the inflow concentration. If
precipitation-dissolution reactions occur, pressure changes
caused by clogging or increased permeability may be
monitored in addition to chemical changes. Chapter Four
describes procedures for flowthrough tests in more detail.
Table 1-8 summarizes information on 11 waste-reservoir
interaction tests reported in the literature. It lists the type of
test, type of waste, geologic-formation lithology, and, where
indicated, the duration and the temperature and pressure
conditions of the experiment.
The following issues should be considered when selecting
a laboratory method for evaluating interactions between
wastes and reservoir materials:
• The results of any test method will contain uncertainties
created by the sample chosen (which may not be
representative of the injection zone) and by the possible
alteration of in-situ properties caused by shaping.
Furthermore, because the duration of such
experiments is usually measured in hours or days, only
those reactions reaching equilibrium quickly will be
measured. Reactions taking years to reach equilibrium
will not be measured.
Tests must simulate temperature and pressures in the
injection zone unless preliminary tests show that these
parameters do not significantly affect the process of
interest. For example, Elkan and Horvath (1977)
performed preliminary tests of microbiological activity at
pressures similar to those in the injection zone being
simulated and found no significant difference between
activity at the elevated pressure and that at normal
atmospheric pressure. Subsequent experiments were
then conducted at atmospheric pressure.
Experimental results from tests using simulated sand
cores or simulated waste solutions have lower
confidence levels than those in which actual cores and
waste streams are used.
Batch experiments using disaggregated material are
likely to overestimate adsorption rates because of the
larger surface area that is created by disaggregation.
Batch experiments using undisturbed cores are more
likely to yield better results, but they still will not simulate
subsurface conditons as effectively as flowthrough
experiments in undisturbed cores.
Flowthrough experiments on subsurface cores at
simulated temperature and pressure conditions will
probably yield the best results, although the
uncertainties that affect all types of waste-reservoir-
interaction experiments still apply.
1.7 Field Case Studies
The most extensive field studies of geochemical fate of
deep-well-injected wastes have taken place at four sites.
Three involved carbonate injection zones in Florida (Pen-
sacola-Monsanto, Pensacola-American Cyanamid, and
Belle Glade, all of which are still active) and one involved an
injection zone of mixed lithology (now abandoned) near
Wilmington, North Carolina. Section 2.7 in this document
provides some information on these field studies. Table 1-9
summarizes information on these field case studies, includ-
ing geochemical processes observed, and some additional
literature references.
The published data for these field case studies are 10 to 15
years old. Additional geochemical field studies of
deep-injection wells are needed in these and other
13
-------
Table 1-8 Summary of Waste-Reservoir Compatibility/Interaction Studies
Waste Type
Batch— Fluids
Acidic, Organic
(diluted)
49 organic
compounds
Time
Formation (days)!
Subsurface bacteria 3
culture
Various bacterial 2-8
cultures
Batch —Disaggregated
Acidic, inorganic St. Peter sandstone 15 '
Alkaline, organic Potosi dolomite
Proviso sittstone
Brine (Devonian)
Acidic, organic
Acidic, ferric
chtoridQ
Cresol, sodium -
borate
Ftoridan limestone —
Dolomite 0.25
Bentonite —
Batch — Undisturbed
Various organtos Cottage Grove —
sandstone
Various organtos
Cottage Grove —
sandstone
Temp
CC)
20
37-56
25-55
—
43
250
60
38-93
Rowthrough — Column
Unspecified Miocene sand — ' —
Acidic, organic
Cretaceous sand 80 :
(simulated) '
Flowthrough — Undisturbed
Acidic (steel) Mt. Simon sandstone — '.
Acidic pickling
liquor
Phenol (in
simulated brine
Dolom'rtic sandstone, — !
dolomite, quartzite ,
Friosand —
20
9
40
38-60
Pressure
(MPa) Source
0.1-27.6 Elkanand
Horvath, 1977
0.1 Grula and
Grula, 1976
0.1-11.7 Royetal.,
1988
5.07 Goolsby, 1972
6.89 Hower et
al., 1972
0.1 Apps et
al., 1988
20.3 Donaldson and
Johansen, 1973
20.7 Donaldson
et al., 1975
— Hower et
al., 1972
0.1 Elkan and
Horvath, 1977
0.1 Bayazeed and
Donaldson, 1973
13.8 Ragone et
al., 1978
24.1 Collins and
Crocker, 1988
14
i
-------
Table 1 -9 Summary of Case Studies
Location
Lithology
Wastes
Processes Observed
Additional
Sources of
Information
Florida
Pensacola
(Monsanto)
Limestone
Pensacola
(American
Cyanamid)
Belle Glade
Limestone
Carbonate
North Carolina
Wilmington
Sand
Silty sand
Limestone
Nitric Acid
Inorganic Salts
Organic
compounds
Acrylonitrile
Sodium salts
(nitrate, sulfate,
thiocyanate)
Hot acid
Organic plant
wastes
Organic acids
Formaldehyde
Methanol
Neutralization
Bacterial dentrif ication
Bacterial dentrification
No retardation of
thiocyanate ions
Neutralization
Bacterial sulfate
reduction
Methane production
Neutralization
Dissolution-precipitation
Adsorption
Bacterial sulfate and
iron reduction
Methane production
Complexation
Barraclough, 1966
Dean, 1965
Goolsby, 1971
Goolsby, 1972
Faulkner and
Pascale, 1975
Pascale and
Martin, 1978
Elkan and
Hovarth, 1977
Willis etal., 1975
Ehrlichetal., 1979
Vecchioii etal., 1984
Kaufman et al., 1973
Kaufman and
McKenzie, 1975
McKenzie, 1976
Garcia-Bengochea
and Vernon, 1970
DiTommaso and
Elkan, 1973
Leenheer and
Malcom, 1973
Peek and Heath,
1973
Leenheer et al.,
1976a,b
Elkan and Horvarth,
1977
geological areas and should be designed to consider current
regulatory requirements.
1.8 Further Research Needs
This section compiles the research recommendations con-
tained in the reports summarized in Chapters Two through
Six and includes additional recommendations arising from
the synthesis in this chapter (see Section 1.8.4). The section
from which the recommendation comes is noted after each
recommendation.
1.8.1 Thermodynamic- and Aqueous-Geochemistry
Modeling
m Continue to refine and add to the thermodynamic
databases required for modeling, and develop data on
thermodynamic properties of water-miscible organic
compounds. (Section 3.8).
15
-------
• Develop solid-solution models and the capability to
model precisely the thermody namic properties of strong
mixed electrolytes (brines) for a diverse range of
injection-zone conditions (Section 3.8). Accurately
determine the activity coefficients of tons in strong
mixed electrolytes (Section 6.8). !
• Develop better data and understanding of tjie
thermodynamic properties of clays, and develop
thermodynamic data for minerals and organic aqueous
species for which no data are currently available
(Section 6.8).
• Perform more field validation studies of geochemical
codes (Section 6.8).
1.82 Adsorption
• Develop an integrated compilation of data in the
extensive literature describing adsorption of inorganic
and organic species on clays (Section 6.8).
• Develop empirical models describing irreversible
adsorption of water-miscible organic compounds on
mineral surfaces in an injection zone (Section 3.8). |
• Develop adsorption or ton-exchange models that can
be used under the conditions of deep-well injection
(Section 6.8).
1.8.3 Specific Laboratory and Field Studies \
m Conduct dynamic coreftood (i.e., flowthrough) studies
of selected phenols (a common hazardous constituent
of injected wastes) and determine their short-term fate
(30 to 60 days) under typical reservoir conditions
created in the laboratory. Such parameters as solution
pH, salt concentration, temperatures, clay composition,
and waste concentration should be evaluated With
respect to precipitation, adsorption, permeability
reduction, and thermal degradation (Section 2.8). !
• Conduct additional dynamic coreftood and/or related
studies of selected hazardous wastes to determine trieir
fate in subsurface environments. These studies might
include: coreftood studies using different cores and
other organic-waste compounds; studies of the
interactions of phenols with confining-layer materials
(using batch reactors rather than corefloods); and
studies of the effects of microorganisms on the
degradation of phenols (Section 2.8). !
• Study further nonequilibrium in in-situ brines. The
comparison of simulated and actual Gulf Coast brines
in Section 3.5 suggest that in-situ brines are not in
homogeneous or heterogeneous! equilibrium (Section
3.8).
• Conduct field studies to compare results of laboratory
experiments and computer modeling reported in
Chapter Three. These studies could involve injecting a
simulated waste stream containing variable amounts of
sodium borate and cresol in an arenaceous
(sandstone) formation. The injected stream could be
left in place for an extended period of time, then
recovered and changes in its composition measured.
The formation fluids could be continually removed and
measured for changes in borate and cresol, allowing
the adsorptive-desorptive capacity of the rock, potential
decomposition products, and various hydrologic
parameters to be determined. The results could be
correlated with laboratory studies and conclusions
drawn regarding the scaling factors and more
fundamental differences in mechanisms between
laboratory and field conditions (Section 3.8).
1.8.4 Other Research Needs
• Measure hydrolysis rate constants of selected
hazardous organic compounds at simulated
deep-well-injection temperatures, pressures, and
salinities. The selected halogenated aliphatic
hydrocarbons and phthalate esters in Table 1-1 would
be good candidates for such studies.
• Develop information on the amount and chemical
characteristics of organic matter in typical
deep-well-injection formations. (The Frio formation in
Texas, which receives more injected wastes than any
other formation, would be a good candidate.) Evaluate
the significance of organic matter in deep-well injection
as it affects adsorption and complexation as compared
with its significance in near-surface environments.
• Perform general studies of the ecology of anaerobic
bacteria in deep-well-injection formations to identify
measurable environmental parameters (pH, Eh,
salinity, inorganic substrates, etc.) that in combination
might be used to predict which type of anaerobic
microorganisms (denitrifying, sulfate-reducing, and
methanogenic) are likely to be most active in degrading
hazardous constituents of injected wastes. Perform
more specific studies of the ability of sulfate-reducing
bacteria to degrade hazardous organic wastes.
References
Apps, J. A. 1988. Current Geochemical Models to Predict
the Fate of Hazardous Wastes in the Injection Zones of
16
-------
Deep Disposal Wells. Draft Report prepared for
Lawrence Berkeley Laboratory, LBL-26007.
EPA,
Apps, J., L. Tsao, and O. Weres. 1988. The Chemistry of
Waste Fluid Disposal in Deep Injection Wells. 2nd Berkeley
Symposium on Topics in Petroleum Engineering, pp. 79-82.
Lawrence Berkeley Laboratory, LBL-24337.
Barnes, 1.1972. Water-Mineral Reactions Related to Poten-
tial Fluid Injection Problems. In Symposiumon Underground
Waste Management and Environmental Implications, T.D.
Cook, ed. Houston, Texas. Am. Ass. Petr. Geol. Mem. 18.
pp. 294-297.
Barraclough, J. T. 1966. Waste Injection into a Deep Lime-
stone in Northwestern Florida. Ground Water4(1):22-24.
Bayazeed, A. F., and E. C. Donaldson. 1973. Subsurface
Disposal of Steel Pickle Liquor. U.S. Bureau of Mines.
Report of Investigation 7804,31 pp.
Bouwer, E. J., and P. L. McCarty. 1984. Modeling of Trace
Organics Biotransformation in the Subsurface. Ground
Water 22(4):433-440.
Callahan, M. A., et al. 1979. Water-Related Environmental
Fate of 129 Priority Pollutants. EPA Report 440/4-79-029a-
b. Washington DC.
Carnahan, C. L. 1986. Simulation of Uranium Transport
with Variable Temperature and Oxidation Potential: The
Computer Program THCC. Lawrence Berkeley
Laboratory, LBL-21639.
Camahan, C. L. 1987. Simulation of Chemically Reactive
Solute Transport under Conditions of Changing Tempera-
ture. In Coupled Processes Associated with Nuclear Waste
Repositories, C.F. Tsang, ed. Academic Press, Orlando,
Florida, pp. 249-257
Cederberg, G. A., R. L. Street, and J. O. Leckie. 1985. A
Groundwater Mass Transport and Equilibrium Chemistry
Model for Multicomponent Systems. Water Resources Re-
search 21:1095-1104.
Chang, H. T, and B. E. Rittmann. 1987a. Mathematical
Modeling of Biofilm on Activated Carbon. Environ. Sci.
Technol. 21(3):273-280.
Chang, H. T, and B. E. Rittmann. 1987b. Verification of the
Model of Biofilm on Activated Carbon. Environ. Sci. Tech-
nol. 21(3) 280-288.
Collins, A. G., and M. E. Crocker. 1988. Laboratory Protocol
for Determining Fate of Waste Disposed in Deep Wells.
EPA Report 600/8-88/008. National Institute for Petroleum
and Energy Research, Bartlesville, OK.
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CHAPTER TWO
RESEARCH SUMMARY NO. 1
STATE-OF-THE-ART REPORT: INJECTION OF HAZARDOUS WASTES INTO DEEP WELLS
\ formed in the compatibility test, may or may not plug
the well depending on the type formed.
2.1 Overview
2.1.1 Origin and Content .
Source: State-of-the-Art Report Injection of Hazardous
V&stes into Deep Wblls. EPA/600/8-87/013, February
1987. Prepared for U.S. Department of Energy and U.S.
Environmental Protection Agency. 55 pages. NTISPB87-
170551.
Authors: Arden Strycker and A. Gene Collins, National
Institute for Petroleum and Energy Research, RO. Box
2128, Bartlesville, Oklahoma 74005.
Contents: Literature review of 82 sources on: (1) proces-
ses affecting the geochemtoalfate of inorganic and organic
hazardous wastes, (2) mathematical models used for
predicting fate, and (3) field case studies involving deep-
well injection of hazardous waste.
2.1.2 Major Conclusions
m Many factors affect the ultimate fate of injebted
wastes. These factors include the pH and Eh of the
waste and reservoir fluids, brine concentration of
the waste fluids, clay type and amount in the reser-
voir, presence or absence of iron oxides, presence
or absence of organic complexing agents,
molecular characteristics of organic materials, and
other factors that determine if the environment is
anaerobic or aerobic. All these factors are interre-
lated and any mixing of different types of hazardous
wastes in the reservoirfurthercomplicatesthe situa-
tion, making it difficultto predict exactly what occurs
after wastes are injected. Relevant research con-
ducted to date concerning this problem has been
limited and is not sufficient to address the problem
of predicting ultimate fate. ;
• The basic compatibility test conducted by mixing
waste fluids and reservoir fluids does not always
give meaningful results. The test must be con-
ducted under reservoir conditions. Precipitates, if
For inorganic wastes, solution pH is critical for
determining the ultimate fate. The identity of
soluble species, solubility products, adsorption
characteristics, and chemical interactions are some
of the variables affected by pH.
The brine concentration, even though not listed as
hazardous, can affect clay stability and adsorption
characteristics.
The presence of organic complexing agents may or
may not affect the mobility of heavy metals in the
reservoir.
Adsorption of inorganic wastes depends on a. num-
ber of factors, such as Eh, pH, clay type, and the
presence or absence of iron oxide and hydroxides.
Hydrolysis is the major mechanism for degradation
of certain halogenated hydrocarbons.
Microbial degradation processes can transform
hazardous wastes after deep-well injection, but
results are not always predictable.
2.2 Processes Affecting the Geochemicat
Fate of Deep-Well-Injected Wastes
2.2.1 Overview of Pate-Influencing Processes
Subsurface reservoirfluids have equilibrated with reservoir
minerals and clays during geologic time. All the minerals,
rocks, hydrocarbons, and gases are interrelated and con-
tribute to the final stable solute/solvent matrix that exists in
the reservoir. On the other hand, waste solutions con-
sidered for deep-well injection are generated in a different
environment and have attained a thermodynamic equi-
librium under different conditions. Consequently, when
22
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wastes are injected into the formation, adjustments must
occur in the reservoir before a new solute/solution equi-
librium is reached.
2.2.2 Partition Processes
Acid/Base Reactions. See discussion of pH in Section
2.3.2.
Adsorption-desorption. Adsorption is a major
mechanism affecting mobility of organic wastes. Factors
affecting the degree of adsorption of a chemical include:
molecular shape and configuration, pH, water solubility,
charge distribution, polarity, molecular size, and
polarizability. Molecular shape may increase or decrease
adsorption energies of any particular compound even
though the other chemical properties may be very similar
(Bailey and White, 1970).
Adsorption mechanisms for organic chemicals include:
ion exchange, protonation at the silicate surface, protona-
tion in the solution phase with subsequent adsorption by
ton exchange, and protonation by reaction with the disas-
sociated protons from residual water present on the sur-
face or in coordination with the exchangeable cation; or-
ganic cations are most easily adsorbed by ion exchange
and the process is similar to that for inorganic materials
(Bailey and White, 1970). Other adsorption mechanisms
include Van der Waals forces, hydrogen bonding, and the
formation of metal complexes.
Ion exchange is a very common adsorption mechanism for
organic wastes (Bailey and White, 1970). At high salinity
levels, ion-exchange rates tend to be slow (Veley, 1969).
Complex polyvalent metal ions tend to adsorb strongly to
clay particles so that ton exchange is less extensive than in
clays with adsorbed calcium and sodium particles (Veley,
1969). The effect of complexing agents on adsorption
depends on energy levels in relationship to ion-exchange
processes; slight difference in conditions may have a major
impact on the overall results (Champlin, 1969).
Adsorption processes in soil depend on: the structural
characteristics of the molecule, organic content of the soil,
pH of the medium, particle size, ion-exchange capacity,
and temperature. Generally, as the solubility of the adsor-
bate decreases, adsorption increases, and as organic
content of soil increases, adsorption increases (Haque et
al., 1980).
Once materials are adsorbed, other processes that lead to
their degradation may take place: microorganisms may
metabolize wastes and clays, or minerals attached to clays
may catalytically initiate other such reactions. Mortland
(1985) discusses organic-adsorption mechanisms that in-
clude: replacement of metals with cationic molecules,
replacement of metals by neutral molecules that are
protonated to become cationic, ton exchange with
polyvalent metals attached to the clay, coordination with
metal cations, and hydrogen bonding. This last occurs
when esters hydrolyze (McAuliffe and Coleman, 1955).
Other examples of adsorption are also discussed bv
Mortland (1985).
Temperature affects adsorption. Since adsorption proces-
ses are generally exothermic and desorption generally
endothermic, an increase in temperature would normally
reduce adsorption processes. EPTC (Theis et al., 1980)
and pentachlorophenol (Choi and Aomine, 1974a,b) are
exceptions.
The relative energies associated with complexing agents,
ion-exchange processes, and adsorption processes will
affect the degree of adsorption. Energies associated with
some chelating agents are approximately the same as
those of cation-silicate interactions, in which case slight
differences in conditions may have a major impact on the
amount of adsorption (Champlin, 1969). Chelating agents
that enhance the solubility of uranium, cobalt, strontium,
and cesium do not necessarily decrease adsorption. In the
presence of clays, cobalt is less strongly adsorbed, stron-
tium and cesium are not affected, and uranium is more
strongly adsorbed. The higher uranium/acid-complex ad-
sorption may be caused by additional electrostatic and
molecular dipole-attractive forces (Means, 1982).
The amount of organic and inorganic materials adsorbed
depends on the amount and type of clay present in the
formation because different clays have different surface
areas and charge densities. Since clays possess an over-
all negative charge, cations such as moderately soluble
metal wastes are attracted to these clays. The more-
soluble ions previously attached to the clays may resolubil-
ize when other, less-soluble ions replace them on the clay
surface, i.e., ion exchange (Wilson, 1980). Two types of
clay, montmorillonite and vermiculite, have very high ad-
sorption capacities, whereas kaolinfte has a very taw
capacity and illite and chlorite have intermediate
capacities. The adsorptive properties of clay have been
attributed to the available surface area for the respective
clays (Bailey and White, 1970). The interactions between
waste fluids and formation clays have been difficult to
characterize.
Clays become saturated with a particular ion when all the
adsorption sites are filled; no further adsorption can occur.
The saturation level depends on the amount and type of
clay and whether iron and manganese oxides are present
as additional adsorption surfaces. The net negative
23
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charge of clays comes from the replacement of aluminum
and silicon tons With other tons having tower oxidation
states within the structure of the clay. Different clays be-
have in different ways because throughout a clay or be-
tween different clays ton replacement is not uniform, with
consequent variations in the degree of negative charge
(Scrivner et al., 1986a,b). Low-salinity or high-pH solutions
cause water molecules to be adsorbed. When adsorbed,
they separate the crystal layers of clay, causing it to swell.
The realigned layers usually will not return to their original
state even if higher-salinity solutions are again added.
Complex (polyvalent) metal tons adsorb strongly to clay
particles. Strong adsorption behavior may immobilize me-
tals and protect clays from swelling and fines migration. At
high salinity levels, complex metal ions that are adsorbed
onto unattached.fine clay particles may migrate in suspen-
sion (Champlin, 1969).
Certain metals (particularly heavy metals) associated with
the clay may bond so tightly that they may be considered
immobile or permanently adsorbed. On the other hand,
metals that do not adsorb tightly may desorb at a later time
when a different waste is injected.
Anumberof experiments showthe variability of adsorption:
• Choi and Aomine (1974a,b) report that ion ex-
change and Van der Waals forces were the adsorp-
tion mechanisms acting on pentachloro- phenol.
The degree of adsorption was highly pH-depehdent
and the temperature effects were not those an-
ticipated.
• O'Connor etal. (1985) report that lesser amounts of
trichloroethylene and pentachlorophenol adsorbed
on Missouri soils as pH increased.
i
• Rogers and McFarlane (1981) report that adsorp-
tion of carbon tetrachloride, ethylene dibromide,
and chloroform on montmorillonite clay depends on
the degree of saturation by such cations as ca)cium
and aluminum.
• Schwarzenbach and Giger (1985) report that at
near-surface conditions adsorption of chlorinated
benzenes increases as organic carbon increases.
See Section 2.5.2 for discussion of methods for predicting
adsorption, and the Wilmington, North Carolina, case
study (Section 2.7.2) foran example of adsorption involving
injected wastes. :
Precipitation-dissolution. Reactions between injected
and interstitial fluids can produce precipitates in deep
wells: (1) alkaline earth metals (calcium, barium, strontium,
and magnesium) can precipitate as insoluble carbonates,
sulfates, orthophosphates, fluorides, and hydroxides; (2)
other metals (such as iron, aluminum, cadmium, zinc,
manganese, and chromium) can precipitate as insoluble
carbonates, bfcarbonates, hydroxides, orthophosphates,
and sulfides, and (3) oxidation-reduction reaction
products, such as hydrogen surfkie with chromium (VI),
may precipitate (Warner, 1966; Selm and Hulse, 1960).
Ferric hydroxide, which is gelatinous, appreciably blocks
the flow of fluids through a porous matrix; barium sulfate
and calcium sulfate, which are finely crystalline, do not
(Warner, 1966). A buffer zone of nonreactive water may
prevent plugging due to precipitation (Warner, 1966).
Precipitation reactions are more sensitive to temperature
than pressure (Grubbs et al., 1972). Reeder et al. (1975)
and Elkan et al. (1975) discuss an injection well in the
Arbuckle formation that is an example of a well becoming
plugged with precipitates. Scale indices have been
developed to predict potential precipitation problems
(Browne, 1984).
Above pH 10, calcium, barium, strontium, magnesium, and
iron will form gelatinous hydroxide precipitates. Lower-pH
solutions containing btearbonates will convert to car-
bonates if the pH is raised and precipitates of iron, calcium,
and magnesium carbonates result.
High-pH solutions can dissolve silica and release fines that
may migrate and plug pores. Re-precipitation of dissolved
silica in another section of the reservoir may reduce per-
meability (Thornton and Radke, 1985; Thornton and
Lorenz, 1987). Certain tow-pH solutions initially may leach
some formation minerals; the solutions may also cause
other minerals to precipitate and reduce permeability
rather than increase it (Grubbs et al., 1972). Low-pH solu-
tions may lead to the formation of silica gels or the dissolu-
tion of some clays and carbonates (either as a matrix or as
cements); these problems are not as evident in carbonate
formations, but later deposition of materials caused by
changes in pH may also be a problem in carbonate forma-
tions.
Toxtoity is a function of solubility, and solubility determines
the relative mobility of materials. The more soluble the
metal, the greater the rate of transport and the greater the
magnitude of toxicity. Simple solution properties such as
pH and Eh affect solubility. Concentrations also affect
solubility; when ferric ton is present in low concentrations,
natural organics such as humic acids form true solutions
due to organometallic complexes. Pentachlorophenol
24
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precipitates in solutions at pH less than 5 (Choi and
Aomine, 1974a,b).
An injection well in the Magothy aquifer in New York ex-
hibited a nearly 10-fold increase in ferrous tons over either
injected or native fluids. Apparently the underground en-
vironment changed from a reducing to an oxidizing en-
vironment, leading to the dissolution of pyrite (iron sulfide).
The ferrous tons produced by this process precipitated as
ferric hydroxide in the presence of oxygen (Ragone et al.,
1973). The chemistry of these reactions is complicated by
the pH and Eh of the solution, the presence of Fe+ 3, partial
oxygen pressure, and the presence of organics. See the
Wilmington, North Carolina, case study (Section 2.7.2) for
examples of precipitation involving injected wastes.
2.2.3 Transformation Processes
Biological Transformation. See discussion of microbial
degradation in this section,
Complexation. Organic chemicals can form complexes
with metals that increase the solubility of the metal (Means
and Hubbard, 1985; Means, 1982; Francis, 1985). The
solubility of most metals is much higher when they are in
the form of organometallic complexes. Naturally occurring
chemicals that can partially complex with metal compounds
and increase their solubility include: aliphatic acids,
aromatic acids, alcohols, aldehydes, ketones, amines,'
aromatic hydrocarbons, esters, ethers, and phenols
(Means and Hubbard, 1985). Natural organics such as
humic acids form solutions of organometallic complexes
when the ferric-ion concentration is low. At higher con-
centrations, colloidal suspensions are formed from the
same humic acids, which may reduce intermolecular repul-
sion forces in the metal-complexed molecule. As a result,
the organic material may recoil and become less hydrated
in the solution, since all of the polar sites are taken up by
the metal, and may remain suspended or precipitate,
depending on the particle size (Means, 1982). Bacteriacan
degrade the organic components of organometallic com-
plexed particles and may also convert some non-com-
plexing materials into complexing agents. Depending on
the conditions, mobility of metals in this situation may be
increased or decreased (Francis, 1985). See the Wil-
mington, North Carolina, case study (Section 2.7.2) for an
example of complexation involving injected wastes.
Cyclization. See discussion of thermal degradation in this
section.
Hydrolysis. Hydrolysis is the chemical process by which
a functional group attached to a molecule is replaced by
an -OH functional group originating from a water molecule.
Potentially, hydrolysis can either detoxify organic hazard-
ous waste, rendering it nonhazardous, or increase the
toxicity of certain wastes. For organic materials, such
factors as pH, temperature, and the presence of other tons
affect the rate of hydrolysis. At tow pH the hydronium ton
predominates and at high pH the hydroxide ton is more
prevalent. The magnitude of temperature effects on dif-
ferent compounds is not always known.
Hydrated polyvalent metal tons hydrolyze and form multiple
associations with other metals. When these complex
polynuclear tons associate with clay particles, a very tight
structure forms around the clay crystal, and months or
years may be required before true equilibrium is reached
among all these different metal associations.
Hydrolysis can be catalyzed by either an acid or a base.
The presence of certain alkaline earth and heavy-metal ions
may also catalyze hydrolysis for a variety of esters (Mabey
and Mill, 1978). Most hazardous wastes that potentially
can undergo hydrolysis reactions are hatogenated hydrocar-
bons. Since these compounds are not normally
biodegradable, hydrolysis is expected to be the main
mechanism of transformation. Hydrolysis half-lives at con-
ditions that exist near the surface range from days to
thousands of years (see discussion of carbon tetrachloride,
ethylene dibromide, and chloroform in Section 2.4.2). Sub-
surface environments, with their increased temperatures
and pressures and reduced Eh, may contribute to shorter
half-lives.
Aliphatic and alkylic halidescan hydrolyze under neutral or
basic conditions to give alcohols, but these compounds
are not likely to undergo the same process under acidic
conditions. Different halides (phenyl dichloromethane,
dichloromethane, and chlorobenzene) have very different
hydrolysis rates. Section 2.5.2 discusses how to predict
hydrolysis half-lives.
Microbial Degradation. Biodegradation can result from a
variety of processes (Alexander, 1980,1981; Crosby, 1973).
Biological transformation may render organic hazardous
wastes nonhazardous but for certain wastes actually in-
crease toxicity. Biodegradation processes include:
mineralization (conversion of organic to inorganic wastes),
detoxification (conversion of toxic compounds to nontoxic
compounds), co-metabolism (conversion of one organic
compound to another without the microorganism's using
this process as a nutrient), activation (conversion of a
nontoxic compound to a toxic compound), and defusing
(conversion of a compound capable of becoming hazard-
ous to another, nonhazardous compound by circumvent-
ing the hazardous intermediate). Defusing has been ob-
served in the laboratory but not identified in the environ-
ment.
25
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Bacteria can degrade the organic components of or-
ganometallic complexed particles and may also covert
some noncomplexing materials into complexing agents.
Depending on conditions, mobility of metals may be in-
creased or decreased (Francis, 1985).
Whether the environment is aerobic determines what
bfodegradatfon process will predominate. Some wastes
are more easily degraded by aerobes (e.g., chtoroben-
zenes) and others are more easily degraded by anaerobes
(e.g., carbon tetrachtoride) (Jackson etal., 1985). Aerobfc
degradation is usually more efficient than anaerobic
degradation, and higher temperatures are not so limiting
for aerobes. Some compounds such as aromatfcs can be
degraded only by aerobes (Grula and Grula, 1976).
Anaerobic degradation is a muftistep process in which
complex compounds are broken down by certain faculta-
tive bacteria, and the resulting short-chain acid anfons are
broken down by methanogento bacteria (Novak ^and
Ramish, 1975). Other mechanisms for anaerobic degrada-
tion have been proposed (Means and Hubbard, 1985). In
anaerobic conditions, pH affects the extent to which
methanogento bacteria or sulfate-reducing bacteria
proliferate. "The two types of bacteria do not degrade the
same compounds (Horvath, 1977).
A number of studies have looked at the effects of environ-
mental conditions on microorganisms. Christofi et al.
