PB87-186433
FATE AND PERSISTENCE IN SOIL OF
SELECTED TOXIC ORGANIC CHEMICALS
PEI Associates, Incorporated
Cincinnati, OH
May 87
U.S. DEPARTMENT OF COMMERCE
National Technical Information Service
N IS,
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EPA/600/6-87/003
Hay 1987
FATE AND PERSISTENCE IN SOIL OF
SELECTED TOXIC ORGANIC CHEMICALS
by
Roxanne Sukol, Edwin Wool son, and Mil 11 am Thompson
PEI Associates, Inc.
11499 Chester Road
P. 0. Box 46100
Cincinnati, Ohio 45246
Contract 68-02-3976
Work Assignment 14
Contract 68-02-4248
Work Assignment 22
Project Officers
Michael H. Shapiro
William M. Burch
Office of Toxic Substances
U. S. Environmental Protection Agency
Washington, D. C. 20460
Task Manager
Charles H. Nauman
Exposure Assessment Group
Office of Health and Environmental Assessment
U. S. Environmental Protection Agency
Washington, D. C. 20460
OFFICE OF HEALTH AND ENVIRONMENTAL ASSESSMENT
OFFICE OF RESEARCH AND DEVELOPMENT
U. S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D. C. 20460
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TECHNICAL REPORT DATA
(Pleoic read laurucliont on the reverie before completing)
1 REPORT NO.
EPA/600/6-87/003
3. RECIPIENT'S ACCESSION NO.
4. TITLE AND SUBTITLE
Fate and Persistence in Soil of Selected
Toxic Organic Chemicals
8. REPORT DATE
May 1987
6. PERFORMING ORGANIZATION CODE
7. AUTHOR(S)
8. PERFORMING ORGANIZATION REPORT NO.
Edwin Wool son, Roxanne Sukol, and William Thompson
, PERFORMING ORGANIZATION NAME AND ADDRESS
PEI Associates. Inc.
11499 Chester Road
P. 0. Box 46100
Cincinnati, Ohio 45246
10. PROGRAM ELEMENT NO.
11. CONTRACT/GRANT NO.
Contract 68-02-3976
68-Q2-424R
12. SPONSORING AGENCY NAME AND ADDRESS
13. TYPE OF REPORT AND PERIOD COVERED
Office of Health & Environmental Assessment
Exposure Assessment Group (RD-689)
U.S. Environmental Protection Agency
Washington. D.C. 20460
14. SPONSORING AGENCY CODE
EPA/600/21
is. SUPPLEMENTARY NOTES EPA Project officers: Michael H. Shapiro and William M. Burch,
Office of Toxic Substances, Washington, DC (382-3667, 3664). EPA Task Manager:
Charles H. Nauman. Office of Health and Fnvirnnmpntal A«t«;pg«;mpnt Uach-inntrm n
16. ABSTRACT
The persistence of toxic and generally refractory halogenated hydrocarbons in
the environment is a key factor in evaluating human exposure. This report summarizes
the chemical and physical properties of some of these compounds and addresses how
these properties can affect their persistence and behavior in various environmental
media. The property that affects persistence and mobility of organic compounds in
soil most directly is water solubility. Within a class of compounds the higher the
degree of halogenation, the lower the water solubility, and thus, the greater the
persistence. Persistence in the environment is dependent also upon several environ-
mental factors, including soil organic matter, total precipitation and intensity,
temperature, intensity of sunlight, and soil texture. In general, the organic carbon
content of soil has the greatest effect on the behavior of hydrophobic organic com-
pounds, as these compounds sorb strongly to the organic matter in the soil. Sorbed
organic compounds in soil are subject to several possible fates in the environment,
including volatilization, microbial degradation, photodecomposition on the soil
surface, translocation to plants, chemical degradation, and leaching to ground water.
Some of these processes are directly related to the degree of sorption. Estimates
of the environmental half-lives of the compounds considered here and others are
uncertain because of the variability in how these fate processes will be influenced
under various environmental conditions.
17.
KEY WORDS AND DOCUMENT ANALYSIS
DESCRIPTORS
b.IDENTIFIERS/OPEN ENDED TERMS C. COSATI Field/Group
IB. DISTRIBUTION STATEMENT
Distribute to Public
19. SECURITY CLASS (ThaReport)
21 NO. OF PAGES
126
20. SECURITY CLASS (This page)
Unclassified
22. PRICE
EPA Form 2220-1 (9-73)
1
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DISCLAIMER
This report has been reviewed in accordance with U.S. Environmental
Protection Agency policy and approved for publication. Mention of trade
names or commercial products does not constitute endorsement or recommenda-
tion for use.
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FOREWORD
The Exposure Assessment Group (EAG) of EPA's Office of Health and Environ-
mental Assessment has three main functions: 1) to conduct exposure assessments;
2) to review assessments and related documents; and 3) to develop guidelines
for Agency exposure assessments. The activities under each of these functions
are supported by and respond to the needs of the various EPA program offices.
As part of the third function, EAG sponsors projects aimed at developing or
refining techniques used in exposure assessments. This study, which is one of
these projects, was done for the Office of Emergency and Remedial Response.
In recent years EPA has focused much attention on halogenated organic
compounds, which tend to be toxic and refractory and to bind tightly to soil.
The degree of persistence of these compounds in soil can be a key variable in
estimating ecological and human exposure when they are disposed of or spilled
in the environment. This report focuses on an evaluation of the sorptive
tendencies for soil of halogenated benzenes, azobenzenes, cyclohexanes,
dibenzofurans, dibenzo-p-dioxins, biphenyls, and naphthalenes. The evaluation
also addresses how soil characteristics and environmental factors may influence
the persistence of these compounds in soil.
Michael A. Callahan
Director
Exposure Assessment Group
iii
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ABSTRACT
This report reviews the environmental fate and behavior of several toxic
organic materials. It summarizes the chemical and physical properties of
these materials and discusses how these properties affect their persistence
and behavior in the soil/water/air systems.
In general, the organic carbon content of a soil has the greatest effect
on the behavior of hydrophobic toxic organic compounds. The organic com-
pounds sorb strongly to the organic matter in the soil. Several equations
have been derived that define water solubility relationships. These are
partition coefficients between octanol/water and organic matter/water.
Persistence of the toxic organic compounds depends on several environ-
mental factors, including soil organic matter, total precipitation and in-
tensity, temperature, sunlight intensity, and soil texture. Organic chem-
icals are subject to one or more of seven possible fates: 1) sorption,
2) volatilization, 3) microbial degradation, 4) photodecomposition on the
soil surface, 5) translocation to plants, 6) chemical degradation, and
7) leaching to ground water. Some of these fates are directly related to the
degree of sorption; i.e., very little of a material that is strongly sorbed
will be in solution and available for degradation or movement by the other
processes.
Some generalities are presented regarding the environmental conditions
and chemical/physical properties that affect persistence and mobility; how-
ever, the reader should bear in mind that there are always exceptions to the
rule.
Disregarding any interactions between environmental conditions, the
following effects might be expected:
1) Temperature—The warmer the temperature is, the greater the vola-
tility, the lower the organic matter content of the soil, the more
active the microbial population, and the higher the rate of evapo-
transpiration. The result is a decrease in pesticide persistence.
2) Moisture—There is an optimum level of soil moisture for microbial
activity. If a soil is too wet or too dry, activity slows down.
Volatility is also affected by moisture content; the nature of the
effect depends on the solubility of the chemical. The total
amount, the intensity, and the frequency cf rainfall or irrigation
water received affect the movement of chemicals in soil (Bailey
1966).
iv
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3) Li ght—Photochemica1 reactions are directly proportional to the
number of photons absorbed by a chemical. Nearness to the equator
or an Increase In altitude will accelerate photochemical reactions.
4) Soil texture—Soil texture Is an important factor. Soil organic
matter is directly influenced by the soil texture. Coarse (i.e.,
sandy) soils will normally be low in organic matter; therefore,
water percolation will be rapid and the leaching potential of
chemical compounds will be high regardless of K or K values.
The opposite is true for heavy (i.e., clayey) soils.
The property that affects persistence and mobility most directly is
water solubility. Within a class of compounds (e.g., dioxins or PCB's), the
higher the degree of chlorination or bromination is, the lower the water
solubility and, therefore, the greater the persistence.
Low-molecular-weight compounds with low chlorine content (e.g., chloro-
benzene, dichlorobenzene, naphthalene) will be subject to a greater degree of
biodegradation, photodecomposition, volatilization, and leaching than will
high-molecular-weight compounds with higher chlorine or bromine content
(e.g., hexachlorobenzene, dibenzodioxins and dibenzofurans, PCB's, PBB's, and
DDT and its related compounds).
The literature search revealed a sparsity of information on many of the
compounds discussed in this document, and gaps were numerous. No information
was found on biphenylenes and azoxybenzenes.
Half-life estimates between compounds were difficult to compare because
of the differences in experimental and/or environmental conditions. If a
standard set of conditions were adopted, and half-life estimates were de-
veloped for a well-studied compound (e.g., DDT) under each set of conditions,
other compounds could be studied under these same standard conditions and
half-life estimates could then be calculated relative to the standard mate-
rials. These relative half-lives could then be compared and used to predict
behavior based on similarities and differences among other compounds of
interest.
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CONTENTS
Page
Disclaimer ii
Foreword i i i
Abstract iv
Figures vii
Tables vii
Acknowledgements ix
1. Introduction 1
2. Soil Properties 4
2.1 Composition 4
2.2 Organic Matter 5
2.3 Water Content 6
2.4 Soil Temperature 6
2.5 Cation Exchange Capacity (CEC) 8
2.6 pH 9
3. Chemical Fates 10
3.1 Soil Sorption 11
3.2 Volatilization 18
3.3 Microbial Degradation 22
3.4 Photodecomposition 24
3.5 Translocation to Plants 25
4. Partitioning of Organic Compounds in the Environment 26
4.1 Partitioning Models 26
4.2 Environmental Fate Models 31
4.3 Effect of Climate on Chemical Fate 32
4.4 Likely Chemical Fates 33
5. Information on Selected Organic Chemicals 36
5.1 Hexachlorobenzene (HCB) 36
5.2 1,2-Dichlorobenzene 40
5.3 3,3'4,4'-Tetrachloroazobenzene (TCAB) 43
5.4 Hexachlorocyclohexane (HCH) 46
5.5 2,3,7,8-Tetrachlorodibenzofuran (TCDF) 54
5.6 2,3,7,8-Tetrachlorodibenzo-p-Dioxin (TCDD) 56
5.7 Polychlorinated Biphenyls (PCB's) 61
5.8 Polybrominated Biphenyls (PBB's) 64
5.9 Polychlorinated Naphthalenes 67
5.10 Toxaphene 69
5.11 Hexachlorobutadiene 73
5.12 DDT (Dichlorodiphenyltrichloroethane) and its Metabolites,
DDE (Dichlorodiphenyldichloroethane) and DDE (Dichlorodi-
phenyldichloroethylene) 76
6. Information Gaps 93
Bibliography B-l
References R-l
vi
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FIGURES
Number Page
1 Biodegradation of DDT 81
TABLES
Number Page
1 Representative Organic Compounds Selected for Review 2
Based on Potential Toxiclty 1n the Environment
2 General Properties of Soil Organic Matter 7
3 Measured K for Environmental Assessment of Toxic
Substances* 15
4 Measured Log K for Selected Compounds 16
5 Various Measured and Estimated Partition Coefficients and
Properties for Environmental Assessment of Selected Toxic
Organic Substances (MEAN) 19
6 Diffusion Behavior of Toxic Organic Compounds In a Soil
with 10 Percent Moisture 21
7 Relative Bioconcentration, Mlcrobial Degradation, and
Photodegradatlon of Selected Organic Compounds In
Activated Sludge 23
8 Chemical Distribution Using Various Models 29
9 Percentage Partitioning of Compounds 1n a Unit World with
the Fugacity I Model 30
10 Likelihood of Environmental Fates of Toxic Organic Com-
pounds In Six Major Climates 32
11 Prospect of Biotransformation of Dichlorobenzenes in
Water-Table Aquifers 42
12 3,3',4,4'-TCAB Produced in Ontario Soils 45
vii
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TABLES (continued)
Number Page
13 Mobility Factors for Lindane Metabolites 50
14 Lindane Recovery After Eight Weeks 52
15 Toxaphene Degradation Under Various Conditions 71
16 Vapor Pressures of DDT and Metabolites 78
17 Potential DDT Volatilization 79
18 Reactions of Environmental Importance for Transformation
of DDT 82
19 Additional Losses of DDT by UV Radiation 83
20 Mobility Factors for DDT and Metabolites 84
21 Half-Life Estimates for Various Soils 91
22 Summary of Properties and Half-Life Estimates for
Toxic Organic Chemicals 92
viii
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ACKNOWLEDGMENTS
The authors wish to thank Dr. Charles H. Nauman and Mr. John L. Schaum
of the Environmental Protection Agency for their assistance and guidance
during the course of this study.
ix
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SECTION 1
INTRODUCTION
The Implementation of environmental programs to clean up organic chem-
icals that have been released onto and contaminated soils requires an under-
standing of the long-term risk associated with leaving the soil In place or
transporting It to an ultimate disposal site. Currently available Informa-
tion on the persistence of highly toxic organic chemicals Is too sparse to
allow development of exposure assessments with the degree of confidence
needed by the regulator or by the public. The U.S. Environmental Protection
Agency (EPA) recognizes the need for guidance materials to assist in the
determination of long-term human health risks posed by persistent toxic
compounds.
The objective of the guidance material in this document is to provide
information needed to support procedures for estimating the environmental
half-life of compounds having a high affinity for soils. Specific compounds
discussed herein were selected from the following groups:
Chlorinated benzenes Halogenated biphenyls
Chlorinated azobenzenes Halogenated biphenylenes
Chlorinated azoxybenzenes Chlorinated naphthalenes
Chlorinated cyclohexanes Halogenated dibenzofurans
Halogenated dibenzodioxins
Also identified for investigation were toxaphene, DDT, and hexachlorobuta-
diene.
Selection of representative compounds or mixtures (e.g., toxaphene,
polychlorinated biphenyls) was based on their potential for human toxicity
and the availability of information on their behavior in soil (Table 1).
Some compounds (e.g., DDT and y-hexachlorocyclohexane) have been studied by
many investigators; therefore, much information is available. Other
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compounds, however, including the entire groups of biphenylenes and azoxyben-
zenes, have not been studied. Because information is sparse on these latter
compounds, they are not addressed here.
TABLE 1. REPRESENTATIVE ORGANIC COMPOUNDS SELECTED FOR REVIEW BASED ON
POTENTIAL TOXICITY IN THE ENVIRONMENT
Group
Compound
Chlorinated benzenes
Chlorinated azobenzenes
Chlorinated cyclohexanes
Halogenated dibenzofurans
Halogenated dibenzodioxins
Halogenated biphenyls
Chlorinated naphthalenes
Related compounds
Hexachlorobenzene
1,2-Dichlorobenzene
3,3',4,4'-Tetrachloroazobenzene (TCAB)
Y-Hexachlorocyclohexane (HCH)
2,3,7,8-Tetrachlorodibenzofuran (TCDF)
2,3,7,8-Tetrachlorodibenzodioxin (TCDD)
Polychlorinated biphenyls (PCB's)
Polybrominated biphenyls (PBB's)
Polychlorinated naphthalenes
Toxaphene
Hexachlorobutadi ene
DDT
Before the primary objective of this guidance material could be achieved,
several lesser objectives had to be identified and addressed. For example,
this study addresses how soil characteristics, physical/chemical processes,
biological processes, chemical structure, microorganisms, and interactions
influence the persistence of compounds in soils. It also evaluates the in-
fluences of various environmental factors (including solar radiation, tempera-
ture, moisture, pH, Eh, and the presence of other chemicals) on persistence
in and affinity for soils.
Wherever possible, the report includes information on the behavior of
toxic organic compounds under varying soil and climatic conditions throughout
the United States. Factors responsible for half-life variability under
various conditions are also identified.
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A method for estimating the half-life of compounds for which no such
data are available is discussed. This method uses the K partition coeffi-
oc
cient to place compounds with unknown half-lives in a position relative to
those whose half-lives are predicted with some degree of certainty. Finally,
the relative soil adsorptive characteristics, persistence, and toxicity of
the selected organic compounds are presented.
Section 2 of the report provides a review of soil and its physical/chem-
ical properties, including composition and texture, water content, pH, or-
ganic matter, cation exchange capacity, and temperature. Section 3 describes
each of the major pathways of chemical loss from the soil. Section 4 pre-
sents an explanation of partitioning in the environment and includes informa-
tion on determining the likelihood of each chemical's fate. Section 5 pro-
vides information on the persistence, toxicity, and half-life of the specific
organic compounds selected for investigation.
The materials used for this report were drawn from an existing litera-
ture data base. The sources are cited in the text and included in a list of
references. A bibliography of other available sources is also included for
the reader's convenience.
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SECTION 2
SOIL PROPERTIES
This section presents a general review of major soil properties and
their interrelated effects. An understanding of the physical and chemical
properties of soil is vital to a discussion of the persistence of toxic
compounds in that medium.
2.1 COMPOSITION
Soil is a three-dimensional product of nature that has resulted from
destructive processes such as the weathering of rock and primary/secondary
minerals, microbial decay of organic material, and synthetic processes (e.g.,
the formation of clays and the development of characteristic compositional
layering patterns) (Brady 1974). Soil consists of four major components:
1) inorganic mineral material, 2) organic matter, 3) water, and 4) air.
These components exist in solid, liquid, and vapor phases and are finely sub-
divided and intimately mixed. The ratio of one component to another differs
substantially from one location to another.
The inorganic mineral material present in soil consists primarily of
quartz, feldspars, silicate clays, and iron and aluminum oxides and hydrox-
ides. Organic matter forms and may accumulate as a result of the decay and
synthesis of plant and animal residues. The water and air components of soil
are held in the maze of pores located between the inorganic and organic
solids. Air moves into those pores that are not filled with water.
Soils that are predominantly mineral in composition generally contain
approximately 1 to 5 percent organic matter. In contrast, soils from bogs,
swamps, and marshes contain 80 to 95 percent organic matter.
The three recognized broad soil textural classes (classified by the
varying size of soil particles) are sand, loam, and clay. Sand refers to
soil containing at least 70 percent sand-sized particles (between 0.05 and
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2.0 mm In diameter). Sand has a low organic matter content and a high bulk
density. (Bulk density Is a weight measurement that takes Into account pore
spaces as well as soil solids.) A high bulk density Indicates that the
particles generally lie in close contact with each other. Clay refers to
soil that has at least 35 percent clay-sized particles (less than 0.002 to
0.005 mm in diameter). Finely textured clay soil has a low bulk density and
a relatively high organic matter content. Loam refers to soil that has a
relatively even mixture of sand and clay particles (Brady 1974).
The voids between individual soil grains are referred to as soil pores.
Their size, shape, and orientation affect the movement of air and water
through the soil. The volume of pores in soil (porosity) is expressed as a
percentage of a given soil volume.
2.2 ORGANIC MATTER
Climate, topography, and vegetation control the levels of organic matter
and the rate of its decomposition in soil. The organic matter content of
soil has widely ranging effects on the soil's capacity to affix or degrade
organic chemical compounds that have been added. Factors affecting the
amount of organic matter in soil include soil temperature, soil moisture, the
presence of oxidizing agents, and the rate of incorporation. For example,
the rate of decomposition of organic matter is restricted in cooler climates;
thus, if all other factors are similar, the organic content of soils in
cooler northern climates tends to be greater than that in wanner areas.
Organic matter content affects soil microbial activity, cation exchange
capacity (CEC), and buffering capacity. Organic soils develop an even lower
pH than do acidic mineral soils. Organic matter accounts for a large portion
(20 to 70 percent) of the CEC of a soil. The magnitude of the CEC, in turn,
determines buffering capacity. Therefore, organic soils show a marked re-
sistance to pH changes compared with that shown by mineral soils.
The introduction of additional organic matter to soil increases soil
porosity and decreases bulk density. The additional organic matter also
increases the soil's water-holding capacity by increasing the number of small
pores, and greater force is necessary to drain small pores than large pores
(Bohn et al. 1985, Brady 1974, Khaleel 1981).
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Organic matter content is the best single indicator of a soil's ability
to adsorb nonionic hydrophobic organic chemicals. It is the primary soil
constituent responsible for chemical sorption, and its role is similar in
different soils. Table 2 summarizes some pertinent information about the
effects of organic matter on the characteristics of soil.
2.3 WATER CONTENT
Cohesion (the attraction of water molecules to one another), adhesion
(the attraction of water molecules to solid surfaces), and soil porosity
affect how a soil retains water and controls its movement and utilization.
Generally, stable, coarse-textured soils with large interconnected pores
(i.e., sand) have lower adhesive capabilities than fine-textured, unstable
soils and thus have a faster water conduction rate. Clay minerals and clay-
sized particles tend to clog connecting pore channels, which increases ad-
hesion and reduces water movement. As water movement is reduced, the water
content of the soil increases.
The rate of water movement in unsaturated soils is slower than in satu-
rated soils because the air that partially fills many pores must be displaced
before the water can move.
