BARR for Decontamination of Soil and Groundwater  Volume 2

This is the TECHNICAL APPENDIX

Copyright 1993 by Larry Dieterich
All Rights Reserved
Duplication and distribution is permitted, but credit must be given to the
author and the document must be distributed in its entire form. No
fragmentation, editing or deletions are permitted on copies for
distribution.

This document is part of a 3 volume set.

Volume 1 BARR report
Volume 2 (this document) BARR Technical Appendix
Volume 3 BARR Bibliography

The BARR process was developed by Larry Dieterich

Larry Dieterich
405 E 7th Street
Davis, California 95616
USA
voice/fax (916) 758-9260
Internet email- [email protected]

This edition of the BARR reports is made available in ASCII format for
distribution over the Internet.

Paper, disk, or formatted Macintosh files with graphics are available upon
request. Contact the author at the above address.


BARR: For Decontamination of Soil and Groundwater

       B                            Bio-
       A                   Anaerobic
       R               Reduction &
       R       Re-oxidation

Bio-Anaerobic Reduction & Re-oxidation

BARR: For Decontamination of Soil and Groundwater

TECHNICAL APPENDIX

BARR  TECHNICAL APPENDIX CONTENTS


TRANSFORMATION PROCESSES / CHEMODYNAMICS
BIOTIC DEGRADATION
ABIOTIC DEGRADATION
SORPTION
BONDING FORCES
FREE ENERGY
INTERFACIAL TENSION
SALT EFFECTS
COSOLVENCY
pH
REDOX
MICELLES
PRECIPITATION
COLLOIDS
CLAY FRACTION
ORGANIC MATTER PROPERTIES
ORGANIC MATTER AND ORGANIC POLLUTANT INTERACTIONS
HUMIC/MINERAL ASSOCIATIONS
COLLOID STABILITY
BIOCOLLOIDS
COLLOIDS IN GROUNDWATER
COLLOIDAL METAL BEHAVIOR
COLLOIDAL BEHAVIOR OF RADIONUCLIDES
COLLOID SUMMARY


TRANSFORMATION PROCESSES / CHEMODYNAMICS



Transformation occurs when a thermodynamically favorable reaction occurs.
Transformation denotes a change in the target chemical. Whether by adding
or removing a substituent group, re-arranging, breaking or forming bonds.
Transformation is not necessarily the same as degradation (although it may
be), but transformation is often a step toward degradation as it
represents a modification of the target chemical as a step toward its
ultimate mineralization.


BIOTIC DEGRADATION

Biological Transformations
Basically 5 processes are involved in microbial transformation:

1. Biodegradation- The contaminant serves as a substrate for growth.

2. Cometabolism- Material is transformed by metabolic reactions but does
not serve as an energy source.

3. Conjugation- The contaminant molecule is joined to another molecule in
the system.

4. Accumulation- In which the material partitions (is incorporated into)
into the organism

5. Secondary effects of microbial activities-In which the pollutant is
transformed because of changes in pH, redox conditions, reactive products,
etc., brought about by microorganisms.

Enzymes in soil can be separated into a number of categories according to
their location within the soil microenvironment. Indeed, the measured
activity of a particular enzyme is usually a composite of activities
belonging to two or more categories:

1. Enzymes associated with living, metabolically active cells.

2. Enzymes associated with viable but nonproliferating cells such as
resting vegetative cells, bacterial endospores, fungal spores, and
protozoan cysts.

3. Enzymes which are associated with, at least briefly, with their
substrates in enzyme-substrate complexes.

4. Enzymes attached to entire dead cells, cell debris or having diffused
away from dead and lysed cells. Many enzymes in this category may have had
an original functional location on or within a cell yet may survive for a
short period when released into the soil aqueous phase.

5. Enzymes which are more or less permanently immobilized on the soil clay
and humic colloids. Expandable clays have a high affinity for enzymes
although this is not always synonymous with the retention of catalytic
ability. Enzymes associated with humates retain their activity for long
periods.

The ubiquitous presence of biological processes and products acts to
influence the overall system. Enzymes, both intact and extracellular, will
catalyze many reactions. Biological lipid membranes will provide a
nonaqueous, nonpolar interface in the aqueous system. This has special
significance for nonpolar pollutants, such as common petrochemicals.

The ability of microorganisms to catalyze transformation of organic
chemicals is considered prerequisite information. There are many types of
microorganisms: fungi, bacteria, actinomycetes, viruses, viroids,
mycoplasmsa-like organisms (MLO's), slime molds protozoans, rotifers,
algae, ... the list is actually endless, both because of the widespread
disagreement between workers in the field as to the definitions and
distinctions between organisms and the fact that the organisms themselves
are continuously c hanging and evolving and mutating. Many microbes have
not been discovered and described and the organisms that have been
described are continuously subject to reclassification as natural
processes of growth and evolution occur.

Regarding the BARR process, the most important microbes appear to be the
bacteria. Bacteria are small, (0.5 - 3 micrometer), unicellular organisms.
Bacteria are the smallest free-living microbes. They are small compared to
the pore sizes in most nonindura ted geological materials and they are
large in relation to the size of hydrated inorganic ions and molecules.

Bacteria have been shown to effect any number of chemical transformations
in both mineral and organic materials.

They multiply by binary fission. (1,2,4,8,16,32,64,128... 2n, n being the
number of divisions) and are capable of extremely large populations in
very short time periods under favorable conditions.

Bacteria catalyze nearly all the important redox reactions that occur in
groundwater. This means that although the reactions are spontaneous
thermodynamically, they require the catalyzing effect of microorganisms
such as bacteria in order to proceed at a significant rate. The catalytic
capability of bacteria is produced by the activity of enzymes that
normally occur within the bacteria. Enzymes are protein substances formed
by living organisms that have the power to increase the rate of redox
reactions by decreasing the activation energies of the reactions. They
accomplish this by interacting strongly with complex molecules
representing molecular structures halfway between the reactant and the
product.

The local molecular environment of many enzyme reactions is very different
from the bulk environment of the aqueous system.  Bacteria and their
enzymes are involved in redox processes in order to acquire energy for
synthesis of new cells and maintenance o f old cells. Some of the energy
obtained from the redox reactions is maintenance energy required by
bacterial cells for such things as mobility, to prevent an undesirable
flow of solutes either into or out of the cell, or for resysnthesis of
proteins that are constantly degrading. For bacteria to be able to make
use of an energy yield from a redox reaction, a minimum free energy change
of approximately 60 kJ/mol between the reactants and the products is
required. The main source of energy for bacteria in the groundwater zone
is the oxidation of organic matter.

Bacteria that can thrive only in the presence of oxygen are called aerobic
bacteria.  Anaerobic bacteria require an absence of dissolved oxygen.
Facultative bacteria can thrive with or without oxygen. The lower limit of
dissolved oxygen for the existence of most aerobic bacteria is considered
to be about 0.05 mg/l, but some aerobic species can persist at lower
levels. Since most methods commonly used for measuring dissolved O2 have a
lower detection limit of about 0.1 mg/l, it is possible that aerobic bac
teria can mediate redox reactions in situations that might appear to be
anaerobic based on the lack of detectable oxygen. Bacteria of different
varieties can withstand fluid pressures of many hundreds of bars, pH
conditions from 1 to 10, temperatures from near 0 C to greater than 75 C,
and salinities much higher than that of seawater. They can migrate through
porous geological materials and in unfavorable environments can evolve
into resistant bodies that may be activated at a later time.

In spite of these apparent characteristics of hardiness, there are many
subsurface environments in which organic matter is not being oxidized at
an appreciable rate. As a result, the redox conditions have not declined
to low levels even though hundreds or thousands of years or more have been
available for the reactions to proceed.

Acclimation Period
Biodegradation of organic compounds in the environment is often preceded
by an extended acclimation period. The acclimation period is taken to mean
the time interval during which biodegradation is not detected, but the
pollutant has been introduced. This period may be significant because a
compound may pass through treatment systems and deeper into the
environment during this time. There are several reasons offered for this
lag time:

(1) The time is needed for small populations to become sufficiently lar ge
to give detectable loss of chemical.

(2) Adverse environmental conditions at the site where the chemical is
discharged.

(3) Rarity of microbes able to degrade specific chemicals.

(4) Presence of compounds inhibiting microbes. The inhibitor need not be
another compound, but the substrate itself, because the concentrations of
many chemicals introduced into the environment are probably high enough to
suppress the microbes having the ability to degrade those compounds. Other
evidence exists that one compou nd may shorten the acclimation period
needed before another is degraded.

(5) An acclimation period may result from the time required for the
appearance of a new genotype after a mutation or genetic exchange, which
presumably occurs during the time microor ganisms are exposed to the
compound. The new organism then grows and mineralizes the compound.
Mutations and gene transfers are rare events, however and appear in only a
small and random percentage of samples.

The acclimation periods prior to detectable dehalogenation of halogenated
benzoates in anaerobic lake sediments ranged from 3 weeks to 6 months.
These acclimation periods are reproducible over time and among sampling
sites and were characteristic of the chemical tested. (LINKFIELD et al.
1989). The lengthy acclimation period appears to represent an induction
phase in which little or no aryl dehalogenation is observed, followed by
an exponential increase in activity typical of an enrichment response. Con
tinuous growth from the time of the first exposure to the chemical is
inconsistent with the extremely low per-cell activities estimated for the
early days of the acclimation period and the fact that the dehalogenation
yields no carbon to support microbial growth.

The finding of a characteristic acclimation time for each chemical argues
against nutritional deficiency, inhibition or predation as an explanation
for this phase of metabolism, while the reproducibility of the findings
with time and space and among replicates argues against genetic changes
as the explanation. The acclimation times correlate with the eventual
dehalogenation rates. This may reflect the general energy limitations in
the anaerobic communities and suggests that those chemicals with faster
deh alogenation rates provide more energy for the induction and growth
phases of the active population.

In aerobic microbial communities the acclimation periods for xenobiotic
compounds typically range from several hours to several days, but for
anaerobic communities these periods are much longer, often from 2 weeks to
6 months or longer.

The concentration and structure of the xenobiotic compound itself probably
influences the acclimation period. In particular, low concentrations and
structure-biodegradability relationships are known to affect aerobic
biodegradation rates, and in some cases they have been shown to affect
the acclimation period as well. The hypotheses posed for aerobic
acclimation periods should be equally valid for the anaerobic ones

Acclimation periods are a function of chemical structure; halogen type and
other ring substituents alter the acclimation period. In the case of the
bromobenzoates, the effect of ring position is clearly demonstrated: the
para isomer exhibits greater acclimation periods than either the ortho or
the meta form.

Increased acclimation periods as a function of high xenobiotic
concentration have been reported for the anaerobic microbial metabolism of
one and two carbon halogenated solvents. The acclimation period observed
in literature reports may be due to either a "true" acclimation i.e., a
period of no biodegradation followed by initiation and acceleration of
degradation, or an "apparent" acclimation, in which biodegradation would
proceed from the time of addition but at rates so slow as to be
nondetectable.

A true acclimation could be caused by the time required for genetic
changes, induction of new protein synthesis, or exhaustion of preferential
substrates.

An apparent acclimation would probably be the case if the initial
population was very small but grew continuously, until, at some point, it
was large enough to result in detectable biodegradation.

Lag periods are reportedly longer for samples collected from field sites
that were low in dissolved inorganic nutrients (N,P). Lag periods
decreased with fertilization with N & P. (LEWIS et al., 1986)

If the redox reactions that require bacterial catalysis are not occurring
at significant rates, a lack of one or more essential nutrients for
bacterial growth is the likely cause. There are various types of
nutrients. Some are required for the incorporation into the cellular mass
of the bacteria. Carbon, nitrogen, sulfur, and phosphorous compounds and
many metals are in this category. Other nutrients are substances that
function as electron donors and energy sources, such as water, ammonia,
glucose, and H2S, and substances that function as electron acceptors,
such as oxygen, nitrate, and sulfate.

Aerobic Metabolism
Aerobic transformation is generally considered to occur in the presence of
oxygen. Oxidizing agents are the chemicals that provide the terminal
electron acceptors in the reaction. These can be any material capable of
donating electrons. Oxygen and hydrogen peroxide are commonly used.

The tendency for reduced hydrocarbons to become oxidized provides the
basis for most of the treatment measures applied to cleanup of deep ground
contamination.

There are a number of limitations and difficulties in the creation and
maintenance of oxidizing conditions for the degradation of chemicals in
the subsurface. Lowering the water table or other measures to maintain
oxygen supply to a contaminated formation is commonly reported as a
remedial measure.

There are reports of more or less success with the oxidative
biodegradation of subsurface contamination. It is frequently found that
the limiting factor in such treatments is the availability of a reduced
carbon source to sustain the levels of microbes av ailable to degrade the
pollutant of concern. The degradation rate declines markedly at about 60%
recovery (BOETHING and ALEXANDER, 1979)

Concentration of the compound may be a significant factor affecting its
susceptibility to microbial attack, and organic chemicals may persist in
some environments as a result of low prevailing concentration or low
solubility in water.  Both of these limitations are removed by the BARR
process.

Anaerobic Transformations
Under certain hydrologic conditions, such as infiltration of
precipitation, anaerobic zones in shallow aquifers may become aerobic as a
result of the rise or decline of the water table, as influenced by the
capillary fringe. Because of possible alternating oxidizing and reducing
conditions, it is reasonable to assume that both aerobic and anaerobic
microbes in these zones of contaminated aquifers have adapted enzyme
systems capable of metabolizing organic compounds, thus influencing their
fate and transport in the subsurface. Under aerobic conditions, molecular
oxygen is required as a terminal electron acceptor during respiration and
also for hydroxylation of aromatic compounds before ring cleavage. These
reactions are mediated by oxygenases. Under anaero bic conditions,
microbial activities tend to attack complex aromatic structures at
substituent groups and convert the parent substrates to hydroxylated,
carboxylated, or amino derivatives. (KUHN et al. 1989). However, in the
absence of molecular oxygen, aromatic compounds may be degraded by
methanogenic consortia in the contaminated aquifer or in the presence of
other electron acceptors such as nitrate of sulfate. It is reasonable that
water can be the source of oxygen for the hydroxylation of
petroleum-derived groundwater contaminants under sufficiently reducing
conditions.

In the absence of molecular oxygen, the catabolism of organic matter by
micro-organisms can be achieved either by fermentation or by anaerobic
respiration. The formal definition of fermentation is that organic
substrates act both as electron donors and acceptors and may concern
large polymers such as starch and cellulose. Fermentative processes do not
require exogenous electron acceptors. The organic substrate plays the role
of an internal electron donor (reducing agent) and one of the metabolites
produced acts as an electron acceptor (oxidizing agent). The organic
substrate is not fully oxidized to CO2 in fermentative processes which
accounts for the build-up of partially reduced compounds (alcohols,
organic acids, etc.) The metabolism of organic substrates by anaerobic
respiration involves inorganic electron acceptors, such as nitrate,
sulfate and CO2 . Such metabolism generally uses small molecules produced
by fermentative metabolism or by the aerobic biodegradation which takes
place in the nearby oxic environment.

Bacteria which use nitrate as an electron acceptor are called
nitrate-reducing bacteria and the process is known as nitrate respiration.
When nitrate is completely reduced to gaseous products, the process is
called denitrification. Many bacteria can grow aerobically with oxygen and
anaerobically with nitrate. These organisms are commonly called
"facultative" organisms.

The strictly anaerobic bacteria which use sulfate as an electron acceptor
form the ecophysiological group of sulfate reducing bacteria. They produce
sulfide as the end product of the dissimilative sulfate reduction.

Carbon dioxide serves as electron acceptor for the strictly anaerobic
methanogenic bacteria which reduce it to methane. Under some
circumstances, methane production may arise from the reduction of some
organic compounds.

All of the above bacterial groups (denitrifiers, sulfate reducers,
methanogens) are able to mineralize organic matter to CO2 . Thus they are
considered as the terminal mineralizers of organic matter in anoxic
environments.

During methanogenesis in anaerobic environments, degradable organic matter
is converted to CH4 and CO2 . During this anaerobic process, about 90% of
the available energy is retained in the methane produced, with a
relatively low yield of microbial cells. (MACKIE and BRYANT 1990). The
relatively small release of energy must be divided and distributed among
the different bacteria involved in the sequential process of anaerobic
degradation. Most of the information on anaerobic degradation comes from
studies on ruminal fermentation, bacterial enrichment cultures and
fermentation of sewage sludge.

The basic difference between aerobic and anaerobic oxidation is that in
the aerobic system, oxygen is the ultimate hydrogen acceptor with a large
release of energy, but in anaerobic systems the ultimate hydrogen acceptor
may be nitrate, sulfate or an organic compound with a much lower release
of energy.

The process of anaerobic decomposition of organic material involves
discrete stages. Briefly, insoluble organics (via hydrolytic bacteria)
give soluble organics (via acid-forming bacteria) give volatile acids (via
methanogenic bacteria) give gases.

Although bacteria are the major group of microbes involved in anaerobic
metabolism, fermentative ciliate and flagellate protozoa and some
anaerobic fungi also occur.

It is convenient to think of these stages as different trophic levels, and
although all three stages are normally occurring simultaneously within an
active system, the micro-organisms involved at each stage are
metabolically dependent on each other for survival. For example, the
methanogenic bacteria require the catabolized end products of the acid
forming bacteria.

The first stage is the hydrolysis of high molecular weight carbohydrates,
fats and proteins that are often insoluble, by enzymatic action into
soluble polymers. In the first stage, the major substrates in the organic
material are hydrolyzed to basic components; proteins to amino acids,
fats to glycerol and long-chain fatty acids, and polysaccharides to mono
and disaccharides. Proteins are hydrolyzed to smaller units such as
polypeptides, oligopeptides, or amino acids by extracellular enzymes
called proteases, which are produced by only a small proportion of the
bacteria. The majority of bacteria are able to utilize these smaller
peptides or the amino acids, which pass through the cell wall and are
broken down intracellularly.  Little is known about the lipolytic
bacteria even though they have been shown to be highly effective in
anaerobic digesters. They are present in densities of up to 700000/ml and
the addition of vegetable oil to digesters to enhance gas production is
commonly practiced in some countries (GRAY 1989).

The second stage involves the acid-forming bacteria which convert the
soluble polymers into a range of organic acids (acetic, butyric and
propionic acids), alcohols, hydrogen and carbon dioxide. Acetic acid,
hydrogen and carbon dioxide are the only end-products of the acid
production that can be converted directly into methane by methanogenic
bacteria. The heterogeneous group of facultative and anaerobic bacteria,
which are responsible for hydrolysis are also responsible for acid
formation. In this second stage, the hydrolyzed substrate is converted to
organic acids and alcohols, with new cells also being produced. Various
biochemical pathways are utilized, including fermentation and
beta-oxidation. There is very little stabilization of the substrate in
terms of BOD or COD removal, with the products of acid fermentation being
large organic molecules. Mono and disaccharides, long-chain fatty acids,
glycerol, amino acids, and short-chain peptides provide the main carbon
source for growth, with saturated fatty acids, carbon dioxide and ammonia
being the main end-products. Alcohols, aldehydes and ketones are also
produced.

The third stage is when the organic acids and alcohols are converted to
acetic acid by acetogenic bacteria.

It is in the fourth and final stage, when methanogenic bacteria convert
the acetic acid to methane.

