BEHAVIOR OF PARTIALLY MISCIBLE ORGANIC COMPOUNDS
IN SIMULATED GROUND WATER SYSTEMS
PATRICIA V. CLINE
A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL
OF THE UNIVERSITY OF FLORIDA IN
PARTIAL FULFILLMENT OF THE REQUIREMENTS
FOR THE DEGREE OF DOCTOR OF PHILOSOPHY
UNIVERSITY OF FLORIDA
i l Sy I OF FORIDA LIBRARIES
I sincerely appreciate the technical and editorial
assistance provided by my research director, Dr. J. Delfino.
I also thank Dr. P. S. C. Rao and Dr. P. Chadik for their
advice and for providing opportunities for challenging
discussions, and Dr. J. Dorsey and Dr. R. Yost for serving
on my committee and reviewing this dissertation. Each
member of my committee has contributed to my graduate career
through excellent teaching and creating a positive
This work was funded by grants from the Florida
Department of Environmental Regulations. Special thanks are
extended to Dr. Geoffrey Watts for his role in securing
funds and providing technical support and comments.
I am grateful to Dr. M. Battiste for discussions of
reaction mechanisms, and for providing the use of his
laboratory for the synthesis of brominated ethanes.
Special thanks go to Angle Harder for her hard work,
Linda Lee for her generosity with analyses and information,
Tom Potter for unselfish computer and mathematical
assistance, and Bill Davis for technical support.
I extend warmest and deepest thanks to my husband Ken
for technical assistance and emotional support, and my son
Brendan for giving me joy.
TABLE OF CONTENTS
Chemistry of Alkyl Halides
Gasoline Partitioning .
MATERIALS AND METHODS .. ..
DEGRADATION OF ALKYL HALIDES
Introduction . . . . .
Degradation of 1,1,1-Trichloroethane . .
Degradation of Brominated Ethanes . .
Halogenated Ethenes . . . .
Structure/Rate Relationships of Alkyl Halides .
Simple SN1/E1 Reactions . . .
Comparisons of Geminal Trihalides . .
Effect of Additional Halogens on the Alpha
Carbon . . . . .
Sediment Matrix Affects . . . .
SOLUBILIZATION AND DEGRADATION OF RESIDUAL TCA .
Behavior of Residual Solvent . .
Aqueous Phase Concentrations . .
Advection . . . . .
Degradation Rate . . . .
Model Parameters and Procedures . .
Limitations of the Model Assumptions ..
. . . ii
. . . 1
. . . . 1 2
. . . . 18
. . 22
GASOLINE IN GROUND WATER
Background . . . . 95
Composition of Gasoline . .. 95
Multicomponent Liquid-Liquid Equilibria 97
Statistics and Pattern Recognition
Applications . . . 102
Partitioning of Gasoline Components into Water 105
Fuel/Water Partition Coefficients . .. 105
Water Soluble Blending Agents .. ... 113
Prediction of Kfw for Other Components 120
Changes in Concentrations with Time .. 122
Differences in Water Extracts of Gasolines 129
Equilibrium Concentrations of Major
Constituents . . . .. 131
Visual Comparison of Water Extracts of
Gasoline . . . .. 131
Preparation of the Data Base for Statistical
Analysis . . . . 135
Basic Descriptive Statistics . .. 138
Bivariate Plots . . . 141
Stepwise Discriminant Analysis . .. 148
Principal Component Analysis . .. 155
Summary . . . . 164
SUMMARY AND CONCLUSIONS . . . 166
APPENDIX A. SOLUBILITY MEASUREMENTS BY LINDA LEE 171
APPENDIX B. FORTRAN PROGRAM FOR MODELING LOSS OF
RESIDUAL TCA . . . .. 173
APPENDIX C. AREA COUNT DATA SET FOR STATISTICAL ANALYSIS
OF WATER EXTRACTS OF GASOLINE . .. 179
REFERENCES . . . . . 187
BIOGRAPHICAL SKETCH . . . . .. 194
Abstract of Dissertation Presented to the Graduate School
of the University of Florida in Partial Fulfillment of the
Requirements for the Degree of Doctor of Philosophy
BEHAVIOR OF PARTIALLY MISCIBLE ORGANIC COMPOUNDS
IN SIMULATED GROUND WATER SYSTEMS
Patricia V. Cline
Chairman: Joseph J. Delfino
Major Department: Environmental Engineering Sciences
Serious ground water contamination problems result from
leaks or spills of organic liquids which are partially
miscible in water. Two important categories of these
liquids include low molecular weight chlorinated solvents
l,l,l-Trichloroethane (TCA) abiotically degrades in
water forming approximately 17-25% l,l-dichloroethene (1,1-
DCE) via an elimination reaction. The substitution product
is acetic acid. The Arrhenius activation energy is 119 +/-
3 kj/mol with an Arrhenius factor of 2 X 1013 s-1, which
results in an estimated half-life for the degradation at
250C of 10.2 months.
Brominated analogs of TCA hydrolyze 11-13 times faster
than TCA. As the number of bromines increase, the percent
of elimination products increases.
These geminal trihalides degrade by a unimolecular
mechanism (E1/SN1). The rate coefficient for TCA
degradation in buffered water at elevated temperature is
approximately six times greater than hydrolysis of 1-
chloropropane (SN2 mechanism) and more than 100 times
greater than l,l-dichloroethane. In the presence of sodium
thiosulfate, the l-chloropropane degradation rate increased
by more than a factor of 100, l,l-dichloroethane rate by 22
and TCA degradation by approximately two.
Halogenated ethenes are stable at various temperatures
and reaction conditions. Trichloroethene degrades in
alkaline solution at elevated temperature.
l,l,l-Trichloroethane and 1,1-DCE form a near ideal
solution in the solvent phase. The solubility of 1,1-DCE at
24C is 3200 mg/l and the solubility of TCA is approximately
The range of concentrations for major components of
gasoline which partition into water was determined for 65
gasoline samples. Benzene concentrations in the water
extracts ranged from 12.3-130 mg/l and toluene
concentrations ranged from 23-185 mg/l.
Fuel/water partition coefficients of seven major
aromatic constituents were measured for 31 gasoline types
and showed a standard deviation of 10-30%. These
coefficients were highly correlated with the pure component
Chemometric techniques were applied to 20 peaks
measured in the aqueous extracts of the 65 gasolines.
Bivariate plots and principal component analyses show
selected brands have distinguishing equilibrium
concentrations, but complete separation of brands was not
Liquids organic compounds ar frequent causes of ground
water contamination. Nonaqueous-phase liquids (NAPL) fall
into two broad categories based on their migration patterns
upon reaching ground water. Mineral oils, including crude
oils as well as various refined products like gasoline, are
less dense than water and move vertically through the
unsaturated (vadose) zone and tend to spread laterally upon
reaching the water table. The majority of spills involving
organic fluids which contaminate ground water result from
this group of compounds (Schwille, 1984).
In many industrialized countries, serious threats to
ground water supplies result from low molecular weight
chlorinated solvents. These anthropogenic substances are
more dense than water and vertical rather than lateral
movement dominates upon reaching the water table. The more
common solvents detected in ground water include 1,1,1-
trichloroethane (TCA), trichloroethene (TCE),
tetrachloroethene or perchloroethene (PCE), and various
dichloroethene isomers. In addition to common usage, the
high frequency of detection is attributed to the compounds'
high mobility and relatively high resistance to degradation.
Decreases in the concentration of contaminants measured
in environmental samples can occur as a result of various
attenuation mechanisms. These include biodegradation,
volatilization, photooxidation, and dispersion. In the
subsurface, losses from pathways like photoxidation are not
important. Other pathways like volatilization occur at
rates slower than those measured from exposed surfaces.
Aerobic biodegradation can occur in the subsurface providing
adequate oxygen and nutrients are available and that the
contaminants are not present in concentrations which are
toxic for microorganisms.
The major objectives of this research include
determining fuel/water partitioning patterns and measuring
chemical degradation rates to aid in the interpretation of
data from contaminated ground water sites. Field
investigations of sites contaminated by gasoline or
chlorinated solvents typically analyze and report the
presence of constituents which are regulated by the state or
federal government (e.g. priority pollutants). These
components are emphasized in my research.
Many chlorinated organic compounds will degrade in
water by hydrolysis or elimination mechanisms. Due to the
extended residence times of organic pollutants in ground
water, this typically slow abiotic degradation within months
or years can be a significant attenuation mechanism. The
focus of my research on the halogenated solvents is on the
transformation processes, and the factors which affect
reaction pathway and the rates of degradation.
Gasoline is a complex mixture of hydrocarbons. Ground
water contamination by gasoline is characterized by elevated
concentrations of the more water-soluble constituents. The
focus of my research on gasoline hydrocarbons is on the
distribution or partitioning of various components of the
gasoline mixtures into ground water and the variability in
the equilibrium concentrations of major constituents.
Chemistry of Alkyl Halides
Abiotic transformation has been reported for TCA, TCE
and PCE, with less work reported on the dichloroethene
isomers. My research reevaluates previous studies and
further examines the chemistry of these compounds.
Mechanisms are evaluated to aid in predicting behavior of
alkyl halides in complex subsurface environments which can
catalyze reactions, lead to the formation of complexes, or
provide localized microenvironments of variable pH or redox
Evidence of the importance of the abiotic
transformation of l,l,l-trichloroethane (TCA) has been
presented (Cline et al., 1986). l,l-Dichloroethene or
vinylidene chloride (1,1-DCE) was one of the five most
frequently detected volatile organic compounds in finished
drinking water supplies, other than trihalomethanes,
according to a survey by the US Environmental Protection
Agency (Westrick et al., 1984). Vinylidene chloride (1,1-
DCE) is a highly reactive, flammable liquid which is
primarily used in the production of copolymers with vinyl
chloride or acrylonitrile. Emissions occur during
manufacturing, shipping and production; however, these
emissions represent less than 1% of the total 1,1-DCE
produced (Environmental Protection Agency, 1985). The
common occurrence of this compound as a ground water
contaminant cannot be entirely explained by its production
and usage patterns.
One source of 1,1-DCE develops during the abiotic
degradation of 1,1,1-trichloroethane (TCA). The production
of TCA is more than three times the production of 1,1-DCE,
and unlike 1,1-DCE, it is an end-use product indicating that
emission to the environment is essentially equivalent to the
production (Environmental Protection Agency, 1985). The
presence of 1,1-DCE is typically associated with the
presence of other alkyl halides. Since 1,1-DCE is more
toxic than TCA, the conversion to 1,1-DCE in ground water
can increase the toxicity of the water supply.
The association of 1,1-DCE with TCA can be seen more
dramatically in field data from sites which show high levels
of chlorinated solvents in ground water. A summary of
volatile organic compounds (VOC's) in Arizona's ground water
(Graf, 1986) states that, of the six most commonly detected
VOC's, only three (1,1,1-trichloroethane (TCA),
trichloroethene (TCE), and tetrachloroethene (PCE)) are used
in large quantities at the industrial facilities. The
presence of 1,2-dichloroethene isomers and 1,1-
dichloroethane, particularly with frequent detections of
vinyl chloride, suggest anaerobic biodegradation (Parsons
and Lage, 1985; Bouwer and McCarty, 1983). Selected
locations show very high levels of 1,1-DCE in association
with TCA, and frequently little evidence of biodegradation
(Table 1). The primary source of 1,1-DCE at these locations
appears to be the chemical degradation of TCA, prompting
questions as to the rate of formation of the 1,1-DCE and its
stability in ground water.