(1985) report that microorganisms exist in water samples
taken from underground coal mines in Germany at all
depths (600 to 3,000 ft) but greatest variety was found in
the least-saline aquifer. Several studies indicate that, in
general, growth and reproduction processes of bacteria
occurring at near-surface conditions decrease with in-
creasing pressures to 600 atmospheres (about 8,800 psi),
whereas bacteria isolated from the marine environment at
depths normal to these same pressures grow wjsll in
laboratory studies (ZoBell and Johnson, 1949; ZoBell and
Oppenheimer, 1950; ZoBell and Cobet, 1962; Morita and
ZoBell, 1956). The effects of high pressures on microor-
ganism metabolic rates are unknown at present. At least
one study has looked at the effects of temperature on
microorganisms. In this study, aliphatic acids (acetate
ions) were found to be degraded by methanogenic, bac-
teria in oilfield waters with temperatures tower than 80°
(Carothers and Kharaka, 1978).
When hazardous-waste injection begins, other microor-
ganisms that can utilize the waste often appear and remain
in the reservoir during injection. After organic wastes are
injected, the reservoir tends to become more anaerobic
and dominated by methanogento and sulfate-reducing
bacteria (Elkan and Horvath, 1975; Elkan, 1975).
Laboratory model studies found microbial populations in-
creased by seven or more orders of magnitude when waste
was introduced into the model. Degradation of formic acid
in the laboratory model increased as pressure increased
to 500, but decreased when pressures were increased to
4,000 psi (Grula and Grula, 1976). Before injection at one
site, aerobic bacteria (3,000 organisms/m) dominated a
saline aquifer at 850-1,000ft. Afterinjection of mostly acidic
wastes (acetic acid, formic acid, and methanol), anaerobic
methanogento bacteria were predominant (DiTommaso
and Elkan, 1972).
Some compounds (certain chlorinated alkanes and
alkenes) are not degraded in the materials normally found
in deep subsurface environmenSs; others are readily
degraded (toluene and styrene), although not equally in
different environments (Wilson et al., 1985). Consequently,
Wilson et al. (1985) recommend that biodegradation
should not be depended on for waste degradation unless
the particular waste has been tested in the materials en-
countered and at the likely downhole conditions. Naph-
thalene and heptaldehyde may be degraded, whereas
hatoforms are not (Rittman et al., 1980). Horvath (1977)
has summarized processes involving biodegradation of
acetate, formate, methanol, formaldehyde, and aromatic
acids. See the Wilmington, North Carolina, case study
(Section 2.7.2) for an example of microbial degradation
involving injected wastes.
Neutralization. See the Wilmington, North Carolina, case
study (Section2.7.2)foran example of neutralization involv-
ing injected wastes. See also pH effects.
Thermal Degradation. Thermal degradation processes
include pyrolysis, condensation reactions, cyclization, and
intramolecular rearrangements. Most of these processes
occuronly under very high temperatures or in the presence
of other chemicals. Reservoirtemperaturesand pressures
commonly existing in the injection zones of hazardous-
waste-injection wells are normally too tow for initiating
high-temperature reactions, but if the right chemicals (not
necessarily hazardous) are present, thermal degradation
might be initiated. For example, phenols can react with
formaldehyde to form phenolic resins. The number and
types of these reactions are almost limitless; each reservoir
and waste should be evaluated individually. Thermal
decarboxylizatfon is probably the mechanism for acetate
degradation in oilfield waters with temperatures greater
than 200°C (Carothers and Kharaka, 1978).
2.2.4 Ttensport Processes
Dilution. Dilution by mixing with other waters contributes
to tow concentrations of aliphatic acids (acetate) in oilfield
26
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waters at temperatures of less than 80°C (Carothers and
Kharaka, 1978).
Dispersion. Mixing of fluids and precipitation reactions
depend on hydrodynamic- dispersion properties (Warner,
1966). Advective and dispersive properties must be
measured in the underground environment to predict ad-
sorption processes (Roberts et al., 1985).
Fluid Migration. There are at least five ways a waste
material may migrate and contaminate potable ground
water (mter VfellJournal, 1974).
1.Wastes may escape through the well bore into a un-
derground source of drinking water (USDW) because of
insufficient casing orfailure of the injection-well casing due
to corrosion, excessive injection pressure, etc.
2.Waste may escape vertically outside the well casing
from the injection zone.
S.Waste may escape vertically from the injection zone
through confining beds that are inadequate because of
high primary permeability, solution channels, joints, faults, or
induced fractures.
4.Wastes may escape vertically from the injectbn zone
through nearby wells that are improperly cemented or
plugged or that have insufficient or leaky casing.
S.Wastes may contaminate a USDW directly by lateral
travel of the injected waste water from a region of saline water
to a region of fresh water within the same aquifer.
Particle Migration. Clay swelling and clay-particle migra-
tion are possible with any injected fluid. In secondary
and/or tertiary petroleum-recovery operations, engineers
usually avoid injecting alkaline solutions and sometimes all
aqueous solutions when water-sensitive clays are present.
Damage to clays can result in drastically reduced per-
meabilities, and to prevent damage to reservoirs, special-
ized products that stabilize clays are often used to treat
injected fluids.
Low concentrations of salts can lead to clay migration, and
high-pH solutions tend to dissolve silica and release fines
that can migrate and plug pores, reducing permeability
(Hower et al., 1972). Low salinity leads to the loosening of
the clay structure, with swelling (usually irreversible) and
migration (Veley, 1969). Sodium ions bind less strongly to
clay than do calcium ions and are more likely to result in
reduction in permeability from clay swelling and migration.
Complex metal ions may bind so strongly with clay particles
that there will be little ton exchange with sodium, so that
swelling and migration are less likely under reduced-
salinity conditions (Veley, 1969). Complex metal ions that
are adsorbed onto very small particles of clay may migrate
as metal-clay particles depending on the physical forces
affecting particle-particle interactions, particle-matrix inter-
actions, and gravitational effects (Champlin, 1969).
Laboratory flow experiments found that at low salinity levels
in a sand core, ton-clay particles were retained by the sand,
whereas at high salinity levels ton-clay particles passed
through the core (Champlin, 1969).
Oxidation-Reduction. Oxidation-reduction (redox) proces-
ses can render organic hazardous waste nonhazardous
but increase toxteity for certain wastes. The exact species
of oxygen radicals in aqueous and soil environments that
initiate oxidation will depend on environmental conditions;
the importance of this process at typical conditions for
hazardous organic-waste injection has not been
evaluated. Changes in oxidation state may render metals
nonhazardous. Compounds such as phenols, aromatic
amines, olefins, dienes, alkyl sulfides, and eneamines are
particularly susceptible to oxidation reactions (Mill, 1980).
Oxidation is more likely to be important in wastes contain-
ing chromium (VI). See the Wilmington, North Carolina,
case study (Section 2.7.2) for an example of reduction of
injected wastes.
2.3 Major Environmental Factors Affecting
Deep-Well-lnjection Geochemical Processes
2.3.1 Geochemical Characteristics of Deep-Well Zones
Atypical injection well might be described as having the
following characteristics. The well is 3,925 ft deep with an
injection zone more than 200 ft thick. The injection zone is
composed of sandstone/sand/silt, the confining zone of
clay/shale. The median wellhead pressure and injection
flow are 285 psig and 150 gallons per minute (Huff, 1986).
Most facilities treat the waste before injection, with common
pretreatment including solids removal, equalization, and
pH adjustment. About 96 percent of the total volume
injected is water.
2.3.2 Specific Environmental Factors
pH. When injected and reservoir solutions have different
pH values, plugging problems can develop. The pH of
these solutions is important because ion concentrations
are linear functions of the fluid proportions, but the equi-
librium constants are not. For example, injected and inter-
stitial fluids that are saturated with carbonate may be
incompatible due to different pH values. Above pH 10,
calcium, barium, strontium, magnesium, and iron will form
gelatinous hydroxide precipitates. Lower-pH solutions
27
-------
containing bicarbonate will convert to carbonates if the, pH
is raised, and iron, calcium, and magnesium carbonates
may precipitate (Barnes, 1972).
Certain tow-pH solutions initially may leach some formation
minerals but may also cause other minerals to precipitate,
actually reducing permeability rather than increasing it
(Grubbs et al., 1972). High-pH solutions tend to dissolve
silica and release fines that may migrate and plug pores
(Hower et al., 1972). Low-pH solutions may produce sjlica
gels or dissolve some clays and carbonates (either in the
matrix or in the well-casing cement). TTie pH of a solution
strongly influences the formation of organometallic cbm-
plexes. High-pH solutions tend to cause clays to swell.
Arsenic cations are more mobile than selenium, cadmium,
and lead under anaerobic conditions when pH is neutral to
alkaline (Fuller, 1977). At higherpH values, cadmium exists
as various hydroxides (Fuller, 1977). In acidic solutions,
pentachtorophenol concentrations decrease by precipita-
tion; when pH values are greater than 5, the concentration
decreases in the presence of clay because of adsorption.
O'Connor et al. (1985) found that lesser amounts of
trichtoraethylene and pentachtorophenol adsorb on Mis-
souri soils as pH increases.
Hydrolysis rates are also affected by pH. Aliphatic jand
alkylto halides can hydrolyze under neutral or basic condi-
tions to give alcohols, but they are not likely to hydrolyze
under acidic conditions. Under anaerobic conditions, pH
strongly affects whether methanogenic or sulfate-reducing
bacteria predominate (Horvath, 1977).
Temperature. Compatibility testing of solutions should be
kept at reservoir temperature. Temperature is generally a
more important factor than pressure in causing plugging
when incompatible fluids are present (Grubbs et al., 1972).
Subsurface temperatures affect the presence or absence
of acetate tons in formation water samples in California and
Texas (Carothers and Kharaka, 1978). The effect of
temperature on the solubility of thorium sulfides depends
on the hydrate being tested (Goldschmidt, 1958). Adsorp-
tfon processes are generally exothermic, so an increase in
temperature normally reduces adsorption, but a number
of exceptions have been found (Theis et al., 1985). The
pesticide EFTC and pentachtorophenol are examples of
substances that are adsorbed more easily at higher
temperatures (Theis et al, 1985; Choi and Aomine, 19j74a).
Temperature also influences hydrolysis rate, but the mag-
nitude of this influence on different compounds is not
always known. Increased temperatures in subsurfacja en-
vironments may lead to shorter hydrolysis half-lives for
organic materials.
Pressure. Compatibility testing of solutions should be
kept at reservoir pressure. Pressure is generally less im-
portant than temperature in causing plugging when incom-
patible fluids are present (Grubbs et al., 1972). Increased
pressures in subsurface environments may lead to shorter
hydrolysis half-lives for organic materials. Degradation
rates of formic acid increase as pressure increases to 500
psi but decrease when pressures are further increased, to
4,000 psi (Elkan and Horvath, 1977; Elkan, 1975). In
general, bacterial growth and reproduction decrease with
increasing pressure to 600 atmospheres (about 8,000 psi)
except for bacteria isolated from deep marine environ-
ments, which are adapted to high pressure (ZoBell and
Cobet, 1962; ZoBell and Johnson, 1949; ZoBell and Op-
penheimer, 1950; Morita and ZoBeSI, 1956).
2.4 Geochemical Characteristics and Fate of
Hazardous Waste
A 1983 EPA survey of 108 active hazardous-waste wells
found that most of the wastes categorized as hazardous
contained either acid solutions or organic materials (U.S.
EPA, 1985). The report by Callahan et al. (1979) provides
a good summary of the expected fate of 129 nonorganto
and organic hazardous-waste compounds. Although that
report addresses the aquatic environment, and not deep-
well-injection zones, the information is useful.
2.4.1 Specific Data on Inorganic Substances
Alkaline Earth Metals. Calcium, barium, strontium, and
magnesium may react with injected fluids and precipitate
as insoluble carbonates, sulfates, orthophosphates,
fluorides, and hydroxides.
Alkaline Solutions. Alkaline solutions injected into reser-
voirs containing water-sensitive clays can drastically
reduce permeabilities.
Arsenic. Arsenic is generally more mobile under
anaerobic than aerobic conditions. It is more mobile than
selenium, cadmium, and lead under aerobic conditions
when the pH is neutral to alkaline. Some microorganisms
can convert arsenic hydroxide to an organic compound
(Fuller, 1977). Certain forms react with limestone to
produce carbon dioxide; certain forms adsorb onto shales
with varying amounts of clay (Stone et al., 1975).
Barium Sulfate. Barium sutfate do as not appreciably block
the flow of fluids through a porous matrix (Warner, 1966).
Cadmium. Cadmium may react with injection fluid and
precipitate as insoluble carbonates, bicarbonates,
28
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hydroxides, orthophosphates, and sulfides. This metal is
less soluble than many others, but more so than lead. At
higher pH values, mixtures of hydroxides will exist in various
concentrations (Fuller, 1977; Jurinak and Santillan-Medran,
1974; Fugii, 1978; Stone et al., 1975). The presence of
oxides and hydrous oxides of iron increases adsorption
properties, and maximum adsorption of cadmium depends
on Eh and pH (Ku et al., 1978). Some forms of cadmium
react with limestone to produce carbon dioxide and adsorb
onto certain shales containing varying amounts of clays
(Stone et al., 1975). Cadmium will precpitate as cadmium
sulfide in the presence of hydrogen sulfide.
Calcium Sulfate. This compound does not appreciably
block the flow of fluids through a porous matrix (Warner,
1966).
Cesium. The solubility of chelated organic compounds
containing cesium is not affected by the presence of clays
(Means, 1982).
Chromium. Chromium (VI) is an excellent oxidizing agent.
This compound reacts with hydrogen sulfide to form a
precipitate. Chromates in a waste stream reacting with
barium sulfate, hydrogen sulfide, and soluble iron form
precipitates that eventually plugged a well in the Arbuckle
formation (Reeder et al., 1975). Under neutral-to-alkaline
conditions, chromium (VI) is more mobile than most of the
metals listed. Under oxidizing conditions in soils,
chromium (VI) will form insoluble precipitates with resident
biological materials. Chromium (III) can adsorb strongly in
acid solutions and will precipitate at pH values above 6 (as
hydroxide, carbonate, or sulfide), whereas chromium (VI)
does not (Fuller, 1977). Chromium (VI) can adsorb onto
oxides and hydrous oxides of iron (Ku et al., 1978).
Cobalt. Solubility of chelated organic compounds contain-
ing cobalt increased in the presence of clays (Means,
1982).
Ferric Hydroxide. This compound appreciably blocks the
flow of fluids through a porous matrix (Warner, 1966).
Lead. The solubilities for lead are lower and the tendencies
to adsorption higherthan for most of the other metals listed
(Fuller, 1977). Typical precipitates are lead hydroxide and
lead carbonate (Jurinak and Santillan-Medran, 1974).
Sodium chloride somewhat increases the solubility of lead
(Stone etal., 1975).
Mercury. Bacteria can convert inorganic mercury com-
pounds to the more toxic and volatile dimethyl mercury
(Fuller, 1977). A major problem in soil environments is the
volatility of mono- and dimethyl-mercury compounds
(Stone et al., 1975). This metal strongly adsorbs onto iron
oxides if present. Whether byconversion and hazardous
volatilization of mercury occur in injection zones is not
known.
Nickel. Nickel adsorbs strongly in the presence of iron and
manganese oxides. It is not very soluble in the presence
of carbonates, hydroxides, or sulfides (Fuller, 1977).
Soluble salts include nickel acetate, chloride, nitrate, and
sulfate (Stone et al., 1975). Nickel oxides may be solubil-
ized in strong acid, but as strong acids are neutralized in
the reservoir, these oxides may then precipitate. Nickel
carbonyl is very toxic and potentially explosive when con-
centrated; it is stable in dilute acid or basic solutions but
will produce carbon monoxide and nickel metal when
heated (Ku etal., 1978).
Selenium. Some selenium compounds adsorb more
strongly in the presence of iron oxides (Nebergall et al.,
1968). Selenium dioxide is readily soluble in water and
forms selenous acid in aqueous solutions (Partington,
1966). Many selenium compounds can be reduced to
produce selenium metal when exposed to organic matter
in the subsurface environment (Goldschmidt, 1958).
Strontium. The solubility of chelated organic compounds
containing strontium is not affected by the presence of
clays (Means, 1982).
Thorium. Thorium salts are not very soluble in neutral-pH
natural waters. Soluble salts include sulfates, chlorides,
and some sulfides, but as the solution becomes basic,
these salts precipitate as hydroxide (Goldschmidt, 1958).
The effect of temperature on solubility of thorium sulfides
depends on the hydrate: some hydrates increase in
solubility with increasing temperature, others decrease
(Goldschmidt, 1958).
Uranium. The solubility of chelated organic compounds
containing uranium decreased in the presence of clays
because of adsorption (Means, 1982).
2.4.2 Specific Data on Organic Substances
Acetic Acid/Acetate. Acetic acid and acetate are present
in oilfield waters (Carothers and Kharaka, 1978). Thermal
decarboxylatfon occurs at temperatures greater than
200°C (Kharaka, 1978). Microbial degradation by meth-
anogenfc bacteria occurs at temperatures less than 80°C
(Means and Hubbard, 1985). Horvath (1977) summarizes
biodegradation processes for these compounds. An-
aerobic methanogenic bacteria replaced aerobic bacteria
after injection of wastes containing acetic acid, formic acid,
and methanol (DiTommaso and Elkan, 1973). See the
Wilmington, North Carolina, case study (Section 2.7.2) for
29
-------
an example of the degradation of acetic acid in an injection
zone.
Acidic Wastes. See Belle Glade, Florida, case study (Sec-
tion 2.7.1).
Alcohols. Alcohols partially complex with metal com-
pounds (Means and Hubbard, 1985). The effects of en-
riched cultures of microorganisms, temperature, pres-
sures, and mixed cultures were studied by Grula and Gryla
(1976).
Aldehydes. Aldehydes partially complex with metal com-
pounds (Means and Hubbard, 1985). The effects of ^n-
riched cultures of microorganisms, temperature, pres-
sures, and mixed cultures were studied by Grula and Grula
(1976).
Aliphatic Acids. Aliphatic acids partially complex with
metal compounds (Means and Hubbard, 1985).
Aliphatic Halldes. Aliphatic halides can hydrolyze under
neutral or basic conditions to give alcohols, but this reac-
tion is not likely under acidic conditions. The hydrolysis
rate depends on the type of halide.
AlkyI Sulfldes. Alkyl sulfides are particularly susceptible to
oxidation (Mill, 1980).
Alkyltc Halldes. Alkylic halides can hydrolyze under
neutral or basic conditions to give alcohols, but this reac-
tion Is not likely under acidic conditions. The hydrolysis
rate depends on the type of halide. !
Amines. Amines partially complex with metal compounds
(Means and Hubbard, 1985). Aromatic amines are par-
ticularly susceptible to oxidation (Mill, 1980). Theeffectsof
enriched cultures of microorganisms, temperature, pres-
sures, and mixed cultures were studied by Grula and Grula
(1965), :
Aromatic Acids. Aromatic acids partially complex with
metal compounds (Means and Hubbard, 1985). Horyath
(1977) summarizes the bfodegradation processes affect-
ing aromatic acids.
Aromatic Compounds. Aromatic compounds can be
degraded only by aerobic bacteria; higher temperatures
are not a limiting factor for bfodegradation of these com-
pounds.
Aromatic Hydrocarbons. Aromatic hydrocarbons partial-
ly complex with metal compounds (Means and Hubbard,
1985).
Benzoic Acids. The effects of enriched cultures of
microorganisms, temperature, pressures, and mixed cul-
tures were studied by Grula and Grula (1976). See also the
Wilmington, North Carolina, case study (Section 2.7.2).
Carbon Tetrachloride. Carbon tetrachtoride, a halo-
genated hydrocarbon, does not adsorb onto calcium-
saturated montmorillonite clay (Rogers and McFarlane,
1981). It is not normally biodegradable, and it has a
hydrolysis half-life of 700 to 7,000 years at near-surface
conditions; subsurface conditions maylead to shorter half-
lives with the increased temperatures, pressures, and Eh
present in this environment.
Carboxylic Acids. The effects of enriched cultures of
microorganisms, temperature, pressures, and mixed cul-
tures were studied by Grula and Grula (1976).
Chlorinated Alkanes and Alkenes. These compounds
are not bfodegraded in materials normally found in deep
subsurface environments (Wilson et al., 1985).
Chlorinated Benzenes. At the concentrations typical in
the natural environment, a linear- adsorption isotherm can
be used to represent the adsorption of chlorinated ben-
zenes in the injection zone. Adsorption increases as or-
ganic carbon increases (Schwarzenbach and Giger, 1985).
Chlorobenzene. Chlorobenzene is virtually resistant to
hydrolysis under normal circumstances.
Chloroform.. Chloroform, a halogenated hydrocarbon,
does not adsorb onto calcium-saturated montmorillonite
clay but showed 17 percent adsotption onto aluminum-
saturated montmorillonite clay. The mechanism for this
difference is not understood (Rogers and McFarlane,
1981); chloroform is not normally biodegradable, and it has
a hydrolysis half-life intermediate between ethylene
dibromide (5 to 10 days) and carbon tetrachloride (700 to
7,000 years) at near-surface conditions. Subsurface con-
ditions, with their increased temperatures, pressures, and
Eh, maylead to shorter hydrolysis half-lives.
Dichloromethane. The hydrolysis rate constant for di-
chtoromethane is about five orders of magnitude tower
than that for phenyl dichtoromethane.
Dienes. Dienes are particularly susceptible to oxidation
(Mill, 1980).
Diethylenetriaminepentaacetic Acid (DPTA). DPTA is a
chelating agent that can increase the mobility of metals in
an underground environment (Means and Hubbard, 1985;
Means, 1982; Francis, 1985).
30
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Enamines. Enamines are particularly susceptible to oxida-
tion (Mill, 1980).
EPIC. EPIC is a pesticide for which adsorption increases
with temperature (Theis et al., 1980).
Esters. Esters partially complex with metal compounds
(Means and Hubbard, 1985). Clay or minerals attached to
the clays can catalytically initiate the hydrolysis of adsorbed
esters (McAuliffe and Coleman, 1955). The presence of
certain alkaline-earth and heavy-metal bns may catalyze
hydrolysis for a variety of esters (Mabey and Mill, 1978).
Ethers. Ethers partially complex with metal compounds
(Means and Hubbard, 1985).
Ethylenediaminetetraacetic Acid (EDTA). EDTA is a
chelating agent that can increase the mobility of metals in
an underground environment (Means and Hubbard, 1985;
Means, 1982; Francis, 1985).
Ethylene Dibromide (EDB). Rogers and McFarlane
(1981) present data on adsorption of ethylene dibromide
on montmorillonite clay. This compound is normally
biodegradable and has a hydrolysis half-life of 5 to 10 days
at near-surface conditions. Subsurface conditions, with
their increased temperatures, pressures, and Eh, maylead
to shorter hydrolysis half-lives.
Formaldehyde. The biodegradation processes affecting
formaldehyde are summarized by Healy and Daughton
(1986). Phenols can react with formaldehyde to form
phenolic resins. This was the only waste organic com-
pound nofadsorbed onto aquifer mineral constituents (see
the Wilmington, North Carolina, case study, Section 2.7.2).
Formate. Biodegradation of formate is summarized by
Healy and Daughton (1986).
Formic Acid. Anaerobic methanogenic bacteria replaced
aerobic bacteria after wastes containing acetic acid, formic
acid, and methanolwere injected (DiTommaso and Elkan,
1973). See also the Wilmington, North Carolina, case
study (Section 2.7.2).
Haloforms. Hatoforms are not subject to biodegradation
(Rittman et al., 1980).
Halogenated Hydrocarbons. These compounds are not
normally biodegradable; hydrolysis is expected to be the
main mechanism of transformation. Most data on
hydrolysis are derived under conditions most likely to occur
in the near-surface environment. See also carbon
tetrachloride, ethylene dibromide, and chloroform in this
section.
Heptaldehyde. Heptaldehyde is subject to biodegrada-
tion (Rittman et al., 1980).
Ketones. Ketones partially complex with metal com-
pounds (Means and Hubbard, 1985). The effects of en-
riched cultures of microorganisms, temperature, pres-
sures, and mixed cultures on ketones were studied by
Grula and Grula (1976).
Methanol. Biodegradation processes affecting methanol
are summarized by Healy and Daughton (1986).
Anaerobic methanogenic bacteria replaced aerobic bac-
teria after wastes containing acetic acid, formic acid, and
methanol were injected (DiTommaso and Elkan, 1973).
See also the Wilmington, North Carolina, case study (Sec-
tion 2.7.2).
Naphthalene. Naphthalene is subject to biodegradation
(Rittman etal., 1980).
Nitrate. SeethePensacola, Florida (American Cyanamid),
case study (Section 2.7.3).
Nitric Acid. See the Pensacola, Florida (Monsanto), case
study (Section 2.7.4).
Nitriles. The effects of enriched cultures of microor-
ganisms, temperature, pressures, and mixed cultures on
nitrites were studied by Grula and Grula (1976).
Nitre-aromatic Compounds. The effects of enriched cul-
tures of microorganisms, temperature, pressures, and
mixed cultures on nitro-aromatic compounds were studied
by Grula and Grula (1976).
Nftrotriacetic Acid (NTA). Nitrotriacetic acid is a chelating
agent that can increase the mobility of metals in an under-
ground environment (Means and Hubbard, 1985; Means,
1982; Francis, 1985).
Olefins. Olefins are particularly susceptible to oxidation
(Mill, 1980).
Organonrtrile Compounds. See the Pensacola, Florida
(American Cyanamid), case study (Section 2.7.3).
Pentachlorophenol. This compound is adsorbed through
a combination of ion-exchange and Van der Waals forces.
A pH greater than 5 results in adsorption; adsorption
decreases with an increase in the concentration of other
salts in solution. Adsorption is greater at higher tempera-
31
-------
tures, e.g.,33'C (Choi and Aomine, 1974a,b). Precipitation
occurs at pH less than 5 (Choi and Aomine, 1974a,b).
Pentachtorophenol adsorbs on several Missouri soils more
readily than trichotorethylene. O'Connoretal. (1985) no|ted
that adsorption of pentachtorophenol decreased as pH
increased. [
Phenols. Phenols partially complex with metal com-
pounds (Means and Hubbard, 1985). They are particularly
susceptible to oxidation (Mill, 1980) and can react with
formaldehyde to form phenolic resins. The effects of en-
riched cultures of microorganisms, temperature, pres-
sures, and mixed cultures on phenols were studied by
Grula and Grula (1976).
Phenyl DIchtoromethane. The hydrolysis rate constant of
phenyl dichtoromethane is about five orders of magnitude
greater than that for dichtoromethane.
PhthaHc Acid. Adsorption of phthalfc acid does not in-
crease with a decrease in the pH of the waste (see the
Wilmington, North Carolina, case study, Section 2.7.2).
p-Tblute Acid. See the Wilmington, North Carolina, case
study (Section 2.7.2).
Sodium Thlocyanate. This compound remained unal-
tered during movementthroughthe injection zone (seethe
American Cyanamid case study, Section 2.7.3).
Terephthallc Add. See the Wilmington, North Carolina,
case study (Section 2.7.2).
Toluene. Toluene bfodegrades readily in materials normal-
ly found in deep subsurface environments; the rate varies
with conditions (Wilson et al., 1985).
Trichtoroethytene. This compound adsorbed less readily
than pentachlorophenol on several Missouri soils.
O'Connor et al. (1985) found that adsorptior? of
trichloroethylene decreases as pH increases.
Styrene. Styrene bfodegrades readily in materials normal-
ly found in deep subsurface environments; the rate varies
with conditions (Wilson et al., 1985).
2.5 Methods and Models for Predicting the
Geochemical Fate of Deep-Well-lnjected Wastes
2.5.1 Basic Approaches
Before the fate of a hazardous waste is assessed, all major
chemical and biological pathways for movement or trans-
formation must be described. The description should
predict concentration as a function of time for the original
chemical and all subsequent products.
Many factors affect the ultimate fate of injected wastes: the
pH and Eh of the waste and reservoir fluids, brine con-
centrations of the waste fluids, clay type and amount in the
reservoir, presence or absence of iron oxides, presence or
absence of complexing agents, molecular characteristics
of organic materials, and other factors that determine if the
environment is aerobic or anaerobic. All these factors are
interrelated, and any mixing of different types of hazardous
wastes in the reservoir further complicates the situation,
making it difficult to predict exactly what occurs after the
wastes are injected. Research is not sufficient to address
the problem of predicting the fate of injected wastes.
Hazardous wastes are complex mixtures and their com-
bination with other mixed waste streams increases fac-
torially the potential number of interactions; knowledge of
these interactions is limited. Further, since subsurface
environments often take many years to reach chemical and
biological equilibrium, prediction may be impossible. Ex-
amples of the difficulties are:
• A model using the simple mixing of injection fluids
and reservoir fluids does not adequately represent
the complexities that often occur. This problem is
illustrated by examples of fluids that appear incom-
patible in the laboratory but cause little trouble in
the field, while apparently compatible fluids have
plugged injection wells.
• Predicting how much waste will be adsorbed, how
long the waste will remain immobile, and underwhat
circumstances the waste will be desorbed is dif-
ficult.
• Theoretical and laboratory studies are not sufficient
to predict the transport of wastes in an underground
aquifer. The underground environment contains
variables that have not been studied extensively.
Consequently, the degree of uncertainty in modeled
predictions is large, and tracer and pilot tests in the
field must still be performed.
• Data on degradation processes are more limited for
organic wastes than for inorganic wastes because
hardly any definitive work has been done and the
number of possible interactions is much greater.
• Data on the origin of bacteria in subsurface environ-
ments, their activity levels, and the importance of
nonbiological processes are not adequate to
32
-------
predict the fate of organic wastes (Healy and
Daughton, 1986). Unless pilot studies have estab-
lished the existence of biodegradation in the sub-
surface environment, the modeler cannot depend
on these processes to detoxify waste. Further, for
each injection system, the relative importance of
organometallic interactions and the possible
presence of bacteria capable of generating com-
plexing materials should be considered in detail.
2.5.2 Specific Methods and Models
Adsorption Methods and Models. Techniques that can
be used to study and demonstrate mechanisms by which
organic chemicals are adsorbed include adsorption
isotherms, calorimetry, X-ray diffraction, UV-visible
spectroscopy, electron-spin-resonance spectroscopy,
and infrared spectroscopy (Mortland, 1985).
The Freundlich isotherm is often used to evaluate adsorp-
tion of chemical compounds to soil particles (Haque et al.,
1980), as given by:
x/m= KC"
where
x =
m =
C =
K =
n =
amount of chemical adsorbed
mass of soil
equilibrium concentration of the chemical
constant describing the extent of adsorption
constant describing the nature of adsorption
tonic compounds or compounds capable of becoming
tonic do not necessarily follow the Freundlich-isotherm
concept.