The tendency of dry soils to retain organic compounds is stronger, and
the amount retained is greater than in wet soils because water molecules
compete with and displace sorbed compounds from organic matter present in the
soil. As a soil dries out, the hydrophobic organic molecules may become
trapped within collapsing clay lattices and not be available for leaching.
The tendency of a compound to be adsorbed is inversely related to its
tendency to be leached. Compounds that are strongly adsorbed are generally
riot leached to a significant degree (Bohn et al. 1985, Brady 1974).
2.4 SOIL TEMPERATURE
The temperature of the soil affects chemical and biological reaction
rates and the absorption and transport of water. Soil temperature depends
directly or indirectly on 1) the net amount of heat absorbed by the soil,
2) the heat energy required to bring about a given change in soil tempera-
ture, and 3) the energy used to evaporate water.
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TABLE 2. GENERAL PROPERTIES OF SOIL ORGANIC MATTER3
Property
Remarks
Effect on soil
Color
Hater retention
Combination with clay
minerals
Chelatlon
Solubility In water
pH relations
Cation exchange
Mineralization
Combination with organic
molecules
The typical dark color of many soils Is caused
by organic matter.
Organic matter can hold up to 20 times Its weight
In water.
Organic matter joins soil particles Into structural
units called aggregates.
Organic matter forms stable complexes with Cu ,
Mn , Zn , and other polyvalent cations.
Insolubility of organic matter results partially
from Its association with clay; salts of divalent
and trlvalent cations with organic matter are also
Insoluble; Isolated organic matter Is partly solu-
ble In water.
Organic matter buffers soil pH In the slightly
acid, neutral, and alkaline ranges.
Total acidities of Isolated fractions of organic
matter range from 3,000 to 14,000 mules/kg
Decomposition of organic matter yields C02, NH,,,
NO;. POi,"3. and SOU"2.
Organic matter affects bloactivlty, persistence,
and blodegradablllty.
May facilitate warming (by solar radiation).
Helps prevent drying and shrinking; Improves
moisture retention In sandy soils.
Permits gas exchange; stabilizes structure;
Increases permeability.
Buffers the availability of trace elements to
higher plants.
Little organic matter is lost by leaching.
Helps to maintain a uniform reaction (pH) 1n the
soil.
Increases the cation exchange capacity (CEC) of the
soil; from 20 to 70 percent of the CEC of many soils
Is caused by organic matter.
Is a source of nutrient elements for plant growth.
Modifies the application rate of pesticides for
effective control.
Stevenson 1982.
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The amount of heat absorbed by the soil is determined primarily by the
quantity of effective solar radiation reaching the Earth. Solar radiation
levels are affected by latitude and climatic conditions. As latitude de-
creases, the Sun's rays become more direct and thus increase the amount of
radiation reaching the soil level. If latitude is equal, climate affects
solar radiation. For example, in a cloud-free arid region, 75 percent of the
available solar radiation reaches the Earth's surface; whereas, in a cloudy
humid region, only about 35 percent of the solar radiation reaches the
Earth's surface.
Soil color, slope of the land, the altitude, and the vegetative cover
also affect the amount of energy the soil absorbs. Dark soils absorb more
energy than light soils. The closer the angle at which light strikes the
Earth is to 90 degrees, the greater the heat that is absorbed. Increased
altitude causes climatic changes similar to increased latitude, whereas
vegetation has an insulating effect on the ground it covers.
Conduction of solar radiation into the soil is controlled principally by
the amount of moisture available. Heat passes from soil to water 150 times
more readily than from soil to air. As water content increases, air content
decreases and heat transfer increases. In poorly drained soil, large amounts
of radiant energy are required to raise the soil temperature. As heat input
is increased, the resulting increase in molecular activity of the soil water
causes the water to evaporate, which actually has a cooling effect at the
soil surface.
Generally, changes in soil temperature occur slowly and lag behind those
occurring at the soil-air interface. For example, in temperate regions,
surface soil is cooler in winter and warmer in summer than the underlying
subsoil.
2.5 CATION EXCHANGE CAPACITY (CEC)
Cation exchange capacity is the ability of soil materials to adsorb
positively charged inorganic or organic chemicals or compounds. The CEC of
different soils ranges from 2 meq/100 g of soil* to as much as 150 meq/
100 g of soil. The CEC differences result primarily from the amount of
Mi Hi equivalents per 100 grams of soil,
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organic matter (20 to 70 percent), crystalline clay structure, and location
of Ionic substitution in the clay lattice.
Most soils have a slight negative imbalance that is offset by soluble
cations. Exchangeable cations compete for sites on the soil solids, and
uncharged hydrophobia organic molecules migrate to the organic matter or
volatilize if the vapor pressure is high enough.
As the pH rises, hydrogen ions held by the soil are displaced by ex-
changeable cations and neutralized. Also, as discussed in the following
subsection, the removal of aluminum hydroxy ions results in the formation of
Al(OH)j and the availability of additional exchange sites on the mineral
colloids (Brady 1974, Theng 1974).
2.6 pH
Soil pH is controlled by the activity of hydrogen, and the hydrogen
activity is influenced by cations present in the soil solution. Soil mate-
rials differ in their abilities to furnish hydrogen ions to the soil solu-
tion. Hydrogen (H ) and aluminum (Al ) ions dominate acid soils. The H
ions contribute directly to the concentration of hydrogen ions in solution.
The Al ions contribute indirectly through hydrolysis, as follows:
Al+3 + H20 -> A1(OH)+2 + H+ (Eq. 2-1)
A1(OH)+* + H0 + AKOH)* + H+ (Eq. 2-2)
Basic pH conditions result from factors that encourage buildup or maintenance
+3 +
of a supply of cations (called exchangeable bases) other than Al and H
[e.g., calcium (Ca), magnesium (Mg), potassium (K), and sodium (Na)]. Leach-
ing encourages acidity by removing cations that would compete with H and
+3
Al on an exchange complex.
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SECTION 3
CHEMICAL FATES
Complex physical and chemical Interactions control the long-term persis-
tence of organic compounds In the soil environment. Persistence Is affected
by the peculiar chemical and physical characteristics of both the compounds
themselves and the soils In which they are found. These characteristics
Include but are not limited to the rate and amount of soil water movement,
soil temperature, soil texture and organic matter content, and the chemical's
water solubility, light Intensity, vapor pressure, character, shape, molecular
size, and configuration.
Organic chemicals entering soil are subject to one or more of seven
possible fates: 1) sorptlon, 2) volatilization, 3) mlcrobial degradation, 4)
photodecomposltlon (on soil surface), 5) translocation to plants, 6) chemical
degradation, and 7) leaching to ground water. The fate of a chemical In an
environmental system depends greatly on its sorptive behavior. Various
processes and factors, such as those listed above, control the extent of
compound's entry Into these pathways.
The single most Important factor governing persistence of organic com-
pounds in soil is the amount of organic matter contained in the soil. Organic
compounds tend to be adsorbed to the organic matter in soil, thereby
decreasing their availability to the six remaining pathways. Adsorption and
desorption of a chemical to organic matter in the soil are referred to collec-
tively as sorption. Factors such as chemical polarity, water solubility, and
hydrophobicity affect the extent of the compound's sorotion. Generally, the
more hydrophobic, less polar, and less water-soluble a compound is, the
greater the extent to which it will adsorb to the organic matter in the soil.
The amount of a chemical undergoing sorption governs the amount that is
available to be subjected to the other six fates. This is because the move-
ment of a chemical away from its initial deposition site is governed by the
amount present in the soil solution and vapor phases. Any molecules that are
10
-------
not adsorbed to soil organic matter remain in the soil solution. Soil solu-
tion is defined as the aqueous liquid phase of the soil and its solutes and
consists of ions dissociated from the surface of soil particles and other
soluble materials. Organic compounds within the soil solution are available
to volatilize, biodegrade, photodecompose, translocate to plants, or leach to
ground water. Table 1 presents a list of the organic chemicals selected for
evaluation in this document. These chemicals are toxic to humans and appear
to have a high affinity for and persistence in soil.
In addition to having an obvious effect on physical transport, sorption
can be involved directly or indirectly in degradation. The chemical reactivity
of a compound in a sorbed state may differ significantly from that in aqueous
solution, both in extent and chemical pathway. Moreover, natural sorbents
may indirectly mediate solution-phase processes by altering the concentration
of the solution-phase compound or by controlling compound release into the
aqueous phase, which may limit the rate of the solution-phase reaction. In
addition, natural sorbents introduce into solution a buffered suite of
inorganic and organic species that may significantly affect compound
reactivity in the aqueous phase.
Because the compounds considered in this report have low polarity and
water-solubility, soil pH, hydrolysis, charge distribution, polarizability,
and cation exchange capacity are not considered important.
This section addresses the likely fate of each chemical. General con-
clusions on the likelihood of sorption, volatilization, and degradation by
light or microorganisms are presented by groups of chemicals. The relative
water-insolubility of these chemicals makes leaching an unlikely fate.
Section 5 contains detailed experimental data on measured concentrations of
the chemicals in soil, partition coefficients, half-life estimates, and
background information, including structure, toxicity, applications, sources,
and metabolites, for each individual chemical.
3.1 SOIL SORPTION
Natural sorbents may be biotic or abiotic, organic and/or inorganic, or
chemical composites thereof. They may range in size from macromolecules to
gravel. All the organic compounds of interest are only slightly water-soluble.
This limited solubility allows the development of simplistic sorption models
with some degree of theoretical legitimacy.
11
-------
It has long been recognized that sorptlon of pesticides and related
organic compounds on soils determines their fate and transport in the envi-
ronment. Thus, a considerable amount of data has been collected over the
past two decades to characterize sorption on soils and sediments. Since the
EPA's development of a list of priority pollutants under the Clean Water Act,
data have been and continue to be collected on sorption of organic compounds
on soils and sediments.
A number of mathematical equations can be used to describe the convec-
tive, diffusive, and dispersive transport of organic compounds during transient
water flow as functions of the soil bulk density, hydrodynamic dispersion
coefficient, pore-water velocity, concentration in solution, concentration in
the sorbed phase, time, and distance moved. Sinks for degradation and plant
uptake are ignored.
Sorption/desorption processes can be described as a function of equi-
librium or nonequilibrium conditions. For equilibrium conditions, linear and
Freundlich isotherm models have been used most commonly for organic chemical
sorption on soils and sediments. These equations can be stated as follows:
S=KC (linear) (Eq. 3-1)
and
S=KfCN when N<1, (Freundlich) (Eq. 3-2)
where S = Adsorbed-phase concentrations at equilibrium
C = Solution-phase concentrations at equilibrium
K, Kf, and N = Empirical constants specific to a sorbent-sorbate
combination
For N=l, Equation 3-2 reduces to Equation 3-1. The values of K, K^, and N
are usually obtained by curve-fitting the two equations to measured data.
The value of the sorption coefficients K or Kr (also referred to as the
partition coefficient) is a measure of the extent of chemical sorption on
soils. These values may vary by one order of magnitude or more among several
soils and sediments for a given chemical. Multiple regression analyses of K
(or K.) with several physical chemical properties of soil suggest that the
soil organic carbon content (OC) may be the single best predictor of sorption
coefficients for nonionic and nonpolar compounds. For a given compound, most
researchers report that the sorption coefficient normalized with respect to
12
-------
soil's organic carbon content Is essentially Independent of soil type. This
normalized sorption coefficient, designated as K , is defined as follows:
Koc = 10° K/OC x 10° for linear 1sotherms
and
KOC = 100 Kf/OC x 100 for nonlinear isotherms. (Eq. 3-4)
The values of K for a broad range of compounds may be found in reviews by
Hamaker and Thompson (1972), Rao and Davidson (1980), Kenega and Goring
(1981), and Karickhoff (1981). Recent work suggests that a KQ(. value for a
given compound is also independent of particle-size fractions within or among
different soils and sediments (Karickhoff et al. 1979, Rao et al. 1982). All
of these findings suggest that organic carbon is the principal sorbent of
organic substances in soils and sediments. This does not imply, however,
that inorganic soil constituents do not act as sorbents. Sorption on clay
minerals is important for cationic pesticides (e.g., the herbicides paraquat
and diquat). Because soil organic matter and clay minerals exist in soils
largely as clay-metal-organic complexes (Stevenson 1976), however, separating
the independent contributions of clay minerals and organic matter to compound
sorption may be difficult. The K concept is a practical and useful simplifi-
cation for estimating sorption of nonionic organic compounds on a broad range
of soils. Its use for estimating sorption coefficients K or Kf in soils with
either very low or very high organic carbon contents, however, may be prone
to introduce considerable errors (Hamaker and Thompson 1972).
Although the K value for a given compound is fairly constant (within a
factor of two) among different soils or sediments, the K values for differ-
ent compounds may vary more than several orders of magnitude. Helling and
Dragun (1981) note that molecular properties that affect KQC values include
1) structure and conformation, 2) acidic or basic dissociation constants,
3) aqueous solubility, 4) charge status, 5) polarity and polarizability, and
6) molecular size. These researchers further state that K values reflect
the relative affinity of the soil's organic carbon content for the compound
and the compound-water interactions. For nonionic compounds, aqueous solu-
bility and K are usually inversely correlated.
Few researchers have investigated the rate at which sorption equilibrium
is attained under nonequilibrium conditions. Under conditions of water flow
13
-------
(in laboratory soil columns or field soil profiles), the contact time may be
insufficient to achieve sorption equilibrium between soil-water or sediment-
water phases. From batch experiments, Rao and Davidson (1980) concluded that
about 60 to 80 percent of the sorption may be complete in less than a minute,
whereas the remainder may take one to a few hours. For some compound-soil
combinations, sorption may continue at a very slow rate for days or weeks.
Although most batch experiments indicate essentially instantaneous or very
rapid sorption, flow experiments with packed soil columns suggest that this
may not be valid.
When the sorption reactions are instantaneous, equilibrium exists between
the solution-phase and sorbed-phase concentrations and is specified by the
adsorption isotherm S=KC. Various rate laws have been proposed to apply when
sorption is not instantaneous, depending on how the processes responsible for
sorption nonequilibrium were conceptualized. Conclusions regarding the
sorption processes, however, are generally applicable to both transient and
steady water-flow conditions. For solution concentrations associated with
agricultural applications, linear sorption isotherms may be adequate. For
applications dealing with waste disposal on land, however, the assumption of
linear sorption isotherms may lead to serious underestimations of toxic
organic compounds leaching in the soil.
Models based on chemical nonequilibrium or physical nonequilibration are
mathematically identical, and the solution to these models would be the same
for given initial and boundary conditions. Nonequilibrium conditions are
expected only when water movement through the soil is greater than 90 cm/h, a
rate that is likely to occur only in sanas.
The K value is the best indicator of adsorption (affinity) across all
soils (except those very low or very high in organic matter). Accordingly,
the information presented in Table 3 indicates that 2,3,7,8-TCDD is very
strongly sorbed, whereas dichlorobenzene may be only weakly sorbed.
The addition of chlorine atoms to a base compound decreases water solu-
bility and increases K . For example, Table 4 shows the effects on log K__
oc oc
of increased chlorination of biphenyls and chlorobenzenes. An estimate of
KQC can be made for a highly chlorinated compound (e.g., 2,3,6,7-tetra-
chloronaphthalene) based on the K of the base compound (naphthalene).
Thus, the K for chloronaphthalenes is probably greater than 10,000 (log K
14
-------
TABLE 3. K VALUES FOR ENVIRONMENTAL ASSESSMENT OF
00 TOXIC ORGANIC SUBSTANCES3
Compound
TCDDb
PBBsc
p.p'-DDT
o.p'-DDT
p.p-DDD
p.p'-DDE
Toxaphene
PCBse
Hexachlorobenzene
Hexachlorocyclohexane
Tetrachlorodibenzofuran
Dichlorobenzenes
Tetrachloroazobenzene
Hexachl orobutadi ene
Other related chemicals
2,4,2' ,4'-Tetrachlorobiphenyl
Naphthalene
2-Chlorobiphenyl
1,2,4-Trichlorobenzene
Chlorobenzene
Koc
808,000
219,000
181,000
_d
80.000
55,000
53,000
5,000
3,900
1,000
-
170 to 251
-
-
32,500
6,000
1,700
501
223
log Koc
5.91
5.34
5.26
4.90
4.74
4.72
3.70
3.59
3.00
2.23 to 2.40
4.51
3.78
3.23
2.70
2.35
a Sabljic 1984, Briggs 1981, Karickhoff 1985, McCall et al. 1983.
2,3,7,8-Tetrachlorodlbenzodioxin.
c Polybromlnated biphenyls.
Information unavailable.
e Polychlorinated biphenyls.
15
-------
TABLE 4. COMPARISON OF CHLORINE CONTENT AND MEASURED LOG K FOR
SELECTED COMPOUNDS3 oc
Compound
2-Chlorobiphenyl
2,2' -Dichlorobi phenyl
2,4'-Dichlorobipheny1
2,4,2' ,4 '-Tetrachlorobi phenyl
2,5,2' ,5' -Tetrachlorobi phenyl
2,3,4,2' ,5 '-Pentachlorobi phenyl
2,4,5,2' ,5'-Pentachlorobiphenyl
2f3,4,2l,3',4l-Hexachlorobiphenyl
2,3,4,5,6,2' ,5 '-Heptachlorobi phenyl
Chlorobenzene
1 , 2-Di chl orobenzene
1 ,3-Di chl orobenzene
1 ,4-D1 chl orobenzene
1 ,2 ,4-Tri chl orobenzene
1,2,3, 5-Tetrachl orobenzene
Pentachl orobenzene
Hexachl orobenzene
1,2-Dichloroethane
1,2-Dibromoethane
Number of
chlorine
atoms
1
2
2
4
4
5
5
6
6
1
2
2
2
3
4
5
6
2
2 (bromine)
log KQC
3.23
3.68
3.90
4.51
4.67
4.50
4.63
5.05
5.95
2.10
2.26
2.23
2.40
2.70
3.20
3.50
3.59
1.28
1.56
Source, Sabljic 1984.
16
-------
>4.0). The higher chlorinated blphenyls are likely to have KQC levels great-
er than 50,000 (log K >4.5) because the log K of the 2,4,2',4'-isomer is
4.51 and the log K for the 2-chlorobiphenyl isomer is 3.23. Brominated
compounds have higher K values than related chlorinated compounds, as
illustrated by the halogenated ethanes shown in Table 4. Therefore,
brominated compounds have a greater affinity for soils or a greater tendency
to leave the aqueous phase.
In addition to K , several other measurements are used to model sorp-
tive behavior of organic compounds in soil. These include K , the octanol-
water partition coefficient; K., the soil- or sediment-water partition coeffi-
cient; US, water solubility; and BCF, the biocentration factor. Several
formulas have been derived empirically to relate K , K , US, and K^ with
compound movement within the soil profile and the potential for bioaccumula-
tion in aquatic organisms (Briggs 1981). These relationships exist because
all are related to hydrophobicity.
log KQC = 0.52 log KQW + 0.52 (Eq. 3-5)*
log WS = 0.01 - log KQw - (0.01 Tm - 0.25)
(Eq. 3-6)
[Tm = melting point, °C]
log (1/R- - 1) = log K + log OM - 1.33
f ow (Eq. 3-7)
[Rf - mobility factor]
log KQC = 0.74 - 0.55 log WS {£q> 3.8)
log BCF = 0.85 log KQw - 0.70 (Eq. 3-9)*
In general, the less water-soluble the compound is, the higher the K . K ,
Kd, and BCF and the lower the potential for the chemical to move into the
ground water phase. Only chemicals with a water solubility greater than 60
ppm are likely to reach ground water (Briggs 1981). Table 5 contains the
values of the partition coefficients for each of the chemicals selected for
investigation.
*Briggs 1981.
17
-------
The equations and observed data were used to assemble Table 5 for the
organic chemicals of Interest. Various observed and/or calculated values
were used to place each chemical In the appropriate position on the chart.
Location is not necessarily precise, as all values for each constant may not
fall into the same classification. The accuracy of the table is also limited
by the quality of the data used.
3.2 VOLATILIZATION
Determining whether the compound of interest is likely to volatilize
(diffuse) rapidly, leach downward, or move up and down in the soil profile in
response to precipitation and evapotranspiration, requires consideration of
the dominant fate or transport path of an organic chemical near the terrestrial-
atmospheric interface via diffusion to the air, along with consideration of
chemical and biological transformation. Simple diffusion transport models
for volatile compounds have been developed with the following important
assumptions (Bomberger et al. 1983):
1) No chemical transport by water movement occurs.
2) The diffusion coefficient is a constant.
3) Porosity correction is constant in both time and space.
4) Chemicals move between the three soil phases much more rapidly than
they diffuse into the air phase. This means that they appear to be
in equilibrium.
5) Sorption is reversible.
6) Soil stays wet.
The fourth and fifth assumptions are the most critical and are probably
violated under field conditions. Smith et al. (1978) found that at least
one-half hour (hours to days is more likely) was required to achieve adsorp-
tion equilibrium between a chemical in the soil water and on the soil solids.
Solution of the diffusion equation (Bomberger et ai. 1983) has shown that
many volatile compounds theoretically have diffusion half-lives of days
(e.g., toxaphene, 9 days), whereas those less volatile have diffusion half-
lives of months to years. Under actual field conditions, the time required
to achieve adsorption equilibrium will retard diffusion, and diffusion
half-lives in the soil will be longer than predicted. Numerous studies have
reported material bound irreversibly to soils, which also would cause
apparent diffusion half-lives in the field to be longer than predicted.