Methanogens are unusual in that they are composed of many species with
very different cell morphology. They require a strict anaerobic
environment for growth with a redox potential below -300 mV (GRAY 1989).
They have simple nutritional requirements; CO2 , NH3 and sulfide. Ammonia
is the essential nitrogen source for growth and no methanogen species are
known to utilize amino acids or peptides. In the overall anaerobic
fermentation of carbohydrate to CO2 and CH4, equal volumes are produced.
The carbon dioxide evolved partially escapes as a gas because it is
soluble in water. It will react with any OH- ions to form bicarbonate.

During methanogenesis in anaerobic environments, degradable organic matter
is converted to CH4 and CO2 . During this anaerobic process, about 90% of
the available energy is retained in the methane produced, with a
relatively low yield of microbial cells.  The relatively small release of
energy must be divided and distributed among the different bacteria
involved in the sequential process of anaerobic degradation.

The bacteria responsible for methanogenesis are similar in different
environments, although little is known about them because of the problems
with isolating and maintaining cultures of bacteria under anaerobic
conditions.  Most of the information on anaerobic degradation comes from
studies on ruminal fermentation, bacterial enrichment cultures and
fermentation of sewage sludge.

Anaerobic digestion and methane production are not unique to anaerobic
digesters, they occur in natural environments including the digestive
tract of most animals, in the sediments of lakes and rivers, and in
estuaries, swamps, marshes and bogs.

In anaerobic subsurface sediments and aquifers, chlorinated alkenes may be
converted by reductive dehalogenation to dichloroethylene and ultimately
to more potent carcinogens such as vinyl chloride. Transformations may be
the result of the microbes actually using the contaminant as a substrate
for growth (direct metabolism) or the transformation may be the result of
some metabolic process which yields a product that cannot be utilized as a
substrate for growth (cometabolism) There may be degradation of a
contaminant due to an enzyme of broad activity that is able to make some
cleavage of a contaminant that results on a non-food end product.

The anaerobic consortium is credited with degradation of a wide variety of
organic chemicals.

Under anaerobic conditions, microbial activities tend to attack complex
aromatic structures at substituent groups and convert the parent
substrates to hydroxylated, carboxylated, or amino derivatives. (KUHN et
al 1989).


Reductive Dehalogenation
Chlorinated compounds are the most extensively studied because of the
highly publicized problems associated with DDT, other pesticides and
numerous chlorinated solvents. Most of the information available on the
biodegradation of chlorinated compounds is on oxidative degradation, since
aerobic culture techniques are relatively simple compared with anaerobic
culture methods. For convenience, the chlorinated hydrocarbons degraded by
microorganisms (bacteria and fungi) are grouped in to three classes (i)
aliphatic, (ii) polycyclic, and (iii) aromatic. Pathways have been
elucidated for a number of degradation processes of these classes of
compounds. A number of organisms have been isolated and, in many cases,
plasmids linked to the metabolic capability have been identified. The
objective being to introduce degradative ability into organisms for use in
bioremedial schemes. To establish the potential applications of the
recombinant strains in the environment, the strains must be stable members
of the indigeno us microflora and the recruitment of catabolic enzymes and
gene regulators with appropriate effector specificities (by natural gene
transfer or by laboratory manipulation) to create new hybrid pathways for
chlorinated compounds must not significantly alte r the host or the
natural ecosystem.

Although abiotic transformations can be significant within the time scales
commonly associated with groundwater movement, the biotic processes
typically proceed much faster, provided that there are sufficient
substrates, nutrients and microbial population s to mediate such
transformations.

It is widely reported that strongly reducing environments are associated
with dehalogenation reactions, although the mechanism appears to be
unknown. There is evidence for the dehalogenation being associated with
the transition from aerobic to anaerobic conditions (KAESTNER 1991). The
reductive dechlorination of PCE (perchloroethylene) via TCE
(trichloroethylene) depended on specific transition conditions after
consumption of the electron acceptor oxygen or nitrate. Transformation
required an additional decrease in the redox potential caused by sulfide.
The decrease must be considered the driving force for the onset of
reductive dechlorination. Repeated feeding of TCE or PCE to cultures after
the change to anaerobic conditions yielded no further dechlorination. In
cultures that were anaerobic from the beginning of incubation, no
transformation of PCE was observed.

The transformation may be catalyzed by the aerobic or facultatively
anaerobic organisms if the redox potential drops to nonphysiological
values. Transformation in the dechlorinating cultures occurred when the
redox potential in the medium decreased to values between -50 and -150 mV
and when carbon sources (electron donors) were present in excess. After
consumption of the electron acceptor oxygen or nitrate by growth of the
aerobic bacteria, the redox potential in the medium reached only to 0 mV.
To reach the low redox potential required for dechlorination, a further
decrease in the redox potential caused by sulfide was necessary. However
dechlorination was also stimulated by the release of sulfide from the
degradation of organic sulfur compounds without the growth of sulfidogenic
bacteria. The number of cells in dechlorinating culture was already
decreasing during dechlorination. This result implies that the release of
cell compounds from the dying cells may also be involved in
dechlorination.(KAESTNER 1991).

The influence of supplemental organic substrates on the degradation of
xenobiotics in anaerobic conditions was reported by GIBSON and SUFLITA
(1990). They noted that degradation of 2,4,5-T by anaerobic bacteria had a
shorter acclimation period at times of the year when the groundwater from
the methanogenic site had dissolved humic material contained in it. The
conclusion was that reductive dehalogenation reactions are limited by the
availability of suitable electron donors. The hypothesis was confirmed by
the inclusion of common fermentation products in their anaerobic mixtures.
Butyrate, propionate, or ethanol additions had the greatest stimulatory
effect on anaerobic 2,4,5-T metabolism and reduced the acclimation time to
less than one month (from 3 months without the added substrate). Acetate
and methanol were less stimulatory and a 2-3 month acclimation was
observed with these treatments. Not only did the carbon amendments
stimulate the onset of dehalogenation, they also increased the extent of
2,4,5-T metabolism.

Biotransformation under methanogenic conditions of several other
chlorinated aliphatics in widespread use (e.g. tetrachloroethylene and
trichloroethylene) occurs primarily by reductive dechlorination. This
process requires the supply of an external electron donor.(FREEDMAN and
GOSSETT, 1991).

Anaerobic removal of the Cl substituents proceeds via reductive
dechlorination before the aromatic ring is cleaved. This dehalogenation
results in the formation of less toxic, more soluble, and less
recalcitrant compounds. (HENDRIKSEN et al. 1991)

In general, reductive dehalogenation reactions are favored under highly
reducing methanogenic conditions. (KUHN et al. 1990)

Recent research has shown that natural gas may stimulate TCE degradation
in aerobic sediment samples. These results suggest that methane-utilizing
bacteria may be able to biodegrade chlorinated alkenes such as TCE.
Methanotrophic bacteria are ubiquitous organisms that posses a unique
methane monooxygenase enzyme system which enables them to utilize methane
as a sole carbon and energy source. The methane monooxygenase enzyme
complex has a low substrate specificity and is able to oxidize or
dechlorinate a wide variety of economically and environmentally important
compounds. The bacteria used was isolated from a waste well where
chlorinated organic solvents were disposed of directly to the groundwater.
Contamination ranges from low to very high (100 mg/l). The inability of
this organism to degrade TCE in the absence of methane or methanol
suggests that TCE biodegradation by methanotrophs is a cometabolic process
which provides little or no metabolic benefit to the organism. The fact
that most aquifers are not well supplied with oxygen will favor anaerobic
reactions. Recharge from stripping towers used to remove volatile organic
compounds from contaminated aquifers will introduce oxygen into the
recharge zone (as well as spores from the surface) and create aerobic
conditions in the immediate area of the recharge well. (FREEDMAN and
GOSSETT, 1991).

Nitrosomonas europea is an obligate chemolithotrophic nitrifying bacterium
which derives its energy for growth exclusively from the oxidation of
ammonia to nitrite.  Evidence indicates that ammonia monooxygenase in
cells of this bacteria is also capable of cooxidizing hydrocarbons,
including DBCP and TCE (RASCHE et al, 1991).

ALLARD, et al, (1991) studied degradation of chlorocatechols by
metabolically stable anaerobic cultures. A high degree of specificity in
dechlorination was observed, and some chlorocatechols were appreciably
more resistant to dechlorination than others: only 3,5-dichlorocatechol,
4,5-dichlorocatechol, 3,4,5-trichlorocatechol, and tetrachlorocatechol
were dechlorinated, and not all of them were dechlorinated by the same
consortium. 3,5-dichlorocatechol produced 3-chlorocatechol,
4,5-dichlorocatechol; tetrachlorocatechol produced only
3,4,6-trichlorocatechol. Incubation of uncontaminated sediments without
additional carbon sources brought about dechlorination of
3,4,5-trichlorocatechol to 3,5-dichlorocatechol. O-demethylation of
chloroguaiacols was generally accomplished by enrichment cultures, except
that catechin enrichment was unable to O-demethylate tetrachloroguaiacol.
None of the enrichments dechlorinated any of the polychlorinated phenols
examined. The results suggested that dechlorination was not dependent on
enrichment with or growth at the expense of chlorinated compounds. They
concluded that the transformations described were mediated by bacterial
reactions.

CRIDDLE, et al (1990) suggest that the formation of radicals from carbon
tetrachloride may explain the product distribution resulting from its
transformation. Radicals formed by reduction of tetrachloromethane (carbon
tetrachloride) presumably react with constituents of the surrounding
milieu to give the observed product distribution. Use of oxygen and
nitrate as electron acceptors generally prevented carbon tetrachloride
metabolism. At low oxygen levels (about 1%) transformation of carbon
tetrachloride t o CO2 and attachment to bacterial cells material did
occur. Under fumarate respiring conditions, the carbon tetrachloride was
recovered as CO2 , chloroform and a nonvolatile fraction. In contrast,
fermenting conditions resulted in more chloroform, more cell-bound
radiolabeled C (from carbon tetrachloride ) and almost no CO2 . Rates of
transformation were faster under fermenting conditions than under fumarate
respiring conditions.

It is also reported that transformation rates decreased over time.
(CRIDDLE et al. 1990)

A plausible hypothesis for the observation is that many halogenated
xenobiotics undergo reductive dehalogenation is that these transformation
are fortuitous, resulting from the inherent activity of reducing agents
created by microorganisms; i.e. cometabolism. If this hypothesis is
correct, then many common and familiar microorganisms may bring about
unexpected and possibly unpredictable transformations when confronted with
chemicals that are foreign to their evolutionary history.

The agents of transformation may be generated by biological activity or
they could result from a change in solution chemistry brought about by
microbial activity.

Trihalomethyl radicals undergo addition reactions at the double bonds of
lipids and unsaturated acids. One explanation for the lack of
transformation under fully aerobic reactions is a paucity of sufficiently
powerful reducing agents. (CRIDDLE, et al. 199 0)

Pentachlorophenol
Pentachlorophenol (PCP), the chlorinated phenols used as wood
preservatives, herbicides, fungicides and general biocides are a large
group of toxic xenobiotics that are serious environmental pollutants.
Reductive dechlorination of PCP has been observed in flooded soils.
Actinomycetes and fungal organisms have also been found to metabolize PCP,
However, little is known about the microorganisms responsible for the
anaerobic degradation of PCP.

MADSEN and AAMAND (1991) report degradation of PCP under methanogenic and
sulfate-reducing conditions with an anaerobic mixed culture derived from
sewage sludge. The consortium degraded PCP via 2,3,4,5-tetrachlorophenol,
3,4,5-trichlorophenol and 3,5-dichlorophenol and eventually accumulated
3-chlorophenol. Dechlorination of PCP and metabolites was inhibited in the
presence of sulfate, thiosulfate, and sulfite. A decrease in the rate of
PCP transformation was noted when the endogenous dissolved H2 was depleted
below 0.11 micromoles per liter in sulfate reducing cultures. The effect
on dechlorination observed with sulfate could be relieved by addition of
molybdate, a competitive inhibitor of sulfate reduction. Addition of H2
reduced the inhibition observed with sulfoxy anions. The inhibitory effect
of sulfuoxy anions may be due to a competition for H2 between sulfate
reduction and dechlorination. When cultured under methanogenic conditions,
the consortium degraded several chlorinated and brominated phenols.

Halogenated phenols are reductively dehalogenated in sewage sludge,
aquatic sediments and soils.  Some of these studies indicate that
reductive dehalogenation reactions may be favored in methanogenic
environments. Experiments with anaerobic groundwater sediment showed that
a dehalogenating potential was present in the methanogenic part of he
aquifer, but this potential was at least partially inhibited by the high
concentration of sulfate at the nearby sulfate-reducing site. Sulfate is
commonly present in anaerobic habitats, such as aquatic sediments, soil
and wastewater sludge.

Bromacil Dehalogenation
The metabolic fate of bromacil in anaerobic aquifer slurries held under
denitrifying, sulfate reducing and methanogenic conditions revealed that
bromacil was debrominated under methanogenic conditions but was not
degraded under other incubation conditions . (ADRIAN and SUFLITA, 1990).
This finding extends the range of aryl reductive dehalogenation reactions
to include nitrogen heterocyclic compounds.

Creosote Breakdown
Unlined surface impoundments at a wood preserving plant in Florida in
direct hydraulic contact with the aquifer resulted in two distinct phases
when the creosote and water mixed. (GODSY et al. 1992). A denser than
water hydrocarbon phase that moved vertically downward, and an
organic-rich aqueous phase that moved laterally with the groundwater flow.
The aqueous phase was enriched in organic acids, phenolic compounds, and
single- and double ring aromatic hydrocarbons. The ground water was devoid
of dissolved O2, was 60-70% saturated with CH4 and contained H2S. Field
analyses documented a greater decrease in concentration of organic fatty
acids, benzioc acid, phenol, 2-,3-,4-methylphenol, quinoline,
isoquinoline, 1(2H)- quinolinone, and 2(1H)-isoquinolinone during
downgradient movement in the aquifer than could be explained by dilution
and/or dispersion. Laboratory studies showed that within the study region,
this effect could be attributed to microbial degradation to CH4 and CO2 .
A small but active methanogenic population was found on sediment
materials taken from highly contaminated parts of the aquifer.

Creosote is a complex mixture of more than 200 major individual compounds
with differing molecular weights, polarities, and functionalities, along
with dispersed solids and products of polymerization. The major classes of
compounds previously identified in creosote show that it consists of
approximately 85% by weight polynuclear aromatic hydrocarbons (PAH), 12%
phenolic compounds and 3% heterocyclic nitrogen, sulfur and oxygen
containing compounds (NSO).

During the first 50 days of residence in the microcosm, C3-C6 volatile
fatty acids were rapidly converted to acetic acid and ultimately to CH4
and CO2 . Benzoic acid, quinoline, and isoquinoline are also biodegraded.
Phenol degradation occurs between days 50 and 99 in the microcosms and
phenol also disappears from the ground water during transit. After 100
days in the microcosm, 2-,3-,4- methylphenol, 2(1H)-quinolinone and
1(2H)-isoquinolinone are biodegraded and are removed from the system after
about 20 0 days. A similar pattern of disappearance is observed for 2-,
3-, and 4-methylphenol. The degradation of 2-methylphenol appears to be
somewhat slower than the other methylphenols in the groundwater. This
compound, which is widely held to be recalcitrant, was readily degraded
during downgradient transport in the aquifer and the microcosm used in
the cited study. (GODSY et al. 1992).

Substituted Indoles
Aromatic N-heterocyclic compounds, including substituted indoles are often
found in aqueous waste effluents associated with oil shale and coal mining
operations. Not surprisingly, the ability of sediment and sewage sludge
microcosms to degrade indole is dependent upon several factors, including
incubation temperature and the amount of sediment or sludge inoculum used.

Degradation of indole by an indole-degrading methanogenic consortium
enriched from sewage sludge proceeded through a two-step hydroxylation
pathway yielding oxindole and isatin. (GU and BERRY, 1991). Under
anaerobic conditions, the source of oxygen for th e hydroxylation reaction
is water (a nucleophile). It is found that an indole degrading
methanogenic consortium is capable of transforming 3-methylindole and
3-indolyl acetate. In neither case were the aromatic ring structures
catabolized.

Cyanide breakdown
Upflow, anaerobic, fixed-bed activated charcoal biotreatment columns
capable of operating at free cyanide concentrations of > 100 mg/l with a
hydraulic retention time of <48h were developed. (FALLON et al. 1991).
Methanogenesis was maintained under a vari ety of feed medium conditions
which included ethanol, phenol or methanol as the primary reduced carbon
source. Under optimal conditions, >70% of the inflow free cyanide was
removed in the first 30% of the column height. Strongly complexed cyanides
were resistant to removal. Ammonia was the nitrogen end product of
cyanide transformation . In cell material removed from the charcoal
columns, bicarbonate was the major carbon end product .

Cyanide spontaneously hydrolyzes to formic acid at a rate positively
correlated with pH, temperature, and trace metal concentration. Therefore,
under the right conditions, with a long enough retention time, the cyanide
spontaneously disappears.


ABIOTIC DEGRADATION
Abiotic Transformation
In real systems, it is generally not possible to distinguish between
biotic and abiotic; or even aerobic and anaerobic transformation events.
Numerous transformations occur in the homogeneous phases, especially in
the liquid phase. Other transformations occur in the interface between
phases. These include reactions that are heterogeneously catalyzed and
those that occur in solution under the influence of the electric field of
charged surfaces. Biologically produced enzyme and other biochemical
compounds in the soil can be involved in transformations of pollutants.
Thus, sterilization, which destroys living organisms, will also affect the
abiotic chemical reactions that are dependent on substances generated by
biological processes. Sterilization processes, furthermore, can alter
nonbiological constituents of the treated systems. Heat and radiation, for
example, can affect the free-radical content of the soil. Finally, many
degradative pathways include both biologically and chemically controlled
steps.

Abiotic degradation occurs when a chemical undergoes a thermodynamically
favorable change without the catalyst of biological enzymes. This includes
a large number of potential reactions, including those catalyzed by
hydrogen ions. Hydrogen ion activity affects transformation kinetics by
two major mechanisms; acid-base mediated hydrolysis reactions and
dissociation of acidic or basic compounds.

In addition to pH induced reactions, other abiotic processes may involve
reactions such as those involving dissolved organics and suspended
particles, metal ions, redox reactions.

Natural sorbents may be biotic and abiotic; may be organic, inorganic, or
chemical composites thereof and may range in size from macromolecules to
gravel. The surface charge and chemical composition may change with
ambient solution pH and redox potential.These changing conditions may
affect the colloidal size, and molecular configuration.

Pollutants vary in water solubility from complete miscibility to virtual
insolubility.

As a function of ambient conditions, any given material can undergo
chemical reaction. If a thermodynamic gradient exists and sufficient
activation energy, or a suitable catalyst is provided, a reaction is
possible. The actual reaction event, or on a larger scale, the rate of
the reaction, may be kinetically limited.

Degradation processes always flow in the direction of the least energy.
Oxidation is a favored reaction because of the stability of the products
in an oxidizing environment. Reduction is a favored reaction in a reducing
environment because of the stability of the products in a reducing
environment.

Sometimes activation energies prevent thermodynamically favorable
reactions from happening. Catalysts, such as enzymes and surface
interfaces, help to allow a favorable reaction to occur by; (1) lowering
the activation energy so the reaction proceeds under ambient conditions
of temperature, pH, pE, etc. ; or (2) by increasing the relative
concentrations of the reactive species at the reactive surfaces.