Aqueous-andSoIution-GeochemistiyModels. Schechter
et al. (1985) present a good review of aqueous-
geochemistry and solution-geochemistry models. If a par-
ticular situation can be defined as containing specified
components, equilibrium constants can be approximated
and overall results predicted. Certain thermodynamic
parameters are known for many of the materials of interest
and provide a reasonable starting point. Some of the
popular models discussed include: WATQF, SOLMNEQ,
PHREEQE, EQ3/EQ6, PATH1, MINEQLI, MINEQLI-STAN-
FORD, PHASEQL/FLOW, and REDEQL.
Biodegradation Models. Although bacteria are docu-
mented to exist in subsurface environments, other factors
necessary for predicting fate remain undetermined. These
factors include: origin of the bacteria, level of activity, and
the importance of other, nonbiological processes.
Laboratory-model studies allow the behavior of microbial
populations to be predicted when certain organic wastes
are introduced. Increases in populations of seven or more
orders of magnitude have been predicted (Elkan and
Horvath, 1977; Elkan, 1975).
A bfofilm model can be used to predict biodegradation of
some compounds. The model presumes the formation of
afilmof microorganisms that degrade the waste chemicals
nearthe point of injection. Correlation of results with model
predictions supports the concept (Rittman et al., 1980).
Hydrolysis Models. Hydrolysis half-lives for organic was-
tes that undergo this process can be estimated fairly ac-
curately if rate constants are known for the compound at
the appropriate temperature, pH, tonic strength, etc. How-
ever, the amount of data available on rate constants under
varying conditions is limited. Extrapolations made from
one temperature to another can introduce large errors
(Mabey and Mill, 1978).
Precipitation Methods and Models. Various methods for
testing the compatibility of injection fluids and reservoir
fluids have been described (Warner, 1977; Wolbach et al.,
1984; Kaufman et al., 1973; Donaldson, 1972). Generally,
two fluids are simply mixed together and allowed to stand.
The formation of precipitate usually indicates fluid-fluid
incompatibility. The solutions should be kept at the
temperatures and pressures of the reservoir. Synthetic
solutions are not as reliable as native reservoir fluids for
accurate results. Sufficient time (hours to days) for incuba-
tion should be allowed; true thermodynamic equilibrium
may be slow for some systems.
The Warner sand-pack model (Warner, 1966) forpredfcting
chemical reactions between injection wastes and reservoir
fluids includes theoretical and laboratory work. The
amount of mixing between different fluids depends on
hydrodynamic dispersion. Once the dispersive property of
aporous medium is characterized, the amount of chemical
reaction can be accurately predicted. The type of
precipitate determines the degree of permeability reduc-
tion (see discussion of ferric hydroxide, barium sulfate, and
calcium sulfate in Section 2.4.1). The laboratory model
showed that a sufficiently large buffer zone of nonreactive
water would effectively prevent a precipitation problem
(Warner, 1966).
Transport Models. Roberts et al. (1985) review mathe-
matical models for simulating waste transport in under-
ground aquifers. These models are designed to predict
the advectfon, dispersion, and adsorption of non-
degradable organic solutes. Actual field cases were ex-
amined and compared with theoretical and laboratory
33
-------
results. The study showed that adsorption can be
predicted with some accuracy, but that advective and
dispersive properties must be measured in the under-
ground environment; theoretical and laboratory studies
were nor sufficient to predict behavior.
Mills et al. (1985) summarize five models that can be used
to predict groundwater contamination. These models cal-
culate contaminant concentrations as afunctfon of time for
a given set of conditions. Two models are radial one-
dirnensfonal, one model is cartesian one-dimensional, and
two models are cartesian two-dimensional. Parameters
such as boundary conditions, aquifer dimensions, disper-
sivity coefficients, porosity, initial contaminant concentra-
tions, and retardation factors are needed.
Kayser and Collins (1986) summarize fourtypes of models
relevant to groundwater contamination from enhanced oil
recovery (EOR) or other fluids: groundwater-flow model,
solute-transport model, heat-transport model, and defor-
mation model. Each type of model is based on different
dependent variables and is particularly useful for answer-
ing different types of questions, i
2.6 Laboratory Procedures and Protocols
Detailed laboratory procedures and protocols are not dis-
cussed in this report. Section 2.5.2 (Precipitation Methods
and Models) lists references describing laboratory
methods for testing compatibility of injected fluids with
reservoir fluids.
2.7 Reid Case Studies
i
2 J.I Field Case Study: Belle Glade, Florida (McKenzie,
1976)
Waste Characteristics. The waste injected at this facility
fs a hot acidic liquid generated at a furfural plant.
Aquifer Geology/Characteristics. The injection zone is a
saline carbonate aquifer.
Injection Activities. In 1966, wastes were injected into the
towerpart of the aquifer, between depths of 1,485 and 1,939
ft. When near-surface contamination was detected, the
injection depth was increased to 2,200 ft, and laterto 3,000
ft.
Processes Observed. The injected acids were neutral-
ized by the limestone formation, resulting in higher con-
centrations of calcium, magnesium, and silica in the waste
solution. Sulfate-reducing bacteria present in the forma-
tion converted sulfates to sulfides. The hydrogen sulfide
produced by the bacteria and the subsequent decrease in
sutfate/chtoride ratio was one indication of fluid migration.
Effects of Injection. The carbonate aquifer could not
contain the hot acid wastes. Within 27 months, the effects
of wastes were detected at a shallow monitoring well in the
upper part of the aquifer. Both upward and lateral migra-
tions were indicated by a decrease in the sulfate/chloride
ratio and a corresponding increase in the hydrogen-sulfide
concentration in the observation well. Increasing the injec-
tion depth to 2,200 ft did not prevent upward migration. The
effectiveness of increasing the injection depth to 3,000 ft
has not been reported. No extensive work was performed
to determine the extent of the contamination zone.
2.7.2 Held Case Study: Wilmington, North Camlina
(Peek and Heatf7,1973; Leenheer and Malcolm, 1973;
Leenheer et al., 1976a,b)
Waste Characteristics. Organic waste derived from the
manufacture of dimethyl terphthalate was injected at this
facility. The waste was composed of acetic acid, formic
acid, p-tolufc acid, formaldehyde, methanol, terphthalic
acid, and benzoto acid, with an average dissolved organic
carbon of about 7,100 mg/L. Before injection, the waste
was neutralized to pH 4 by adding lime, resulting in a
calcium content of about 1,300 mg/L.
Aquifer Geology/Characteristics. The injection zone was
a sedimentary aquifer with saline water. Sodium chloride
was the major dissolved-solid constituent in the native
ground water, and average dissolved-solids concentration
was about 28,800 mg/L.
Injection Activities. From May 1968 to December 1972,
waste was injected at a rate of about 300,000 gal./day. The
injection zone consisted of multiple :zones ranging in depth
from about 850 to 1,000 ft. Injection was discontinued in
1972 after the operators determined that waste disposal
into the reservoir was not desirable. Monitoring of the
waste movement and subsurface environment continued
into the mid-1970s. Samples were taken from three obser-
vation wells located 1,500 to 2,000 ft from the injection
wells.
Processes Observed. A number of processes were ob-
served:
• The waste organic acids dissolved carbonate
minerals, aluminosilicate minerals, and the sesqui-
oxide coatings on the primary minerals in the injec-
tion zone.
34
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• The waste organic acids dissolved and formed
complexes with iron and manganese oxides.
These dissolved complexes reprecipitated when
the pH increased to 5.5 or 6 because of neutraliza-
tion of the waste by the aquifer carbonates and
oxides.
• The aquifer mineral constituents adsorbed all waste
organic compounds except formaldehyde. Ad-
sorption of all organic acids except the phthalic acid
was increased with a decrease in waste pH.
• Phthalic acid formed complexes with dissolved iron.
The concentration of this complex decreased as pH
increased because the complex coprecipitated with
the iron oxide.
• Biochemical-waste transformation occurred at low
waste concentrations, resulting in the production of
methane. Additional microbial degradation of the
waste resulted in the reduction of sulfates to sul-
f ides and the reduction of ferric ions to ferrous ions.
Effects of Injection. Wells became plugged after a few
months of injection because of waste reactivity. The plug-
ging resulted from precipitation of the initially dissolved
minerals and from the formation of such gases as carbon
dioxide and methane. The combination of plugging in the
formation and the dissolution by the organic acids of the
bond between the cement grout surrounding the well
casing and confining beds resulted in leakage of waste
upward into the shallower zone.
2.7.3 Field Case Study: Pensacola, Florida (American
Cyanamid)(Birlichetal., 1979; \fecchiolietal., 1984)
Waste Characteristics. Industrial waste liquid containing
organonitrile compounds, nitrate, and sodium thiocyanate
was injected at this facility.
Aquifer Geology/Characteristics. The injection-zone
aquifer was a limestone formation.
Injection Activities, No details were given on the injection
activities themselves. One observation well was con-
structed in the zone.
Processes Observed. Microbiological degradation con-
verted organic compounds to carbon dioxide, and nitrate
was reduced to elemental nitrogen. These transformations
were virtually complete within a short distance from the well.
Sodium thiocyanate remained unaltered during movement
through the injection zone and was used to detect the
degree of mixing of waste liquid with native water at an
observation well. An 80% reduction in chemical oxygen
demand was observed.
Effects of Injection. The waste liquid was free from or-
ganonitriles and nitrate by the time it reached the monitor-
ing well. Sodium thiocyanate remained as a contaminant.
2.7.4 Field Case Study: Pensacola, Florida (Monsanto)
(Pascale and Martin, 1978)
Waste Characteristics. Liquid waste containing nitric
acid, inorganic salts, and numerous organic compounds
was injected at this facility.
Aquifer Geology/Characteristics. The injection-zone
aquifer was saline limestone.
Injection Activities. Data on injection rates, volumes, pres-
sures, water levels, and laboratory analyses of waste
samples taken from three monitoring wells were collected
between 1970 and 1977. Wellhead pressure averaged 180
psi in March 1977, and the hydraulic pressure gradient was
0.53 psi/ft of depth at the top of the injection zone.
Processes Observed. Microbial degradation of waste in
the injection zone is inferred from observed increases in
bicarbonate, dissolved organic carbon, and gas content in
the deeper monitoring well.
Effects of Injection. No effects were observed in the
shallower observation well. The deeper monitoring well to
the south showed an increase in concentrations of bicar-
bonates, dissolved organic carbon, and gas.
2.8 Further Research Needs
All areas need to be investigated further. The following
could be the most productive in the near-term:
• Conduct dynamic coreflood studies of selected
phenols to determine their short-term fate (30 to 60
days) under typical reservoir conditions created in
the laboratory. Such parameters as solution pH,
salt concentrations, temperatures, clays, and waste
concentration should be evaluated with respect to
precipitation, adsorption, permeability reduction,
and thermal degradation.
• Conduct additional dynamic coreflood and/or re-
lated studies of selected hazardous wastes to
determine their fate in subsurface environments.
These studies might include: coreflood studies
35
-------
using different cores and other organic waste com-
pounds; studies of the interactions of phenols with
confining-layer materials (using hydrothermal reac-
tors rather than corefloods); and studies of fhe
effects of microorganisms on phenols. |
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39
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CHAPTER THREE
RESEARCH SUMMARY NO. 2
THE CHEMISTRY OF WASTE FLUID DISPOSAL IN DEEP INJECTION WELLS
3.1 Overview
3.1.1 Origin and Content
Source: The Chemistry of Waste Fluid Disposal in Deep
Injection Wells. Second Berkeley Symposium on Topics in
Petroleum Engineering, March 9-10, 1988, pp. 79-82.
Lawrence Berkeley Laboratory LBL-24337. Prepared for
U.S. Department of Energy. (
Authors: J. Apps, L. Tsao, and O. Weres. Lawrence
Berkeley Laboratory, University of California, Berkeley,
California 94720.
Contents: This paper focuses on chemical aspects of
deep-well injection of hazardous wastes. It includes (1)an
overview of types of models for predicting fate and of
deficiencies in available models, (2) a comparison of the
laboratory simulation of the evolution of Gulf Coast brines
with actual brines, and (3) the results of laboratory expiari-
ments studying the interactions between bentonite clay and
a simulated waste of sodium borate and cresol.
3.13 Major Conclusions
• Not much is known about the chemical consequences
of injecting dilute toxic waste streams into deep
sedimentary formations.
• Computer simulations of chemical processes currently
model only the simplest systems, and laboratory and
field studies are required for more realistic predictions.
Semiempirical techniques will continue to be necessary
for the foreseeable future. Any studies for organic
waste disposal must be site-specific, using rock frjom
the proposed injection zone. ;
• A computer simulation of the evolution of Gulf Coast
brines and experimental studies at 250°C, with a simu-
lated waste stream containing creosol and sodium
borate interacting with bentonite clay, illustrates the
potential complexities and uncertainties in attemptingto
predict quantitatively the fate of waste in the injection
formation.
3.2 Processes Affecting the Geochemical
Fate of Deep-Well-Injected Wastes
3.2.1 Overview of Fate-Influencing Processes
Problems arising from deep-well injection may be classified
as mechanical, hydrological, or chemical. This paper con-
centrates on the chemical problems.
The broad scope of chemical processes that must be un-
derstood includes (1) the chemical evolution of the ground
water in migrating from its source to the injection zone, (2)
the interaction of the ground water with the injected waste
stream, and (3) the interaction of the injected waste stream
with the host rock.
Interaction of the waste stream with the ground water and
host rock may have both deleterious and beneficial effects.
Deleterious effects include (1) formation of high gas over
pressures, (2) hazardous-daughter products resulting from
decomposition of constituents in the waste stream, (3) the
partition and concentration of hazardous constituents in a
more highly mobile form (e.g., vapor phase), and (4)
precipitation of reaction products that could seal the injection
zone. Beneficial effects would lead to the attenuation of
toxic constituents through neutralization, precipitation,
decomposition, adsorption, oxidation or reduction, or bac-
terial decomposition.
Many detailed processes with homogeneous and hetero-
geneous reactions may be involved in the interaction of a
waste stream with ground water and host rock. Homo-
geneous reaction processes may include complexation,
oxidation/reduction, hydrolysis, and polymerization. Hetero-
geneous reaction processes may include nucleation, colbid
formation, precipitation/ dissolution, adsorption, ton ex-
change, immiscible phase separation (i.e., formation of non-
aqueous gas or liquid phases), and bacterial decomposition.
40
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3.2.2 Partition Processes
Adsorption is the only partition process specifically covered
by this report. An extensive literature detailing the transport
of organic waste compounds in soils provides ample
evidence that many organic compounds are effectively ad-
sorbed onto the organophilic surface of organic detritus.
Smectite clays also adsorb organic compounds possessing
hydrophilic ligands. Models describing adsorption onto
natural materials are in preliminary development (see Sec-
tion 3.5.1).
3.2.3 Transformation Processes
Catalysis is the only transformation process specifically
covered by this report. Bentonite clay can serve as a
catalyst for several types of reactions involving cresol. A
small percentage of cresol will be demethylated to produce
phenol. With an acid-washed clay, about 75 wt.% of the
cresol reacted to form 1-methoxy-4-methyl benzene, with
less than 1% of the initial cresol remaining. See discussion
of cresols in Section 3.4.2.
3.2.4 Transport Processes
Transport processes are not specifically discussed in this
report.
3.3 Major Environmental Factors Affecting
Deep-Well-Injected Waste
3.3.1 Geochemteal Characteristics of Deep-WelUnjectton
Zones
Gulf Coast injection wells are typically between 4,000 and
7,000 ft. deep with temperatures up to 80°C. Atypical injec-
tion zone is an arenaceous horizon containing up to 70 wt.%
detrital quartz, together with 15 wt.% of detrital plagioclase
and potash feldspars, with the remainder clay minerals with
secondary cateite.* Confining shale horizons typically con-
sist of about 70 wt.% clays with smaller amounts of other
detrital minerals and secondary pyrite. A significant al-
though minor amount of organic detritus is present in both
shales and sandstones.
3.3.2 Specific Environmental Factors
Brines. Brines in Gulf Coast injection wells typically contain
between 30 and 80 g/L (30,000-80,000 mg/L) of a mixture
of sodium and calcium chlorides. The salinity is attributed
to the dissolution of sodium chloride from evaporites. Table
3-1 shows average concentrations of aqueous species in
Gulf Coast brines at 80°C and Table 3-2 lists minerals in
saturation with the simulated brines in Table 3-1.
Table 3-1 Comparison of Predicted and Measured
Aqueous Species in Gulf Coast Brines, 80°C
Species
Na+
K+
Mg2"1"
Ca2+
Fe2+
Al3*
C032-
SC-42-
S2~
SiC>2(aq)
PCO2
pCH4
PH2
pH2S
Eh
pH
Concentration,
Simulation
9.1 x10"1
2.5 x10"2
1.6X10"3
4.0 x10'2
1.6x10"7
3.4 X10"4
6.3 X10"4
1.6 x10'7
1.0 X10"6
6.3 X10"4
1.6x10'2bar
1.0x10° bar
3.2 x 10"^ bar
LOxlO"5
-300 mv
6.8
Mole (kg hfeO)"1
7 Aquifers'*
7.8 x10"1
8.0 X10"3
1.2 x10"2
5.0 x10'2
1.0 xlO"3
<1.0x10"3
7.0
aKreitlerandRichter
Table 3-2 Minerals Predicted to be in Equilibrium with
Gulf Coast Brines at 80°C
Observed
Predicted Minerals
Primary
Secondary
Calcite
Dolomite
Hematite
Kaolinite
LowAlbite
Pyrite
Pyrrhotite
Quartz
Smectite
The paper states 40% quartz, 15% fedlspar, and 40% clay minerals and cafcite. A high percentage of clay minerals
was used for geochemical modeling to reflect the fact that the high surface area of clay results in a reactivity that is
higherthan a simple weight percentage woudl indicate (communication from John Apps, Lawrence Berkeley Laborato
Berkeley, California, August 7,1989.
41
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Low concentrations of magnesium and sulfate tons com-
pared w'rth simulated values suggest that brines in the field
are not in homogeneous or heterogeneous equilibrium.
Clays. Claytypescommonlyfound in Gulf Coast deep-well-
injectfon formations include smectite, illite, mixed-layer
clays, chlorite, and kaolinite. Most of the clays in both
the arenaceous and argillaceous horizons of Gulf Coast
deep-well formations are detrital in origin. Smectite clays
absorb organic compounds possessing hydrophilic ligands.
See Section 3.2.3 for a discussion of the catalytic effects of
bentonfte clay.
3.4 Geochemical Characteristics and Fate of
Hazardous Wastes
3.4.1 Specific Data on Inorganic Wastes
Sodium borate is the only inorganic discussed in this report.
This compound (27.1 g/L) catalyzed the condensation of
cresol either directly or indirectly by its influence on pH in
laboratory-simulated interactions with cresol and bentohite
clay.
3.42 Specific Data on Organic Wastes Cresols.
Laboratory-simulated interactions between a waste stream
of sodium borate and the paracresol isomer with bentonite
clay in various concentrations and combinations showed
that 1.4 wt.% of cresol was demethylated to produce phenol
in 72 hours, and a series of largely unidentifiable condensa-
tion products or dimers of cresol were formed during initial
experiments. The clay catalyzed the demethylation of
cresol. The presence of sodium borate catalyzed the con-
densation of cresol either directly or indirectly through its
influence on pH. With an acid-washed clay, approximately
75 wt.% of the cresol reacted to form 1-methoxy-4-meihyl
benzene, w'rth less than 1% of the initial cresol remaining.
The remainder of the reaction products were methoxyben-
zene and unidentifiable reaction products.
l-Methoxy-4-methyl benzene. This compound was jthe
primary product of catalytic reactions involving cresol in the
presence of acid-washed bentonite clay.
Methoxybenzene. This compound was a minor product of
catalytic reactions involving cresol in the presence of acid-
washed bentonite clay.
3.5 Methods and Models for Predicting the
Geochemical Fate of Deep-Well-Injected Wastes
357 Basic Approaches
Because of the chemical complexity of waste streams,
aquifer brines, and host rocks, many interactions are pos-
sible. Unfortunately the present state of knowledge is quite
insufficient to make useful predictions based on computer
simulations alone. Laboratory and field experiments must
supplement these modeling efforts.
Computer codes used to predict chemical processes may
be divided into five broad categories:
• Models used to reduce and evaluate experimental data.
• Models used to calculate the thermodynamic properties
of phases or species at temperatures and pressures
other than standard-state conditions.
• Models used to determine the distribution of species at
equilibrium, given the principal extensive and intensive
parameters of the system.
• Models used to predict the evolution of a chemical sys-
tem, either as a function of reaction progress or as a
function of time. The spatial distribution of reactants or
products is not predicted by these models.
• Models used to predict the chemical evolution of a sys-
tem in which both chemical reactions and transport
proceed simultaneously.
The thermodynamic properties of participating minerals and
aqueous species are needed for most types of modeling.
Substantial thermodynamic data bases have been com-
piled, and detailed evaluations have been conducted to en-
sure that the data are both internally consistent and correct.
Serious deficiencies remain.
• The thermodynamic properties of many relevant water-
miscible organic species are either incomplete or un-
available.
• Many minerals are solid solutions (e.g., clays, am-
phiboles, and plagioclase feldspars). Either solid-solu-
tion models remain to be worked out or appropriate al-
gorithms have not been incorporated into computer cod
42
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• Models describing the adsorption of water-miscible or-
ganic compounds on natural materials are in the
preliminary stages of development and have not been
correlated with field observations under typical injec-
tion-zone conditions. Few computer codes contain al-
gorithms permitting the distribution of species between
the adsorbed and aqueous state to be calculated.
• Calcium-sodium chloride brines (which typically occur
in deep-well injection zones) require sophisticated
electrolyte models to calculate their thermodynamic
properties. Many parameters for characterizing the
partial molal properties of the dissolved components in
such brines have not been determined. Precise model-
ing is limited to relatively tow salinities, where many
parameters are unnecessary, or to chemically simple
systems near 25°C.
• Current computer codes usually calculate only the ther-
modynamically most stable configuration of a system.
Modifications can simulate nonequilibrium conditions,
but there are limitations to the extent to which codes can
be manipulated to simulate processes that are kineti-
cally controlled.
• Little is known about the kinetics of dissolution,
precipitation, and oxidation-reduction reactions in the
natural environment. Therefore, attempts to simulate
the kinetics of the evolution of the more complicated
injection-zone chemistry must remain in the develop-
mental stage for some time.
The most sophisticated computer codes are those which
combine transport and chemical processes. Few such
codes have been developed and successfully tested (see
Section 3.5.2, Transport).
The experimental study reported in this paper shows clearly
that the reactions observed could not have been anticipated
by a priori computer simulations. Furthermore, it will not be
easy to simulate the complex reaction paths observed
without a more fundamental understanding of the
mechanisms involved. Thus, computer simulations of the
waste stream reacting with the injection-zone environment
will provide only limited insight into the consequences of
waste-stream injection.
At present, computer models incorporating kinetics are very
limited in their applicability and usefulness for predicting in-
jection-zone conditions. The availability of computer codes
for modeling complex kinetic systems will probably precede
the availability of suitable data or techniques for correlating
theoretical concepts with the reactivity of an injection zone.
Therefore, semi-empirical techniques will continue to be
necessary.
3.5.2 Specific Methods and Models
Aqueous- and Solution-Geochemistry Models. Although
most available models calculate only the thermodynamically
most stable configurations of a system, it is possible to
manipulate their operation to simulate metastable condi-
tions. Simple input modifications to include metastable or
unstable compounds can produce fairly realistic simulations
of nonequilibrium systems.
Where kinetic data are not available, the evolution of a
chemical system can be simulated using the reaction-
progress variable developed by de Donderand Van Ryssel-
berghe (1936). Helgeson et al. (1970) developed algo-
rithms using this variable. In its simplest form, a reaction-
progress code simulates changes in a chemical system in
terms of the amount of material reacted rather than as a
function of time. Usually, one kilogram of aqueous phase
waste is allowed to react with an assemblage of minerals.
The aqueous phase is assumed to be always in internal
equilibrium and the reactant minerals dissolve in proportion
to their initial mole fractions. Product phases are in revers-
ible equilibrium with the aqueous phase.
The paper contains results of a simulation showing the
evolution of a typical Gulf Coast brine using reaction-
progress variables in the EQ6 code developed by T. J.
Wolery and his associates at the Lawrence Livermore Na-
tional Laboratory. The reaction of a 1 molal sodium-chloride
brine (representing dissolution of an evaporite) reacting with
a representative detrital mineral assemblage at 80°C was
simulated to see whether it would evolve into a sodium-cal-
cium chloride brine and produce the secondary minerals
observed in the field. Table 3-1 compares the resulting brine
composition, when the reaction progress variable equals
10", with average values for Gulf Coast brines; Table 3-2
lists the minerals in saturation with the brine.
With some notable exceptions, results are generally consis-
tent with what is observed in the field. The predicted value
for potassium is too high and can be explained by the failure
of the simulation to saturate with respect to illite, probably
due to erroneous thermodynamic properties for that clay.
The tow magnesium and sulfate concentrations suggest that
the brines in the field are not in equilibrium. Predicted and
observed minerals are in good agreement.
Transport Models. White et al. (1984) have developed a
groundwater-contarnination model that integrates TRUMP
(Edwards, 1972), a transport model, with PHREEQE
(Parkhurst et al., 1980), a distribution-of-species model.
Another model, taking a more fundamental approach, is
CHEMTRN (Miller and Benson, 1983).
43
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3.6 Laboratory Procedures And Protocols
Experiments involving a simulated waste stream containing
sodium borate and cresol (up to 500 ppm) interacting with
bentonile were conducted in 600 ml_ Parr Minireactors. in-
itial runs were made at 250°C with 14.3 g/L of cresol and a
solution of sodium borate at 27.1 g/L, to obtain measurable
product yields in a reasonable time. Even though the waste
stream being simulated contained mixed isomers of cresol,
the paracresol isomer was chosen because it was expected
to be more reactive and a single isomer would make iden-
tifying reaction products simpler. Benton'rte was used at a
concentration of 143 g/L, later decreased to 100 g/L. After
closure, the vessel was flushed with helium and pressurized
to approximately 16 psig.
At the conclusion of the run and after cooling the vessel,
head-space gases were collected overwater in a separately
funnel. These gases were analyzed on a Consolidated En-
gineering Corp. 21-102 mass spectrometer. The vessel
was then opened and a spike of perdeuterophenol solution
was added to the product solution. After mixing the spikjed
solution, it was decanted into plastic jars and centrifuged to
separate the clay. The supernatant solution was removed
and an aliquot was extracted with methylene chloride in a
separatory funnel. The methylene-chloride extract was
washed once with distilled water and dewatered by filtering
through anhydrous sodium sulfate.
The reaction product extract was analyzed by two methods:
(1) a gas chromatograph with flame-ionizatfon detector to
screen product solutions for further analysis, and (2) a Fjn-
nigan 4000 gas chromatograph/mass spectrometerto iden-
tify reaction products and determine the absolute quantity
of phenol produced. Results of the experiments are dis-
cussed in Section 3.4.
3.7 Field Case Studies
No field case studies are included in this report.
3.8 Further Research Needs
3.8.1 General
• Continue to refine and add to the thermodynamic
databases required for modeling. '
• Develop solid-solution models and the capability to
model precisely the thermodynamic properties of strong
mixed electrolytes for a diverse range of injection-zone
conditions.
• Develop data on thermodynamic properties of water-
miscible organic compounds.
• Develop empirical models describing irreversible ad-
sorption of water-miscible organic compounds on
mineral surfaces in an injection zone.
3.8.2 Specific
m Study further the issue of noriequilibrium in actual
brines. The comparison of simulated and actual Gulf
Coast brines suggests that the actual brines are not in
homogeneous or heterogeneous! equilibrium.
• Conduct field studies to compare results of laboratory
experiments and computer modeling reported in the
paper. These field studies could involve injecting a
simulated waste stream containing variable amounts of
sodium borate and cresol in an arenaceous formation.
The injected stream could be left in place for an ex-
tended period of time, then recovered, and changes in
its composition measured. The formation fluids could
be continually removed and measured for changes in
borate and cresol, allowing the adsorptive-desorptive
capacity of the rock, potential decomposition products,
and various hydrologte parameters to be determined.
The results could be correlated with laboratory studies
and conclusions drawn regarding the scaling factors
and more fundamental differences in mechanisms be-
tween laboratory and field conditions.
References
de Donder, T. H., and P. Van Rysselberghe. 1936. Ther-
modynamic Theory of Affinity. Stanford University Press,
142 pp.
Edwards, A. L 1972. TRUMP: A Computer Program for
Transient and Steady State Temperature Distributions in
Multi-Dimensional Systems. Lawrence Livermore National
Laboratory. Rep. 14754, Rev. 3, Livermore, California.
Helgeson, H. C., T. H. Brown, A. Nigrini, and T. A. Jones.
1970. Calculation of Mass Transfer in Geochernical
Processes Involving Aqueous Solutions. Geochimica et
Cosmochimica Acta 34:569-592.
Kreitler, C. W., and B. C. Richter. 1986. Hydrochernical
Characterization of Saline Aquifers of the Texas Gulf Coast
Used for Disposal of Industrial Waste. Bureau of Economic
Geology, University of Texas at Austin, 164 pp.
44
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Miller, C.W., and L.V. Benson. 1983. Simulation of Solute
Transport in a Chemically Reactive Heterogeneous System:
Model Development and Application. Water Resources /?e-
searc/7,19(2)381-391.
Parkhurst, D. L, D. C. Thorstensen, and L. N. Plummer.
1980. PHREEQE—A Computer Program for Geochemical
Calculation. U.S. Geological Survey Water Resource In-
vestigation 80-96.
U.S. Environmental Protection Agency. 1985. Report to
Congress on the Injection of Hazardous Wastes. EPA
Report 570/9-85-003. Office of Drinking Water, U.S. En-
vironmental Protection Agency, Washington, D.C.
White, A. R, J. M. Delaney.T. N. Narashimhan, and A. Smith.
1984. Groundwater Contamination from an Inactive
Uranium Mill Tailings Pile, I., Application of a Chemical
Mixing Model. Water Resources Research 20:1743-1752.
45
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CHAPTER FOUR
RESEARCH SUMMARY NO. 3
LABORATORY PROTOCOL FOR DETERMINING FATE OF WASTE DISPOSED IN DEEP
WELLS
4.1 Overview
4.1.1 Origin and Content
Source: Laboratory Protocol for Determining Fate of Waste
Disposed In Deep Wells. EPA/600/8-88/008, February
1988. Prepared for the U.S. Department of Energy and U.S.
Environmental Protection Agency. 63 pages. NTIS PB88-
166061.
Authors: A. Gene Collins and M. E. Crocker, National
Instftute for Petroleum and Energy Research, P.O. Box
2128, Bartlesville, Oklahoma 74005.
