7. TOXIC ORGANIC CHEMICALS

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7. TOXIC ORGANIC CHEMICALS

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Title: 7. TOXIC ORGANIC CHEMICALS

1
7. TOXIC ORGANIC CHEMICALS

If we live as if it matters and it doesn't
master, it doesn't matter. If we live as if it
doesn't matter, and it matters, then it matters.

There are 4 million organic chemicals (IUPAC).
1000 new organic chemicals are synthesized
each year.
A fraction of these is toxic or
carcinogenic, and the vast majority of them break
down in the environment.
If organics are persistent as wel1 as toxic, we
may need to use mathematical models to determine
if they pose an unreasonable risk to humans or
the environment.
Organic chemistry is the chemistry of compounds
of carbon. Organic chemicals are obtained from
material produced originally by living organisms
(petroleum, coal, and plant residues) or they are
synthesized from other organic compounds or
inorganics (carbonates or cyanides).

3
7.1 NOMENCLATURE

Figure 7.1 shows some classes of organic
compounds that are widely used. The left-hand
side of the figure gives some general classes of
compounds and the right-hand side is a specific
example of each.

Figure 7.1 Some common classes of organic
compounds (left) and examples (right). R and R
indicate different alkyl group.
4

In the environment, alkanes ? alcohol.
Enzymes catalyze the reactions, but other
abiotic processes such as photolysis, hydrolysis,
chemical oxidation or reduction may also be
important.
Microbial "infallibility" would state that all
organic chemicals that are synthesized can be
mineralized all the way to carbon dioxide and
water as shown above.
Microbes are not infallible, although given
the proper conditions, enough time, and in
concert with other physical and chemical
reactions, they can often help to break down most
organic chemicals. On the other hand, microbes
and plants can sometimes synthesize chemicals in
nature that are quite toxic and rather slow to
degrade.
Chlorinated organic chemicals are not purely
man-made (xenobiotics), but now we know that some
chlorinated organic chemicals are synthesized by
plants and quite common in nature.

Figure 7.2 shows some examples of cyclic organic
chemicals that are sometimes difficult to degrade
in the environment.

To oxidize benzene to carbon dioxide and water
requires that the very stable benzene ring must
be cleaved. Under anaerobic conditions this can
be a difficult task.

Figure 7.2 Examples of cyclic organic compounds
(including alicyclic, aromatic, and heterocyclic
compounds).
6

Drinking water standards. Organic chemicals for
which maximum allowable drinking water standards
have been established are shown in Figure 7.3.

Figure 7.3
(a) Volatile organic compounds that have maximum
contaminant level (MCL) drinking water standards.
(b) Some synthetic organic chemicals for which
maximum contaminant levels (MCLs) have been
established.

7
7.2 ORGANICS REACTIONS

The types of reactions biological
transformations, chemical hydrolysis,
oxidation/reduction, photodegradation,
volatilization. sorption, and bioconcentration
are among the important reactions that organic
chemicals undergo in natural waters.
Biological transformations - the microbially
mediated transformation of organic chemicals,
often the predominant decay pathway in natural
waters. It may occur under aerobic or anaerobic
conditions, by bacteria, algae, or fungi, and by
an array of mechanisms (dealkylation, ring
cleavage, dehalogenation, etc.). It can be an
intracellular or extracellunar enzyme
transformation.
The term "biodegradation" is used synonymously
with "biotransformation," but some researchers
reserve "biodegradation" only for oxidation
reactions that break down the chemical. Reactions
that go all the way to CO2 and H2O are referred
to as "mineralization." In the broadest sense,
"biotransformation" refers to any microbially
mediated reaction that changes the organic
chemical. It does not have to be an oxidation
reaction, nor does it have to yield carbon or
energy for microbial growth or maintenance.

7.2.1 Biological Transformations
8

The term "secondary substrate utilization" - the
utilization of organic chemicals at low
concentrations in the presence of one or more
primary substrates that are used as carbon and
energy sources.
"Co-metabolism" - the transformation of a
substrate that cannot be used as a sole carbon or
energy source but can be degraded in the presence
of other substrates.
Many toxic organic reactions in natural waters
are microbially mediated with both bacteria and
fungi degrading a wide variety of pesticides.
Dehalogenation, dealkylation, hydrolysis,
oxidation, reduction, ring cleavage, and
condensation reactions are all known to occur
either metabolically or via co-metabolism (see
Table 7.1). In co-metabolism, the microbe does
not even derive carbon or net energy from the
degradation rather, the pesticide is caught up
in the overall metabolic reactions as a
detoxification or other enzymatic reaction.
Several bacterial genera are known that are
capable of utilizing certain organics as the sole
carbon, energy, or nitrogen source. Pseudomonas
(with 2,4-D and paraquat), Nocardia (with dalapon
and propanil), and Aspergillus species (with
trifluralin and picloram) are poignant examples.

9
Table 7.1 Biological Transformations Common in
the Aquatic/Terrestrial Environment
10

It is convenient when possible to express rate
expressions for organic transformations as
pseudo-first-order-reactions, such as equation
(1) below. The reaction rate expression is then
(1)
where C is the toxic organic concentration in
solution and kb is the pseudo-first-order
biotransformation rate constant.
Table 7.2 is a summary of pseudo-first-order and
second-order rate constants kb for the
disappearance of toxic organics from natural
waters and groundwater via biotransformation.
The actual microbial biotransformation rate
follows the Monod or Michaelis-Menton enzyme
kinetics expression
(2)
Where kb pseudo-first-order biological
transformation rate constant,T-1
µ maximum growth rate, T-1
X viable microbial biomass
concentration, M L-3
Y cell yield, microbial cell
concentration yield/ organic concentration
utilized
KM Michaelis half saturation
constant, M L-3.
Typical cell concentrations in surface waters
would be 106 107 cells mL-1 and less in
groundwater

11
Table 7.2 Selected Biotransformation Rate
Constants.
12

Under typical environmental conditions, the
concentration of dissolved organics (C lt 10 µg
L-1) is less than that of the Michaelis
half-saturation constant (KM 0.1-10 mg L-1).
Therefore the equation becomes
(3a)
where kb µ/YKM.
Sometimes organic chemicals that are adsorbed to
suspended particulate matter are biodegraded in
addition to soluble chemical. Equation (3a) must
be rewritten in terms of both dissolved and
adsorbed chemical concentrations
(3b)
where CT is the total whole water chemical
concentration, C is the dissolved phase
concentration and Cp is the particulate adsorbed
concentration.
If the substrate concentration C is very large
such that C gtgt KM , then the microorganisms are
growing exponentially, and the rate expression in
equation (2) reduces to
(4)
which is a zero-order rate expression in C and
first-order in X.