SORPTION
Phase interfaces are important. Most chemical reactions that occur in
water takes place at phase discontinuities, such as the air-water or
solid-water interfaces (STUMM and MORGAN, 1981). Sorption onto a surface
can alter the configuration or energy status of a molecule in such a way
as to enable a reaction. The physical process of adsorption onto a surface
causes changes in the conformation or arrangement of the bonds in the
adsorbed species (BARROW, 1973). Such changes may increase the rate of a
reaction and thus be considered a catalytic effect. The catalysis of
chemical reactions by certain surfaces is an important process. Mineral
clays are reported to catalyze some reactions involving organic chemicals
(MORTLAND, 1985). In addition to catalysis, concentration of material at
a surface can increase the effective concentration of reactants and thus
enable reactions that might not be possible in dilute systems.

Adsorption is the concentration of a component on the external surface at
an interface while absorption is usually meant to describe the movement of
something into the interior of a matrix.

Because of the difficulties in discerning the boundaries of the
solid-water interfaces, the more general term "sorption" has frequently
been adopted to describe both adsorption and absorption. Sorption is a
more generally applicable term which encompasses both processes and
simply relates interfacial flux. In practice, sorption is usually meant to
indicate the movement from the free, or mobile phase (gas or liquid) into
or onto the fixed phase. Desorption is used to denote the movement from
the fixed phase back into the mobile phase. In actuality, sorption
involves two processes; the movement from one phase to another involves
changes in both phases and the overall systems of both phases will reflect
the event. In the case of adsorption from aqueous solution, it is usually
considered that the process is competitive and that something must be
removed (desorbed) or re-arranged to accommodate the newly sorbed species.
Absorption, on the other hand, can involve the movement from one liquid
into another with out the necessity of removing a sorbed species or
competing for a site on the surface (MINGELGRIN and GERSTL, 1983).

BONDING FORCES
A number of different attractive forces may be involved in sorption and
different mechanisms have been proposed to explain the process by which a
given species can partition between phases. The mechanisms act between all
of the system constituents; i.e., species in solution can interact with
each other, they can also react with the water molecules or with the solid
phase to influence the overall system behavior.

Covalent bonds are thought to involve the sharing of electrons between
adjacent atoms in a molecule. The formation of such bonds between species
is a possible form of partition interaction. In such an arrangement, the
bonding electrons of the reactants share orbitals and form relatively
strong bonds. These types of bonds are typically exothermic, largely
irreversible, and frequently require an activation energy or catalyst to
form.

Uneven distribution of electrons due to differences in electronegativity
between molecular substituents can result in the formation of a dipole
such as water. Molecules with a net dipole moment can undergo mutual
bonding interactions by virtue of localized areas of charge resulting
from uneven concentrations of electrons relative to the positively charged
nuclei of the constituent atoms. Because water is polar, polar materials
tend to be more or less soluble in water. In much the same way as the
dipoles in the water milieu can form positive associations with polar
contaminants, so can polar molecules other than water form polar
associations with polar contaminants. Hence, polar interactions can serve
to partition contaminants into or out of the aqueous phase. Molecules in
the solid phase can, by virtue of unequal distribution of electrons,
affect molecular orientation and distribution by polar interaction.

Hydrogen bonding is a very important polar interaction in water. The
importance of hydrogen bonding in aqueous systems is such that there is
often a tendency to consider hydrogen bonding as some sort of special or
unique bond type. It may well be considered as an extreme manifestation
of dipole-dipole interactions which typically arises when hydrogen is
attached to very electronegative atoms. Hydrogen bonding also occurs in
some other polar liquids such as alcohols.

Molecular excess or deficiency of electrons relative to protons can
produce anions or cations which posses a static charge. Electrostatic
charge can influence the distribution or orientation of charged or polar
species in solution. Species which have net charge can be dissolved by the
water matrix in a manner somewhat akin to the polar mechanism of solution
described above. Water can act to dissolve ions in a crystal, when the
force of attraction of the water dipoles is greater than the force of
attraction of the bonds in the crystal lattice of the solid phase. The
presence of an ion in solution will cause a certain amount of "ordering"
or nonrandom arrangement of the water molecules as the dipoles orient
their polar ends in response to the standing charge on the central ion,
coordinated as ligands in a sphere of hydration around the ion.

The formation of neutral ion pairs in solution can occur when oppositely
charged ions become attached to each other and the attraction or motion of
the water molecules is sufficient to maintain the ion pair in the
dissolved state against the forces of gravity or attraction to the
crystalline solid phase.

Ions in solution can affect the distribution and status of other dissolved
ions as well. An ion in solution may be surrounded by neutral or
oppositely charged ligands to form a complex.

Chelation is an especially strong sort of complex whereby a ligand forms
two or more bonds with the central coordinating ion. A single ligand
capable of forming more than one association is termed multidentate or
polydentate.

In addition to the interaction of charged species in solution, sorption
onto the solid phase is also possible due to the interaction of a charged
surface with the dissolved ions in solution. A charged surface will
attract oppositely charged counterions and result in a relative
concentration of counterions in the interfacial region. The interfacial
region between a charged surface and a solution containing ions is
considered to be "ordered", relative to the ionic distribution in the bulk
phase, as counterions in solution are distributed in response to the
static charge on the surface. This has been called an electrical double
layer. There have been several descriptive models advanced to relate the
distribution of ions in solution in response to the existence of a
charged interface (STUMM and MORGAN, 1981).

The thickness of the double layer will diminish with the ionic strength of
the solution. As the concentration of charge in solution increases (i.e.,
the ionic strength of the solution increases), the total concentration of
countercharges adjacent to the charged surface increases, and the
thickness of the layer of counterions decreases or collapses in proportion
to the ionic strength of the bulk solution. An increase in solution ionic
strength may have an impact on the processes of adsorption at the surface,
not only by decreasing the thickness of the electrical field surrounding
the charged surface, but by increasing the competition for adsorption by
other solution components.

Ionic species can induce a dipole in a nonpolar molecule over a short
range. London forces exist between instantaneous and induced dipoles, and
are operative between all bodies when they are close together. They are
also commonly called van der Waals attr active forces after the Dutch
physicist (J.D. van der Waals) who described these forces as being active
in crystals (PAULING, 1957). The London/van der Waals force is also
frequently referred to as the dispersion force. This force is very
important in solution phase as well.

The nature of the London force is that it is proportional to the molecular
volume and the number of polarizable electrons of the species experiencing
the force. Even nonpolar neutral species undergo momentary imbalances in
electron distribution. The forces which exist between instantaneous
dipoles are responsible for much of the interactive cohesion in solutions
of nonpolar liquids. The impact of the London force on sorption from
solution tends to become pronounced when large molecules are involved;
larger molecules have a larger molecular volume and more electrons. It is
thought that the essence of the van der Waals force is the attraction of
electrons of one molecule for the atomic nuclei of another (PAULING,
1957). The ability of species to engage in van der Waals bonding is
related to the number of electrons and to the ability of those electrons
to accommodate the close approach of the bonding partner's electrons. This
latter ability is called polarizability and may be thought of as the ease
of inducing a dipole moment in a species.

As a result of the nature of the intermolecular interaction which gives
rise to van der Waals force, this force is only active at very close
range. The molecules must approach one another closely before the
attraction which results in sorption can exert itself. It is generally
believed that the force of the van der Waals attraction between two
molecules is proportional (a) to the square of the polarizability and
varies inversely with the sixth power of the distance between the
molecules.

               Q (proportional)   n2/ r6       (1)
where
Q = force of the attraction between molecules
n = polarizability
r = distance between the molecules

The variation of the energy of attraction attributed to van der Waals
force as a function of distance between sorbate and sorbent may be
described graphically with a hypothetical plot of potential energy vs
distance.

(ascii format does not support this graphic- see original report for
graph) Figure 1. Hypothetical plot of van der Waals attraction as a
function of distance between interacting molecules

At distances greater than a few molecular diameters, the energy of
attraction is negligible. As the molecules approach, the force of
attraction increases (the potential energy decreases) as natural or
induced dipoles begin to interact. As the molecules grow even closer,
steric factors come into play and the potential increases dramatically.
The point of minimum potential energy, then, is the point of maximum
attraction and relates to the point of closest approach.

These bonding interactions described are frequently considered to be
representative of the major types of forces which exist between species,
although there is some disagreement about the nature and magnitude of the
forces involved. It is probable that combination or hybrid forces come
into play in real material interactions. It is also probable that multiple
types of attractive and repulsive forces act simultaneously in many
complex systems (e.g., WERSHAW and PINKNEY, 1973).

FREE ENERGY
Partitioning is governed by free energy change. The net free energy
describes the overall tendency of the system to make a specific change.
The concept is in accord with the laws of thermodynamics and assumes that
it is the natural tendency of a system to spontaneously seek a condition
of minimum energy and maximum disorder. The most common form of the
equation is

               delta G = delta H - T delta S   (2)
where
delta G is the change in free energy associated with the event
delta H is the change in enthalpy
T is the absolute temperature
delta S is the change in entropy which accompanies the event.

The consideration of net free energy is associated with a specified change
and demands clear definitions of the system under consideration, both
before and after the change.

The value of the free energy relation is that spontaneous reactions must
always be associated with a negative change in free energy i.e., delta G <
0. If delta G is greater than zero, the reverse reaction is
thermodynamically favored.

The free energy of a sorption process can, in principle, be determined
from K, the slope of the linear isothermal plot according to the equation

                       delta G = RT ln K       (3)
where
R = the gas constant
T = the absolute temperature

Quantitative application of free energy data requires rigorous definitions
of the system. Since the equilibrium constant for the distribution between
the bulk and surface phases (K) is not well defined due to the uncertainty
in the thickness (volume) of the adsorption layer, the values of delta G
are only approximate (MINGELGRIN and GERSTL, 1983).

The tendency of the system to minimize its energy is accounted for by
considering the energy (enthalpy, H) contained in the bonds or forces of
association between the system components before and after the specified
change. If the net energy of bonds is lower in the system after the
change, the change is considered to be favorable from the aspect of net
enthalpy.

The free energy concept accounts for the tendency of the system to
maximize disorder through the entropy term; S. The entropy of the system
is directly related to the numbers of system components and the freedom of
random motion of the system before and after the specified change.

It is incorrect to assume that adsorption always represents a decrease in
system entropy. Adsorption at the surface by a solute component may
require the removal of another species which is adsorbed to the surface,
hence the increased order or disorder of the system accompanying
competitive adsorption from solution is not so clear cut as might be the
case of adsorption of a gas molecule from a near-vacuum.

The transfer of a hydrophobic solute from aqueous solution across a phase
boundary into an immiscible liquid phase is reported to represent an
increase in entropy. Two major sources of entropy increase have been
suggested. One is that hydrophobic solutes lead to increased structuring
of water. Decreased structuring when the solute leaves the aqueous phase
would increase randomness in water and therefore increase entropy. Another
cause of increase in entropy is greater conformational freedom of
hydrophobic molecules in non-aqueous media than in water. The increase in
structural conformations leads to an increase in randomness and an
increase in entropy (HASSETT and ANDERSON, 1982). Entropy changes in
complex systems may be difficult to enumerate. In fact, spontaneous events
(i.e. those with delta G < 0) are observed to display variations in both
magnitude and sign for enthalpy (H) and entropy (S) changes (MINGELGRIN
and GERSTL, 1983; OPPERHUIZEN et al., 1988). It is the combination of
these two parameters, along with the consideration of the temperature (T),
which describes the net free energy, and hence the opportunity for a
spontaneous event.

In any case it is well to remember that the existence of a favorable free
energy gradient (delta G < 0) does not guarantee that an event will occur
within any time frame. Kinetics are not considered in the free energy
determination, nor is the existence of an activation energy. An event may
have a favorable free energy gradient and yet be limited by the kinetics
or activation energy requirements.

INTERFACIAL TENSION
Water molecules at the air-water interface experience unbalanced
attraction for the water and the air. This is a manifestation of the polar
nature of water in contact with a nonpolar phase (the air). The water
molecules are drawn together, resulting in a phenomenon called "surface
tension". The contact area between the water and the nonpolar phase is a
region of relatively high interfacial tension and the system will
naturally tend to minimize such contact. This polar structure of water
will tend to make the aqueous medium relatively inhospitable to nonpolar
neutral (uncharged) molecules as well (HORVATH, et al.,1976).

A nonpolar neutral species in a polar medium such as water experiences
interfacial tension. Solvophobic theory is a general statement of
hydrophobic theory which has been developed to explain the tendency of
neutral organic species to flee the water phase.

It has been reported that the solution of nonelectrolytes in water is
attended by a net decrease in entropy (EGANHOUSE and CALDER, 1976). This
has been attributed to an increased structuring of water molecules in the
vicinity of the solute. The process may be conceptually rationalized by
considering that a solute must occupy space in a cohesive medium. The
solute must create a "cavity" in the water milieu and then occupy that
cavity. (e.g., McDEVIT and LONG, 1952; AMIDON et al., 1974; YALKOWSKY et
al.,197 5; BRIGGS, 1981). The very high cohesive density of water creates
considerable interfacial tension in the region of contact with a nonpolar
solute and is responsible for the magnitude of the hydrophobic effect.
This interfacial tension has also been called the internal pressure
(GORDON and THORNE, 1967) and it creates a driving force for the
nonelectrolyte to flee the solution as the system tries to minimize the
area of contact between the water and the nonpolar solute. This is a more
rigorous way of saying that oil and water don't mix. The hydrophobic
concept has been of great utility in explaining the behavior of organic
chemicals in water. Hydrophobic forces can drive nonpolar neutral solutes
across an interfacial boundary into an adjacent immiscible nonpolar
liquid. A substantial part of the driving force of this reaction may be a
positive entropy change which was described above.

What is sometimes called "hydrophobic bonding" is largely the extension of
solvophobic behavior to create a partitioning event such as adsorption
onto a solid material. The so-called hydrophobic bond is not so much a
special type of bond as it is a way for the system to minimize the area
of the polar and nonpolar interface (HORVATH et al.,1976). If the site of
sorption is itself hydrophobic, sorption of a nonelectrolyte onto such a
site will be attended by a proportionally greater reduction in the overall
system interfacial tension and the driving force will be that much
greater. Upon sorption, London forces are certainly involved and so
bonding per se is occurring, but the solvophobic tendency is providing a
considerable gradient for the sorption event.

A direct consequence of hydrophobic theory is manifest in Traube's rule,
which states that the water solubilities of an homologous organic series
decrease as the length of the carbon chain increases. As the length of the
nonpolar carbon chain increases, so does the nonpolar surface area of the
molecule. While a functional group may be relatively polar, the nonpolar
surface area creates the interfacial tension in aqueous solution and thus
the water solubility will decrease as the chain length increases. Traube's
rule accommodates the balance between hydrophobicity and hydrophilicity.

Traube's rule has been somewhat extended and formalized with the
development of a quantitative methods to estimate the surface area of
molecules based on their structures (e.g., YALKOWSKY et al.,1972,1975;
AMIDON et al.,1974;  HORVATH et al.,1976; BRIGGS 1981). The molecular
surface area approach suggests that the number of water molecules that can
be packed around the solute molecule plays an important role in the
theoretical calculation of the thermodynamic properties of the solution.
Hence, the molecular surface area of the solute is an important parameter
in the theory. In compounds other than simple alkanes, the functional
groups will tend to be more or less polar and thus relatively compatible
with the polar water matrix. Hence, the total surface area of the
molecule can be subdivided into "functional group surface area"  and
"hydrocarbonaceous surface area". These quantities may be determined for
simple compounds as an additive function of constituent groups with
subtractions made for the areas wh ere intramolecular contact is made and
thus no external surface is presented (AMIDON et al.,1975).

It is found that for molecules with longer carbon chains, the ability to
predict solubility based on calculated molecular surface area is
diminished. It has been suggested that, as molecular size increases,
coiling and self-association of the flexible chains, or perhaps
multimolecular aggregate formation spontaneously occurs to minimize the
nonpolar surface area (AMIDON et al.,1974).

The solubility of organic chemicals in water ranges from complete
miscibility to near insolubility. Many natural aqueous systems contain a
nonpolar phase such as soil organic matter or lipids in organisms.

The octanol/water partition coefficient has arisen as a measure of the
water/nonpolar partitioning behavior. This parameter has gained wide use
as a key indicator of the environmental fate of organic chemicals. The
octanol/water partition coefficient is the ratio of a chemical's
concentration in octanol phase to its concentration in the aqueous phase
of a two phase octanol/water system. The octanol/water partition
coefficient is not the same as the ratio of a chemical's solubility in
octanol to that in wa ter, because the organic and aqueous phases of the
binary octanol/water system are not pure octanol and pure water (CHIOU and
FREED, 1977b; CHIOU et al.,1982). The octanol water partition coefficient
is a useful indicator of the hydrophobicity of an orga nic chemical.

SALT EFFECTS
Natural water is not pure and it has been observed that solution
modifications affect the behavior of solutes. The presence of dissolved
salts in solution has been observed to both increase and decrease the
solubility of neutral nonelectrolytes.

"Salting out" is the term that has been used to describe the decreased
solubility of nonelectrolytes as a function of concentration of simple
salts in solution (McDEVIT and LONG, 1952; AQUAN-YEUN et al.,1979). Such
behavior is consistent with the hydrophobic theory discussed above. The
addition of simple salt to water increases the interfacial tension between
water and nonpolar phases and will thus tend to force nonelectrolytes out
of solution.

The tendency to "salt-out" is apparently countered when salts containing
large complex ions are used for the electrolyte. In such cases, the
solubility of the nonelectrolyte is frequently seen to increase, in a
phenomenon that has been logically called "salting-in". The tendency to
salt-in is seen to increase with the size of the ions involved, whether
they are cations or anions. This seems to indicate that additional
interaction terms of the van der Waals type must be considered (LONG and
McDEVIT, 1952) . Such interactive forces involve the polarizability of
salt ions, solvent molecules and non-electrolyte solute molecules as well
as forces between any component dipoles that may be present (SAYLOR et
al.,1952; STEIGMAN et al.,1965).

The effect of salts on the solubility of nonelectrolytes may be described
by the Setschenow equation;

                       log f = ksCs    (4)


where
f  = So/S = molar activity coefficient of the nonelectrolyte
So = molar solubility of the nonelectrolyte in pure water
S  = molar solubility of the nonelectrolyte in salt solution
ks = parameter dependent on the particular salt
Cs = the molarity of the salt

In any case the effectiveness of different electrolytes to create these
effects in solution is considerable.

(ascii format does not support this graphic- see original report for
graph) Figure 2. Effect of salts (approximate) on the activity coefficient
of benzene (after McDevit and Long, 1952)

It is important to note that the effect of salt on nonelectrolyte
solubility (benzene in this case) is significant only at salt
concentrations that exceed those typical in fresh waters.


COSOLVENCY
Polar neutral organics can be very miscible in water by virtue of
compatibility with the polar water molecules; for example dipole-dipole
interactions such as those interactions between short-chain alcohols and
water give rise to essentially complete miscibility.

In contrast to the increase in surface tension accompanying solutions of
salts, miscible organics in solution tend to decrease the surface tension
of the aqueous medium.

(ascii format does not support this graphic- see original report for
graph) Figure 3. Approximate surface tension as a function of the
composition in mixed solvents and salt solutions (after Horvath et
al.,1976).


Miscible organic solutes modify the solvent properties of the solution to
decrease the interfacial tension and give rise to an enhanced solubility
of organic chemicals in a phenomenon often called "cosolvency".

According to theory, a miscible organic chemical such as a short chain
alcohol, will have the effect of modifying the structure of the water in
which it is dissolved. On the macroscopic scale, this will manifest as a
decrease in the surface tension of the solution. It is noted that surface
tension is a gross parameter and experimentally determined interfacial
tensions (that which exist between the solute and solvent species) are
generally less than would be predicted based on surface tension
measurement ( YALKOWSKY et al.,1976).

