PB85-214310
DIOXIN TRANSPORT FROM CONTAMINATED SITES TO EXPOSURE LOCATIONS:
A METHODOLOGY FOR CALCULATING CONVERSION FACTORS
Office of Hazardous Waste Management
Richland, WA
Jun 85
U.S. DEPARTMENT OF COMMERCE
National Technical Information Service
NTIS
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EE85-21U310
(EPA/600/8-85/012
June 1985'
OIOXIN TRANSPORT FROM CONTAMINATED SITES TO EXPOSURE LOCATIONS
A METHODOLOGY FOR CALCULATING CONVERSION FACTORS
Prepared by
Gaynor W. Dawson
Ji'11 M. Meuser
Mary C. Lilga
Battelle Project Management Division
Office of Hazardous Waste Management
Richland, MA 99352
EPA Project Officer
John Schaum
EPA Contract 68-01-6861
OFFICE OF HEALTH AND ENVIRONMENTAL ASSESSMENT
OFFICE OF RESEARCH AND DEVELOPMENT
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON,.DC 20460
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TECHNICAL REPORT DATA
(Please read Instructions on the reverse before completing)
1. REPORT NO.
EPA/600/8-85/012
2.
3. RECIPIENT'S ACCESSION NO.
5 27. >o 10 /AS
4. TITLE AND SUBTITLE
Dioxin Transport from Contaminated Sites To Exposure
Locations: A Methodology for Calculating Conversion
Factors
5. REPORT DATE
June 1985
6. PERFORMING ORGANIZATION CODE
7. AUTHOR(S)
Gaynor W. Dawson
Jill M. Meuser
Mary C. Lilga
8. PERFORMING ORGANIZATION REPORT NO.
B556-21101
9. PERFORMING ORGANIZATION NAME AND ADDRESS
Office of Hazardous Waste Management
Battelle Project Management Division
601 Williams Blvd.
Richland, WA 99352
10. PROGRAM ELEMENT NO.
1 1
CT/G RANT NO.
Subcontract EPA 38-9
Unrk flcginnmonl" 11
12. SPONSORING AGENCY NAME AND ADDRESS
Exposure Assessment Group
US EPA, Office of Research and Development
Washington, DC 20460 •
13. TYPE OF RETORT AND PERIOD COVERED
Final 7/1/84 - 1/15/85
14. SPONSORING AGENCY CODE
EPA/600/21
16. SUPPLEMENTARY NOTES
16. ABSTRAC
Procedures have been developed by the US EPA for estimating the risk
associated with exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin (dioxin).
Cdncentrations of dioxin at the contaminant source are usually known, but
exposure may occur at locations away from the source where concentrations are
usually unknown. As a result, a need was identified for estimating dioxin
concentrations away from the source.
This report discusses the transport of dioxin from a source and presents
methods for estimating dioxin concentrations at potential points of exposure
away from a source. The transport pathways that were considered to be important
were volatilization, suspension and depositon of windblown particles, overland
sediment runoff, and in-stream sediment transport. Concentrations at locations
away from a source can be estimated using conversion factors for air, soil, and
sediment. Concentrations in these media at potential points of exposure can be
estimated using the source concentration and factors that describe the physical
characteristics of the source and the transport pathways.
Because ingestion of contaminated foodstuffs will result in exposure to
dioxin, an example is provided for estimating the amount of dioxin in beef.
Missouri beef distribution patterns and a market dilution concept were used to
estimate potential chronic exposure to contaminated beef products within the
17.
KEY WORDS AND DOCUMENT ANALYSIS
DESCRIPTORS
b.lDENTIFIERS/OPEN ENDED TERMS C. COS AT I Field/Group
2,3,7,8-Tetrachlorodibenzo-p-dioxin (Dioxii
Contaminant Transport
Conversion Factors
Environmental Contamination
18. DISTRIBUTION STATEMENT
Distribute to Public
19. SECURITY CLASS /This Report/
Unclassified
21. NO. OF PAGES
89
20. SECURITY CLASS (This page 1
Unclassified
22. PRICE
EPA Form 2220-1 (R«». 4-77) PREVIOUS EDITION is OBSOLETE
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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 conmercial products does not constitute endorsement or
recommendation for use.