Contents: Describes laboratory procedures for: (1) core
analysis, (2) brine analysis, (3) a dynamic fiuid-f tow system
that simulates the interaction of hazardous organic wastes
with Injection-zone rock, and (4) a static waste/rock interac-
tion test that simulates longer-term degradation processes.*
Protocol testing resulted in some data on the mobility,
adsorption, and degradation of phenol and 1,2- dichloro-
ethane in simulated subsurface conditions for the Frio
sandstones; also presented were some data from easier
adsorption experiments using the Cottage Grove
sandstone.
4.1£ Major Conclusions
m Laboratory simulation of the interaction between in-
jected hazardous wastes and reservoir rocks can help
evaluate the mobility, adsorption, and degradation of
organic hazardous wastes in the deep-well environ-
ment.
• Dynamic fluid-ftow interactions in the laboratory cai^i be
used to evaluate the adsorption and desorption
behavior of organic hazardous wastes in reservoir rock
at simulated subsurface temperature and pressure con-
ditions. Static waste-reservoir-rock interaction tests
can be used to evaluate adsorption behavior and
degradation products over longer periods of time
(months) under simulated temperatures.
Preliminary tests of interactions between simulated
phenol-brine and 1,2-dichloroethane-brine solutions
and two sandstones (Frio formation in Texas and Cot-
tage Grove sandstone in Oklaho ma) indicate the follow-
ing:
Adsorption rates of the two compounds are proportional
to their concentration in solution and inversely propor-
tional to the temperature. The adsorption process is
exothermic, and low isosteric heats of adsorption indi-
cate that Van der Waals, or physical adsorption, is the
dominant type of process.
Adsorption rates for 1,2-dichloroethane were higher in
the Cottage Grove sandstone than in the Frio
sandstone. The reverse was true for phenol; adsorption
rates for phenol on the Cottage Grove sandstones were
tower than for the Frfo sandstone by a factor of four or
five.
Flushing a Frio core that had attained adsorption equi-
librium for phenol with a phenol-free brine resulted in no
desorption.
No phenol degradation products were observed either
in effluents generated in the fluid-flow experiments or in
Specific laboratory procedures are briefly described in this summary. The original report should be obtained for
detailed descriptions. See in particular Appendix B of that report (Experimental Procedures).
46
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static tests where the brine-phenol solution was mixed
with crushed Frto rock at 140° F and held for 83 days.
4.2 Processes Affecting the Geochemical
Fate of Deep-Well-injected Wastes
4.2.1 Overview of Fate-Influencing Processes
Adsorption is a significant process that can affect the migra-
tion of injected-waste constituents. Little laboratory data
exist on the equilibrium adsorption properties of subsurface
formation rocks with respect to organic compounds (Stryck-
er and Collins, 1986).
4.2.2 Partition Processes
Adsorption-desorption. Many organic constituents are
hydrophobic, and their movement in groundwater systems
is affected by adsorption phenomena, where the solute is
transferred from the liquid or solvent phase to the solid
phase. Adsorption is a phase-partitioning process and it
may be fully or partly reversible (desorption). Many chemi-
cal and physical properties of both the solute and the solid
phases affect adsorption-desorption, and the physical
chemistry is complex.
Adsorption can be categorized into three classes: ex-
change, physical, and chemical. Exchange occurs from the
electrostatic attraction of tons or charged functional groups
from the solute phase to the portions of the solid phase that
possess charge locations. Physical adsorption occurring
between the phases is motivated by Van der Waals forces
or electrostatic interaction between atoms and molecules
(Miller, 1984). Chemical adsorption (chemisorption) is
caused by a reaction between the solute phase and the
solid phase. These classes of adsorption rarely act singly.
Phase partitioning of a solute- solvent- solid system in which
two or all three of the adsorption classes operate probably
occurs.
Smith (1968) categorizes the adsorption overall into several
physical processes that represent the steps occurring when
a molecule is transferred from a solution to an active adsorp-
tion site on a solid surface. These steps are: (1) transfer of
a molecule from the bulk phase near to the solid surface by
molecular or convective diffusion; (2) diffusion of the
molecule into the pores of the solid to the site of adsorption;
(3) adsorption of the molecule onto the site. Where a
chemical reaction takes place, Steps 2 and 3 occur in
reverse for the products.
The rate at which adsorption at the site takes place (Step 3)
for aqueous solutions is very rapid, especially when it is
accompanied by a "loose" bond, which is indicated by a low
heat of adsorption. Heats of adsorption that are less than 8
kcal/gmole indicate weak Van der Waals (physical) adsorp-
tion. The high rate of adsorption in Step 3 indicates that this
step does not influence the overall kinetics of the adsorption
process. In physical adsorption, bulk and internal pore
diffusion (Steps 1 and 2) are the steps controlling the
kinetics. Since diffusion is the controlling mechanism, it is
not unusual for organic compounds, where physical adsorp-
tion is the dominant process, to exhibit adsorption rate
constants that are numerically close.
In general, the rate of adsorption of organic waste com-
pounds is proportional to their concentration in solution and
inversely proportional to the temperature. The adsorption
process is exothermic for phenol and 1,2-dichloroethane,
and tow numeric values for the isosteric heats of adsorption
(> 8 kcal/gmole) indicate that physical adsorption is the
dominant process.
4.2.3 Transformation Processes
None discussed.
4.2.4 Transport Processes
The only transport process covered specifically in this report
is solute migration. Possible mechanisms whereby injected
hazardous wastes might contaminate aquifers include: (1)
a surface spill followed by migration of the waste into ground
water; (2) unplugged or incompetently plugged wells that
penetrate the geologic zone into which the waste is injected,
providing a route whereby the waste can enter an overlying
potable aquifer; (3) vertical fracturing of the injection and
confining strata caused by excessive injection pressures,
whereby a communication channel allows the injected
wastes to migrate to a freshwater aquifer; and (4)
mechanical failure of the injection system such as corrosion
of surface pumps or pipes, or subsurface tubing or casing,
which allows the waste to escape and migrate to an aquifer
(Collins, 1975).
The U.S. EPA(1985) described four major ways subsurface
injection can cause fluids to migrate into underground sour-
ces of drinking water (USDWs): (1) faulty well construction;
(2) improperly plugged or completed wells in the zone of
endangering influence; (3) faulty or fractured confining
strata; and (4) lateral displacement.
An injected fluid moving through a porous system spreads
into the reservoir fluid being displaced by the simultaneous
actions of convective and molecular diffusion. Convective
dispersion is influenced by the density and viscosity dif-
ference between the two miscible fluids and by local varia-
tions of the fluid velocity, pore-size distribution, pore constric-
tions, and the tortuosity of the flow path.
47
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Molecular diffusion can occur by bulk diffusion and i by
Knudsen diffusion. Wheeler (1951) suggests that bulk dif-
fusion atone will occur at a pore radius of 10"4 cm or greater.
The pore-size distributions measured for five sandstones
(Donaldson et al., 1975) show that most of the pores have
radl greater than this value. Consequently, Knudsen dif-
fusion can be considered negligible.
Other modes of mass transfer that affect dispersion in
porous media are eddy diffusion and transverse molecular
diffusion. If laminar flow exists (generally the case in
sandstone deep-well-injection zones), eddy diffusion is not
a contributing factorto dispersion. Transverse diffusion can
be neglected if the ratio of the radius to the length of
sandstone samples is small and the viscosities of the dis-
placed and injected fluids are equal. Hassinger and v^on
Rosenberg (1968) and Grane and Gardner (1961) provide
additional information on transverse dispersion. :
4.3 Major Environmental Factors Affecting
Deep-Well-Injection Geochemical Processes
4.3.1 Geochemical Characteristics of Deep-Well-
Infectton Zones I
Current knowledge of deep subsurface geologic reservoirs
was established primarily by petroleum-related sciences.
Subsurface reservoirs into which hazardous wastes are
injected consist of sedimentary deposits, the same types
from which petroleum is generated and produced. :
Water Chemistry. Brines found in the Frio formation vary.
The salinity (total dissolved solids) can be as tow as 16,000
mg/L in some of the geopressured wells and more than
100,000 mg/L in normally pressured wells. The dominant
tons are sodium and chloride; alkalis, alkaline earths, and
halide tons plus trace elements are present also.
Sandstones. Most deep geologic formations used for dis-
posal of hazardous wastes consist of unconsolidated sands
and sandstones, as exemplified by the Frfo formation. The
minerals in these rock formations serve as sites for the
adsorption of numerous organic compounds and thus retard
or attenuate the migration of the waste from the injection
zone.
The Frio formation is the dominant geologic repository used
for hazardous-waste disposal in the Gulf Coast area be-
cause of its lithologto characteristics. It is of Oligocehe-
Mfocene age and has produced about 6 billion barrels of oil
and 60 trillion cubic feet of gas. Typical subsurface condi-
tions in the Frio formation are 3,000 psi (20.7 MPa) with
temperatures up to 140'F (60°C). The core used in the
protocol experiments came from a depth interval of 7,154 to
7,155 ft. Porosity was 36% and permeability was 1.7 dar-
cies. Composition of the core was as follows: quartz
(52.0%), feldspar (17.0%), calcite (28.0%), pyrite (trace),
kaolinite (3.0%), and illite (trace).
The Cottage Grove sandstone is another formation that is
representative of formations into which waste solutions are
injected for disposal. The core used in experiments related
to the project had 26% porosity and a permeability of 284
millidarcies. Composition of the core was as follows: SiOa
(75.4%), AlaOa (5.8%), KaO (5.7%), kaolinite (6.0%), chlorite
(1.0%), and illite/mica (6.0%).
Confining Beds. Shale- and/or clay- or silt-dominant for-
mations, which overlie many formations such as the Frio,
serve as hydrocarbon traps and also should retard the
vertical migration of injected hazardous wastes out of the
zone.
4.3.2 Specific Environmental Factors
Temperature. In general, an increase in temperature
decreases the rate of adsorption.
Pressure. Pressure affects degree of permeability but
should have no effect on adsorption or degradation.
pH. The pH of thef lowthrough solutions before reaction with
the Frfo core was 5.7. After reaction, the solutions were at
pH 8.1, indicating that they were probably reacting with
minerals such as calcite.
4.4 Chemical Characteristics and Fate of
Hazardous Wastes
4.4.1 Chemical Properties of Inorganic Hazardous
Wastes
None discussed.
4.4.2 Chemical Properties of Organic Hazardous
Wastes
1,2-Dichloroethane. Static adsorption tests of brine solu-
tions containing 1,2-dichloroethane, a halogenated aliphatic
hydrocarbon, using Cottage Grove sandstone showed in-
creased rates of adsorption with increased concentrations
(about 50 mtorograms [ng]/gram [g] of rock at 1,000 ppm to
about 340 ug/g at 5,000 ppm at 100T). Adsorption rates
decreased as temperature increased. For example, the rate
of 340 ug/g at 100T declined to about 300 ug/g at 140T.
Similar relationships were found for adsorption onto the Frfo
sandstone except that the total rate was considerably tower
(maximum of about 140 |ig/g of Frio sandstone at 1 DOT vs.
48
-------
about 340 ug/g of Cottage Grove sandstone). The adsorp-
tion process was exothermic, and tow isosteric heats of
adsorption indicate that Van der Waals (physical) adsorption
is the dominant type acting on this waste.
Phenol. Phenol, a monocyclic aromatic, is the predominant
organic hazardous waste injected into the Frfo formation.
Flowthrough experiments with phenol (500 mg/Lto 10,000
mg/L) in a simulated sodium-chloride brine (10,170 mg/L)
injected through Frfo sandstone indicated the following:
• Adsorption increased as concentration increased (from
9 ug/g adsorbed of rock at 500 mg/Lphenol to 312 jjg/g
at 10,000 mg/L and 100T).
• Adsorption decreased as temperature increased. For
example, the rate of 312 ug adsorbed/g of rock at
10,000 mg/L phenol and 100T decreased to 276 jjg/g
at 140°F.
• The adsorption process was exothermic, and tow
isosteric heats of adsorption indicate that Van der Waals
(physical) adsorption is the dominant type.
• Flushing a core that had attained adsorption equilibrium
for phenol with brine containing no phenol resulted in
no desorption.
• No phenol degradation products were observed in
effluents generated in the fluid flow experiments.
Similarly, no desorption occurred when the brine-
phenol solution was mixed with crushed Frfo rock at
140°F and held for 83 days.
Static adsorption tests of phenol-brine solution with Cottage
Grove sandstone showed similar relationships to those
described above for the Frfo sandstone except the total rate
was considerably lower (maximum of about 80 ug/g of rock
at 100T, and about 50 ug/g at 140T).
4.5 Methods and Models for Predicting the
Geochemical Fate of Deep-Well-Injected Wastes
4.5.1 Basic Approaches
The fate of wastes injected into subsurface rock can be
demonstrated only if the behavior of the waste after injection
is known. The interactions of the waste with other reservoir
fluids and reservoir-rock constituents such as silicates, car-
bonates, sulfates, and clays must be determined. Hazard-
ous wastes are complex mixtures, and when these mixtures
are combined with other complex mixtures, the numbers of
possible interactions increase factorially.
4.5.2 Specific Methods and Models
Adsorption. Adsorption rate constants can be obtained
from flowthrough experiments by sampling the effluent at
specified time intervals for each concentration level
analyzed. After the samples are analyzed by GLC for the
specific concentration, the concentrations are then plotted
as a function of time.
The Adsorption-FIuid-Flow Method. Waste fluids are
injected through a core at temperature and pressure condi-
tions that simulate the deep-well environment. Effluent from
the core is sampled periodically to determine when the
concentration of the organic hazardous waste is equal to the
concentration of the injected fluid. When this equilibrium
point is reached, injection stops. Total adsorption is then
calculated in ng of waste/g of rock. The equilibrium amount
adsorbed is plotted versus the concentration at equilibrium
with the rock surface, to obtain the constants for the
Freundlich isotherm equation:
= FCnflnal
where
Ai = Amount of solute adsorbed by the rock, mg/kg of
rock
F = Constant of the Freundlich isotherm (from the
intercept with the vertical axis)
Cfinai = Final solution concentration in equilibrium with
the solid, mg/L
n = Freundlich isotherm constant (from the slope of
the line)
The constants F and n indicate the probable nature of
adsorption. The relationships of F to the free-energy chan-
ges that occur in adsorption are discussed by Haque and
Coshow (1971) and Crisp (1956). When n=1, the intercept
F can be used as an indirect measurement of the surface
free-energy change. A decrease in the value of F as
temperature increases is characteristic of the exothermic
nature of adsorption.
Static adsorption curves are developed by plotting the
amount of hazardous organic waste adsorbed as a function
of time. The adsorption rate constant is calculated using the
equation:
k = 1/Hn(Cfinal/Cfinal-Cl)
49
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where
k * Adsorption rate constant, L/hr
t « time, hours
Cfmai » Final solution concentration in equilibrium with
the solid, mg/L !
Ci a Solution concentration at any time, mg/L [
If diffusion is suspected as the controlling mechanism! for
adsorption (i.e., values of k do not change greatly; as
temperature increases), the coefficients of molecular dif-
fusion for the compound(s) for each temperature can be
calculated for verification using the equation developed by
WKe and Chang (1955).
When adsorption experiments are conducted at two
temperatures, the isoteric heats of adsorption can be calcu-
lated. These values provide information on what type of
surface bonding is occurring. Values that are less than 8
kcal/gmole indicate that the most probable mechanism is
Van der Waals (physical) bonding.
Desorptfon experiments can be performed by injecting or
pumping brine solutions containing no organic hazardous
waste compounds through the core and analyzing the
effluents produced for the compounds that were previously
adsorbed.
The adsorption-static method evaluates the degree of
potential adsorption. In this method, wastes are mixed with
crushed samples of reservoir rock and the fluid analyzed at
time intervals. This method can be used as a supplement
for validating results of flowthrough experiments and can be
used to evaluate longer-term interactions.
Reservoir Characterization. Core and reservoir brine
samples can be characterized using a variety of methods.
The inductively coupled plasma (ICP) spectrometer can be
used for elemental analysis of core and brine samples (U.S.
EPA, 1979). With ICP, atomic emissions are measured
using an optical spectrometerforsimultaneous orsequerjtial
multi-element determination of trace elements in solution
(U.S. EPA, 1979).
Scanning-electron microscopy (SEM) provides visual im-
ages and semiquantitative elemental analysis of core
samples. Postek et al. (1980) describe procedures for SEM
analysis. The method is useful in determining sample mor-
phology, surface mineral composition and type, and location
of clays. When the core is compared before and afteri an
experiment, the potential rock-fluid interactions can be more
readily determined and reasons for permeability tosses may
be indicated.
X-ray-diffraction analysis (XRD) allows further mineral iden-
tification and semiquantitative interpretation of mineral
abundance (Carroll, 1970). This procedure is particularly
valuable for identifying clays in the reservoir rock.
Waste-Reservoir Compatibility. Organic hazardous was-
tes are usually mixed with an aqueous brine solution before
injection into a deep subsurface reservoir. The amount of
waste that can be mixed depends on its solubility in the
aqueous brine. The information needed to evaluate pos-
sible concentration ranges for the injected waste include (1)
density (g/mL), (2) boiling point, and (3) solubility in water
(g/100 ml_). These data can be used to determine ranges
of concentrations for use in the adsorption-f luid-f low method
described above.
Transport. Dispersion experiments using a tracer must be
performed on the assembled injection core system in the
adsorption-fluid-flow method to ensure that the core is an
integral unit and does not contain channels, bypasses, or
other inconsistencies. Two simplifying conditions are set for
determining the coefficient of linear dispersion (D) in the
method: (1) variations of density and viscosity between the
two miscible phases are eliminated by using aqueous-brine
solutions so that transverse diffusion does not have to be
considered; (2) use of laminar flow eliminates diffusion as
a factor. Under these conditions, D is a complex function
of molecular diffusion, velocity, and tortuosity. The report
provides documentationforacomputercodeusingforrnulas
from Bear (1972) and Satter et al. (1977) that tests multiple
values of D until the computed curve matches the ex-
perimentally determined effluent-response curve.
4.6 Laboratory Procedures And Protocols
4.6.1 Waste/Reservoir Characterization
The sample reservoir rock and brines were analyzed using
an ICP spectrometer, a scanning electron microscope, and
X-ray diffraction (see Section 4.5.2). The specific protocols
were summarized in Appendix B of the report and can be
found in U.S. EPA (1979), Postek et al. (1980), and Caroll
(1970), respectively.
The porosity of the core was determined in a 7-step procedure:
1. Determine dimensions of the core.
2. Mount the core in the flowthrough-test-cell apparatus
(see Section 4.6.3).
50
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3. Determine dry weight of core and fittings.
4. Saturate core with brine.
5. Determine saturated core weight.
6. Determine weight and volume of brine reservoir.
7. Calculate porosity and pore volume.
Dispersion experiments using a tracer were carried out as
described in Section 4.5.2 (Transport).
Wastes (phenol and 1,2-dichloroethane) were reagent-
grade chemicals in a stock solution containing 10.17 g/L
sodium chloride in Milli-Q water. Waste/brine solutions of
500, 5,000, and 10,000 ppm solutions of phenol and 500,
2,500, and 5,000 ppm solutions of 1,2-dichloroethane were
created for the experiments. The lowest concentration for
each chemical was selected based on a series of tests run
with a gas-liquid chromatograph using a hydrogen-flame-
tonization detector, which identified each chemical's tower
detection limit. The highest concentrations were selected
based on solubility limits or on the fact that single organic
components of injected wastes do not usually exceed
10,000 ppm. Data on density, boiling point, solubility in
water, and concentration ranges for these wastes were
obtained from the literature.
4.6.2 Static Interaction Tests
This static test method was used to determine the adsorp-
tion and desorption rates and degradation potentials of the
organic wastes. It was also used to evaluate the core di-
mensions and adsorption potentials forthe fluid-flowthrough
tests.
Two series of tests were run, each of which consisted of one
blank sample plus the experimental samples. A commer-
cially available Paarbomb unit was used in the analysis.
In the first series, consolidated Frio core rock was crushed
and sized for passing through a No. 20 (u. 841) mesh sieve.
One gram of this crushed core material was then mixed with
each waste sample, which was prepared as discussed
above, sealed into the Paarbomb unit and placed in an oven
at the test temperature. The individual samples were
evaluated for adsorption at 2-, 4-, 6-, or 24-hour intervals
using a gas chromatograph. The results were then com-
pared with the chromatograph of the hazardous-waste
standard. Finally, the blanks were analyzed to validate the
test's integrity.
In the second test series, all procedures were the same
except that the samples were placed in glass vials in a 1:1
rock-to-liquid ratio, shaken, and the amount of adsorption
determined by comparing the waste concentrations at each
sample time with the initial test concentration.
Static degradation tests were also run using crushed Frfo
rock mixed with a brine/phenol solution sealed in a unit
heated to 140T for 83 days. The fluid was then analyzed
for degradation products.
4.6.3 Dynamic Flowthrough Tests
The dynamic ftowthrough system was designed to estimate
adsorption, desorption, and degradation potentials of or-
ganic wastes under conditions simulating the temperatures,
pressures, overburden pressures, and linear-flow rates
found in subsurface injection systems.
General Apparatus. The f lowthrough system resembled a
high-pressure liquid chromatograph except that the chro-
matographic column is replaced by a core sample
(Donaldson et al., 1980). Figure 4-1 presents a diagram of
the system. The heart of this system is the test cell, which
is machined from 321 stainless-steel tubing stock (2.5-in. ID
and 3-in. OD). It was tested to 10,000 psi (68.9 MPa), twice
the normal operating pressure. The flow lines were made
from stainless-steel tubing and the connections were 316
stainless-steel Swagelokfittings, allowing reactive materials
(such as brines and corrosives) to be analyzed.
A5,000-psi hydraulic pump was used to simulate overbur-
den pressure, and a heating mantle around the cell simu-
lated reservoir temperatures. Dead volume in the system
was minimized using capillary tubing where possible. Total
system volume was 3.86 ml.
The sample-collection system included an automatic frac-
tion collector, as well as a UV spectrophotometer/ refrac-
tometer for onstream analysis. This arrangement allowed
the researchers to choose the most suitable method of
detection. Brine fractions, for example, cannot be analyzed
using UV but must be collected for analysis by another
method.
Pressure Simulations. The core was placed in the high-
pressure cell and the cell then filled with fluid, which can be
pressurized using the overburden pump. The pressure
applied to the outside of the core, which was separated from
the overburden fluid by a rubber Hassler sleeve, simulated
the overburden pressure on an aquifer system and pre-
vented the injected fluid from leaking when the injection
pressure was increased.
The experimental fluids were injected using a high-pressure,
constant-rate pump, and the internal pressure of the system
was adjusted using a pressure control valve.
51
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Figure 4-1 Schematic of fluid flow apparatus.
FLUID
CONTAINERS
COOLING
PRESSURE
RELIEF
VALVE
PUMP
CLOSED SYSTEM
INJECTION VALVE
7)
FILTER
U.V.
INTEGRATORS
r
) REFRACTOMETER
rmmnii
COLLECTOR
RECORDER
Temperature Simulations. The heating mantle allowed
the core to be heated to reservoir temperatures.
Linear-How-Rate Simulations. The high-pressure, con-
stant-rate pump (a Waters HPLC pump with a 6,000 psi
pressure and reservoir-flow rate) provided the needed 1
ft/day linear flow rate found in subsurface waste-injecption
systems. To simulate these conditions, experimenters must
ensure that the pump can operate at this rate with negligible
pulsation at pressures approaching 3,675 psi. The pump
must also have an inlet manifold capable of switching
between fluids with no mixing or disruption (i.e., it must have
zero dead volume).
Fluld-Ftow-System Operation. Dynamic flow experiments
were performed using 500,5,000, and 10,000 ppm phenol
solutions. The overburden pressure was simulated at 3,400-
3,500 psi; the internal (injection) pressure was 2,900-3,000
psi; the pump pressure was 2,900 psi; the injection rate was
0.3 mL/min., and the temperature was set at either 10OTF or
140T.
The core was mounted in the cell, the constant-rate pjump
primed, and the overburden pressure applied around the
core. The flowing pressure of the system, controlled by the
spring setting of the pressure-control valve, was set to the
desired pressure, and the initial solution was pumped
through the core. Simultaneously, the cell temperature was
gradually increased to the desired level. When the predeter-
mined temperature and pressure settings reached equi-
librium, the sample was injected into the core using the
injection valve and the sample collector and recorder were
activated.
Desorption Experiments. Brine containing no phenol was
flushed through the core and effluents were analyzed forthe
presence of phenol and degradation products.
4.6.4 Quality Assurance/Quality Control Procedures
Waste Selection, Handling, and Analysis. Wastes were
purchased as pure, reagent-grade chemicals and solutions
stored at 40T to preserve sample integrity.
Gas chromatography was performed in accordance with
procedures established by the EPA Environmental Monitor-
ing and Support Laboratory.
To evaluate the waste samples, ASTM Method D2580-83,
Phenols in Water by Gas Liquid Chromatography, and
52
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ASTM Method 2908-74, Measuring Volatile Organic Matter
in Water by Aqueous-Injection Chromatography, were used
as guidelines. These methods incorporated other methods
and guidelines:
• ASTM D1129, Definition of Terms Relating to Water
(ASTM, 1987a)
• ASTM D1193, Specification for Reagent WaterfASTM
1987a)
• ASTM E200, Methods for Preparation, Standardization,
and Storage of Standard Solutions for Chemical
Analysis (ASTM, 1987a)
• ASTM E260, Recommended Practices for General
Gas Chromatograph Procedures (ASTM, 1987b)
• ASTM E355, Recommended Practices for Gas
Chromatography Terms and Relationships (ASTM,
1987b)
• Section III of the 1987 Annual Book of ASTM Stand-
ards-Saline and Brackish Waters, Seawaters, and
Brines (ASTM, 1987a)
• Properties of Reservoir Rocks: Core Analysis
(Monicard, 1980)
• API Recommended Practices for Analysis of Oil-Field
Waters (API, 1968)
Instalment Maintenance. To ensure proper operation
of the instruments, a detailed preventive-maintenance
schedule was followed.
Corrective Action. Laboratory personnel were responsible
for noting the need for corrective actions when tempera-
tures, pressures, resolution, or fluid toss varied to an unex-
pected degree. The managers of the project were respon-
sible for noting major problems and defining corrective
procedures.
Documentation. All aspects of each individual experiment
and analytical test were monitored and logged. Each sample
tested was assigned a unique number to establish a chain
of custody for future reference.
Data Evaluation. Where practical, standard analytical pro-
cedures were used to expedite comparison of results from
other experiments.
4.7 Case Studies
No case studies were reported.
4.8 Further Research Needs
No research needs were noted.
References
American Petroleum Institute (API). 1968. API Recom-
mended Practices for Analysis of Oil-Field Waters. API R.P.
45, Second Edition, API, Washington, D.C.
American Society for Testing and Materials (ASTM). 1987a.
1987 Annual Book of ASTM Standards. Water and En-
vironmental Technology, Vol. 11.01 and 11.02 (Water)
ASTM, Philadelphia, Penn.
American Society for Testing and Materials (ASTM). I987b.
1987 Annual Book of ASTM Standards, Vol. 14.01, ASTM,
Philadelphia, Penn.
Bear, J. 1972. Dynamics of Fluids in Porous Media, Chap-
ter 10, American Elsevier Publishing Co., New York.
Carroll, D. 1970. Clay Minerals: A Guide to Their X-Ray
Identification. USGS Special Paper No. 126.
Collins, A. G. 1975. Geochemistry of Oilfield Waters. El-
sevier Scientific Publishing Co., New York, p. 434.
Crisp, D. J. 1956. The Adsorption of Alcohols and Phenols
from Nonpolar Solvents onto Alumina. J. Colloid Sci
11:356-376.
Donaldson, E. C., M. E. Crocker, and F. S. Manning. 1975.
Adsorption of Organic Compounds on Cottage Grove
Sandstone. ERDA/BERC/RI-75/4.
Donaldson, E. C., R. F. Kendall, E. A. Pavelka, and M. E.
Crocker. 1980. Equipment and Procedures for Fluid Flow
and Wettability Tests of Geologic Materials. U.S. Depart-
ment of Energy Report DOBBETC/IC-79/5.
Grane, F. E., and G. H. F. Gardner. 1961. Measurements
of Transverse Dispersion in Granular Media. J. Chem. and
Eng. Dafa6(2):283-287.
53
-------
Haque, R., and W. R. Coshow. 1971. Adsorption of Ispcil
and Bromacil from Aqueous Solution onto Some Mineral
Surfaces. EnvironmentalSci.& Tech. 5(2):139-141.
Hasslnger, R. C., and D. U. von Rosenberg. 1968. A
Mathematical and Experimental Examination of Transverse
Dispersion Coefficients. Soc. Pet. Eng. J. 8(2):195-204.
Miller, C.T. 1984. Modeling of Sorption and Desorptbn
Phenomena for Hydrophobia Organic Contaminants in
Saturated Soil Environments. PhD dissertation, Univ. of
Michigan, p. 402. '
Montcard, R. P. 1980. Properties of Reservoir Rocks: Gore
Analysis. Gulf Publishing Co., Houston, Texas.
Postek, M.T., K. Howard, A. Johnson, and K. L. McMichael.
1980. Scanning Electron Microscopy. Ladd Research In-
dustries, Inc.
Satter, A., Y M. Shum, W. T. Adams, and L. A. Davis. 1977.
Chemical Transport in Porous Media. Presented at the
52nd Annual Technical Conference of Society of Petroleum
Engineers, Denver, Colorado, Oct. 9-12, SPE Paper 6847.
Smith, J. M. 1968. Kinetics of Adsorption. In Adsorption
from Aqueous Solution. ACS Advances in Chemistry Series
79, pp.8-22.
Strycker, A., and A. G. Collins. 1986. Injection ofHazaitfous
Waste in Deep Wells: State-of-the-Art Report. Dept. of
Energy Report No. NIPER-230. (Note: Chapter One of the
Research Summaries document is the summary of an
updated version of this report.)
U.S. Environmental Protection Agency. 1979. Methods for
Chemical Analysisof Water and Wastes. U.S. EPAEnviron-
mental Monitoring and Support Laboratory, Cincinnati, Ohio,
EPA Report 600/4-79-020.
U.S. Environmental Protection Agency. 1985. Report to
Congress on Injection of Hazardous Waste. ERA Report
570/9-85-003.
Wheeler, A. 1951. Advances in Catalysis and Related
Subjects. Academic Press, New York Vol. Ill, pp. 249-327.
Wilke, C. R., and P. Chang. 1955. Correlation of Diffusion
Coefficients in Dilute Solutions. AlChEJ. 1(2):264-270.