It has been generally considered that there is an exponential increase in
the solubility of a solute as the fraction of the cosolvent increases
linearly. The only requirement for the log linear relationship seems to be
that the solute must be less polar than the mixed solvent (YALKOWSKY et
al.,1976). The validity of the log-linear nature of the cosolvent process
has been well validated in the literature (e.g., NKEDI-KIZZA et al.,1985,
1987; RAO et al.,1985; WOODBURN et al.,1986,1989; WALTERS and GUISEPPI-
ELIE, 1988; WALTERS et al.,1989).

The effect of a cosolvent on solubility has been figured according to the
following equation:

               ln (Sm) = fc ln (Sc) + (1 - fc) ln (Sw) (5)

where
Sm = molar solubility of a nonpolar solute
fc = nominal cosolvent volume fraction
Sc = molar solubility in pure cosolvent
Sw = molar solubility in pure water

This model assumes the absence of specific solute-solvent interactions and
is based upon a linear relationship between the free energy of solution
and solute surface area. It assumes that the overall solubility is simply
the sum of the solubilities in the individual solvent components. This
model treats the cosolvent and the water as distinct entities and neglects
any interaction between them.

More recent work with cosolvency in dilute systems seems to indicate that
the magnitude of the solubility enhancement is linear up to some 10-20%
cosolvent fraction (BANERJEE, 1985; BANERJEE and CASTROGIVANNI, 1987;
BANERJEE and YALKOWSKY, 1988; ZACHARA et al.,1988). At very low
concentrations of cosolvent, the assumption of non-interaction between the
cosolvent and water cannot hold. In dilute solutions the individual
cosolvent molecules will be fully hydrated, and as a result, will disrupt
the water network structure. If the total volume disrupted is regarded as
the extended hydration shell, and if Sc* is the average solubility within
this shell, then the overall solubility Sm in the water-cosolvent mixture
will be approximated by

               Sm = fcVH Sc* + (1- fcVH) Sw  ;  fcVH < 1       (6)

where VH is the ratio of the hydration shell volume to the volume of the
cosolvent.

In dilute solutions, the solute will, on average, contact only one
hydrated cosolvent molecule at a time, and the degree of solubilization
should be a linear rather than a logarithmic function of cosolvent
content.

Thus, it is expected that the log-linear relationship between Sm and fc
that applies at high cosolvent concentrations will become linear at low
cosolvent levels due to a change in the mechanism of solubilization. If S+
is defined as solubility enhancement , (Sm - Sw), then the relative
solubility enhancement at low cosolvent concentration will be given by

               S+/Sw = fc VH(Sc*/ Sw -1)       (7)

While cosolvency has been applied to environmental water chemistry
discussions, it is important to point out that the principle was
originally described by pharmaceutical chemists interested in solubilizing
nonpolar drugs. The presence of 20% miscible cosolvent in a therapeutic
treatment for a mammal is somewhat different from 20% cosolvent in a
natural water system. In a natural water system where cosolvent was
present at sufficient levels to influence contaminant solubility, the
cosolvent itself would probably constitute a contaminant. In a
contaminated groundwater, however, such a cosolvent concentration may be
realistic to create, thereby significantly enhancing the degradation of
the target pollutant. If the cosolvent were itself biodegradable, the
resulting effect would be the removal of the pollutant without adverse
long-term effects on the resource.

The log-linear solubility enhancement by cosolutes may be important in
characterizing concentrated leachate plumes or chemical spills, but will
be of little importance in characterizations of the dilute aqueous systems
that predominate in nature (NKEDI-KI ZZA et al.,1985; FU and LUTHY, 1986).

MICELLES
Organic chemicals can be quite variable in structure and properties. Many
organic molecules have both polar and nonpolar moieties, and the
solubility of the material in water will be the result of a balance
between the hydrophobic and hydrophilic tendencies. If a molecule
containing a hydrophilic region also has a significant hydrophobic region,
such as a long carbon chain, the water solubility will be diminished. This
diminished solubility can be manifest in several ways. The chemical can
sorb onto a surface and thereby diminish the interfacial tension with the
water or it can form a separate, immiscible bulk phase. A third
possibility exists, whereby the nonpolar moiety can undergo association
with the nonpolar regions of other molecules to form smaller subunits
within the water matrix. Such an organizational arrangement minimizes the
contact between the hydrophobic moieties and the water while allowing the
hydrophilic (polar/ionic) moieties to contact the water. Such an aggregate
arrangement is frequently referred to as a micelle.

Typically, organic chemicals having both polar and nonpolar moieties can
form micelles. Such chemicals are often referred to as "amphiphiles" or
described as being "amphipathic", which refers to the dual affinity of
such species for both polar and nonpola r media. Surfactants and soaps are
amphiphiles. They are often characterized by having a polar or ionic end
(or "head") and a nonpolar "hydrocarbonaceous" end (or "tail"). These
molecules in solution will to be subjected to the forces of interfacial
tension or polar affinity as have been so far delineated. The polar or
ionic end will be readily solvated by water, which will repel the nonpolar
end. Micelles arise when these molecules undergo intermolecular
association of the hydrophobic moieties and form a droplet of material
which has a hydrophobic interior and a hydrophilic exterior. The
interfacial tension between the water and the hydrophobic end is thus
minimized and the droplet may be solvated by its outer "shell" of polar or
charged ends in association with the polar water phase. This arrangement
has been called a "pseudophase" denoting the existence of a hydrophobic
interior of the droplets suspended by the interaction of the hydrophilic
moieties with the polar water.

It has been observed that the association of homogeneous surfactant
monomers to form micelles is characterized by some critical concentration
of dissolved monomers before true micelle formation (micellization)
occurs. A commonly described parameter associated with micelle formation
is the critical micelle concentration- or CMC. The onset of micellization,
which occurs at CMC, is typically accompanied by some well-defined or
observable change at that point. For example, a visible turbidity may
accompany the CMC. It is commonly reported that the addition of surfactant
monomer to water can cause the surface tension of the solution to decline
steadily until CMC is attained, after which continued addition of monomer
produces no more drop in the measured surface tension. The transition is
typically a sharp one. Experimentally, it is often found that micelles are
undetectable in dilute solutions of the monomers, and become detectable
over a narrow range of concentrations as the total concentration of solute
is increased, above which nearly all additional solute material forms
micelles. The concentration at which the micelles become first detectable
depends on the sensitivity of the experimental apparatus used to observe
the change in surface tension. The concentration range over which the
fraction of additional solute which forms micelles changes from nearly
zero to nearly unity depends on such factors as the number of monomers in
the micelle, the chain length of the monomer, the properties of
counterions and other details affecting the monomer-micelle equilibrium.
An approximate rule is that the higher the CMC value, the broader is the
concentration range over which this transition takes place, in absolute
value as well as in relative value in comparison with the CMC (MUKERJEE,
1971). Since different experimental methods may reflect this transition to
different extents, some systematic variations in operationally defined
CMC's are expected.

The impact of salt concentration on the formation of micelles has been
reported and is in apparent accord with the interfacial tension model
discussed above, where the CMC is lowered by the addition of simple
electrolytes (STEIGMAN et al.,1965; GORDON and THORNE, 1967).

The existence of a micellar phase in solution is important not only
insofar as it describes the behavior of amphipathic organic chemicals in
solution, but the existence of a nonpolar pseudophase can enhance the
solubility of other hydrophobic chemicals in solution as they partition
into the hydrophobic interior of the micelle. A general expression for the
solubility enhancement of a solute by surfactants has been given by KILE
and CHIOU (1989), in terms of the concentrations of monomers and micelles
and the corresponding solute partition coefficients, giving


               Sw*/ Sw = 1 + Xmn Kmn + Xmc Kmc         (8)


where
Sw*     = apparent solute solubility
X       = total stoichiometric surfactant concentration
Sw      = the intrinsic solubility in "pure water"
Xmn     = concentration of the surfactant as monomers
Xmc     = concentration of the surfactant in micellar form
Kmn     = partition constant between monomers and water
Kmc     = partition constant between micelles and water

The separation of the concentration terms (Xmn and Xmc) accounts for
differences in the partition efficiency of the solute with monomers and
micelles. By equation (8), one would expect a plot of the apparent solute
solubility (Sw*) versus the total concentration of surfactant to be
bilinear; giving a straight line with slope of Kmn from X = 0 to X = CMC,
followed by another straight line with slope of Kmc at X greater than or
equal to CMC. Because of the markedly greater organic environment of
micelles relative to monomers, the increase in slope of the plot on
exceeding the CMC should be very sharp. The two distinct slopes define the
values of Kmn and Kmc for a given solute-surfactant system.


(ascii format does not support this graphic- see original report for
graph) Figure 4. Hypothetical plot of surface tension as a function of
surfactant concentration for three molecularly homogeneous surfactants
(after Kile and Chiou, 1989).

While CMC is assumed to be an observable and definite value in the case of
surfactant monomers, there are frequent reports in the literature of the
formation of "aggregates" or micelle-like associations in solutions of
organic solutes so dilute as to apparently preclude the formation of
micelles.

Work with different types of commercial surfactants has indicated that
molecularly nonhomogeneous surfactants do not display the sharp inflection
in surface tension associated with CMC in molecularly homogeneous
monomers, rather the onset of aggregation is broad and indistinct (KILE
and CHIOU, 1989). The lack of well-defined CMC's for nonhomogeneous
surfactants is speculated to result from the successive micellization of
the heterogeneous monomers at different stoichiometric concentrations of
the surfactant, which results in a breadth of the monomeric-micelle
transition zone.

(ascii format does not support this graphic- see original report for
graph) Figure 5. Hypothetical plot of surface tension as a function of
surfactant concentration for three molecularly heterogeneous surfactants
(after Kile and Chiou, 1989).

It is observed that molecularly nonhomogeneous surfactants are able to
enhance the solubility of very hydrophobic chemicals such as DDT at
surfactant concentrations well below the CMC. This is attributed to the
successive micellization of the heterogeneous monomer species.

Examination of the solubility enhancement with different types of
commercial surfactants reveals that molecularly homogeneous surfactants
show relatively insignificant (but linear) solubility enhancement below
CMC. Molecularly nonhomogeneous surfactants, on the other hand, show a
much greater solubility enhancement at concentrations below the CMC.

The plot of the apparent solubility of DDT as a function of the ratio of
surfactant concentration (X) to critical micellar concentration (CMC)
shows a smooth, gradual upward curvature below the nominal CMC that
becomes increasingly steeper near and beyond the CMC as the pseudophase
grows.

(ascii format does not support this graphic- see original report for
graph) Figure 6. Hypothetical plot of apparent solubility enhancement of
DDT by surfactants at concentrations approaching nominal CMC (after Kile
and Chiou, 1989).

These characteristics are not predicted by the conventional theory for
homogeneous surfactants; the nonlinear relation near the nominal CMC is
strongly indicative of a continuous aggregate formation. This effect may
be attributed to a sequential micelliza tion of the heterogeneous monomers
because of their unequal solubilities in water. As a result, the
monomer-micelle transition of a heterogeneous surfactant may be expected
to be considerably less sharp than that of an homogeneous surfactant. This
reasoning is in accord with the relatively smooth solubility enhancement
curves over the region of the nominal CMC for the surfactants with
nonhomogeneous molecular composition.

The conventional methods for CMC determinations are not sensitive to
incipient formation of aggregates. In comparison, measurement of the
solubility enhancement appears to be much more sensitive to surfactant
aggregates. This can be partially rationalized by the fact that the change
in surface tension (as related mainly to monomers) of a surfactant
solution on exceeding the CMC is generally less than a factor of 3,
whereas the associated change in the apparent solubility of DDT (which is
strongly a function of the concentration of micelles) can be more than 2
orders of magnitude. For this reason, determination of the solubility
enhancement data of DDT or other extremely water-insoluble compounds may
prove the be the most sensitive method to date for detecting the
association transition of the surfactant in water (KILE and CHIOU, 1989).

The presence of water-soluble macromolecules in solution at submicellar
concentrations has been reported to enhance the water solubility of
hydrophobic organic chemicals in several instances (e.g., MADAN and
CADWALLADER, 1970; ROHMER et al.,1972).

Sorptive interactions or molecular aggregate formation in solution can
alter the reactivity of solutes. DUYNSTEE and GRUNWALD (1965) studied the
rate of alkaline hydrolysis of methyl-1-naphthoate in a solution of
different salts and observed the effect of the additions on the rate of
reaction to form methanol and the naphthoate ion. The results suggested
that methyl-1-naphthoate formed complexes with the added organic species.
If the complexes had a negative charge (organic anions), attack by the
hydroxide was impeded; in those with a positive charge, attack was
facilitated. The specific manner in which the neutral-salt effects varied
with the structure of the organic species suggested that the dominant
interactions leading to complex formation involve London or van der Waals
force.

This suggests that the organic-ion salt effects involve molecular
interactions similar to those that cause micelle formation in aqueous
solutions of detergent salts. This is stated in spite of the observation
that there did not appear to be the actual formation of micelles per se .
It appears that the actual formation of micelles created a phase in which
the micellar components were unreactive to hydrolysis.

The presence of macromolecules in solution can enhance the apparent
solubility of solutes by sorptive interactions in the solution phase. The
processes by which macromolecules enhance the solubility of contaminants
are probably variable as a function of the particular physical and
chemical properties of the system. A macromolecule possessing a
substantial nonpolar region can sorb a hydrophobic molecule thereby
minimizing the interfacial tension between the solute and the water.

PRECIPITATION
In a process somewhat akin to the formation of micelles, dissolved
inorganic solution constituents may precipitate and form a solid phase.
This solid phase may form as a bulk phase or exist as dispersed particles
in the solution phase with a continual gradation in between the two
extremes. At one end of this spectrum is the formation of neutral ion
pairs which exist in true solution. The settling of a solid precipitate or
the growth of crystals as a discrete solid phase may be considered as the
other end of the spectrum. The process of precipitation occurs as a
function of the intrinsic properties of the materials involved. The event
of precipitation or dissolution may be considered to be governed by the
free energy relation described above. The tendency of a material, AB, to
dissolve or precipitate may be described in an equilibrium mass-action
expression such as

                       Adissolved + Bdissolved = ABsolid       (9)

The relative tendencies of the reaction to proceed forward or backward as
written may be described by measuring the concentrations of all species at
equilibrium. The ratio of the products to the reactants may be figured to
yield an equilibrium constant according to the following form

                       [{A} {B}] / {AB}  = Kequilibrium        (10)


Where the brackets indicate activities of the reactants. Since the
activity of the solid phase is taken to be unity (by defining it as the
reference state), the denominator can be eliminated to yield the
solubility product constant, Ksp (STUMM and MORGAN, 1981; p.231).

                               {A} {B} = Ksp           (11)

This seemingly straightforward principle is somewhat complicated in
natural systems by the existence of changing conditions, kinetic
limitations and coincident or competing reactions.

Changing conditions may include changes in oxidation state of the
reactants (which may alter the reaction entirely) or changes in
temperature. As previously stated, the equilibrium considerations
undertaken here do not consider reaction rates or physical limitations
which may affect precipitation events. Simply attaining reactant
activities exceeding to the solubility product does not guarantee a
precipitation event.

When it does occur, precipitation can sometimes result in different
allotropic modifications, ranging from amorphous to crystalline and can
have variations within each form (e.g., WHITTEMORE and LANGMUIR, 1975;
STUMM and MORGAN, 1981).

Coprecipitation can occur when materials in solution get trapped or
caught-up in a precipitation event. This can cause scavenging of solution
constituents when a precipitate forms (e.g., JACKSON, 1975). Scavenging
can occur on different scales. On the molecular level, dissolved species
can become entrapped or bonded in the crystalline lattice if it forms.
This may result in phenomena such as isomorphic substitution in clay
minerals or simply the existence of "impurities" in the resulting solid.
On the macroscopic level, scavenging can occur when dissolved or
suspended solution components are taken out of solution by becoming
entrapped in a precipitate. This sort of coprecipitation or scavenging
event is intentionally created in the use of coagulants in water
treatment operations where a slightly soluble salt is rapidly added to
water in sufficient amounts to create a saturated solution. When
solubility is exceeded, the precipitation event scavenges materials in
solution which are responsible for undesirable turbidity. These are
trapped in the amorphous matrix and settled out by gravity, thus removing
them from solution.

pH
pH is a fundamental quantity. The pH of the solution can have an impact on
the solubility of organic and inorganic solutes. The pH can have an effect
on a reaction equilibrium if the reaction, or a related reaction, consumes
or produces H+ or OH-. pH relates the activity of protons in solution,
i.e.,

                       pH = -log {H+}  (12)

Hydroxide and carbonate typically form insoluble precipitates with
polyvalent cations in natural waters. The activity of both of these
species increases with pH.

The presence of surface functional groups which are capable of exchanging
a proton creates pH dependent charge, whereby the ionic character of the
surface increases with pH.

The molecular configuration of polyelectrolytes may be influenced by pH as
the molecules coil and uncoil as the pH decreases or increases. In such a
situation, charged sites such as acidic hydroxyl groups or amines can lose
or acquire charge as a result  of changes in solution pH. In a large,
flexible, polyfunctional molecule, intramolecular self-association is
thought to occur in the absence of electrostatic repulsion. The tendency
to form such intramolecular bonds will vary as charged sites are created
or satisfied by pH changes. In such a situation, decreases in pH will
satisfy the charge on the surface of the molecule, thereby lowering the
hydrophilicity of the surface and also decreasing the coulombic repulsion
of the molecular chain for itself and pe rmitting intramolecular bonding
(e.g., MATSUDA and SCHNITZER, 1971).


REDOX
Redox is another fundamental quantity. The redox status of a solution can
have an impact on the solution behavior by affecting the oxidation states
of the species in the system. Elements form different compounds as a
function of differing oxidation states . These compounds frequently differ
in their solubilities. Redox relates the activity of electrons in the
system. This can be measured using a galvanic cell consisting of two
electrodes connected by a conducting solution: oxidation occurs at the
negative electrode (anode) and electrons are produced, whereas electrons
are consumed and reduction takes place at the positive electrode
(cathode). The redox potential is quantified by comparison with a standard
redox couple. By convention, the standard redox couple is that present at
a hydrogen electrode consisting of a platinum electrode, with hydrogen
ions in solution. In the presence of platinum as the catalyst, the
reaction is: H2 to 2H+ + 2e-, and the tendency to donate reducing
equivalents, as electrons in this case, is measured as the voltage
(potential) of the electrical current generated, when the electrode is
coupled in series with another redox couple electrode. Under standard
conditions, 25C, 1 atm of H2 and pH0, the redox couple H2/2H+ + 2e- is
-420 mV. The symbol Eh is used for the redox potential under standard
conditions and is measured in volts.

Somewhat analogous to pH, pE gives the hypothetical electron activity at
equilibrium.

The two quantities are related to each other by

                       pE = -log {e-} = Eh /2.3 RTF-1  (13)

where
R = the gas constant
T = temperature (absolute)
F = the Faraday constant

pE is a measure of the relative tendency of the solution to accept or
transfer electrons. In a highly reducing solution the tendency to donate
electrons, that is, the hypothetical "electron pressure", or electron
activity is large. Just as the activity of hydrogen ions is very low at
high pH, the activity of electrons is very low at high pE. Thus a low pE
(or a low Eh) indicates reducing conditions.