Reproduced from
best available copy.
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TABLE OF CONTENTS
Page
Tables and Figures iv
Foreword v
Abstract vi
Acknowledgements vii
1.0 Introduction 1
2.0 Summary 8
3.0 Atmospheric Concentrations 13
3.1 Introduction . 13
3.2 Particulate Exposure ..... 13
3.3 Gaussian Dispersion . . 22
3.4 Vapor Exposure . . '. 24
4.0 Soil Concentrations ....... 27
4.1 Introduction . . . . 27
4.2 Oioxin Behavior in Soil 27
4.3 Photodegradation 28
4.4 Volatilization 30
4.5 Approach 30
4.6 Average Soil Losses. . . 33
4.7 Soil Deposition 41
4.8 Wind Deposition . 46
5.0 Sediment Concentrations 49
5.1 Introduction . . . 49
5.2 Dioxin Behavior in Water 49
5.3 Sediment Transport in Streams. 50
6.0 Missouri Beef Distribution Patterns 56
References 63
Appendix A - Example Site 1 " 68
Appendix B - Example Site 2 74
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TABLES AND FIGURES
Table 1. Physical Properties of 2,3,7,8-Tetrachlorodibenzo-
p-dioxin 2
Table 2. Average K Values for Soils on Erosion
Research Stations 37
Table 3. General Magnitude of the Soil/Erodibility Factor,
K, When Organic Content Data are Available 38
Table 4. Values of the Topographic Factor, LS, for Specific
Combinations of Slope Length and Steepness 39
Figure 1. Dioxin Structure 1
Figure 2. Pathways- for Exposure from Contaminated Soil 4
Figure 3. Relationship Between Conversion Factors . . 8
Figure 4. The Product of ayaz as a Function of
Downwind Distance from the Source for each
of the Six Stability Classes 10
Figure 5. Relative Resuspension Factors Under Various
Site Conditions 16
Figure 6. Average Annual Values of the Rainfall
Erosion Index 35
Figure 7. Generalized Relationship Between Size of
Drainage Basin Area and Sediment Delivery Ratio .... 43
Figure 8. Origin of Beef Consumed in Missouri 60
Figure A-l. Example Site 1 69
Figure B-l. Example Site 2 75
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FOREWORD
The Exposure Assessment Group (EAG) of EPA's Office of Research and
Development 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.
In relation to the third function, EAG sponsors projects aimed at developing
or refining techniques used in exposure assessments. This study is one of
these projects and was done for the Office of Solid Waste and Emergency
Response.
Dioxin problems first surfaced in the U.S. in the early 1970's with
Agent Orange and the Missouri Horse Arenas. Since then, dioxin contamination
has been found elsewhere in Missouri, Arkansas, Michigan, New York, and New
Jersey. EPA has become increasingly involved in the discovery, assessment,
and clean-up of these sites. The purpose of this document is to provide
methods to use in conducting exposure and risk assessments of dioxin
contamination sites.
James W. Falco, Director
Exposure Assessment Group
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ABSTRACT
Procedures have been developed by the U.S. EPA for estimating the risk
associated with exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin (dioxin).
Concentrations of dioxin at the contaminant source are usually known, but
exposure may occur at locations away from the source where concentrations are
usually unknown. In response to this problem,, a need was identified for
estimating dioxin concentrations away from the source.
This report discusses the transport of dioxin from a source and presents
methods for estimating dioxin concentrations at potential points of exposure
away from a source. The transport pathways that were considered to be
important were volatilization; suspension and deposition of windblown
particles; overland sediment runoff; and in-stream sediment transport.
Concentrations at locations away from a source can be estimated using
conversion factors for air, soil, and sediment. Concentrations in these
media at potential points of exposure can be estimated using the source
concentration and factors that describe the physical characteristics of the
source and the transport pathways.
Because ingestion of contaminated foodstuffs results in exposure to
dioxin, the report includes an example of how to estimate the amount of
dioxin in beef- Missouri beef distributic- patterns and a market dilution
concept were used to estimate potential chronic exposure to contaminated beef
products within the state.