Biotransformation experiments are conducted by
batch, column, and chemostat experimental
methods. Other fate pathways (photolysis,
hydrolysis, volatilization) must be accounted for
in order to correctly evaluate the effects of
biodegradation.
It is incumbent on the fate modeler to understand
the range of breakdown products (metabolites) in
biological transformation reactions. Metabolites
can be as toxic (or more toxic) than the parent
compound.
Following all the metabolites and pathways in the
biological degradation of organic chemicals can
be complicated. Polychlorinated biphenyls (PCBs)
are mixtures of many isomers - the total number
of different organic chemicals is 209 congeners.
Figure 7.3b shows the structures, where x and y
represent the combinations of chlorine atoms (one
to five) at different positions on the biphenyl
rings. Each congener has distinct properties that
result in a different reactivity than the others.
Both the rate of the biological transformation
and the pathway can be different for each of the
congeners.

There are several basic types of biodegradation
experiments. Natural water samples from lakes or
rivers can have organic toxicant added to them in
batch experiments. Disappearance of toxicant is
monitored.
Organic xenobiotic chemicals can be added to a
water-sediment sample to simulate in situ
conditions, or a contaminated sediment sample
alone may be used with or without a spiked
addition. Primary sewage, activated sludge, or
digester sludge may be used as a seed to test
degradability and measure xenobiotic
disappearance.
Radiolabeled organic chemicals can be used to
estimate metabolic degradation (mineralization)
by measuring CO2 off-gas and synthesis into
biomass. These experiments are called
heterotrophic uptake experiments.
The organic chemical may be added in minute
concentrations to simulate exposure in natural
conditions, or it may be the sole carbon source
to the culture to determine whether
transformation reactions are possible.

Biodegradation is affected by numerous factors
that influence biological growth
Temperature effects on biodegradation of toxics
are similar to those on biochemical oxygen demand
(BOD) using an Arrhenius-type relationship.
Nutrients are necessary for growth and often
limit growth rate. Other organic compounds may
serve as a primary substrate so that the chemical
of interest is utilized via co-metabolism or as a
secondary substrate.
Acclimation is necessary for expressing
repressed (induced) enzymes or fostering those
organisms that can degrade the toxicant through
gradual exposure to the toxicant over time. A
shock load of toxicant may kill a culture that
would otherwise adapt if gradually exposed.
Population density or biomass concentration
organisms must be present in large enough numbers
to significantly degrade the toxicant (a lag
often occurs if the organisms are too few).

16
7.2.2 Chemical Oxidation

Chemical oxidation takes place in the presence of
dissolved oxygen in natural waters. Oxygen is
reduced and the organic chemical is oxidized, but
the reaction can be slow. Alternatively, chemical
oxidation can be triggered by photochemical
transients that may have considerable oxidizing
power but low concentrations.
Oxidants such as peroxyl radicals ROO, alkoxy
radicals RO, hydrogen peroxide H2O2, hydroxyl
radicals OH, singlet oxygen O2, and solvated
electrons are produced in low concentrations and
react quickly in natural waters. Because of their
large oxidizing power, they may react with a
variety of trace organics in solution, but each
transient reacts rather specifically with certain
trace organic moieties.
It is better to determine the relevant oxidant
chemistry and to measure the oxidant
concentration when possible. Since the transient
chemical oxidants are often generated
photochemically, light-absorbing chromophores,
such as humic and fulvic acids and algal
pigments, and sunlight intensity will influence
oxidation rates.

Alkyl peroxyl radicals (1 10-9 M in sunlit
natural waters) react rapidly with phenols and
amines in natural waters to form acids and
aromatic radicals
Singlet oxygen reacts specifically with olefins
Singlet oxygen concentrations in sunlit natural
waters are on the order of 110-12 M. All of
these oxidation reactions may be assumed to be
second-order reactions
(5)
where C is the organic concentration and Ox is
the oxidant concentration.
Table 7.3 the second-order rate constants for
chemical oxidation of selected priority organic
chemicals with singlet oxygen and alkyl peroxyl
radicals.

18
Table 7.3 Second-Order Reaction Rate Constants
for Chemical Oxidation Summary Table of
Oxidation Data with Singlet Oxygen O2 and Alkyl
Peroxyl Radicals ROO
19

Free radical oxidation requires a chain or series
of reactions involving an initiation step,
propagation, and subsequent termination. We will
illustrate the free radical reaction using the
alkyl peroxyl radical ROO as an example.

The chemical is represented as an arbitrary
organic, RH. A-B is the initiator, which is any
free radical source including peroxides, H2O2,
metal salts, and azo compounds. Investigators
have utilized a commercially available azo
initiator to estimate the reactivity of
pesticides to ROO in natural waters.
If no initiators are available in the water, then
reaction (c) represents the probable oxidation
pathway, a slow reaction with dissolved oxygen.
Otherwise steps (a) and (b) lead to peroxide
formation, step (d). Once the highly reactive
peroxide radical is formed, it continues to react
with the organic chemical, RH, and regenerates
another free radical, R', as given in reaction
(e).
This step may be repeated thousands of times for
every photon of light absorbed. Chance collisions
between free radicals can terminate the reaction,
reactions (f), (g), and (h). At the low pollutant
concentrations found in natural waters, reaction
(f) is the most likely termination step. Hydrogen
peroxide may also be formed, especially when
natural dissolved organic matter (DOC) and
humates are present. H2O2 is a powerful oxidant
in natural waters.