Oxidation-reduction reactions occur by the transfer of electrons between
the reactants. Different materials differ in their tendency to accept
electrons in reactions, hence preferential reduction of easily reduced
materials occurs. Under equilibrium conditions with an abundance of
reducing agent, sequential reduction of electron acceptors takes place
with those most easily reduced accepting electrons until the concentration
of that species is depleted. Under such idealized conditions, the redox
potential of the solution will be constant during this process. When the
supply of this most-easily reduced reagent is exhausted, the pE will drop
to the threshold of the next most easily reduced material and the process
will continue. In real systems, kinetic considerations may limit the
ability of the redox system to express its equilibrium status.

The redox condition of the natural water system will have profound impact
on the nature of the overall system and the solubilities of aqueous
constituents. In well-aerated systems, oxygen usually serves as the
terminal electron acceptor in most oxidation reactions. In the absence of
sufficient oxygen, the pE drops until another electron acceptor becomes
available. This is commonly NO3-, Mn+4 or Fe+3, although other materials
may predominate depending on the geology or other conditions unique to the
specif ic situation. Under very reducing conditions in natural waters,
sulfide minerals are often formed from the reduction of sulfur containing
compounds. Redox changes can also induce changes in the ionic strength of
a solution by making relatively insoluble materials into more soluble
ones. An example of this is the reduction of iron or manganese. Both of
these compounds are more soluble in the reduced form.

The association between solubility and pE is linked with the pH of the
system. pE and pH affect one another as redox-driven processes produce or
consume hydrogen or hydroxide ions. The presence of water itself sets
limits to the extremes of pE which can o ccur in water by virtue of
water's ability to be oxidized or reduced.

Oxidation-reduction (or redox) reactions in biological reactions are
normally defined in terms of loss and gain of hydrogens or electrons. Each
oxidation is accompanied by a reduction. When both couples combine in a
complete redox reaction, the net flow of the reaction can be determined
by the relative tendency of each couple to donate or accept reducing
equivalents, which is the redox potential. A couple of lower redox
potential will always donate reducing equivalents to a couple of higher
potential and, during the oxidation of a substrate, reducing equivalents
are transferred in the direction of increasing potential. This transfer is
accompanied by the release of free energy, the magnitude of which is given
by the standard free energy change, � G = -nF delta Eh, where n is the
number of electrons transferred in the reaction, F is the charge on one
mole of electrons, which is Faraday's constant (96.649 kJ/V mol) and delta
Eh is the standard electrode potential. Biological systems have evolved to
conserve this energy and to convert it into biologically useful forms.

In practical terms, the redox potential can be used to indicate which
redox reactions will occur within a system. The redox potential gives a
measure of the general condition of the liquid.

Anaerobic processes will have low values of Eh (<-200 mV), where as
aerobic processes will have higher values (>+50mV). More precisely, values
of Eh -150mV to -420mV are found in anaerobic environments, whereas
aerobic environments vary between -200mV and +420 mV. Facultative
environments change from aerobic to anaerobic systems at about +100mV.
(GRAY. 1989)

In complex natural systems, there are numerous redox couples which are not
necessarily in equilibrium with each other. As a consequence, it is not
possible to define a unique pE to characterize the whole redox system. In
addition, it is generally accepted that there are no free electrons in
aqueous solution which could react with a redox electrode like H3O+ ions
do with a pH electrode. Despite these facts, redox electrodes are
frequently used to characterize the redox state by a single measurement
presuming that the solution is in equilibrium.

COLLOIDS
The presence of colloids in natural aqueous systems will act to influence
the distribution and behavior of contaminants. Colloids are formed by some
of the physical and chemical processes described in the theoretical
section above. These same physical and chemical forces govern the
processes by which natural (and perturbed) systems distribute pollutants
to and from the colloidal phase. The composition and behavior of colloids
are complex and difficult to rigorously define. In the absence of an
ability to effectively model colloidal systems in the natural world, we
must rely on descriptions of the processes in a more conceptual sense.

Colloids are particles in a solution that will not settle out. They are
common in natural waters and can enhance the apparent solubility of a wide
range of water pollutants, both organic and inorganic. Colloids may be
considered as an extension of the solid and aqueous phases and are formed
by conditions that can be quite variable in time and space; hence colloids
can be dynamic and relatively ephemeral. The composition of colloids can
vary with the composition of the solid and aqueous phases. Colloids can
be made up of organic, inorganic, or a mixture of material.

A colloidal solution has been defined as a solution intermediate in
character between a suspension and a true solution. Particles with
diameters less than 10 micrometer are usually called colloids (STUMM and
MORGAN, 1981), although the distinction based on size is arbitrary. The
size of particles is a continuum, and the point at which large
macromolecules end and small colloids begin is subject to judgment as is
the upper end of the size continuum, where colloids and suspended
particles merge. The tendenc y of suspended particles to settle out of
solution is not really a function of size alone, rather the relative
density of the particles and the motion of the water will determine what
is suspended and what settles. The very use of the term "colloid", which
defines a behavior and an approximate size, begs the more rigorous
definition of the chemical composition of the particles, yet this is a
realistic descriptor in common usage.

Colloids are present in surface and groundwaters. Surface systems receive
terrestrial input as runoff carries soil-derived materials into the
streams, rivers, lakes or estuaries. Groundwater receives leachate and
percolation water and is frequently well-connected with surface water
bodies. Colloids may also be formed in situ by native processes of
precipitation and dissolution, suspension or biological activity.

Colloids in solution represent a highly disperse solid phase. Because of
the sorptive behavior of interfaces, the higher surface area of dispersed
colloids tends to make colloids a more effective adsorbent on a mass basis
than an equivalent mass of precipitated or solid material. Colloids act
to enhance the solubility of slightly soluble contaminants, whether they
be organic or inorganic. Hydrophobic organics, and slightly-soluble
inorganics, including radionuclides, have been associated with colloids in
apparent solution.

Colloids as a Third Phase
The importance of the colloidal phase in the distribution of water
pollutants is a relatively recent issue in the environmental literature.
The phenomenon of colloidal solubility enhancement was detected by workers
in several fields and was largely unexplained. The concept was apparently
developed and forwarded by working with partitioning behavior of water
pollutants in water/sediment systems. It was observed that the amount of
sediment used in batch isotherms influenced the percentage of pollutant
which was sorbed onto the solid phase.

Work cited by O'CONNOR and CONNOLLY (1980) found that equilibrium sorption
partition coefficients of several radioactive materials into Texas river
sediments declined as sediment concentration increased in isothermal
studies. This has been interpreted as an indication that colloids in
solution were competing with the sediment for sorbate and that the
concentration of colloids increased as the concentration of sediment
increased.

The importance of colloids was recognized by VOICE et al. (1983) when they
discussed what they called the "particle concentration effect"; a term
coined to describe the observation that the partition coefficient for
strongly sorbed or slightly soluble solutes varied with the concentration
of the soil/sediment used in the experimental work. They proposed that the
observed change in partitioning behavior due to solids concentration could
be attributed to a transfer of sorbing, or solute binding, material from
the solid phase to the liquid phase during the course of the partitioning
experiment. This material, whether dissolved, macromolecular, or
microparticulate in nature, was not removed from the liquid phase during
the separation procedure and was capable of stabilizing the compound of
interest in solution. The amount of material contributed to the liquid
phase was thought to be most likely proportional to the amount of solid
phase present, and thus the capacity of the liquid phase to accommodate
solute would depend upon the concentration of solids in the system. The
overall effect can be viewed either as a two-phase system; where the
properties of one phase (liquid) vary with the mass of the other (solids),
or as a three-phase system consisting of water, solids, and a third phase
that is not separated from the water but possesses a higher capacity for
the solute than the water itself. This is the colloidal phase.

Measurements of "dissolved" sorbing phase (weight of dissolved solids,
turbidity, and dissolved organic carbon) demonstrate the increased loading
of nonsettling microparticles or macromolecules in the supernatants of
batch equilibrium experiments as the solids-to-water ratio increases. It
is clear that nonsettling microparticles or macromolecules vary regularly
with suspended solid concentration.

The observation that dissolved colloidal material was increasing the
apparent solubility of analytes in laboratory studies led to the attempt
to wash the sediment to try to remove these materials. Successive washings
reduced the amount of material in solution, but failed to remove it.
After five successive washes, the nonsettling microparticles or
macromolecule content dropped about an order of magnitude, yet remained at
an amazingly high level of 100 mg/l even after five washes (GSCHWEND and
WU, 1985). WALTERS et al.,(1989) confirmed the report that aqueous
colloids couldn't be removed satisfactorily by washing or centrifugation.

That this occurs should not be particularly surprising. Particle size
distributions of natural sediments and soils are undoubtedly continuous
and do not drop to zero abundance in the region of typical centrifugation
or filtration capabilities. Additionally, there is some evidence to
indicate that dissolved and particulate organic carbon in natural waters
are in dynamic equilibrium, causing new particles or new dissolved
molecules to be formed when others are removed. Experiments with soil
columns have shown that natural soils can release large quantities of
dissolved organic carbon into percolating fluids (GSCHWEND and WU, 1985
and references therein).

In work by BEASLEY and JENNINGS (1984), transport in the solution phase
was indicated in the removal of radionuclides from the Columbia River.
This river received apparently substantial inputs of radioactive materials
as it was used to cool single-pass reactors engaged in weapons
production. An inventory of the river was made to determine the
radioactivity of the buried sediment. It was observed that the sediment
contained levels of radionuclides which were much lower that what was
expected from accepted partition models based on a two-phase system. It
was suggested that the removal of these contaminants was accomplished in
the absence of any major erosional events; hence the transport was assumed
to have been accomplished in the solution phase.

It has been observed (ALBERTS and WAHLGREN, 1977) that storm-caused
turbidity in lake Michigan was associated with elevated levels of
plutonium and americium in finished drinking water. The implication is
that radionuclides that are strongly sorbed to the sediment are
re-suspended into the water as the sediment is stirred by wave action.

The apparent solubility of many radionuclides is enhanced through
complexation with naturally occurring humic and fulvic acids (NELSON et
al.,1985), or through electrostatic binding to colloid clay, metal oxides
or other inorganic colloids.

Colloids have been repeatedly shown to be important in enhancing the
apparent solubility of hydrophobic organic chemicals (e.g., HASSETT and
ANDERSON, 1979, 1982; MEANS and WIJAYARATNE, 1982; VINTEN et al.,1983;
HASSETT and MILICIC, 1985; WHITEHOUSE, 1985 and others).

The solid phase is the source of dissolved or suspended colloidal material
that is acting as the third phase. It is observed that the solution phase
is in dynamic equilibrium with the solid phase (soil/sediment).

The discussion of colloid formation requires a description of the material
from which they are formed. To this end, a brief discussion of the nature
of the inorganic and organic components of the natural world will be
undertaken to provide a background and reference to the discussion of the
formation and behavior of colloids.

CLAY FRACTION
Much of the early work in characterizing the environmental behavior of
chemicals was accomplished in the area of agricultural chemistry. Work
surrounding the behavior of plant nutrients in the soil has provided a
large base of information about the processes of environmental chemistry.
Workers investigating the effectiveness of soil-applied herbicides
determined that the herbicidal activity of organic chemicals varied with
soil properties. It was determined the clay fraction and the organic
matter content of the soil were related to the ability of a soil to
diminish the effectiveness of an organic herbicide applied to the soil.

The clay fraction, which has long been considered as a very important and
chemically active component of the soil, has both textural and mineral
definitions.

In its textural definition, clay is generally assumed to be the mineral
fraction of the soil that is smaller than about 0.002 mm in diameter. The
small size of clay imparts a large surface area for a given mass of
material. This large surface area of the clay textural fraction of the
soil makes it very important in processes involving interfacial phenomena
such as sorption or surface catalysis.

In its mineral definition, clay is composed of secondary minerals such as
carbonate and sulfur minerals, layer silicates and various oxides. Layer
silicates are perhaps the most important component of the clay mineral
fraction.

Because of isomorphic substitution of ions in the crystalline lattice of
layer silicates, many clay surfaces have a net negative charge which
results in the ability of such minerals to exchange cations from the soil
solution.  The cation exchange capacity varies from about 3 to 200
meq/100g (e.g., THOMPSON and TROEH, 1973). In addition to the existence of
a static charge on the clay surface resulting from intracrystalline charge
imbalances (isomorphic substitution), soil minerals may acquire charge
from the pH-influenced dissociation of surface hydroxyl groups. The
magnitude of this type of cation exchange capacity will tend to increase
with pH.

The sphere of influence or extent of impact of the charged clay surface on
the structure of ions of the solution will be, to some extent, determined
by the ionic strength of the solution according to the double-layer
effects discussed above.

In addition to the presence of phyllosilicate minerals which exist in
crystalline layers and frequently possess a net surface charge from
isomorphic substitution, other products of mineral weathering and
dissolution may be present in the clay fraction which do not exist in
layers and do not possess an intrinsic charge. These are sometimes called
accessory minerals and some of these minerals, such as allophane, may have
pH dependent charge. These minerals may exist as uncharged oxides,
hydroxides and hydroxyoxides of aluminum, iron and titanium. Finely
divided grains of these accessory minerals coat the surface of other
mineral grains in the soil.

The results of mineral weathering form particles with a size continuum
from ions to grains. Mineral dissolution and precipitation occur more or
less continuously as a function of ambient conditions. Particles of the
clay textural fraction may be suspended in solution as colloids as well as
existing as part of the stationary solids.

For example, iron in groundwater is often present both in solution and as
suspended ferric oxyhydroxides. A significant percentage of the iron in
many groundwaters (30-70% is not uncommon) is present as suspended ferric
oxyhydroxides. Solubilities of the oxyhydroxides vary greatly depending on
such factors as dissolved Fe (II), initial precipitation pH, the base used
for hydrolysis, surface area of the precipitate, and time. For a given Eh
and pH, waters in equilibrium with freshly precipitated amorphous material
can contain approximately seven orders of magnitude more dissolved iron
than those in equilibrium with coarsely crystalline goethite or hematite
(WHITTEMORE and LANGMUIR, 1975).

It is reasonable to assume that clay colloids would exhibit similar
surface chemistry to that clay sorbed, bonded or precipitated in the
stationary solid phase. Mineral colloids may be formed when precipitation
or dissolution forms particles which are resistant to settling. These
particles may be formed by any number of conditions whereby the solubility
of a particular solute is exceeded (e.g., WHITTEMORE and LANGMUIR, 1975;
GSCHWEND and REYNOLDS, 1987) or a stable solid is disrupted mechanically
or chem ically (NIGHTINGALE and BIANCHI, 1977; BUDDEMEIER and HUNT, 1988)

The composition of the mineral fraction of the soil, being extensively
composed of oxygen and silicon bonded with various metals, will lend a
relatively polar nature to the surface of most of the inorganic soil
components. This polar/ionic nature will cre ate a natural affinity
between soil and ionic or polar solutes.

In addition to sorption of ionic and polar solutes onto clay in the
solution and solid phases, the clay fraction has been shown to be
important in the sorptive behavior of neutral hydrophobic organic
compounds from the water (KHAN et al.,1979; HASSETT et al.,1981). The
large surface area of the clay fraction offers a large sorptive interface
upon which hydrophobic bonding may occur (MINGELGRIN and GERSTL, 1983;
BANERJEE P. et al.,1985). For such nonpolar compounds, polar/ionic
attraction is generally secondary to hydrophobic effects in sorption on
most sediments and soils.

ORGANIC MATTER PROPERTIES
It appears that for very hydrophobic molecules, the organic carbon content
of the sorbent is of greater importance than the mineral surface itself
(KHAN et al.,1979). However, in the low carbon environments characteristic
of the subsurface, mineral sorption may play an important role in
affecting hydrophobic pollutant mobility (BANERJEE P. et al.,1985;
McGINLEY et al.,1989). It has been reported that if the soil contains more
than about 0.2% organic carbon, all of the sorption of hydrophobic
organics appears to be due to the organic carbon. If the solid phase
contains less than about 0.2% organic carbon, the sorption of hydrophobic
chemical from the aqueous phase may be attributed to the clay fraction
(BANERJEE P. et al.,1985).

In work with sorption of aromatic hydrocarbons by sediments, it was
reported by KARICKHOFF et al.,(1979) that when the individual partition
coefficients for the sorption of the compounds (Kp) were divided by the
organic carbon contents of the sediments (% OC), a unique constant, (Koc)
was generated that was independent of sediment properties and dependent
only upon the nature of the organic analytes.

                       Koc = Kp/(%OC)  (14)

These authors reported a significant correlation between the Koc values
obtained from the sorption of organic compounds on three local sediments
and the partition constants (Kow) for the partitioning of the compounds
between octanol and water.

                       log Koc = 1.00 log Kow - 0.21   (15)

There have been a number of empirical expressions of this sort developed
relating the partitioning behavior of a chemical between water and organic
carbon to the octanol-water partition coefficient for the chemical (e.g.,
MEANS et al.,1980; KARICKHOFF, 1 981,1989; SCHWARZENBACH and WESTALL,
1981). It has been noted that the tendency of a chemical to partition into
the organic phase of the soil or sediment is inversely related to the
water solubility of the chemical (CHIOU et al.,1979,1982). Hence, the ten
dency of an organic chemical to be sorbed by soil or sediment organic
matter will be a function of its hydrophobicity. The octanol-water
partitioning behavior of a material has also been related to its intrinsic
hydrophobicity. Organic solids with high melting-points are reported to
behave anomalously in such considerations (BANERJEE et al.,1980; MACKAY et
al.,1980).

The organic fraction has been determined to be of considerable importance
in the environmental behavior of pollutants. The importance of organic
matter in the processes of pollutant partitioning warrants some brief
description of this material as a preface and background for discussing
some of the properties it displays which influence the apparent solubility
of water contaminants.

HUMUS
Humus is considered to be the remains of living things that are no longer
visually recognizable as to their origins. The physical nature of humus is
that of an amorphous, brownish material with a density somewhat lower than
that of mineral soil. In its natural state, humic material is somewhat
variable in composition and form.

The processes of formation of organic matter are more or less unique to a
particular geographic environment on the large scale. Since the material
is derived from plant remains which have been more or less degraded by
detritovores, it is reasonable to assume that there would be some
differences between humic material from different biomes. The vegetation,
soil minerals, climate, and microbial population are some of the variables
which might act to create differences in the organic matter of a different
area. In spite of reported differences, the variations are not so great
as to preclude comparisons between humus from one biome to another. The
process of plant growth is essentially one of photosynthesis; i.e., the
sunlight-driven reduction of oxidized carbon. The reduced carbon takes
many forms and combines with many elements in a very complex array of
chemicals. When this living material dies, the chemical energy it contains
is exploited by heterotrophic organisms for their own life processes.
Humus is considered to be produced from that portion of the reduced
carbon which was resistant to degradation either as a function of
intrinsic nature (e.g., polyphenols such as lignins and tannins) or of
ambient conditions which restrict the oxidative processes of degradation
(e.g., cold, anaerobic environments).

In addition to the reported differences between and within bioregions, it
is reported that humic material of terrestrial origin is different from
the humic material in freshwater streams. The humic material in aqueous
sediments is reported to be less polar (as judged by a lower oxygen
percentage) than the soil organic matter (LEE et al.,1981).

Humus undergoes changes as it ages. The humus which exists in the soil is
the result of extensive alteration of the original component materials and
is subject to degradation. Under different conditions, humus undergoes
diagenesis and transformation in response to the ambient conditions.
Humus buried deep in the subsurface is subject to different forces and
will be accordingly different after the passage of time. Coal, peat and
oil are examples of organic matter that has undergone extensive
diagenesis. Sometimes organic material undergoes diagenesis with earth
minerals to form mineral-organic composites such as oil-shale.