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ACKNOWLEDGEMENTS
The authors would like to thank the following individuals who provided
valuable comments during the peer review process:
Stuart M. Brown (CH2M HILL)
Jerald L. Schnoor (University of Iowa)
Louis J. Thibodeaux (Louisiana State University)
The guidance, aid, and encouragement of the Project Officer, John Schaum, are
gratefully acknowledged.
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DIOXIN TRANSPORT FROM CONTAMINATED SITES TO EXPOSURE LOCATIONS:
A METHODOLOGY FOR CALCULATING CONVERSION FACTORS
1.0 INTRODUCTION
A great deal of literature has been published recently concerning
2,3,7,8-tetrachlorodibenzo-p-dioxin. This compound is one extremely toxic
member of a class of compounds containing the basic dioxin nucleus
(Figure 1).
O'
Dioxin Nucleus _2(3,7.8-Tetr»chlorod1benzo-p-dioxin
Figure 1. Dioxin Structure
There are 75 possible chlorinated dioxins, including 22
tetrachlorodibenzo-p-dioxins. However, the 2,3,7,8-tetrachloro isomer is one
of the most toxic substances known. Throughout this report, the term
"dioxin" has been used to refer to the 2,3,7,8-tetrachloro isomer, the
properties of which are given in Table 1. Although other isomers may be
transported by- the mechanisms described, in this report, the other isomers
have different physical properties which may render inapplicable the
concentration relationships derived for the 2,3,7,8-tetrachloro isomer.
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Table 1. Physical Properties of 2,3,7,8-Tetrachlorodibenzo-p-dioxin
Esposito et Mabey et Freeman and Perkaw et
al.. 1980 a!.. 1981 Schroy. 1984 al.. 1980
Molecular Weight 322 322 — 322
Melting Point, °C 305 — — 303-305
Vapor Pressure at 25°C, , o c 7
mm Hg - 10'6 1.5xlO"9 lO^-NT'
Water Solubility (ug/L) 0.2 . 0.2 0.317 0.2
Octanol-Water . _
Partition Coefficient — 6.9x10° -- 1.38x10'
In response to the discovery of a growing number of dioxin-contaminated
sites, the Exposure Assessment Group within the U.S. Environmental Protec-
tion Agency Office of Research and Development has drafted procedures for
estimating the human health risk associated with these sites (Schaum, 1984).
The procedural algorithm developed for .calculating exposure is of the form:
Dioxin
Concentration
lifetime ln ^°^ at Y Conversion Contact Exposure Absorption
Average = Source x Factor x Rate Rate x Fraction
Exosure Bod Weiht x 70 r Lifetime
Average = ource acor ae
Exposure Body Weight x 70 yr Lifetime
The algorithm contains a conversion factor that relates contamination at a
point of exposure to contaminant levels at a site (e.g., conversion of a
dioxin concentration in soil to a downwind airborne dioxin concentration).
All conversion factors are based on dioxin concentrations in soil at the
primary contaminant source. The purpose "of this report is to describe and
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quantify the conversion factor. The other factors in, and use of, the
exposure algorithm are described by Schaum (1984).
Depending on exposure mechanisms, it is apparent that several conversion
factors must be considered. Exposure to contaminants at a point away from
the source may occur as a result of contaminant transport to receptors by a
number of potential routes, as illustrated in Figure 2 for a generic site
contaminant. Due to the intrinsic properties of dioxin, there is limited
pdtential for transport in the dissolved phase. Thus, certain pathways in
Figure 2 may be ignored. While dioxin may be transported by these routes,
the concentrations at any given time would be low and would not result in
high exposure risks.
The pathways of primary interest are those associated with the transport
of solid particles containing adsorbed dioxin. The following mechanisms are
*•
considered to be important in the transport of dioxin from a site:
• resuspension and deposition of windblown particles;
• sediment runoff; and
t sediment transport in streams.
In addition, recent literature suggests that volatilization of dioxin from
contaminated soils may occur, despite the very low vapor pressure of the pure
compound (Thibodeaux, 1983; Freeman and Schroy, 1984). Consequently, vapor
transport is a fourth mechanism that must be considered when evaluating the
transport of dioxin from a source area.