18
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TABLE 5. VARIOUS MEASURED AND ESTIMATED PARTITION COEFFICIENTS
AND PROPERTIES FOR ENVIRONMENTAL ASSESSMENT OF
SELECTED TOXIC ORGANIC SUBSTANCES
Compound
Hexabromobi phenyl
Hexachloroblphenyl
Tetrachlorodibenzo-
p-furan
Tetrachlorobi phenyl
Tetrach 1 orod i benzo-
p-dioxin
p.p'-DDT
o.p'-DDT
p.p'-DDD
p,p'-DDE
Hexachl orobenzene
Tetrachloroazo-
benzene
US
mg/ liter
0.02-0.006
0.06
-
0.017
0.002
0.02-0.085
0.085
0.02-0.10
0.01-0.12
0.006-0.32
Insol.
Tetrachlorobiphenylene 0.6
Hexachl orocyclo-
hexane
Tetrachloro-
naphthalene
Toxaphene
2-Chlorobi phenyl
Trlchl orobenzene
B1 phenyl
Naphthalene
Dichl orobenzene
0.15-17
-
0.4-7.0
6
19
-
31
100
Kd
500
150
-
-
-
-
-
-
-
6.0
-
50
12.4
-
-
15
-
-
-
14.8
log KQC
5.34
6.61
-
4.51
5.91
5.26
-
4.90
4.74
3.59
-
3.30
3.11-3.59
-
4.72
3.23
-
-
3.78
2.26
log KOW
7.00
6.34
6.95
6.44
6.15
3.98-6.19
-
5.06-5.99
5.43-5.69
5.57
4-5 est.
5 est.
3.80
-
3.30-5.43
4.72
3.97
-
2.38-3.36
-
Soil thin
layer
chroma to-
graphy,
RF
0.002
0.006
-
-
-
-
-
-
-
_
0.02
0.06
Upper limit for systemic activity in plants on KQW scale
Chi orobenzene
4
2.35
60
2.84
0.19
Upper limit for leaching of stable chemicals to ground water
Benzene
Dibenzofuran
1
1.70
600-1791
1
0.44
aSabljic 1984, Briggs 1981; Karickhoff 1985; McCall et al. 1983; Verschueren
1983; Stratten et al. 1979; Nash CP File 1985; Callahan et al. 1979.
19
-------
Another important underlying assumption is that the soil stays wet. It
is well established that organics bind much more strongly to dry soil than to
wet soil; thus, the K of a compound increases when the soil dries out.
Therefore, in a dry soil, the relationship defining the volume fraction of
the compound in air and water is too low and the estimated diffusion coef-
ficient is too large. Spencer and Cliath (1970) measured dramatic increases
in sorption of lindane (y-hexachlorocyclohexane) as soil moisture was de-
creased, and Ehlers et al. (1969) showed decreases in the effective soil
diffusion coefficient as the soil dried. Because the soil can dry out under
field conditions, predicted soil diffusion half-lives will almost always be
longer than those that actually occur.
Because the compounds under consideration in this report have low vapor
pressures, the following equation can be used to examine diffusion:
w w ^
soil w '
where Dw = compound's diffusion coefficient in free water
T = correction factor for the soil porosity
H = Henry's constant
a = constant relating the volume fraction of soil that is surrounded
by air or water, the fraction of organic carbon, Henry's
constant, and soil bulk density.
Diffusion coefficients in water are much smaller than diffusion coeffi-
cients in air. For example, for oxygen, Da (in air) = 1.75 x 10" cm /s, but
Dw (in water) = 2.1 x 10 cm /s. Consequently, the predicted soil diffusion
half-lives for volatile compounds range from hours to days, whereas the
predicted half-lives for nonvolatile compounds range from weeks to months.
This long diffusion half-life for nonvolatile compounds results in a viola-
tion of one of the derivation assumptions—that no transport occurs by water
movement. Over several weeks a significant fraction of soil water will
evaporate, and the movement of water through the pores becomes the principal
transport mechanism for dissolved nonvolatile organics. The water evapora-
tion rate determines the compound's mass transport rate, and overall volatil-
ization rates will be slow.
In comparing the leaching potential with the volatilization potential
for lindane and toxaphene, Bomberger et al. (1983) stated that lindane does
not volatilize by diffusion, but rather by mass transport as a result of
Bomberger et al. 1983.
20
-------
water evaporation at the soil surface. Toxaphene, on the other hand, is
strongly adsorbed and does not move in the soil profile. For toxaphene which
is in solution, however, volatilization is predicted to be rapid. The frac-
tion of toxaphene in air is greater than that of lindane in air (Table 6).
TABLE 6. DIFFUSION BEHAVIOR OF TOXIC ORGANIC COMPOUNDS IN A SOIL
WITH 10 PERCENT MOISTURE3
Toxaphene
Lindane
log KQC
5.32,,
3.11d
Hb
9'° -5e
5 x 10 be
>»c
0.998
0.994
»ac
1.4 x 10"^
1.2 x 10"b
Dsoil,
cm2/s
3.0 x 10"5
f
Half -life,
days
9
a Taken from Bomberger et al. 1983.
Henry's constant, estimate based on vapor pressure and solubility.
**
Fraction of compound in either soil(s) or air(a) phase.
Hamaker and Thompson 1972.
e Hamaker 1972.
Not shown. The material cannot diffuse significantly in soil air because
of low volatility.
Environmental conditions in and at the soil surface affect the volatility
of organic compounds. Drying out the soil surface can cause the solubility
of the diffusing organic to be exceeded and crystallization to occur. An
increase in soil temperature increases random movement of molecules. In
water or when sorbed to the soil surface, these molecules increase in energy
as the temperature rises. As their energy increases, the amount escaping
from the water or solid phase into the air in the soil also increases. In
addition, the activity of the molecules already in the soil's air increases,
and they escape into the atmosphere more rapidly.
Soil moisture affects volatility in that the pore volumes of the air and
water in the soil change. Changing soil moisture from 10 to 25 percent
increases soil half-lives by a factor of 6 for the volatility loss (Bomberger
et al. 1983).
The loss by volatility of the organic compounds studied in this report
will vary according to compound. Most loss by volatilization comes from the
fraction of chemicals in the equilibrated water phase. As indicated earlier,
water solubility for most of these chemicals is quite low; thus, little
chemical is available for loss by volatization. Significant amounts of DDT
and its congeners are lost through volatilization, however.
21
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Because the system Is dynamic and only a small amount of a compound may
be in the aqueous phase at any time, that mass in the aqueous phase that is
lost will be replenished from the sorbed material. Hence, total mass trans-
ferred over a long time period may be significant.
3.3 MICROS IAL DEGRADATION
In general, microbial degradation is not significant among the compounds
under consideration. The compound may degrade to an intermediate compound in
the process and then ultimately degrade to carbon dioxide (C02). As with
many of the environmental reactions, water solubility seems to influence,
either directly or indirectly, the extent of microbial degradation possible.
Lindane and dichlorobenzene, two of the compounds with the greatest
water solubility, lowest K , and lowest sorption to organic carbon, degrade
much more rapidly than compounds with extremely low water solubility, such as
DDT and TCDD.
Among the several factors affecting the potential of a compound to
biodegrade are the organic matter content of the soil, additions of organic
matter, moisture content, temperature, and degree of chlorination of the
molecule. Addition of organic matter to a soil may dramatically increase the
microbial population, and organic compounds may be co-metabolized as a result
of this growth. Degradation rates are slow, however, and they do not increase
with time (Alexander 1981).
Chemical and biological reaction rates are slow in cold soils. The rate
of biodegration also decreases in flooded soils because of lower soil tempera-
ture, slower decomposition of organic matter, less availability of nutrients,
and slower absorption and transport of water and nutrients into plants. Some
compounds, however, degrade under anaerobic conditions, and flooding enhances
the degradation rate of these compounds.
Table 7 demonstrates the effect of chlorination on biodegradation rates.
Monochlorobenzenes biodegrade rapidly, whereas other chlorobenzenes degrade
less quickly. Hexachlorobenzene is particularly resistant to degradation
(Freitag et al. 1974).
Microbial degradation of propanil results in the formation of 3,4-di-
chloroaniline. At high treatment rates, the dichloroaniline can form tetra-
chloroazobenzene (TCAB) or the oxygenated tetrachloroazoxybenzene (TCAOB) by
22
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TABLE 7. RELATIVE BIOCONCENTRATION, MICROBIAL DEGRADATION, AND .
PHOTODEGRADATION OF SELECTED ORGANIC COMPOUNDS IN ACTIVATED SLUDGE0
Compound
Benzene
Chlorobenzene
1,2-Dichlorobenzene
1,3-Dichlorobenzene
1,2,4-Trlchlorobenzene
Hexachlorobenzene
Biphenyl
2 ,2 ' -Dichl orobi phenyl
2, 4 '-Dichl orobi phenyl
2,5,4' -Tri chl orobi phenyl
2,4,6,2' -Tetrachl orobi phenyl
2 ,4 ,6 ,2 ' ,4 ' -Pentachl orobi phenyl
DDT
Hexachl orocycl ohexane
Hexachl orocycl opentadi ene
BCFb
1,700
1,700
14,300
560
1,400
35,000
2,600
6,300
9,800
32,000
6,500
27 ,800
14,000
820
2,400
Microbial deg-
radation, % co2
29
31.5
0.1
<0.1
0.1
<0.1
15.2
0.1
<0.1
<0.1
<0.1
0.3
<0.1
0.1
-
Photodeg-
radation,
% co2
15
18
2
5
10
2
10
4
3
6
4
5
-
50
Source: Frietag et al. 1974.
Bioconcentration factor.
0 C0? represents the ultimate breakdown product, by microbial degradation
or photodegration, of the toxic organic compounds in soil. Percent C02
evolved from the soil represents a relative measure of the extent of
degradation of the organic compounds. Stable intermediates may be present
when the parent is gone and no C02 has been evolved.
23
-------
bimolecular condensation. These compounds are similar in structure to TCDD
and are resistant to further microbial degradation (Bordeleau and Bartha
1972). Tetrachloroazobenzene also may be present in technical propanil as a
contaminant (Bunce et al. 1979). The PCB's, PBB's, and other organic compounds
with high K or K values are also quite resistant to microbial degradation;
2,2'-dichlorobiphenyl degrades so slowly that no metabolites can be detected
(Vockel and Korte 1974). Major degradation products of many chlorinated
organic compounds are hydroxy analogs. These metabolites may react with
existing organic matter to form bound residues. If the metabolite can never
be removed, it becomes a part of the organic pool and eventually oxidizes to
CCL in a manner analogous to all soil organic matter.
3.4 PHOTODECOMPOSITION
Many organic chemicals introduced into the environment absorb sunlight
and undergo transformations to new molecular species. Because ozone in the
atmosphere absorbs ultraviolet light less than 285 nm, only those molecules
that absorb energy above 285 nm are expected to undergo photochemical decom-
position in sunlight.
The principle behind photodegradation is that it occurs only when a
chemical (in its ground electronic state) absorbs a photon or an equivalent
energy packet that converts it into a more energetic electronic state having
a different electron distribution and thus a different chemistry. The ener-
gized molecule may lose energy by going completely or partially back to its
initial ground state. Even though the partially energized states are less
energetic, many photoprocesses occur from these partially energized states
because their longer lifetimes (milliseconds) afford a better opportunity for
bimolecular encounters.
Other chemicals can alter the course of photochemical reactions by
creating other activated species that may interact with or transfer energy to
the compound of interest. The reactions that occur depend on the physical
state of the compound, the solvent, and the presence of other reactants
(i.e., activators, oxygen, free radical donors, etc.). For example, 2,3,7,8-
tetrachlorodibenzodioxin content in soil decreased 70 percent after irradia-
y
tion with 2 mU/cm . The addition of xylene-ethyl oleate to the soil led to
complete TCDD decomposition after 9 days. The amendments reduced TCDD
24
-------
content by 50 percent In the subsoil (Liberti et al. 1978). Polychlorinated
dibenzodioxins and polychlorinated dibenzofurans undergo photoreductive
dechlorlnation with ultraviolet light irradiation in the presence of an
effective hydrogen donor.
Photodecomposition reactions depend greatly on latitude, time of year,
climate, and temperature, as these factors can affect the amount of energy
that reaches the soil surface to initiate photodecomposition reactions. In
soil, photochemical processes are generally not significant (Mill and Mabey
1985), especially if the compound is incorporated in the soil and none is on
the surface that is exposed to sunlight.
3.5 TRANSLOCATION TO PLANTS
Translocation into plants is a function of the amount of a compound
present in the soil solution. Because the compounds considered in this
document are highly insoluble, only a small amount is present in soil solution
to be absorbed by plant roots. Some solid/solid transfer is possible (i.e.,
soil organic matter directly to plant root). The amount found on the plant
root and ultimately in the plant is a direct function of the KQC and an
inverse function of soil organic matter content. Inasmuch as only small
amounts of these compounds have been found in plants, however, this process
is not considered significant.
25
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SECTION 4
PARTITIONING OF ORGANIC COMPOUNDS IN THE ENVIRONMENT
Having an understanding of sorption processes 1s an Important key In the
description of pollutant fate because sorption may significantly alter the
physical transport and chemical reactivity of pollutants. Hydrophobic inter-
actions are dominant in the sorption of uncharged organic chemicals to natural
sorbents. For composite particulates (i.e., sediments/soils), organic matter
is the primary sorbing constituent. Sorption partition coefficients, indexed
to organic carbon (K ), are relatively invariant for natural sorbents. The
K 's can be estimated from other physical properties of pollutants (aqueous
oc
solubility or octanol/water partition coefficients). Sorption tends to be
affected by hydrophilic contributions under one or both of the following
conditions: 1) high sorbate polarity, and 2) low organic carbon content of
the sorbent, especially with coincident high clay content. Although a priori
estimation techniques comparable to hydrophobic sorption are not currently
available, estimates of hydrophilic contributions relative to K can be
based on chemical class and sorbent composition. Although sorption to sediment
or soils is frequently viewed as a rapid process in environmental modeling,
true sorption equilibrium may require weeks or months to achieve, as pollutant
uptake and release kinetics are highly dependent on molecular size, sorbent
cohesive properties, and solids concentration.
4.1 PARTITIONING MODELS
*
Distribution of organic chemicals among environmental compartments can
be defined in terms of simple equilibrium expressions. Partition
coefficients between water and air, water and soil, and water and biota can
be combined to construct model environments that will provide a framework for
Air, water, soil, sediment, suspended sediment, biota.
26
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preliminary evaluation of expected environmental behavior. This approach Is
particularly useful when few data are available, as partition coefficients
can be estimated with reasonable accuracy from correlations between
properties. In addition to Identifying those environmental compartments In
which a chemical Is likely to reside (which can aid in directing future
research), these models can provide a base for more elaborate kinetic models.
Because environmental partition coefficients are largely a measure of a
chemical's tendency to partition between aqueous and organic media, correla-
tions can be made between various combinations of partition coefficients.
Water solubility has been related to n-octanol/water ratios, bioconcentration
factors, and soil sorption constants. The ratio of a compound's solubility
in n-octanol to its solubility in water (octanol/water partition coefficient,
K ) has been correlated with bioconcentration factors and soil sorption
constants. More recently, Kenega and Goring (1978) have given correlation
equations for all combinations of these parameters.
The following correlation equations can be used in the estimation of
partition coefficients (McCall et al. 1983):
In WS(ppm) = -1.7288 In KQC - 0.01(MP-25) + 15.1621 (Eq. 4-1)
In Koc = In KQW - 0.7301 (Eq. 4-2)
In BCF = 0.935 In KQW - 3.443 (Eq. 4-3)
The advantage of developing such correlations is that, once any of the
parameters is known, it becomes a simple process to estimate the others.
In its simplest form, a partitioning model evaluates the distribution of
a chemical between environmental compartments based on the thermodynamics of
the system. The chemical will interact with its environment and tend to
reach an equilibrium state among compartments.
A model is selected that predicts distribution patterns of chemicals in
a simulated environment representative of a segment of the world. The goal
is not to predict actual expected environmental concentrations, but to pre-
dict expected behavior by answering such questions as: To which phase is the
substance likely to migrate? Will a compound applied to or spilled on soil
leach or be volatile? Will a chemical accumulate in the biotic compartment?
and so on.
27
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A terrestrial model, a pond model, and an ecosystem model (which combines
the first two models) can be described in terms of equilibrium schemes and
compartmental parameters. The selection of a particular model will depend on
the questions asked regarding the chemical. For example, if the partitioning
behavior of a soil-applied pesticide were the item of interest, one would use
the terrestrial model. The pond model would be selected if questions concerned
aquatic partitioning, and the ecosystem model would be the choice if the
overall environmental distribution were to be considered. The selection
would also depend on whether the assessment is to be made on- or off-site.
The problem of pesticide runoff would be assessed offsite, whereas the ground
water impact would be assessed on site.
Partition coefficients can then be combined to describe the ecosystem,
assuming all the compartments are well mixed so that equilibrium is achieved
between them. This assumption is generally not true of an environmental
system because exchange rates between compartments may be slower than trans-
formation rates within compartments. Therefore, equilibrium is never really
approached, except perhaps with very stable compounds. Such simplifications,
however, indicate the compartments into which a chemical will tend to migrate
and can provide a mechanism for ranking and comparing chemicals.
Table 8 presents a comparison of various models. Based on these models,
if applied to soil, the three compounds shown would tend to remain in the
soil; however, if they were applied to or reached an aqueous system, lindane
would tend to stay in the aqueous phase, whereas the tetrachlorobiphenyl and
DOT would migrate to the soil. In the ecosystem model, the DDT's sink is
soil and sediment, whereas the tetrachlorobiphenyl and lindane tend to be
volatile.
Fugacity is frequently used to estimate the behavior of organic compounds
in model environments. Fugacity can be regarded as the "escaping tendency"
of a chemical from a phase, and it is expressed in units of pressure. Equi-
librium between phases is defined as occurring when their fugacities are
equal. Fugacity, f(Pa), can be related to concentration, C(mol/m ), by using
a fugacity capacity, Z (mol/m Pa), such that:
C = fZ (Eq. 4-4)
The Z values of a chemical can be estimated from knowledge of vapor pressure
and solubility or the octanol/water partition coefficient.
28
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TABLE 8. CHEMICAL DISTRIBUTION USING VARIOUS MODELS3
(percentage of chemical In each compartment)
Compound
Environmental compartment
Air
Water
Soil
Sediment
Suspended
sediment
Biota
Terrestrial Model
DDT
Tetrachlorobiphenyl
Lindane
0.000014
(0.015)c
0.014
(14.6)
0.0026
(2.68)
0.0067
(0.0007)
0.031
(0.003)
0.76
(0.080)
99.9
(2.11)
99.9
(2.10)
99.2
(2.09)
ro
\O
Pond Model
DDT
Tetrachl orobi pheny 1
Lindane
1.31
(0.000263)
5.77
(0.00115)
60.6
(0.0121)
98.6
(1.58)
93.8
(1.50)
39.4
(0.63)
0.099
(0.0000197)
0.094
(0.0000188)
0.039
(0.00000788)
0.081
(16.2)
0.42
(84.2)
0.020
(3.94)
Ecosystem Model
DDT
Tetrachlorobiphenyl
Lindane
4.72
(0.00943)
97.9
(0.196)
83.2
(0.166)
0.65
(0.00042)
0.065
(0.0000435)
7.48
(0.00499)
44.8
(1.86)
0.96
(0.023)
4.41
(0.11)
49.8
(2.65)
1.06
(0.057)
4.86
(0.26)
0.05
(0.0000332)
0.001
0.04
(27.3)
0.0048
(0.000000071) (3.18)
0.004
(0.0000032)
0.0024
(1.62)
a McCall et al. 1983.
DDT = dichlorodiphenyltrichloroethane.
c ( ) = concentration in compartment; air - yg/m3; others - ppm.
-------
Several levels of tugacity models have been used to calculate the distri-
bution between phases. Level I, the simplest, calculates the equilibrium
distribution (Patterson 1985). The Level II model also calculates the distri-
bution at equilibrium, but it is capable of including reaction or transforma-
tion and advection. Reaction includes photolysis, hydrolysis, biodegradation,
and oxidation. The Level III fugacity model depicts a steady-state nonequili-
brium system. This system, in which the fugacities in each phase are different,
is produced by including transport between the compartments. The Level IV
model is a dynamic version of Level III, in which emissions (and concentrations)
vary with time. Several other models also have been developed to simulate or
predict the behavior of toxic organic compounds in segments of the ecosphere.
The Inherent error 1n most models is that they assume instantaneous equilibrium
between components of the system.
Table 9 shows the distribution of several organic compounds when the
Fugacity I model was used. Compounds such as chlorobenzene and trichloroben-
zene, which have a high Henry's constant and hence a low Z for water,
partition mainly into air; however, the percentages in the soil and sediment
increase with increasing K. Chemicals with low H and low K remain primar-
ily in water, whereas those with low H and high K (e.g., DDT) partition
mainly into the soil and sediment.