Diagenesis is reported to increase with depth and time of burial.
Maturation is the result of mild heat and pressure; it is possible that
interactions with mineral surfaces and complexed metals are also involved
(JACKSON, 1975).

Thermodynamic stabilization occurs in diagenesis. The least stable and
most reactive components or their substituents are gradually eliminated.
This process leads, with increasing age and depth of burial, to a gradual
stabilization, not necessarily of each individual compound but the
sedimentary organic matter as a whole. In terms of structures the
transformation of open chains to saturated rings and finally to aromatic
networks is favored; hydrogen becomes available for inter or
intramolecular reduction processes. Eventually, highly ordered, stable
structures of graphite may be formed. It is pointed out the most
characteristic feature of organic diagenesis is the appearance of extreme
structural complexity and disorder at an intermediate stage, interposed
between the high degree of biochemical order of the starting material, and
the even greater crystallographic order of graphite, the end product of
diagenesis (STUMM and MORGAN, 1981).

The long-recognized complexity of organic matter has generally confounded
accurate and detailed description of the material and has instead spawned
qualitative divisions of the natural material which have been adopted by
workers in the field to allow for some agreement on methodology at least.
The nature of humus has been studied exhaustively and, in spite of some
conflicting reports, a number of points have been agreed on; some of these
will be related as they apply to the material at hand.

The humic substances are organic polyelectrolytes which are most commonly
identified with the organic material present in soils. However, humic
substances are also present in practically all of the suspended and bottom
sediments of rivers, lakes and estua ries. Humic materials are also
apparently soluble in water and are present in surface and groundwaters.

This group of compounds enters into a wide variety of physical and
chemical interactions, including sorption, ion exchange, free radical
reactions and solubilization. The water holding capacity and buffering
capacity of soils and the availability of nutrients to plants are
controlled to a large extent by the amount of humus in the soil. Humus
also interacts with soil minerals to aid in the weathering and
decomposition of silicate and aluminosilicate minerals. It is also
adsorbed by some soil minerals.

The chemical nature of humus is the subject of variable and sometimes
conflicting reports in the literature. ALEXANDER (1977) sums up the issue
by stating that "humus should be considered as a portion of the soil that
is composed of a heterogeneous group of substances, most having an unknown
parentage and an unknown chemical structure".

For the purposes of this technical appendix, it is important to point out
that humus is chemically reactive and has variable chemistry, manifesting
both polar and nonpolar tendencies. In general, humus contains a number of
chemical functional groups assoc iated with a polycyclic aromatic matrix
of varying size.

In term of types of compounds, humus contains a number of polymerized
substances; aromatic molecules, polysaccharides of several kinds, bound
amino acids, polymers of uronic acids and phosphorous containing
compounds. Chemical degradation has shown that the basic building blocks
of humic acids are benzene carboxylic acid groups, substituted phenolic
groups and quinone groups.

KHAN and SCHNITZER (1972) proposed that humic material exists in a
molecular sieve or clathrate (lattice-like) structure joined by polar
bonding mechanisms. They proposed this because they observed that the
material would sorb hydrophobic organic compounds and not release them to
extraction with organic solvents. Methylation of the mixture enabled
extraction of the sorbed analyte. They supposed that the methylation step
disrupted hydrogen bonding within the humic material.

It is worthwhile to discuss several of the more likely possibilities of
mechanism of aggregation of humic materials. The mechanism or mechanisms
that bring about the aggregation or disaggregation of humic substances
will be determined by charge distributi on and functional-group
distribution on the exposed surfaces. A number of workers have shown that
humic materials contain abundant polar functional groups. The highly polar
nature of some of the functional groups makes dipole bonds and hydrogen
bonds probable active mechanisms of structural change (e.g., WERSHAW and
PINKNEY, 1973).  Reports that humic materials contain both electron rich
and electron deficient sites gives evidence that polar bonding will be
likely to occur (MELCER et al.,1989).

In addition to the hydrogen bonding, coulombic attraction of charged
particles will also create bonds in humus. The charged sites on a
polyelectrolyte molecule may arise in several different ways. Ionic
compounds will dissociate in solution, producing molecules with charged
sites. These charged sites may also result from charge�transfer reactions
such as the transfer of an electron from a carbanion or a radical anion to
another molecule (WERSHAW and PINKNEY, 1973).

Humus is also capable of forming covalent bonds with aqueous solutes
(PARRIS, 1980). Humus is the site of considerable microbial activity.
Living and dead organisms and extracellular enzymes are typically
associated with humus as part of the material (e.g ., ALEXANDER, 1977).
The presence of enzymes can catalyze reactions. It is also reported that
humus contains stable free radicals which make it very reactive and able
to form covalent bonds or create ions. It is reasonable to assume that
charge-transfer reactions between free radicals are important in the
aggregation of humic materials in light of the high concentrations of free
radicals that have been detected in both soils and aqueous humic
preparations. The free radicals detected in soils and humic acids may
arise from the reduction of a diamagnetic molecule by a solvated electron,
enzymatic reactions or photolysis (WERSHAW and PINKNEY, 1973).

The diverse nature of chemical bonding arrangements exhibited by humus
enables the formation of associations both with nonhumic materials and
with other humic materials to create a dynamic structure in what WERSHAW
and PINKNEY (1973) term a "living polymer". Such a system is capable of
undergoing inter and intra molecular bonding to add or lose constituents
or change configuration in response to ambient conditions. The chemically
diverse and highly reactive nature of the humic matrix imparts the ability
of humus to both lose and acquire molecular moieties in a dynamic manner.

A fractionation procedure has been derived and widely applied to studies
of humic material. The procedure begins with natural organic matter
(humus) and uses a basic solution to solubilize a fraction of the
material. The basic extract is then acidified which causes a precipitate
to form. The precipitate is named humic acid. The fraction which remains
in solution is called fulvic acid. Humin is the name given to the
insoluble fraction that remains after extraction of humic and fulvic
acids.

At near-neutral pH, which is characteristic of most natural water, the
fulvic acid is the most water soluble of these three fractions. Humic acid
is somewhat less soluble, with its solubility increasing as the pH
increases. Humin is normally considered in soluble at all pH values.

Fulvic acids are soluble in water and so are the majority of the salts of
these acids (e.g., OGNER and SCHNITZER, 1970b). The aquatic fulvic acid
fraction contains substances with molecular weights ranging from 500 to
2000 and is monodisperse. Aquatic fulvic acids are dissolved rather than
colloidal (THURMAN et al.,1982).  Fulvic acid has been found to contain
branched, cyclic and linear alkanes, as well as fatty acids (OGNER and
SCHNITZER, 1970b). It can combine with insoluble organic compounds such as
alkanes, fatty acids and dialkyl phthalates to form stable "complexes"
that are soluble in water (OGNER and SCHNITZER, 1970a; MATSUDA and
SCHNITZER, 1971, BOEHM and QUINN, 1973).

Humic acid is pictured as being made up of a hierarchy of structural
elements. At the lowest level in this hierarchy are simple phenolic,
quinoid and benzene carboxylic acid groups. These groups are bonded
covalently into small particles.

The molecules of humic acid are reported to be nonspherical, or more
probably, nonspherical and hydrated (PIRET et al.,1960). Other workers
have reported that humic materials are rigid spherocolloids in solution
(KONONOVA, 1961; and references contained therein). Work by GHOSH and
SCHNITZER (1980) makes the overall conclusion that the configurations of
humic and fulvic acid molecules are not unique; they vary with changes in
the environment. These authors report that both humic and fulvic acid
molecules are flexible linear colloids at low concentrations, provided
hydrogen ion and neutral salt concentrations are not too high. As these
factors increase, the macromolecules assume coiled configurations similar
to those of uncharged polymers or rigid spherocolloids. It has been
reported by KONONOVA (1961) (and references contained therein) that humic
acids may have an amorphous structure and furthermore that the size and
weight of the humic acid molecules may vary as a function of ambient
solution conditions.  It has been postulated that the molecular weight of
the humic species may vary from 1000 to 50,000 (PIRET et al.,1960). They
may consist of particles capable of aggregation or dissociation.

Humic acids are larger than fulvic acids and form polydisperse systems
(THURMAN et al.,1982). Precipitation is used to isolate humic acid from
soil. The humic acid must aggregate to precipitate, therefore, it gives
polydisperse systems showing that it exists as aggregates of various
sizes. WERSHAW et al.,(1977) contend that humic acids are a mixture of a
limited number of more or less chemically distinct fractions of relatively
low molecular weight that form molecular aggregates in solution. Particles
of similar chemical structure are thought to be linked together by weak
bonds to form "homogeneous" aggregates. Two or more different types of
aggregates may be linked together to form mixed aggregates. While there is
no general agreement on the matter, some workers in the field contend that
humic acid is really simply an aggregate of common humic materials and so
is chemically similar to fulvic acid and to humin (WERSHAW et al.,1977).

The distinctions of humic and fulvic acids and humin are purely
operational definitions. WERSHAW (1990) termed the distinction between
humic and fulvic acids and humin "a totally artificial division that tends
to obscure the close interaction between the organic constituents of
natural water systems"

Humus can exist in solution as well as in the solid phase. The behavior of
water-soluble humic materials is of great relevance to the discussion of
solubility enhancement of aqueous pollutants, both organic and inorganic.

Freshwater aquatic humic substances originate from soil humic material and
terrestrial and aquatic plants. In surface waters these compounds
generally account for 30 to 50% of the dissolved organic matter (THURMAN
et al.,1982).

Systems that contain naturally high levels of dissolved organic matter
include bogs, swamps, and interstitial waters of sediments. Interstitial
water (porewater) is formed by the entrapment of water during
sedimentation, which isolates it from the overlying water. Porewater is
considered to be in equilibrium with the sedimentary solid phase and
separate from the overlying water column, or bulk water. In high carbon
sediments, dissolved organic carbon in porewater can exceed 100 mg/L,
whereas overlying waters typically contain less than 5 mg/L of dissolved
organic carbon (THURMAN and MALCOLM, 1981; CARON and SUFFET, 1989).

The molecular weight of most of humic substances in water is less than
10,000. Although some studies have found humic substances with molecular
weights greater than 100,000 (LEE et al.,1981). The ability of humus in
solution to form extensive aggregates w as discussed above.

INTERACTIONS BETWEEN ORGANIC MATTER AND ORGANIC POLLUTANTS
It is suggested that binding to dissolved humic materials could
significantly affect the environmental behavior of hydrophobic organic
compounds. The rate of chemical degradation, photolysis, volatilization,
transfer to sediments, and biological uptake may be different for the
fraction of pollutant that is bound to dissolved humic materials (CARTER
and SUFFET, 1982). If this is the case, the distribution and total mass of
a pollutant in an ecosystem would depend, in part, on the extent of humic
material-hydrophobic binding.

Agricultural chemists found that organic herbicide activity was inversely
related to soil organic matter content (e.g., UPCHURCH and MASON, 1962;
LAMBERT et al.,1975). The volatility of organic pesticides was found to be
diminished in the presence organic colloids isolated from soil organic
matter (PORTER and BEARD, 1968). The relationship between the organic
matter content and the behavior of organic chemicals in soils and
sediments has been documented extensively (e.g., KHAN and SCHNITZER, 1972;
KARICK HOFF et al.,1979; MEANS et al.,1980; BROWN and FLAGG, 1981). It is
commonly reported that solutions of soil or sediment-derived organic
matter increase the solubility of hydrophobic organic chemicals (e.g.,
WERSHAW et al.,1969; BALLARD, 1971; HASSETT and ANDERSON, 1979; OGNER and
SCHNITZER, 1970a; MATSUDA and SCHNITZER, 1971). The presence of dissolved
organic matter in sorption studies will be manifest as a decreased
soil/sediment partition coefficient at equilibrium, because the humic
material in solution competes with the stationary solid phase for
sorption of the analyte. Other workers have reported similar observations
of enhanced solubility by organic carbon in the bulk aqueous phase (e.g.,
BOEHM and QUINN, 1973; CARTER and SUFFET, 1982; LANDRUM et al.,1984; HAAS
and KAPLAN, 1985; GSCHWEND and WU, 1985; and others).

Pollutants may be bound to humic materials through abiotic or biological
processes whereby the formation of bound residues usually results in
detoxification of the pollutant. Therefore, enhancing the binding of
xenobiotic chemicals to humic materials can serve as a means to reduce
toxicity as well as migration of the toxic compounds.  Complex formation
can occur by an oxidative coupling reaction leading to oligomeric and
polymeric products. BOLLAG and BOLLAG, (1990) report the effect of
phenoloxidazes (peroxidases, tyrosinases, and lacases) on the binding of
substituted phenols and aromatic amines to humus monomers as well as to
humic substances. Copolymerization largely depends on the chemical
reactivity of the substrates involved. Certain phenolic humus
constituents, such as guaiacol or ferulic acid are highly reactive in the
presence of phenoloxidases. When one of these compounds was incubated
together with a phenoloxidase with less or even non-reactive phenols,
anilines or other chemicals, a synergistic reaction took place, resulting
in increased formation of bound residues of these compounds.
Phenyloxidases are able to catalyze the polymerization and/or binding of
numerous organics to humic constituents.  The inclusion of phenolic humus
constituents, such as syringic acid, vanillic acid or vanillin, resulted
in the enzyme induced formation of various cross coupling products. A wide
variety of xenobiotics can become cross coupled to naturally occurring
humic monomers by the action of phenyloxidases.  These xenobiotics include
phenols such as various mono-, di- and tri- substituted chlorophenols and
2,6-xylenol, and anilines such as 4-chloroaniline, 3,4-dichloroaniline and
2,6-diethylaniline.  This enhanced removal of a xenobiotic is by no means
unique to fulvic acid. It has been shown that the addition of a highly
reactive humic monomer, such as syringic acid, to a phenoloxidase
containing system can initiate the effective polymerization and/or binding
of a molecule which by itself is only poorly transformed, if at all.
There seems to be no shortage of examples of enzyme induced polymerization
and binding of xenobiotics. It is thought that the enzyme induced
oxidation of naturally occurring phenols yields free radical quinonoid
structures. This is a common pathway in the phenoloxidase catalyzed
polymerization and binding of both naturally occurring and man made
compounds. Another pathway is the decarboxylation of a highly reactive
compound such as sytingic acid and the formation of a covalent bond at
that site to generate phenolic oligomers. Binding of a pollutant to humic
acids, clays, or other materials would be expected to decrease its toxic
effects. Binding can reduce the amount of a compound available to the
biota and as the quantity of an available xenobiotic is reduced, toxicity
also declines.

The distribution of hydrophobic organic pollutants between sediments and
water has typically been viewed as a surface adsorption phenomenon and, as
such, has been studied with batch sorption isotherm techniques. Adsorption
isotherms of nonpolar organic co mpounds on a number of soils and
sediments are linear over a wide range of equilibrium solute
concentrations (e.g., CARON and SUFFET, 1989).

If the mechanism of solubility enhancement of hydrophobic organics is one
of surface sorption, it might be expected that partition coefficients of
aquatic humic substances may be less than those of the organic matter on
particles, since macromolecules in solution must be relatively hydrophilic
(GSCHWEND and WU, 1985). This view is supported by the reports describing
heteroatom compositional differences between fulvic and humic acids
recovered from natural water. The smaller, more water soluble fulvic
acids have higher oxygen-to-carbon ratios compared to the larger humic
compounds (THURMAN and MALCOLM, 1981). Thus, smaller, more water soluble
macromolecules may be expected to be more polar sorbents (i.e., exhibit
relatively lower Koc's) than related larger macromolecules and
particulate matter.

It has been proposed that the association between nonpolar compounds and
the organic carbon fraction of sediments, soils, and natural waters is
better described as a liquid-liquid partitioning phenomenon than as a
surface adsorption process (CHIOU et al., 1979, 1983; KILE and CHIOU,
1989a). An organic matter partitioning process is supported by a number of
observations, including;
- linear sorption isotherms to near aqueous saturation concentrations of
nonpolar organic substances, with no evidence of isotherm curvature at the
higher concentration range;
-isotherm curvature at higher concentrations is predicted by adsorption
theories ;
- small temperature effects on solute sorption;
- absence of competition in experiments using binary solute systems;
- data covering seven orders of magnitude in which sediment-water
partition coefficients were inversely proportional to aqueous solubility
and well correlated to octanol-water partition coefficients.

CHIOU et al.,(1987) considered the mechanism for water solubility
enhancement of nonionic organic solutes by dissolved organic matter of
soil and aquatic origins. Such enhancement effects were effectively
explained in terms of a partitionlike interaction of solutes with
dissolved high molecular weight humic materials on the basis of the
properties of the solutes and humic materials. The observed solubility
enhancement of the solute by dissolved organic matter (DOM) can be
expressed by

                       Sw* = Sw(1 + X KDOM)    (16)

where
Sw*     = apparent water solubility in the solution
Sw      = apparent water solubility in pure water
X       = concentration of dissolved organic matter
KDOM    = partition coefficient between DOM and water

The difference in values of KDOM for a solute with different types of
fractionated humic materials has been explained in terms of the polarity,
molecular size, and molecular configuration of the humic materials (based
on elemental data analysis) gives a reasonable estimate of the relative
enhancing effects among humic extracts.

Conceivably, the compositions and structures of humic materials in
different aquatic systems can be significantly different because of such
environmental factors as water pH, biological processes, and the presence
of other chemical species that affect the concentration (solubility) of
humic materials. In more acidic streams or rivers, there appears to be a
tendency for the humic material to contain a larger percentage of oxygen
compared to samples from a neutral or basic water. A decrease in oxygen
content of the humic materials from acidic to neutral water can also be
accompanied by an increase in carbon content. The solubility enhancement
effects of individual humic samples appear to be closely correlated with
the polarity of the materials (using elemental data as approximate
indices), suggesting that differences in molecular sizes of humic
materials are not as much a critical factor as the polarity in affecting
the partition interaction with organic solutes (CHIOU et al.,1987).

It is supposed that the solubility enhancement cannot be explained by the
cosolvency theory forwarded by Yalkowsky and others (e.g., YALKOWSKY et
al.,1972,1975; AMIDON et al.,1974) because the magnitude of solubility
enhancement is greater than that which would be predicted from cosolvent
effects alone. This was investigated by CHIOU et al.,(1986) who used
phenylacetic acid, synthetic organic polymers (poly(acrylic acid)) and
dissolved humic and fulvic acids to assess the solubility enhancement
effects on different chemicals. They found significant solubility
enhancements of relatively water-insoluble solutes by dissolved organic
matter of soil and aquatic origins. The concentrations of the humic
materials varied from 0 to 94 ppm. They observed that the apparent solute
solubilities increased linearly with dissolved organic matter
concentration and showed no competitive effect between solutes. With a
given dissolved organic matter sample, the solute partition coefficient
increased with a decrease of the solute's water solubility or with an
increase of the solute's octanol-water partition coefficient. The
partition coefficient values of solutes with soil-derived humic acid were
approximately 4 times greater than with soil fulvic acid and 5-7 times
greater than with aquatic humic and fulvic acids. The effectiveness of
dissolved organic matter in enhancing solute solubility appeared to be
largely controlled by the molecular size and polarity of the material.

The organic acid and polymer (molecular weight varied in separate
experiments from 2000 to 90,000) created no observable solubility
enhancement. The investigation of phenylacetic acid as a cosolute, with
the concentration exceeding 600 mg/l, shows slight enhancement for DDT,
which was the most hydrophobic analyte in the experiment. The magnitude of
DDT solubility enhancement per unit mass of phenylacetic acid was much
smaller than with the humic or fulvic acids.