Based on the considerations discussed above, the development of
conversion factors for use in the exposure algorithm was limited to five
cases as follow, where [dioxin] refers to concentration of dioxin:
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SOURCE AREA
CONTAMINATED SOIL
ATMOSPHERE
en
c
CT
c
Q>
u
in
O
Q.
SURFACE WATER
Figure 2. Pathways for Exposure from Contaminated Soils
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• [dioxin] in Air at Point of Exposure (uq/m^)
air " [dioxin] in Soil at Original Source (ug/Kg)
2\ £p ., _ [dioxin] in Soil at Point of Inqestion (ug/Kg)
50 [dioxin] in Soil at Original Source (ug/Kg)
3) CF -i - [diPX"*"] in Soil at Point of Contact (ug/Kg) ,.\
5011 [dioxin] in Soil at Original Source (ug/Kg) ( '
4) CF _ [dioxin] in Pasture Soils (ug/Kg) /5v
5011 " [dioxin] in Soil at Original Source (ug/Kg)
5\ CF [dioxin] in Sediment where Fish are Caught (uq/Kg) ,fiv
' sediment - [d1oxin] 1n Soi1 at Original Source (ug/Kg) { '
CF • is dependent on both particulate and vapor emissions from the
source. Section 3.0 describes the conversion factor for particulates and
discusses recent research concerning possible vapor emissions.
The second, third, and fourth factors (CFsoi-i) differ only in the mode
of exposure. The physical processes that transport contaminated soil from
the source to the point of exposure are the same in all three scenarios.
Thus, the CF -j-, for all three is the same. The treatment of the different
exposure routes is discussed in the predecessor paper on appropriate exposure
algorithms (Schaum, 1984). The derivation of CF$oil is discussed in Section
4.0.
The fifth factor, CF ^. ^, relates in-stream sediment concentrations
to source strength. It is dependent on runoff-derived particles that
accumulate in the stream bed. Estimation of CF .. t is discussed in
Section 5.0. A summary of the three factors (CFair, CFsoil, and CFsediment)
is provided in Section 2.0.
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Each of the concentration factors presented in this report is derived
from generalized ranges of environmental observations. The relationship
between anthropogenic, topographic, hydrologic, climatic, and vegetative
influences at a particular site will be very complex. However, consideration
of these influences generally requires computer modeling, and for some
influences the mathematical models do not exist.
The concentration factors developed in this report are based primarily
on average, gross transport. Airborne contaminant transport is assumed to be
downwind from a source. Areas of concern from overland transport will be
natural drainageways and topographic breaks. Finally, catastrophic events
such as hurricanes, tornadoes, and floods are not included in the derivation
of conversion factors, although such events may effect significant
environmental transport of contaminants. For example, Collier (1963)
reported that the sediment yield from a single-day storm exceeded 40% of the
yield for that year and exceeded the annual sediment yield for the previous
three years. Despite the potential importance of catastrophic events in
transporting contaminated material, the factors that describe the events can
only be considered for specific sites. These events are not easily reduced
to generic description due to their intensity, variability, and irregular
occurrence. When estimation of acute event transport is necessary, ASCE
(1975) should be consulted.
It must be emphasized that the conversion factor approach is a survey
method to provide rapid but approximate estimates of dioxin transport and its
implications. Ideally, empirical monitoring data or sophisticated numerical
modeling approaches would be employed for more quantitative estimates, as
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recommended by Schaum (1984). Conversion factors are useful to screen
candidate source sites and to help prioritize those sites for which more
quantitative estimates are needed.
The final section of this report is concerned with Missouri beef
distribution patterns. The discussion in Section 6.0 is intended to provide
a means for determining the extent to which Missouri inhabitants may be
exposed to dioxin through consumption of contaminated beef. The emphasis is
placed on developing a means- of estimating the fraction of beef consumed in
the state that is likely to have come from contaminated herds. Schaum (1984)
has developed a method for calculating dioxin concentrations in beef as a
function of soil levels in the pasture area. These concentrations are
determined by using the conversion factors (CF .,) derived in this work.
Hence, this report complements Schaum's work by providing a complete
methodology for determining, on a qualitative basis, possible dioxin exposure
at sites in Missouri.