TABLE 9. PARTITIONING OF COMPOUNDS INA UNIT WORLD WITH THE
FUGACITY I MODEL3
(percentage of chemical in each compartment)
Compound
DDT
2,4,2',4'-Tetra-
chlorobiphenyl
2-Dichlorobiphenyl
1,2,4-Trichloro-
benzene
Chlorobenzene
Air
0.275
5.10
69.6
95.9
99.1
Water
0.281
0.300
4.78
1.40
0.815
Soil
51.4
48.9
13.2
1.41
0.045
Sedi-
ment
48.0
45.6
12.4
1.32
0.42
Suspended
sediment
0.080
0.076
0.021
0.0022
0.00005
Biota
0.032
0.031
0.0083
0.00088
0.000028
Patterson 1985.
30
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4.2 ENVIRONMENTAL FATE MODELS
Numerous environmental fate models have been developed for specific
situations that are more complex than the simple partitioning models just
discussed. They generally treat situations where steady-state conditions are
not applicable (e.g., rivers, ponds, transport in soil or ground water). A
review of these models has been compiled (Dickson et al. 1983).
Two models, EXAMS and HydroQual, represent a complexity beyond the
models considered in the preceeding subsection and are thus of interest in
situations where the single partitioning models are not applicable.
4.2.1 EXAMS
The Exposure Analysis Modeling System (EXAMS), developed by a group at
the U.S. Environmental Protection Agency (EPA) laboratory in Athens, Georgia
(Burns et al. 1981), is one of the most successful and widely used models for
assessing the behavior of chemicals in aquatic environments. The model
system allows the user to specify volumes and compositions of water and
sediment; equilibria are expressed by conventional partition coefficients;
reactions such as photolysis, hydrolysis, and biolysis can be included; and
transfer rates between compartments are quantified. The model can be run to
calculate steady- or unsteady-state conditions. It can be applied to ponds
or rivers, and validation can thus be obtained by comparing actual and
calculated behavior.
4.2.2 HydroQual
HydroQual, Inc., in association with the Chemical Manufacturers Associa-
tion, has developed a modeling system to describe the behavior of chemicals
in lakes and rivers (Di Toro et al. 1982). The model is particularly strong
in its treatment of interactions between the water column and sediments.
Considerable effort has been devoted to expressing the key algebraic and
differential equations in a simple form.
4.2.3 U.S. EPA Models
McDowell-Boyer and Hetrick (1982) developed the TOX-SCREEN model for the
EPA as a multimedia screening-level model for assessing the potential for
human exposure to chemicals released in the environment. The intent of this
model is to identify chemicals that are most unlikely to present a problem.
It is simplified and conservative in nature and makes use of data that are
31
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typical of large areas of the United States rather than site-specific data.
It Includes intercompartmental transfer processes and equations for disper-
sion in air and water.
4.2.4 ENPART
The Environmental Partitioning Model (ENPART) developed by Wood et al.
(1982) is based on the fugacity equations of Mackay. This interactive or
batch model has the ability to receive a minimum set of input data for the
physical and chemical properties of a compound and to calculate other
properties required by using correlations. It also has the ability either to
receive input of degradation and transfer rates or to supply default values.
This model thus provides a first-level screening analysis of a chemical and
its partitioning in the environment. The interactive mode allows sensitivity
analyses of various parameters of a substance to be performed quickly and
easily.
4.3 EFFECT OF CLIMATE ON CHEMICAL FATE
Climatic factors that affect the fate of toxic organic compounds are not
simple; one factor influences another. Soil moisture and wind speed influence
soil temperature, light intensity influences soil temperature, and many
environmental influences affect soil organic matter. Table 10 shows the
primary dissipation routes of toxic organic compounds in each of the six
major climates of the United States.
TABLE 10. LIKELIHOOD OF ENVIRONMENTAL FATES OF TOXIC ORGANIC
COMPOUNDS IN SIX MAJOR CLIMATES
Fate
Soil sorption
Biodegradation
Photodecomposition
Volatilization
Leaching
Major climates3
1
Medium
High
High
High
Low
2
High
Medium
Medium
Medium
Low
3
Medium
Medium
Medium
High
Low
4
Low
Low
High
High
Low
5
High
Medium
Low
Low
Low
6
Medium
High
High
High
Low
1 - Subtropical (Southern Coastal Pacific and Southeast).
2 - Temperate (Midwest, Mid-Atlantic, New England, Northern Coastal Pacific)
3 - Dry (Great Plains, Intel-mountain West).
4 - Desert (Southwest).
5 - Boreal (Rocky Mountains, Alaska).
6 - Tropical (Hawaii, island territories).
Not significant in dissipation of hydrophobia toxic organic compounds.
32
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Without regard for any Interactions that may occur, the following rough
effects might be expected:
1) Temperature—The warmer the temperature is, the greater the volatil-
ity, the lower the organic matter content of the soil, the more
active the microbial population, and the higher the rate of evapo-
transpiration. The result is a decrease in pesticide persistence.
2) Moisture—There is an optimum level of soil moisture for microbial
activity. If a soil is too wet or too dry, activity slows down.
Volatility also is affected by moisture content; the nature of the
effect depends on the solubility of the chemical. The total
amount, intensity, and frequency of rainfall or irrigation water
received affect the movement of chemicals in soil (Bailey 1966).
3) Light—Photochemical reactions are directly proportional to the
number of photons absorbed by a chemical. Nearness to the equator
or an increase in altitude will speed the rates of photochemical
reactions.
4) Soil texture—Soil texture is a most important factor. Soil organic
matter is directly influenced by the soil texture. Coarse (i.e.,
sandy) soils will normally be low in organic matter; therefore,
water percolation will be rapid and the leaching potential of
chemical compounds will be high regardless of K or K values.
The opposite is true for heavy (i.e., clayey) soils.
4.4 LIKELY CHEMICAL FATES
The fate of any particular chemical is governed by an interactive combi-
nation of water-solubility, K , and Henry's Law Constant, H. Because
chemicals with high K or K values are strongly sorbed onto the soil, they
generally are not available for biodegradation, chemical decomposition,
photodecomposition, volatilization, or leaching. However, they may adsorb on
algae surfaces in aquatic systems where they are then subjected to photo-
decomposition. The following compounds fall into this category: hexabromo-
biphenyl; hexachlorobiphenyl; tetrachlorodibenzodioxin; p.p'-DDT; o,p'-DDT;
p.p'-DDE; p.p'-DDD; and polybrominated biphenyls.
At the other extreme of the chemical scale, substances with higher water
solubility and low K are not sorbed strongly in soil and thus are slightly
more available for biodegradation, photodecomposition, volatilization, and
leaching. Chlorobenzene, dichlorobenzene, and naphthalene could fall into
this category.
Toxaphene is apparently strongly adsorbed and does not move in the soil
profile. This strong sorptive interaction with soils, however, may cause
33
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some material to erode Into surface waters during irrigation or precipitation
events. There is no evidence to indicate the occurrence of oxidation, hydro-
lysis, or biodegradation; and photolysis is unlikely. Rapid volatilization
(a characteristic that would normally cause toxaphene to be considered
nonpersistent in the soil environment) is predicted, however. Nevertheless,
field studies show that toxaphene is persistent. The difference between
prediction ana reality probably results from the fact that toxaphene is not a
single compound, but a mixture of hundreds of component compounds. Although
the overall mixture is strongly sorbed and not mobile in the soil, some
individual components may be mobile and may leach deep into the soil, where
diffusion to the surface would be slow. In addition, although the mixture is
highly volatile, some components may be relatively nonvolatile and would
therefore persist in the soil.
Volatilization calculations predict that lindane that has entered the
soil profile does not volatilize by diffusion, but rather by mass transport
as a result of water evaporation at the soil surface. On the other hand, the
leaching calculations show that lindane is not highly mobile in the soil
profile. There is no evidence to indicate oxidation, and because the aquatic
hydrolysis half-lives are at least 6 months, hydrolysis rates in the soil
will probably be slow. Photolysis is not expected, but biodegradation does
occur. The results of studies of the biodegradation rates vary widely, and
reported half-lives range from weeks to months. If lindane were applied to
the surface in a climate where precipitation exceeded evapotranspiration, it
could leach deeply into the soil over time and could contaminate ground water
(if the biodegradation rate was slow). If evapotranspiration and precipita-
tion or irrigation were in balance, lindane could move up and down in the
soil profile with little net transport in either direction. The major loss
route would be by biodegradation to y-pentachlorocyclohexane (PCCH), followed
by volatilization of the PCCH. Actual field experiments have confirmed this
behavior (Cliath and Spencer 1971).
Tetrachlorobiphenyl was predicted to partition into air (Table 8).
Hexachlorobenzene, which is more water-soluble than lindane, might be expected
to move within the soil profile in an analogous manner.
34
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Given the interrelationships between the various aspects of behavior and
the good general agreement between individual calculated and measured values,
it is clear that chemicals having undesirable properties can be identified
without extensive testing.
35
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SECTION 5
INFORMATION ON SELECTED ORGANIC CHEMICALS
Preceding sections have presented general information about the chemical
properties and environmental fates of several toxic organic compounds identi-
fied as having high affinity for soils. They also have attempted in impart
knowledge to assist in an understanding of how certain properties can be used
to determine long-term residence in various types of soils in various locations.
This section provides detailed information on each of the compounds of concern.
The data on each chemical are organized in the following manner. Each
subsection begins with background information on the compound's structure,
sources, metabolites, applications, and toxicity. The subsection continues
with information on water-solubility, volatilization, biotic and abiotic
degradation, photodecomposition, partition coefficients, adsorption, and
translocation to plants. Each subsection then concludes with discussions of
experimental degradation measurements, detected levels in soil, and half-life
estimates.
5.1 HEXACHLOROBENZENE (HCB)
Structure and Physical Properties
Cl
The molecular weight of hexachlorobenzene is 284.79, and its melting point
and boiling point are 230°C and 322°C (sublimate), respectively (CRC 1984).
36
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Sources
Large quantities of hexachlorobenzene, a highly stable compound, are
formed as a result of the commercial production of chlorinated solvents and
pesticides, including perchloroethylene, carbon tetrachloride, atrazine,
propazine, simazine, and dacthal (Farmer et al. 1978). This compound is also
a technical impurity in quintozene (Hankawa 1978, Beck and Hansen 1974).
Metabolites
Hexachlorobenzene degrades to pentachlorophenol (Ballschmiter et al.
1977).
Applications
Hexachlorobenzene has commercial application as a fungicide and is used
as a seed dressing and wood preservative [World Health Organization (WHO)
1979]. It is also a byproduct of tetrachloroethylene production.
Toxicity
Hexachlorobenzene is moderately toxic by ingestion, and it is a suspected
carcinogen [U.S. Department of Health and Human Services (HHS) 1983, Sax
1984]. The LU5Q are presented below. The suggested allowable daily intake
for man is 1 g/kg body weight (Verschueren 1972).
.n Single dose
LU50 mq/kq
Mouse 4000
Rat 3500
Rabbit 2600
Cat 1700
Water Solubility
Hexachlorobenzene is nearly insoluble in water—6.2 ug/liter at 23.5°C.
Because of its low water solubility, HCB movement through the soil via mass
flow is negligible (Farmer et al. 1978, 1980). Any factor that increases the
water content of soil will decrease the amount of HCB flux through the soil.
37
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Volatilization
The vapor pressure of HCB Is 1.91 x 10 mm Hg at 25°C (Farmer et al.
1980). Small amounts may be able to move through the soil by vapor phase
diffusion and volatilization (Farmer et al. 1980).
Vapor-phase diffusion of HCB would occur through air-filled pores in the
soil and would depend on the porosity of the soil. Increasing the bulk
density of a soil will decrease the soil porosity and, consequently, HCB flux
through the soil (Farmer et al. 1980).
Results indicate that air-filled porosity has an exponential effect on
HCB diffusion through the soil. Farmer et al. (1978) showed that increasing
the air-filled porosity of soil by 13.4 percent increased the apparent HCB
diffusion coefficient by 45 percent.
Volatilization of HCB is a diffusion-controlled process. The rate of
removal of HCB by volatilization is determined by the speed with which mole-
cules diffuse upward through soil to the air. According to Farmer et al.
(1978), volatilization increased exponentially with rising temperature. Each
10°C rise in temperature caused HCB volatilization to multiply by a factor of
3.5.
Farmer et al. U978) used the following flux equation for steady-state
diffusion to calculate the length of time required for 54.9 percent HCB by
weight in soil to diffuse to the air-soil interface from a depth of 122 cm:
J = DS(C2-CS)/L (Eq. 5-1)
p
where J = Volatilization, or vapor flux, r.g/cm /day ~
D = Apparent steady state diffusion coefficient, cm /day
Cy - Air concentration at soil surface, ng/cm -
C* = Air concentration at bottom of soil layer, ng/cm
L = Depth of soil layer, cm.
According to this equation and assuming no degradation within the soil,
HCB would continue to volatilize from the soil at a maximum rate for several
million years (Farmer et al. 1978). Volatilization is obviously not a signifi-
cant pathway for HCB dissipation from the soil.
38
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Microbial Degradation
Hexachlorobenzene is resistant to microblal degradation (Farmer et al.
1978, Freitag et al. 1974). In a study by Haider (1980), HCB remained stable
under both aerobic and anaerobic conditions and showed only slight dechlorina-
tion of 3 to 4 percent when Incubated with Clostrldlum under anaerobic condi-
tions and Pseudomonas under aerobic conditions. Clostrldlum and Pseudomonas
are common, highly adaptive soil microbes.
Photodecompos1tion
No Information was available on photolysis of HCB.
Partition Coefficients
Log KQw's for HCB have been estimated at 5.50, 5.57, 6.22, and 4.99
(NASH CP File 1985). The log KQC is 3.59.
Adsorption
Evidence indicates that a very high direct correlation exists between
soil organic matter content and amounts of HCB adsorbed (Griffin and Chou
1980, 1981; Uallnoefer et al. 1975).
Translocation to Plants
Significant amounts of HCB were taken up by carrots, spinach, radishes,
young sugar beets, lettuce, and wheat (Freitag et al. 1974, Ogiso and Tanabe
1982, Wallnoefer et al, 1975, Scheunert et al. 1983b). German researchers
claim that barley and wheat can metabolize hexachlorobenzene by enzymatic
activity (Klein et al. 1983).
Experimental Degradation Measurements
Several investigators found no significant decreases in the concentra-
tion of HCB in soil samples over periods of time ranging from 6 to 20 months
(Hankawa 1978, Griffin and Chou 1980, Beck and Hansen 1974).
Detected Levels in Soil
Hexachlorobenzene has been detected throughout the United States at 0.19
mg/kg, 0.11 mg/kg, 0-900 yg/kg (wet weight basis), and 1677 wg/kg (dry weight
basis); in Belgium at 0.44 and 0.85 mg/kg; in Germany at 0.002-1.003 mg/kg;
and in Italy at 40 yg/kg (MO 1979).
39
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Half-Life Estimates
No direct estimates have been made of the half-life of HCB In soil;
however, based on Its strong adsorption to soil organic matter and relative
Insolubility In water, it Is expected to be strongly persistent, especially
In soils with a high percentage of organic matter. Under most environmental
conditions, It has a very low rate of degradation (HHS 1983).
5.2 1,2-DICHLOROBENZENE
Structure and Physical Properties
The molecular weight of 1,2-dichlorobenzene Is 147.01, and Its melting point
and boiling point are -17°C and 180.5°C, respectively (CRC 1984).
Sources
1,2-Dichlorobenzene (o-dichlorobenzene) Is produced as an Intermediate
in herbicide manufacture (WHO 1982).
Metabolites
o-Dichlorobenzene degrades to a chlorophenol (Ballschmiter et al. 1977,
Ballschmiter and Scholz 1980).
Applications
o-Dichlorobenzene has commercial applications as an insecticide fumigant
(WHO 1982).
Toxicity
o-Dichlorobenzene is moderately toxic by inhalation and ingestion. It
is the most toxic of all three dichlorobenzene isomers (Gosselin et al.
1984). In guinea pigs the LD50 is estimated to be 1.5 g/kg.
40
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Water Solubility
The water solubility of o-dichlorobenzene is 100 mg/liter at 20 C°.
Therefore, movement through soil in the water phase is likely to be
significant.
Volatilization
The vapor pressure of o-dichlorobenzene is a moderate 1.5 mm at 25°C
(WHO 1982). Therefore, vapor phase diffusion and volatilization are probably
significant pathways for the dissipation of this compound in soil.
Microbial Degradation
Evidence suggests that o-dichlorobenzene is susceptible to microbial
degradation. In one study, a mixed culture of soil bacteria was used to
transform chlorobenzenes into chlorophenols (Ballschmiter et al. 1977). In
another study, incubation of all three isomers of dichlorobenzene with soil
species of Pseudomonas yielded dichlorophenols and dichloropyrocatechols
(Ballschmiter and Scholz 1980). In a third study, dichlorobenzene was degraded
under both aerobic and anaerobic conditions.
Wilson and McNabb (1983) explain that at reasonably high concentrations
(e.g., greater than 100 pg/liter), microbial utilization of a pollutant can
provide an ecological advantage in that it increases the number of microbes
that metabolize the organic pollutant. At concentrations lower than about 10
ug/liter, use of the pollutant does not provide enough advantage to lead to
enrichment of the active organisms. At concentrations greater than 1,000 to
10,000 pg/liter, metabolism of the pollutant may be limited by availability
of oxygen and other metabolic requirements. Table 11 shows the likelihood of
biotransformation of dichlorobenzenes in water-table aquifers (Wilson and
McNabb 1983). This information was not available for pollutants in soils.
-------
TABLE 11. PROSPECT OF BIOTRANSFORMATION OF DICHLOROBENZENES IN
WATER-TABLE AQUIFERS a
Compound
o-Dichlorobenzene
p-Di chl orobenzene.
m-Di chl orobenzene
Concentration of pollutant
> 100 ug/liter
Probable
Probable
Improbable
< 10 ug/liter
Possible
Possible
Improbable
Anaerobic
water
None
None
None
a Wilson and McNabb (1983).
The researchers did not explain why the likelihood of biotransfor-
mation of m-dichlorobenzene differs from that of o- and p-dichloro-
benzene.
Photodecomposition
No information was available on photolysis of the dichlorobenzenes.
Partition Coefficients
D1chlorobenzene has an octanol/water partition log coefficient of 2.26
and a log KQC of 2.23 (NASH CP File 1985).
Adsorption
Although no research data were found, the direct correlation shown
between adsorption of hydrophobia organic compounds and soil organic matter
content probably applies to dichlorobenzenes as well.
Translocation to Plants
No information was available on the uptake of any of the dichloroben-
zenes by vegetation.
Experimental Degradation Measurements
No information was available on degradation measurements of the
dichlorobenzenes.
Detected Levels in Soil
No information was available on detected levels of any of the
dichlorobenzenes in soil.
42
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Half-Life Estimates
No direct estimates have been made for the half-life of o-dichloroben-
zene in soil; however, based on its relative adsorption to soil organic
matter and relative solubility in water, it is expected to be persistent,
though not as persistent as compounds with higher octanol/water and organic
carbon partition coefficients. Volatility, biodegradation, and leaching will
be major loss routes for o-dichlorobenzene in solution.
5.3 S.S'.M'-TETRACHLOROAZOBENZENE (TCAB)
Structure and Physical Properties
The molecular weight of TCAB is 320.02, and its melting point and boiling
point are 68°C and 293°C, respectively (CRC 1984).
Sources
Numerous researchers have shown that 3,3',4,4'-tetrachloroazobenzene is
a condensation product of 3,4-dichloroaniline (OCA). Burge (1972) and Bunce
et al. (1979) have indicated that 3,4-DCA transforms microbially to TCAB. It
further has been indicated that 3,4-DCA is itself a microbial breakdown
product of the herbicide propanil |.N-(3,4-dichlorophenyl)(propionamide)]
(Helling 1971a, hughes and Corke 1974). The electron-releasing substituents
of anilines are highly susceptible to biochemical transformations, but they
yield predominantly higher polymers rather than azobenzenes (Bordeleau and
Bartha 1972). Researchers have also detected 3,3',4,4'-TCAB in soil treated
43
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with the herbicide Swep [methyl-N-(3,4-dichlorophenyl)-carbamate] (Sprott and
Corke 1971).
According to Canadian researchers, azobenzene formation is of greatest
concern In sandy loam soils (Sprott and Corke 1971).
Metabolites
The probable decomposition of 3,3',4,4'-TCAB to organic radicals with
the liberation of nitrogen, coupled with its insolubility in water, is
suggestive of a chemical degradation process.
Evidence suggests that 3,3',4,4'-TCAB is a microbial breakdown product
of 3,4-dichloroaniline, but no data are available to show the biodegradation
of 3,3',4,4'-TCAB.
Applications
No information was available on applications of 3,3',4,4'-TCAB.
Toxicity
It has been confirmed that 3,3',4,4'-TCAB is a mutagen (Sax 1984).
Uater Solubility
Because 3,3',4,4'-TCAB is relatively insoluble in water, its movement
through the soil in the aqueous phase is negligible. Research data support
this claim (Sprott and Corke 1971). Helling (1971b) used soil thin-layer
chromatography studies to rate 3,3',4,4'-TCAB as completely immobile.