Table 1. Maximum Concentrations (jg/L) of VOC's Detected
at Selected Sites in Arizona (Graf, 1986)
Site TCA 1,1-DCE TCE 1,2-DCE
1 630 3320 13000 20
2 490 1320 9
3 9800 10400 410 933
4 98 206 139 106
Two products are formed during the abiotic degradation
of TCA. The elimination product is 1,1-DCE, while the
substitution or hydrolysis product is acetic acid (Figure
1). Previous research (Cline, 1987) described the rate of
degradation of TCA and formation of 1,1-DCE in dilute buffer
solutions (pH 4-10) at temperatures from 27 to 700C.
H3C-CQ'+ + CI
Figure 1. Abiotic degradation pathways for 1,1,1-trichloroethane.
Transformation processes are most evident in field data
when the degradation products accumulate and are analyzed
and reported, as shown for TCA. The slow degradation of
priority pollutants to products which are not analyzed or
reported (alcohols, aldehydes, or acids which are not
regulated substances) is not as easily characterized in
field investigations. This may occur during degradation of
chlorinated ethenes. The common occurance of TCE and PCE,
as well as the formation of dichloroethene isomers during
degradation, suggest additional study of pathways of the
The relative importance of anaerobic biodegradation
versus chemical degradation on a site (Table 1) may be
inferred by observations of the amount of the biodegradation
product of TCE (cis-1,2-DCE) or of TCA (l,l-dichloroethane)
as compared to the chemical degradation product of TCA, 1,1-
DCE. Specific site conditions can affect the relative rates
of these attenuation mechanisms. Abiotic degradation rates
increase as the ground water temperature increases.
Biodegradation rates may be influenced by many factors
including presence of other organic, redox potential,
oxygen concentration, and nutrients.
The specific objectives of my research are to examine
the degradation rate and pathways for halogenated ethanes
and ethenes and determine factors which may affect these
Gasoline contamination of ground water has become a
major environmental concern. Documented cases of
contamination from underground storage tanks (Florida
Department of Environmental Regulation, 1985) have prompted
enactment of additional legislation, the "State Underground
Petroleum Environmental Response Act of 1986" (SUPER Act),
to protect the ground water and surface waters of the state
of Florida. The SUPER act was designed to maximize ground
water protection, encourage early detection, reporting, and
clean-up of leaking underground storage tanks.
Issues relating to the behavior of gasoline components
in ground water are diverse and complex. Gasoline itself is
a complex mixture of hydrocarbons and some of the factors
which affect the concentration of these constituents in the
subsurface environment (vadose zone and ground water)
include solubility, biodegradability, volatility, soil
sorptive capacity, and dilution.
Components of gasoline may undergo abiotic chemical or
photochemical oxidations through free radical formation.
Thermal degradation is negligible at environmental
temperatures below 800C. Since free radicals are limited in
the subsurface environment, chemical degradation is not
expected to play a significant role there (Bossert and
Aerobic biodegradation will be an important attenuation
mechanism provided that sufficient oxygen and nutrients are
present, and these components typically become limiting
after a spill or leak. Attempts to stimulate aerobic
biodegradation of underground petroleum need to remedy both
nutrient and oxygen deficiencies. In addition, hydrocarbons
in the C5-C9 range (which are typical of gasoline) have
relatively high solvent-type membrane toxicity which will
reduce the number of microorganisms and therefore, decrease
the amount of biodegradation following a gasoline spill
(Bossert and Bartha, 1984).
Sites which have been contaminated by gasoline spills
occasionally report results of the analysis of the "floating
layer." Recovery wells to remove the residual organic
liquid are typically installed as an early remediation
measure. Ground water is typically analyzed for benzene,
toluene, and the xylenes (BTX) and more recently for the
oxygenated gasoline additive methyl tertiary butyl ether
Concentrations of the BTX or oxygenated constituents
will vary spatially and temporally. At the source, changes
in relative concentrations of hydrocarbon components occur
through weathering, primarily volatilization and
solubilization of the liquid residual organic constituents,
resulting in increasing concentrations of the least mobile
constituents. Compounds detected in ground water
downgradient from the spill occur as a result of transport
from the source, and therefore show higher concentrations of
the more mobile constituents.
The downgradient aqueous concentrations are dependent
on the initial partitioning of the gasoline components into
water at the source. The presence of the residual
hydrocarbon will dominate the partitioning process, with
soils playing an increasing role as the residual hydrocarbon
is depleted. Field data are complex to interpret. This is
due to many factors, including site heterogeneities, well
construction and sampling variables, and lack of detailed
information which can provide estimates of the rates of
partitioning and transport. However, patterns resulting
from physical processes, i.e. partitioning and transport,
may be observed. In Table 2 are summarized the highest
concentrations of BTX components measured in monitoring
wells at various gasoline spill/leak sites in Florida.
Table 2. Maximum concentrations (mg/L) of BTX components
in monitoring wells at selected gasoline
contamination sites in Florida.
County Benzene Toluene Xylenes
Hillsborough 24 64 16
11 46 15
Volusia 10 28 11
8 46 9
Desoto 0.8 60 9
These concentrations are similar to those measured in
laboratory gasoline-water partitioning experiments in this
study in spite of differences which exist in the age of the
spills and various physical and biological factors. The
time component for the weathering of gasoline at the source
is dependent on many site-specific factors. Even the
relative contributions of volatilization and solubilization
will depend on conditions like the depth of the water table
at the time of the spill and subsequent water table
A simplification of the complex problem of determining
patterns of gasoline constituent concentrations following a
spill is to initially focus on the partitioning of gasoline
components from the fuel to water. This allows estimations
of equilibrium concentrations of different components from a
fresh.spill in contact with water. Different brands and
grades of gasolines may then be evaluated to determine if
differences among the source types are distinguishable, and
how differences in composition affect the partitioning
The major objectives of the gasoline study include
determination of the variability in the fuel/water partition
coefficients for aromatic constituents. Factors which may
affect the partitioning (concentration, cosolvents) will be
evaluated. Chemometric analyses on hydrocarbon components
present in the aqueous solution in equilibrium with gasoline
will be performed to evaluate similarities and differences
in various brands and grades of gasolines.
MATERIALS AND METHODS
Reagent grade chemicals (Fisher Scientific) were used
to prepare buffers and standard solutions. Phosphate
solutions (0.05 M) were prepared at pH 4.5, 7.0 and 8.5 by
mixing stock solutions and monitoring the pH with a Fisher
Accumet model 230A pH meter. Solutions of 0.05 M potassium
dihydrogen phosphate and 0.05 M potassium hydrogen phosphate
were prepared using distilled deionized water. Equal molar
volumes were used for the pH 7.0 buffer. The phosphate
solutions at pH 4.5 (potassium dihydrogen phosphate) and pH
8.5 (potassium hydrogen phosphate) required minor pH
adjustment using 0.05 M phosphoric acid or potassium
Stock standard solutions of TCA and 1,1-DCE were
prepared in methanol at concentrations of approximately 1
mg/mL. Working standards were prepared by spiking
approximately 5 pL of the stock standard solution into 10 mL
of distilled deionized water. Aliquots of 100-500 pL of.the
working standards were used to prepare standard curves for
the response of the gas chromatograph (GC) to the
concentration of analyte.
Seawater samples were obtained from the coastal
Atlantic Ocean near Ormond Beach, Florida. Samples were
filtered and subsequently handled similar to the phosphate
Ground water samples from two monitoring wells were
obtained from a site in Orlando, Florida, which had been
contaminated by chlorinated solvents. These samples were
purged to remove existing solvents and interfering
substances, then filter (10 pm) sterilized.
Approximately 6.6 mL of the phosphate solutions,
seawater or distilled deionized water were added to 5 mL
(nominal volume) glass ampules (Wheaton Scientific). The
ampules were plugged with cotton and autoclaved for 15
minutes at 1210C.
These ampules were then aseptically spiked with 10 pL
of the stock solution of TCA in methanol and flame sealed
using a Model 524PS sealing unit manufactured by O.I.
Corporation. Final concentrations were approximately 1-3
mg/l. Approximately 0.5 to 1 mL of air space was present in
the ampules after sealing.
Ampules were incubated at 280C (Precision Scientific
Model 6) and at 370C (Precision Scientific Model 4).
Experiments at higher temperatures were performed in a
Magna-Whirl constant temperature water bath (Blue M).
Samples were analyzed using a purge and trap device
(Tekmar LSC-2), interfaced with a Perkin Elmer Model 8410 GC
with flame ionization detector (FID) which employed a 30 m
J&W DB-I, 0.53 mm i.d. wide bore capillary column with a 3
pm stationary film thickness. The temperature program
included a 10 minute hold time at 300C and temperature
ramping of 5C/min to 800C. The helium flow was 2.5 mL/min.
Selected analyses were performed by gas chromatography/mass
spectrometry (GC/MS) for quantification and confirmation of
the formation of 1,1-DCE.
The brominated analogs of TCA were not commercially
available. These compounds were synthesized according to
the protocol described by Stengle and Taylor (1970). Two
hundred and fifty milliliters of carbon disulfide (CS2) were
added to a 500 mL, 3-neck flask that was saturated with HBr
vapors at OOC. Excess vapors were trapped over aqueous KOH.
Five milliliters of TCA were added. Five grams of aluminum
bromide (AlBr3) were added to 100 mL anhydrous CS2, placed
in a dropping funnel, and gradually added to the TCA/CS2/HBr
solution over a period of one hour.
This solution was extracted with ice water made basic
with ammonium hydroxide. The solvent was then removed by
distillation and the residue was filtered. An aliquot of
the mixture was added to methanol and spiked into ampules
containing water. Analysis by purge and trap GC showed two
primary peaks and a secondary peak. The major peaks were
determined by GC/MS to be tribromoethane and
dibromochloroethane. A smaller peak was shown to be
bromodichloroethane. Trichloroethane was below detection
levels ( <30 pg/L )in these analyses.
Some of the spiked ampules were heated for a few hours
to determine if halogenated ethenes would be formed, and if
so, to subsequently determine their corresponding retention
times. Two major peaks were identified by GC/MS to be 1,1-
dibromoethene and l-bromo-l-chloroethene. A sample GC
chromatogram containing reactants and products is shown in
Figure 2, with mass spectra of TBA and DBE in Figure 3.
The same analytical conditions were used for the
brominated compounds as were used for TCA, although the
final temperature was slightly increased.
Pure standards of these compounds were not available
for quantification. The degradation rate was determined
directly from the decrease in area counts, since the
response of the external standard remained consistent during
the time of the experiments. However, determination of
molar concentrations was required to determine the percent
transformation to the elimination product.
The response on an FID is generally related to the
number of carbons and can be affected by functional groups.
To determine if the molar response on the FID was affected
by the type of halogen on the molecule (bromine or
chlorine), I examined the response of the trihalomethane
series for which standards were available (Table 3). The
molar response on the FID was the same for this series of
10. 1 .
Figure 2. Sample chromatogram of partially degraded geminal
a t-: r PGN
73S RCT. TIHC: 18.55 TOT AIJV4D- 132141. BASE PK/MrUtli: 136.9 34530.