54
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CHAPTER FIVE
RESEARCH SUMMARY NO. 4
GEOCHEMICAL INTERACTIONS OF HAZARDOUS WASTES
WITH GEOLOGICAL FORMATIONS IN DEEP-WELL SYSTEMS*
5.1 Overview
5.1.1 Origin and Content
Source: Geochemical Interactions of Hazardous Wastes
with Geological Formations in Deep-mil Systems.
Prepared forthe U.S. Environmental Protection Agency and
the Illinois Hazardous Waste Research and Information
Center. HWRIC Project Number 86015, November 1988.
Available from HWRIC, 1808 Woodfield Drive, Savoy, Illinois
61874. 61 p.
Authors: W. R. Roy, S. C. Mravik, I. G. Krapac, D. R.
Dickerson, and R. A. Griffin, Illinois State Geological Survey,
615 E. Peabody Drive, Champaign, Illinois 61820.
Contents: This report includes (1) a description of
laboratory procedures for batch-type waste-rock-brine in-
teraction tests at simulated subsurface temperature and
pressure conditions;** (2) data on geochemical interac-
tions at different temperatures and pressures between
two types of hazardous waste (acidic and alkaline) with
material from two injection-zone formations and one
confining formation that occur in the upper Midwest
(Mt. Simon sandstone, Potosi dolomite, and Proviso
siltstone); and (3) a comparison of the empirical data
with predictions using two solution-geochemistry
models (WATEQ2 and SOLMNEQF).
5.1.2 Major Conclusions
• Batch-type-interaction studies of hazardous
waste and reservoir brines and rocks performed to
assess geochemical interactions in deep-well systems
should be conducted in a range of temperatures
and pressure simulating subsurface conditions. Both
gaseous and aqueous samples should be collected
and analyzed.
Thermodynamic models of dissolution-precipitation of
mineral phases (such as WATEQ2 and SOLMNEQF)
can predict some solution equilibria of waste-brine-rock
interactions but have too many limitations to predict all
interactions. Consequently, empirical, laboratory-based
investigations are needed to assess the chemical inter-
actions among injected wastes, injection formations,
and associated formation waters.
The acidic waste was neutralized when it reacted with
the dolomite and siltstone through carbonate dissolu-
tion, generating dissolved carbonate species and carb-
on dioxide. The waste was partially neutralized by the
sandstone through dissolution of clay minerals and ton
exchange augmented by the dissolution of a minor
amount of calcareous material.
At temperatures and pressures lower than those
simulating temperatures and pressures at a depth of
3,000 ft, the alkaline waste was hazardous, whereas it
was not hazardous under the proper depth-simulating
conditions. The silica solid phases of the Mt. Simon
sandstone and Proviso siltstone dissolved in the
alkaline waste.
5.2 Processes Affecting the Geochemical Fate of
Deep-WeH-lnjected Wastes
5J2.1 Overflew of Fate-lnftuencingPtocesses
The geochemical interactions among liquid wastes, forma-
tion waters, and formations have not been well researched
i1 ?t?S,Lep0rt W3S US6d to prepare this summarV. Changes in the reference section have been
r2t ™° pUmmafy'A S'!9o % reVI'Sed Versfon Of this report was Published in 1989 bv «ie Illinois State
Geological Survey as Environmental Geology Note 130.
A detailed description of laboratory products can be found in Appendix B of the full report.
55
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(Brower et a!., 1988; Warner, 1965; van Everdingen and
Freeze, 1971; Gordon and Bloom, 1984; LaMoreaux and
Smith, 1985; Sullivan et a!., 1986).
Achemically incompatible waste-formation system is one in
which precipitates or gases evolve during injection. Chemi-
cal precipitates can accumulate in the void spaces that make
up the formation and reduce its porosity. The reduced
porosity will lead to a reduction in injection efficiency, buildup
of pressures, and possibly complete well failure. The ac-
cumulation of gaseous components, formed as reaction
products, can also plug formation voids, reduce the per-
meability of the receiving formation, and create pressures in
the formation leading to blowouts.
Liquid wastes injected into chemically incompatible forma-
tions (including in-situ brines) may cause the confining strata
to deteriorate and wastes to migrate out of the injection zone
(Gordon and Bloom, 1984).
5.2.2 Partition Processes
Adsorptlon-desorptlon. Butanol, n-hexylamine, i and
phenol were adsorbed on Cottage Grove sandstone in
laboratory tests at elevated temperatures and pressures
(Donaldson and Johansen, 1973). The presence of organic
compounds adsorbed on solid surfaces can retard the rate
of calcite crystallization (Skirrow, 1975). Rapid calcium-car-
bonate precipitation from sea water does not occur until
much of the dissolved organic matter is removed (Skirrow,
1975). The major organic compound in the alkaline waste,
hexachtorocyctopentadiene, can be adsorbed by clay
minerals (Chou and Griffin, 1983). Consequently, the or-
ganfc solutes in the alkaline waste may have retarded the
development of calcite equilibrium.
Precipitation-dissolution. The amount of limestone that
dissolved during laboratory tests with mixed acidic organic
wastes increased with pressure (Goolsby, 1972). Adding
hydrochloric acid (HCI) to a ferric-chloride waste solution
mixed with dolomite resulted in ferric-hydroxide precipita-
tion; adding acetic or citric acid in lieu of the HCI resulted in
no iron precipitation (Hower et al., 1972).
In this study, calcium dissolution increased somewhat at
higher temperature and pressure in the acidto-siltstone sys-
tem, and more so in the acidic-dolomite system. Mag-
nesium showed relatively greater dissolution with increased
temperature and pressure in the acid-dolomite system.
Other studies have demonstrated that calcite is more
soluble when Mg2* is present in the solution. Mg * derived
from the dissolution of dolomite can form Mg-COa com-
plexes, which reduce the activity of CO32- in solution,
inducing further calcite dissolution (Hassett and Jdrinak,
1971; Bemer, 1975). Silica dissolution increased substan-
tially as temperature and pressure increased in the acid-St.
Peter sandstone system. The lack of thermal response in
the St. Peter sandstone indicated that ton-exchange
mechanisms, not just dissolution, contributed to the ap-
pearance of aluminum in the solution. Mixing of the connate
brine, which contained 117 mart, magnesium, 'with the
alkaline waste resulted in the Mg2* disappearing from solu-
tion, apparently due to brucite (Mg[OH]2) precipitation. In
an unrelated study, Mehnert et al. ("1988) found that brucite
precipitated near the injection zone of the Velstool facility at
Marshall, Illinois.
In each alkaline-rock system an increase in temperature and
pressure was associated with tower quantities of calcium in
solution. The reverse was true for sulfates. The solubility
of calcite decreased with temperature and increased with
pressure, although temperature had more effect on solubility
than did pressure. The sulfate equilibria of the alkaline-rock
system could not be resolved using computer models.
Immiscible Phase Separation. In this study under am-
bient conditions, carbon dioxide gas was generated in
acid-rock systems. The relative amount of carbon dioxide
was reduced by an order of magnitude at the higher
temperatures and pressures, presumably due to the greater
gas pressure exerted on the liquid, preventing degassing.
A portion of the dissolved carbon dioxide converted to
carbonates, depending on the pH of the solution.
A major problem that can occur when acidic wastes are
injected into carbonate formations is well blowout, where
gaseous COa escapes to the surfaice. The Cabot Corpora-
tion initially injected a 32% HCI solution and in 1975 their
well erupted. The amount of COa generated far exceeded
its solubility in the reservoir fluids at that HCI concentration.
Since the 1975 incident, the Cabot Corporation reduced the
HCI concentration to avoid this problem. The dilute nature
of the sample used in this project (0.09% HCI) reflected this
concern. No further problems with blowout have been
reported, nor would they be expected based on the
laboratory results presented in this study.
5.2.3 Transformation Processes
Neutralization. Carbonate dissolution is the main con-
tributor to neutralization of acidic wastes. Neutralization
may also be facilitated by dissolution of aluminosiltoate solid
phases in the Mt. Simon sandstone, but dissolution
decreased as temperature and pressure increased. The
neutralization of HCI by a base i« an exothermic reaction.
At tow HCI percentages (<1%), the temperature increase is
negligible. However, thermochernical calculations indicate
that temperature increases can become significant (10°C)
when the amount of HCI is greater than about 8%
(Panagiotopoulos and Reid, 1986).
56
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The dominant mechanismof HCI-waste neutralization in Mt.
Simon sandstone was the dissolution of clay minerals and
ion exchange augmented by the dissolution of a minor
amount of calcareous material. After 15 days of contact, the
waste was rendered nonhazardous by the pH criterion but
remained acidic. The extent of reaction progressed slightly
with an increase in temperature.
The injection of waste acids, particularly inorganic acids in
carbonate formations, has been widely practiced for years.
The chemical interactions in such systems have been dis-
cussed by Kamath and Salazar (1986). The practice has
the obvious attraction of neutralizing a hazardous waste via
acid-base chemistry, and the process may increase the
capacity of the formation to receive injected wastes.
Oxidation-Reduction. See discussion of measuring
oxidation-reduction potential (Eh) in Section 5.6.1,
Waste/Reservoir Characterization.
5.2.4 Transport Processes
Not discussed.
5.3 Major Environmental Factors Affecting
Deep-Well-Injection Geochemical Processes
5.3.1 Geochemical Characteristics of Deep-Well-
Injection Zones
General Data. Sandstones and carbonates are the two
major types of injection materials in Illinois as well as in the
United States (Warner and Lehr, 1977). Injection zones in
Illinois range in depth from approximately 1,550 to 5,540 ft.
Subsurface geologic formations have been used in Illinois
for waste disposal for about 20 years. Currently, nine
Class I injection wells are operating, including two standby
wells, at seven industrial sites. The following describes
some of the geochemical characteristics of the injection
zones used in Illinois.
Water Chemistry. A connate-formation brine sample was
collected from the injection zone (observation well at about
2,400ft in Devonian limestone) at the Velsicolsfle. The brine
was very reduced and tended to oxidize quickly when
removed from the pressure canister. Total dissolved solids
measured about 22,000 mg/L with a pH of 9.07. Brine data
from other sources indicate a range of about 12,000 to
28,100 mg/L (Meents et al., 1952, Illinois State Geological
Survey files, 1977-1985).
Carbonates. The Cambrian-age Potosi dolomite is used
for deep-well injection in Illinois. It is a finely crystalline, pure
to slightly argillaceous dolomite that ranges in thickness
from 100 to 300 ft. The sample used was 95% dolomite with
approximately 5% quartz. Alkaline waste from the Velsicol
pesticide plant at Marshall, Illinois, is injected into a
Devonian limestone.
Sandstones. The Cambrian-age Mt. Simon sandstone is
also used for deep-well injection in Illinois. It ranges from
less than 500 ft to approximately 2,600 ft in thickness. The
formation consists of fine- to coarse-grained, partly pebbly,
friable sandstone. The bulk sample was approximately 90%
quartz, 6% potassium feldspar, and less than 5% clay. The
clays were composed of about 75% illite, 25% expandable
clays, and a trace of chlorite.
Confining Beds. The Proviso siltstone member (Eau
Claire formation) overlies the Mt. Simon sandstone. The
Eau Claire formation is used as the upper confining layer at
two deep-well-injection facilities in Illinois (Brower et al.,
1988). The siltstone is approximately 150 to 300 ft thick and
is predominantly a dolomitic, sandy, feldspathic, slightly
glauconitic siltstone (Willman et al., 1975). The Proviso
sample was composed of approximately 50% quartz, 25%
potassium feldspar, 15% dolomite, and 10% clay. The clays
were composed of approximately 87% illite, 7% chlorite, and
6% expandable clays.
The New Albany shale (upper) and Maquoketa shale (lower)
serve as confining units for injected wastes at the Velsicol
pesticide plant at Marshall, Illinois.
5.3.2 Specific Environmental Factors
pH. The pH of the Velsicol alkaline waste did not react
strongly with the sandstone, siltstone, or dolomite, nor did it
appearto correlate with formation type, time, ortemperature
and pressure.
Eh. Reducing conditions (-154 mV) existed in brines from
Devonian limestone (at about 3,200 ft). The Eh of the
Proviso siltstone-acidic waste system rapidly decreased
during the first 3 days (from +800 to +300 mV at ambient
temperature and pressure). More-reducing conditions were
generally associated with an increase in temperature and
pressure (around +100 mV at 55°C and 11.7 MPa). In the
Proviso siltstone-alkaline waste system, Eh showed little
change at ambient temperature and pressure, but dropped
significantly at 55°C and 11.7 MPa (from +600 to around
+200 mV). See also discussion on measuring redox poten-
tial in Section 5.6.1, Waste/Reservoir Characterization.
Salinity. See the discussion of ton concentration in Section
5.5, Aqueous and Solution Geochemistry.
57
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Reservoir matrix (clays, Iron oxides). The dissolution of
feldspars and clay minerals is characteristically a very stow
reaction.
Temperature. Temperatures of 40°C simulate an [ap-
proximate depth of 1,500 ft, and 55*C, an approximate depth
of 3,000 ft. At a depth of 3,000 ft temperatures range from
50' to 100'C (Roedder, 1959; Bayazeed and Donaldson,
1973).
Time. Some chemical reactions require considerably
longer to go to completion than others (Ostroff, 1965). The
time required to establish dolomite equilibrium in neutralized
acidic wastes is not known. Quartz dissolution-precipitation
at temperatures less than 50'C is extremely slow (Stumm
and Morgan, 1981). Rfteen days was not long enough for
silica dissolution in the the Proviso siltstone to reach equi-
librium, even though the system appeared to be stable at
ambient temperatures.
Pressure. Estimates vary, but injected liquids at a depth of
3,000 ft are subjected to hydrostatic pressure exceeding 10
MPa (Roedder, 1959; Bayazeed and Donaldson, 1973). A
pressure of 6 MPa (871 psig) simulates the pressure at
approximately 1,500 ft, and 11.7 MPa (1,697 psig) a depth
of about 3,000 ft. A few investigators have conducted
laboratory experiments at elevated pressures and tempera-
tures to simulate subsurface conditions. Goolsby (1972)
found that the amount of limestone that dissolved in
laboratory tests with an acidic, mixed organic waste in-
creased with pressure with a concqmitant decrease in
solution pH. Higher temperature and pressure tended to
increase the dissolution of calcium in the acidic waste-jrock
interactions.
5.4 Geochemical Characteristics and Fate of
Hazardous Wastes
5.4.1 Chemical Properties of Inorganic Hazardous
Wastes
This report discusses the chemical properties of extremely
acid/alkaline wastes. Laboratory compatibility studies in-
clude Goolsby (1972) and Bayazeed and Donaldson
(1973).
In this study, an acidic inorganic liquid waste was collected
from the Cabot Corporation plant nearTuscola, Illinois. The
waste liquid is a byproduct of the production of a high-purity,
amorphous silica and contains 0.09% HCI (Bergohson,
1988).
An alkaline liquid waste was collected from the Velsicol
Chemical Corporation at Marshall, Illinois. The brine-like
solution (200,000 mg/L) was the caustic process waterf rom
pesticide manufacturing. It did not react strongly with the
three core samples except for the dissolution of silica. The
pH did not appear to correlate with formation type, tempera-
ture, or pressure.
5.4.2 Chemical Properties of Organic Hazardous
Wastes
The Velsicol alkaline waste contained numerous organic
compounds. Volatile organic hazardous wastes detected in
the solution included six halogenated aliphatic hydrocar-
bons (chloroform, 1,1-dichloroethane, carbon tetrachloride,
trichloroethene, bromoform, and tetrachloro-ethene) and
monocyclic aromatics (benzene, toluene, ethylbenzene,
and o-xylene). Interactions of these materials with reservoir
rock were not evaluated.
Phenol. Phenol was adsorbed on Cottage Grove sand-
stone at elevated temperatures and pressures in static
laboratory tests (Donaldson and Johansen, 1973).
Butanol. Butanol was adsorbed on Cottage Grove sand-
stone at elevated temperatures and pressures in static
laboratory tests (Donaldson and Johansen, 1973).
n-Hexylamine. n-Hexylamine was adsorbed on Cottage
Grove sandstone at elevated temperatures and pressures
in static laboratory tests (Donaldson and Johansen, 1973).
5.4.3 Mixtures of Hazardous Wastes
Laboratory experiments on interaction of acidic wastes
containing adiponitrile, hexamethylenediamine, alcohols,
ketones, and esters with limestone under pressure have
been conducted by Goolsby (1972).
5.5 Methods And Models For Predicting The
Geochemical Fate Of Deep-Well-lnjected
Wastes
5.5.1 Basic Approaches
Accurate thermodynamic data are lacking for many
heterogeneous solid wastes and chemical interactions.
Kinetic data for many reactions under elevated temperature
and pressure are unknown. Because of the lack of data, the
application of thermodynamic models to predict the fate of
injected wastes is tenuous (Sullivan et al., 1986). Brower et
al. (1988) concluded that laboratory compatibility studies on
58
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chemical interactions have betterpredictive value than equi-
librium models.
To avoid using laboratory studies to determine the possible
chemical interactions of specific waste-formation systems,
some investigators have attempted to use computer-
assisted transport and chemical (thermodynamically
based) models to assess the fate and distribution of chemi-
cal constituents in injection scenarios. Scrivner et al. (1986)
concluded that the chemical fate of injected wastes can be
determined by "standard chemical engineering techniques"
and that the concentrations of hazardous constituents are
typically reduced by reactions within the waste or injection
environment. Scrivner and his co-workers at E.I. Du Pont
de Nemours and Company have used computer simula-
tions to model the fate of injected hazardous wastes. The
models considered reaction rates and the equilibrium con-
stants for the dominant reactions. The Du Pont model has
not been validated, and was not publicly available at the time
this summarized research report was published.
5.5.2 Specific Methods and Models
Waste-Reservoir Compatibility. Very few studies of
chemical compatibility have been conducted. Some of
these studies consisted of mixing a sample of the liquid
waste with a synthetic formation-water sample for 4 to 6
hours at room temperature and pressure. Since the mixture
remained free of precipitates, the waste was considered
compatible with the formation fluid (Warner, 1965). Suffi-
cient time is necessary to allow chemical reactions to go to
completion (Ostroff, 1965), and past studies have tended to
ignore the effects of formation temperatures and pressures
on chemical reactions (see, for example, Headlee, 1950;
Bayazeed and Donaldson, 1971; and Barnes, 1972).
Studies involving elevated temperatures and pressures in-
clude Goolsby (1972) (static, limestone and acidic mixed
organic waste); Hower et al. (1972) (static, dolomite and
ferric chloride mixed with several acids); and Donaldson and
Johansen (1973) (static, Cottage Grove sandstone with
various organic compounds: butanol, n-hexylamine, and
phenol). Studies involving flowthrough experiments at am-
bient temperatures with simulated acid wastes from steel
processing and Mt. Simon sandstone were performed by
Bayazeed and Donaldson (1973).
The change in calcium concentrations in the alkaline waste-
rock systems with increased temperature and pressure
suggests that fate or compatibility-type demonstrations con-
ducted under ambient conditions may not generate data that
simulate subsurface conditions. The Proviso siltstone ap-
peared to be stable at ambient conditions, but dissolution of
silica occurred at elevated temperatures and pressures, and
apparently the system had not reached equilibrium after 15
days.
Aqueous and Solution Geochemistry. In some cases,
solution equilibria could be modeled using the ther-
modynamic principles of dissolution-precipitation of mineral
phases. In other cases, empirical, laboratory- based inves-
tigations are needed to assess the interactions among
injected wastes, injection formations, and associated forma-
tion waters.
The thermodynamic model WATEQ2 (Truesdell and Jones,
1974; Plummeretal., 1976; Ball and Jenne, 1979) was used
to help understand the geochemical interactions between
the liquid hazardous wastes and the core samples. This
computer program is based on the equilibrium-constant
approach that predicts the distribution of aqueous species
based on the input chemical data. The program simul-
taneously solves several nonlinear equations by successive
approximation using the continued-fraction approach.
WATEQ2 has a temperature range of applicability of 0°C to
100°C. Equilibrium constants are calculated at a given
temperature using empirical regression depending on the
availability of data for a specific solid phase, or they are
interpolated using a Vant Hoff equation.
The equilibrium constants for calcite dissolution at 6 MPa
and 11.7 M Pa pressure were calculated from changes in the
partial molar volume of the reaction, using the method given
in Skirrow (1975). Dolomite equilibria could not be corrected
for pressure effects because of a lack of reliable data.
The chemical data were also treated by the thermodynamic
model SOLMNEQF (Kharaka and Barnes, 1973).
SOLMNEQF is similar to WATEQ2 in structure and
database, although fewer solid phases are considered. Like
WATEQ2, SOLMNEQF calculates equilibrium constants as
a function of temperature using a Van't Hoff equation, but it
also corrects for pressure over a range of 1 to 1,000
atmospheres (atm) (101.3 MPa). Pressure-corrected equi-
librium constants of solid phases are approximated using
the coefficient of expansion, isothermal compressibility, and
molar volume of the mineral phase. The effects of pressure
on ionic species are not considered. As with all ther-
modynamic models, the results must be interpreted
cautiously owing to discrepancies in reported values for
equilibrium constants, heterogeneous redox equilibria, and
kinetically inhibited reactions. Both models were used in this
study for the sake of comparison.
The only equilibrium relationship indicated by WATEQ2 and
SOLMNEQF was the hydrolysis of chalcedony, a fibrous
form of silica. Chalcedony occurs in sedimentary rocks,
possibly forming from the dissolution of clay minerals
(Jenne, 1988). At higher temperatures and pressures, the
increased dissolution of silica was due to the increased
59
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solubility of chalcedony. The waste-brine system was also
equilforated with chalcedony. It appears that the amount of
silfca In solution in this system could be estimated using
equilibrium constants for the hydrolysis of chalcedony.
Therefore, this approach could be used in the modeling to
predict the dissolution of chalcedony in similar deeprwell
scenarios.
The Potosi acidic system was neutralized by the dissolution
of dolomite, but dolomite equilibrium was not attained at any
temperature; the solution was supersaturated with respect
to a dotomitto-carbonate phase. The time required to estab-
lish dolomite equilibrium is not known.
Nordstrom and Ball (1984) concluded that ton-association
models (such as WATEQ2 and SOLMNEQF) cannot be
used to predict mineral solubilities or solute activities at tonic
strengths exceeding 0.6 mole/L. The methods and concepts
used by these programs, such as estimating activity coeffi-
cients, were based on solutions with much tower tonic
strengths than those of the alkaline waste-rock systems
(about 4.6 mole/L). Consequently, for some reactions, the
tonic strength of waste sample was beyond the range where
reliabledataforfon associations and hydrolytto reactions are
available. Although these ton-association models have
limitations when applied to concentrated solutions, they
were used in this study to help explain observed trends.
The inability to determine very tow solution concentrations
(such as those of magnesium) is a limitation in using equi-
librium models coupled with laboratory studies to predict
chemical interactions. ,
Thermodynamics and Kinetics. Heats of reaction deter-
mined under ambient conditions provide an indicator of
chemical reactivity. Such measurements represent the
summing of many exothermic and endothermic reactions
and must be interpreted cautiously. !
As with all thermodynamto models, the results must be
interpreted cautiously because of discrepancies in reported
vatuesforequilibrium constants, heterogeneous redox gqui-
libria, and kinettoally inhibited reactions. Both models were
used in this study for the sake of comparison. !
5.6 Laboratory Procedures And Protocol^
5.6.1 Waste/Reservoir Characterization
General Characterization. The brine sample from Velsfeol
was stored under nitrogen in the field at 345 Pa in a
stainless-steel pressure canister to minimize oxidation and
degassing. Solution concentrations of chloride, fluoride,
nitrate, sulfate, and iron (Fe43) were determined using an
ton chromatograph. Solution concentrations of metals and
other cations were determined using an inductively coupled
argon-plasma (ICAP) emission spectrophotometer. Sam-
ples of the alkaline liquid were analyzed by gas chronnatog-
raphy/mass spectroscopy (GC/MS). Volatile organic com-
pounds were determined by direct GC injection. Gas
samples were characterized with a gas chroma-tograph.
The pH and electrical conductivity of the solutions were
measured by electrode (American Public Health Associa-
tion, 1985). Mineratogical composition of the three core
samples was determined by x-ray diffraction using proce-
dures discussed in Russell and Rirnmer (1979).
Heats of reaction between the liquid wastes and the disag-
gregated core sample were determined with a Parr 1451
solution calorimeter (see Ramette [1984] for details on
procedures and theory of operation).
Measurement of Eh. The oxidation-reduction potential (Eh)
of a solution is usually measured using a platinum electrode
or a saturated catomel electrode. In this study, Eh electrode
measurements were made using procedures described by
the American Public Health Association (1985) and the Eh
reference solution described by ZoBell (1946) and Wood
(1976). Such measurements are difficult to duplicate, and
readings of a sample solution with two identical electrodes
may vary by as much as 50 mV depending on the stability
of the solution, the electrode, and the skill of the analyst
(Nordstrom et al., 1979). Some investigators (summarized
by Nordstrom et al., 1979) have found good agreement
between measured Eh and a dominant redox couple, while
others regard Eh measurements as only a qualitative indica-
tion of redox potential.
In this study, ton chromatography (1C) was applied to deter-
mine the concentrations of Fe2+ and Fe3+ to derive Eh
values for comparison with the electrode-based observa-
tions. As an alternative method, ferrous ton was determined
by titrating the solutions with dichromate using a combina-
tion platinum electrode (Skoog and West, 1976). The IC-
based Eh values showed poor agreement with the
electrode-based observations, and the investigators con-
cluded that ton chromatography is unlikely to yield reliable
iron-couple data without considerably more research. Con-
sequently, the redox potentials reported in the study should
be considered qualitative values.
When a redox electrode is immersed in a solution, the
observed reading may reflect the summation of different
individual redox couples, and this combined potential may
differ greatly from that of any known potential (Bonn et al.,
1979). Lindberg and Runnells (1984), inan analysis of more
than 600 groundwater samples from diverse geographic
60
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areas, concluded that aqueous oxidation-reduction reac-
tions in natural groundwater systems are generally not at
equilibrium. Multiple redox couples present in individual
samples yielded computed Eh values spanning as much as
1,000 mV. Thus, the potential of a platinum electrode in a
redox mixture may be a poorly defined average of the
potential of redox couples present. Furthermore, the point
at which the electrode has equilibrated with a solution is not
always certain; often the reading may slowly drift rather than
stabilizing on a value. The contribution of each couple to
the observed measurement is an unknown function of its
concentration. For example, nitrate stabilizes potentials at
200 to 400 mV and prevents the formation of Fe2+ in soils
{Ponnamperuma, 1972).
5.6.2 Static Interaction Tests
Batch pressure reaction vessels were operated in tempera-
ture-controlled water baths to represent subsurface tempera-
ture-pressure regimes. Vessels could not be pressurized to
exceed 11.7 MPa without risking failure of the rupture disks.
5.6.3 Dynamic (Flowthrough) Tests
These tests were not used.
5.614 Quality Assurance/Quality Control Procedures
Quality-control standards obtained from the U.S. EPA were
used to verify results at the time of analysis.
5.7 Case Studies
This study presents the results of laboratory waste-reservoir
interaction tests of acidic, inorganic waste from the Cabot
Corporation nearTuscola, Illinois, and alkaline liquid waste
from the Velsicol Chemical Corporation in Marshall, Illinois.
See Section 5.4.1 for additional information on the chemical
properties of the wastes.
5.8 Further Research Needs
No further research needs were discussed.
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Gordon, W., and J. Bloom. 1984. Deeper Problems: Limits
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Undberg, R. D., and D. D. Runnelis. 1984. Ground Water
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225:925-927.
Meents, W. F., A. H. Bell, O. W. Rees, and W. G. Tilbury.
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Nordstrom, D.K..E. A. Jenne, and J.W.Bali. 1979. Redox
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Ponnamperuma,F.N. 1972. The Chemistry of Submerged
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Roedder, E. 1959. Problems in the Disposal of Acid
Aluminum Nitrate High-Level Radioactive Waste Solutions
by Injection into Deep-Lying Permeable Formations. U.S.
Geological Survey Bulletin 1088,65 pp.
Russell, S.J., and S.M.Rimmer. 1979. Analysis of Mineral
Matter in Coal, Coal Gasification Ash, and Liquefaction
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Skoog, D. A., and D. M. West. 1976. Fundamentals of
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CHAPTER SIX
RESEACH SUMMARY NO. 5
CURRENT GEOCHEMICAL MODELS TO PREDICT THE FATE OF HAZARDOUS
WASTES IN THE INJECTION ZONES OF DEEP DISPOSAL WELLS
6.1 Overview
6.1.1 Origin and Content
Source: Current Geochemical Models to Predict the Fate
of Hazardous Wastes in the Injection Zones of Deep \Dis-
posal Wells. Draft Report. January 1988. Lawrence
Berkeley Laboratory Report LBL-26007. Draft Report
Prepared for U.S. Environmental Protection Agency. 118
pages. *
Authors: John A. Apps, Earth Sciences Division, Lawrence
Berkeley Laboratory, University of California, Berkeley,
California 94720.
Contents: This report contains eight major sections: (1)an
Introduction to the EPA regulations covering no-migration
petitions for deep-well injection of hazardous wastes and
how the petitions relate to geochemical modeling;** (2) a
discussion of the reactions that must be modeled giver) the
chemical conditions expected in the injection zone; (3) the
equations of state that must be used;*** (4) the availability
of thermodynamic data; (5) modeling of nonequilibh'um
systems; (6) the availability of geochemical-modeling com-
puter codes; (7) criteria affecting the satisfactory chemical
modeling of waste injection; and (8) conclusions and recom-
mendations.
6.1J2 Major Conclusions
• The geochemical modeling of the fate of hazardous
wastes in saline aquifers contained in deep sedimen-
tary formations is in a preliminary state of development.
Geochemical modeling has not been adequately tested
in the field or the laboratory to show that it can be used
to make quantitative predictions under all conditions.
• Many diverse mechanisms affect the fate of hazardous
constituents in deep aquifers. Mo single computer code
is currently capable of modeling all processes. Avail-
able codes must be selected by how they perform on a
particular application.
• The state of the art in many areas of geochemical
research is embryonic. For example, activity coeffi-
cients of ions in strongly mixed electrolytes (i.e., brines),
the thermodynamic properties of clays, and the ther-
modynamics of adsorption have yet to be accurately
determined. Thermodynamic data for many minerals
and organic aqueous species are unavailable. There-
fore, much preparatory research must be done before
suitable simulations can be conducted. Existing ther-
modynamic databases used with geochemical model-
ing codes will require close scrutiny before they are
used.
• Although the literature describing the adsorption of
inorganic and organic species on clays is substantial,
integrated compilations of data are not available for
practical applications. Additionally, suitable adsorption
or ton-exchange models that can be used for the diverse
range of conditions expected in deep-well environ-
ments are also not available.