Volatilization
The vapor pressure of 3,3',4,4'-TCAB is moderately low (MHO 1975).
Therefore, the mechanisms of vapor phase diffusion and volatility will not
move large amounts of this compound through air-filled pores in the soil.
Photodecomposition
No information is available on photolysis of 3,3',4,4'-TCAB.
Partition Coefficients
No partition coefficients are available for 3,3'4,4'-TCAB. Because
:nzene has a log K of :
TCAB is likely to be 4 to 5.
azobenzene has a log K of 3.8 (Verschueren 1983), however, the log K for
44
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Adsorption
The retarded mobility of 3,3',4,4'-TCAB may reflect the preferential
adsorption of 3,3',4,4'-TCAB to active clay sites that become available
through decreases in soil organic matter (Helling 1971a). This hypothesis
has not been tested. It does partition readily into soil organic matter.
Translocation to Plants
No information is available on the uptake of 3,3',4,4'-TCAB by vegetation.
Experimental Degradation Measurements
In one study conducted in Canada, 3,3',4,4'-TCAB production was measured
as a percentage conversion of dichloroaniline (Sprott and Corke 1971).
Maximum concentrations of TCAB were detected at the end of 1 week; at the end
of 3 weeks, TCAB could no longer be detected. These data indicate that when
small amounts of TCAB are formed in Guelph loam, concentrations decrease
rapidly after peak levels are reached. From this observation, researchers
concluded that TCAB does not persist in Guelph loam. Table 12 shows the
maximum concentrations of TCAB produced in the four different soils.
TABLE 12. 3,3',4,4'-TCAB PRODUCED IN ONTARIO SOILS3
Guelph loam
Peel clay
Lincoln clay
Vasey sandy loam
3,4 DCA added
100 ug/g soil
100 ug/g soil
100 ug/g soil
100 ug/g soil
TCAB produced
0.075 ug/g soil
o.oo ug/g soil
0.024 ug/g soil
0.60 ug/g soil
Percentage
conversion
of DCA to TCAB
0.08
0.00
0.02
0.60
a Sprott and Corke (1971).
Detected Levels
Detected concentrations of 3,3',4,4'-TCAB ranged from 0.01 to 0.05 ppm
in 6 of 99 soil samples from rice-growing areas in Arkansas, California,
Louisiana, Mississippi, and Texas (Carey et al. 1980).
45
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Half-Life Estimates
The studies of Sprott and Corke (1971) suggest that TCAB does not
persist In certain kinds of soil when present in small concentrations. It
probably is dissipated by biodegradation. No other information is available
on the half-life of 3,3',4,4'-TCAB in soil.
5.4 HEXACHLOROCYCLOHEXANE (HCH)
Structure and Physical Properties
With regard to the persistence of HCH in soils, the gamma isomer has
been by far the most widely studied of all the HCH isomers. Thus, this
discussion is limited to f-hexachlorocyclohexane (also known as lindane).
The molecular weight of lindane is 290.83, and its melting point and
boiling point are 112°C and 323.4°C, respectively (CRC 1984).
Sources
Hexachlorocyclohexane is a manufactured pesticide and scabicide.
Applications
Lindane (Y-Hexachlorocyclohexane) is an insecticide.
Toxicity
Hexachlorocyclohexane is highly toxic by ingestion and moderately toxic
by inhalation or skin absorption. All the isomers of HCH are highly suspect
as carcinogens (HHS 1983). The LCgQ for 96-hour exposures of fish are >1.4
mg/liter.
46
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Water Solubility
The water solubility of HCH is 0.15-17 mg/liter (Guenzi et al. 1974;
Verschueren, 1983). It is more mobile than most other chlorinated hydro-
carbons. Because solubility rises with increasing temperature, HCH adsorp-
tion to soil decreases as it moves into the water phase.
Volatilization
The vapor pressure of HCH is measured at 9.4 x 10" mm Hg at 20°C
(Guenzi et al. 1974), and 1.3 x 10~4 mm Hg at 30°C (Cliath and Spencer 1971).
It is one of the more volatile organochlorine insecticides.
Vapor diffusion is a significant pathway of HCH loss from the soil.
Within an optimal range of soil moisture content (from about 4 to 20 per-
cent) (Guenzi et al. 1974), 68 percent of an initial amount of lindane
introduced could be lost in about 30 days (Jury et al. 1983).
Because small amounts of water can greatly increase the vapor density of
HCH in soil, the volatility of HCH is much greater in wet soils than in dry
soils. The displacement of the pesticide from clay surfaces increases the
vapor density (Spencer and Cliath 1970). Very dry soil surface layers
restrict HCH volatilization, which results in the creation of a relatively
high concentration of pesticide in the upper 2-mm surface layer of the soil
(Cliath and Spencer 1971).
The rate of volatilization decreases with time as the concentration of
HCH rapidly decreases to a concentration below that required to give a maximum
vapor density. Maximum volatilization rates obtained by Fanner et al. (1972:)
show losses of 202 kg/ha per year.
Microbial Degradation
Evidence indicates that Bacillus coli and Clostridium sporogenes degrade
HCH to benzene and monochlorobenzene, and that microbial activity is also
responsible for the metabolism of HCH to Y-pentachlorocyclohexene (PCCH)
(Guenzi et al. 1974, Cliath and Spencer 1971, Yule et al. 1967, Sethunathan
et al. 1969).
47
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14 14
Detection of the evolution of C(L from soils treated with C-labeled
lindane indicates that a ring cleavage mechanism is involved (Guenzi et al.
1974, MacRae et al. 1967).
One metabolite of lindane, PCCH, is probably more mobile than lindane,
and a relatively large proportion of lindane lost to the atmosphere may be
volatilized as PCCH.
Other evidence indicates that lindane stimulated the soil fungal popula-
tion and did not inhibit dehydrogenase activity, which is thought to reflect
the total range of oxidative activities of soil microflora (Tu 1981). Low
levels of HCH, however, could not be used as a carbon source in microbial
metabolism of soil organic matter (Suzuki et al. 1975).
Degradation of HCH by Flavobacterium sp., Agrobacterium sp., and Alcaligenes
sp. was reported by Castro and Yoshida (1974) and degradation by a mixed
anaerobic microbial culture was reported by Haider (1979, 1980). Other
researchers observed that HCH degradation rates were quicker in nonsterilized
soil than in sterilized soil and that degradation rates could be increased
further by adding organics to the soil (Castro and Yoshida 1974). Higher
degradation rates were noted in submerged unsterile soil (anaerobic) than in
moist aerated soil (aerobic) or sterilized moist soil (aerobic) (Kohnen et
al. 1975). Aerobic degradation occurred slowly, primarily by the formation
of chlorinated aromatic compounds; however, rapid dechlorination and degrada-
tion occurred in anaerobic soils enriched with anaerobic bacteria such as
Clostridium sp., Bacillus sp., and Enterobacteriacae (Haider 1979). Japanese
researchers observed growth of Clostridium rectum as HCH degraded (Ohisa and
Yamaguchi 1978). Canadian researchers also noted biodegradation of lindane
in an anaerobic sandy loam soil (Mathur and Saha 1975).
Degradation products include trichlorobenzenes, tetrachlorobenzenes,
benzene, monochlorobenzene, y-pentachlorocyclohexene, and trichlorophenol. A
very small amount (less than 1 percent) has been shown to isomerize to the a,
B, 6, and e forms of hexachlorocyclohexane (Waliszewski 1980).
Photodecomposi tion
The limited available information on photolysis of HCH indicates that it
is stable in light under atmospheric conditions (wHO 1979), but is oxidized
by ultraviolet light at 90°-95°C (Verschueren 1983).
48
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Partition Coefficients
The octanol/water partition coefficient, log Kow' of HCH is 3>8° and the
log organic carbon partition coefficient is 3.00 (NASH CP File 1985).
Adsorption
Lindane diffusion in Gila silt loam is strongly influenced by soil water
content, bulk density, and temperature. Ehlers et al. (1969) and Spencer and
Cliath (1970) found that decreasing bulk density and increasing temperature
caused a greater proportion of the lindane to be in the vapor phase. No
diffusion occurred when soil water content equaled zero percent, and the
optimal water content was approximately 4 percent (Scheunert et al. 1983a).
Adams and Li (1971) found that variability in lindane sorption is due
almost entirely to organic carbon. Evidence of lindane adsorption to organic
matter is shown by the fact that movement is more pronounced in sandy loam
than in silty clay (Guenzi and Beard 1967, Cliath and Spencer 1971). Desorp-
tion isotherms for lindane on Gila silt loam fit the Freundlich equation
(Spencer and Cliath 1970). Adsorption of lindane on three different soil
types increased with increasing concentration, following the Freundlich
isotherm (Chopra and Goel 1971). In agreement with calculated results,
when larger amounts of lindane were added, its concentration increased at
each point in the soil column, but these larger amounts had no influence on
the depth of maximum concentration or depth of penetration (Huggenberger
1972, Huggenberger et al. 1973).
Table 13 shows mobility factors developed by Fuhremann and Lichtenstein
(1980) for lindane and metabolites. Only one metabolite, 2,4,5-trichloro-
phenol, was less mobile than lindane in a developing system that used a
hydrophobic carrier.
49
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TABLE 13. MOBILITY FACTORS FOR LINDANE METABOLITES3
Compound
1,2,4,5-Tetrachlorobenzene
1,3,5-Tri chlorobenzene
1,2,3,4-Tetrachlorobenzene
1,2,4-Tri chlorobenzene
1,2,3-Tri chlorobenzene
Lindane
2,4,5-Trichlorophenol
Mobility factor, Rf
0.92
0.92
0.82
0.82
0.74
0.22
0.10
a Fuhremann and Lichtenstein (1980).
Adams and Li (1971) developed an equation to predict percent sorption of
lindane in soil:
Percent sorption = 0.057+ 0.1607x - 0.0077x2 (Eq.5-2)
where x = organic carbon content of soil
x_ = square of the organic carbon content
These researchers showed that desorption of lindane is complete and that it
is independent of all soil variables measured and the amount of lindane
sorbed. Organic matter and moisture-holding capacity had the greatest influ-
ence on the sorption of organochlorine pesticides, including lindane. The
findings of Wahid and Sethunathan (1979) support this conclusion. The influence
of soil water is attributed to the inability of less polar compounds to com-
pete with water for sorption sites. Adams and Li (1971) also showed that
coarse-textured soils generally had the least affinity for lindane.
El Beit et al. (1981) suggested three possible operative factors in
their observation that lindane adsorption decreased with increasing tempera-
ture:
1) Increasing the temperature increases the rate of water movement
through soil and causes increased losses by leaching.
50
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2) Increasing the temperature Increases the rate of evaporation and
degradation caused by decreased adsorption.
3) Adsorption Is an exothermic reaction evolving heat. As temperature
Increases, the degree of adsorption decreases, particularly If
adsorption mechanisms are physical.
Lindane adsorption may be affected by physical forces or hydrophobic
bonds, or both. Decreased adsorption resulting from increased temperature
and contact time and the rapid occurrence of adsorption after lindane is
applied to soil support the physical theory. The hypothesis that lindane
adsorption occurs through weak inter-molecular forces at lipophilic sites on
soil particles would support the hydrophilic theory. High unaccountable
losses noted at high temperatures suggest that processes such as degradation
and volatilization may be more operative at high temperatures and may cause
decreased adsorption (El Beit et al. 1981).
Lindane is particularly prone to removal by leaching and runoff. The
total proportion adsorbed at different concentrations was low compared with
other organochlorine compounds investigated by El Beit et al. (1981). Lindane
is moderately mobile, and these researchers suggested that physical leaching
is probably more pronounced in light mineral soils than in heavy clay or
organic soils. Brazilian investigators showed that lindane was more persistent
in high-organic soils than in low-organic soils (Lord et al. 1978) and that
it degraded more quickly in soil poor in organic matter than in organic-rich
soil (Flores-Ruegg et al. 1980).
Translocation to Plants
Lindane has been detected in oat roots and tops and in potatoes (at
0.001 to 0.020 mg/kg), chickpeas, peas, carrots, corn, wheat, and soybeans
(Fuhremann and Lichtenstein 1980, Uhnak et al. 1979, Kathpal et al. 1984,
Lichtenstein 1975, Oloffs et al. 1971, Nash and Harris 1973).
Experimental Degradation Measurements
Numerous researchers have measured the degradation of HCH in a variety
of soils and under various circumstances. One year after treatment, 70 and
82 percent of the original amounts of lindane applied to the soil were
recovered from mixed and undisturbed soils, respectively (Guenzi et al.
1974). Of 117.3 kg/ha added to the soil between 1950 and 1953, 7.7 percent
51
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remained In 1968 (Chisholm and McPhee 1972). Of lindane applied at 10 or 100
pounds per acre, 0.2 percent was recovered after 15 years. Eleven years
after application, 94.7 and 99.5 percent of the original dosages had been
lost. By the end of a single growing season, 57 and 45 percent of the
applied lindane were lost from soils treated with 10 and 100 pounds per acre,
respectively (Lichtenstein et al. 1971).
One month after spraying at 11.2 kg/ha, lindane residues were 31 ppm in
the soil and 585 ppm in the surface litter. Lindane in the surface litter
moved slowly into the soil, and the residue level in soil reached a high of
58 ppm 1.5 years after application. Five years after application, residues
were 32 and 27 ppm in the soil and litter, respectively (Jackson et al.
1974).
Other investigators showed an average residue of 10 percent 14 years
after two applications of 56 and 224 kg/ha; an average residue of 41 percent
11 years after applications of 0.3, 2.8, 5.6 and 11 kg/ha; and an average
residue of 0.2 percent 15 years after applications of 11 and 112 kg/ha
(Jackson et al. 1974).
Under favorable moisture conditions, all HCh added to soil at 700 pg/kg
was completely degraded within 56 days (Zonglian et al., undated). After the
use of lindane was prohibited in Japan, total HCH residue in one location
decreased from 7.9 mg/kg in 1970 to 0.4 pg/kg in 1973. In another location,
levels decreased from 0.06-1.12 mg/kg in 1970 to 0.013-0.848 mg/kg in 1972
(WHO 1979).
Table 14 shows the results of an 8-week-long radioactive lindane tracer
degradation study. The least lindane was lost in moist organic soil, and the
most was lost in submerged mineral soil.
TABLE 14. LINDANE RECOVERY AFTER EIGHT WEEKS3
Soil type
Moist organic soil
Submerged organic soil
Moist mineral soil
Submerged mineral soil
Percentage recovered
89
86
84
70
WHO (1979).
52
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Soil application of 10 kg/ha lindane resulted in losses of 99.4 percent
in 180 days (Miles et al. 1978). Fifteen years after the last application of
lindane to a sandy loam soil, 7.5 percent remained of the total amount applied
(Stewart and Fox 1971). In 1972, seven air-dried soil samples contained 21.0
to 75.5 percent of the amount of lindane found in 1970 (Stewart and Chisholm
1971).
Lindane losses of 28.6 to 33.4 percent and 41.7 to 45.4 percent were
noted in normal and alkaline soils, respectively, at the end of a 9-month
period (Chawla and Chopra 1967). In India, initial residues of 2.89 ppm
dissipated to 0.03 ppm in 100 days in chickpea soils; dissipation was "almost
complete" in 148 days (Kathpal et al. 1984). Lindane loss was 53 and 88
percent at 1 and 5 months, respectively, after application (Pal and Kushwaha
1977). Indian researchers also measured lindane loss of 99.4 percent after
180 days (Agnihotri et al. 1977).
German researchers found 62 percent of the applied lindane remained
after 105 days' incubation in moist aerated soil; 58 percent remained after
77 days in autoclaved submerged soil; and 76 percent remained after 105 days
in autoclaved aerated soil (Kohnen et al. 1975).
In Nova Scotia, Stewart and Fox (1971) detected no lindane 12 years
after its application in a sandy loam soil with 5.5 percent organic matter.
Suzuki et al. (1975) observed rapid lindane decreases from spring to summer
in flooded, biologically active conditions; high temperatures; and high
humidity.
Detected Levels
In South Bohemia, lindane has been detected in soil at 1 ug/kg (WHO
1979), and in the United States, 13 out of 41 soil samples showed levels
ranging from 0.001 to 0.35 ppm and averaging 0.01 ppm. (MHO 1974). Lindane
was found in 8 of 50 agricultural soils tested in Colorado (Mullins et al.
1971). In Spain, levels of O.OC3 to 0.0024 mg/kg were found (Simal et al.
1977); and in Taiwan, a level of 34.2 ppb was detected (Wang et al. 1981).
Half-Life Estimates
Because of its moderate water solubility, vapor pressure, and degree of
adsorption to soil organic matter, lindane is not likely to persist in soils
as long as TCOO, DOT, and hexachlorobenzene. Experimental results support
53
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estimates that lindane is moderately persistent, especially under favorable
moisture conditions. Jury et al. (1963) estimate a half-life of 260 days;
Guenzi et al. (1974) estimate a half-life of 2 years. Volatilization and
degradation are probably the primary pathways of loss.
5.5 2,3,7,8-TETRACHLORODIBENZOFURAN (TCDF)
Structure and Physical Properties
There are 135 possible chlorinated isomers of dibenzofuran. No informa-
tion is available on their behavior in soils. Because they are structurally
and physicochemically similar to polychlorinated dibenzodioxins, however, one
can draw tentative conclusions about dibenzofurans by extrapolating known
information about dioxins.
The molecular weight of TCDF is 305.98. Although its melting and
boiling points have not been measured directly, its melting point has been
estimated by different methods at 107°, 139C, and 248°C; and its boiling
point, at 346° and 510°C (EPA 1985).
Sources
Chlorinated dibenzofurans are a pyrolysis breakdown product of PCB's and
are often found as unintentional impurities in products derived from
chlorophenols and chlorobenzenes (EPA 1982, 1983; WHO 1978).
Applications
Polychlorinated dibenzofurans have no commercial applications.
Toxicity
The three isomers considered to be the most toxic are
2,3,7,8-tetrachlorodibenzofuran; 2,3,4,7,8-pentachlorodibenzofuran; and
54
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1,2,3,7,8-pentachlorodibenzofuran. Only 10 to 12 of the 135 possible poly-
chlorinated dibenzofurans are expected to have significant toxicity (EPA
1982).
Furan itself has a toxicity rating of 6 out of a possible 6, which puts
it into the class of the most severely toxic compounds (Gosselin et al.
1984). The toxicity of 2,3,7,8-tetrachlorodibenzofuran (TCDF) is believed to
be one-fiftieth that of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD).
Water Solubility
The water solubility of TCDF is estimated to be 0.2 yg/liter at 25°C
(EPA 1983). Dibenzofuran itself is slightly water-soluble at 3 mg/liter at
25°C, but increasing halogenation promotes lipid solubility and diminishes
water solubility (EPA 1982). Although both TCDD and TCDF are comparatively
insoluble in water, TCDF is slightly more soluble and it may be more readily
leached from soils than TCDD is (EPA 1983).
Volatilization
No vapor pressures of chlorinated dibenzofurans have been measured
directly. The vapor pressure of TCDF is estimated to be 2.0 x 10" mm Hg at
25°C (EPA 1982, 1983). with such a low vapor pressure, volatilization is not
likely to be a major dissipation route from the soil.
Microbial Degradation-Metabolites
No direct evidence exists concerning microbial degradation of chlori-
nated dibenzofurans, but only 5 of the 100 strains tested showed any ability
to metabolize TCDD (EPA 1982). Polychlorinated dibenzofurans can undergo
photoreductive dechlorination in the presence of ultraviolet light and an
effective hydrogen donor (EPA 1985).
Photodecomposition
Higher chlorinated dibenzofurans can be degraded by ultraviolet light.
Polychlorinated dibenzofurans undergo photoreductive dechlorination in the
presence of ultraviolet light and an effective hydrogen donor (EPA 1985).
Partition Coefficients
The log octanol/water partition coefficient of TCDF is estimated to be
6.95 (EPA 1985).
55
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Adsorption
No direct evidence was found regarding TCDF adsorption to soils. Even
though TCDF is slightly more soluble than TCDD, it will not leach from soil.
The more chlorinated a dibenzofuran, the less polar it becomes; and
the less polar it is, the stronger are the forces of adsorption to soil
organic matter. Inasmuch as TCDD is strongly adsorbed to soil organic matter
(Young 1980), it is logical to assume that TCDF also adsorbs strongly to soil
organic matter.
Translocation to Plants
No information is available on the uptake of chlorinated dibenzofurans
by vegetation.
Experimental Degradation Measurements
No experimental results are available on degradation of polychlorinated
dibenzofurans in soil.
Detected Levels
No information is available on levels of chlorinated dibenzofurans
detected in soils.
Half-Life Estimates
Based on the documented persistence of TCDD in soil, coupled with the
structural similarity, low water solubility, low volatility, and high
capacity for adsorption to organic matter of both chlorinated dioxins and
chlorinated dibenzofurans, one can conclude that a strong possibility exists
that research will eventually show that chlorinated dibenzofurans are
strongly persistent in soils.