C 100 26.1
( 93 10 172 251 266
( D) O 1 , i- r ,,
32 80 100 120 140 60 18 200 22 240 260
p m/z *
1MLPVYC/TBA/KCL*e0i1nt4T ,45-450,2:0~OP.24HOV8 7,unD ~ 13S66,~ I 31
DiCS,30M,IUn,8030eS-2SO,S.2D,3SR,SCRYO 874 SCaNS ( 874 SCANS, 15.88 MIMS)
S1.0 MAftS RACCE: 44.0, 269.8 TOTAL 3SUNDs 323S689.
SS 110 16S 220 276 331 384 439 494 550 65S 660 715 77 82S
2* 29 RET. TIME: 10.2 TOT GEUII[- 199?422. EFSE PK/ACiUUI': 10S.O/ 46640.
I *" o I- -
Figure 3. Total ion chromatogram (center) for partially
degraded geminal trihalide mixture, with mass spectra for
1,1,1-tribromoethane and 1,1-dibromoethene.
compounds. Therefore, the molar response factor for TCA was
used to quantify the ethanes containing bromine and the
molar response factor for 1,1-DCE used to quantify the
Table 3. Relative response
e of trihalomethanes on
ng nmoles Counts
616 5.15 24.18
924 3.65 15.19
924 7.73 43.98
1386 5.48 22.02
502 3.06 18.47
386 1.85 9.89
1255 7.65 38.79
965 4.63 21.56
Response Factor, nmoles/area counts.
Analyses for gasoline constituents were also performed
by GC/FID, using a Perkin-Elmer Model 8410 gas chromatograph
with a 30 m wide bore capillary column (J&W, DB-1) having a
3 pm film thickness. The neat gasoline samples were
analyzed by direct injection of 0.05 pL of the fuel.
Gasoline components dissolved in water were determined by
sparging volatiles from water using a Tekmar LSC-2 Purge and
Trap instrument interfaced to the Perkin-Elmer GC. The
temperature program for both neat gasolines and water
extracts included a 13-minute hold time at 350C, temperature
ramping of 30C/min to 90C, then 5C/min to 2000C. The
helium flow rate was 3.0 mL/min.
Between August and December 1986, subsamples of
gasolines were obtained from the Department of Agriculture
and Consumer Services (DACS) Petroleum Laboratory in
Tallahassee, Florida. These samples were originally
collected by field inspectors and shipped for analysis to
assess compliance with ASTM guidelines and represent various
terminals in northern and central Florida. These samples
represented both summer and winter blends. Subsamples were
collected into 40 mL VOA screw cap vials with Teflon lined
septa and stored on ice prior to analysis.
Local samples were also collected from selected gas
stations in Gainesville. Samples were obtained from the
pump in gasoline safety containers, then a subsample was
transferred to a VOA vial and cooled.
Procedures for evaluating the partitioning of gasoline
into the aqueous phase were reported by Coleman (1984) and
Brookman et al. (1985a). Brookman et al. (1985a) measured
concentrations of aromatic compounds in the aqueous phase
with varying rotation contact times and found a maximum
concentration after two hours. Samples were then
centrifuged to separate the two phases. Coleman et al.
(1984) determined that a rotation contact time of 30 minutes
and an equilibration period of approximately 1 hour produced
consistent results and that longer periods had little effect
on the final concentrations.
Saturated, equilibrated solutions of neat gasolines in
contact with distilled, deionized, organic-free water were
prepared. Two mL of gasoline were added to 40 mL water in
VOA vials having Teflon septa. Samples were mixed on a
rotating disk apparatus for 30 minutes at room temperature
(generally 21-230C). The vials then sat undisturbed for one
hour, in an inverted position. Each separated water phase
was removed through the septum at the bottom of the VOA
bottle using a 5 mL syringe. A separate needle was inserted
to allow air to enter the vial so that a vacuum did not form
preventing withdrawal of the water.
Triplicate samples of each water phase were then sealed
in 2 mL crimp-seal vials and refrigerated until the GC
analysis was performed, typically within 2 days. Replicate
extractions, and replicate analyses of extracts were
performed for quality control.
Some overlap or incomplete peak resolution occurred in
the early eluting compounds for both the neat gasoline
samples and the water extracts. Enhancement of the more
water soluble components occurred following aqueous
extraction, making it easier to identify compounds like
benzene and MTBE in the water extract. Toluene was easily
identified in both the neat and water fractions.
When the objective of comparing gasoline samples
involved identification and quantitation of MTBE, analysis
of the water extract provided the most straightforward
interpretation. Although MTBE may be present in gasoline in
quantities approaching 11%, it was more commonly present at
about 5%. MTBE has a lower FID response than the
hydrocarbons, and eluted early in the chromatogram where
several other components also eluted. In samples that did
not contain MTBE, hydrocarbon peaks were present at lower
concentrations at MTBE's retention time. Since MTBE has a
much greater water solubility than these other constituents,
the relative proportion of MTBE to hydrocarbons was
increased in the water extract.
DEGRADATION OF ALKYL HALIDES
In this section the degradation kinetics for 1,1,1-
trichloroethane (TCA) and other 1,1,1-trihaloethanes will be
presented and discussed. These compounds degrade in water
forming both elimination and substitution products.
Specific experiments were performed to determine the
mechanism of this reaction and to describe factors which may
effect the rate or pathway of the degradation.
Mechanisms of hydrolysis/elimination have been studied
for many years and numerous reviews, textbook chapters and
empirical concepts have been developed to describe the
chemical degradation of alkyl halides in water. The
following review provides the framework for subsequent
discussions of alkyl halide structure and reaction
mechanisms where specific examples will be presented. The
information was synthesized from several sources (March,
1985; Carey and Sundberg, 1984; Mabey and Mill, 1978;
Bentley and Schleyer, 1977).
Classical SN1, SN2, El and E2 mechanisms have been
defined as early as 1933 (Figure 4). The distinction
between SN1 and SN2 is whether or not the nucleophilic
-C-C+ + OH-
-C-C+ + X
OH- + -C-X D- HO C X-- HO-C- + XSN2
i I I
+ H20 +
Figure 4. Classical substitution and elimination reaction
mechanisms for degradation of alkyl halides in water.
/C = C
attack at the alpha carbon (carbon containing the halogen
leaving group) occurs before the transition state in the
rate determining step, not the extent to which the bond to
the leaving group is broken. Clear cut differences in
substitution reaction mechanisms are apparent in many
reactions. In practice, there is a spectrum of SN2
mechanisms involving varying amounts of nucleophilic attack,
with SN1 being the limiting case where nucleophilic attack
does not occur before the transition state of the rate
Unimolecular (SN1 or El) processes are favored by
systems that form stable carbocationsI. A classic example
would be the hydrolysis of t-butyl bromide. The more polar
the solvent, the faster the reaction. An increase in ionic
strength will typically increase the reaction rate, unless
the anion is the leaving group ion (common ion effect). The
reaction is independent of the concentration of nucleophile.
The classic SN2 case occurs in molecules with low
steric hindrance and low carbocation stability. Simple
primary halides react by the SN2 mechanism, while secondary
halides react by an SN2 or intermediate mechanism. Solvent
1 For years these were called "carbonium ions".
Recently, it was determined that the term "carbonium ions"
more accurately refers to pentacoordinated positive ions
(e.g. CH5+) and the more typical positive ion intermediates
(R3C ) are "carbenium ions". The term "carbocation"
includes either type and is generally used to describe any
of these intermediates (March, 1985, p. 141-142).
polarity has less effect on the reaction rate than is
observed for SN1 reactions, but the rate is more sensitive
to changes in concentration or strength of nucleophiles.
The E2 reaction occurs when base attacks the hydrogen
at the carbon adjacent to the carbon containing the leaving
group (beta carbon). This reaction occurs at higher pH and
is more rapid for molecules containing a more acidic
Degradation of 1.1.1-Trichloroethane (TCA)
The abiotic degradation of TCA was the subject of my
master's thesis (Cline, 1987) which included a detailed
discussion of related degradation studies and illustrations
of the first order decay of TCA in aqueous solution.
Additional data were collected subsequent to those studies.
This included additional concentration measurements in long
term degradation studies and measurements of rate
coefficients in additional matrices. In this section, a
concise comprehensive summary of these data are presented.
A brief synopsis of previous degradation studies of TCA
which have been reported in the literature is summarized
here. Dilling et al. (1975) performed reactivity studies on
selected chlorinated solvents, including TCA. Estimated
rate coefficients were based on four measurements over a
period of one year for each of two sets of reaction ampules;
one set was maintained in the laboratory and a second set
kept outdoors in Midland, Michigan. The same estimated rate
was reported for each experiment, with half-lives of
approximately six months. Reaction products were not
The hydrolysis of TCA in seawater was reported by
Pearson and McConnell (1975). A half-life of 39 weeks (9
months) was estimated for TCA at 100C with the predominant
reaction being dehydrochlorination to 1,1-DCE. Walraevens
et al. (1974) examined the degradation of TCA in 0.5, 1.0
and 2.0 M sodium hydroxide solutions. The elimination
reaction was not observed, and sodium acetate was shown by
infrared analysis to be the sole reaction product. The
elimination product, 1,1-DCE, was assumed to be stable under
all experimental conditions.
Vogel and McCarty (1987) monitored the degradation of
TCA and formation of 1,1-DCE in water at pH 7 and a
temperature of 20C. The TCA half-life at 200C was
estimated to be between 2.8 and 19 years. Haag and Mill
(1988) report approximately 22% conversion of TCA to the
elimination product, with an extrapolated half-life of 350
days (11.5 months) at 250C.
Degradation experiments were performed at various
temperatures and in different sample matrices. The results
of these experiments are summarized in Table 4. First order
degradation kinetics were observed (Figure 5) in the
data as verified by plotting In [TCA] versus time. Linear
regression analyses were performed on each data set. All
Table 4. Summary of TCA Degradation Rates
and Product Formation
70 pH 4
62 pH 13
53 pH 4.5
39 pH 4.5
28 pH 4.5
1390 +/- 85
1530 +/- 90
1400 +/- 95
1480 +/- 90
1400 +/- 80
565 +/- 35
140 +/- 12
140 +/- 15
144 +/- 20
145 +/- 16
155 +/- 18
133 +/- 14
25 +/- 1.2
24 +/- 1.1
24 +/- 1.2
4.4 +/- 0.2
3.9 +/- 0.2
4.2 +/- 0.2
26 +/- 1
25 +/- 1
26 +/- 2
38 +/- 1
25 +/- 2
24 +/- 2
24 +/- 2
25 +/- 2
23 +/- 3
19 +/- 1
22 +/- 1
17 +/- 1
23 +/- 2
19 +/- 2
21 +/- 2
DW, Distilled organic free water
GW, Ground water matrix
6.8 "" --
0 100 200 300
0 i i -O- -- -- --
0 0.2 0.4 0.6 0.8 1
Figure 5. First order kinetic data for the degradation of
1,1,1-trichloroethane at 280C and pH 4.5, with the
corresponding data for the formation of the elimination
rate constants were based on reactions showing a minimum of
75% degradation of the initial concentration of TCA.