If the initiation step is rapid, then the
rate-limiting step is the rate of oxidation of
the organic in reaction (e)
(6)
Provided that reaction (d) is more raped than
reaction (e), the rate of peroxide formation is
(7)
and assuming steady state, the rate of radical be
equal to the rate of termination
(8), (9)
Substituting equation (9) into equation (6), we
find the final reaction rate for the oxidation of
the organic chemical is
(10)

The rate of reaction is a pseudo-first-order
reaction, where k3 is the overall reaction rate
constant which is a function of rf, the rate of
peroxide formation. If the rate of peroxide
formation is relatively constant (as expected in
natural waters), then the free radical oxidation
of the toxic organic can be computed as a
pseudo-first-order reaction.
First-order oxidations of pesticides and organic
chemicals have been reported in natural waters.
However, these oxidations are often microbially
mediated. Strictly chemical free radical
oxidation of toxic organics in natural waters
remains important for a few classes of compounds.
Free radical oxidation is often a part of the
photolytic cycle of reactions in natural waters
and atmospheric waters.
Oxidations of organic chemicals by O2(aq) is
generally slow, but it can be mediated by
microorganisms. Cytochrome P450 monooxygenase is
a well-studied enzyme with an iron porphyrin
active site. Methanotrophs and other organisms
can use this pathway to oxidize organics in
natural waters, a type of biological
transformation.

23
7.2.3 Redox Reactions

Electron acceptors such as oxygen, nitrate, and
sulfate can be reduced in natural waters while
oxidizing trace organic contaminants. Oxidation
reactions of toxic organic chemicals are
especially important in sediments and
groundwater, where conditions may be anoxic or
anaerobic. The general scheme for utilization of
electron acceptors in natural waters fort lows
thermodynamics (Table 7.4).
The sequence of electron acceptors is
approximately
The organic chemical in Table 7.4 is represented
as a simple carbohydrate (CH2O such as glucose
C6H12O6) but other organics may be important
reductants in natural waters and groundwaters.

Strict chemical reduction reactions that do not
involve a biological catalyst (abiotic reactions)
are common in groundwater but less important in
natural waters and sediments, where a great
complement of enzymes are available for redox
transformations. In groundwater, H2S is a common
reductant. It can reduce nitrobenzene to aniline
in homogeneous reactions.

Table 7.4 Redox Reactions in a Closed Oxidant
System at 25ºC and pH 7.0 and Their Free Energies
of Reaction.
25

Likewise, humic substances and their decay
products (natural organic matter, NOM) are good
reductants in homogeneous systems.
Figure 7.4 is a structure-activity relationship
demonstrating that, in homogeneous solution, the
second-order kinetic rate constant kAB is
directly proportional to the one-electron
reduction potential of the redox couple.
(11)
where H2X is the reductant.
Schwarzenbach et al. have shown that, in the case
of juglone, it is not the diprotic dihydroquinone
H2JUG that is the reactant with nitroaromatics,
but rather the anions HJUG- and JUG2-. Reductants
in natural waters include quinone, juglone (oak
tree exudate), lawsone, and Fe-porphyrins.
Nitroreduction is a two-electron, two-proton
transfer reaction.
The reduction of nitroaromatic compounds in
natural waters and soil water may be viewed as an
electron transfer system that is mediated by NOM
or its constituents.

26
Figure 7.4 Liner free-energy relationship
between second-order rate constant and the one
electron potential for reduction of substituted
nitrobenzenes with natural organic matter
(Juglone). From Schwarzenbach, et al..
27

Natural organic matter contains electron transfer
mediators such as quinones, hydroquinones, and
Fe-porphyrin-like substances.
These mediators are reactants that are
regenerated in the process by the bulk reductant,
which is in excess.

28
Table 7.5 Redox Half-Reactions Pertinent in
Wastewater, Groundwater, and Sediment Reactions
29
7.2.4 Photochemical Transformation Reactions

Direct photolysis, a light-initiated
transformation reaction, is a function of the
incident energy on the molecule and the quantum
yield of the chemical.
When light strikes the pollutant molecule, the
energy content of the molecule is increased and
the molecule reaches an excited electron state.
This excited state is unstable and the molecule
reaches a normal (lower) energy level by one of
two paths
- it loses its "extra" energy through energy
emission, that is, fluorescence or
phosphorescence
- it is converted to a different molecule
through the new electron distribution that
existed in the excited state. Usually the organic
chemical is oxidized.
Photolysis may be direct or indirect. Indirect
photolysis occurs when an intermediary molecule
becomes energized, which then reacts with the
chemical of interest.

The basic equation for direct photolysis is of
the form
(12)
Where C is the concentration of organic chemical,
and kp is the rate constant for photolysis.
Photo1ysis rate constants can be measured in the
yield with sunlight or under laboratory
conditions.
The first-order rate constant, kp can be
estimated directly
(13)
where kp photolysis rate constant, s-1
J 6.02 1020 conversion constant
f quantum yield
I? sunlight intensity at wavelength ?,
photons cm-2 s-1
e? molar absorbtivity or molar extinction
coefficient at wavelength ?,
molarity-1 cm-1.
The near-surface photolysis rate constants,
quantum yields, and wavelengths at which they
were measured are presented in Table 7.6.
Photolysis will not be an important fate process
unless sunlight is absorbed in the visible or
near-ultraviolet wavelength ranges (above 290 nm)
by either the organic chemical or its sensitizing
agent.

The quantum yield is defined by
(14)
An einstein is the unit of light on a molar basis
(a quantum or photon is the unit of light on a
molecular basis). The quantum yield may be
thought of as the efficiency of photoreaction.
Incoming radiation is measured in units of energy
per unit area per time (e g., cal cm-2 s-1). The
incident light in units of einsteins cm-2 s-1
nm-1 can be converted to watts cm-2 nm-1 by
multiplying by the wavelength (nm) and 3.03
1039.
The intensity of light varies over the depth of
the water column and may be related by
(15)
where Iz is the intensity at depth z, I0 is the
intensity at the surface, and Ke is an extinction
coefficient for light disappearance.
Light disappearance is caused by the scattering
of light by reflection off particulate matter,
and absorption by any molecule. Absorbed energy
can be converted to heat or can cause photolysis.
Light disappearance is a function of wavelength
and water quality (e.g., color, suspended solids,
dissolved organic carbon).