They found that the solubility enhancement exhibited by the dissolved
humic material may be described in terms of a partition-like interaction
of the solutes with a "microscopic nonpolar organic environment"
associated with the high-molecular weight humic species. The relative
inability of high-molecular-weight poly(acrylic acids) to enhance solute
solubility was attributed to their high polarities and extended chain
structures that do not permit the formation of a sizable intramolecular
nonpolar environment.

This observed "partition-like" interaction between hydrophobic organic
solutes and dissolved humic material has led to the proposition that humic
micelles may exist in solution. WERSHAW (1986) has proposed an elaborate
model for this micelle structure. In this model, humic materials are
pictured as existing as membrane-like aggregates which are made up of
partially decomposed plant-derived compounds, which are held together in
the aggregates by weak bonding mechanisms, such as pi bonding, hydrogen
bonding and hydrophobic interactions. The humic membrane-like structure
consists of polar hydrophilic exterior surfaces with hydrophobic
interiors. Polar compounds will interact with the exterior polar groups of
the humic structures, while hydrophobic compounds will partition into the
hydrophobic interiors of the structures.

This model is consistent with much of the reported information in the
literature, especially with regard to the "membrane-like" behavior. Such a
structure might explain the report of HASSETT and ANDERSON (1979), wherein
it was reported that the solubility of cholesterol was enhanced by
high-molecular-weight dissolved organic matter in river water. These
authors found that solvent extraction of the radiolabeled cholesterol was
ineffective as a means of recovery unless the organic matter was destroyed
by U V radiation. The existence of a polar interface like the outside of a
"membrane-like" structure, as proposed by Wershaw (1986), would explain
the inability of a nonpolar solvent to recover the cholesterol. It might
be envisioned that such a polar or ionic region might be a zone of high
interfacial tension which would be relatively inhospitable to transversal
by nonpolar molecules. Similar results were reported by FISH et al.(1989),
who observed that solvent recovery of sorbed hydrophobic organics was
enhanced by a digestion technique which degraded the dissolved organic
matter. Simple adsorption onto a the nonpolar region of humic molecules by
van der Waals forces and hydrophobic interfacial tension would probably
not impede solvent recovery of adsorbed hydrophobic organics.

Acceptance of the micelle model for dissolved humic material has been slow
to arrive (e.g., MINGELGRIN and GERSTL, 1983; MaCINTYRE and SMITH, 1984;
MACKAY and POWERS, 1987; and others).  PIRET, et al.,(1960) reported that
peat-derived humic acid had a critical micelle concentration of about 18
g/L. Since most of the observed solubility enhancement has been associated
with humic concentrations orders of magnitude lower, it has been assumed
that micelles were not present. More recent work with micelle formation
has indicated that molecular aggregate formation may occur below the
critical micelle concentration with molecularly nonhomogeneous
surfactants. Natural humic materials may be considered as molecularly
heterogeneous amphiphiles.

It has also been shown (KILE et al.,1990) that a commercial surfactant
which consists of a diverse admixture of monomers (made by reacting
petroleum with concentrated sulfuric acid) does not exhibit behavior
typically associated with micelle formation i.e ., a sharp inflection of
solvent properties as the concentration of surfactant reaches CMC. These
surfactants exhibit gradual change in solvent behavior with added
surfactant. It is proposed that this gradual solubility enhancement
indicates that micelle formation is a gradual process instead of a single
event i.e., CMC does not exist as a unique point, rather it is a
continuous function of molecular properties. This type of surfactant may
construed as more realistic representation of humic material in water and
may indicate that dissolved humic substances form molecular aggregates or
colloids in solution; which comprise a third phase in aqueous environment.
This phase will have the effect of increasing the apparent solubility of
very hydrophobic chemicals. Polar solutes will not be noticeably affected
by these colloids because they are already pretty soluble.

These conclusions are consistent with the observations of BOEHM and QUINN
(1973) who observed that the solubility of hydrophobic hydrocarbons was
increased by humic material dissolved in sea water. There was evidence to
suggest that the mode of solubilization of the hydrocarbons was by
incorporation into micelles formed by intermolecular association of the
surface active humic-type monomers. The solubilized hydrocarbons were
determined to exist in a semicolloidal or micellar state formed by
interaction of humic-like monomers in solution. The presence of salts in
solution appeared to be prerequisite for the formation of the aggregates
responsible for the solubility enhancement. The authors also observed that
addition of the hydrocarbon being solubilized appeared to lower the
concentration at which humic micelles formed.

HUMIC/MINERAL ASSOCIATIONS
In addition to the ability of humic substances to form associations with
hydrophobic organic species, humic material also reacts readily to form
associations with inorganic minerals and polar and ionic organic materials
as well. These sorts of association s are involved in colloid formation
with a wide variety of materials.

It is reported that thorium(IV) was strongly bound to colloidal humic
materials via metal-polyelectrolyte binding. This was determined to be
electrostatic in nature, with complex stability increased with increasing
ionization of the colloidal organic matter (NASH and CHOPPIN, 1980). The
dissociation of thorium bound to humate in aqueous solution has further
been studied, (CHOPPIN and NASH, 1981) and it is suggested that Th(IV) is
bound by at least four types of sites with different basicities and
different local polymer structure. CACHERIS and CHOPPIN (1981) reported
that thorium(IV) was bound to humic material in solution and reported that
the complex appeared to be characterized by two mechanisms with different
dissociation tendencies.

Fulvic acids can interact with clay minerals (OGNER and SCHNITZER, 1970a)
and are known to form stable complexes with metal ions and hydrous oxides
(e.g., JACKSON, 1975).

The operational technique of isolation of humic acid involves a pH-induced
precipitation and it is likely that accessory minerals may be associated
with the precipitation process. Complexes of humic acid and clay minerals
are also formed, the increased ash content of humic acid suggests that
amorphous silica and clay may aggregate with the humic acid fraction
(WERSHAW et al.,1977).

Humin is that part of humus which is not solubilized by alkaline solution.
It is thought to comprise an amorphous aromatic matrix interlinked by
strong bonds. Humin is the least polar of the commonly studied fractions.
Because of its extensive nonpolar ar omatic network, humin is probably the
best sink for hydrophobic organic chemicals (MANAHAN, 1989). Extraction of
the humin fraction of soils and sediments with methylisobutyl ketone
(MIBK) demonstrates that the humin fraction also appears to be composed of
several different components which can be separated by relatively gentle
techniques. The humin can be fractionated into a hydrophobic fraction and
a hydrophilic fraction. The hydrophobic fraction is white in color and it
appears to be a lipid-like material, being at least partially soluble in
solvents such as hexane, chloroform and benzene-methanol mixtures. This
lipid-like fraction may be a mixture of plant lipids. The hydrophilic
fraction is composed of two subfractions: an inorganic subfraction which
settles out with time and a brown organic subfraction which remains in
solution (WERSHAW, 1986). Alternate treatment of the humin fraction with
strong mineral acids and strong bases generally renders most of the humin
soluble in basic solutions. It was concluded by KONONOVA (1961), that
humins are humic acids that are bound to the mineral constituent of soils.

The amphiphilic nature of dissolved humic substances lends them the
ability to associate with both hydrophobic organics and polar or ionic
species (e.g., WERSHAW et al.,1977). Inorganic ions or mineral colloids in
solution will interact with the electrically active surface of humic
material in solution or in the solid phase according to the same bonding
forces which lead to the association between soil organic matter and the
soil mineral matrix. Humic matter in water is associated with various
metal ions, clays and amorphous oxides of iron and aluminum (e.g., DAVIS,
1982). In aqueous environments, oxide mineral surfaces are generally
covered with hydroxyl groups. Organic macromolecules can sorb onto these
surfaces both by ligand exchange and by van der Waals forces to create a
very strong association.

In his model of "membrane-like" humic micelles, WERSHAW (1986) points out
that metal ions cause aggregation of humic materials and suggests that
inorganic ions may be associated with such structures. There are frequent
reports of polyvalent cations causing humic material to aggregate (e.g.,
LEE et al.,1981; ARES and ZIECHMAN, 1988).The associations between humus
and mineral matter is manifest as residue upon ashing of humic materials.

Humic materials may be bound to the clay surfaces by the amino acids of
proteins. This binding is most likely due to electrostatic charge effects.
Metal ions could also act as bridges between the humic materials and the
clay mineral surfaces. Oxides with relatively acidic surface hydroxyls,
e.g. silica, do not react strongly with the organic matter (DAVIS, 1982)
via coulombic attraction, however van der Waals forces will certainly
exist between the large humic species and the silica surface.

Humic materials are polyfunctional macromolecules with a number of surface
groups capable of dissociating as ions. These groups will be more or less
dissociated to impart some measure of ionic character to the material.
This will tend to attract polar species such as water as well as
counterions (e.g., JACKSON, 1975; LYTLE and PERDUE, 1981; BARKER et
al.,1986).

Humus is frequently considered to be able to form stable complexes such as
chelates with polyvalent cations. Soil organic matter is capable of strong
polydentate binding to transition metals in a chelate (e.g., DALANG et
al.,1984; BOHN et al.,1985).

The speciation of trace metals in natural waters is controlled by the
interaction of the metals with a complex and varying mixture of inorganic
anions, organic ligands, reducible or oxidizable dissolved chemical
species, reactive surfaces, and organisms.  Filterable concentrations of
metals may include fine colloidal particles as well as organic and
inorganic metal complexes (HERING and MOREL, 1990). In many natural waters
high concentrations of colloidal organic material are frequently
associated with high concentrations of iron (KNOX and JONES, 1979).

Humus has been reported to act as a reducing agent. VISSER (1964) measured
the formal redox potentials of neutral humic acid derived from tropical
sphagnum peat and reported that the mean value of normal potentials was
between +0.32 and +0.38 volts and decreased with increasing depth. More
recent measurements in different soil reports that humic material is an
active redox system with an E value of + 0.70 V (MANAHAN, 1989). Sometimes
humus will reduce a metal and then bind it in a polydentate complex. For
example, Cr2O7-2 is reduced by humic acids to chelatable cationic Cr3+
(MANAHAN, 1989).  DOUGLAS and QUINN (1989) showed a very stable complex
with chromium(III) and humic materials in reduced sediment. The
accumulation of vanadium and molybdenum in peats has been attributed to
reduction and chelation of soluble oxides of these elements. The reduction
of iron(III) to iron(II) and subsequent retention in the reduced form has
been demonstrated on the oxidized surface of coal (MANAHAN, 1989).

In summary, the chemical and structural nature of humus makes it very
active in the environmental behavior of many types of pollutants. The
presence of bound enzymes and free radicals in the material allows it to
form covalent bonds with a variety of molecules. The existence of
nonpolar regions of the humus introduces the possibility of intramolecular
sorptive partitioning of nonpolar organic materials into the humic matrix.
The extent and polarizability of the humic surface enable it to bind to
materials by the van der Waals force. The existence of electrostatic
charges on the surface of the substance make it reactive with respect to
water, ions and mineral surfaces. The nature of the surface chemistry
grants humus a surface charge which is pH dependent, hence the tendency to
coil or uncoil, to flocculate or disperse, will be more or less a function
of pH and ionic character of the solution.

COLLOID STABILITY
It is commonly reported that dissolved humic material tends to coat
mineral particles and thereby affect the surface chemistry of those
materials. Dissolved organic matter coats the surfaces of soil particles
even when it is present at very low concentrat ions (DAVIS, 1982). It
furthermore imparts a negative charge to the surfaces which it coats. The
organic coating is expected to have a great significance on subsequent
adsorption of inorganic cations and anions (DALANG et. al.,1984). The
importance of adsorbed organic material on trace metal uptake will be
considerable because of the cation exchange capacity of the organic
matter. Anion adsorption will also be greatly influenced by surfaces
coated with organic material. This may be due to competition for the
adsorption sites or possibly by electrostatic repulsion (DAVIS, 1982).

In addition to the ability of organic matter to coat mineral particles and
thus enhance the cation exchange capacity of the soil minerals, a thin
organic coating may tend to increase the disperse nature of small mineral
particles by imparting a net negative charge and creating a repulsion
between the particles (GIBBS, 1983). The pH dependent nature of the charge
on such coated particles may create a pH-dependent dispersion tendency; as
the pH drops and the surface functional groups of the organic matter
become electrically neutral, the particles coated with this organic matter
would become less mutually repulsive and intraparticle collisions might
result in the formation of van der Waals bonds. Such an event might result
in flocculation of the particles.  The intraparticle repulsion of such
coated minerals will also diminish as the ionic strength of the solution
increases. Experimental evidence has verified this (GIBBS, 1983). This is
in accord with the model of double layer compression at higher ionic
strength, which allows closer approach between particles. Organic coated
particles coagulate much slower than the particles with the coatings
removed. They will also resist sorption onto the stationary phase (as in
saturated groundwater flow) if the stationary phase is also coated with
negatively-charged organic matter (GSCHWEND and REYNOLDS, 1987). Such
behavior is important in environmental management. Solid surfaces in
natural aqueous systems are the sites of important geochemical phenomena.
Coagulation, sedimentation, adsorption, and other processes are usually
controlled by physical chemistry of the solid/liquid interface. Most
models for these processes are based on studies of colloidal systems in
the absence of organic matter. However, almost all part icles are
negatively charged due to adsorbed organic material. This raises the
question of whether adsorption models based on clean oxide surfaces are
useful for a description of natural systems (DAVIS, 1982).

The dispersal and sedimentation of clay minerals and other mineral
colloids may be influenced appreciably by sorbed humic matter. While humic
matter may keep clay particles in a dispersed state under conditions
otherwise conducive to flocculation, humic matter could conceivably
"cement" clay particles together, as a polyelectrolyte bridge, to form
stable aggregates, as in soil, thereby promoting deposition of clay in a
hydraulic regime in which individual colloids would be kept in suspension
(JACKSON, 197 5). The sorptive nature of the colloidal surface creates the
possibility for aggregation between colloids. Aggregation (or coagulation
or flocculation) can cause settling of the colloids as the particle
densities increase. The tendency of colloids to coagulate is a function of
conditions such as pH, ionic strength, solution composition and, as
discussed above, repulsion between colloids. In natural (and polluted)
waters, these conditions causing flocculation can change and the
aggregated particles can disperse back into the solution. The stability of
colloids in natural waters cannot be explained by electrostatic theory
alone, but must be considered as a combination of electrical, kinetic and
purely chemical forces (STUMM and MORGAN, 1981, p. 660).

Dissolved organic matter in solution will influence the sorption chemistry
and aggregation behavior of mineral particles in aqueous systems. The
presence and nature of suspended and dissolved minerals, in turn, will
influence the behavior of the dissolved organic matter. The aqueous phase
will thus contain suspended and dissolved mineral/organic colloids at
greater or lesser concentration as a function of ambient chemistry and
physical conditions. Organic material can form colloids when aggregates or
micelles form. Mineral/organic colloids can exist when mixed aggregates
coprecipitate or agglomerate in solution, or when conditions bring mixed
material into apparent solution.


BIOCOLLOIDS
The term "biocolloids" is frequently applied to microbes in solution.
Bacteria, algae, protozoans and many other biological agents present in
the aqueous phase may be considered to exhibit colloidal behavior. Insofar
as these species are able to sorb contaminants like other colloids, the
distinction between living and nonliving colloids is relatively
unimportant. It is also known that biological exudates or subcellular
fragments may exist in solution (AWWA COMMITTEE, 1981; XUE et al.,1988).

The sorptive nature of bacterial or algal exterior membranes is well
documented. Biological particles can influence the distribution of heavy
metals in natural waters because the functional groups on the cell
surfaces are able to bind metal ions (XUE et a l.,1988). The mechanism of
sorption of metals onto biological surfaces seems to be of different
sorts. CRIST et al.,(1990) report that adsorption of Sr on Vaucheria
released an equivalent amount of Ca and Mg, indicating that metal
adsorption by alkali and alkaline-earth metals is an ion-exchange
phenomenon based on electrostatic interactions. Release of protons when Cu
was adsorbed demonstrated additional covalent bonding for this transition
metal. Protonated ethylenediamine is adsorbed both as a cation similar to
metals and as a neutral species, indicating the presence of additional
bonding sites. Anions such as carboxylate groups of pectin, the polymer of
galacturonic acid, are the most likely sites for electrostatic bonding.
These substances are found in microbes (CRIST et al.,1990).

Microbes are ubiquitous in the subsurface and as such may play an
important role in groundwater solute behavior. Microbes in the subsurface
can influence contaminants by solubility enhancement, precipitation or
transformation (biodegradation) of the contaminant species.

Microbes in the groundwater can act as colloids or participate in the
processes of colloid formation. Bacterial attachment to saturated granular
media can be reversible or irreversible and it has been suggested that
extracellular enzymes are present in the system. Extracellular exudates
(slimes) can be sloughed-off and act to transport sorbed materials (AWWA
COMMITTEE, 1981). The stimulation of bacterial growth in the subsurface
may be considered as in situ formation of colloids.

In the same way as described for surface water, inputs of dissolved
organic matter from the surface tend to stimulate microbial activity
because they constitute reduced carbon which can be utilized as a
substrate. Subsurface microbial activity associated with inputs of organic
substrate will consume oxygen and create reducing conditions if oxygen
demand exceeds supply.

The availability of oxygen in the groundwater can influence the redox
status of system and hence the water chemistry and colloidal status.
Biological activity surrounding soil deposits of oxidizable organic matter
has been found to create sharp and highly localized drops in the redox
potential (PARKIN 1987). Such an event could cause a reduction of a metal
in complex and result in a change in the status of the complex. Ion
complexes such as chelates may be altered by changes in the redox
potential (DOUGLA S and QUINN, 1989). If such an event was truly
localized, a contaminant metal ion might undergo reduction and then
re-oxidation as it was transported out of the localized reducing zone and
into a zone of higher pE.

In addition to the colloidal behavior of microbial species acting to
transport or influence the availability of pollutants, biotransformation
or degradation may occur as well. Biodegradation can occur even when
concentrations are very low. It is reported that compounds such as
chlorinated diphenylamines (products of microbial degradation of
pesticides), naphthalene, styrene, chlorobenzenes, 2,4-D, and Sevin are
biodegradable at concentrations below 100 �g/L (AWWA COMMITTEE, 1981).

COLLOIDS  IN GROUNDWATER
Groundwater was long considered to be relatively invulnerable to
contamination from surface activities; a consideration which has been
widely reevaluated as water quality concerns and analytical capabilities
have advanced (WILSON et al., 1981). Contamination of groundwater has
been widely documented and is of great concern. Groundwater is frequently
valuable and groundwater contamination is often very difficult to remedy.
The role of colloids in groundwater contamination is a topic of current
interest and development.

Groundwater which underlies the soil surface receives input from the soil
as the soil solution migrates downward in the percolating water column
under the influence of gravity. Because of variations in physical
properties such as porosity and permeability involved in geologic
depositional patterns, the hydraulic connections between the groundwater
and the surface can be rather complex. The directions and velocities of
subsurface flow are also subject to variation for the same reasons.

Groundwater is connected to surface through unsaturated flow from human
surface activities such as agriculture or surface landfills (WILSON et
al.,1981; BARKER, et al.,1986).

In addition to unsaturated flow through the vadose zone making inputs to
groundwater, surface water is frequently well-connected with the
groundwater and polluted surface water can move into the subsurface
through saturated flow (SCHWARZENBACH and WESTALL , 1981; SCHWARZENBACH et
al.,1983).