5.6 2,3,7,8-TETRACHLORODIBENZO-P-DIOXIN (TCDD)
Structure and Physical Properties
Cl
56
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The molecular weight of TCOD is 321.97. Its melting point has been
measured directly at 305° to 306°C and has been estimated by different
methods to be 305°C (Verschueren 1983) and 263°C (EPA 1985). Its boiling
point has been estimated at 348° and 533°C (EPA 1985) by differing methods.
Sources
Found primarily as a contaminant in products derived from trichloro-
phenols (including 2,4-D and 2,4,5-T), TCDO is also formed as a byproduct of
low-temperature incineration of wastes containing chlorinated precursors
(Young 1980).
Applications
This compound has no commercial applications.
Toxicity
Known to be a carcinogen (MHO 1977), TCDD also is extremely toxic. Its
toxicity rating of 6 out of a possible 6 indicates that it is in the class of
most severely toxic compounds (Gosselin et al. 1984). It is estimated to be
50 times more toxic than tetrachlorodibenzofuran (ETA 1985).
The acute LDgg is estimated at:
Guinea pig 0.6 ug/kg
Male rat 22 ug/kg
Female rat 45 ug/kg
Monkey <70 ug/kg
Rabbit 115 ug/kg
Water Solubility
The water solubility of TCDD is estimated to range from 0.0002 ug/g at
an unspecified temperature to O.OG6 ug/g at 30°C (NASH CP File 1985).
Helling (1971) reports the water solubility of TCDO to be 0.2 ppb at 25°C.
Because of its low water solubility, movement of TCDD through the soil via
mass flow is expected to be negligible (Young et al. 1976). According to the
results of a long-term study in a natural ecosystem, no TCDD movement was
detected despite an annual rainfall of 150 cm (Young 1980).
Volatilization
Researchers who conducted a long-term study in a natural ecosystem
estimated that approximately 10 percent of TCDD volatilizes from the soil and
57
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an estimated 5 percent is lost by volatilization each year (Young 1980). The
estimated vapor pressure of TCOD is listed as 0.2 mPa (NASH CP File 1985).
Microbial Degradation - Metabolites
Although stable in the presence of heat, acids, and alkalis, TCDD
undergoes photoreductive dechlorination in the presence of ultraviolet light
and an effective hydrogen donor. Loss of chlorine atoms from the
polychlorinated dibenzo-p-dioxins occurs preferentially from lateral
positions on the carbon ring (EPA 1985). Two breakdown products of TCDD are
2,3,7-trichlorodibenzo-p-dioxin and dichlorodibenzo-p-dioxin (isomer unspeci-
fied) (MHO 1977).
Researchers disagree about the biodegradability of TCDD. One group has
hypothesized that strong adsorption of TCDD to soil particles limits its
bioavailability. They have concluded that situations in which herbicide
residues are negligible, TCDD residues are in the form of a precipitate, and
the overall soil organic matter is less than or equal to 1 percent are not
conducive to TCDD biodegradation (Young 1980). One study that supports this
view showed that only 5 of 100 soil microbe strains tested showed any ability
to metabolize TCDD (MHO 1977); another showed that all efforts to biodegrade
TCDD in soil were unsuccessful (Pocchiari 1978). Meanwhile, other researchers
have concluded that TCDD is degradable by soil microbes, especially
in the presence of other chlorinated hydrocarbons (Young et al. 1976). In
two studies, application of 5,000 to 40,000 mg/kg of the herbicides 2,4-D and
2,4,5-T, which contained unknown amounts of TCDD, stimulated the growth of
species of actinomycetes and fungi. It seems probable that the herbicides
and TCDD were the carbon sources (MHO 1977).
Philipp et al. (1981) observed that in some experimental microbial
cultures approximately 1 percent of TCDD transformed into a polar metabolite
that may have been a hydroxylated derivative.
Massive concentrations of 2,4-D and 2,4,5-T did not sterilize soils,
which indicates that microbial degradation of TCDD, a contaminant of these
herbicides, remained a possibility (Young et al. 1983).
Photodecomposition
Some researchers suggest that TCDD can be broken down by ultraviolet
light. Published data show a 70 percent decrease in soil TCDD content after
53
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irradiation at an intensity of 2mW/cm2 (Young et al. 1976, Liberti et al.
1978). Other results show photodecomposition of TCDD only when the TCDD is
dissolved in an organic solvent, not in soil or water (Liberti et al. 1978,
MHO 1977). Contaminated soils sprayed with xylene-ethyl oleate and exposed
to solar radiation showed approximately 100 percent destruction of TCDD on
the surface layer in 9 days and 40 to 60 percent destruction in the subsoil
(Liberti et al. 1978). These investigators concluded that no increase in the
decay rate of TCDD was observed when naturally occurring soil was exposed to
sunlight.
Partition Coefficients
The logio octanol/water partition coefficient for TCDD has been
calculated at 6.15, and the estimated organic carbon partition coefficient is
5.91 (NASH CP File, 1985).
Adsorption
Researchers believe TCDD is stable and immobile in soil and that it has
a high capacity for adsorption to organic matter (Helling 1971, Briggs 1981).
Soil thin-layer chromatography (TLC) results support this hypothesis (Helling
1971, Briggs 1981). The adsorption of TCDD to organic matter significantly
increases its persistence in the soil (Young 1980). Philipp et al. (1981)
demonstrated that a decrease 1n TCDD extractability over a long-term period
of incubation with high microbial activity was a result of adsorption of the
TCDD to soil material.
Distributed mostly in the upper 15 cm of soils, TCDD concentration
decreases sharply with increasing depth (Di Domenico et al. 1980). The
highest levels have been found between 0.5 and 1.5 cm and at not less than
0.5 cm (Di Domenico et al. 1980).
Little downward leaching of TCDD occurs, but lateral movement may occur
by surface erosion (WHO 1977).
Translocation to Plants
Indications are that trans!ocation of TCDD is either minimal or
nonexistent. A maximum of 0.15 percent of TCDD in soil was translocated to
aerial parts of oats and soybeans (WHO 1977).
59
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Experimental Degradation Measurements
Various measurements have been made of TCDD degradation In soils. Young
(1980) found 10 to 1500 parts per trillion (ppt) In the top 15 cm of a soil
14 years after the last herbicide application at a site where 2.6 kg of TCDO
had been applied to 37 ha between June 1962 and July 1964. Analysis of 61
soil samples showed that less than 1 percent of the original amount of TCDD
remained on the test area.
Young et al. (1983) measured TCDD concentrations at 0.5 ppm (mg/kg),
which tended to decrease over 3 years; however, no accurate estimate of
persistence was possible.
Seventy-one percent of 100 mg/kg of TCDD in silty clay loam was recovered
after 1 year; 56 percent was recovered from sand after the same period (WHO
1977). In another study, 1 percent of an initial amount of TCDD deposited at
a depth ranging from 0 to 15 cm was detected at a depth below 30 cm after 414
to 557 days.
Detected Levels
In soil samples of varying depths, TCDD was detected at concentrations
ranging from less than 10 to 1500 ppt. The median concentration was 30 ppt
and the mean was 165 ppt (Young et al. 1981). No TCDD was detected in 3-foot
core samples from an area that received 2,4,5-T annually for 7 years (limit
of detection equal to 1 ppb) (Woolson et al. 1973).
In Vietnam, TCDD concentrations equal to 1.2 to 2.3 yg/kg were detected
in soils sprayed aerially with 2,4,5-T. Of an estimated total of 100 kg of
TCDD sprayed in Vietnam during the years 1966 through 1969, the average
concentration in the top centimeters of soil averaged 40 ng/kg (WHO 1977).
half-life Estimates
In a soil with a high herbicide concentration that provided microorganisms
with a readily available carbon source, the half-life of TCDD was measured at
less than 200 days (Young 1980). In a dry climate (Utah), the measured
half-life was 330 days; whereas in a warm and humid area (Florida), the
half-life was approximately 190 days.
60
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One month after an accidental release of TCDD In Seveso, Italy, the
half-life was calculated to be approximately 1 year. Sixteen months later
the half-life was recalculated to be more than 10 years. The TCDD concentra-
tion dropped to about half its original concentration in the first 5 months,
after which no further decreases were detected (Di Domenico 1980). The 50
percent degradation of nanogram amounts of TCDD in 5 hours indicated that low
levels of TCDD may degrade more quickly than high levels (WHO 1977).
In Missouri, TCDD concentrations of 0.12 to 0.85 ng/kg after 3 years
indicated a half-life of 6 months. Estimates of TCDD half-life in soil fall
consistently in the range of 0.24 to 1 year (WHO 1977).
These varying results indicate disagreement regarding the actual half-life
of TCDD in soil. Perhaps the differences result from highly varying environ-
mental conditions among experiments. The high K of TCDD, however, suggests
that it adsorbs strongly to soil organic matter and that it is extremely
persistent.
5.7 POLYCHLORIMATED BIPHENYLS (PCB's)
Structure and Physical Properties
A total of 209 possible PCB isomers exist (WHO 1978). Isomers are
normally found in mixtures; therefore, PCB's are a mixture of many isomers,
all with different melting and boiling points.
Source
Polychlorinated biphenyls are manufactured. They are often found as a
blend with trichlorobenzenes.
61
-------
Metabolites
Although PCB's are virtually chemically inert, they form chlorinated
dibenzofurans under extremely high temperatures (WHO 1978).
Because of their thermal stability, PCB's were once used as insulating
fluid in electrical equipment (such as transformers and capacitors).
Toxicity
Polychlorinated biphenyls are highly suspect as carcinogens (HHS 1983,
Sax 1984), and they are severely toxic to animals and humans (Brinkman et al.
1981).
Mater Solubility
Most PCB's are virtually insoluble in water. Water solubility ranges
from 0.04 to 0.2 mg/liter for the more common isomers (Table 5) and decreases
with the addition of chlorines.
Volatilization
No information is available on the vapor pressure of PCB's. Vapor
pressure values of 1 mm Hg at 89.3°C and 1 mm Hg at 96.4°C were found for two
monochlorinated biphenyls: 2-chlorobiphenyl and 4-chlorobiphenyl, respectively
(Carey et al. 1980).
Microbial Degradation
Polychlorinated biphenyls are highly resistant to biodegradation, and
the higher the chlorine content of the PCB mixture, the greater the
resistance (Kirk-Othmer 1984). When various microorganisms (including soil
microbes, Azotobacter, Pseudomonas, lactic acid bacteria, yeast, fungi, and
Neurospora) were incubated with 2,2'-dichlorobiphenyl, no metabolites were
detected (Vockel et al. 1974). Microbial metabolism of certain dichloro-
substituted rings of PCB's by biphenyl-oxidizing bacteria has been noted,
however. These bacteria require an adjacent 2-3 position to form the first
biphenyl metabolite, cis-2,3-dihydro-2,3-dihydroxybiphenyl (Jacobs et al.
1976).
Photodecomposi ti on
No information is available on photodecomposition of PCB's.
. 62
-------
Partition Coefficients
Log1Q K values of 5.99, 6.34, and 6.73 have been calculated for
2,2',4,41,5,51-hexachlorobiphenyl (NASH CP File 1985). Briggs (1980)
obtained log K values of 6.30 and 6.84 (experimentally and by calculation,
respectively) for pentachlorobiphenyl.
Adsorption
Polychlorinated biphenyls adsorb strongly to soils (Brinkman et al.
1981). A very high direct correlation exists between soil organic matter
content and the amount of RGB's adsorbed (Griffin and Chou 1981).
Translocation to Plants
Research data show that PCB's are taken up by grass and beets, carrots,
and sugar beets (WHO 1978, Monza et al. 1976).
Experimental Degradation Measurements
According to some data, PCB's are very environmentally persistent.
Norwegian researchers observed no change in PCB levels in orchard soils over
a 2-year period (Kveseth et al. 1977). According to other data, however,
some of the less heavily chlorinated biphenyls may be subject to loss by
14
volatilization. German researchers applied C-labeled 2,2'-dichlorobiphenyl
to a soil in which they grew carrots during one growing season and sugar
beets during the following year's growing season. In the first year, 53.5
percent of the applied radioactivity was lost by volatilization; in the
second year, an additional 25.2 percent was lost (Monza et al. 1976). In
other work, Scheunert et al. (1983a) showed that the percentage of conversion
products for 2,5,4'-trichlorobiphenyl and 2,2',4,4',6-pentachlorobiphenyl
detected after one week was nearly the same as that detected after an entire
growing period. From these data, Scheunert et al. (1983a) concluded that the
conversion rate of PCB's is slow and time-dependent.
Detected Levels
Various soil concentrations of PCB's have been reported. In Japan, they
have been detected at levels ranging from 0.01 yg/g to greater than 100 wg/g.
In Finland, levels were measured at 0.1 yg/g (Rautapaa et al. 1976). Near a
PCB manufacturing plant in Illinois, levels ranging from 130 to 20,700 ug/kg
63
-------
were measured in the soil. At a capacitor manufacturing site, levels ranged
from 2.4 to 7.3 yg/kg, and levels of 0.3 yg/kg and 21 vg/kg were measured at
this site's new and old dumps, respectively. At another capacitor manufactur-
ing site, levels ranged from 2.7 to 470 pg/kg (Erikson et al. 1978). Brinkman
et al. (1981) reported PCB levels of 0.160 wg/g in one soil sediment sample
and less than 0.030 ug/g in four other samples. A study conducted by the EPA
revealed that PCB levels measured near five metropolitan areas were higher
than those measured in rural areas. The highest concentration reported in
this study was 11.94 wg/g (Carey et al. 1979).
Half-Life Estimates
No direct estimates have been made of the half-life of PCB's in soil.
Based on their strong adsorption to soil organic matter and their relative
insolubility in water, however, they are expected to be strongly persistent,
especially in soils with a high percentage of organic matter. The higher the
degree of chlorination is, the more persistent the compound.
5.8 POLYBROMINATED BIPHENYLS (PBB's)
Structure and Physical Properties
There are 209 possible polybrominated biphenyl isomers. These isomers
are usually found in mixtures. The major component isomer of PBB mixtures is
2,2',4,4',5,5'-hexabromobiphenyl (HBB) (WHO 1978, Jacobs et al. 1976). This
isomer, in combination with two isomers of pentabromobiphenyl (unspecified),
three additional isomers of hexabromobiphenyl (unspecified), and two isomers
of heptabromobiphenyl (unspecified) account for 98 percent of the PBB mixtures.
The molecular weight of 2,2',4,4',5,5'-hexabromobiphenyl is 627.59. Its
melting and boiling points were unavailable.
64
-------
Sources
Polybrominated biphenyls are manufactured.
Metabolites
Certain dibromo-substituted aromatic rings may be subject to attack by
biphenyl-oxidizing bacteria that require an adjacent 2-3 position to form the
first biphenyl metabolite, cis-2,3-dihydro-2,3-dihydroxybiphenyl (Jacobs et
al. 1976). Generally, however, PBB's are considered highly stable chemically
although some bromine replacement is possible abiotically.
Applications
Polybrominated biphenyls have a commercial application as a flame
retardant (Jacobs et al. 1976).
Toxicity
Polybrominated biphenyls are highly suspect as carcinogens (HHS 1983).
Water Solubility
Polybrominated biphenyls are relatively insoluble in water. Solubility
is believed to be in the ppb range and decreases with the addition of
bromine. Results of leaching studies on four Michigan soils amended with 100
ppm HBB showed that less than 0.6 percent of the compound was lost when
washed with quantities of leachate equivalent to 20 times the average annual
rainfall in Michigan (Filonow et al. 1976). Therefore, movement through soil
in the water phase can be expected to be negligible.
Volatilization
No information is available on the vapor pressure of PBB's.
Microbial Degradation
Evidence has shown microbial metabolism of certain dichloro-substituted
aromatic rings of PCB's. Therefore, certain dibromo-substituted aromatic
rings of PBB's could be similarly attacked by biphenyl-oxidizing bacteria.
These bacteria require an adjacent 2-3 position to form the first biphenyl
metabolite, cis-2,3-dihydro-2,3-dihydroxybiphenyl (Jacobs et al. 1976).
65
-------
Photodecomposition
Two independent researchers have shown that PBB's are sensitive to
light. According to Hill et al. (1982), the observed pattern of PBB degrada-
tion indicates a photochemical decomposition mechanism. The work of Jacobs
et al. (1978) indicates that photodecomposition of PBB isomers is greater
than microbial degradation, but that photodecomposition does not degrade
significant amounts of PBB. In addition, the photodecomposition products
that do exist appear to be soil-bound and mostly nonextractable.
Partition Coefficients
The log K for PBB's is 5.34. No measurements of the octanol/water
partition coefficient are available.
Adsorption
A very high direct correlation exists between the organic matter content
of soil and amounts of PBB adsorbed (Griffin and Chou 1981). Filonow et al.
(1976) showed that 2,2',4,4',5,5'-hexabromobiphenyl adsorption to soils
conforms to the Freundlich isotherm, r » 0.87 to 0.96. This team of researchers
also showed that neither the percentage of clay nor the pH correlated well
with hexabromobiphenyl adsorption.
Translocation to Plants
Orchard grass grown in PBB-contaminated soil showed no uptake, and
carrots showed only minor uptake (Jacobs et al. 1976).
Experimental Degradation Measurements
In an experiment conducted on Brookston loam, only one isomer of penta-
bromobiphenyl showed any significant disappearance after a 24-hour incubation
period (Jacobs et al. 1976). Other results showed that PBB's were detected
in contaminated manure 1 year after their application to soil (Jacobs et al.
1976). Griffin and Chou (1980) found no measurable degradation in PBB-contami-
nated soils over a 6-month period.
Detected Levels
The highest levels of PBB's found in soil, sediment, and water in a
survey made near manufacturing and use sites in northeast New Jersey were 4.6
ppm, 60 ppb, and 9.8 ug/liter, respectively (Stratton et al. 1979).
66
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Half-Life Estimates
No direct estimates of the half-life of PBB's in soil have been made;
however, based on their strong adsorption to soil organic matter and relative
insolubility in water, PBB's are expected to be strongly persistent, especially
in soils with a high percentage of organic matter.
5.9 POLYCHLORINATED NAPHTHALENES
Structure and Physical Properties
The chlorinated naphthalenes are structurally and chemically similar to
the polychlorinated biphenyls (PCB's) (Erickson et al. 1978). The molecular
weight of naphthalene itself is 128.2, and its melting point and boiling
point are 80°C and 218°C, respectively (CRC 1984).
Sources
Polychlorinated naphthalenes are manufactured.
Metabolites
A proposed metabolic pathway for naphthalene is hydroxylation, followed
by ring cleavage. Naphthalene is oxidized to 1,2-dihydroxynaphthalene
through l,2-dihydro-l,2-dihydroxynaphthalene before ring cleavage (Griffin
and Chou 1981).
Applications
Chlorinated naphthalenes have been used in electric wire insulation and
as additives to engine oil (Erickson et al. 1978).
67
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Toxicity
The penta- and hexachloronaphthalenes are the most highly toxic chlori-
nated naphthalene isomers (Erickson et al. 1978). They are toxic by
ingestion, inhalation, and skin absorption. They also are known to cause
chloracne, perhaps because they contain dioxin and/or furan contaminants.
Water Solubility
The chlorinated naphthalenes are slightly soluble in water. Naphthalene
itself is water-soluble at 31.7 ppm (unspecified temperature) (Briggs 1981),
and 2-chloronaphthalene is soluble at 6.74 ppm (Callahan et al. 1979).
Volatilization
No information was available on the vapor pressure or volatility of
chlorinated naphthalenes.
Microbial Degradation
The chlorinated naphthalene 1-chloronaphthalene is metabolized by
hydroxylation and ring cleavage of the unchlorinated ring (Verschueren 1983)
to form 3-chlorosalicylic acid. No information is available on microbial
degradation of the higher chlorinated naphthalenes.
Photodecomposition
The adsorption band of 1-chloronaphthalene is in the 300-nm region, and
this chlorinated naphthalene is susceptible to direct photolysis (Callahan et
al. 1979). No information is available for other chlorinated naphthalenes.
Partition Coefficients
The log K for naphthalene itself has been calculated at 3.36 (Briggs
1981). Partition coefficients are not available for chlorinated naphtha-
lenes, but they will be higher as a function of the degree of chlorination.
The log K of 1-chloronaphthalene is 4.12 (Callahan et al. 1979).
Adsorption
Naphthalene has a high affinity for organic substances. Using Freundlich
adsorption isotherms, Rippen et al. (1982) demonstrated that its adsorption
to soil is strongly correlated with the organic matter content of the soil.
68
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Translocatlon to Plants
No information is available on the uptake of naphthalene by vegetation.
Experimental Degradation Measurements
No information is available on experimental measurements of chlorinated
naphthalene degradation in soil.
Detected Levels
Soil samples collected in the vicinity of a polychlorinated naphthalenes
manufacturing plant contained predominantly tri-, tetra-, and pentachlorinated
isomers. Concentrations ranged from 130 to greater than 96,000 wg/kg, with
an average of 9400 wg/kg (Erickson et al. 1978). At another site, near a
drainage ditch, sediment samples contained 1250 to 5000 wg/kg (Erickson et
al. 1978).
Half-Life Estimates
Low to moderate water solubility, moderate volatility, and strong adsorp-
tion to soil organic matter indicate that chlorinated naphthalenes are highly
persistent in soil. The octanol/water partition coefficient of naphthalene,
however, is significantly lower than those of DDT, hexachlorobenzene, and
TCDD. Based on this information, naphthalene is probably less persistent
than these compounds and will probably move in the water phase and out of the
soil in the vapor phase.