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2.0 SUMMARY
Human exposure to dioxin may occur through contact with soil particles
on which dioxin is adsorbed. Contaminated soil may be transported from a
dioxin-contaminated source area by several mechanisms: suspension and
deposition of windblown particles; overland transport and deposition of .soil;
and in-stream sediment transport. The relationship between conversion
factors is illustrated in Figure 3.
Figure 3. Relationship Between Conversion Factors
The conversion factor for atmospheric concentrations of dioxin,
depending on wind speed, is determined by both the particulate and vapor
levels of dioxin, as follows:
CF . =-= (l x 1(T7 [W2(W"- 9)]}
Soil
(7),
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o
where C . = dioxin concentration in air above soil (ug/m )
a 11
W = average wind velocity (m/sec)
CSQ.-| = dioxin concentration in soil at point of interest "(
P
°"y°"z = product of the Gaussian dispersion coefficients (m ),
from Figure 4
F = fetch or downwind dimension of the source (m)
x = distance from source boundary to the point of interest (m)
If W < 10 m/sec, the expression reduces to:
CFair -f^ • l x 10~7
usoil
(8)
These expressions can be used to estimate atmospheric concentrations of
dioxin, Ca^r>. at the source if contaminated source conditions are used for
C .,. To determine atmospheric concentrations downwind from the source,
soil concentrations, either calculated from conversion factors or measured,
are used for C$oil. If Csource is used for C$oil, the degradation and
dispersion factors become unity.
Dioxin concentration in the surface soil in the vicinity of a
contaminated source is a function of the amount of sediment- delivered to the
point of exposure. The conversion factor for overland transport is:
PC Csoil Lsource x Asource ,Q,
CFsoil=p =-r — .(9)
source basin x basin
where CS01--| = (fioxin concentration in soil at point of interest (ug/Kg)
C = dioxin concentration in soil at the source (ug/Kg)
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Reproduced from ^f
best available copy, ^m
Figure 4. The Product of OyOz as a Function of Downwind
Distance from the Source (from Turner, 1970)
for each of- the Six Stability Classes A-F
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LSQurce = estimated soil loss from the , Universal Soil Loss Equation
(USLE) for the source (tons/acre/yr)
Asource = source area (acres)
Lbasin = estimated soil loss from the USLE for the watershed upstream
of the point of interest (tons/acre/yr)
Abasin = watershed area upstream of the point of interest (acres)
The conversion factor for windblown contributions to soil contamination is:
PF -, .
CFsoil =
-source
o
where ay0z = product of the Gaussian dispersion coefficients (m ),
from Figure 4
x = distance from source boundary to the point of interest (m)
F = fetch or downwind dimension of the source (m)
If runoff patterns and wind direction are coincident, the contributions to
soil contamination are additive and the expression for the conversion factor
becomes :
Csoi1 Lsource x Asource .
-
^source Lbasin x Abasin yzx+F
In order to determine the dioxin concentration on sediment delivered to
a stream in the vicinity of a contaminated source, the procedure is the same
as that used to determine surface soil concentrations. The conversion factor
for estimating stream sediment concentrations at distances downstream from
the source is:
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f I v A
rr sediment source source
u sediment = ~c = L . x A,
source basin basin
where Csediment = dioxin concentration in sediment at the point of
interest (ug/Kg)
s estimated USLE soil loss for source (tons/acre/yr)
Asource = source area
h>asin s estimatecl s01^ l°ss ^rom tne U^LE for the watershed
upstream of point of interest { tons/acre/yr)
\asin a watershed area upstream of the point of interest
(acres)
The conversion factors developed for the three exposure modes have
limitations. Because chronic, long-term exposure rates are of concern, the
•methods developed utilize average or estimated values for parameters such as
wind velocity, precipitation, runoff, soil erodibility, topography,
vegetative cover, and stream characteristics. Because of these
generalizations, the methods must be applied with caution, particularly in
areas such as the western United States where little experimental work on
sediment yields has been conducted.
Use of conversion factors is further limited by the absence of actual,
comprehensive site data with which to test the methods developed by this
study. Further site studies, designed to collect the necessary data on
dioxin concentrations and environmental characteristics at a source and at
potential points of exposure, will be necessary in order to verify these
methods.