Statistical analyses were performed to assess if the
slopes measured at any given temperature were significantly
different, thus determining the extent to which the sample
matrix, or pH affected the rate constant. The reaction
rates in the buffer solutions (pH 4.5, 7 and 8.5) were not
significantly affected by pH (p < 0.01). In addition, the
rates measured in ground water matrices at 70C (GW1, GW2)
were not significantly different from rates measured in the
buffer solutions at the same temperature.
The spiking solutions typically were prepared with
methanol, which resulted in approximately 0.1% methanol in
the final solution. Separate experiments were conducted
without the use of methanol with no apparent affect on the
rates. The use of methanol decreased the variability in
concentrations observed among ampules, apparently due to the
decreased volatility of TCA in the methanol spiking
Reaction rates at 530C in seawater, distilled deionized
water and 0.05 M phosphate buffer solutions showed that the
ionic matrix affected the rate of reaction. The fastest
rate was observed for seawater, while the rate in distilled
deionized water (DW) was 14% lower and those in the buffer
solutions were approximately 10% lower. The rates measured
in the distilled deionized water and the buffer solutions
were not significantly different; however, the rate in the
seawater matrix was higher than these at the p<0.01 level.
The 10-14% increase in reaction rate observed in the
seawater matrix at this temperature may be due to the
catalytic influence of some component of that matrix, or to
the increase of ionized species concentration in the
The relationship between the rate coefficient, k, and
temperature is expressed by the Arrhenius equation,
In k In A EA/RT, where EA is the Arrhenius activation
energy, R is the gas constant, T is the temperature and A is
the Arrhenius pre-exponential factor. The plot of the data
from this and other studies is shown in Figure 6. The plot
includes rates for a variety of matrices including seawater
and sodium hydroxide solutions. Since two products were
formed, the degradation process was complex, but the overall
linearity of the Arrhenius plot implies that a single rate-
determining step is involved in the degradation. Based on
these results, an activation energy of 119+/-3 kJ/mol and an
Arrhenius (A) factor of 2.0x1013 s'1 were calculated.
Extrapolated rate constants and estimated half-lives are
shown in Table 5.
Table 5. Extrapolated Half-Lives for the Degradation of TCA
Temperature (C) Half-life (years)
15 4.5 +/- 0.8
20 2.0 +/- 0.3
25 0.85 +/- 0.13
-10 This dissertation.
v Pearson and McConnell, 1975.
x Vogel and McCarty, 1987.
-12 o Haag and Mill, 1988.
O Dilling et al., 1975.
+ Wolroevens et al., 1974.
2.8 3 3.2 3.4 3.6
Figure 6. Arrhenius plot for the abiotic degradation of 1,1,1-trichloroethane.
Included in the Arrhenius plot are the degradation rate
coefficient for TCA in a pH 13 buffer and also the rate
coefficients calculated by Walraevens et al. (1974) for the
sodium hydroxide solutions. The rates for these high pH
solutions were within the confidence interval for the
regression line, indicating the reaction rate was not
significantly accelerated in alkaline media. The lack of
change in the rate in the presence of a high concentration
of a strong nucleophile (i.e. OH') suggested that the
reaction with the nucleophile occurs after the rate
determining step, characteristic of SN1 reactions.
Similarly, the increase in base strength did not shift the
elimination to an E2 mechanism through a large rate increase
and/or increase in formation of the elimination product.
The rate data which exceeded the confidence interval of
the regression line (Figure 6) were from studies (Vogel and
McCarty, 1987; Pearson and McConnell, 1975) which estimated
the rates of the slow reactions with less than 50%
degradation of the parent compound occurring. Rate
constants calculated for low conversion are more variable
than rates established based on higher amounts of conversion
(Levenspiel, 1972, p. 85). The strong linear Arrhenius
relationship between temperature and rate observed between
25 and 800C, regardless of sample matrix, suggests that
reaction rates at temperatures below 250C can be estimated
The elimination product, 1,1-DCE, was measured to
establish the factors which influenced the reaction pathway
(substitution versus elimination). Degradation of 1,1-DCE
was observed only at very high pH and even under those
conditions the rate was slow compared with the degradation
of TCA. Therefore, the ratio of the rate for elimination
(ke) to the total rate of degradation (k) was estimated by
plotting the concentration of 1,1-DCE versus (l-e-kt) where
t is time. The slope of the line equals ([TCA]o (ke/k)),
where (TCA]o is the concentration of TCA at time zero.
This calculation required an estimate for the starting
concentration of TCA. For most experiments, multiple
analyses were performed for the estimate of the initial
concentration. Other authors (eg. Vogel and McCarty, 1987)
have used the intercept in the regression analysis of the
degradation, and this value was used as the estimate of
initial concentrations in this study.
Increases in pH and/or temperature theoretically favor
elimination over substitution. The elimination pathway
(Table 4) ranged from 17 to 38% of the total degradation
rate of TCA. Higher temperatures showed slightly more
transformation to 1,1-DCE over the temperature range
evaluated in these experiments. The percent of TCA
degradation due to elimination was not affected by matrix in
the pH range of 4.5 to 8.5. Seawater had no apparent effect
on the relative proportion of products. The highest percent
elimination pathway was measured in the strongest sodium
hydroxide (pH 13) solution.
Qualitative observations (GC and GC/MS) of TCA
degradation at approximately 600C in 0.5, 1.0 and 2.0 molar
sodium hydroxide solutions, showed the presence of 1,1-DCE,
and separate experiments indicated that 1,1-DCE also slowly
degraded under those conditions. These findings contradict
the results reported by Walraevens et al. (1974) in which
1,1-DCE was not detected in TCA degradation experiments at
high pH. This may be due to differences in analytical
methods, or the slow degradation of 1,1-DCE under their
Degradation of Brominated Ethanes
The degradation rates of brominated versus chlorinated
1,1,1-trihaloethanes were compared to provide insight into
the mechanisms and overall behavior of these compounds.
Since bromine is a better "leaving group" than chlorine,
brominated compounds typically degrade faster than their
chlorinated counterparts. In reviewing hydrolysis
degradation processes, Mabey and Mill (1978) concluded that
Br is more reactive than Cl by a factor of 5 to 10.
In a search of Chemical Abstracts, fewer than 20
references were reported for the brominated analog of TCA,
1,1,1-tribromoethane (TBA). Most of the papers addressed
spectra and bond energy studies, while no information on the
hydrolysis of this compound was reported.
Brominated analogs of TCA were not commercially
available. Therefore, TBA was synthesized according to the
methods reported by Stengle and Taylor (1970). The
procedure for the synthesis of 1,1,1-tribromoethane (TBA)
produced a mixture of brominated analogs of TCA. The
primary components were TBA and l,l-dibromo-l-chloroethane
(DBCA), while smaller quantities of l-bromo-l,l-
dichloroethane (BDCA) were present. Kinetic data for
abiotic degradation of TBA and DBCA were measured for
several temperatures while data for BDCA were obtained in
only selected experiments conducted at higher overall
concentrations. Compound structures are illustrated in
Figure 7. The elimination pathway involved loss of HBr to
form the corresponding alkene, the dominant elimination
product was the ethene formed by loss of a bromine. The
substitution pathway forms acetic acid.
Initial degradation experiments involving the
synthesized brominated mixture were conducted in reagent
grade (Milli-Q) water to obtain preliminary data on the
transformation process. Subsequent experiments were
conducted in buffer solutions at pH 4, 7, and 10. The
results of these experiments are summarized in Table 6.
First-order kinetics of degradation were observed, as
were also seen for TCA. Rate constants were calculated
from the linear regression analysis of the plots of the
1,1-Dibromo-1 -chloroethane (DBCA)
1 -Bromo-1.1 -dichloroethone (BDCA)
1-Bromo-1 -chloroethene (BCE)
Figure 7. Brominated analogs of 1,1,1-trichloroethane and
corresponding elimination products. Since bromine is a
better leaving group than chlorine, the predominant pathway
is elimination of HBr.
Table 6. Summary of Brominated Compound Degradation
Rate Coefficients and Product Formation
20 pH 4
20 pH 7
20 pH 10
30 pH 4
30 pH 7
30 pH 10
37 1 M KC1
20 pH 4
20 pH 7
20 pH 10
30 pH 4
30 pH 7
30 pH 10
37 1 M KC1
DW, Distilled organic free water
Na2S203, 1 M Sodium thiosulfate
Extrapolated Half-Lives for Degradation of TBA and DBCA
natural log of the concentrations versus time. All rate
constants were based on reactions showing a minimum of 75%
The results of the degradation of TBA, DBCA and BDCA at
65C are illustrated in Figure 8. The differences in
slopes for the degradation of these compounds were not
statistically significant indicating that the rate
determining step was similar for each compound.
The formation of products (Figure 9) was calculated as
discussed previously for the formation of 1,1-DCE. The
percent elimination (ke/k) was the slope of the regression
line divided by the initial concentration of the parent
product. The smaller slope for 1,1-DCE, and its lower
maximum concentration, was a function of both lower initial
concentration of reactant (BDCA) and lower percent of BDCA
degradation which occurred through the elimination pathway.
The Arrhenius plot for TCA as determined in this study
is compared in Figure 10 with that of the brominated
compounds, TBA and DBCA. The Arrhenius plot for the two
brominated compounds was represented by a single regression
line. The regression line for TCA was essentially parallel
to that of the brominated compounds. The Arrhenius
activation energy (EA) for all of these compounds was almost
identical, since EA is a function of the slope of this line.
The rate of degradation of TCA at 25C was
approximately a factor of 11 to 13 times slower than for the
0 2 4 6
Figure 8. First order kinetic data for the abiotic
degradation of TBA, DBCA and BDCA in water at 650C.
Figure 9. Formation of the elimination products (BCE, DBE,
DCE) in water at 650C from the abiotic degradation of the
+ BCE :
* DBE ,
* DCE ++
2.8 3 3.2 3.4 3.6
Figure 10. Arrhenius plot for the abiotic degradation of 1,1,1-trihaloethanes.
brominated analogs. The observed rate constants for the
various pH values were not significantly different, as was
also observed for the degradation of TCA.
Experiments were performed to determine the reaction
mechanism for these 1,1,1-trihaloethanes. First, the
degradation experiment was conducted in a 1 molar sodium
thiosulfate solution. Sodium thiosulfate is a much stronger
nucleophile than water or hydroxide (Swain and Scott, 1953)
and a dramatic increase in degradation rate in this solution
is indicative of an SN2 reaction in which the nucleophile is
directly involved in the rate determining step. The
degradation rates measured for the brominated 1,1,1-
trihaloethanes increased less than a factor of 2 in the
thiosulfate solution, which may be attributed to the
increased ionic strength of the solution. The percent
elimination was also unaffected by this sample matrix.
To further characterize the mechanism for these
degradation reactions, the brominated geminal trihalides
were placed in a 1 M KC1 solution at 370C. High ionic
strength solutions generally increase the rate of SN1 or El
reactions. When a common ion is present, the rate of the
reverse reaction is enhanced. In the presence of high
concentrations of chloride, chloride may be exchanged for
bromide when an ion pair forms. If BDCA forms an ion pair
and chloride is exchanged, TCA will be formed (Figure 11)
providing evidence of a carbocation intermediate. Even
though TCA will degrade, it is more stable than the
brominated compounds and it may accumulate to detectable
The BDCA compound had the lowest concentration in the
mixture of the three geminal trihalides in reaction
solution, and TCA concentration was less than 40 ug/l.