5.10 TOXAPHENE
Structure and Physical Properties
The exact composition of toxaphene is unknown. It is a mixture of chlo-
rinated camphenes that contains more than 170 components having a wide range
of gas chromatographic retention times and 4 to 12 carbon atoms per molecule.
It is 67 to 69 percent chlorine and has an approximate formula of C10H.0Clg
(Willis et al. 1983, LaFleur 1974, WHO 1979).
The molecular weight of toxaphene is approximately 413.85, and its
melting point ranges from 65° to 90°C (Kirk-Othmer 1984).
Sources
Toxaphene is manufactured.
69
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Applications
Toxaphene Is an Insecticide used primarily on cotton and soybean crops
In their early stages of growth (LaFleur 1974).
Toxicity
Toxaphene Is toxic by ingestion, inhalation, and skin absorption. It is
also a suspected carcinogen (HHS 1983).
Water Solubility
The water solubility of toxaphene ranges from 0.4 to 7.0 ug/g (NASH CP
File 1985) and 0.4 ppm (Guenzi et al. 1974). Because it is relatively
water-insoluble, toxaphene's movement through soil in the water phase will be
negligible.
Volatilization
Toxaphene has a relatively low vapor pressure. Nevertheless, results
obtained by Seiber et al. (1979) showed that a decline in toxaphene
concentration in the aerated top soil of a cotton field environment occurred
primarily by vaporization. According to Willis et al. (1983), vaporization
is the major route of toxaphene loss from foliage. The vapor pressure of
toxaphene has been variously calculated to be 0.2 to 0.4 mm Hg at 20°C and 1
x 10"6 mm Hg at 25°C (Willis et al. 1983). This great difference is most
probably a result of the heterogeneous nature of the mixture of 170 components
that comprise toxaphene.
Microbial Degradation
Parr and Smith (1975) showed that toxaphene degraded rapidly in silt
loam soil under anaerobic conditions; the rate of degradation increased with
decreasing oxidation-reduction potential. No degradation occurred under
aerobic conditions for 6 weeks, and little degradation occurred in soil that
had been autoclaved before anaerobic incubation. Table 15 provides the
percentage of degradation that occurred after 6 weeks under various condi-
tions.
70
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TABLE 15. TOXAPKENE DEGRADATION UNDER VARIOUS CONDITIONS'
Percent
degradation
98
90
50
0
Presence of oxygen
Anaerobic
X
X
X
Aerobic
X
Moisture
Moist
X
X
Flooded
X
X
Stirred
X
Unstirred
(unspecified)
X
(unspecified)
a Parr and Smith (1975).
Similarly, Smith and Willis (1978) noted extensive toxaphene disappear-
ance in both amended and unamended anaerobic environments. Degradation was
most complete in flooded, stirred, and anaerobic environments; less complete
in flooded, unstirred, and anaerobic environments; and was not detected in
aerobic environments (Smith and Willis 1978).
In another experiment, extensive toxaphene degradation was thought to
have occurred by anaerobic reduction of the toxaphene components (Seiber et
al. 1979).
Williams and Bidleman (1978) showed that toxaphene degraded in anoxic
salt marsh sediments to compounds with shorter gas chromatographic retention
periods than toxaphene. This degradation occurred in both sterile and
nonsterile sediments.
These results indicate that microbial degradation of toxaphene in an
anaerobic environment is a likely mechanism of toxaphene loss from soil.
Photodecomposltion
The limited evidence suggests that toxaphene dehydrochlorinates in
prolonged sunlight (WHO 1979). It has a photolysis half life greater than 10
years (Callahan et al. 1979).
Partition Coefficients
The login K for toxaphene has been measured at 3.23 to 5.43. The
log1Q KQC is 4.72 (NASH CP File 1985).
Adsorption
Toxaphene has an affinity for hydrophobic solvents and substrates (La
Fleur 1974). Measured concentrations have been higher in soils with high
71
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organic matter content than In soils free of organic matter (Gallagher et al.
1979). A linear relation has been determined between toxaphene yield and
sediment yield in runoff from silty clay soil (McDowell et al. 1981). These
results indicate the strong affinity of toxaphene for soil organic matter, a
finding that corroborates the conclusion of the U.S. Department of Health and
Human Services (1983) that toxaphene is extremely persistent in soil.
Trans location to Plants
Gallagher et al. (1979) report the uptake of toxaphene into unspecified
plant tissues in both sandy and silty clay loam marshes; concentrations were
highest in the roots.
Experimental Degradation Measurements
In a series of experiments conducted in the top 5 feet of Houston black
clay, less than 22 percent of the toxaphene applied over 10 years was
recovered. Of the toxaphene that was recovered, 90 to 95 percent was in the
top 12 inches of soil (Swoboda et al. 1971). Forty-five percent of the
toxaphene applied to a sandy loam soil in 1951 was still there 20 years later
(WHO 1979). One-half of the toxaphene applied to a topsoil at a rate of 100
kg/ha remained after approximately 100 days. Toxaphene was then found in the
underlying ground water within 2 months (LaFleur et al. 1973).
These results seem to conflict. The varying composition of toxaphene in
combination with numerous interacting environmental factors may contribute to
the differing behavior of toxaphene under various conditions.
Detected Levels
In a study conducted in Colorado, toxaphene was discovered in 1 of 50
samples taken from agricultural soils (Mullins et al. 1971). Soil samples
taken from three municipal locations in 1969 contained toxaphene concentrations
of 0.11, 12, and 15-53 mg/kg. Three of 28 samples taken from one location in
1970 contained concentrations ranging from 7.7 to 33.4 mg/kg, and 1 of 27
samples taken at a second location contained a concentration of 16.1 mg/kg
(WHO 1979).
A sediment sample taken at the soil surface 0.2 mile from a toxaphene
plant outfall contained 1858 mg/kg, and one taken 1.4 miles away at a depth
of 70 to 80 cm contained 5.27 mg/kg (WHO 1979).
72
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Half-life Estimates
One estimate of toxaphene half-life in soil is 11 years (Guenzi et al.
1974). Seiber et al. (1979) noted "extensive" toxaphene component degradation
after 1 year.
Agreement has not yet been reached on the half-life of toxaphene in
soil. Its strong adsorption to soil organic matter, coupled with its
relative insolubility in water, indicates strong persistence in soil.
Degradation of toxaphene under anaerobic conditions does not indicate
biodegradation in the normally aerobic environment of soil pore spaces.
These three factors, in combination, increase the likelihood of toxaphene's
strong persistence in soil.
5.11 HEXACHLOROBUTAOIENE
Structure and Physical Properties
V ^
c = c - c = c
Cl
The molecular weight of hexachlorobutadiene is 260.76, and its melting
point and boiling point are -21°C and 215°C, respectively (CRC 1984).
Sources
Hexachlorobutadiene is manufactured.
Metabolites
No information is available on the breakdown products of hexachlorobuta-
diene.
Applications
Hexachlorobutadiene is used as a heat-transfer liquid and as a trans-
former and hydraulic fluid (Kirk-Othmer 1984).
73
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Toxicity
hexachlorobutadiene Is a suspected carcinogen, and It is toxic by
inhalation and ingestion (HHS 1983).
Mater Solubility
The water solubility of hexachlorobutadiene is 2 ug/g. Because of this
relative insolubility, its movement through the soil in the water phase is
probably limited, but slight movement is possible. Results published by
Litvinov and Gorenshtein (1982) indicate that rain stimulates the penetration
and decreases the volatilization of hexachlorobutadiene.
Volatilization
Hexachlorobutadiene has a moderate vapor pressure of 22 mm Hg at 100°C.
3
In experiments where measurements in air of 0.08 and 0.03 mg/m were taken
one day and one month, respectively, after application of 250 kg/ha, investi-
gators also showed that vaporization of hexachlorobutadiene is quicker from
light than from heavy soils, and that increasing the temperature from 13° to
15°C multiplied the vaporization rate fivefold (Litvinov and Gorenshtein
1982).
Microbial Degradation
No evidence was found to indicate microbial degradation of hexachloro-
butadiene. In fact, the evidence indicates that hexachlorobutadiene may be
toxic to microflora (Mirzonova and Perov 1963, Drui et al. 1969). Doses of
150 to 250 kg/ha decreased the nitrifying power of the soil slightly; it was
restored 2 to 5 months after treatment. Doses of 100 mg/kg did not affect
the nitrate content of the soil, whereas doses greater than 100 mg/kg decreased
nitrate levels. The actual numbers of microbes were unaffected, even at
doses of 1000 mg/kg. Above 1000 mg/kg, the nitrification process was inter-
rupted at the nitrate formation step, possibly as a result of oxygen
deficiency in a soil saturated with hexachlorobutadiene vapors (Drui et al.
1969). In another set of experiments, soil nutrient content and hyarolytic
enzyme activity decreased and were restored 3 to 4 months later (Mirzonova
and Perov 1968).
-------
Photodecomposition
No Information Is available on the photodecomposition of hexachlorobuta-
dlene.
Partition Coefficients
The log KCW of hexachlorobutadiene is 3.4 (Callahan 1979).
Adsorption
Investigations show a direct correlation between retention of hexachloro-
butadiene and soil organic matter content (Gorenshtein and Litvinov 1983).
This is to be expected in view of the relative insolubility of the compound
in water. Some movement will occur.
Translocatlon to Plants
No information is available on the uptake of hexachlorobutadiene by
vegetation.
Experimental Degradation Measurements
Vorob'eva (198G) discovered that hexachlorobutadiene had disappeared
completely 4 years after several aerial applications of 250 kg/ha. The
compound was applied 30 cm deep and accumulated to depths of 50 to 75 cm near
the plants and 75 to 100 cm in the rows between the plants.
Detected Levels
In one experiment conducted in Russia, concentrations of 0.08 and 0.03
Q
mg/m were detected one day and one month, respectively, after aerial
application of 250 kg/ha; 0.06 and 0.001 mg/ha were detected one day and one
month, respectively, after aerial application of 150 kg/ha (Litvinov and
Gorenshtein 1982).
Half-Life Estimates
No direct estimates have been made of the half-life of hexachlorobutadiene
in soil. The results of a few Russian researchers indicate that although
hexachlorobutadiene is moderately persistent, it will dissipate from the soil
within a few years after aerial application of moderate amounts. This dissipa-
tion will occur despite its relative insolubility in water and toxicity to
certain soil microbes.
75
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5.12 DDT (DICHLORODIPHENYL TRICHLOROETHANE) AND ITS METABOLITES, ODD (DI-
CHLORODIPHENYLDICHLOROETHANE) AND DDE (DICHLORODIPHENYLDICHLORO-
ETHYLENE)
Structure and Physical Properties
p.p'-DDT
Cl
CC1.
r
C
I
M
Cl
The molecular weight of p.p'-DDT Is 354.5. Its melting point and
boiling point are 108°C and 260°C, respectively (CRC 1984).
o.p'-DDT
CC1,
Cl
Cl
The molecular weight of o.p'-DDT is 354.5. Its melting point is 74°C
(CRC 1984).
p.p'-DDD (TDE)
HC Cl.
Cl
Cl
76
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The molecular weight of p.p'-DDD is 320.0. Its melting point is 109°C
(CRC 1984).
p.p'-DDE
The molecular weight of p.p'-DDE is 318.0. Its melting point is 88°C
(CRC 1984).
Sources
DDT (Dichlorodiphenyltrichloroethane) is manufactured. Technical DDT is
a mixture of p,p'- and o,p'-DDT in a ratio of approximately 3:1. It includes
traces of some other isomers and analogs as well (Terriere et al. 1965).
Metabolites
The primary metabolites of DDT are DDD (TDE) and DDE. Like DDT, they
are highly soluble in organic solvents and slightly soluble in water (WHO
1974). Water solubilities range from 0.002 to 0.06 mg/liter.
Dicofol [4,4'-dichloro-o-(trichloromethyl)benzhydrol] is another degrada-
tion product of DDT that is sometimes found in trace amounts.
The primary breakdown product of DDT is p.p'-DDE; and the secondary
breakdown product is p.p'-DDD (Forsyth et al. 1983, Stewart and Chisholm
1971, Stringer et al. 1974).
Applications
The pesticide DDT was used primarily on cotton and tobacco. It is no
longer in common use because it adversely affects wildlife.
Toxicity
Toxic by ingestion, inhalation, and skin absorption, DDT is also carcin-
ogenic (WHO 1977).
Water Solubility
This compound is practically insoluble in water. The water solubility
of o.p'-DDT is 0.085 mg/liter at 25°C, and that of DDT (isomer unspecified)
77
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Is 0.001 to 0.085 ug/liter (Guenzl et al. 1974, MHO 1974, Fuhremann and
Lichtenstein 1980, NASH CP File 1985, Callahan et al. 1979). When water is
added to soils that have adsorbed chlorinated hydrocarbons such as DDT, they
displace the pesticide and cause an increase in the vapor pressure and subse-
quent vaporization of the displaced compound (Guenzi et al. 1974). The high
persistence and immobility of DDT are attributed, in part, to its very low
solubility in water. Amounts adsorbed are inversely related to water
solubility.
The extreme hydrophobicity of DDT and its low solubility in water indicate
limited leaching. Nevertheless, DDT molecules tend to migrate to the surface
and evaporate with water molecules more quickly than the extremely low vapor
pressure would predict (Guenzi et al. 1974).
Volatilization
The vapor pressures of DDT and its primary metabolites are recorded in
Table 16. The vapor pressure of DDT is low, and its metabolite DDE is almost
eight times more volatile than DDT itself (Forsyth et al. 1983, Spencer et
al. 1974).
TABLE 16. VAPOR PRESSURES OF DDT AND METABOLITES
(mm Hg)
Compound
p.p'-DDT
o.p'-DDT
p.p'-DDD
p,p'-DDF
Vapor pressure, mm Hg
1.5 x 10"7 at 20°Ca>b
1.9 x 10"7 at 20°Ca>c
7.3 x 10"7 at 30°Cd
5.5 x 10"6 at 30°Cd
1.0 x 10"6 at 30°Cd
6.5 x 10"6 at 30°Cd
d Willis et al. (1983)
b Sleicher and Hopcraft (1984)
c Fuhremann and Lichtenstein (198C)
d Guenzi et al. (1974)
73
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When the soil moisture Is In a certain optimum range, DDT volatilizes.
Guenzi and Beard (1970) showed that as soil moisture dropped, DDT was lost at
a constant rate until the soil contained less than a monolayer of water on
soil surfaces (Guenzi and Beard 1970). Increasing temperature causes an
Increase In the rate of volatilization (Guenzi and Beard 1970, Baker and
Applegate 1970).
Some data suggest that DDT sublimes directly without prior degradation.
Test results have shown that sublimation alone can account for the disappearance
of low-volatility pesticides, even If they are strongly adsorbed to the soil
(Sleicher and Hopcraft 1984). Calculated sublimation rates based on mass
transfer through the boundary layer were sufficient to account for the disap-
pearance of DDT.
In other research, DDT losses were attributed primarily to volatiliza-
tion because the losses in runoff, emigrating insects, and small animals were
negligible.
Farmer et al. (1972) obtained maximum volatilization rates, which repre-
sented DDT losses of 5 kg/ha per year. Table 17 gives potential volatil-
ization rates of DDT from Gila silt loam at 10 percent soil water content and
100 percent relative humidity.
TABLE 17. POTENTIAL DDT VOLATILIZATION3
Soil Concentration, yg/g
DDT loss, kg/ha per year
1
5
10
50
0.28
1.3
2.9
4.7
a Farmer et al. (1972).
Volatilization rates are partially dependent on the vapor pressure. A
comparison of high and low vapor pressure compounds showed that the volatil-
ization rates of DDT from soil were low at first and decreased slowly, whereas
compounds of the former group had high initial volatilization rates that
decreased rapidly. Both groups were volatilizing at the same rate by the
79
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ninth aay, at which point more than half the DDT was still In the soil (Nash
1983). Volatilization may be a significant mechanism for DDT loss from the
soil, particularly where soil organic matter content is low.
Microbial Degradation
Although the results of Forsyth et al. (1983) indicate that DDT is
resistant to microbial degradation, other researchers have published
contrasting results.
Test results show that DDT degrades much more quickly anaerobically than
aerobically, although the path of DDT degradation is not fully understood
(Guenzi et al. 1974). The compound DDT degrades anaerobically to DDD, and
many microbes can dehydrohalogenate DDT to DDE. Further degradation has been
difficult to establish. The compound o,p'-DDT (15% of DDT) converts to
o.p'-DDD and o.p'-DDE. Figure 1 illustrates a potential microbial degrada-
tion pathway.
A Hydrogenomas sp. degraded DDM by ring cleavage to p-chlorophenylacetic
acid. The proposed metabolic pathway included ring hydroxylation followed by
ring cleavage (Guenzi et al. 1974). Metabolism of p-chlorophenylacetic acid
has been reported, but no metabolic products have been identified. Another
possible DDT degradation product is p-Chlorobenzoate.
Sethunathar, et al. (1969) noted anaerobic degradation of DDT by Clostri-
dium, possibly by reductive dechlorination (Sethunathan et al. 1969). Some
evidence suggests that biodegradation of DDT to DDD is anaerobic and rapid,
whereas biodegradation of DDT to DDE is aerobic and slow (Brady 1974).
Other work showed that DDT degraded more quickly in flooded, temporarily
anaerobic soils than in aerobic soils (Guenzi et al. 1974). This may be due
to reductive degradation.
When soil was flooded for 7 weeks, DDT degraded primarily to DDD; there
was little effect on DDE isomers. This degradation was observed with and
without the addition of organic matter to the soil. Flooding caused a decrease
in the quantity of DDE volatilizing from the soil because the DDT that
volatilized was degraded to DDD instead of DDE (Spencer et al. 1974).
Alexander (1981), Castro and Yoshida (1974), and Anderson et al. (1970)
showed that DDT was modified in nonsterile environmental samples, but not in
sterile environmental samples. Degradation rates were increased by adding
more organic matter to the soil (Castro and Yoshida 1974). Table 18 shows
the important reactions in the transformation of DDT.
80
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ODE
ODD
dlcofol
ODMU
DDKS
DOOM
ODNU
DDA
OBP
DDK
OBH
Key
l.l-dtchloro-2.2-b1s(£-chlorophenj'l)ethylene
l.l-d1chloro-2.2-bis(p-chlorophenyl)ethane
l.l.l-trichloro-2.2-bTs(])-chlorophenyl)ethanol
l-chloro-2,2-b1s(£-chlorophenyl)etliylene
l-chloro-2,2-bis(p-chlorophenyl)ethane
2.2-bts(p.-chloropKenyl(ethanol
I.l-b1s(e-chlorophenyl)ethylene
2.2-b1s(p-chlorophenyl)acet1c acid
4.4'-dlcnlorobeniaphenone
b1s(p-chlorophenyl)methane
duhTorobenzhydrol
/
DDE
DDT * dicofol
1
ODD * DDMS * DDOH •» DDNU
DDA * DDM - DBH * DBP-
p-chlorophenylacetic acid
p-chlorobenzoate
Figure 1. Biodegradation of DDT.
81
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TABLE 18. REACTIONS OF ENVIRONMENTAL IMPORTANCE FOR TRANSFORMATION OF DDTa
Dehalogenation
bAr2CHCH2Cl + Ar2C=CH2
Ar2CHCHCl2 * Ar2C=CHCl2
Ar2CHCCl3 + Ar2CHCHCl2
Ar2CHCCl3 * Ar2C=CCl2
Decarboxylation
Ar2CHCOOH * Ar2CH2
Reduction of double bond
Ar2C=CH2 * ArzCHCH3
Ar2C=CHCl * Ar2CHCH2Cl
Hydration of double bond
Ar2C=CH2 •» Ar2CHCH2OH
a Alexander (1981).
Ar » Aromatic chlorinated ring.
Evidence shows that the addition of DDT to soil did not adversely affect
populations of bacteria or actinomycetes, and the application of DDT at
normal rates and 10 times the normal rate did not significantly affect oxygen
uptake by soil microbes (Ruffin 1974, Talekar et al. 1978, Hubbell et al.
1973). Harris (1969) noted no significant decrease in biological activity
over 48 weeks. Numbers of fungi in soil did increase significantly after
soil treatment with a combination of DDT, parathion, and zineb (Hubbell et
al. 1973).
Mucor alternans. a fungus isolated from agricultural loam soil, partially
degraded DDT in 2 to 4 days into three hexane-soluble and two water-soluble
metabolites. None of the metabolites were identified as DDE, DDD, DDA, DBF,
82
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dlcofol, or l.l-bis(p-chlorophenyl) ethane. In the laboratory, DDT degrada-
tion was shown with live fungi myeelia only, but dead mycelia did absorb the
insecticide. Addition of Mucor alternans spores to soil containing DDT
caused no degradation (Anderson et al. 1970).
These results indicate that microbial degradation of DDT is clearly one
mechanism by which DDT is lost from the soil. The rate at which biodegrada-
tion occurs is not clear.