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3.0 ATMOSPHERIC CONCENTRATIONS
3.1 Introduction
Human exposure to dioxin may result from inhalation of contaminated soil
particles.suspended in the atmosphere or vapors resulting from volatilization
from contaminated soil. In order to quantify exposure, it is necessary to
derive a means of estimating atmospheric dioxin levels in the environs of a
contaminated source.
3.2 Particulate Exposure
Particulate exposure results from the presence of dioxin-contaminated
soil in the atmosphere. Suspension of soil particles in the atmosphere may
result from the erosive action of wind or from activities which disturb the
soil, such as plowing or excavation. The concentration of particles in the
air column and their residence time are highly dependent on particle size and
atmospheric conditions. Ideally, the conversion factors should be derived
from studies in which' the soil size fractions were isolated, and in which the
concentration versus size fraction in the air column was known at various
distances from a source. Unfortunately, no data were found on particle size
distribution of dioxin-contaminated soils. However, data do exist for dioxin
concentrations on fly ash and larger ash particles caught in precipitators
(Fred C. Hart, 1984). These data indicate that dioxin levels may increase by
a factor of 2 to 12 on smaller size.particles compared to larger particles.
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If this relationship were found to be true for soils, it would provide some
guidance for adjusting airborne particulate level estimates based on soil
conditions. In the" absence of data to confirm increased concentrations on
small soil particles, selection of exposure factors has been based on the
assumption that concentrations of dioxin adsorbed on suspended particles will
be the same as dioxin concentrations in the bulk soil source. This
assumption, in turn, assumes that dioxin will be adsorbed on smaller-size
particles, which are subject to resuspension and respiration, at levels
comparable to concentrations in bulk soil. The assumptions conform to work
by Thibodeaux (1983), who found dioxin levels on dust in the air at the
Vertac site to be 1.1 ug/Kg compared to soil levels of 1.3 ug/Kg.
Conversion factors can be derived by deterministic or empirical
approaches. Deterministic approaches use numerical models to simulate the
basic phenomena involved. The models mathematically describe the physical
processes that effect transport. Many models have been developed to describe
resuspension and deposition; however, inadequate data are available to
validate these models (Sehmel, 1980).' For those models that do exist, the
large number of input requirements can be severely limiting. Sehmel (1980)
lists over 40" factors which influence resuspension, although the
relationships between these factors are not thoroughly understood. This
degree of complexity is too great for derivation of simple site
characteristic guides for predicting atmospheric contaminant or particulate
levels. Gillette (1973) employed a simplified relationship to describe
horizontal flux (Fh) of particulates:
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Fh = Ch W*(W*-Wt) (13)
where W* = wind shear velocity (m/sec)
Wt = a threshold velocity (m/sec)
Ch = empirical constant
However, because Ch is empirically derived, site-specific data are required
to calibrate the algorithm.
The empirical approach is based on correlation analysis of data from
actual sites, with subsequent selection of a factor that best matches the
relationship between soil contamination and atmospheric contamination levels.
As noted previously, this approach is required to calibrate Gillette's
simplified model. A single datum has been found for atmospheric levels of
dioxin at, near, or downwind from contaminated sites. Thibodeaux (1983)
reviewed monitoring data from the Vertac site and found atmospheric dust
concentrations of 54 ug/m , dioxin concentrations on the dust of 1.1 ug/Kg,
and soil dioxin levels of 1.3 ug/Kg over a 37 day period. These
o i
concentrations yield a CF,. of 4.6 x 10 Kg/m . Based on only this datum,
air
an empirical approach specific to dioxin cannot be determined at this time.
However, data are available on the relationship of particulate contaminant
levels and surface contaminant levels for a range of wind conditions and for
a variety of soil disturbing activities. These values are summarized in
Figure 5. The resuspension factors in Figure 5 were determined by comparing
atmospheric contamination (volumetric) to surface soil contamination (areal)
for particulate-based contaminants. The areal measurements were possible
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REFERENCE
STEWART, 1967
STEWART. 1967
STEWART, 1967
LANGHAM, 1971
STEWART. 1967
STtWART. 1967
STEWART, 1967
CALC. FROM MILHAM et al.