After three days of incubation at 370C, the concentration of
TCA rose to approximately 200 ug/l. This was a minor
pathway (less than 5% of the BDCA degraded forming
detectable TCA) in the overall degradation process. 1,1,1-
Trichloroethane was not detected in other sample matrices
during the degradation experiments of the brominated
compounds, indicating that its presence in this solution was
a result of the reverse reaction of carbocation with the
chloride in the solution.
Increasing the extent of bromination increased the
percent of the degradation resulting in the elimination
product (Table 6). The proportion of the total degradation
which resulted in elimination for BDCA at 650C was within
the error estimate for the percent elimination of TCA at
elevated temperatures, and both of these parent compounds
produced 1,1-DCE. The highest percent elimination was
observed for TBA which formed approximately 60% 1,1-
dibromoethene (Figure 12). This may be due to an increase
in steric hindrance in carbocations containing bromine
rather than chlorine, slowing the substitution pathway.
Figure 11. Reaction pathways for BDCA in 1 M KC1 solution. The exchange reaction of Cl
with the ion pair forms TCA, which degrades more slowly than the brominated compound.
Figure 12. Comparison of the percent of the elimination pathway for 1,1,1-trihaloethanes.
These experiments provided evidence that the abiotic
degradation of 1,1,1-trihaloethanes occurred by SN1/El
rather than SN2 and E2 mechanisms. The trihaloethanes
containing one or more bromine atoms degraded at similar
rates, approximately a factor of 11-13 faster than TCA,
reflecting that bromine was a better leaving group. As the
number of bromines present on the trihaloethanes increased,
the percent of the degradation occurring through the
elimination pathway increased.
Degradation of Halogenated Ethenes
One of the primary objectives of examining the behavior
of halogenated ethenes was to provide an accurate evaluation
of their formation and stability during degradation of the
corresponding ethanes. The literature provided some
evidence that slow degradation of these ethenes may occur at
a rate of interest for ground water studies.
Supporting the possibility of degradation, Dilling et
al. (1975) reported half-lives for the abiotic degradation
of trichloroethene (TCE) of 10.7 months (0.002 day-1) and
for tetrachloroethene (PCE) of 9.9 months at 250C.
Molecular oxygen was present and the degradation rates were
suggested to result from oxidation as well as hydrolysis.
In this often referenced work, it was suggested that
mechanisms of degradation at lower temperatures may differ
from rates extrapolated from studies at higher temperatures.
In a study of hydrolytic decomposition by Pearson and
McConnell (1975), volatilization was extrapolated to zero
and a degradation half-life for TCE of 30 months was
Roberts (1985) examined field evidence for the
degradation of various chlorinated organic and estimated
rate constants for both TCE and PCE of approximately 0.003
day-1, which may be due to a variety of factors including
sorption and dilution.
Wilson et al. (1985) studied the aerobic degradation of
TCE, PCE and other compounds in actual aquifer materials
from two sites in Oklahoma and Louisiana. No detectable
biodegradation of these compounds was observed under the
experimental conditions. Since degradation was noted in
autoclaved samples, the authors postulated that TCE and PCE
degradation was likely due to abiotic processes with rates
similar to those reported by Dilling et al. (1975).
The dehydrochlorination reaction of TCE occurs under
basic conditions and generates dichloroacetylene and
hydrogen chloride. This reaction of TCE with base is
spontaneous at room temperature and was responsible for
dichloroacetylene intoxication observed in patients inhaling
TCE-containing air in closed systems equipped with alkali
absorbers (Environmental Protection Agency, 1979).
Dichloroacetylene was detected in the gas phase above
aqueous alkaline solutions with pH 11 to 13 and upon
incubation with moderately alkaline material such as
concrete (Greim et al., 1984). They concluded
dehydrohalogenation can occur under these relatively mild
conditions resulting in toxicity from exposure to the
Many substitution and addition reactions of TCE have
been carried out in the presence of base. What initially
appeared to be a direct substitution reaction may in fact
have been multistep processes involving intermediates like
carbanions, chloroacetylenes, or carbenes. Rappaport (1969)
reviewed the mechanisms for nucleophilic vinylic
substitution processes in alkaline solutions at elevated
Mechanisms may differ for chemical studies performed
under extreme conditions of temperature and high pH compared
to reactions occurring under more typical environmental
conditions. The possibility of slow nucleophilic attack in
aqueous solution was considered because March (1985) reports
that although vinyl halides are generally considered
resistant to nucleophilic attack, the presence of electron-
withdrawing groups like halogen lower the electron density
of the double bond enhancing nucleophilic substitution or
In ground water, even very slow degradation may be an
important attenuation mechanism. Since environmental
studies report slow degradation of TCE or PCE in water and
the chemical studies show presence of electron withdrawing
groups like chlorine increases the susceptibility of an
olefin to nucleophilic attack, experiments to evaluate
possible reactions were performed.
References to possible hydrolysis reactions of 1,1-DCE
or its brominated analogs were not found upon review of the
literature. The 1,1-dihaloethenes would be less susceptible
to nucleophilic attack than TCE since fewer electron
withdrawing groups are present. The pure compounds however,
are very reactive and polymerize readily. Their
reactivities in dilute aqueous solution have not been
The focus of my research with halogenated ethenes was
to examine the stability of these compounds in relatively
dilute aqueous solutions and to determine their
susceptibility to nucleophilic attack. Autooxidation or
other reactions of the pure liquid compounds which may be
present in the vadose zone following a spill could occur,
but these reactions are not addressed here.
There were two major purposes for the examination of
the degradation behavior of halogenated ethenes. First, the
stability of the ethene products formed during the
transformation of the geminal trihalides needed to be
determined to accurately describe the kinetics of the
appearance of these elimination products. Secondly,
previous studies which indicated that halogenated ethenes
like trichloroethene (TCE) and tetrachloroethene (PCE) may
undergo slow abiotic degradation in water at room
temperature with a half-life of less than one year were
reevaluated. The question of possible nucleophilic attack
by water, hydroxide ion, or other nucleophiles must be
addressed to understand the stability of these commonly
detected ground water contaminants.
The stability of 1,1-DCE was evaluated in experiments
that were performed concurrently with the evaluation of TCA
degradation. In the buffer solutions, seawater, and
distilled deionized water, no significant degradation of
1,1-DCE occurred during the course of the evaluation of the
degradation of TCA.
The formation of ethenes containing bromine was
monitored during the degradation studies of the brominated
ethanes, and their concentrations were continually monitored
for some time after the ethane degradation was completed.
Trichloroethene was studied in separate experiments
performed at various temperatures selected to repeat the
experiments conducted by Dilling et al. (1975). In addition
to buffer solutions, one set of ampules was prepared with a
nutrient solution which was not autoclaved, and to which
ground water known to show biological activity was added.
This was done to determine if any degradation which might
have occurred during the long term studies could have been
due to biological activity.
A summary of the results of these experiments is
presented in Table 7. No significant degradation of these
compounds was found in the experimental matrices during the
indicated reaction times, as evidenced by the slopes of In C
vs time which were not significantly different from zero.
The overall coefficient of variation for the observations is
similar to values obtained for simple replicate analyses.
Experiments were also conducted to evaluate the overall
behavior of these compounds under more rigorous conditions.
The literature indicated that halogenated ethenes such as
TCE can undergo elimination to form chloroacetylenes at
elevated pH (Rappaport, 1969). This reaction was verified
by using GC/MS to confirm the formation of dichloroacetylene
from TCE and also chloroacetylene from 1,1-DCE by analysis
of the headspace vapor above an alkaline (1 M NaOH) aqueous
solution of the halogenated ethene which was warmed to
The rate of degradation of components in a mixture of
1,1-DCE, TCE and PCE in sodium hydroxide solutions was
examined at 600C (Table 8). These were the matrices used by
Walraevens et al. (1974) in their examination of the
degradation of TCA, wherein they did not observe formation
of 1,1-DCE. One objective was to establish if the
elimination product was stable under their reaction
Summary of Experimental Conditions for which
Halogenated Ethenes were Stable.
Cmpd. Days Obs.
DW Distilled organic free water
Thio 1 M Sodium thiosulfate solution
Table 8. Second Order Degradation Rates (1 mole-1 hr-1)
of Halogenated Ethenes at 600C
in Sodium Hydroxide Solutions
NaOH TCE 1,1-DCE PCE
0.1 M 0.6 0.02 nd
0.5 M 0.28 0.01 nd
1.0 M 0.17 0.01 nd
2.0 M 0.12 0.004 nd
nd no significant degradation occurred after 260 hours.
The rate of degradation of TCE was the greatest among
the tested compounds due to the presence of an acidic
hydrogen (a hydrogen present on a carbon containing a
halogen). The elimination reaction was also an available
pathway for the degradation of 1,1-DCE, although the rate of
degradation was approximately 30 times slower than for TCE
in all solutions except for the 1.0 M NaOH.
Tetrachloroethene (PCE) did not degrade since the
dehydrohalogenation reaction could not occur, and apparently
conditions were not favorable for an addition process.
For environmental applications, there are concerns with
the mildest conditions (temperature and pH) which may still
result in degradation of these compounds. The pathway for
the degradation of ethenes at elevated temperature could
differ from reactions at lower temperatures where the
elimination reaction would be less favorable and a possible
addition reaction could occur instead. Therefore, TCE was
incubated at 200C in a solution at pH 12.5. No degradation
was observed during four months of incubation (Table 7).
The experiments demonstrate the resistance of the
halogenated ethenes to degradation in dilute aqueous
solution. Reports of the degradation of these compounds
with half-lives of less than 1 year appear to represent a
process other than abiotic degradation in water. In the
same way as Dilling et al. (1975), my experiments were
conducted in sealed ampules containing a headspace, however
degradation was not observed as reported in their study. I
believe their results may be a result of analytical error.
The half-lives for each experiment were based on four
measurements. The results showed a chemically diverse group
of compounds had similar decreases in concentration and
temperature had little effect on these decreases. A
possible explanation for these results would be a decrease
in instrument response over the year of the study.
The halogenated ethenes generally showed very little
degradation, with the exception of the rapid degradation of
TCE at high pH and temperature. It appears that any
degradation of these compounds in aqueous solution which
occurs, does so under rather extreme conditions and is not
expected to be a dominant process.
Structure/Rate Relationships of Alkyl Halides
In the previous sections degradation patterns and
kinetics were evaluated for various 1,1,1-trihaloethanes in
aqueous solution. A broader perspective on hydrolysis /
elimination reactions can be obtained by comparisons with
other haloalkanes reported in the literature. The
1. To compare degradation rates measured for
trihaloethanes of other simple alkyl halides which react by
an SN1/E1 mechanism.
2. To compare degradation rates of trihaloethanes with
other geminal trihalides reported in the literature to
determine structure/activity relationships with changes in
the substituents on the beta carbon, and describe shifts in
mechanisms which may occur for these trihalides.
3. To compare degradation rates and pathways of 1-
chloropropane and l,l-dichloroethane with TCA to show the
effects of increasing number of chlorines on the alpha
The classic reaction mechanisms for substitution and
elimination reactions are SN1, SN2, El and E2, as previously
discussed. The presence of various functional groups can
effect the rate and pathway of degradation of an alkyl
halide. For example, rates of hydrolysis are greater for
alkyl halides containing Br rather than for Cl by a factor
of 5 to 10. The rates also increase as the alkyl group goes
from primary to secondary to tertiary in the ratio of
1:10:1000 for chloride. Allyl groups enhance the rate of
hydrolysis of a primary halide by a factor of 5 to 100,
while benzyl groups enhance the rate by a factor of 50.