Indirect or sensitized photolysis occurs when a
nontarget molecule is transformed directly by
light, which, in turn, transmits its energy to
the pollutant molecule. Changes in the molecule
then occur as a result of the increased energy
content.
The kinetic equation for indirect photolysis is
(16)
where k2 is the indirect photolysis rate
constant, X is the concentration of the nontarget
intermediary, and kp is the overall
pseudo-first-order rate constant for sensitized
photolysis.
The important role of inducing agents (e.g.,
algae exudates and nitrate) has been
demonstrated.
Inorganics, especially iron, play an important
role in the photochemical cycle in natural
waters. Hydrogen peroxide, a common transient
oxidant, is a natural source of hydroxyl radicals
in rivers, oceans, and atmospheric water
droplets.

Direct photolysis of H2O2 produces OH, but this
pathway is relatively unimportant because H2O2
does not absorb visible light very strongly. The
important source of OH involves hydrogen
peroxide and iron (II) in a photo-Fenton
reaction.
Hydroxyl radicals are a highly reactive and
important transient oxidant of a wide range of
organic xenobiotics in solution. They can be
generated by direct photolysis of nitrate and
nitrite in natural waters, or they can be
generated from H2O2 in the reaction shown above.
Nitrobenzene, anisole, and several pesticides
have been shown to be oxidized by hydroxyl
radicals in natural waters.

34
7.2.5 Chemical Hydrolysis

Chemical hydrolysis is that fate pathway by which
an organic chemical reacts with water.
Particularly, a nucleophile (hydroxide, water, or
hydronium ions), N, displaces a leaving group, X,
as shown.
Hydrolysis does not include acid-base, hydration,
addition, or elimination reactions. The
hydrolysis reaction consists of the cleaving of a
molecular bond and the formation of a new bond
with components of the water molecule (H, OH-).
It is often a strong function of pH (see Figure
7.5).
Three examples of a hydrolysis reaction are
presented below.

Types of compounds that are generally susceptible
to hydrolysis are
- Alkyl halides
- Amides
- Amines
- Carbamates
- Carboxylic acid esters
- Epoxides
- Nitriles
- Phosphonic acid esters
- Phosphoric acid esters
- Sulfonic acid esters
- Sulfuric acid esters
The kinetic expression for hydrolysis is

A summary of these data is presented in Table
7.7.
Hydrolysis experiments usually involve fixing the
pH at some target value, eliminating other fate
processes, and measuring toxicant disappearance
over time. A sterile sample in a glass tube,
filled to avoid a gas space, and kept in the dark
eliminates the other fate pathways. In order to
evaluate ka and kb, several non-neutral pH
experiments must be conducted as depicted in
Figure 7.5.
Often, the hydrolysis reaction rate expression in
equation (17) is simplified to a
pseudo-first-order reaction rate expression at a
given pH and temperature (Table 7.7, 298 K and pH
7).
(18)
where kh kb OH- ka H kn and kh is the
pseudo-first-order hydrolysis rate constant, T-1
kb is the base-catalyzed rate constant,
molarity-1 T-1 ka is the acid-catalyzed rate,
polarity-1 T-1 and kn is the neutral rate
constant, T-1.

37
Table 7.7Selected Chemical Hydrolysis Rate
Constants, at 298 K and pH 7.
38
Figure 7.5Effect of pH on hydrolysis rate
constants.
39
7.2.6 Volatilization/Gas Transfer

The transfer of pollutants from water to air or
from air to water is an important fate process to
consider when modeling organic chemicals.
Volatilization is a transfer process it does not
result in the breakdown of a substance, only its
movement from the liquid to gas phase, or vice
versa.
Gas transfer of pollutants is analogous to the
reaeration of oxygen in surface waters and will
be related to known oxygen transfer rates. The
rate of volatilization is related to the site of
the molecule (as measured by the molecular
weight).
Gas transfer models are often based on two-film
theory (Figure 7.6). Two-film theory was derived
by Lewis and Whitman in 1923. Mass transfer is
governed by molecular diffusion through a
stagnant liquid and gas film. Mass moves from
areas of high concentration to areas of low
concentration. Transfer can be limited at the gas
film or the liquid film.
Oxygen, for example, is controlled by the
liquid-film resistance. Nitrogen gas, although
approximately four times more abundant in the
atmosphere than oxygen, has a greater liquid-film
resistance than oxygen.

Volatilization, as described by two-film theory,
is a function of Henrys constant, the gas-film
resistance, and the liquid-film resistance. The
film resistance depends on diffusion and mixing.
Henry's constant, H, is a ratio of a chemical's
vapor pressure to its solubility. It is a
thermodynamic ratio of the fugacity of the
chemical (escaping tendency from air and water).
(19)
where pg is the partial pressure of the chemical
of interest in the gas phase
Csl is its saturation solubility.
Henry's constant can be "dimensionless" mg/L (in
air)/mg/L (in water) or it has units of atm m3
mol-1.

Figure 7.6 Two-film theory of gas-liquid
interchange.
41

The value of H can be used to develop simplifying
assumptions for modeling volatilization. If
either the liquid-film or the gas-film controls -
that is, one resistance is much greater than the
other - the lesser resistance can be neglected.
The flux of contaminants across the boundary can
be modeled by Fick's first law of diffusion at
equilibrium,
(20)
where D is the molecular diffusion coefficient
and dC/dx is the concentration gradient in either
the gas or liquid phase.
If we consider the molecular diffusion to occur
through a thin stagnant film, the mass flux is
then
(21)
where k D/?z in which ?z is the film thickness
and k is the mass transfer coefficient with units
of LT-1.
At steady state, the flux through both films of
Figure 7.6 must be equal
(22)

If Henry's law applies exactly at the interface,
we can express the concentrations in terms of
bulk phase concentrations, which are measurable
by substitution below
(23)
(24)
(25)
By rearranging equation (25), we can solve for N
in terms of bulk phase concentration, mass
transfer coefficients for each phase, and Henry's
constant
(26)
where KL is the overall mass transfer coefficient
derived for expression of the gas transfer in
terms of a liquid phase concentration.
(27)

We may think of the first term on the right-hand
side of the equation as a liquid-film resistance
and the second term as a gas phase resistance
using an electrical resistance analogy.
We can compare the two resistances to determine
if the
(28)
gas phase resistance, rg, or the liquid phase
resistance, rl, predominates.
Equivalently, we could choose to write the
overall mass transfer in terms of the buck gas
phase concentration.
(29), (30)
If the gas is soluble, then H is small and the
gas-film resistance controls mass transfer.
In terms of a differential equation, the overall
gas transfer
(31)
where Csat pg/H, A is the interfacial surface
area, and V is the volume of the liquid.

