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).

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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.

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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

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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

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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

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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.

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                                                     16

-------
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                                                  21

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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

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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

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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

-------
 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

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     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

-------
 •   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

-------
 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

-------
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

-------
  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

-------
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

-------
  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

-------
(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

-------
  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

-------
 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

-------
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

-------
  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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Register 53:28-[ 18-28157.
 Van Everdingen, R. O., and R. A. Freeze. 1971. Subsur-
 face Disposal of Waste in Canada.  Department of the
 Environment, Inland Waters Branch, Technical Bulletin No.
 49, Ottawa, Ontario, 64 pp.

 Warner, D. L 1965. Deep-Well Injection of Liquid Waste:
 A Review of Existing Knowledge and an Evaluation  of
 Research Needs. U.S. Department of Health, Education,
 and Welfare, Public Health Service Publication No. 999-
 WP-21,55p.

 Warner, D. L., and J. H. Lehr.  1977.  An Introduction to the
 Technology of Subsurface Waste  Water Injection.  U.S.
 Environmental Protection Agency, EPA Report 600/2-77-
 240,319 pp.

 Willman, H. B., E. Atherton, T. C. Buschbach, C. Collinson,
 J. C. Frye, M. E. Hopkins, J. A. Lineback, and J. A. Simon.
 1975. Handbook of Illinois Stratigraphy. Bulletin 95, Illinois
 State Geological Survey,  Urbana, Illinois, 261 pp.

 Wood, W. W. 1976.  Guidelines for Collection and Field
 Analysis of Ground-Water Samples for Selected Unstable
 Constituents.   U.S. Geological Survey,  Techniques of
 Water-Resources Investigations, Chap. D-2,24 pp.

ZoBell, C.  E. 1946. Studies on Redox Potential of Marine
Sediments. American Association of Petroleum Geologists
 Bulletin 30:477-513.
                                                   63

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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
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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
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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.
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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-
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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
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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
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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

-------
 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-
                                                    81

-------
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.
                                                    86

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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.
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