(Mabey and Mill, 1978)
The formation of stabilized carbocations by electron
donation from the non-bonded electron pairs of halogens
adjacent to the cationic carbon center have been reported
(Olah, 1974). The stabilizing effect was enhanced when two
or even three electron-donating heteroatoms coordinate with
the electron-deficient carbon atom as illustrated in Figure
13. Specific examples, designated as "chlorocarbenium
ions" by Olah (1974), have been identified and are
illustrated in Figure 14.
Simple SN1/E1 Reactions
My data suggested 1,1,1-trihaloethanes form carbocation
intermediates. The intermediate would contain two halogens
and one methyl group. The observed rates and pathways are
compared (Table 9) to compounds containing two methyl groups
and one halogen (2,2-dihalopropanes) and three methyl groups
Degradation of tertiary halides like t-butyl chloride
occurs with a carbocation intermediate and these compounds
are resistant to bimolecular nucleophilic displacement. The
half-life for the aqueous degradation of t-butyl chloride is
approximately 23 seconds at 25C with about 19% of the
degradation occurring through the elimination pathway. The
carbocation intermediate is stabilized by the three methyl
Figure 13. Stabilization of carbocations by halogen (Olah, 1974).
;-< ^ cl
Figure 14. Examples of "chlorocarbenium ions" (Olah, 1974)
Table 9. Summary of Degradation Rate Coefficients and
Pathways for Tertiary and Secondary Halides
1. March, 1985.
2. Queen and Robertson, 1966.
3. This dissertation.
groups. The rate coefficient at 25C is approximately 106
faster than for TCA.
Queen and Robertson (1966) examined the hydrolysis of
2,2-dihalopropanes. These compounds form carbocation
intermediates with two methyl and one halogen group. The
rate coefficients for the degradation of 2,2-dihalopropanes
are intermediate between t-butyl chloride and the 1,1,1-
trihaloethanes. The mechanism was reported to be SN1/El
based on results of experiments with deuterated gem-
dihalides. The degradation rates of these compounds were
10-50 times higher than of the corresponding secondary
halides (e.g., 2-chloropropane).
The degradation rates were affected by the leaving
group, bromine or chlorine. Also, the structure and
stability of the resulting carbocation affected the rate and
pathway (elimination and/or substitution) of the reaction.
Since bromine was a better leaving group than chlorine,
there was a rate increase when bromine was present as
compared to the corresponding chlorinated compound. 2,2-
Dibromopropane degraded 19 times faster than 2,2-
dichloropropane, while 2-bromo-2-chloropropane degraded 5
times faster than the dichloro compound (Queen and
Robertson, 1966). The 1,1,1-trihaloethanes containing
bromine degraded 11-13 times faster than TCA.
Rates were also increased as the number of methyl
groups present on the carbocation increased. The t-butyl
chloride degraded approximately 3000 times faster than 2,2-
dichloropropane and 106 faster than TCA.
There were two major differences between my results and
those reported by Queen and Robertson (1966). First, they
reported a rate nearly four times higher for 2-bromo-2-
chloropropane than for 2,2-dibromopropane, while the rate
coefficients I measured for the trihaloethanes containing at
least one bromine were approximately equal (within 20%).
Secondly, they report only formation of the elimination
product for all 2,2-dihaloethanes, while the percent
elimination in my experiments was a function of the number
of bromines and was always less than 60%. The percent
elimination for t-butyl chloride was less than the value
obtained for the trihaloethanes.
The effect of alpha halogen is complex, "combining a
negative inductive effect and an electron-releasing
resonance effect" (Queen and Robertson, 1966, p. 1364).
Based on my results and the results for t-butyl chloride,
elimination was not expected as the primary pathway nor the
large difference in rates observed for the two
dihalopropanes which contained a bromine. The rate data
were determined for the dihalopropanes by a conductance
method. Extraction of the products of solvolysis of 2,2-
dibromoethane with CC14 and analysis by vapor phase
chromatography (GC) and nmr showed 2-bromopropene was the
only product in other than trace amounts. It may be that
the substitution product, acetone, would not have
partitioned and been measured using that analytical
Comparisons of Geminal Trihalides
A number of compounds in the literature contain a
geminal trihalide group (R-CX3), and many of these compounds
have environmental implications. My experiments on 1,1,1-
trihaloethanes indicated that the -CX3 group was sterically
hindered and resistant to attack by an SN2 mechanism, and
that the halogens could help to stabilize the formation of a
carbocation. The overall rate of degradation of other
geminal trihalides will increase if R also stabilizes the
carbocation. If the beta carbon contains an acidic hydrogen
the mechanism may shift to E2 at elevated pH.
A summary of degradation rates (expressed as reaction
half-lives) of various geminal trihalides is presented in
Table 10. The simplest compounds, trihalomethanes, were
very resistant to hydrolysis. The R- consists only of
hydrogen, which was inadequate to stabilize a carbocation.
The mechanism for this degradation has been determined to be
a base catalyzed process with a carbanion intermediate
(Hine, 1950). The extremely low reactivity also suggests
that steric hindrance may prevent SN2 attack.
By contrast alpha,alpha,alpha-trichlorotoluene has a
half-life of 19 seconds at 250C, which corresponds to a rate
of a factor of 106 greater than for TCA. Therefore, the
Table 10. Half-lives for Abiotic Degradation of
cI C H CC[3
Mabey and Mill,
Mabey and Mill,
12 yr Wolfe et al., 1977
1 yr Wolfe et al., 1977
(CH 0 CH CCI3
a,a,a-Trichlorotoluene / CC a
19 s Lyman et al., 1982
rate increase was much greater than the factor of 50
reported by Mabey and Mill (1978).
Quemeneur et al. (1971) determined that tri-chloro
compounds of the type p-RC6H4-CC13 (R is OMe, Me, H, Cl, or
NO2) were hydrolyzed in neutral or acidic medium via a
cationic transition state for all types of R substituents.
The hydrolysis of the p-substituted alpha,alpha-
dichlorotoluenes reacted via a cationic mechanism when R is
an electron-donor, and a bimolecular mechanism when R is an
electron-attracting group. These results also supported the
observation that halogens contributed to the stability of
the carbocation. Monochlorotoluene reacts nearly 3000 times
more slowly by an SN2 mechanism than the trichlorotoluene
reacts by the SN1.
Methoxychlor and DDT are two environmentally important
pesticides which contain a geminal trihalide functional
group. Wolfe et al. (1977) provided an in depth examination
of the degradation of these compounds. There is a beta
hydrogen on each of these compounds. At elevated pH the
degradation rate increased as a function of pH and the
elimination products were dominant, suggesting these
structures were more susceptible to degradation by the E2
mechanism than is TCA. While the elimination product, DDE,
was the major product of DDT hydrolysis even at lower pH,
the major product of methoxychlor at pH 7 was the hydrolysis
product, with minor amounts of the elimination product,
DMDE. The hydrolysis products formed were anisoin and
anisil, which were explained by phenyl group rearrangement
after the formation of the carbocation.
Mochida et al. (1967) showed 1,1,1,2-tetrachloroethane
and pentachloroethanes reacted more slowly than TCA under
lower pH conditions, which indicated that chlorines on the
beta carbon decrease the stability of the carbocation. The
presence of these chlorines on the beta carbon however,
increased the acidity of the hydrogens, with enhanced
degradation rates for the tetra and pentachloroethanes by an
E2 mechanism at elevated pH.
There is considerable evidence that geminal trihalides
can form carbocations in the presence of an appropriate
neighboring group. Subsequent reaction pathways may vary
according to the structure of the carbocation resulting in
elimination, substitution, or rearrangements. An E2
reaction may also occur for compounds containing an acidic
hydrogen on the beta carbon.
Effect of Additional Halogens on the Alpha Carbon
The hydrolysis of a simple primary halide, 1-
chloropropane, was compared with the reactivity of 1,1-
dichloroethane and TCA in experiments I performed at
elevated temperature. As the number of hydrogens on the
alpha carbon decrease, steric hindrance can increase and
result in a shift in reaction mechanism. The experiments
were designed to demonstrate the relative rates of
hydrolysis in aqueous solution, and the response to an
increase in concentration of a strong nucleophile whose
effect would be a function of the mechanism.
Based on the literature, simple primary alkyl halides
like l-chloropropane are expected to degrade by an SN2
mechanism. Therefore, l-chloropropane should show an
increase in degradation rate in the presence of a strong
nucleophile, since the nucleophile is involved in the rate
Predicting the degradation rate of l,l-dichloroethane
is more difficult. Secondary chlorides, like isopropyl
chloride, have been shown to degrade more quickly than the
primary alkyl halides, possibly by an intermediate
mechanism. Chloride can contribute somewhat to the
stability of a carbocation, however, it is not as effective
as a methyl group as discussed previously. In addition, the
presence of a halogen can increase the steric hindrance at
the alpha carbon.
Comparisons of the degradation rates of these compounds
were made at elevated temperature (650C) in pH 7 buffer
solution, and in a 1 M thiosulfate solution. In the buffer
solution the degradation of TCA was approximately 6 times
faster than the hydrolysis of 1-chloropropane. Degradation
of 1,1-dichloroethane was less than 6% of the rate of 1-
chloropropane degradation. This rate comparison is
illustrated in Figure 15.
0 200 400 600
Figure 15. Pseudo-first-order kinetic data plots for hydrolytic degradation of TCA,
1-chloropropane, and 1,1-dichloroethane in pH 7 buffer solution at 650C.
The degradation of 1-chloropropane was enhanced by more
than a factor of 100 in the thiosulfate solution, 1,1-
dichloroethane degraded approximately 22 times faster, and
TCA degradation rate increased less than a factor of 2. The
differences in rate enhancement among these compounds is
attributed to differences in mechanism. Part of the
increase in rate of degradation of TCA in thiosulfate is
attributed to the increasing ionic strength, and TCA
degradation rate was clearly less affected by the presence
of thiosulfate than the other compounds. The rate
enhancement for l,l-dichloroethane was similar to the type
of rate increase which would be observed for secondary
halides which react by an intermediate mechanism.
The thiosulfate solution was used as a matter of
convenience as a strong nucleophile to assist in
demonstrating how knowledge of mechanism may be necessary in
estimating degradation rates as matrices change. Greatest
changes in rates in the presence of sulfur nucleophiles may
be expected for simple primary alkyl halides, and the least
effect occur with compounds which react via an SN1 or El
Sediment Matrix Effects
There is considerable interest in possible effects of
solid surfaces on rates of hydrolysis. Most hydrolysis
experiments are performed in simple buffered aqueous
solution. Contaminants in the vadose zone or ground water
have considerable contact with a variety of aquifer
materials which could potentially affect degradation rate.
Hydrolysis reactions may be affected by factors like acid or
base catalysis, sorption and ionic strength. Since
compounds which react by different mechanisms may be
impacted differently by these solid surfaces, both 1-
chloropropane and TCA were used in degradation experiments
performed in various matrices.