Photodecomposi t i on
The only information available on the photodecomposition of DDT was
published by Baker and Applegate (1970). They maintained that ultraviolet
radiation catalyzed or accelerated the breakdown of DDT to DDE. Exposure to
ultraviolet radiation consistently resulted in additional loss of DDT over
the losses obtained without radiation from three different soils, at two
different temperatures, and at two different application rates. Table 19
presents these data.
TABLE IS. ADDITIONAL LOSSES OF DDT BY UV RADIATION3
Conditions
5 ppm,
50 days elapsed
20 ppm,
60 days elapsed
Percent loss
at 30°C
4-8
3-11
Percent loss
at 50°C
7-9
29-32
a Baker and Applegate (1970),
Partition Coefficients
The following log K values have been measured for p,p'-DDT: 5.57,
5.99, 6.19, and 6.36 In KQW. Log KQW values for p,p'-DDD and p.p'-DDE are
5.06 and 5.43, respectively. Log KQC values for p,p'-DDT, p.p'-DDD, and
p.p'-UDE are 5.26, 4.90, and 4.74, respectively (NASH CP File 1985).
Adsorption
The extreme hydrophobicity of DDT and its strong affinity for soil
organic matter account for its very low leaching rate ana relative immobility
in soil (Guenzi and Beard 1967, Guenzi et al. 1974, Reed and Preister 1969).
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Kiigemagi and Terrlere (1972) found that, over a 20-year period, about
one-third of the decrease in the amount of DDT and its analogs in the top 6
inches of soil was a result of downward movement. Terriere et al. (1965)
found that less than 5 percent of the remaining pesticide was below a depth
of 1 foot after 20 years. These results indicate a very slow leaching rate.
Lichtenstein et al. (1971) found that although DDT was applied to a
depth of 5 inches, 30 percent of the DDT recovered 10 years later was at a
depth of 6 to 9 inches.
A number of researchers have published mobility factors for DDT and its
metabolites (Helling 1971, Fuhremann and Lichtenstein 1980). All results
indicate the relative immobility of DDT. Table 20 provides relative mobility
factors for DDT ana some of its metabolites. These factors show that DDE is
more mobile than DDT.
TABLE 20. MOBILITY FACTORS FOR DDT AND METABOLITES3
Compound
p.p'-DDE
o.p'-DDT
p.p'-DDT
DBP
TDE (ODD)
dicofol
DDA
Mobility factor, Rf
0.81
0.70
0.62
0.42
0.24
0.10
0.00
a Fuhremann and Lichtenstein (1980).
Owen et al. (1977) saw little vertical movement of DDT into lower soil
horizons over a 10-year period. Sharom et al. (1980) found that DDT was the
most highly adsorbed of 12 insecticides tested. A strong correlation was
found between adsorption and soil organic matter. Desorption and mobility
decreased with decreasing solubility of the insecticide in water. Insecticides
were desorbed most easily from sand, less easily from sandy loam, and the
least easily from sediment. Almost no DDT was desorbed from soil with four
rinses, and results showed 100, 90, 98, and 97 percent adsorption to organic
soil, creek sediment, sandy loam, and sand, respectively (Sharom et al.
1980).
84
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Mobility factors were calculated by the following equation:
n = 10
Mobility Factor (MF) = Z (H-n)(xn/y) (Eq. 5-3)
n = 1 n
where n = Number of 200-ml fractions
x = Amount of insecticide in fraction
y = Amount of insecticide in treated soil
According to this equation, DDT mobility factors were equal to zero in
both sand and organic soil. In spite of different chemical structures, the
researchers showed a strong correlation of adsorption and desorption with
water solubility (Sharom et al. 1980, Miles et al. 1978).
In other work, adsorption of DDT to four different soils was shown to be
directly proportional to the percentage of organic matter in the soil. No
DDT was leached from sand or muck with ten 200-ml water rinses (Miles et al.
1978). Using DDD, Mueller-Wegener (1981) also demonstrated a linear relation
between soil organic matter content and adsorption.
Lichtenstein et al. (1977) showed that 25 percent of 14C- labeled DDT
became bound to an agricultural soil, which left about 70 percent that was
extractable. Stream water and bottom mud contained less than 1 ppm DDT
residue, which signified little or no movement of DDT from soils into nearby
waterways (Kuhr et al. 1974).
Translocation to Plants
Residues of DDT have been detected in oat roots (but not oat tops), in
grass roots at concentrations 5 to 15 times higher than those in surrounding
soil, and in aerial grass parts at low concentrations (Fuhremann and Lichtenstein
1980, Voerman and Besemer 1975). Residues of DDT have also been found in
potatoes (at 0.079 to 0.094 mg/kg), in peanut hulls and forage, and in turnip
greens (Uhnak et al. 1979, Young 1969). Tobacco leaves and soybeans took up
0.70 and 0.65 ppm, respectively, from sandy loam soil containing 2, 8, and 16
pounds per acre of DDT (Reed and Preister 1969). Residues of DDT have also
been founa in sugar beets at 5.5 percent of the soil concentration at
planting time (Onsager et al. 1970).
85
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Experimental Degradation Measurements
Numerous results have been published on measurement of DOT degradation
in soil. The amount of DDT converted to DDE ranged from 6.7 percent in
Valentine loamy sand to 21.2 percent in Raber silty clay loam (time period
unspecified) (Guenzi and Beard 1970).
Over a period of 5 years, 18 percent of the DDT applied to an unculti-
vated sandy loam and 17 percent of the DDT applied to a cultivated clay was
lost (Kiigemagi and Terriere 1972). The decline in total residues was proceed-
ing at a rate of 3.5 percent per year. These investigators referred to
another study in which the initial DDT concentration was 2300 ppm; a decline
of 30 to 50 percent was noted over a 20-year period.
In a study performed by Lichtenstein et al. (1971), the total amount of
DDT recovered 15 years after soil treatment was 10.6 percent of the combined
dosage of 10 pounds per acre and 17 percent of the combined dosage of 100
pounds per acre. Soils treated with higher DDT dosages showed higher residue
levels, which indicated the relatively higher stability of insecticide in
soil treated at a higher dosage. When present in high concentrations, DDT
did not metabolize to p,p'-DDE and other residues as quickly as when present
at lower concentrations. Fifteen years after application of technical DDT at
10 or 100 pounds per acre, 18 percent and 24 percent, respectively, of the
applied aosages were present in the upper 6 inches of soil as p,p'-DDT,
o,p'-DDT, and p,p'-DDE. During the 15-year period, only 75.2 percent of the
o,p'- isomer was lost, compared with 84.5 percent of the p,p'- isomer,
which indicates that the former was more persistent. In other work (Terriere
et al. 1965), the ratio of p.p'-DDT to o,p'-DDT ranged from 7 to 14 at different
soil levels, which indicates that the o,p'- isomer is less persistent in soil
than is the p,p'- isomer.
According to Baker and Applegate (1970), at an application rate of 5
ppm, 21 to 24 percent of DDT was lost from three soils after 50 days at 30°C,
and 43 to 50 percent was lost at 50°C. At an application rate of 20 ppm, 22
to 27 percent was lost after 60 days at 30°C and 35 to 44 percent was lost at
50°C. Ultraviolet radiation of the soils increased these DDT losses, as
shown earlier in Table 19.
86
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In the fall of 1974, Forsyth et al. (1983) found that the total quantity
of DDT residues in the top 12 cm of soil was approximately 38 percent of the
total amount applied in 1969. Fifteen years after the last application, 57.1
percent of a total of 585 kg/ha DDT applied remained in the soil (Chisholm
and McPhee 1972).
Guenzi et al. (1974) estimated that DDT persisted in the soil for 4
years. The DDT was incorporated into Carrington silt loam at a depth of
13 cm on 5 consecutive days each week for 3 months. Four months after treat-
ment, soil analyses showed that 55.9 percent of the applied DDT remained in
disked soil and 74.2 percent remained in undisturbed soil (Guenzi et al.
1974).
Twelve years after application, Owen et al. (1977) showed that soil from
areas sprayed with 1.12 kg/ha still showed residues of 4.5 ppm and that 1973
and 1976 levels were not significantly different from 1967 levels. Their
conclusions that higher dosages decay more slowly on a percentage basis
corroborate the earlier findings of Kiigemagi and Terriere (1972).
Another researcher found that over 7 months (April through October) the
ratio of DDE to DDT increased sevenfold, whereas the ratio of DDD to DDT
increased fivefold (Miles 1980). In related work, no significant decrease in
total amounts of p,p-' and o.p'-DDT was noted 4 years after application to a
light sandy soil with 3 percent organic matter; however, over that time the
p,p'-DDE concentration more than doubled (Voerman and Besemer 1975).
In a study by Yule et al. (1967), no DDT degradation was noted over a
6-month period. In another study, 25 to 46.3 percent of the p,p'-DDT
remained in seven containers of air-dried soil 2 years after application
(Saito and Kitayama 1973). In still another study, p.p'-DDT residues
exceeded 50 percent of the original dose 3 years after application, and no
degradation at all was noted in the winter when the soil temperature was less
than 5° to 8°C (Czaplicki 1979). In Norway, roughly 45 percent of DDT
sprayed in eight orchards since 1945 remained in 1968 (Stenersen and Friestad
1969). The upper 5 cm was most heavily contaminated; only minor amounts were
present in the 15- to 20-cm layer. In another investigation, levels of DDT,
DDD, and DDE in orchard soil were unchanged 2 and 4 years after termination
of the use of DDT (Kveseth et al. 1977).
87
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Less than 16 percent of the DDT applied to Houston black clay over 10
years was recovered in the top 5 feet of soil, and 60 to 75 percent of the
amount recovered was in the top 12 inches (Swoboda et al. 1971). This informa-
tion indicates that DDT does not leach considerably, as expected from its
relative water insolubility. The small amount of downward movement that did
occur was thought to be caused by the washing of top soil into large cracks.
In a study by Stewart and Chisholm (1971), 16 percent of the amount of
DDT originally applied was recovered from a sandy loam 15 years after its
last application, compared with the recovery of less than half that amount
(7.5%) of lindane, which is a more water-soluble pesticide.
Application of 10 kg/ha p,p'-DDT to an organic soil in Ontario, Canada,
showed a loss of 94.8 percent after 180 days (Miles et al. 1978). Agnihotri
et al. (1977) corroborated these findings exactly. In India, Chawla and
Chopra (1967) showed p,p'-DDT losses of 36.5 to 38.1 percent and 46.7 to 51.3
percent after 9 months in normal and alkaline soils, respectively. Chisholm
and McPhee (1972) showed that 57.1 percent of the applied DDT remained in
soil after 16 years, and Stewart and Chisholm (1971) showed that 55 percent
of the DDT remained in a sandy loam soil after 15 years.
Over a 7-year period, Ware et al. (1978) found that soil residues of
DDT-related degradation products declined an average of 23 percent. Residues
in desert soil declined 60 percent.
From 1953 to 1969, Stringer et al. (1974) annually measured DDT levels
at an orchard sprayed with technical DDT (77 percent p,p'- and 22 percent
o,p'-). After 17 years, 21 percent of the p,p'-DDT and 7 percent of the
o,p'-DDT remained. Eighty percent of the recovered DDT was detected in the
top 10 cm; none was detected below SO cm.
In five orchards, DDT residues of 258.8 pounds per acre and 7.3 pounds
per acre were detected under trees and between tree rows, respectively, 12
years after application. Concentrations decreased with depth. Seventy to 90
percent of the total residue was DDT (Kuhr et al. 1974).
In a Canadian study, the proportions of p,p'-DDT decreased between 1968
and 1971, whereas the relative content of p,p'-DDE and, to a lesser extent,
o.p'-DDT, increased (Yule 1973). In a study by Kuhr et al. (1972) the propor-
tion of DDE in the recovered residue increased from 12 percent after 12 years
to 27 percent after 24 years; the percentage of DDT recovered as DDT and DDE
-------
decreased from 50 to 33 to 22 at 6, 12, and 24 years, respectively (Kuhr et
al. 1972).
Yeadon and Perfect (1981) found that the final soil residue of plots
treated with 70 kg/ha DDT over 4 years In the subhumid tropics was 2 to 3
percent of the amount applied. In Taiwan, 6.6 percent of the DDT applied to
sllty loam soil at 5 kg/ha remained after 5 years; residues degraded rapidly
after spring application (hot and rainy), but accumulated after fall applica-
tion (cool and dry); and cultivated soil retained more Insecticide than did
fallow soil (Talekar et al. 1983).
Chapman et al. (1981) showed that DDT degraded more quickly In mineral
soil than in organic soil. Reed and Priester (1969) noted a "small" loss of
DDT from planting to harvesting time.
These results point toward the strong persistence of DDT in soil.
Although some researchers noted relatively rapid dissipation of DDT under
certain circumstances, the preponderance of evidence indicates that DDT in
soil does not degrade quickly and may last for many decades, or possibly
longer.
Detected Levels
A World Health Organization report (1974) indicates average soil
concentrations of DDT plus DDE were 0.5 ppm in soybean soil, 10.1 ppm in
orchard soil, 0.85 ppm in alfalfa soil (including DDD), and 15.10 ppm in
onion soil (including DDD). The overall U.S. average equals 0.168 gram DDT
per square meter.
On 14 farms, Miles and Harris (1978) found average DDT concentrations
of 29 ppm, with a maximum concentration of 60 ppm.
In a study by Steriersen and Friestad (1969), 32 of 55 orchards contained
DDT concentrations ranging from 2.3 to 10 mg/kg in the upper 15 cm of soil.
Four of the 32 orchards had more than 30 ppm in the upper 15 cm. The average
value was 10.1 ppm.
Residues of DDT, DDE, and DDD were detected in Taiwanese asparagus
fields at 127.1, 252.2, and 66.9 ppb, respectively (Wang et al. 1981). In
Norway, 6.5 ppm DDT was detected in orchard soil (Kveseth et al. 1977). In
Finland, 0.02 ppm DDT was detected (Rautapaa et al. 1976).
89
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In a study of soils in five U.S. cities, high DDT levels were found in
all samples (Carey et al. 1979). The DDT levels were higher in cultivated
areas than in undeveloped areas, and higher in urban soils than in cropland
soils. Twenty-seven of 50 samples of Colorado agricultural soils contained
DDT (Mullins et al. 1971). In southwest Ontario, DDT levels ranged from less
than 0.1 to 29 ppm (Miles and Harris 1978). Apple and peach orchard soils in
Ontario had DDT levels of 43.3 and 9.22 ppm, respectively. In West Germany,
out of 1035 soil samples from 222 sites with at least one DDT treatment in
the preceding 5 years, 4.6 percent had no DDT and 10 percent had more than 2
ppm DDT (Heinisch et al. 1968). In western Oregon, DDT was detected in four
different areas, although only one ever received a direct application of DDT
(Moore and Loper 1980). Levels in Maryland were measured at 1.5 ppm (Gish
1970). Levels in an estuary in New York were measured at an average of 15
kg/ha and a maximum of 36 kg/ha (Woodwell et al. 1967).
Half-life Estimates
According to Edwards (1966), the average time for 95 percent disappear-
ance of DDT is 10 years, and an average of 50 percent remains after 3 years.
Chisholm and McPhee (1972) show a half-life estimate of 15 years. Other
half-life estimates include 110 days, 2 to 4 years, 10.5 years, 3 years, 1.5
years (for o.p'-DDT only), approximately 12 years (in cultivated soil), 7
years (in desert soils), and 22.9 months (Sleicher and Hopcraft 1984,
Stringer et al. 1974, Buck et al. 1983, Onsager et al. 1970, Guenzi et al.
1974).
According to Baker and Applegate (1970), DDT may persist for 10 or more
years. Cooke and Stringer (1982) published half-life estimates of 11.7
years, 3.3 years, and 7.1 years for p.p'-DDT, p,p'-DDD, and o.p'-DDT, respectively.
Owen et al. (1977) published a list of half-life estimates for various
soils. Table 21 shows these estimates.
90
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TABLE 21. HALF-LIFE ESTIMATES FOR VARIOUS SOILS*
Soil
Oregon forest
New Brunswick forest
Agricultural
Light sandy
Maine forest
Years
Less than 10
Average 10.5,
maximum 35
No loss after 4
Little loss after 9
a Owen et al. (1977).
The soils of northern Maine contain many factors that enhance persistence,
Including high organic content, low temperatures, low microblal populations,
slow Incorporation of organic matter, and low pH. These factors Interact to
result In slow decay of DOT. The data Indicate that DDT residue decay,
especially In forest soils such as those In northern Maine, Is slow and may
approximate a half-life of up to 35 years.
A summary of properties ano resulting half-life estimates Is presented
In Table 22. The listed values were compiled from throughout this document.
91
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TABLE 22. SUMMARY OF PROPERTIES AND HALF-LIFE
ESTIMATED FOR TOXIC ORGANIC CHEMICALS
Hexachlorobenzene
1.2-Ofchlorobenzene
3.3>.4.4l-Tetrachloro-
azobenzene
Hexachlorocyclohexane
2.3.7.8-TCDF
2.3.7.8-TCDD
PCB's
PBb's
Polychlorlnated
naphthalenes
Toxaphene
Hexach 1 orobu tad < ene
p.p'-DDT
o.p'-DDT
p.p'-OOD
p.p'-ODE
Hater solubility og/1
.0062 (23.5°C)
49-123
Relatively insol.
0.15-17
0.0002 (25°C)
0.0002 (25°C)
0.006 l30°C)
0.007-S.9
0.0008
1-C1; 6.74
Slightly soluble
0.4-7.0
2
0.002 to 0.085
0.065
0.020-0.100
0.01 to 0.12
Vapor pressure
1.91 x 10-5 roiHg (2S°C)
1.5 ornHg (25°C)
Moderately 1m
9.4 x 10~5 mnH9 (20°C)
1.3 x 10"* omHg (30°C)
2.0 x 10'6 moHg (25"C)
0.2 mPa
1 omHg (89.3°C) 2 chloro-
blphenyl
1 omHG (96.4°C) 4-chloro-
blphenyl
-
1-C1; 0.017 (20°C)
-
0.17-0.40 mnHg (20°C)a
1 x 10"6 mnHg (20-25'C)
22 antig (100°C)
1.5 to 7.3 x 10"7
5.5 x 10"6
10.2 x 10"7
6.5 x 10'6 (20«C)
"ow ""S'
5.57a
2.26-3.40
4 to 5
3.80
6.95d
6.15*
6.00 to 6.84d
-
1-C1; 4.12
-
3.3
5.3
3.4
3.98 to 6.19
.
5.99
5.43 to 5.69
3.59
2.23-3.40
_b
3.00
-
5.91
3.70
5.34
_
-
4.74
-
-
5.26
.
4.90
4.74
t. estimates (in soil)
None (strongly persistent)
None (persistent)
Not persistent
Not highly persistent
260 days <• t. estimate
2 years - t. 'estimate
Strongly persistent
<200 ddys; 330 days; 190
days; 1 year; 180 days.
(indications are a rapid
rate for the first 6 mo.
then a much longer t.)
Very environmentally
persistent
Strongly persistent
Strongly persistent
Highly persistent
11 yr, extensive after 1 yr
Moderately persistent
3 yr, 15 yr, 110 days
.
2-4 yr, 10.5 yr, 3 yr
I.Syr
10
ro
a Estimated.
Information unavailable.
-------
SECTION 6
INFORMATION GAPS
During the collection of the information for this report, several data
gaps were identified. The behavior of many compounds in soils has not been
examined. Among these are the entire classes of biphenylenes and azoxy-
benzenes. In other cases, the behavior in soils of only one or two representa-
tive compounds in a class has been investigated to a significant degree; for
example, hexachloro- and 1,2-dichlorobenzene in the chlorinated benzene
group, and 3,3',4,4'-tetrachloroazobenzene in the azobenzene group. Compared
with the study of chlorinated organic compounds, the brominated, iodinated,
and fluorinated species of these compounds have not been studied significantly.
Only polybrominated biphenyls (PBB's) were covered sufficiently in the litera-
ture for inclusion in this report. The fate of compounds with applications
as pesticides or herbicides has been studied extensively, whereas most other
toxic organic compounds generally have not. The highly toxic compound TCDD,
which has no known applications, is one notable exception.
It may not be necessary, however, to study each compound in depth to
draw conclusions about its probable fate in soils. Development of octanol/water
partition coefficients (KQw) plus the available information on the effects of
the degree of halogenation and water solubility may provide sufficient
information to predict (by comparison with well-studied compounds) the extent
to which a compound will adsorb to soil organic matter. Knowledge of the
degree of adsorption of a compound to soil organic matter will give strong
clues as to the fate and persistence (and thus the half-life) of a particular
compound.
A major difficulty exists in interpretation of half-life estimates.
Because experiments are conducted under nonstandard conditions, a comparison
of results between different compounds tested in different climates and
different soils is not highly meaningful. If several standard sets of condi-
tions were adopted and half-life estimates were developed for a well-studied
93
-------
compound (e.g., DOT) under each set of conditions, other compounds could also
be studied under these standard conditions. In much the same way that specific
gravity values for all compounds are related to the density of water at a
specified temperature, half-life estimates generated for any test compound
could be related to the behavior of the standard compound under the same
conditions as the test compound. Relative half-life estimates could be used
to generate more meaningful data than currently exist about the similarities
and differences among compounds of interest.
94
-------
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