CALC. FROM MIUIAM ot al.
CALC. FROM MILHAM et al.
. , 1976
. , 1976
. , 1976
STEWART, 1967
STEWART, 1967
STEWART, 1967
STEWART. 1967
CALC. FROM BENNETT, 1976
CALC. FROM MILHAM et al., 1976
CALC. FROM IRANZO AND SALVADOR,
STEWART, 1967
CALC. FROM MILHAM et al., 1976
MILHAM et al., 1976
CALC. FROM MYERS et al., 1976
CALC. FROM MILHAM et a]., 1976
STEWART, 1967
ANSI'AUGH Pt al. , 1970
STEWAIIT, 1967
STEWART. 1967
STEWART, 1967
SEHMEL AND ORGII.L, 1973
CALC. FROM IRANZO AND SALVADOR,
LANGHAM, 1971
STEWART, 1967
CALC. FROM BENNETT, 1976
STEWART, 1967
SEHMEL AND LLOYI), 1976
CAIC. BY RENftETT, 1976
HAMILTON CALC. BY BENNETT, 1976
KRF.Y et al., 1976
ANSPAUGH et al., 1975
LOCATION
MARAL1NGA TRIALS
MARALINGA TRIALS
MARALINGA TRIALS
NEVADA TEST SITE
MARALINGA TRIALS
MONTI HELLO ISLANDS
MONTE BELLO ISLANDS
FIELD
FIELD
FIELD
MONTE BELLO ISLANDS
C. D. TRIALS
AUSTRALIAN DESERT
MARALINGA TRIALS
NEW YORK
FIELD
1970 PALMARES, SPAIN
AUSTRALIAN DESERT
FIELD
FIELD
SLUDGE
FIELD
MARALINGA TRIALS
NEVADA TEST SITE
SANDY-GRASS
SANDY-DEBRIS
MONTE BELIO ISLANDS
ROCKY FLATS
1970 PALMARES, SPAIN
NEVADA TEST SITE
PAVING STONES
HEW YORK-FALLOUT
SANDY-CLEARED
HANFORD
HEW YORK-U
UNITED KINGDOM-U
ROCKY FLATS
NEVADA TEST SITE
STRESS
STIRRED DUST
VEHICLE, 0.3 m
WALKING
VEHICULAR
CAB. LANDROVER
VEHICLE, 7TII DAY
VEHICLE, 4TH DAY
MOWING
PLANTING. DISKING
SUBSOIL ING
VEHICLE, 7TH DAY
WORK, OPEN
VEHICLE TAILBOARD
VEHICLE, 1-2 DAY
FALLOUT CONCS.
TRACTOR. DOWNWIND
YEARLY FARMING
WALKING
TRACTOR CAB
TRACTOR
ROTOTILLING
FERTILIZING
WIND
WIND
WIND
WIND
WIND
WIND
WIND
WIND
WIND
WIND
WIND
WIND
WIND
WIND
WIND
WIND
1ECHANICAL RESUSPENSIOM STRESSES
t-
MIND RESUSPENSIOH STRESSES
1 . . .1^ , , ,I..J . . .I.-J . . .1 I . . .1_J . . ,l.fl i I llalJ
10-10 ui-9 jo-8 10-7 i0-6 jo-b i0-4 10'
RESUSPEHSIOH FACTOR, m-1
10
-2
Figure 5. Relative Resuspension Factors* Under Various Site Conditions (After Sehmel, 1980)
*Resuspension Factor (RF) = Ratio of Contaminant per Volume Air to Contaminant Per Unit Area of Soil (nr* units)
-------
because radio-contaminant levels can be measured in situ without regard to
sample depth. For chemicals such as dioxin, concentrations would be
required.