Catalysis of hydrolysis or elimination reactions of
alkyl halides by saturated aquifer materials has not been
demonstrated. Because high concentrations of 1,1-DCE have
been observed in Florida and Arizona at solvent spill sites
contaminated with TCA, the role of sand or other materials
which may influence the degradation of TCA was evaluated.
The nonbiological degradation of pesticides in the
unsaturated zone was shown to play an important role for a
few groups of pesticides, mainly organophosphates and s-
triazines. Clay mineral surfaces have shown catalytic
activity, correlated to their acid strength. This catalytic
process is most important at low moisture content, and
therefore is more important in the vadose zone than beneath
the water table (Saltzman and Mingelgrin, 1984).
Haag and Mill (1988) did not observe significant
differences in the kinetics or products of TCA in contact
with sediment pore water. Epoxide hydrolysis was
accelerated by a factor of four in sediment as compared to
rates in buffered water.
Mabey and Mill (1978) indicated that acid promotion of
the aqueous hydrolysis of halogenated aliphatic hydrocarbons
has not been observed. March (1985) stated that gem-
dihalides can be hydrolyzed in water with either acid or
basic catalysis to give aldehydes or ketones, although the
strength of acid was not addressed.
In a review of elimination reactions in the presence of
polar catalysts, Noller and Kladnig (1976) stated that
"interaction of X with an acid is probably as indispensable
as the reaction of H with base in liquid-phase elimination
reactions, but this function is probably taken over by the
solvent and is less pronounced than the base promoted
Clarification of interactions with polar surfaces may
provide insight into possible effects of sediments or soil
on reaction rates. Clays, for example, contain polar
surfaces which have been shown to catalyze degradation of
some pesticides (Saltzman and Mingelgrin, 1984).
Noller and Kladnig (1976) illustrated elimination
reaction products were a function of the specific catalyst
Cl C1 C2 Cl
as reactant. Basic catalysts (e.g., KOH-SiO2) attack the
most acidic H, that at C1, forming more 1,1- than 1,2-
dichloroethene. Acidic catalysts (e.g., silica-alumina)
attack C1 on Cl because the formation of the carbocation is
facilitated by the other Cl on that carbon resulting in the
formation of much more of the 1,2-dichloroethene isomer.
The choice of catalyst will determine the predominant
product giving selectivity to the reaction.
Mochida et al. (1967) reported that the reactivity of
TCA on solid acids was greater than that for other
chlorinated ethanes (mono-, di-, tri- and tetra- chloro
compounds). On solid bases it was less reactive than
penta-, tetra-, and 1,1,2-tri- chloroethanes. The shift in
reactivity of the ethanes with change in catalyst showed
enhanced ability of TCA to form a carbocation by
accelerating the reaction on an acid surface as compared to
the other chlorinated ethanes. There was also the lack of
an acidic beta hydrogen to permit catalysis by base.
Possible catalysis would be compound- and mechanism-
specific. Degradation experiments were performed on 1-
chloropropane (SN2) and TCA (SN1,E1) at 650C in 5 ml
distilled deionized water, with a final concentration of
approximately 2 mg/l. Separate ampules were prepared with
the addition of 0.4 g bentonite clay, 1 g limestone, 1 g
sand, and 0.2 g silica gel.
Similar trends were observed for both compounds (Table
11). The slowest rates relative to water were obtained for
both compounds in the sample containing clay, while the
fastest rates were observed in the sand.
The data generally showed greater variability in the
samples containing the solids as compared to the DW system
(Figures 16 and 17) as evidenced by correlation coefficients
less than 0.99. However, the rates of l-chloropropane
degradation in ampules containing solids differed by less
than 10% of the rate obtained for Milli-Q water.
The relative degradation rates for TCA differed more as
a function of matrix than observed for chloropropane,
however, there was also greater variability as evidenced by
the correlation coefficients. In the case of TCA, the
formation of 1,1-DCE was similar in all matrices suggesting
the ratio of products was not affected by the presence of
The relatively small differences in rates measured in
these matrices may be due to a variety of factors including
sorption, however significant surface catalysis was not
observed. For this type of saturated system, the amount of
alkyl halide in contact with the surface would be small.
Differences may be attributed to normal variability and
differences in ionic strength or composition of the aqueous
phase in contact with the solids.
Table 11. Matrix Effects for Degradation Rates of
1-Chloropropane and 1,1,1-Trichloroethane at 700C.
Linear Regression Output for the Plot of in C (ug/1)
vs. Time (hours).
Regression Output: MQ
Clay Limestone Sand
Std Err of Y Est
No. of Observation
Degrees of Freedom
X Coeff. (Rate) -0.0102 -0.0094 -0.0098 -0.0113
Std Err of Coef. 0.0004 0.0009 0.0004 0.0004
Std Err of Y Est
No. of Observations
Degrees of Freedom
MQ Clay Limestone Silica Gel Sand
X Coeff. (Rate)
Std Err of Coef.
-0.027 -0.020 -0.038
0.0008 0.0009 0.0010
Figure 16. Effect
rate of hydrolytic
40 60 80 100 120 140
of the presence of
degradation of TCA
+ Silica Gel
4 Milli-Q Water
1 + 0
Figure 17. Effect of the presence of solid material on the
rate of hydrolytic degradation of l-chloropropane at 650C.
These experiments do suggest that TCA in sand aquifers
may show a slightly increased rate as compared to low ionic
strength buffered water experiments. The rate coefficient,
however, will fall within the error limits for the rate
estimate for the degradation of TCA based on the experiments
in buffered distilled water.
SOLUBILIZATION AND DEGRADATION OF RESIDUAL TCA
A computational model was constructed to describe the
attentuation of TCA beneath the water table in the presence
of multiple phases. This simplified scenario for a TCA
spill considered the chemical transformation of TCA to 1,1-
DCE along with advective transport resulting from ground
water flow, of TCA and 1,1-DCE out of this zone containing
the residual solvent. Biodegradation of TCA in this highly
contaminated zone was considered negligible.
The major objective in developing this model was to
describe the relative concentrations of the major
constituents and how their concentrations may change with
time. These trends are illustrated for various ground water
flow rates, change in initial concentrations, and initial
Behavior of Residual Solvent
The migration pattern of chlorinated hydrocarbons
following a spill is illustrated in Figure 18. These dense
nonaqueous phase liquids (NAPL) will infiltrate the porous
media, with some of the NAPL retained in residual
concentration. The retention capacity for these NAPL in the
unsaturated zone may range from 5 L m-3 (approximately 12
mL/L of pore space) in highly permeable media to 30-50 L m-3
.. ..... ................................................................
G r o u n d ... ..... ............................................
W a te r .........................
----...::.---- II ....i....iKE.....uo4 S
S.of.. e:.Pore.sp e...:.....
::::::::::::::::: ::: :: ::::: :::::::::: :::::::::::::::::::::::T C A
:- : --" ' "," 1 D' '
W at.................................................................a t e r
Chlorinated Solvent Pool
at impermeable layer
Figure 18. Equilibrium model for the attenuation of
residual TCA present beneath the water table.
in media of lower permeability (Schwille, 1984). Additional
factors which influence whether the NAPL will reach the
water table include the spilled volume and infiltration
If sufficient volume of dense NAPL reach the water
table, it will sink into the saturated zone and continue to
migrate downward as long as the retention capacity of the
zone is exceeded. Wilson and Conrad (1985) reported
residual hydrocarbon occupying 15-40% of the pore space in
the saturated zone.
Water continues to flow laterally through the water
saturated zone containing residual NAPL. The globules of
NAPL provide a large interface with the water providing a
solution zone, where the initial concentration of a given
component is proportional to its aqueous solubility as
determined by the NAPL composition. These globules are
generally trapped in the larger pore spaces and are being
prevented from entering the smaller pores due to the high
capillary entrance pressure. There is a reduction in
permeability to water where the residual NAPL is present, as
the largest channels become blocked at several places by
discontinuous solvent ganglia. This forces water to flow
around the solvent in fairly thin films and/or be diverted
into the smaller channels whose carrying capacity
(conductivity) is low (Jones, 1985).
In laboratory experiments, the initial concentration of
chlorinated solvent was at saturation concentration even
when the layer of sand with residual solvent was thin
(Schwille, 1988). The concentration gradually decreased
until the levels in the water were sufficiently low that
further removal of solvent was slow. At this point
approximately 86% of the residual had been removed.
My model was developed assuming that equilibrium
saturation was maintained, the dissolution of residual
solvent being faster than the degradation or advective
transport of components. Diffusion or hydrodynamic
dispersion was not considered to be a limiting factor in
maintaining equilibrium. The solvent-contaminated zone was
then treated similar to a well-mixed flow reactor.
Interactions of the solutes in the water with the solid
matrix of the saturated zone were considered minimal
providing residual solvent was present; the porous medium
was assumed to provide a matrix in which the residual
solvent was retained.
Once the flow of the NAPL stopped, the subsequent
losses were assumed to occur through degradation or
advection of the compound in the aqueous phase.
Hydrolysis/elimination of TCA occurs much faster in dilute
aqueous solution than would occur for water dissolved in the
TCA solvent phase (Walraevens et al., 1974). Ground water
continues to flow through this zone, although at somewhat
reduced velocities, carrying dissolved components out of
Aqueous Phase Concentrations
The quantity of solvent lost each day by advection or
degradation is a function of the concentration of each
component (TCA and 1,1-DCE) in the aqueous phase, which in
turn depends on the composition of the residual NAPL. The
solvent phase may contain TCA and/or 1,1-DCE, or another
solvent which may have been spilled with the TCA.
The distribution of a component between the two liquid
phases can be expressed in terms of fugacity. For ideal
mixtures, the solubility of the solute at any composition is
estimated by multiplying the unit solubility by the mole
fraction of the component in the solvent phase at
equilibrium. Nonideal mixtures form deviations from
linearity. Estimates of aqueous concentrations resulting
from a nonideal solvent mixture requires knowledge of the
activity coefficients at the various mole fraction
compositions. For the simpler ideal case,
[TCA]w x STCA
[DCE]w (l-x) SDCE
x TCAs / (TCAs + DCEs)
where TCAs and DCEs are the number of moles of that compound
in the solvent phase at equilibrium, x is the mole fraction
of TCA in the solvent phase, STCA and SDCE are the pure
component solubilities, and [TCA]w and [DCE]w are the
aqueous phase concentrations at equilibrium. The total
number of moles of TCA in a unit volume of porous media is
the sum of the moles present in the aqueous and solvent
The model describes changes for TCA spilled on a high
permeability material like sand. As TCA degrades and forms
1,1-DCE, the degradation product partitions into the NAPL
affecting the aqueous phase concentration of TCA (and DCE).
Both the individual solubilities and the solubility of
a mixture of TCA and 1,1-DCE are required in the model and
it was also necessary to assess if mixtures of TCA and 1,1-
DCE deviate significantly from ideality. Literature values
for the solubilities of these constituents vary widely
(Table 12). The solubility data for 1,1-DCE reported by
Lyman (1981), showed as much as a 700% error from a
predicted concentration based on regression relationships.
That estimated concentration is much closer to the
concentrations reported by Verschueren (1977).
Table 12. Solubilities of TCA and 1,1-DCE (mg/L)
Temp (0C TCA 1.1-DCE Source
20 480 400 Pearson and McConnell (1975)
20 4400 2640 Verschueren (1977)
30 1088 3675 Verschueren (1977)
25 273 Lyman (1982)
4 1700 4200 This study.