Groundwater contamination from surface activities such as sewage
infiltration, agricultural practices or waste disposal is becoming well
known. Colloids may be transported directly from the surface through the
matrix of soil pores. The soil will tend to filter out some materials,
but since many colloids are smaller than soil pores through which water
moves, the filtration effect will probably be relatively small. It is
likely that the processes of surface sorption, whereby colloids are sorbed
onto the solid phase are more important in removal of colloids during
infiltration through the surface than are processes of physical sieving or
straining by the soil media. Organic macromolecules can move with the
regional groundwater flow as demonstrated by ROBERTSON et al.,(1984) who
found macromolecular tannins and lignins transported 1000 m from a waste
pulp liquor lagoon.

A change in the salt balance of percolation water can cause deflocculation
of the surface soil and result in the transport of mineral colloids into
the groundwater. This was documented by NIGHTINGALE and BIANCHI (1977) who
described groundwater turbidity which became evident in wells near a
groundwater recharge facility in Fresno, California. These workers
reported that the colloidal turbidity was transported from the surface and
through the aquifer to the wells; a distance of several miles in some
cases . The turbidity was determined to be caused by surface application
of recharge water with low ionic strength which caused clay in the surface
soil to disperse and form colloids which moved with the bulk flow.

Metals can be transported to the subsurface in colloidal state. Colloidal
transport of metals, especially polyvalent metals able to form complexes
with organic polyelectrolytes, has been suggested as a possible
explanation of groundwater contamination by cobalt and uranium at the Oak
Ridge National Laboratory in Tennessee (MEANS et al.,1978).

It is reported that dispersed colloids have been observed to transport DDT
and paraquat through vertical soil columns. The DDT was sorbed onto sewage
sludge and the paraquat onto montmorillonite. Transport was enhanced by
water of low ionic strength (VINT EN et al.,1983).

Colloidal transport in the saturated subsurface occurs with the bulk flow
of the groundwater. The potential for colloid transport is suggested
because the colloid particles are far smaller than the pores in permeable
and fractured media, and their high surface area per unit mass means that
they will be effective sorption substrates. Colloid removal from solution
by capture onto fixed media surfaces is controlled by the Brownian motion
of the colloids and the attachment efficiency following collision. In
general, for natural waters with low ionic strengths, colloid attachment
to surfaces is hindered by electrostatic repulsion, but predictions based
on double-layer theory underpredict observed attachment by orders of
magnitude (BUDDEMEIER and HUNT, 1988).

In unconfined, shallow aquifers, unpolluted groundwater has been reported
to be well-supplied with oxygen, i.e., oxygen levels were measured and
found to be near saturation, while pollutant plumes containing dissolved
organic matter are strongly reducing in the interior of the plume where
oxygen diffusion rates are inadequate to meet the demands of microbes
oxidizing the organic matter (THURMAN et al.,1986). Such plumes are
considered to have an oxic interface where the contaminant plume meets the
uncontaminated groundwater (ROBERTSON et al.,1984). Precipitation events
might be expected at such an interface where oxidation of reduced iron or
other elements occurs. These reduced species such as Fe(II) which are
formed in the reducing environment of the pollutant plume will be
oxidized at the oxic interface to form less soluble species such as
Fe(III). The precipitation event might serve to scavenge dissolved
solution components and remove them from solution if the precipitate was
settled and became associated with the solid phase. It has been suggested
that such a precipitation event might serve to modify the hydraulic
conductivity of the aquifer as small particles deposited in the
interstitial pores prevented water from flowing through the pores. If, on
the other hand, the precipitated mineral was of small enough size or low
enough density to be suspended in the groundwater, the resulting colloid
might serve as a vehicle for transport of the pollutant.

There is good evidence that colloids are formed in situ in the subsurface
as changes in water chemistry perturb the local equilibrium and cause
precipitation of dissolved minerals. This was pointed out by GSCHWEND and
REYNOLDS (1977), who analyzed colloid s from groundwater near a secondary
sewage infiltration site. It was determined that these microparticles
consisted primarily of iron and phosphorous. The authors concluded that
these microparticles were formed by sewage-derived phosphate combining
with ferrous iron released from the aquifer solids, and that these
colloids may be moving in the groundwater flow.

The chemical nature of the colloids recovered from the groundwater was
different from that of the colloids in the wastewater influent, indicating
that the colloids were formed in the subsurface and did not simply move
into the aquifer from the infiltration ponds. These authors described a
scenario of a drop in the redox potential of the groundwater driven by
dissolved organic matter in the sewage. This solubilized iron from the
aquifer matrix which combined with the phosphate in the sewage to
precipitate as an insoluble colloidal suspension in the groundwater. They
suggested that such a subsurface transport process could have major
implications regarding the movement of particle-reactive pollutants
traditionally viewed as non-mobile in groundwater.

This same process was envisioned by PENROSE et al.,(1990) who were
investigating groundwater contamination by radionuclides in a shallow
unconfined aquifer. Their water analyses revealed a drop in dissolved
oxygen in the groundwater relative to the surface water.  Under anoxic
conditions, they expected that ferrous iron would be leached from the
aquifer matrix. It would reprecipitate upon encountering oxygen, producing
colloidal materials and particulates.

Confined aquifers (those which are overlain by an aquiclude and are thus
relatively isolated from the surface) are observed to sustain
comparatively reducing conditions as the water moves downslope from the
recharge area. Presumably, organic matter contained in the recharge water
contributes to the oxygen demand in excess of supply. This results in a
depleted oxygen condition which the confining conditions sustain by
preventing oxygen diffusion from the terrestrial atmosphere. A steady
source of organic m atter into a confined aquifer can produce very
reducing conditions. (CHAMP et al.,1979) Under extremely reducing
conditions, sulfide minerals may form as insoluble precipitates with
colloidal dimensions (ROBERTSON et al.,1984). Co-precipitation might
involve otherwise soluble contaminant species in these colloids by the
mechanisms described previously.

COLLOIDAL METAL BEHAVIOR
In addition, dynamic interactions at the solution-solid interfaces
determine the transfer of metals between aqueous and solid phases. Thus,
trace metals may be in a suspended, colloidal, or soluble form. The
suspended and colloidal particles may consist of (1) compounds or
heterogeneous mixtures of metals in forms such as hydroxides, oxides,
silicates, or sulfides or, (2) clay, silica, or organic matter to which
metals are bound by sorption, ion exchange or complexation. The soluble
forms are usually ions , simple or complex, or unionized organometallic
chelates or complexes. Most highly charged metal ions (e.g. Th+4, Fe+3,
Cr+3) are strongly hydrolyzed in aqueous solution.

Fe(H2O)6+3 + H2O= Fe(H2O)5OH2+ + H3O+

Hydrolysis may also proceed further by the loss of one or more protons
from the coordinated water.

Fe(H2O)5OH2+ + H2O= Fe(H2P)4(OH)2+ + H30+

Many divalent metals (e.g. Cu+2, Pb+2, Ni+2, Co+2, and Zn+2) hydrolyze
within the pH range of natural waters.

The hydrolysis of aqueous metal ions can also produce polynuclear
complexes containing more than one metal ion, for example, 2FeOH2+ =
Fe2(OH)24+

Polymeric hydroxo forms of metal ions (e.g., Cr+3) may condense slowly
with time to yield insoluble metal oxides or hydroxides. Polymeric species
are important in moderate to high concentrations of metal salt solutions.

Metal ions also react with inorganic and organic complexing agents present
in water from both natural and contaminant sources. Dominant inorganic
complexing ligands include Cl-, SO4-2, HCO3-, F-, sulfide, and phosphate
species. These reactions are somewhat similar to the hydrolysis reactions
of metal ions in that sequences of soluble complex ions and insoluble
phases may result depending on the metal and ligand concentrations and pH.
Inorganic ligands are usually present in natural waters at much higher
concentrations than the trace metals they tend to complex. Each metal ions
has a speciation pattern in simple aqueous solutions that it dependent
upon (1) the stability of the hydrolysis products and (2) the tendency of
the metal ion to form complexes with other inorganic ligands. For example,
Pb(II), Zn(II), Cd(II) and Hg(II) each form a complex series when in the
presence of Cl- and/or SO4-2 at concentrations similar to those of
seawater. The pH at which a significant proportion of hydrolysis products
are formed is dependent upon the concentration of the ligand, for
example, Cl- competing with OH- for the metal ion.

Metals can also bond to natural and synthetic organic substances by way of
(1) carbon atoms yielding organometallic compounds, (2) carboxylic groups
producing salts of organic acids, (3) electron donating atoms O, N, S, P,
and so on forming coordination complexes, or (4) pi-electron donating
groups (e.g. olefinic bonds, aromatic ring, etc.). It is reported that
under aerobic conditions free metal ions occur mainly at low pH, and with
increasing pH the carbonate and then the oxide, hydroxide, or even
silicate solids precipitate. Metal speciation is also controlled by
oxidation-reduction conditions.

COLLOIDAL BEHAVIOR OF RADIONUCLIDES
Transport of colloidally bound material has been indicated in the movement
of radionuclides in groundwater in several instances. This has been
well-documented (e.g., SHORT et al.,1988; McCARTHY and ZACHARA, 1989 and
others. The contemporary relevance of environmental contamination by
radionuclides warrants special consideration of this topic.

In their 1977 report by a Group of Experts, the OECD NUCLEAR ENERGY AGENCY
reports that plutonium hydrolyzes in the pH range of natural waters. The
hydrolysis products are stated to exhibit colloidal behavior and be easily
adsorbed on particle surfaces. The solubility of uranium and plutonium is
enhanced by the presence of bicarbonate ions, with which they form stable
carbonate complexes. This report mentions difficulties of forecasting the
behavior of plutonium in the aqueous environment due to the various
possible oxidation states of the element and the great chemical
variability of natural waters.

The work cited above reports at the burial ground at Maxey Flats,
Kentucky, plutonium has been detected in surface soil, in soil cores 90 cm
deep, in monitoring wells, and in streams which drain the site. It is not
clear if ground water has been the main dispersing medium, since spreading
of contamination above ground has been suggested. However, in case of
underground transfer, the data would indicate migration of tens and
possibly hundreds of meters in less than 10 years. The most likely
migration paths would be along the fractures and joints that are fairly
well developed in the burial formation. In a case like this the high
adsorption capacity of the shales containing the waste would not
constitute an effective barrier since only a small fraction of the water
moves through the intergranular interstices.

MEANS et al.,(1978) found that radionuclides (cobalt and uranium) in
nuclear waste buried at the Oak Ridge National Laboratory facility were
migrating in chelated form with EDTA, which is used to cleanup accidental
spills. The chelated material was buried and is migrating into groundwater
as a stable complex. The same report also states that radionuclide
complexes with humic material were found, although this was not thought to
be responsible for the initial movement of the contaminant from the site
of burial. It is suggested that the humic/radionuclide complex was formed
in the environment after the EDTA/radionuclide complex moved away from its
place of burial.

This suggests the importance of the stability of the complex; the
dissociation tendency of the ligand from the bound material, and the
degradability of the chelating agent must be considered. Many humic or
biologically produced macromolecules may be able to complex contaminants
and may also be susceptible to degradation in the environment. Such an
event would release the bound pollutant, perhaps resulting in the
formation of another complex. The reported presence of humate-bound
radionuclides would seem to indicate that the EDTA/radionuclide complex
was dissociating to some degree to enable the humate/radionuclide complex
to form.

SHEPPARD et al.,(1980) used gel filtration to determine the particle size
distribution of radionuclides in association with the colloids. They
report that particles of colloidal dimensions are shown to be important
potential vehicles for the transport of radionuclide elements in soils and
groundwater. For the soils studied, the distribution of radionuclides
between the soil and aqueous phases is determined by a characteristic
particle size spectrum of radionuclide-bearing particles. These spectra,
which are related to the physical and chemical composition of the soils,
include uncomplexed ions, complexes of fulvic and humic acid polymers, and
larger radionuclide-bearing particles, such as clay.

For the soils and experimental techniques used, the authors report that
three broad classes of particles determined the radionuclide distribution
ratios; (1) ionic particles containing radionuclides which have radii less
than 1 nm, (2) complexes of humic matter, possibly humic acid polymers
with molecular weights between 8000 and 50,000 (2-3 nm radii), and (3)
larger soil particles bearing radionuclides and with radii in the 10-60 nm
range.

In this particular study, the authors found a maximum in radionuclide
capacity occurring at sizes corresponding to molecular weights of
8000-50,000 (2-3 nm.). This is in the range of mean molecular weights of
soil humic acids reported by KONONOVA (1961).

Strontium is reported to be strongly bound by humic matter. The reported
distribution ratio of Sr at 3 nm is three orders of magnitude larger than
its ionic component (SHEPPARD 1980).

Uranium is often found in natural deposits of organic matter (JACKSON,
1975). This indicates an affinity of uranium or its geologic precursors
for organic matter.

Workers in Canada (CHAMP et al.,1984) investigated radionuclide
contamination of groundwater. The authors suggested that organic ligands
enhanced the mobility of Co, Ce, Cs, Eu, Sb and Zr.

An in situ glass-block-leaching experiment is one of the four study sites
at the Chalk River Nuclear Facility in Canada. Field measurements of 137Cs
migration indicated transport four times farther than predicted from Kd
values. Soil columns prepared from undisturbed, uncontaminated cores
showed Cs transport by 0.2 - 1.0 micrometer particles. There was also an
indication that microorganisms were involved in particulate Cs release and
transport.

Work by BUDDEMEIER and HUNT (1988) conclusively demonstrates radionuclide
transport by colloids in groundwater flowing through fractured media at
the Nevada Test Site. Groundwater samples collected from within a nuclear
detonation cavity and from approxim ately 300 meters outside the cavity
were analyzed for chemical composition and radionuclide activity. In
samples from both locations, approximately 100% of the transition element
(Mn, Co) and lanthanide (Ce, Eu) radionuclides were associated with
colloids . Their presence outside the cavity indicates transport in the
colloidal form. Equilibrium distribution coefficients calculated for Ru,
Sb, and Cs nuclides from both field sample locations indicate partitioning
on the 0.05-0.003 micrometer colloids. Calculation of transport
efficiencies relative to colloid mass concentrations and dissolved neutral
or anionic nuclides indicates that both the cations and the radiolabeled
colloids appear to experience capture by, or exchange with, immobile
aquifer surfaces.

The following conclusions are offered from their work:

(1) There is a strong association between colloidal solids and
radionuclides, particularly in the case of elements that are strongly
sorbing and/or insoluble under groundwater conditions.

(2) Both the dissolved and colloidal radionuclide species undergo
hydrological transport through the fracture-flow system.

(3) A substantial proportion of the radionuclides associated with
suspended colloids pass through the conventional filters traditionally
used to distinguish between "particulate" and "dissolved" species.

(4) Measurements indicate that most of the colloidal material consists of
natural minerals. There is some evidence that the high colloid
concentrations are not unique to the specific study site. (It was observed
that no noticeable turbidity accompanied the colloids).

(5) Chemical analyses, isotopic comparisons and distribution coefficient
calculations all suggest that the radionuclides have chemically
equilibrated with the groundwater system, and that their behavior is
representative of their stable analogs and of trace element geochemistry
in general.

(6) Colloidal nuclides and soluble cations are transported less
efficiently than soluble neutral or anionic species, indicating that both
colloids and cations experience some additional processes of loss to or
exchange with the aquifer during hydrogeological transport.

The work of PENROSE et al.,(1990) describes the release of treated aqueous
wastes containing traces of plutonium and americium into a desert canyon,
within the site of the Los Alamos National Laboratory, N.M. The wastes
infiltrate a small aquifer within the canyon. Although laboratory studies
have predicted that the movement of actinides in subsurface environments
will be limited to less than a few meters, both plutonium and americium
are detectable in monitoring wells as far as 3390 m downgradient from the
discharge. Investigation of the properties of the mobile actinides
indicates that the plutonium and part of the americium are tightly or
irreversibly associated with colloidal material between 25 and 450 nm in
size. A fraction of the americium exists in a low molecular weight form
(diameter less than or equal to 2 nm) and appears to be a stable, anionic
complex of unknown composition.


COLLOID SUMMARY
Application of contaminant behavior models which neglect the colloidal
phase may result in inaccurate and unfortunate estimations of apparent
solubility and transport.

The impact of colloidal solubility enhancement will be the most pronounced
for the least water soluble solutes. The affinity of a solute for a
colloid is a function of the same tendencies which drive a material to
sorb onto the stationary solid phase; namely bonding interactions and
hydrophobicity.  Hence, colloids will manifest the greatest solubility
enhancement for those materials which are the least soluble in water or
the most attracted to the solid phase. Materials which are soluble in
water are less likely to be sorbed onto the solid or colloidal phase in
the absence of specific bonding interactions. A 1985 report by CARON et
al., reported that dissolved organic matter increased the apparent
solubility of DDT significantly, while the solution behavior of lindane
was virtually unchanged. Lindane is reported to be 3 orders of magnitude
more soluble in pure water than DDT. The colloidal carbon effects on the
apparent solubility of neutral organics is the most significant for low
solubility, highly sorbed, solutes. It is for these hydrophobic organics
that the relative increases in solubility due to sorption on colloidal
material is the greatest (BOUCHARD et al.,1989).

Because the affinity of organic solutes for binding to humic substances is
related to their hydrophobicity, the potential impact of organic colloids
may be assessed from the octanol-water partition coefficient of the
solute. Several different estimates ha ve been made as to what constitutes
hydrophobicity. Where one draws the line is a matter of judgment and
experience. In general, it may be assumed that the larger the Kow, the
greater the tendency of the solute to partition out of the aqueous phase.
Octanol-water partition coefficient values are widely cataloged.

In a similar manner to the assessment of the hydrophobicity of organic
material, the water solubility of inorganic solutes may be considered as
an important parameter in estimating the apparent solubility of metals due
to the presence of colloids in solut ion. Because of the potential
diversity of the solution and colloidal composition, assessing the
behavior of metal-colloids in solution may be very challenging. The
presence of insoluble counterions and complexing agents may act to
influence the status of a metal ion.

At any rate, the composition and concentration of the colloidal phase will
obviously have an impact on the ability of such a phase to transport
slightly-soluble contaminants.

Colloids are more or less an extension of the solid phase, i.e.,
biological, mineral and organic matter, and they will vary in composition
in similar ways. Research has shown that colloidal organic matter does
differ in its ability to sorb neutral organics, and that organic colloid
sorptivity increases with size and hydrophobicity.

The concentration of colloids in the water will have an effect on the
degree of solubility enhancement. It is often reported that dissolved
organic carbon contents of natural waters are typically in the region of 5
mg carbon/L (THURMAN and MALCOLM, 1981). Given the proven importance of
molecular size and polarity, such a number really doesn't mean much in an
absolute sense (CHIAN and DeWALLE, 1977). Not surprisingly, dissolved
organic carbon levels are reported to have a very wide range. A reported
higher carbon-to-oxygen ratio of groundwater than surface water humic
material suggests a lower polarity and therefore greater sorptivity for
neutral organics by groundwater humic material (BOUCHARD et al.,1989).

While it is possible that colloids may be involved with water-soluble
contaminants as well, it is doubtful if the role of colloids in
transporting such materials will be significant relative to transport in
the water. It is possible that sorption of any material onto colloids
will affect the degradation or transformation of that material. The water
soluble nature of a colloid would seem to indicate a hydrophilic surface
which could conceivably sorb a polar or ionic species. It has been
frequently reported that sorption/association with a surface affects the
chemical transformation of the sorbed/associated species. The increased
surface area of colloids in solution will introduce increased interfacial
area into the system and hence will tend to enhance surface-related
reactions. This aspect of colloidal behavior makes colloids an important
consideration in the BARR process.


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