• Information on the kinetics of both heterogeneous and
homogeneous reactions relevant to the fate of hazard-
The report was undergoing peer review at the time this summary was prepared and had several minor errors in the
text and references section. The author was consulted where questions arose, and corrections were incorporated
into this summary.
This part of the report is not summarized here because it deals with regulatory rather than technical matters.
Except for simple equations, mathematical discussions in the report are not reproduced here.
64
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ous compounds is fragmentary and often inadequate.
Careful evaluation of the literature and experiments will
often be required.
A substantial number of computer codes are available
to evaluate the distribution of chemical species in solu-
tions. Computer codes that model mass transfer or
mass transport with simultaneous chemical reactions
are currently limited in availability and/or scope. Codes
should be selected with caution to ensure aptness.
There are two types of distribution-of-species codes—
the equilibrium-constant codes and the Gibbs free-
energy-minimization code. Both types are mathemati-
cally equivalent, but in principle the latter is more flexible
and easier to use. Most available codes are equi-
librium- constant codes; additional effort is required to
modify these types for use in deep-well injection
studies.
Chemical-transport codes fall into two categories—the
"one-step" or the "two-step" codes. For the one-step
codes, highly nonlinear computationally intensive equa-
tions must be calculated. The two-step codes have had
problems with drift, but tend to be more flexible in a
diverse range of problems.
Some geochemical codes have been verified, but few
have been used to test the validity of the underlying
models. There is an urgent need to conduct additional
field- validation studies under any conditions, especially
those in deep disposal wells.
6.2 Processes Affecting the Geochemical
Fate of Deep-Weil-lnjected Wastes
6.2.1 Overview of Fate-Influencing Processes
Many waste streams contain organic constituents, which
may be partly or completely miscible in water. These
streams adsorb on minerals and can hydrolyze, decarboxy-
late, or be destroyed through bacterial action. The fate of
those containing toxic heavy metals (such as cadmium,
hexavalent chromium, mercury, nickel, and lead) must be
considered in light of potential interactions between the host
rock and the waste stream and the presence of complexing
agents.
Immobilizing Reactions. The mechanisms that can lead
to immobilization of hazardous constituents in the waste
stream in the injection zone are precipitation/co-precipita-
tion, ion exchange, or adsorption. Total immobilization is
impossible because all precipitates have a finite solubility.
Also, all ion-exchange and adsorption processes must have
finite, though sometimes very small, reversible-exchange or
adsorption coefficients. The immobilization of wastes may
be further complicated by (1) metastability and the slow
kinetics of heterogeneous processes, and (2) waste
streams with more than one hazardous constituent. To
obtain an injection-well permit based on geochemical immo-
bilization of wastes, an operator must show that the hazard-
ous waste will be immobilizedforthe next 10,000 years (i.e.,
the concentration in solution decreases well below
regulatory standards or three orders of magnitude below
current detection limits). In reality, many hazardous con-
stituents, particularly water-miscible organic species, are
not immobilized effectively.
The effects of immobilization or decomposition may be
enhanced by viewing the injection zone as a reactor in which
constituents added to the waste stream can react benefi-
cially with the hazardous constituents. Examples of such
innovative techniques include (1) wet combustion (Smith
and Raptis, 1986) and (2) addition of a fixation agent such
as tetramethyl ammonium ion to a waste stream to enhance
sorption of organic wastes on smectite clays (Barrer and
MacLeod, 1955). Much research is needed to demonstrate
that these techniques are both effective and reliable; the
added cost of such research and treatment procedures may
exceed the benefits of the enhanced containment potential.
Reactions for Which Kinetics May Be Important. Many
reactions in an injection zone proceed rapidly enough that
a local, reversible equilibrium may be assumed with
reasonable confidence. Most homogeneous reactions in
the aqueous phase (with the exception of some oxidation-
reduction reactions), surface-adsorption reactions, and ion-
exchange reactions are of this type. At the other end of the
spectrum, some reaction rates are so slow that equilibrium
is not attained even after 10,000 years. In this case,
hydrologic arguments alone must be used to demonstrate
containment. Several categories of chemical reactions
proceed sufficiently slowly that local equilibrium cannot be
assumed, yet substantial progress toward equilibrium is
achieved withjn the 10,000-year time frame. Types of reac-
tions that may require analysis using a kinetic equation
include:
• Heterogeneous precipitation of secondary minerals or
solid phases from solution
• Oxidation/reduction reactions in the aqueous phase
• Hydrolysis, decarboxylation, dechlorination, etc., of or-
ganic compounds
65
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• Bacterial destruction of nitrogen compounds, hydrocar-
bons, and hatogenated hydrocarbons
For predictive modeling, the range of magnitudes forthe rate
constants ki(s"1) of chemical reactions that fall between the
two reaction rate extremes must be determined.
The actual attenuation of a species in solution to nonhazar-
dous levels, whether through adsorption, precipitation, or
decomposition, depends on the following factors:
• Rate constant
• Initial concentration of the species and other participat-
ing species
• Substrate area or nucleatfon sites if a heterogeneous
reaction is involved
• Concentration levels below which the chemical is no
longer considered hazardous
Figure 6-1
-4
Plot of first-order decomposition rate constant
vs. time for given attenuation ratios, Cj,t/Cj,o
CT -10 =.
Cl
.2
234
log t (yrs)
Figure 6-1 is a graph of the simple case in which removal of
the hazardous constituent is a first-order equation with
respect to time for various attenuations in concentration,
from10~1 to10~10. For reaction half-lives greaterthan 1 year,
rate constants greater than 10"V1 will not be of concern,
because the hazardous constituent will be immobilized or
destroyed near the injection well (although formation of
pore-blocking precipitates might adversely affect well opera-
tions). At the other extreme, rate constants less than 10~11 s"
1 will not result in significant attenuation in 10,000 years.
Consequently, for "irreversible" first-order reactions of inter-
est, the rate constant generally falls between 10~11 and
10V.
6.2.2 Partition Processes
Three groups of partition processes are discussed In this
report:
• Adsorptfon-desorption
• Precipitation-dissolution
• Immiscible phase separation
These mechanisms are discussed In the sections below.
Adsorption-desorption. One of the most important
mechanisms by which small amounts of hazardous con-
stituents can be removed from awasrte stream is adsoiption.
Adsorption is a process in which chemical species in solu-
tion attach to the surface of a solid substrate. The process
should be thermodynamically reversible, but many adsorp-
tion experiments have demonstrated apparent irre-
versibility.* Some researchers consider adsoiption an irre-
versible phenomenon (Van Genucliten et al., 1974). The
meaning of adsorption is further confused, since man-
ganese oxide, clay interlayer, and zeolite-ion exchange are
often defined as adsorption mechanisms. Thus, parallel
treatments have been developed for ion exchange and
adsoiption, which are essentially the same phenomenon.
For further information, see the discussion of Reservoir
Matrix (Section 6.3.2) and Adsorption (Section 6.5.2.3).
Precipitation-dissolution. Any waste stream can react
with the host rock or ground water to produce undessirable
side reactions; separate gas phases can form or pore-clog-
ging solids can precipitate. Several conditions contribute to
Editor's note: Rao and Davidson (1980) have suggested three major causes of apparent irreversibility in
adsorptfon-desorptfon experiments: (1) artifacts created by some aspect of the experimental method, (2) failure to
establish complete equilibrium during the adsoiption phase, and (3) chemical and/or microbial transformations
during the experiment.
66
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precipitation. For example, even before injection begins,
waste streams can be supersaturated with either benign or
hazardous constituents; precipitation in the zone near the
well bore could seriously decrease injection rates unless
pumping pressure is increased. Increases in pumping pres-
sure can lead in turn to hydrofracturing and escape of the
injected waste outside the injection zone. Neutralization of
exceedingly acid or alkaline wastes involves substantial
hydrolysis and dissolution of host-rock minerals, which may
also lead to precipitation.
During the late 1950s, investigators focused on slow, sur-
face-controlled dissolution reactions involving the oxidation
of suifides in abandoned coal and nonferrous-metals mines
and the waste and spoils heaps from such operations. Also,
at this time, geochemists conducted the first tentative
studies on the dissolution kinetics of silicates (Garrels and
Howard, 1957). Studies of the hydrometallurgical treatment
of ores to recover metals concentrated on many matters
identical to present concerns (Burkin 1966, 1983). New
techniques for characterizing and observing surfaces, such
as ESCA, SIMS, and SEM, were introduced and provided
an added incentive to interpret correctly the dissolution or
precipitation mechanisms involved in hazardous-waste in-
jection.
Except for carbonate dissolution by strong acids, most
heterogeneous reactions involving dissolution or precipita-
tion of solid phases occur slowly (in relation to the 10,000-
year-containment standard) at injection-zone temperatures.
Unfortunately, information on the kinetics of precipitation or
dissolution of relevant phases is sparse. For example, a
review paper onthis subject by Lasaga (1981 a) summarized
investigations of the kinetics of only 14 minerals, and studies
of only a few more minerals have been completed since that
time. Geochemical research has tended to address specific
issues, such as:
• Feldspar weathering (Wollast, 1967; Helgeson, 1974;
Weare et al., 1976; Petrovich, 1976; Petrovich et al.,
1976; Berner and Holdren, 1977, 1979; Holdren and
Bemer, 1979; Fleer, 1982; Helgeson et al., 1984)
• Quartz-dissolution kinetics (Apps et al., 1975; Rimstidt
and Barnes, 1980; Bird and Boon, 1984; Bird et al.,
1986)
• Carbonates (Sjoberg, 1976; Plummer and Wigley,
1976; Plummer et al., 1978)
• Sulfates such as gypsum and anhydrite
Dissolution or precipitation studies of other minerals are
uncommon (Lasaga, 1981b, 1984).
Only during the last decade have researchers attempted to
understand mineral dissolution and precipitation kinetics in
multicomponent heterogeneous systems (Dibble and Tiller,
1981; Lasaga, 1981 a,b,c; Petrovich, 1981 a,b; Aangaard and
Helgeson, 1982; Helgeson and Murphy, 1983; Helgeson et
al., 1984). Absolute-rate theory or transition-state theory
(Glasstone et al., 1941) is often used to interpret heter-
ogeneous reactions. The general theory proposed by Aar-
gaard and Helgeson (1982) is complicated, yet does not
quantify actual dissolution or precipitation mechanisms,
which must be determined by experiment. Unfortunately,
incongruent dissolution or precipitation and secondary
nucleatfon effects, as well as mechanisms in which H+, OH-,
or other dissolved species catalyze reactions, may compli-
cate the interpretation of experiments.
When a solution is grossly undersaturated or super-
saturated with a given mineral, precipitation-dissolution
rates are a function of the surface area of the mineral (or,
more accurately, the growth of dissolution sites on the
surface of the mineral) and whether the rate is zero-order
(independent of the concentrations of reactants in solution).
Wood and Walther (1983) evaluated dissolution rates of a
number of rock-forming aluminosilicates, quartz, and corun-
dum, and found that reaction rates were consistent with the
equation:
log k =-2,900/1-6.85
when all data were normalized to the number of gram-atoms
of oxygen per square centimeter (cm2) of mineral surface.
Wood and Walther (1983) proposed that zero-order dissolu-
tion kinetics of aluminosilicate minerals might be determined
by a simple linear equation when their surfaces are normal-
ized to the number of gram-atoms of oxygen per cm2.
Unfortunately, more recent studies (Murphy and Helgeson,
1984) have questioned the validity of this simple model.
When gross supersaturation occurs, the thermodynamically
most stable phase does not necessarily precipitate. In fact,
the nucleating and precipitating phases are functions of the
degree of supersaturation, the surface tension of the
precipitating phase (Sohnel, 1982), and the growth
mechanism (Stranski and Totomanow, 1933; Gutzow and
Toschev, 1968).
As equilibrium is approached, back reactions, which depend
on the activities of the species in solution, begin to retard the
overall rate. The participating species and the associated
rate laws cannot be identified readily except by direct ex-
perimentation. Original investigations may be necessary,
however, to clarify the kinetics of reactions unless literature
describing the kinetics for the phase of interest is available.
67
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For example, Petrovich(1986) has developed a generalized
scheme for explaining the kinetics of aluminosilicate
precipitation and dissolution.
Immiscible Phase Separation. Most deep aquifers con-
tain such carbonates as calcite, dolomite, ankerite,' or
sWerite; thus, substantial quantities of carbon dioxide may
be generated and form a separate CCte-rich phase, i
6.2.3 Transformation Processes
The two transformation processes discussed in this report
are: \
• Hydrolysis
• Thermal degradation
Hydrolysis. Mabey and Mill (1978) reviewed hydrolysis
rates of organic compounds. Their results indicate that
hydrolysis involves the participation of either H+, OH", or
water. The overall rate is the sum of the three mechanisms:
KL= kB[OH1 + Ka[H+] + kN
The (pseudo) first-order hydrolysis constants for different
organfcs vary widely, from 10~2 to 10"1VS. The region of
concern covers a broad range of constants, as indicated in
Section 6.2.1 (i.e., 10"6 to 1(r1V1).
Thermal Degradation. Smith and Raptis (1986) sug-
gested using the injection zone as a reactorfor wet combus-
tion.
6.2.4 Transport Processes
Transport processes were not addressed specifically.
6.3 Major Environmental Factors Affecting
Deep-Well-Injection Geochemical Processes
63.1 Geochemical Characteristics of Deep-Well-lnjeclion
Zones
General Data. Most deep-well-injection facilities' are
operated by the chemical industry, and nearly two-thirds of
the injected-waste volume originates from manufacturers of
organic chemicals. Another 25% of the volume originates
from petroleum-refining and petrochemical industries. The
balance is generated by the metals, minerals, and
aerospace industries, and commercial facilities receiving
wastes from many industrial sources. Most large plants
have onsite injection wells (U.S. EPA, 1985).
It is difficult to generalize about waste-stream compositions
because of their diversity. Most are relatively dilute, i.e.,
greater than 90% of the stream by weight is water. Waste
streams can be either exceedingly acid or alkaline, can
contain organic constituents, which can be partly or com-
pletely miscible in water, or can contain a variety of toxic
heavy metals. Table 6-1 summarizes the most likely range
of chemical characteristics of injected fluid.
Table 6-1 Variations in Chemical Parameters for Injected
Fluid
Parameters
TDS
pH
EH
Organic
Inorganic
Heavy metals,
e.g., Cd, Cr6+,
Hg, Ni, Pb
Unit
mg.kg"1
mol.kg"1
mV
wt%
molal
ppm
Range
1,000—300,000
2—12.5
-1,000— f- 1,500
0—5 or more
0—30
0—500
Class I wells are located where deep sedimentary basins
exist. Most are concentrated along the Gulf Coast of Texas,
Louisiana, Arkansas, and Mississippi and in states overlying
the Illinois Basin, i.e., Illinois, Indiana, Ohio, and Michigan
(U.S. EPA, 1985). Approximately two-thirds of the wells are
located in Texas and Louisiana, which receive 90% of the
injected wastes. Injection sites are also concentrated in
Texas north of the Oachita uplift and in Oklahoma and
Kansas. A relatively small number of wells islocated in
California and Florida. Table 6-2 summarizes the range of
environmental and reservoir conditions most likely to be
encountered in a typical Class I zone.
Water Chemistry. Table 6-2 presents the most likely con-
centration ranges of major ions in solution.
Lithology. The injection zones of Class I wells are normally
arenaceous (sands and sandstones) aquifers, confined by
argillaceous (clay and shale) horizons that can be tens to
hundreds of feet thick. Occasionally, the injection zone is
limestone, a beneficial choice for the disposal of acid wastes
provided precautions are taken.
68
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Table 6-2 Chemical Parameter Variation in the Injection
Zone Environment
Parameter
Unit
Range
T
P
Formation Fluid
TDS
PH
Eh
Na
K
Mg
Ca
Al
SiOa
HCO3-(alkalinity)
SC-42-
F
cr
Mineral Assemblage
Quartz
Plagioclase
K Feldspar
Carbonate, e.g.,
Calcite
Dolomite
Siderite
Clays
Smectites
Illite
Chlorite
Mixed-layer clays
°C
bar
mg.kg'1
mol.kg
mV
mg.kg"1
1-250
0.1-1.0
wt%
20-50
50-300
10,000-350,000
5.0-8.5
0.0-500
5,000-70,000
1-2,400
20-36,000
5-50
10-6,000
1-3,000
5,000-150,000
30-85
3-10
3-10
0-15
0-15
0-5
0-10
0-10
0-10
0-10
6.3.2 Specific Environmental Factors
pH/Eh. The pH range in the deep-well environment
generally ranges between 5.0 and 8.5, and typical Eh from
about 0 to -500 mV. See Table 6-2.
Salinity. The maximum salinity range likely to be found for
total dissolved solids (TDS) is from 10,000 to 350,000
mg/kg. See Table 6-2.
Reservoir Matrix. Table 6-2 shows the breakdown of
mineral assemblage. In particular, exchange sites on clays,
and perhaps zeolites, affect adsorption reactions. Many
cations can be readily exchanged on natural clays and
zeolites, thereby making them very effective scavengers for
certain heavy metals and some organic species. Because
smectites, illites, and kaolinite make up a substantial portion
of clays in typical sedimentary injection zones, the cation-
exchange properties of such clays are of particular interest.
Aargaard and Helgeson (1983) used the site-mixing theory
(see Section 6.5.2.1) to investigate the assumed ideal
mixing of atoms on sites in montmorillonites, illites, and
mixed-layer days.
Kent et al. (1988) classified minerals into fourcategories that
have distinctly different properties when adsorption
phenomena are considered:
• Simple oxides, such as silica (SiO?) or corundum
(AlaOa), with crystal lattices that are electrically neutral,
i.e., they possess no residual charge except on surface
sites resulting from the discontinuity caused by the
surface itself.
• Multiple-surface-site minerals, such as silicates,
aluminosilicates, and complex oxides, which also have
no residual charge except on surface sites, but the
presence of more than one type of atom in the crystal
lattice means that at least two different types of adsorp-
tion sites exist at the surface. Among these are
feldspars, which commonly occur as detrital or secon-
dary minerals in deep aquifers. For example, potash
feldspar, KAISiaOs, would expose both SiOH and AIOH
surface groups.
• Fixed-charge minerals, which have a permanent
negative charge due to the presence of substitutions in
the lattice that are compensated by electrostatically
bound exchangeable cations. Clays, zeolites, and
manganese oxides fall into this category. Clays, with
their large surface areas and cation-exchange capacity,
are important adsorbents in the deep-well environment,
and they are expected to dominate the total adsorption
capacity of the injection zone. Zeolites and manganese
oxides are normally less important. Deep sedimentary
aquifers, unless contaminated by ash-falls, are unlikely
to contain zeolites, and manganese oxides are present
in minor amounts or have been replaced by man-
ganese-containing carbonates. Both mineral types can
be artificially created in an injection zone, however,
depending on the composition of the waste stream.
• Salt-type minerals, such as calcite (CaCOs) and gyp-
sum (CaSO4.H2O), which are formed from ionically
bonded anions and cations. Consequently, the surface
area of these minerals is arranged in a grid of negatively
and positively charged sites corresponding to sections
through the tonic lattices of the solid. Carbonates are
frequently an important constituent of saline-aquifer
host rocks. Therefore, the ion-exchange mechanisms
69
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discussed in Section 6.5.2.3, for example, could be
effective in removing heavy metals.
The adsorption properties of organic matter must also be
considered because the specific surface area of organic
detritus could be quite large and could significantly con-
tribute to the total adsorption capacity of the sedimentary
rock contained in the injection zone (Sposito, 1984; Karick-
hoff, 1984). Kerogen, often a major constituent of shales,
can have a wide variety of nitrogen, oxygen, or sulfur (NOS)
functional groups attached to its surface, such as -SQsH,
-NH2, -COOH, -OH, as well as saturated hydrocarbons and
aromatic rings. NOS functional groups, in particular, can be
reactive adsorbents.
Temperature. The maximum temperature range likely to
be found in injection zones is 20'C to 150°C. See Table 6-2.
Pressure. The most likely pressure range in injection zones
is 50 to 300 bars. Higher pumping pressures required as a
result of precipitation reactions caused by waste-reservoir
interactions could exceed EPA-mandated limits and lead to
hydrofracturing and bypassing of the zone by the injected
fluid.
6.4 Geochemical Characteristics and Fate of
Hazardous Wastes i
6.4.1 Geochemical Characteristics and Fate of Inor-
ganic Hazardous Wastes
The chemical characteristics and fate of inorganic hazard-
ous wastes were not specifically discussed.
6.4.2 Geochemical Characteristics and Fate of Organic
Hazardous Wastes
The chemical characteristics and fate of organic hazardous
wastes were not specifically discussed.
6.5 Methods and Models for Predicting the
Geochemical Fate of Deep-Well-lnjected Was-
tes
i
5.5.? Basic Approaches
Geochemical modeling uses information previously ac-
quired through theory, experiment, and testing to predict the
geochemical evolution of a system. It must be coordinated
with hydrotogfc modeling, laboratory studies, and field tests,
all of which are essential in the no-migration petition
process. Most geochemical modeling performed has not
been relevant to hazardous-waste injection. Therefore,
whether modeling can convincingly demonstrate how a
particular injection scenario will perform is largely unproved.
Geochemical modeling of deep-well injection of hazardous
wastes requires an understanding of physicochemical
phenomenathat have been studied as components of many
disciplines, including soil chemistry; clay chemistry and
mineralogy; aqueous geochemistry; fooiler-water chemistry;
hydrometallurgy; the physical chemistry of strong
electrolytes; process chemistry; engineering; physical or-
ganic chemistry; and environmental chemistry. Scientists
are rarely required to integrate so many fields to address a
technological problem of such complexity. Each subdis-
cipline has its own specialized literature, so it is difficult to
determine whether even state-of-the-art techniques are
sufficiently well developed that practical answers can be
found for chemical problems relating to hazardous-waste
disposal.
The geochemistry of injected waste is modeled to provide
answers to such questions as:
• What type of compounds and what amount will
precipitate or be adsorbed from a solution of a given
composition?
• What will be the saturation concentration of a hazard-
ous constituent after precipitation?
• What will be the effect of the dissolved constituents of
a hazardous waste on the saturation concentration?
• How will the concentration of a hazardous waste in
solution be affected by reactions of the waste stream
with ground water or host rocks?
• Are there unforeseen chemical consequences of inject-
ing a waste stream?
• How fast will a given substance precipitate or be ad-
sorbed from solution?
A number of computer codes, discussed below, have been
developed to try to answer these questions.
5.5.1.1 Geochemical Computer Codes
Existing geochemical modeling codes can often provide
answers to the questions listed above if thermodynamic
data relating to the participating species and phases are
available (see Thermodynamic Data Bases in Section
6.5.1.3). However, no available computer code can solve
all the problems of predicting waste migration in subsurface
aquifers. For all but extremely simple systems, waste
70
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migration must still be predicted using several techniques,
and computer codes only facilitate the calculations.
Computer codes may be used to:
• Reduce and evaluate experimental data.
• Calculate thermodynamic properties of phases or
species at temperatures and pressures other than
those at standard-state conditions using equations of
state algorithms (see discussion in this subsection).
• Determine the distribution of species at equilibrium,
given the principal extensive and intensive parameters
of the systems (distribution codes—see Aqueous and
Geochemistry Models, Section 6.5.2.4.
• Predict the evolution of a chemical system, either as a
function of the reaction progress (the amount of material
reacted) or as a function of time (mass-transfer codes).
Such codes do not account for the spatial distribution of
reactants to products (see Aqueous and Geochemistry
Models, Section 6.5.2.4).
• Predict the chemical evolution of a system in which the
chemical reactions and chemical transport proceed
simultaneously (chemical-transport codes)—see Trans-
port Models, Section 6.5.2.5.
Equations of State. An equation of state relates the
measurable physical and compositional properties of a
phase with its thermodynamic properties as a function of
parameters such as pressure and temperature. An equa-
tion is usually valid only over a specified range of tempera-
tures, pressures, or compositions. Outside this range it may
prove seriously defective. The principal equations of state
relevant to deep-well injections can be subdivided into four
categories:
• Pure and multicomponent solids (see Reservoir Char-
acterization, Section 6.5.2.1).
• Pure and multicomponent fluids (see Waste-Reservoir
Interactions, Section 6.5.2.2).
• Component behavior in liquid aqueous systems (see
Aqueous and Solution Geochemistry, Section 6.5.2.4).
• Systems involving interfacial phenomena (see Adsorp-
tion, Section 6.5.2.3).
Development of Computer Codes. To model geochemi-
cal transport, the researcher must create a mathematical
model of the physical and chemical environment in which
hazardous waste transport occurs. To simulate the behavior
of the system in response to various perturbations overtime,
the system must be represented numerically. This process
usually involves writing algorithms representing the evolu-
tion of the system over time in coded form as a set of
instructions to a computer, which performs the calculations.
Errors can occur at all stages of the modeling and simula-
tions process. To help reduce error:
• The physical and chemical conditions and processes
must be correctly identified to create a realistic mathe-
matical model.
• Even if the physical and chemical conditions are cor-
rectly identified, they must be represented in the correct
mathematical form.
The mathematical model must be represented by algo-
rithms acceptable for digital computation and the correct
boundary conditions identified.
• The algorithms must be correctly coded for use by a
digital computer.
• In conducting a computer simulation, the correct initial
conditions and parameter values must be identified and
chosen.
• Laboratory or field data must be correctly measured
and/or analyzed.
• Apredictive simulation must be made overatime period
for which the model assumptions and parameters are
valid.
6.5.1.2 Model Verification, Calibration, and Validation
An important aspect of quality assurance in geochemical
modeling is the verification, calibration, and validation of
computer codes.
Verification. In verification, the code is tested to ensure that
it yields a correct or nearly correct answer with specified
input data. Models are verified by comparing the numerical
with the analytical solutions, by back-calculating a result to
ensure that it is consistent with the algorithm and input data
used, or by directly comparing the results with the results of
another code capable of solving the same problem. In most
cases, verifying geochemical codes is relatively straightfor-
ward.
Several verification studies of geochemical models have
been conducted. In these studies, the output of one code
was compared with another and the same problem was
71
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initially proposed for solution. These studies include one by
Nordstrom et al. (1979) on 14 distribution-of-species codes,
a comparison of EQ3/6 with PHREEQE (Intera Environ-
mental Consultants, 1983), and a study by Kincaid and
Morrey (1984). Verification studies involving chemical-
transport codes include those by Miller and Benson (1983),
Cederberg et al. (1985), and Liu and Narasimhan (1989b).
Calibration. Calibration of transport or chemical-transport
codes is performed when some but not all of the parameters
used have been assigned correct initial values. Correct
values are estimated by comparing the computer simulation
with observed field data and then adjusting the unknown
parameters until a good fit is obtained. The simulation might
be tested against several sets of field data to obtain a range
of values for the parameters in question. The calibration
process typically is used when certain parameters cannot
be measured directly, either in the laboratory or in the field.
Validation. Codes are validated by testing the model on
which the code is based. For simple distribution-of-species
problems, validation attempts often reveal serious dis-
crepancies between the real and simulated environments.
The discrepancies can be traced to three sources: (1) use
of an inappropriate model for the problem; (2) errors in|the
thermodynamic or kinetic data, or other parameter values;
or (3) misinterpretation of laboratory or field data.
In many cases, all three types of errors are present but the
principal contributor is usually not readily apparent. Thus
when undertaking validation studies, the investigator must
ensure that the chemical system under scrutiny is complete-
ly and unambiguously characterized.
Validation studies using chemical-transport codes have typi-
cally been restricted to simple or partial systems in which
only qualitative validation was achieved. For example, the
results of a simulation correlated well with the gross features
of the system's chemical response, but quantitative agree-
ment was lacking or unavailable. The best quantitative
agreement between simulation and laboratory or field ob-
servation has been demonstrated by Valocchi et al. (1981),
Miller and Benson (1983), and Cederberg et al. (1985), all
using the same data from the Palo Alto Baylands project. In
this study, simple reversible ion-exchange processes on
alluviumwere assumed, resulting in agood agreement with
the response of a sedimentary aquifer to injected fluid of a
different composition from the preexisting ground water.
Grove and Wood (1979) studied a slightly more complex
system involving the dissolution of gypsum as well as ion
exchange. The results were compared with data obtained
from a laboratory experiment and a shallow test aquifer at
Lubbock, Texas; field observations agreed reasonably well.
In general, the work developing and verifying computer
codes for geochemical modeling has exceeded validation
efforts. Because validation studies are lacking, the reliability
of thermodynamic data and the understanding of geochemi-
cal processes occurring in the field are questionable. Even
if predicted mineral stabilities correlate with those observed,
the simulation may still be only a very rough approximation.
Validation of more elaborate geochemical codes, such as
those simulating reaction progress, is even less secure.
These codes may help calibrate the evolution of a chemical
system during advective transport in groundwater systems.
Consequently, their use should probably be expanded
provided that problems related to thermodynamic data can
be resolved. These problems are discussed below.
6.5.1.3 Thermodynamic Data Bases
To model a geochemical problem, researchers must often
use experimental data to derive thermodynamic data. Sug-
gested steps are to:
• Derive reference-state properties of a phase (heat of
formation, free energy, entropy, heat capacity, and
specific volume) or the corresponding partial molal
properties for an aqueous species at the reference 25°C
and one atmosphere pressure.
• Use equations of state to predict the properties at the
conditions required for the simulation (see Section
6.5.1.1).
Differences between temperatures and pressures in the
injection zone and those in the reference state lead to small
but significant changes in the thermodynamic properties of
participating species, which for precise work must be cor-
rected.
Thermodynamic data can be retrieved through a variety of
techniques:
• The heat capacity, heat content, and heats of reactions
in which the phase or species participates, and the
specific volume of a phase or species as a function of
temperature and pressure, can be directly measured to
derive the values of entropy and heat of formation and
their functional dependence on pressure and tempera-
ture.
• The thermodynamic properties of a phase or species
can be computed indirectly from phase equilibria,
solubility (miscibility), electrochemical or spectrometric
data or phase relations and groundwater compositions
observed in the field.
72
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Robinson et al. (1982) details these techniques. Often
these techniques are combined, resulting in the need to
reconcile the results from different sources of information
and reach "best value" consensus for a particular ther-
modynamic parameter. Although best values have been
agreed on for a number of minerals and species, they may
not be available for those present in the injection zone; in
this case, the available data must be critically evaluated to
choose a best value.