24 1580 3200 This study.
Measurements (Figure 19) were made on the solubility of
the individual components (TCA and 1,1-DCE) and on the
3000- + + DCE
0 0.2 0.4 0.6 0.8 1.0
Equilibrium Mole Fraction of TCA
Figure 19. Aqueous solubilities of a binary mixture of TCA and 1,1-DCE as a function of
mole fraction composition in the solvent phase (24 C).
solubility of each with varying compositions of the binary
mixture. Mixtures were at room temperature, approximately
The pure component solubility of TCA (1580 mg/L or 11.8
mmoles/L) and the solubility of 1,1-DCE in the aqueous phase
(3200 mg/L or 33 mmoles/L) measured at 24C were within the
concentration range listed by Verscheuren (1977) who
reported solubilities at 20 and 300C. This is significantly
higher than solubilities reported by Lyman (1981) and
Pearson and McConnell (1975). The solubility for 1,1-DCE
reported in this dissertation was verified independently by
solubility measurements performed using high performance
liquid chromatography (HPLC) (Linda Lee, University of
Florida, Personal communication, 1988). She measured an
average for the solubility of 1,1-DCE at 24C as 2990 mg/L.
Her report is included in Appendix B.
Verscheuren (1977) reported that the solubility of TCA
at 200C was four times greater than at 3000, a value
approximately three times greater than our result at 240C.
Since the mass lost per unit time from degradation is a
function of aqueous concentration and the first-order
degradation rate coefficient, higher aqueous concentrations
at lower temperatures could compensate for the lower
degradation rate. The solubility of TCA at 40C was measured
to verify this trend. As shown in Table 12, a significant
increase in solubility of TCA at lower temperatures was not
The linearity of the change in solubility with
increasing mole fraction for these two compounds suggested
that 1,1-DCE and TCA form a near-ideal solution in the
solvent phase. Based on these measured data, I assumed that
mole fraction in the solvent phase multiplied by the aqueous
solubility of the pure compound provided a reasonable
estimate of aqueous phase concentration of TCA and 1,1-DCE.
Loss of TCA from this hypothetical contaminated zone
occurs via advection and degradation, both of which are a
function of the aqueous phase concentration. The relative
importance of these two mechanisms is a function of the flow
velocity advectionn) and the temperature (solubility and
degradation rate). Observations of selected field data
suggest higher concentrations of 1,1-DCE appear in southern
state aquifers where the ground water temperatures are
higher. The model therefore, assigns a temperature of 250C.
The volume of water exchanged through the contaminated
zone is a function of the ground water flow velocity and the
length of the contaminated zone. Fresh water upgradient of
the spill enters the contaminated zone while an equal volume
of water at equilibrium saturation of the contaminants is
displaced. Velocities for the model are expressed as the
per cent of the volume of contaminated water exchanged per
day. These values include the "no flow" or "low flow" (0.1%
per day) cases, in which the dominant loss occurs through
degradation. At 0.25% per day, the rate of advection is
comparable to the rate of degradation. Finally, a flow rate
of 0.5% per day represented the case in which the loss of
TCA is primarily due to advection. At flows greater than
0.5% per day the losses would be dominated by the advective
term. These volume exchange rates represent slow flows
and/or very large spill areas. An exchange of 0.5% per day
represents an approximate flow through 5 meters of
contaminated porous media at a rate of 2.5 cm/day.
The solubility of TCA affects not only its rate of
advection from the contaminated zone, but also the total
mass of TCA degraded per unit time. The first-order rate
constant at 250C is approximately 0.00226 day-1 as measured
in this study. In a contaminant plume, the half-life for
the degradation of TCA is approximately 10.2 months.
Although the first-order rate coefficient remains constant,
the mass of TCA converted per unit time decreases as the
concentration of TCA in the aqueous phase decreases.
In the model, it was assumed that the TCA concentration
remained at saturation within the zone containing residual
solvent since the TCA that degraded was replaced by
dissolution of the residual solvent. The amount of TCA
degraded per unit time follows zero-order kinetics. The
zero-order rate equals the mass converted per unit time in
the first-order equation as the time increment approaches
zero. This becomes 0.00226 day-1 multiplied by the aqueous
concentration of TCA. A 50% decrease in the solubility
would therefore result in a corresponding 50% decrease in
the mass of TCA degraded per unit time.
Model Parameters and Procedures
Initial conditions for the model include a unit volume
of water (1 liter) in contact with 100 mmoles of TCA. After
equilibrium 11.8 mmoles of TCA will be in the aqueous phase
leaving 88.2 mmoles (approximately 11.8 grams or 8.5 mL) in
the residual solvent phase. The changes in concentration of
TCA or 1,1-DCE in this unit volume are displayed graphically
illustrating the effects of different flow rates, higher
initial mass of TCA, and effect of the presence of an inert
solvent mixed in the residual phase.
Iterative calculations (Appendix C) are made in the
model for advection and degradation in relatively small time
increments, with subsequent reequilibration of the solvent
remaining in the zone of residual contamination. The
residual solvent mass will continue to decrease until at
some point a separate solvent phase does not exist.
Calculations become more difficult (smaller time increments
must be used to attain convergence of the iterative
mathematical solution) and other factors would become more
important as the NAPL is depleted. Therefore, the
calculations are stopped when amounts of TCA in the residual
NAPL are less than 10 mmoles. At lower levels of residual
NAPL, the process may become diffusion limited as the NAPL
is trapped in regions of the soil matrix removed from the
aqueous flow. The results of the model are shown in Figures
The total mass of TCA in the NAPL showed zero-order
decay with flows from 0.1-0.5% per day (Figure 20). As the
flow rate decreases, slight nonlinearity is observed. This
reflects the slow accumulation of 1,1-DCE in the solvent
phase which begins to decrease the aqueous concentration of
The decrease in aqueous concentration of TCA (Figure
21) as the total mass of TCA in the system goes from 100
mmoles to approximately 15 mmoles (slightly in excess of the
solubility) is dependant on the flow. The larger decrease
is observed for the case of no-flow, which results in a 45%
decrease in the aqueous phase concentration after 10 years.
The major reason 1,1-DCE fails to accumulate
significantly in the solvent phase is its higher water
solubility. Having a solubility twice that of TCA, 1,1-DCE
is advected from the zone containing residual solvent more
readily. In the special case of no flow through the system,
1,1-DCE is not advected and begins to accumulate in the
solvent phase affecting the aqueous phase concentration of
TCA. However, since only approximately 20% of the TCA is
0 2 4 6 8 10
Figure 20. Model results: Decrease in total TCA mass in
the residual zone as a function of flow.
Figure 21. Model results: Change in aqueous concentration
of TCA as a function of flow.
converted to 1,1-DCE, the effect of the accumulation is not
observed until substantial degradation has occurred. If all
the TCA degraded in this closed system, 20 mmoles of 1,1-DCE
would be produced, which is 60% of the pure component
aqueous solubility of 1,1-DCE. Therefore, for the initial
conditions of the model, a residual NAPL will exist only
when excess TCA is present.
A comparison of different initial conditions for a
constant flow (0.25%) is shown in Figure 22. With an
increase in amount of residual TCA, the same zero-order
decay rate is observed, indicating that doubling the amount
of TCA in the solvent phase doubles the time needed for
removal of the residual.
In addition, Figure 22 illustrates the rate of loss of
TCA when the initial 100 mmoles is mixed with another
solvent, a hypothetical mixture in which the mole fraction
of the "inert" compound remains at 0.5 in the solvent phase.
This represents a case where a compound with solubility
similar to TCA (like TCE) is present in the residual. The
presence of this other compound causes a 50% reduction in
the aqueous phase concentration of TCA, and therefore the
rate of loss of TCA, doubling the time to remove the TCA
from the residual phase.
The patterns of change in mass of 1,1-DCE in the
solvent or aqueous phase over time are more complex when
there is advection from the system Figure 23. The aqueous
YEARS (Flow, 0.25%/Day)
Figure 22. Model results: Change in total mass of TCA as a
function of initial mass of TCA and composition of the
YEARS (Flow, 0.25%/Day)
Figure 23. Model results: Pattern of 1,1-DCE formation and
advection as 100 mmoles of TCA in the residual zone
concentration of 1,1-DCE continues to increase for some time
as the mass of 1,1-DCE in the solvent phase begins to
decrease because its mole fraction continues to increase in
the solvent phase.
The total mass of 1,1-DCE in the zone of residual
contamination increased over time, reaching a maximum as the
TCA mass in the solvent phase approached zero. Increasing
the flow rate not only shortened the time in which 1,1-DCE
was accumulating, but decreased the maximum amount of 1,1-
DCE present in that zone. This is true for the aqueous
phase concentrations (Figure 24) and amount in the solvent
phase (Figure 25). The maximum concentration of 1,1-DCE in
the aqueous phase for a flow of 0.5% per day is
approximately 1 mmole/L (100 mg/L) at the point where some
residual phase is still present. The concentration of TCA
at that time is nearly at saturation (approximately 1500
The changes in aqueous concentration of 1,1-DCE for
larger amounts of TCA originally present or in the presence
of an inert solvent as previously discussed, are shown in
Figure 26. The changes in the amount of 1,1-DCE in the
solvent phase is shown in Figure 27. The inert solvent
increases partitioning into the organic phase, keeping the
aqueous concentration low.
The model illustrates factors which affect the time for
removal of a residual phase under varying conditions, and
0 2 4 6 8
Figure 24. Model results: Increase in aqueous
concentration of 1,1-DCE forming from degradation
a function of flow.
of TCA as
Figure 25. Model results: Pattern of accumulation of 1,1-
DCE in the solvent phase as TCA degrades.
0 2 4 6
YEARS (Flow, 0.25%/Day)
Figure 26. Model results: Change in aqueous concentration
of 1,1-DCE as a function of initial mass of TCA and
composition of the solvent phase.
---- 100 mmole
-- 200 mmole and
100 mmole TCA +
inert organic solvent
0 2 4 6 8 10
Figure 27. Model results: Change in total mass of 1,1-DCE
in the residual zone as a function of initial mass of TCA
and composition of the solvent phase.
the different concentrations of 1,1-DCE which would result.
Given a constant initial mass of TCA, the maximum
concentration of 1,1-DCE in the aqueous phase occurs at the
lowest flow rates. For flow rates higher than the 0.5%
volume exchange per day the advective term is dominant and
concentrations of 1,1-DCE in the residual zone remain
As long as a residual NAPL is present, aqueous
concentrations are dominated by TCA. Equal concentrations
of TCA and 1,1-DCE in the water from monitoring well data
from various sites would occur according to the model
primarily in the plume of dissolved constituents
downgradient from the residual zone, or in the original
spill area after all residual solvent was dissolved or
degraded. The presence of a low solubility compound in the
solvent phase with the TCA will considerably slow TCA rate
of advection and degradation.
First-order degradation will continue in the ground
water plume downgradient from the source and this process
could be modeled (Kinzelbach, 1985). Evidence of the
formation of 1,1-DCE would support the assignment of a
degradation rate. Assuming similar retardation factors for
TCA and 1,1-DCE, equal concentrations of TCA and 1,1-DCE
would occur after approximately 3 half-lives, approximately
2.5 years at 250C.