In streams, A/V is the reciprocal depth of the
water and the equation can be expressed as
(32)
where Z is the mean depth and kli is termed the
volatilization rate constant (T-1).
Equations (31) and (32) apply for either gas
absorption or gas stripping from the water body.
It is a reversible process.
The mass transfer coefficients are dependent on
the hydrodynamic characteristics of the air-water
interface and flow regime. For flowing water, we
may write
(33)
where u is the mean stream velocity and Z is the
mean depth.
For smooth flow (no ripples or waves) and wind
speed less than 5 ms-1, 1/Kd predominates.
(34), (35)
where CD is the dimensionless drag coefficient,
W is the wind speed, and v is the kinematic
viscosity.

The transfer term for aerodynamically rough flow
with wave is
(36)
where d is the diameter or amplitude of the
waves, u is the surface shear velocity and a is
a constant dependent on the physics of the wave
properties.
The diffusion coefficients in water and air have
been related to molecular weight
(37)
where Dl is the diffusivity of the chemical in
water and MW is the molecular weight, and
(38)
where Dg is the diffusivity of the chemical in
air.
The mass transfer rate constant, kli, can then
be related to the oxygen reaeration rate, ka, by
a ratio of the diffusivity of the chemical to
that of oxygen in water
(39)

The reaeration rate, ka, can be calculated from
any of the formulas available. In addition, the
overall gas-film transfer rate may be calculated
from
(40)
where vg is the kinematic viscosity of all (a
function of temperature) as presented in Table
7.8, Z is the water depth, and W is the wind
speed in m s-1 kgi has units of T-1.
Solubility, vapor pressure, and Henry's constant
data are presented in Table 7.9.
Dimensionless Henry's constant refer to a
concentration ratio of mg/L air per mg/L in the
water phase.
Yalkowsky measured the solubility of 26
halogenated benzenes at 25 ºC and developed the
following relationship
(41)
Where Sw is solubility (mol L-1), MP is the
melting point (ºC), and Kow is the estimated
octanol/water partition coefficient.

47
Table 7.8 Kinematic Viscosity of Air
48
Table 7.9 Summary Table of Volatilization Data at
20 ºC
49
Table 7.9 (continued)
50

Lyman et al. compiled solubility data on 78
organic compounds and presented estimation
methods based on Kow for different classes of
compounds. They also included a method based on
the molecular structure.
Mackay measured Henry's constant for 22 organic
chemicals as part of a study of volatilization
characteristics.
Transfer coefficients for the gas and liquid
phases were correlated for environmental
conditions as
(42)
(43)
Where U10 is the 10-m wind velocity (m s-1), ScL
and ScG are the dimensionless liquid and gas
Schmidt numbers.
Volatile compounds such as those shown in Figure
7.3a are easily removed from water and wastewater
by purging with air or by passing them through an
air stripping tower. In natural waters, they are
removed by stripping from the atmosphere.
The overall mass transfer coefficient KL can be
related to that of oxygen (Table 7.10) because so
much information exists for oxygen transfer in
natural waters.

51
Table 7.10 Estimated Henrys Constant and Mass
Transfer Coefficients for Selected Organics at 20
ºC
52
7.2.7 Sorption Reaction

Soluble organics in natural waters can sorb onto
particulate suspended material or bed sediments.
The mechanism and the processes by which this
occurs include
- physical adsorption due to van der Waals
forces
- chemisorption due to a chemical bonding or
surface coordination reaction
- partitioning of the organic chemical into the
organic carbon phase of the particulates.
Physical adsorption is purely a surface
electrostatic phenomenon. Partitioning refers to
the dissolution of hydrophobic organic chemicals
into the organic phase of the particulate matter
it is an absorption phenomenon rather than a
surface reaction, and it may occur slowly over
time scales of minutes to days.
Adsorption isotherms refer to the equilibrium
relationship of sorption between organics and
particulates at constant temperature. The
chemical is dissolved in water in the presence of
various concentrations of suspended solids. After
an initial kinetic reaction, a dynamic
equilibrium is established in which the rate of
the forward reaction (sorption) is exactly equal
to the rate of the reverse reaction (desorption).

The sorption of toxicants to suspended
particulates and bed sediments is a significant
transfer mechanism. Partitioning of a chemical
between particulate matter and the dissolved
phase is not a transformation pathway it only
relates the concentration of dissolved and sorbed
states of the chemical.
The octanol/water partition coefficient, Kow, is
related to the solubility of a chemical in water.
Tables 7.9 and 7.11 provide log Kow values for a
number of organic chemicals of environmental
interest.

Table 7.11 Octanol/Water Partition Coefficients
of Selected Organics, 298 K
54

The laboratory procedure for measuring Kow is
given by Lyman.
1. Chemical is added to a mixture of pure
octanol (a nonpolar solvent) and - pure water (a
polar solvent). The volume ratio of octanol and
water is set at the estimated Kow.
2. Mixture is agitated until equilibrium is
reached.
3. Mixture is centrifuged to separate the two
phases. The phases are analyzed for the chemical.
4. Kow is the ratio of the chemical
concentration in the octanol phase to chemical
concentration in the water phase, and has no
units. The logarithm of Kow has been measured
from -3 to 7.
If the octanol/water partition coefficient cannot
be reliably measured or is not available in
databases, it can be estimated from solubility
and molecular weight information,
(44)
where MW is the molecular weight of the pollutant
(g mol-1) and S is in units of ppm for organics
that are liquid in their pure state at 25 ºC.

For organics that are solid in their pure state
at 25 ºC,
(45)
where MP is the melting point of the pollutant
(ºC) and ?Sf is the entropy of fusion of the
pollutant (cal mol-1 deg-1).
The octanol/water partition coefficient is
dimensionless, but it derives from the
partitioning that occurs in the extraction
between the chemical in octanol and water.
Octanol was chosen as a reference because it is a
model solvent with some properties that make it
similar to organic matter and lipids in nature.
For a wide variety of organic chemicals, the
octanol water partition coefficient is a good
estimator of the organic carbon normalized
partition coefficient (Koc).