Because the thermodynamic properties of chemicals are
important to industry and the sciences, many compilations
have been published over the years:
• JANAF (Joint Army, Navy, and Air Forces) (Chase et
al., 1978)
• Thermal Constants of Compounds (Glushko, 1965)
• Robie et al. (1978) (restricted to geochemistry)
• Naumov et al. (1971) (restricted to geochemistry)
These compilations referto original source material, and the
published data have been carefully scrutinized. Even so,
the information in these sources may have to be reconciled
with more recent material or may require an independent
reevaluation in response to specialized needs. In particular,
recent geochemical studies have focused on statistical and
linear programming methods to ensure internal consistency
of thermodynamic data sets.
In the following sections, the availability, reliability, and
suitability of reference-state thermodynamic data are dis-
cussed in relationship to the need to model minerals,
aqueous species, adsorbed species, and other phases.
Availability, Reliability, and Suitability of Thermodynamic
Data. Minerals. In recent years, internally consistent ther-
modynamic data for mineral phases have been compiled
and evaluated, using information derived both from
calorimetric and phase-equilibrium measurements. Some
partial or comprehensive compilations include those by
Helgeson et al. (1978), Perkins et al. (1979), Hemley et al.
(1980), Haas et al. (1981), Robinson etal. (1982), Chatterjee
et al. (1984), Halbach and Chatterjee (1982,1984), Wood
and Holloway (1984), Berman et al., (1985) and Berman
(1988). These compilations serve to highlight the inconsis-
tencies between the independent sets of experimental data
and underline the need to resolve inconsistencies.
The thermodynamic properties of minerals are typically
estimated using cycles (Robinson et al., 1982) or paths for
constructing internally consistent thermodynamic data sets,
which can be linked to reference phases whose properties
have universal acceptance. Helgeson et al. (1978) used the
traditional method in which various data sets are compared
and judged for correctness and/or quality. The data are
refined statistically or subjectively to obtain "best values" for
the thermodynamic parameters, usually starting with a ref-
erence standard such as corundum or quartz, and progress-
ing from the well established mineral thermodynamic
properties to those less well characterized. Using this ap-
proach allows the investigator to start from a secure base
and concentrate on evaluating specific subsets of ex-
perimental data; however, errors may be propagated rather
than distributed.
Inconsistencies in eariierthermodynamic-data compilations
are beginning to be mitigated. New techniques use inter-
nally consistent functions to analyze related thermodynamic
parameters and analyze large numbers of data sets simul-
taneously to find those which are consistent. These techni-
ques include a multiple-regression approach in which func-
tions for molar volume and heat capacity are fitted to the
data and appropriate integration constants to compute the
associated entropy, enthalpy, and Gibbs free energy (Haas,
1985). This technique was described first by Haas and
Fisher (1976) and used by them in several compilations
(Haas et al., 1981; Robinson et al., 1982).
A technique suggested by Gordon (1973) and employed by
Halbach and Chatterjee (1982, 1984), Chatterjee et al.
(1984), Berman et al. (1985), and Berman (1988) uses
linear-programming methods, which compare all data sets
simultaneously. Haas (1985) lists the relative advantages
of the two methods. Regardless of the approach, the
investigator must ensure that the maximum number of
independent checks are available to verify the selected
thermodynamic parameters. Additionally, combinations of
data sets in which one or more inconsistencies are detected
must be tested to determine why they are present. If
inconsistencies are overlooked, these data can become a
major source of error.
Data sets typically used to compute the thermodynamic
properties of minerals are the currently available calorimetric
and phase-equilibrium data. Solubility data, with the excep-
tion of those for quartz and such readily soluble salts as the
carbonates and sulfates, are typically unavailable. This
deficiency is serious because mineral thermodynamic
databases are often used extensively to predict mineral
solubilities in groundwater systems despite not having been
adequately verified through independent tests.
When independent thermodynamic databases such as
those developed by Helgeson et al. (1978) and Berman et
al. (1985) are compared, major inconsistencies often ap-
73
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pear. The researchers either interpreted the same data in
different ways or used different techniques to estimate
thermodynamfc parameters for minerals when no ex-
perimental data were available. .
Often the minerals under consideration have been inade-
quately characterized, usually when researchers interpret
laboratory phase-equilibrium studies that exhibit relatively
short equilibrium times or that have been performed at
relatively tow temperatures. In such situations, fine crystal-
lite size, site disorder, metastable polymorphs, or unknown
contaminants can drastically affect the perceived equi-
librium state without the investigators' being aware. Oc-
casionally, the experimental setup can produce erroneous
results—for example, phase-equilibria data obtained using
a uniaxial press when a vapor phase participated in the
reaction (see Delany and Helgeson [1978] regarding the
work of Kennedy [1959]).
Despite these difficulties, increasingly reliable and more
comprehensive thermodynamic databases for minerals are
becoming available. Thermodynamicdatafor many mineral
phases relevant to host-rock/groundwater interactions,
however, remain either poorly determined or totally lacking.
Data are particularly scarce for solid solutions such as clays,
and satisfactory models to describe the thermodynamics of
solid solutions are formative (see Section 6.5.2.1). Further-
more, researchers modeling the thermodynamics of rock-
water interactions usually overlook the fact that clays are
finely crystalline, and consequently their surf ace free energy
contributes significantly to their thermodynamic stability
(Stumm and Morgan, 1970).
Many other detrital minerals and alteration products
(feldspars, pyroxenes, amphiboles, zeolites, carbonates,
and sulfates) are also solid solutions, and the ther-
modynamic properties of some of the end members remain
in doubt, e.g., anorthite (Chatterjee et al., 1984; Berman,
1988).
Aqueous Species. As with other data bases, ther-
modynamic databases for aqueous species are either un-
critical compilations or contain data that have been selected
as best values through critical review. During a critical
review, incomplete or questionable data may be aug-
mented, rejected, or accepted using correspondence plots
(Criss and Cobble, 1964; Helgeson and Kirkham, 1976;
Helgeson et al., 1981; Tanger and Helgeson, 1988;iand
Shock and Helgeson, 1988, 1989). Data at different
temperatures can be reconciled using such techniques as
those adopted by Cobble et al. (1982) and Phillips and
Silvester (1984) or using more elaborate models such as
the HKF equation of state (Tanger and Helgeson, 1988;
Shock and Helgeson, 1988).
Many techniques are used both to identify the species and
to determine the thermodynamic properties of complexes
present in the aqueous phase. Among these are the titration
techniques developed by Sillen et al. (1960). The user
should be fully aware of the techniques, assumptions, and
models used in the critical review. Critically reviewed data
should be in the standard state and not in an uncorrected
form at the tonic strength of a supporting electrolyte such as
a noncomplexing perchlorate.
Despite their lack of review, uncritical compilations, such as
by Sillen and Martell (1964), may be very comprehensive
and contain a wealth of information. They serve to assess
both the level of knowledge regarding a particular complex
and the level of agreement among different investigators
and are important sources of information on which to base
a critical review.
The thermodynamic properties of organic molecules in the
aqueous phase are of particular interest when evaluating
the injection of hazardous wastes. Calculating such proper-
ties usually requires some knowledge of hydration reac-
tions. From solubility studies, the Gibbs free energy of such
a reaction can be computed, which, when combined with
that of the pure substance in its gaseous, liquid, or solid
state, will yield the Gibbs free energy of the molecular
species in solution. The thermodynamic properties of hydra-
tion have already been compiled (Cabani et al., 1981;
Abraham, 1982,1984). Shock and Helgeson (1989) have
recently compiled an extensive listing of thermodynamic
properties of aqueous organic species for use with the HKF
equation of state and embodied in the SUPCRT code.
Literature search and evaluation will be required for species
not included in these references; additional sources of
solubility or miscibility data for specific organic compounds
might be found in the Chemfate data base compiled by
Syracuse Research Corporation (1986) or in the Arizona
data base compiled by Yalkowsky et al (1987).
Adsorbed Species. The thermodynamics of adsorbed
species are limited primarily to reference temperature and
pressure values (25°C and 1 atm pressure). Compilations
based on the triple-layer model have been summarized
(Kent et al., 1988), but as noted in Section 6.5.2.3, most such
data pertain only to oxide substrates. Properties are usually
given as dissociation constants.
Organic and organometallic species adsorb on clays
through a variety of mechanisms (see MacEwan and Wil-
son, 1984; Laszlo, 1987). Avast bady of literature describ-
ing adsorption on clays and organophilic substrates exists
in uncompiled form. Many studies have been conducted
using specially cleaned and characterized clays whose
adsorption properties bear little relation to those found in
74
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situ. Information is available in the form of various adsorp-
tion isotherms, exchange constants, and Kd or Kp values
(Karickhoff, 1984). This type of information and the proce-
dures for presenting adsorption data in a usable form should
be surveyed. Since such a survey is currently unavailable,
investigators wishing to model adsorption must do their own
literature survey, data evaluation, and derivation of ther-
modynamic data.
Other Phases. Waste streams may contain substantial
concentrations of dissolved salts that do not occur naturally
or that exceed naturally occurring concentrations. Thus,
unusual phases never observed to occur naturally can
possibly precipitate, substantially affecting the evolution of
the waste stream in the injection zone and markedly affect-
ing the behavior of coexisting hazardous constituents.
The thermodynamic properties of these unusual phases are
not as easily characterized as those covered in the preced-
ing three sections. These phases, as well as those which
can develop from other types of interactions, include
kerogens, colloids, micelles, organic liquids, amorphous
gels, adsorbed surface phases, and slimes of ill-defined
composition or structure. Such substances may play a
dominant role in mobilizing or containing hazardous-waste
constituents through coprecipitation or adsorption. Geo-
chemical models generallyfailto accountforsuch materials,
except in specialized applications. The computer code
RELAX (Weres et al., 1986) is an example of a model
designed to determine the partitioning of liquid hydrocar-
bons between aqueous, oil, and gaseous phases for a
specified bulk composition, temperature, and pressure. Ad-
ditional literature search and calculation of thermodynamic
data for such phases may be required.
Calculating Thermodynamic Data. Typically, specialists
calculate thermodynamic data by writing computer codes.
Furthermore, the sources of thermodynamic data for model-
ing purposes is so diverse that computer codes must often
be used opportunistically rather than systematically. One
particular code, SUPCRT, includes the revised HKF equa-
tion of state (Tanger and Helgeson, 1988) for computing the
partial-molal properties and molecular species and is par-
ticularly valuable for extracting thermodynamic properties of
solid phases from phase equilibrium and solubility data
Another, PHAS20, is used by the USGS (Haas, 1974; Haas
and Fisher, 1976).
Using systematic procedures (see Shock and Helgeson,
1988,1989), partial-molal properties of ions or molecular
species from raw data can be calculated for insertion into
the SUPCRT data base. SUPCRT can then be used to
calculate the dissolution reaction constants at any specified
pressure and temperature from 1 to 5.5 kbar and between
0° and 800°C. Such information is necessary for all distribu-
tion-of-species codes using the equilibrium- constant ap-
proach (see Section 6.5.2.4). The author recommends
using SUPCRT as a tool for computing equilibrium con-
stants whenever temperatures in an injection zone differ
significantly from 25°C. In this way, data treatment can be
standardized more readily and dissociation constants and
solubility products extrapolated to higher temperatures with
greater accuracy than by other methods, such as the Criss
and Cobble approach (Criss and Cobble, 1964; Cobble et
al., 1982). If fluid is injected at temperatures greater than
200°C or if wet combustion in the injection zone is being
considered (see Smith and Raptis, 1986),*then use of the
modified HKF model and the SUPCRT code is almost
mandatory.
6.5.2 Specific Methods and Models
6.5.2.1 Reservoir Characterization
Injection-zone host rocks are composed of a wide variety of
heterogeneous components. The most important equation
of state for solids is the Gibbs free-energy equation. The
phase distribution (minerals) in this system at ther-
modynamic equilibrium can be predicted provided the ther-
modynamic properties of the individual phases can be
specified. The free energy of a phase at a certain tempera-
ture and pressure determines whether it will be stable or
unstable. If the free energy is negative, chemical reactions
occur spontaneously; if it is positive, energy must be added
to the system for changes to occur.
The fundamental properties of the solid phase are (1) the
specific volume, (2) the heat capacity as a function of
temperature, and (3) the heat of formation (the energy
required or given off when a compound is formed from its
elements, also called the enthalpy of formation). The
entropy (degree of randomness) of a phase may be corrh
puted from tow-temperature heat-capacity measurements.
If the heat of formation and entropy of a phase are known,
the Gibbs free energy can be calculated. (These properties'
are always calculated at a reference temperature and pres-
Editors s note: The process described by Smith and Raptis (1986) does not involve geochemical interactions with
the reservoir rock or fluids. The injected wastewater returns to the surface for additional treatment or disposal
when combustion is completed. Consequently, no migration occurs in the injection zone
75
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sure so that values for different substances can be com-
pared.)
Equations of state incorporatir^ certain model assumptions
must be applied to calculate tKe thermodynamic properties
of phases at conditions other than the reference tempera-
ture and pressure. Such equations have been developed
by Delany and Helgeson (1978), Helgeson et al. (1978),
Robinson et al. (1982), Berman et al. (1985), and Berman
(1988). Two simplifying assumptions are used to predict
entropy, heat of formation, andfree energy to pressures and
temperatures at or above those necessary for deep-well-in-
jection calculations: (1) the heat capacity of a mineral can
be represented by the Maier-Kelley equation (Maier and
Kelley, 1932), which describes heat capacity as a function
of temperature with the use of three empirical coefficients,
and (2) the molal volume of a mineral is independent of
pressure and temperature (i.e., compressibility and thermal
expansivity are of minor importance). ;
The thermodynamic properties of many pure minerals have
been determined, but natural systems are composed
primarily of minerals that are multicomponent solid solutions
(i.e., different cations substitute for each other in the crystal
lattice structure). Equations of state forthese minerals must
incorporate the effects of solid solutions.
Wood and Frazer (1977) and Helgeson and Aargaard
(1985) reviewed the literature on predicting the behavior of
mineral solid solutions. Currently there are two general
approaches to such modeling. The first considers that the
solid solution is made up of components representing the
end members (Bourcier, 1985). The second uses the site-
mixing model approach (Helgeson and Aargaard, 1985). In
the second approach, individual structural sites in the crystal
lattice of a solid solution are considered to be those in which
tonic substitution or ordering can occur. Helgeson and
Aargaard (1985) derived the general thermodynamic rela-
tions for site mixing. Aargaard and Helgeson (1983) used
site-mixing theory to investigate the assumed ideal mixing
of atoms on sites in montmorilionites, illites, and mixed-layer
clays.
6.52.2 Waste-Reservoir Interactions
The thermodynamfc parameters used to predict reactions
among minerals and water for the range of conditions
expected in the deep-wel! environment can be obtained
using the computer code SUPCRT (Kirkham et al., 1978).
The aqueous phase possesses a significantly varying heat
capacity, thermal expansivity, and compressibility, even
beyond the range of temperatures and pressures expected
in an injection well. This variability must be accommodated
with a carefully calibrated equation of state. One equation
frequently used in geochemical evaluations comes from
Keenan et al. (1969). It is well suited to modeling the
properties of pure water under injection-well conditions and
has been adapted for use with aqueous electrolytes by
Helgeson and Kirkham (1974a) in the SUPCRT code.
If a COa-rich or organic phase partitions in the subsurface
environment, it will be necessary to calculate its stability in
relationship to the aqueous phase and the manner in which
hazardous constituents partition between the two phases.
Unlike solid phases, polar and non|x>lar liquids and dense
fluids possess significantly varying molal volumes, heat
capacities, and derivative properties, i.e., compressibility
and thermal expansion as a function of pressure and
temperature. Therefore, complex equations of state are
sometimes required to describe their thermodynamic
properties. Prausnitz (1969) reviewed equations of state to
describe the behavior of COa-rich phases. For more recent
developments, the literature must be reviewed.
An equation of state for an aqueous solution is necessary
to establish thermodynamic relations between solid-phase
assemblages and the aqueous phase. This might, for
example, involve combining an aqueous-electrolyte model
(see Section 6.5.2.4, Aqueous and Solution Geochemistry)
and a solid-solution model (see Section 6.5.2.1, Reservoir
Characterization).
6.5.2.3 Adsorption
In order of increasing sophistication, adsorption of a given
species can be measured as follows:
• The distribution coefficient (Kd) can be measured either
in a reaction vessel containing a disaggregated or
comminuted (crushed or ground) mass of sorbent in
suspension, or by flowing a waste stream through an
intact core of porous injection-zone rock under simu-
lated downhole conditions and examining the degree of
retardation.
• Adsorption of a species on injection-zone rock can be
measured under simulated subsurface conditions as a
function of concentration. Such experiments establish
whether reversible adsorption occurs, over what range
of concentrations, and whether the adsorption follows
a linear (i.e., Langmuir) or logarithmic (i.e., Freundlich)
isotherm.
• For ion exchange, Scrivner et al. (1986) have success-
fully applied regular solution theory described by Gar-
rels and Christ (1965) with a mixing rule by Hildebrand
et al. (1970) to analyze binary and ternary ion-exchange
data on illites. When exchange data are available for
specific metals or organic compounds of interest, this
76
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approach could be used to predict hazardous waste
uptake on clays (and zeolites if present).
• Experimental data can be collected over a wide range
of parameter values, and triple-layer or related theory
can be used to calculate surface-adsorption constants
for the hazardous species adsorbing on a variety of
natural substrates.
These approaches range from the naive but practical
(measuring the Kd) to the sophisticated but impractical
(triple-layer models). The first approach is the one most
likely to be used to evaluate the adsorptivity of an injection
zone and its ability to retard the transport of hazardous waste
constituents. Much research would be needed to charac-
terize the adsorption process in terms of fundamental ther-
modynamic parameters and would not necessarily be suc-
cessful in predicting the behavior of a hazardous waste
constituent during transport.
Use and Abuse of the Distribution Coefficient Kd. Ad-
sorption of metal tons or organic compounds on soil is well
known, and soil scientists have developed an empirical
method for quantifying the process using the distribution
coefficient, Kd. The term is defined as the distribution of a
given species between solution and solid, expressed as a
ratio of mass of adsorbent per unit mass of solid.
The inherent weaknesses in using a Kd value to characterize
adsorption of a species have long been recognized (Apps
et al., 1977). One problem is that the method used to
measure it (see Section 6.6.2) is severely limited to the
specific conditions under which the test was performed. Any
deviation from experimental conditions is very likely to cause
changes in the Kd value, sometimes by orders of magnitude.
Table 6-3, adapted from Apps et al. (1977), lists some
variables that would affect Kd.
A more specific problem with applying Kd measures to
underground injection zones is that it is difficult to compare
a dispersed solid such as a soil with a coherent rock. Clearly,
crushing and grinding the rock will lead to a much larger
specific surface area than that present in the injection zone.
Furthermore, the newly exposed surfaces will not be repre-
sentative of injection-zone material, and may adsorb dif-
ferently from those normally irt contact with the ground water.
A final problem is that the adsorption experiment may be
completed within a few days, whereas the hazardous waste
stream may be in contact with the host rock for tens of
thousands of years, allowing slower reactions to proceed.
These slower reactions may eventually completely
dominate any short-term adsorption reactions reflected in
the Kd term.
In spite of these problems, Kd measurements are used
frequently by hydrotogfc modelers because of their
simplicity. The Kd values can be used to calculate a retar-
dation factor, which in turn can be used in a one-dimen-
sional differential equation to describe solute transport
(Javandal et al., 1984). By assuming that the conditions
postulated in the model are identical to those in the Kd
measurement experiment and that the adsorption reaction
is locally reversible, the modeler may be able to estimate
the degree of retardation expected for a given adsorbent.
The adsorption capacity and potential retardation in the
injection zone can be estimated by measuring adsorption of
a hazardous constituent in a core from the injection zone
under simulated downhole conditions using core-
ftowthrough experiments. The results will probably be con-
servative because the test will be of short duration; more
long-term processes involving diffusion into micropores or
solid-state ton exchange will not be observed. Also, the
injected formation will probably be relatively permeable and
therefore will contain minor concentrations of such adsor-
bents as smectites, which are more likely to be found in the
adjacent confining zone. If such an experiment is properly
designed, a Kd value that can be incorporated into transport
models can be calculated.
Langmuir and Freundlich Isotherms. The use of Kd
measurements assumes that no complicating factors, such
as supersaturatfon with respect to a solid phase or the
complexation of variable degrees of adsorbent in solution,
will affect adsorption. If side reactions or extensive mixing
and dilution of the waste stream occur in the injection zone,
experiments should be conducted over a range of adsorbent
concentrations to determine an adsorption isotherm. As
with Kd, such measurements may be taken using cores or
disaggregated material. The difference is that separate
measurements are made at different concentrations (see
Section 6.6.2).
The Langmuir and Freundlich equations are both well-ac-
cepted representations of adsorption behavior. The
Langmuir adsorption isotherm is measured by assuming
that the surface of the adsorbing substrate is covered by a
monolayerof solvent and adsorbent molecules in competi-
tion with each other. This formulation predicts a linear
change in adsorption in response to increased concentra-
tions and was applied originally to gas adsorption on a metal
surface (Langmuir, 1918). The linear isotherm response is
observed widely in the adsorption of trace concentrations of
compounds on soils and other heterogeneous substrates
(Chtou et al., 1983; Schellenberg et al., 1984). Kincaid and
Morrey (1984) show that the Langmuir adsorption equation
is valid only at extremely tow concentrations of adsorbing
solute and at fixed pH. Karickhoff (1984) has developed a
77
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Table 6-3 Factors Influencing Kd for an Adsorbent, Based on Conditions Expected in a Subsurface
Injection Zone
Parameters
Principal
Effect
Injection-
Zone
Conditions
Potential
Effect on Ka
1a. Major Components
b. Minor Components
c.Ph
d.Eh
2. Hazardous constituent
concentration and
speciatlon
3. Flow rate
4. Permeability
5. Surface Area
6. Temperature
Solution Chemistry
Ionic strength
Activity coefficients
Complexing
Complexing
Complexing
Chemical potential
Chemical potential
Supersaturation
Polymerization
Metastable equilibrium
Metastable equilbrium
transport mechanisms
Changes in apparent
surface area contacted
Flow rate (see above)
Adsorption
Complexing
Solubility
Adsorption
Determined by host-
rock chemistry and
other factors including
leaching chemistry of
waste product..
As above
5.0-8.5
Buffering of hetero-
and homogeneous
equilibria keep pH
within narrow limits
Variable, over a narrower
range, usually reducing
Very variable concentration
Could range from 10,000
ppm to 0 near the isoelectric
point (-106)
cm/sec
~10"2cm2/g
fractures, microfractures
irrtergranular pores
100to150"C
Unpredictable—
probably 10"3 to
As above
Up to 1010 or
even more
Difficult to estimate
but could be very large
for Inorganic amphoteric
species
Slow flow rates could
lead to different rate-
controlling transport
mechanisms, (e.g., tonic
or molecular diffusion) and
to different thermodynamic
controls (0 to 106)
Same as above
~103
Upto10d
78
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simplified approximate linear-adsorption-isotherm equation
that is essentially a form of Henry's Law constant. It appears
to be valid for natural sediments, as long as the concentra-
tion of hydrophobic adsorbent is less than 10"5 M, or less
than one-half its solubility in water.
The Langmuir adsorption equation inherently assumes that
the surface activity coefficients of the adsorbing species are
unity, i.e., adsorption behavior is ideal. Experience has
shown that the equilibrium constant in the Langmuir equa-
tion varies with the fraction of surface occupied, suggesting
that this assumption is not valid. Adamson (1982) points out
that if this variation is attributed entirely to a variation in the
heat of adsorption, then observed adsorbence changes
logarithmically.
This logarithmic adsorption isotherm is named after
Freundlich (1926), who first used it extensively in many
applications. The Freundlich equation has two empirical
constants: the intercept and the slope. Because the con-
stant for the slope represents a number of parameters that
cannot easily be derived independently, the Freundlich
isotherm equation remains a purely empirical formulation,
albeit with some theoretical basis. If the adsorption process
does not show reversibility, then the implicit assumption that
the Freundlich isotherm is an approximation of a ther-
modynamic process is incorrect and has no theoretical
basis. Van Genuchten et al. (1974) suggest that it would be
preferable to use kinetic rate equations to describe adsorp-
tion (e.g., Lapidus and Amundson, 1952; Lindstrom et al.,
1971; Oddson et al., 1970).
Each Freundlich isotherm must be measured keeping all but
the adsorbent concentration parameter (pH, Eh, tempera-
ture, and solution composition) constant. Kincaid and Mor-
rey (1984) found that the Freundlich isotherm operates over
a far wider range of solute concentrations at a given pH than
the Langmuir isotherm, and thus the former has been used
successfully to measure adsorption of a variety of metal tons
on clays as part of the Civilian Radioactive Waste Manage-
ment Project of the U.S. Department of Energy (1987) and
in the adsorption of organic species on clays and soils
(Means et al., 1982; Sheindorf et al., 1982).
Triple-Layer Adsorption Models. More complex adsorp-
tion models are required to model host-rock or groundwater
interaction with the injected waste stream when substantial
changes in pH, EH, tonic strength, or tonic concentrations
are expected. Unfortunately, complex models require that
more parameters be measured and the model conditions
defined more rigorously.
One of the most sophisticated model developments for
describing adsorption phenomena in aqueous solutions is
known as the triple-layer model (TLM). This model
describes the surface of a mineral in terms of a site-binding
model, developed by Yates et al. (1974), combined with a
surface-binding model, by Davis and Leckie (1978). It has
been refined, developed, and tested overa number of years
by faculty and research staff at Stanford University (Davis
and Leckie, 1978, 1980; Kent et al., 1988) and is also
referred to as the Stanford General Model for Adsorption
(SGMA).
TLM separates the aqueous interface with the adsorbent
surface into three layers: surface, inner diffuse, and outer
diffuse. Each layer has an associated electrical potential,
charge density, capacitance, and dielectric constant.
Protons (hydrogen tons) are assumed to bind at the surface
plane whereas electrolyte tons bind at the inner, diffuse
plane. The surface is assumed to be coated with hydroxyl
groups (OH), with each surface site associated with a single
hydroxyl group. The surface sites may either react with
other tons in solution(s) or dissociate, according to a series
of reactions, each reaction being identified with an as-
sociated equilibrium constant (Kent et al., 1988). Ex-
perimental terms relate the concentrations of the tons at their
respective surface planes to those in the bulk solution. The
sum of charges in the three layers is assumed to be zero
(i.e., the triple layer is electronically neutral).
Of the four groups of minerals discussed in Section 6.3.2
(simple oxides, multiple-surface-site minerals, fixed-charge
minerals, and salt-type minerals), simple oxides have been
the only group successfully modeled using TLM. Oxide
substrates have been studied extensively and their adsorp-
tion behavior modeled over a range of environmental con-
ditions. Unfortunately, simple oxide surfaces (represented
by one adsorption site) are not the only surfaces a waste
stream can encounter in an injection zone, and adsorption
capacity due to the presence of simple oxides may repre-
sent only a small fraction of the total capacity of the zone.
TLMs representing two or more sites have not yet been
developed because data sets adequate to characterize the
adsorption properties of aluminosilicates do not exist (Kent
et al., 1988). To characterize the surface chemical proper-
ties of fixed-charge minerals, both ton-exchange and ad-
sorption reactions must be measured. Although significant
achievements have been made inthis area, a data base that
would allow clay-mineral adsorption to be predicted over a
wide range of clay compositions and environmental condi-
tions does not exist. Similar problems apply to zeolite and
manganese oxide; further, the latter group is sensitive to the
oxidation state of the system.
The surface charge and potential of salt-type minerals are
controlled by the relative abundance of the constituent tons
79
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in the associated solution and only indirectly by the pH.
Adsorption on salts has not been investigated with any
thoroughness, and potential adsorption mechanisms are
speculative. Anionto adsorption, as well as cationfc or
anfonlc exchange, appearsto be possible (Kent et al., 1988).
Carbonates are often important constituents of saline-
aqu'rfer host rocks. Therefore, catfonfc adsorption on such
minerals could be important in removing heavy metals from
the injection zone.
The TLM formulation assumes that adsorbent properties
are experimentally characterized as a suspension or slurry
in the aqueous phase. However, with varying suspension
concentrations, particle interference can take place, which
modifies the apparent surface properties of the adsorbent.
Still greater effects may be observed in the properties of two
adsorbents when mixed. Ideally, the adsorption properties
of an aggregate of different adsorbents would be additive,
obeying what is called the linear adsorptivity model (LAM).
However, experimental studies reported by Honeyrhan
(1984) and unpublished studies conducted by Myer (per-
sonal communication) of ORNL reveal that major nonlinear
adsorption effects are sometimes observed with mixtures
and that LAM may represent the exception rather than the
rule.
Any sophisticated adsorption model requires a correspond-
ing electrolyte model that accurately replicates the ther-
modynamic properties of the solute species. Conventional
models using a single parameter extension of the Debye-
Huckel equation for individual ion-activity coefficients, as in
MINEQL (Westall et al., 1976), are limited to an ionic
strength of 0.1 molal for accurate modeling (about 5,800
mg/Lfor NaCI solutions), or 0.5 molal (about 29,900 mg/L)
for less precise interpretations. Any treatment of brines with
higher tonic strength would require an entirely different
formulation of the adsorption model, using, for example, the
Pitzer interaction model (Pitzer, 1973; Pitzerand Kim, 1974).
Any further devetopmerrts of TLM must be designed such
that the solubility product with respect to any solid phase
involving the participating components in solution is not
exceeded. Researchers can use an appropriate distribu-
tfon-of-species computer code such as EQ3 (Wolery, 1983)
to ensure that the solubility product is not exceeded. How-
ever, the effectiveness of the code in identifying super-
saturating phases depends on the availability of ther-
modynamic data for them.
For all its sophistication, TLM is of limited value in
demonstrating the containment of hazardous-waste con-
stituents in an injection zone for the following reasons:
• Site-binding constants have been determined for only
a limited range of simple oxides; with only one type of
surface site.
• Extensive fundamental studies of common clays and
other minerals found in the typical injection-zone host
rock would be required. The successful outcome of
such studies is uncertain.
• No satisfactory model has been developed that would
permit predictions of adsorption properties of mixtures
of adsorbents based on the properties of individual
adsorbents.
• No satisfactory means of measuring and interpreting
the adsorptive properties of intact host rock in relation
to TLM has been developed.
• The present TLM is restricted in its application to
coexisting electrolytes with tonic strengths much less
than unity.
For those situations where none of these problems applies,
the SGMAmodel has been incorporated into several codes,
including MINEQL (Westall et al., 1976) and the MINTED
code, which combines MINEQL with WATEQ (Felmy et al.,
1984; Krupka and Morrey, 1985). Otherwise, less-sophisti-
cated models using the Langmuir or Freundlich isotherms,
or even the distribution coefficient Kd, would be required.