Karickhoff et al. and Schwarzenbach and Westall
have published useful empirical equations for
predicting Koc as a function of Kow
(46)
(47)
Once an estimate of Koc is obtained, the
calculation of a sediment/water partition
coefficient suitable for natural waters is
straightforward because the
(48)
where foc is the decimal fraction of organic
carbon present in the particulate matter
(mass/mass).
Figure 7.7 is a schematic of how Kp, Koc, and Kow
are interrelated. Figure 7.7 is the Langmuir
adsorption isotherm for sorption of one chemical
on particulate matter.

57
Figure 7.7Relationship between the
sediment/water partition coefficient Kp, the
organic carbon partition coefficient Koc, and the
octanol/water partition coefficient Kow.Plot (a)
and (b) are for only one chemical and (c) is for
many chemicals.
58

Kp is a measure of the actual partitioning in
natural waters.
The linear portion of the adsorption isotherm
(Figure 7.7a) can be expressed by equation
(49)
The Langmuir isotherm in Figure 7.7a is derived
from the kinetic eqn for sorption-desorption
(50), (51)
where C is the concentration of dissolved
toxicant, Cp is the concentration of particulate
toxicant, Cpc is the maximum adsorptive
concentration of the solids, and k1 and k2 are
the adsorption and desorption rate constants,
respectively.

At steady-state, eqn (51) reduces to a Langmuir
isotherm in which the amount adsorbed is linear
at low dissolved toxicant concentrations but
gradually becomes saturated at the maximum value
(rc) at high dissolved concentrations.
(52)
Generally, the adsorption capacity of sediments
is inversely related to particle size clays gt
silts gt sands. Sorption of organic chemicals is
also a function of the organic content of the
sediment, as measured by Koc, and silts are most
likely to have the highest organic content.
Sometimes a Freundlich isotherm is inferred from
empirical data. The function is of the form
(53)
where n is usually greater than 1. In dilute
solutions, when n approaches 1, the Freundlich
coefficient, K, is equal to the partition
coefficient, Kp.

The partition coefficient is derived from
simplification of the kinetic eqns (50) and (51)
if rc gtgt r (the linear portion of the Langmuir
isotherm). In this case, we may write
(54a), (54b)
Where kf is the adsorption rate constant and kr
is the desorption rate constant.
The total concentration of toxicant
(55)
Where fd and fp are the dissolved and particulate
fractions, respectively
and the ratio of the reaction rate constants is
related by
(58)
Where the 8 subscripts indicate chemical
equilibrium.

(56), (57)
61

From kinetics experiments where dissolved and
particulate concentrations are monitored over
time, the ratio of steady-state concentrations
can be read from the graph (Figure 7.8).
Sorption reactions usually reach chemical
equilibrium quickly, and the kinetic
relationships can often be assumed to be at
steady-state. This is sometimes referred to as
the "local equilibrium" assumption, when the
kinetics of adsorption and desorption are rapid
relative to other kinetic and transport processes
in the system.
O'Conner and Connolly first reported that, for
organics and metals alike, the sediment/water
partition coefficient Kp declines as sediment
(solids) concentrations increase. It is a
consistent phenomenon in natural waters that is
particularly important for hydrophobic organic
chemicals. For example, the Kp for a chemical in
sediments is much lower than that observed in the
water column. Most researchers attribute this
fact to artifacts in the way that one attempts to
measure Kp, including complexation of a chemical
by colloids and dissolved organic carbon that
pass a membrane filter.

62
Figure 7.8 Kinetic sorption experiment in a
batch reactor
63
7.2.8 Bioconcentration and Bioaccumulation

Bioconcentration of toxicants is defined as the
direct uptake of aqueous toxicant through the
gills and epithelial tissues of aquatic
organisms. This fate process is of interest
because it helps to predict human exposure to the
toxicant in food items, particularly fish.
Bioconcentration is part of the greater picture
of bioaccumulation and biomagnification that
includes food chain effects. Bioaccumulation
refers to uptake of the toxicant by the fish from
a number of different sources including
bioconcentration from the water and biouptake
from various food items (prey) or sediment
ingestion. Biomagnification refers to the process
whereby bioaccumulation increases with each step
on the trophic ladder.
The terms bioconcentration, bioaccumulation, and
biomagnification are sometimes mistakenly used
interchangeably. It is useful to accept the
following definitions for the sake of discussion.

Bioconcentration the uptake of toxic organics
through the gill membrane and epithelial tissue
from the dissolved phase.
Bioaccumulation the total biouptake of toxic
organics by the organism from food items
(benthos, fish prey, sediment ingestion, etc.) as
well as via mass transport of dissolved organics
through the gill and epithelium.
Biomagnification that circumstance where
bioaccumulation causes an increase in total body
burden as one proceeds up the trophic ladder from
primary producer to top carnivore.
Bioconcentration experiments measure the net
bioconcentration effect after x days, having
reached equilibrium conditions, by measuring the
toxicant concentration in the test organism. The
BCF (bioconcentration factor) is the ratio of the
concentration in the organism to the
concentration in the water.

The BCF derives from a kinetic expression
relating the water toxicant concentration and
organism mass
(59)
where e efficiency of toxic absorption at the
gill
k1 (L filtered/kg organism per day)
k2 depuration rate constant
including excretion and clearance of metabolites,
day-1
C dissolved toxicant, µg L-1
B organism biomass, kg L-1
F organism toxicant residue (whole
body), µg kg-1
Steady-state solution is
(60)
where BCF has units of (µg/kg)/(µg/L).
Bioconcentration is analogous to sorption of
hydrophobic organics. Organic chemicals tend to
partition into the fatty tissue of fish and other
aquatic organisms, and BCF is analogous to the
sediment/water partition coefficient, Kp.
Bioconcentration also can be measured in algae
and higher plants, where uptake occurs by
adsorption to the cell surfaces or sorption into
the tissues.

An empirical relationship for bioconcentration
(BCF-Kow) in bluegill sunfish in 28 days exposure
for 84 organic priority pollutants was
(61)
and for rainbow trout with ten chlorobenzenes it
was
(62)
for low-level exposures typical of natural
waters. Fathead minnow, bluegill, rainbow trout,
brook trout, and mosquito fish are the species
most frequently involved in bioconcentration
tests.
Bioconcentration experiments, per se, do not
measure the metabolism or detoxification of the
chemical. Chemicals can be metabolized to more or
less toxic products that may have different
depuration characteristics. The bioconcentration
experiment only measures the final body burden at
equilibrium (although interim data that were used
to determine when equilibrium was reached may be
available).
The fact that a chemical bioaccumulates at all is
an indication that it resists biodegradation and
is somewhat "biologically hard" or "nonlabile."

The kinetics of bioaccumulation are shown
schematically in Figure 7.9.
Fish can lose unmetabolized toxics via biliary
excretion or "desorption" through the gill. On
the other hand, toxic organics can undergo
biotransformations and be eliminated as metabolic
products.
The rate constant, k2, includes total depuration
(both excretion of unmetabolized toxics, k2, and
elimination of metabolites, k2). Only a
fraction of this elimination is returned to the
water column as dissolved parent compound,
designated as k2 in Figure 7.9.
Hydrophobic organics tend to accumulate in fatty
tissue of animals. Lipid normalized
bioconcentration factors both in the laboratory
and in the field have been correlated
successfully with the hydrophobicity of toxic
organics as measured by the octanol/water
partition coefficient, Kow (Table 7.12).
Biomagnification occurs in lake trout for PCBs in
the Great Lakes due to the contribution of
alewife and small fish to the diet of these top
carnivores.

68
Figure 7.9 Bioaccumulation kinetics for
hydrophobic organic chemicals in fish
69
Table 7.12Bioconcentration Factor (BCF) for
Selected Organic Chemicals in Fish (Units µg/kg
fish- µg L water)
70
7.2.9 Comparison of Pathway

Most of the transformations discussed in Section
7.2 are expressed as second-order reactions. It
is difficult to compare the magnitudes of these
reactions-the rate constants all have different
units. Each of the transformations can be written
as pseudo-first-order reactions assuming that the
second concentration in the reaction rate
expression can be assumed to be relatively
constant.
The overall reaction rate
(63)
where C dissolved organic concentration, ML-3
t time, T
kb biotransformation rate constant,
T-1
ko oxidation rate constant, T-1
kr reduction rate constant, T-1
kp photolysis rate constant, T-1
kh hydrolysis rate constant, T-1
kv volatilization rate constant, T-1

Equation (63) includes an assumption that the
atmosphere has a neg1igible concentration
(partial pressure) of the organic, so only
volatilization occurs (stripping out of the water
body).
For first-order reactions in a batch reactor
without transport, the reaction rate
(64)
Solving for the concentration as a function of
time
(65), (66)
Taking the natural logarithm of both sides of
equation (66) and solving for time (half-life)
yields the well-known relationship below
(67)
where t1/2 overall half-life of the chemical
due to all transformation reactions
the sum of all the
pseudo-first-order reaction rate constants
Individual half-lives may be compared to
determine which reaction predominates (gives the
shortest half-life).

72
7.3 ORGANIC CHEMICALS IN LAKES 7.3.1
Completely Mixed Systems

As an approximation, lakes can be represented as
ideal completely mixed flow through reactors (CMF
systems) or a network of CMF compartments.
A mass balance system of equations
Figure 7.10 a schematic of the various reactions
in the lake water column and sediment.
An assumption of local equilibrium may be used to
relate the particulate adsorbed concentration to
the dissolved concentration through the partition
coefficient Kp.
(68)
where Kp sediment/water partition
coefficient, L kg-1
C dissolved organic chemical
concentration, µg L-1
r mass sorbed, µg kg-1
M suspended or bed solids
concentration, kg L-1
Cp particulate adsorbed concentration, µg L-1
CT total (dissolved plus particulate)
concentration, µg L-1

73
Figure 7.10 Schematic of a fate model for
organic chemicals in water and sediment
74

Sorptive equilibrium is usually a valid
assumption in natural waters because the time
scale for most sorption reactions (minutes to
hours) is small compared to the time scale for
reactions and transport (days to years).
Figure 7.10 indicates a rapid local equilibrium
assumption for bioconcentration. If uptake and
depuration kinetics (hours to days) are fast
relative to other reactions and time scales, this
is a valid assumption. Use of the
bioconcentration factor (BCF) helps to simplify
the equations, and it is another partitioning
coefficient that we may use similar to Kp.
(69)
where BCF bioconcentration factor, L kg-1
C dissolved chemical
concentration, µg L-1
F residue concentration in whole
fish, µg kg-1
The total concentration of chemical CT may be
larger or smaller in the sediment than the
overlying water depending on whether the water
column or sediment was contaminated first.
Partitioning of the chemical between the
dissolved pore water C2 and adsorbed sediment
Cp2, may also be different due to the dependence
of Kp2, on solids concentration. Generally, Kp2 lt
Kp1, because the sediment has a much higher
solids concentration.

A framework for a mass balance model for an
organic chemical in a lake is given by Figure
7.10. Waste inputs, their fate and effects, can
be assessed in this context.
Anthropogenic inputs may also enter the water
body from the atmosphere via wet precipitation
and dry deposition. The concentration in rainfall
is related to the gas phase concentration and
Henry's constant, so the deposition mass is equal
to the volume of rainfall times the aqueous phase
concentration
(70)
Where Cprecip is the precipitation
concentration, Cg is the gas phase concentration,
and H is Henry's constant with the appropriate
units.
The flux of contaminants due to dry deposition is
related to the depositional velocity and the
gaseous concentration
(71)
where vd is the deposition velocity (LT-1), Cg is
the gas phase concentration (ML-3) and Jd is the
areal mass flux due to dry deposition (ML-2T-1).
Equation (71) is empirical. Both gases and
aerosol particles may contribute to dry
deposition but the gas phase concentration should
be proportional in either case, vd serving as the
empirical proportionality constant.