6.5.2.4 Aqueous and Solution Geochemistry
Two types of codes are used to model aqueous and solution
geochemistry: distribution-of-speeies codes, which rep-
resent the thermodynamics of a static system, and reac-
tion-progress codes, which examine the consequences of
an evolving system in which various phases in a system
reactwithone another. Each code uses equations of state,
which establish the thermodynamic relationships between
the solid- and aqueous-phase assemblages. Equations of
state are discussed below; discussions of the two types of
code follow.
Equations of State. There are a number of approaches for
developing equations of state. These approaches can be
divided into two basic types: an approach that calculates
the partial molal properties of the standard state, and an
approach that uses an aqueous elertrolyte model. The HKF
equation represents the first approach and the Debye-
Huckel and its B-dot extension, the Davies equation, and
the P'rtzer model represent the second.
HKF Equation of State. The HKF equation of state (named
after its principal authors, Helgeson, Kirkham, and Flowers,
1981) predicts the standard-state, partial-molal, ther-
80
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modynamic properties of tonic and molecular species in
aqueous solution over a range of temperatures and pres-
sures suitable for modeling deep-well conditions. The equa-
tion has been under development for 15 years (Helgeson
and Kirkham, 1974a,b, 1976; Helgeson et al., 1981), and
Tanger and Helgeson (1988) have eliminated some serious
deficiencies in earlier versions.
The HKF equation estimates entropy and Gibbs free energy
as well as the standard-state properties over a range of
pressures and temperatures, provided that (1) the partial-
molal volume, the partial-molal heat capacity, and the par-
tial-molal heat of formation of an ionic species at the refer-
ence temperature and pressure (298.15°K and 1 bar) are
known and (2) a series of model parameters are evaluated
for that particular species. Shock and Helgeson (1988)
provide guidelines and summarize the values and the model
parameters for a large number of common inorganic tonic
species at reference temperatures and pressures. Shock
and Helgeson (1989) also provide similar data for many
water-miscible, tow-carbon-number organic species. A
modified HKF equation of state is incorporated into a revised
version of the SUPCRTcode, which is currently undergoing
prerelease testing and evaluation. (Documentation for an
earlier version is given in Kirkham et al. [1978]).
The HKF equation is highly rated because of its general
precision and its versatility for modeling injection-zone con-
ditions. However, simpler equations of state may be used
instead. One simplification assumes that the heat capacity
of a reaction is constant over the range of pressure and
temperature under consideration. This simplification was
popularized by Criss and Cobble (1964). Ingeneral, assum-
ing constant heat capacity of reaction is a reasonable
approximationforextrapolatingfrom25°Cto200°C provided
the total pressure of the system remains less than 300 bars
and precise predictions are not required (Cobble et al.,
1982).
Aqueous-Electrolyte Models. Aqueous-electrolyte models
can be used to relate the measured concentrations of
dissolved constituents and waste and reservoir fluids to the
thermodynamic properties of the dissolved species using
distribution-of-species codes (discussed later). General
reviews of the theory and development of these models can
be found in textbooks by Lewis and Randall (1961) and
Robinson and Stokes (1959) and a paper by Pitzer (1977).
Aqueous-electrolyte models provide important means for
relating the measured concentrations of dissolved con-
stituents and waste and reservoir fluids to the ther-
modynamic properties of the dissolved species.
Aqueous-electrolyte models use the concept of the activity
coefficient, which determines the concentration of a sub-
stance in solution. This coefficient varies depending on the
nature and total concentration of dissolved constituents,
temperature, and whetherthe species is charged or neutral.
The Debye-Huckel equation (Debye and Huckel, 1923;
Debye 1923, 1924) can be used to predict the activity
coefficient of species in dilute solutions less than 0.1 molal
but is too imprecise for solutions with concentrations of
dissolved constituents in the molal range, as is typical of
most waste streams and saline ground water.
Several methods have been used to extend the range of
applicability of the Debye-Huckel equation. Most important
are (1) the Debye-Huckel B-dot extension, (2) the Davies
equation, and (3) the Pitzer interaction model. Several other
extensions of the Debye-Huckel equation have also been
formulated to compute activity coefficients in high-tonic-
strength, mixed electrolytes. Zemaitis et al. (1986) exten-
sively review and compare various equations for computing
the individual ton activity coefficients or mean activity coef-
ficients of salts in single and multicomponent strong
electrolytes.
The B-dot extension adjusts the Debye-Huckel equation
(Guggenheim, 1935; Guggenheim and Turgeon, 1955)
using empirical data, and its value has been demonstrated
by Pitzer and Brewer in Lewis and Randall (1961). The
B-dot parameter is adjustable and normally set so that the
activity coefficient best reflects that measured in NaCI solu-
tions. Because NaCI is usually the dominant salt in deep
ground waters, this setting is appropriate for many situa-
tions. Hazardous waste streams with significant concentra-
tions of dissolved salts other than NaCI, however, would
require an empirical adjustment of B-dot for each composi-
tion, which is inconvenient. Furthermore, the B-dot exten-
sion becomes unreliable at tonic strengths greater than
about 0.5 molal, considerably less than the tonic strength of
many deep saline aquifers and waste streams. The B-dot
extension is used in several distributfon-of-species com-
puter codes, e.g., EQ3/6 (Wolery, 1983), PATH (Helgeson
et al., 1969), FASTPATH (Schlitt and Jackson (1981), and
PHREEQE (Parkhurst etal., 1980).
The Davies equation (Davies, 1962) is another empirical
extension of the Debye-Huckel equation that gives fairly
good estimates to about 0.5 molal. It suffers, however, from
the same disadvantages as the B-dot parameter extension.
It has been incorporated in the GEOCHEM (Sposito and
Mattigod, 1980), MINTEQ (Felmy et al., 1984), MINEQL
(Westall et al., 1976), and PHREEQE (Parkhurst et al.,
1980) distributfon-of-species codes and will be an option for
the latest version of EQ3/6.
The Pitzer interaction model (Pitzer, 1977) uses a more
complex equation that contains several empirical coeffi-
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cients for application to a variety of solutions. It also contains
three parameters that reflect electrostatic interactions be-
tween ions of like and opposing charges specific to each
ionic interaction and unchanging with variations in the bulk
composition of the electrolyte. Pitzer and Mayorga (1973)
have obtained interaction parameters by fitting osmotic
coefficients for about 200 electrolytes. To determine inter-
actions among three ormore unlike ions, osmotto-coeffiqent
data must be fitted to data on mixed electrolytes with the
binary-interaction parameters fixed, resulting in somewhat
more complicated expressions for activity coefficients (Pit-
zer and Kim, 1974).
The Pitzer interaction equations can model multicomponent
electrolytes to very high tonic strengths accurately—often to
saturation limits. Several studies have validated the Pitzer
interaction model by simulating brine-salt equilibria in natural
evaporite systems (Eugster et al., 1980; Harvie and Weare,
1980; Harvie et al., 1982,1984; Felmy and Weare, 1986).
The Pitzer equation has several disadvantages:
• Data are lacking on the interaction parameters and their
temperature dependence for many important
electrolytes and minor components.
• It is premised on electrostatic interactions between ionic
species, thereby eliminating the need to account ex-
plicitly for ton-pair formation. Consequently, it cannot
account for strong, covalently bonded complexatfon. If
covalent bonding of aqueous complexes occurs, some
hybrid model must be formulated.
Despite these limitations, the Pitzer interaction model is very
promising.
DIstributlon-of-SpecIes Codes. All equations of state
require information on how the concentrations of varipus
elements or constituents measured analytically in solution
are distributed among the tonic or molecular species in
solution. This information is developed using distribution-of-
species codes. These codes also determine which species
represent the most stable configuration for the system, i.e.,
the minimum Gibbs free energy of the system at a given
temperature and pressure. Two approaches have evolved:
the equilibrium-constant approach and the Gibbs free ener-
gy minimization method. As pointed out by Zeleznik and
Gordon (1968), the methods have a common origin and their
content is the same. Thus, the choice of method becomes
one of convenience and traditional use.
Most distribution-of-species codes use the equilibrium-con-
stant approach, primarily because equilibrium constants
relating to aqueous species can be determined directly and
are better-known than the underlying thermodynamic
properties of the participating species (Nordstrom et al.,
1979). Almost all codes employ the equilibrium-constant
approach, including:
• EQUILIB (Shannon et al., 1977; Morrey, 1981; Money
and Shannon, 1981)
• EQ3/EQ6 (Wolery, 1983; Wolery et al., 1985)
• GEOCHEM and REDEQL (Morel and Morgan, 1972;
McDuff and Morel, 1978; Sposito and Mattigod, 1980)
• MINEQL (Westall et al., 1976; James and Parks, 1976)
• MINTEQ (Felmy et al., 1984; Krupka and Morrey, 1985)
• PHREEQE (Parkhurst et al., 1980; Plummer and
Parkhurst, 1985)
• WATEQ and related codes (Truesdell and Jones, 1973,
1974; Plummer et al., 1976; Ball et al., 1981; Krupka
and Jenne,1982)
When complex multicomponent systems are evaluated,
however, several serious disadvantages are apparent:
• The measured equilibrium constants are rarely
measured at a standard temperature and pressure and
must always be corrected to standard-state conditions
for subsequent modeling purposes.
• The temperature and pressure at which a system is to
be modeled usually differ from the temperature and
pressure at which the equilibrium constant was
measured. The equilibrium constant must be recalcu-
lated either using some equation of state or interpolating
between corrected experimental values.
• The chemical reactions needed to define the equi-
librium constants are usually arbitrary and their choice
often has no effect on the final results. Therefore, code
writers have developed little uniformity in their use.
Further, the reactions connote a mechanistic interpreta-
tion that often has no basis in reality.
• Most equilibrium constants, especially those for solid
phases, are not obtained by direct measurement.
Thus, they must be calculated from Gibbs free-energy-
of-specification data after the reaction equation has
been specified.
82
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• Because equilibrium constants are usually computed
from the Gibbs free-energy data, the equilibrium con-
stants of all reactions in which a species participates
must be recalculated whenever the thermodynamic
properties of that species change.
• Equilibrium constants must be calculated for a range of
pressures and temperatures and stored in a database.
The pressure and temperature ranges covered may not
be appropriate for the user.
The Gibbs free-energy-minimization approach could be
more convenient than the equilibrium-constant method.
Once thermodynamic data for a given mineral or aqueous
species were stored in a data base, the Gibbs free-energy
minimization method could be used to calculate the stand-
ard-state Gibbs free energies of the participating species at
a given pressure and temperature for entry into a distribu-
tion-of-species code. The distribution-of-species code
could thereby be made more flexible, and refinements in the
thermodynamic properties of participating species could be
easily incorporated. Unfortunately, no such method exists,
although it appears that one such as this eventually will
become available.
The equilibrium-constant method solves a set of nonlinear
mass-action equations and linear mass-balance equations
and can calculate the charge balance constraint. The mass-
action equations for aqueous species and solids are usually
written as dissociation or dissolution equations, each break-
ing down into its "basis" species. These species are usually
the simple charged ionic species of an element, such as
Na+, Mg , Al , cr, or commonly occurring, simple mol-
ecular or ionic species such as SiOa(aq), SO*-, COa-, or
NO3-.
The Debye-Huckel equation, or some empirical or semi-em-
pirical extension, is always used to calculate the activity
coefficient of charged species. For neutral species, the
activity coefficients are often determined by empirically cor-
relating the activity of carbon dioxide in a solution of cor-
responding ionic strength. To calculate the activity of water,
various equations are used (see, for example, Wolery, 1983,
p. 22). Ideally, the Gibbs-Duhem relationship between the
solvent and solute species is used, butthis is not doneoften.
For solutions with many components and many species and
complexes in solution, the number of linear and nonlinear
equations can often exceed 100. The mathematical tech-
nique used to solve the set of linear mass-balance and
nonlinear mass-action equations incorporates algorithms
for calculating the activity coefficient, converts the nonlinear
equation to an infinite sequence of linear equations, and
then solves the equations iteratively.
Zeleznik and Gordon (1968) and Van Zeggeren and Story
(1970) reviewed the various methods for solving these
equations. The Newton-Raphson method is most often
used in such schemes. Various techniques are used to
ensure rapid convergence. Although not essential, the
system should be electrically neutral. Chemical analyses
always contain some analytical error, usually (ess than 5%
by weight of the total charge. This error can be eliminated
by "balancing" on a predominant species in solution, such
as Cr or Na+. A species should be chosen such that the
addition or subtraction of a small quantity of the balancing
element does not significantly affect the thermodynamic
properties of the system.
Not all iterative methods will converge for all types of
problems. Some codes may hang-up or crash, and
problems without solution can be specified. It is helpful if
the code writer has incorporated diagnostic statements to
identify potential problems. Wolery (1983) has made par-
ticular efforts in this direction with the EQ3/6 code.
Most codes require data from a typical chemical analysis
(i.e., the concentrations of various elements or species).
Other inputs such as alkalinity, pH, Eh, temperature, and
solution density may also be required, as well as specifica-
tion of the aqueous species needed to balance the charge
and ensure electrical neutrality. Some codes allow other
features: for example, a solution with a particular mineral
can be modeled. The partial pressure of a gas phase can
be controlled, and the oxidation state can be specified in
terms of various redox pairs (e.g., H2S(aq)/SO4-,
CH4(aq)/CO2(aq), NH4+/NO3-, etc.). The actual codes that
allow temperature of the sample to be specified, rather than
an approximate temperature, are particularly valuable.
The output of a distribution-of-species computer code can
include the distribution of the species (usually listed in order
of abundance), specifying the concentrations, activities, and
logarithms of activities. The code can also show the relative
abundance of species containing a particular element, cal-
culate saturation indices with respect to minerals and gases,
or even conduct a mass-transfer calculation to show what
and how much material must precipitate to prevent super-
saturation in solutbn. Three codes-^he latest version of
EQ3/6 (Jackson and Wolery, 1985), TRANSCHEM (Scriv-
ner et al., 1986), and PHREEQE (Crowe and Longstaffe,
1987)—-incorporate the Pitzer interaction-parameter elec-
trolyte model, which is required when highly saline waste
streams or ground waters are modeled.
The choice of distribution-of-species code depends on its
use. The following comparative reviews may be useful in
selecting an appropriate code: Nordstrom et al. (1979),
Jenne (1981), Kincaid and Money (1984), Carnahan
83
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(1987b), and Mangold and Tsang (1987). Intera (1983)
assessed computer codes specifically for radioactive-waste
containment.
Reaction Progress and Kinetics. Distribution-of-species
codes represent a static system, reaction-progress codes a
dynamic system. The original concept of reaction progress
was developed by deDonder (1920) and deDonder and Van
Rysselbeurghe (1936) and subsequently adopted by Hel-
geson et at. (1969). The essential modeling premise made
by Helgeson et al. (1969) is that 1 kg of water and the
dissolved species are allowed to react with a defined quan-
tity and numberof solid orgaseous phases. Progress of the
chemical reaction is monitored as a function of the amount
of material that reacts and usually is expressed in terms of
gram-moles of the normalized stoichiometry of the reactant
assemblage. In its simplest form, the model assumes the
following:
• The aqueous phase is always in internal equilibrium.
• Product phases are in reversible equilibrium with the
aqueous phase.
• The system is isothermal.
• The reactant phases always dissolve in proportion to
their initial stofchfometric ratios.
These assumptions are reasonable approximations of
natural hydrothermal environments where time spans are
relatively long and higher temperatures promote rapid equi-
libration with product or secondary minerals. They break
down, however, when applied to problems involving rela-
tively slow reactions such as those which occur during
deep-well disposal of hazardous wastes at relatively low
temperatures. Such systems are rarely in equilibrium. For
example, natural tow-temperature systems, such as brines
In Gulf Coast sediments, may not be in internal equilibrium
even though the brines may be millions of years old (Apps
et al., 1988; Shock, 1988). Lindberg and Runnels (1984)
have questioned whether redox pairs can predict system
EH, since coexisting aqueous redox species are often not
in equilibrium and yield different Eh values. These re-
searchers estimate that homogeneous equilibrium with
respect to Eh in ground waters may be achieved only after
thousands of years. Other investigators (Palciauskas and
Domentco, 1976; James and Rubin, 1979) have also ques-
tioned whether adsorption or ton-exchange reactions reach
tocal equilibrium. Attempts to model chemical transport
systems that assume tocal equilibrium with respect to ad-
sorption have demonstrated that kinetic factors must play a
role (Van Genuchtenetal., 1974;Vallette-Silveretal., 1981).
Helgeson et al. (1969) also applied the reaction-progress
variable in monitoring geochemical reactions so as to avoid
the explicit use of chemical kinetics, because so little is
known about the dissolution and precipitation kinetics of
minerals in aqueous solution (see discussion of precipita-
tion-dissolution, Section 6.2.2). No unifying method that
quantitatively predicts mineral-dissolution kinetics is cur-
rently available.
Few computer codes have been wrilten to simulate reaction
progress. The best-known and documented are EQ3/6
(Wolery, 1983) and PHREEQE (Parkhurst et al., 1980).
Others cited in the literature include SOLVEQ (Reed, 1982)
and the now obsolete PATH1 (Helgeson et al., 1969).
Several codes such as REDEQL (Morel and Morgan, 1972;
McDuff and Morel, 1973) and MINEQL(Westall et al., 1976)
contain options for reaction-progress simulations, but these
codes are very limited in scope compared with EQ3/6 and
PHREEQE.
Two greatly improved codes based on PATH1 principles
were written during the 1970s—FASTPATH (Schlitt and
Jackson, 1981) and one developed by C. Herick at Los
Alamos National Laboratory—but they are now obsolete
and not available to the general user. EQ3/6 is currently the
most versatile code, and it has quality-assurance documen-
tation (Wolery, 1986), but a revised version of PHREE:QE is
to be released soon.
The assumptions inherent in computer codes such as
PATH1 (Helgeson et al., 1969), EQ3/6 (Wolery, 1983), and
PHREEQE (Parkhurst et al., 1980) are generally not valid
for predicting the fate of hazardous; wastes in the injection
zones of deep wells when reaction rates are slow in relation-
ship to groundwater movement (i.e., if tocal homogeneous
or heterogeneous reversibility within about a meter of the
point of injection is not attained). If the assumptions implicit
in the model are good approximations (i.e., acid or alkaline
neutralization or system reduction), reaction-progress
codes are convenient and permit powerful simulations of
reaction chemistry.
Reaction-progress codes can incorporate the net enthalpy
"generated" by the simulated chemical reactions. For
simple adiabatic systems, reaction-progress codes could
monitor the resulting temperature excursion as a function of
reaction progress. This feature may be useful where par-
ticipating reactions are strongly endothermic or exothermic,
as in the case of wet combustion.
Despite their lack of applicability for systems in which reac-
tions are slow compared with grouncJwater movement, reac-
tion-progress simulations have sometimes provided in-
sights into the processes occurring in natural hydrochemtoal
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systems. Thus Plummeret al. (1983) reconciled the evolu-
tion of ground water with the mineralogy of the limestone
and dolomite aquifer between Polk City and Wachula,
Florida, using the reaction-transport code PHREEQE. Apps
et al. (1988) simulated the evolution of Gulf Coast brines
using EQ3/6, which was consistent with many but not all
field observations. Solomon (1986) has demonstrated that
EQ3/6 simulations of the evolution of ground waters in
basalts correspond well to observations in the Grande
Ronde formation in Washington. Kerrisk (1983) has simu-
lated the evolution of groundwater chemistry and mineral
formation at Rainier Mesa, Nevada.
Reaction-progress codes have also been used to interpret
laboratory studies (Knauss et al., 1984). Apart from these
examples, few other reaction-progress simulations have
been rigorously correlated with field or laboratory observa-
tions, and to the author's knowledge none has been applied
to the analysis of a hazardous-waste plume.
Reaction-progress codes for geochemical simulations are
readily modified to incorporate chemical kinetics. Aargaard
and Helgeson (1982) use the term reaction velocity (the
change in material reacted over a specified period of time)
to relate reaction progress with the kinetics of the process.
The EQBcode can simulate reaction kinetics (Delany et al.,
1986), and can be used to simulate the evolution of systems
where several heterogeneous reactions compete (Hel-
geson amd Murphy, 1983).
6.5.2.5 Transport
Currently, two approaches are available for modeling chemi-
cal transport. In the first, all mass, momentum, and energy-
transfer equations, including those in which chemical reac-
tions participate, are solved simultaneously for each time
step in the evolution of the system (one-step approach). In
the second, two linked but separate subroutines are used
(two-step approach).
The one-step approach solves for mass, momentum, and
energy balances for the fluid at each time step. Then the
chemistry is reequilibrated through a distribution-of-species
code. This type of approach has been applied for a number
of years and includes studies by Lai and Jurinak (1971),
Rubin and James (1973), Valocchi et al. (1981), Jennings
et al. (1982), Miller and Benson (1983), Noorishad and
Carnahan (1985), Carnahan (1986), Noorishad et al.
(1987), Willis and Rubin (1987), Merino et al. (1986) and
Carnahan (1987a). In most early studies, relatively primitive
isothermal systems were investigated. These systems
usually involved simple ion-exchange formulations, some-
times including complexation; system pH was assumed
constant. Under such conditions, even the ion-activity coef-
ficients changed little and could be ignored without major
error.
Many reactions controlling groundwater composition are
slow in relation to groundwater movement, and chemical
kinetics must be introduced into the transport models.
Therefore, the conditions of local reversibility, or instan-
taneous local equilibrium as assumed in early simulations
involving ton exchange or adsorption (e.g., Rubin and
James, 1973; Valocchi et al., 1981; Jennings et al., 1982),
cannot be used in many realistic simulations. A recent
development is the application of CHEMTRANS (Noorishad
et al., 1987). This code can simulate in one dimension both
homogeneous aqueous-phase and heterogeneous tempera-
ture-dependent reaction kinetics. It has been applied to a
variety of simple problems involving both reversible and
irreversible dissolution, oxidation/reduction, and carbon-
isotope fractionation in ground water. If the kinetic and
thermodynamic parameters are known for a particular prob-
lem, it is relatively easy to modify the code to simulate a
problem in which heterogeneous and homogeneous reac-
tions and equilibria play a part. This code is, however, difficult
to use and computationally intensive, and has not been
tested with complex multicomponent systems.
In addition to the approaches discussed above, the
Freundlich isotherm can be substituted into a differential
equation to describe solute transport. Often desorption
experiments do not indicate that the Freundlich adsorption
process is reversible, and it therefore is necessary to incor-
porate separate expressions for adsorption and desorption
in transport equations (e.g., see Van Genuchten et al.,
1974). In general, substituting chemical reactions into
transport equations results in equations that are very non-
linear and difficult to solve numerically, and entail extended
execution times.
The second approach appears to have been used first by
Grove and Wood (1979) and subsequently adopted by the
following researchers:
• Reardon(1981)
• Walsh et al. (1982) used a distribution-of-species code
developed by Morel and Morgan (1972)
• Cederberg et al. (1985) incorporated the distribution-of-
species code MICROQL(Westall, 1976) into atransport
code TRANQL (Cederberg, 1985)
• Theisetal. (1983); Kirkneretal. (1984,1985)
• EPRI (Electric Power Research Institute) combined the
transport code SATURN (Huyakorn et al., 1983) with
85
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MINTED (Feltnyeta!., 1984;KrupkaandMorrey, 1985)
to produce FASTCHEM
• Narasimhan et al. (1986) united TRUMP (Edwards,
1972), a three-dimensional-transport model, with
PHREEQE (Parkhurst et al., 1980) to produce
DYNAMIX. Liu (1988) developed refinements, and the
most recent version of DYNAMIX (Liu and Narasimhan,
1989b) can handle the thermodynamics of hydrolysis,
aqueous complexatfon, redox reactions, precipitation-
dissolution, and the kinetics of mineral dissolution. It
can model systems in three dimensions and accom-
modate large arrays of both solid and aqueous species.
An earlier version successfully modeled the behavior of
groundwater contamination from an abandoned
uranium-mill-tailings dam (White et al., 1984)
The few cases in which the one-step and two-step ap-
proaches have been compared by modeling identical
problems indicate that they give comparable results, with
the two-step approach requiring less computertime. Cecler-
berg et a!. (1985) modeled the one-dimensional Pato Alto
Baylands groundwater transport with ion-exchange reac-
tions using TRANQL and compared the results with the
earlier one-step analysis by Vatacchi et al. (1981). The
results were the same but execution times were faster in the
two-step analysis. Liu and Narasimhan (1989a) compared
DYNAMIX with THCC (Camahan, 1986), a model similarto
CHEMTRNS (Noorishad et al., 1987). In this study, |the
one-dirnensional transport of uranium in the presence of
redox reactions was simulated in a hypothetical 3-meter
column. Local equilibrium was assumed, and no kinetically
controlled reactions were included. The resulting outputs
from both simulations were similar.
Which of the two type codes is more useful for modeling
transport in underground-injection environments has not
been established. If more computationally efficient algo-
rithms are developed to solve a one-step code, the differen-
ces In execution times between the two may not be of
concern. Although the methods have been shown to yield
similar results, the fundamental differences in method may
lead to extremely different results for certain problems. For
example, a two-step code may have difficulty maintaining
mass balance when rapid precipitation and dissolution
occur. Drift with respect to the mass balance of elements
present in the system can occur when the method uses
linearized ordinary differential equations that approximate
nonlinear differential equations. Codes should thus be
selected for aptness.
6.6 Laboratory Procedures and Protocols
6.6.1 Waste-Reservoir Characterization
Laboratory procedures for waste-reservoir characterization
were not specifically discussed.
6.6.2 Static and Ftowthrough-tntemction Tests
Static and f towthrough tests are often used to determine the
Kd or adsorption isotherms of a substance. The experimen-
tal procedure for determining Kd usually involves agitating a
suspension of a known paniculate mass of soil or other solid
in a solution of defined volume and known concentration of
adsorbent. After a specified time, the concentration of the
adsorbent in solution is measured, and the amount ad-
sorbed is calculated by comparing the initial and final con-
centrations in the solution. Similar procedures are used for
determining Langmuir or Freundlichi adsorption isotherms
except that experiments are repeated at various con-
centrations.
Flowthrough tests approximate adsoirptfon characteristics of
injection-well reservoirs more closely than batch-type tests.
A suitable core material must be selected, however, and the
core should be installed in the measuring apparatus such
that the injected fluid is not bypassed.
If aflowthrough system is too costly or suitable core material
is not available, rock wafers can b3 used. Using a rock
wafer, the researcher allows the hazardous-waste solution
to permeate a saturated wafer and measures the concentra-
tion of the adsorbent as a function of time. The total surface
area exposed may be insufficient to obtain quantitative
measurements, however, and the process of shaping or
disaggregating the sample will also introduce major uncer-
tainties.
Researchers must characterize very carefully the material
under study when performing laboratory studies to support
more sophisticated adsorption models, such as the triple-
layer model. The specific surface area must be measured,
usually by the BET method. The surface-site density and
number of site types must also be measured. As with
simpler adsorption experiments, the exposed surfaces must
be representative of the undisturbed material and should not
be disaggregated. The use of undisturbed material can
present problems in some cases, however, because it is
difficult to conduct adsorption experiments with consolidated
materials like sandstone or shale.
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Even if the surface area and the density of material can be
properly characterized, a lengthy series of experiments
would be required to analyze adsorption characteristics
completely. To compute the appropriate binding constants,
the adsorption capacity of the material must be examined
over a wide range of pH, tonic strengths, and coexisting
solution compositions, including variations in the total con-
centration of adsorbates.
6.6.3 Quality-Assurance/Quality-Control Procedures
EPA regulations require a quality-assurance and quality-
control plan that covers all aspects of a no-migration
demonstration. Only the EQ3/6 code (Wolery, 1986) sup-
ports the kind of stringent quality-assurance documentation
acceptable to EPA because such documentation is expen-
sive to produce. One benefit of quality-assurance
documentation is the traceability of sources of information
from which the computer code or its associated ther-
modynamic database is derived. This traceability facilitates
checking the source material and judging the reliability of the
package for an application.
If the petitioner must derive original thermodynamic and
kinetic data to conduct the simulations for submission to
EPA, then all such derivations should be completely docu-
mented, and all experimental data should be obtained using
analytical methods and procedures acceptable to EPA. All
standards should be calibrated and referenced to National
Bureau of Standards (NBS) guidelines or other acceptable
standards. Deviations from acceptable procedures must be
fully documented and include demonstrations that alterna-
tive procedures yield resultsthe same as orbetterthanthose
accepted by EPA.
6.7 Case Studies
No case studies were discussed.
6.8 Further Research Needs
The following areas need to be developed to advance the
state of the art of geochemical modeling:
• Accurate determination of activity coefficients of ions in
strong mixed electrolytes (brines).
• Better data on and understanding of the ther-
modynamic properties of clays, and thermodynamic
data for minerals and organic aqueous species for
which none are currently available.
An integrated compilation of data in the extensive litera-
ture describing the adsorption of inorganic and organic
species of clays.
Adsorption or ton-exchange models that can be used
for the diverse range of conditions expected in deep-
well injection.
More field validation studies of geochemical codes.
References
Aagaard, P., and H. C. Helgeson. 1982. Thermodynamic
and Kinetic Constraints on Reaction Rates Among Minerals
and Aqueous Solutions, I. Theoretical Considerations.
American Journal of Science 282:237-285.
Aagaard, P., and H. C. Helgeson. 1983. Activity/Composi-
tion Relations Among Silicates and Aqueous Solutions, II.
Chemical and Thermodynamic Consequences of Ideal
Mixing of Atoms Among Energetically Equivalent Sites in
Montmorillonites, Illites, and Mixing Layer Clays. Clays and
Clay Minerals 31:307-317.
Abraham, H. H. 1982. Free Energies, Enthalpies, and
Entropies of Solutions of Gaseous Nonpolar Non-
electrolytes in Water and Nonaqueous Solvents. The
Hydrophobto Effect. Journal of the American Chemical
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Abraham, H. H. 1984. Thermodynamic of Solution of
Homologous Series of Solute in Water. Journal of the
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181.
Adamson,A.W. 1982. PhysicalChemistry of Surfaces, 4th
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Apps, J., L. Tsao, and O. Weres. 1988. The Chemistry of
Waste Fluid Disposal in Deep Injection Wells. Second
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LBL-24337. Lawrence Berkeley Laboratory, Berkeley,
California, pp. 79-82.
Apps, J. A., J. Lucas, A. K. Mathur, and L. Tsao. 1977.
Theoretical and Experimental Evaluation of Waste
Transport in Selected Rocks. 1977 Annual Report of LBL
Contract No. H5901AK, Lawrence Berkeley Laboratory
LBL-7022,139pp.
Apps, J. A., E. L. Madsen, and R. L. Hinkins. 1975. The
Kinetics of Quartz Dissolution and Precipitation. Annual
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Ball, J. W., E. A. Jenne, and M. W. Cantrell. 1981.
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