The mass balance equation for a lake with toxic
organic chemical inputs can be written assuming
complete mixing, steady flow conditions,
instantaneous local sorption equilibrium, and no
atmospheric deposition.
(72)
Equation (72) has three unknown dependent
variables CT, and C - but the assumption of
local equilibrium allows us to write the equation
entirely in terms of total (whole water,
unfiltered) concentration.
(73)
where CT total concentration C Cp, ML-3
V volume of the lake, L3
t time, T
Q flowrate in and out, L3T-1
fp particulate fraction of total
chemical concentration, dimensionless
Cp/ CT KpM /(1 KpM)
fd dissolved fraction of total
chemical concentration, dimensionless
C/ CT 1 /(1 KpM)
C dissolved chemical concentration, ML-3
Cp particulate chemical concentration,
ML-3
ks sedimentation rate constant,T-1

Equation (73) is an ordinary differential
equation with constant coefficients. It is
solvable by first-order methods such as the
integration factor method. Dividing through by
the constant volume and rearranging, we have
(74)
The final solution is
(75)
where CTo initial total input concentration,
ML-3
a integration factor
t mean hydraulic detention time
V/Q, T
We see that the solution to a continuous input of
organic chemical to a lake is composed of two
terms in equation (75) the first term is the
die-away of initial conditions, and the second
term is the asymptotic "hump" (the shape of a
Langmuir isotherm), which builds to a
steady-state concentration as t ? 8.
(76)
The steady-state concentration is directly
proportional to the total concentration of
organic inputs to the lake.

Because it takes an infinite amount of time (or
the lake to reach steady state in the strictest
sense, we speak of time to 95 of steady state,
that is, the length of time required for the
concentration in the lake to reach 95 of the
value that it will ultimately achieve.
(77)
or
(78)
By inspection, one can prove that equations (75)
and (78) are equal when
(79)
Equation (79) gives the time to 95 of steady
state. For the simplest case of a nonadsorbing
dissolved chemical undergoing first-order
reaction decay, a k 1/t.
The greater is the flushing rate ( 1/t) and the
reaction rate constant, the less is time required
to achieve steady state. Conservative substances
(k 0) take the longest time to reach steady
state after a step function change in inputs.

79
Figure 7.13Schematic of lake recovery from a
persistent hydrophobic pollutant
7.3.2 Dieldrin Case Study in Coralville
Reservoir, Iowa

The following case study is used to illustrate
aspects of ecosystem recovery from a persistent
hydrophobic organic pollutant. It also
demonstrates the use of compartmentalization
within a lake to simulate transport.
Figure 7.13 a schematic of water column,
sediment, and fish concentrations following a
period when large discharges of chemical were put
into the system. Because the contaminant is
hydrophobic and persistent, it remains in the
system for a long time, accumulating in fish
tissue and sediment. It disappears by washout
(advection), burial into the deep sediment, and
slow degradation reactions. Depending on the
sediment dynamics of the system and the rate of
chemical degradation, these can be slow processes
taking years to decades.

80
Figure 7.14Selected insecticides used in the
past in the midwestern United States

Figure 7.14 some persistent insecticides (e.g.
chlorinated hydrocarbons) used in the Midwest.
These chemicals were banned in the 1970s and
early 1980s because of their persistence and
propensity to bioaccumulate in fish and wildlife.
Also shown are two replacement insecticides
(ester compounds), which hydrolyze and break down
in the environment. They are toxic but much less
persistent.

Agricultural usage of pesticides in Iowa is
widespread, particularly grass and broadleaf
herbicides and row crop soil insecticides. One of
the insecticides widely used for control of the
corn rootworm and cutworm from 1960 to 1975 was
the chlorinated hydrocarbon, aldrin.
Aldrin is microbially metabolized to its
persistent epoxide, dieldrin. Dieldrin is itself
an insecticide of certain toxicity and is also a
hydrophobic substance of limited solubility in
water (0.25 ppm) and low vapor pressure (2.7
10-6 mm Hg at 25 ºC). It is known to
bioaccumulate to levels as high as 1.6 mg/kg wet
weight in edible tissue of Iowa catfish.
Coralville Reservoir is a mainstream impoundment
of the Iowa River in eastern Iowa. It drains
approximately 3084 square miles (7978 km2) of
prime Iowa farmland and receives extensive
agricultural runoff with 90 of its drainage
basin in intensive agriculture. It is a
variable-level, flood control and recreational
reservoir, which has undergone considerable
sedimentation since it was created in 1958.
At conservation pool (680 ft above mean sea
level, msl), the reservoir has a capacity of
38,000 acre-ft (4.79 107 m3), a surface area of
4900 acres (1.98 107m2), a mean depth of
approximately 8 ft (2.44 m), and a mean detention
time of 14 days. In 1958, the capacity at
conservation pool was 53,750 acre-ft (6.63 107
m3).

The total pesticide concentration is the sum of
the particulate plus the dissolved
concentrations, with instantaneous sorptive
equilibrium assumed
(80)
where fd C/ CT 1/(1 KpM) fraction of
dissolved pesticide
fp Cp/ CT KpM/(1 KpM) fraction
of particulate pesticide
W(t) time-variable loading of
pesticide, M/T
CT total concentration in the water
column, ML-3
sum of the pseudo-first-order
degradation rate constants
t mean hydraulic detention time
V reservoir volume, L3
ks sedimentation rate constant, T-1
The fish residue equation is
(81)
where k1 pesticide uptake rate by fish, T-1

Equations (80) and (81) may be solved
analytically for constant coefficients and simple
pesticide loading functions, W(t), or they may be
integrated numerically. In the case of a
pesticide ban, the W(t) might typically decline
in an exponential manner due to degradation by
soil organisms or a ban on application.
For an exponentially declining loading function
at rate ?, the analytical solutions to equations
(80) and (81) are
(82)
(83)
where CTo initial total pesticide
concentration in lake, ML-3
CTin,o initial total pesticide inflow
concentration, ML-3
? rate of exponentially declining
inflow concentration,T-1

84
Figure 7.15Compartmental configuration for a
two-box pond model or an eight-box lake model

Figure 7.15 is a schematic diagram of
hypothetical pond or lake configurations that are
possible for this problem. Each box is assumed to
be completely mixed with bulk exchange between
water compartments. There is dispersion in
Coralville Reservoir that seems to be simulated
best by the eight-compartment model based on dye
studies.

85
Figure 7.16Result of model and field data for
dieldrin in Coralville Reservoi

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