An alternate method of calculating particulate levels in the atmosphere
is to compare atmospheric monitoring data to average soil levels for
conservative contaminants other than dioxin. Data were collected on
fluorides, chromium, copper, manganese, nickel, vanadium, and arsenic for
this approach (National Research Council, 1971; 1974a; 1973; 1975; 1974b;
Versar, 1976; Sullivan, 1969; and Nriagu, 1979, 1980). Atmospheric values
from rural areas were used to minimize the influence of anthropogenic
materials emitted from stacks. Values for each contaminant were combined to
derive a conversion factor, CF , which is defined as the ratio of atmospheric
A
concentrations (ug/m) to soil levels (ug/Kg). Assuming atmospheric
particulates are derived from nearby soils, the predicted average contaminant
level in ambient air at the site can be obtained by multiplying CF by the
A
average concentration of the contaminant in soil. Additionally, CF is an
/\
estimate of particulate levels in air (Kg/m ). The values for CFX determined
for each of.the seven contaminants are provided below:
Fluorides: CFF = °'02 to °'05 U9 F/m3 air (14)
F 20,000 to 500,000 ug F/Kg soil
= 3 x 10'6 to 4 x 10"8 Kg soil/m3 air
(Data from National Research Council, 1971)
-17-
-------
Chromium: CFr = °-01 U9 Cr/m air (15)
O 37,000 ug Cr/Kg soil v '
= 3 x 10"7 Kg soil/m3 air
(Data from National Research Council, 1974a)
Copper: CFr • °-005 to °-05 US Cu/m air ( .
™ Cu 20,000 ug Cu/Kg soil •
= 3 x 10'6 to 3 x 10'7 Kg soil/m3 air
(Data from Nriagu, 1979)
Manganese: CFMn = 0.08 ug Mn/m3 air (1 }
Mn 800,000 ug Mn/Kg soil
= 1 x 10"7 Kg soil/m3 air
(Data from National Research Council, 1973)
Nickel: CF,. = - 0.006 ug Ni/m3 air - ( }
Nl 30,000 to 80,000 ug Ni/Kg soil v '
= 2 x 10'7 to 8 x 10'8 Kg soil/m3 air
(Data from National Research Council 1975 and Nriagu, 1980)
Vanadium: CFV = 0.002 ug V/m3 air
v 200,000 ug V/Kg soil ^ '
= 1 x 10"7 Kg soil/m3 air
(Data from National Research Council, 1974b)
-18-
-------
Arsenic: CF.C = °-001 U9 As/m air ( }
As 5,000 ug As/Kg soil
= 2 x 10"7 Kg soil/m3 air
(Data from Versar, 1976 and Sullivan, 1969)
In general, CFX values are in the range of 1-3 x 10"7 Kg/m3. This range
is comparable to 100 ug/m , a level commonly found in polluted air and in
excess of the Federal Ambient Air Quality Standard of 75 ug/m3. If a typical
surface soil density of 1600 Kg/m3 and soil depth of 1.0 cm are assumed,
these CF¥ values would convert to resuspension factors (RF) of 1 x 10 Kg/m
A
x 1.6 x 10"3 m3/Kg x 1 x 102 ttf1 = 2 x 10"8 nf1. This value is consistent
with the work from which the values of RF were derived; in his work, Sehmel
(1980) determined that the tracer had mixed to a'depth of 1.0 cm. The value
O 1
-of 2 x 10 m corresponds to the RF values in Figure 5 for the lowest wind
or activity stresses. Hence, the CFX values derived from metal concentration
ratios agree with empirical data for particulates in general. Then, as a
rule of thumb: Cair in ug/m3 = (1 x 10 ~7) Csoji in ug/Kg. This
Q 1
relationship compares favorably with the 4.6 x 10 Kg/m value calculated
from the Thibodeaux (1983) data for the Vertac site. The relationship
represents average conditions over a year rather than those that would
prevail during storm events or in areas with high soil disruption activity
levels. The relationship addresses atmospheric levels at the perimeter of
the source. For downwind concentrations, the dispersion factor developed in
Sections 3.3 can be used, or CFai-r can be calculated based on soil dioxin
concentrations at the point of exposure (measured or calculated using CF .,
from Section 4.0).
-19-
-------
It has been observed that soil erosion is a function of wind velocity
cubed (Sehmel, 1980). More specifically, erosion is proportional to
W2(W -Wt) (Gillette, 1973) where W is wind velocity and Wt is a threshold
value of 6 - 13 m/sec. For calculations in wind, the median value (9) for
the threshold wind velocity was used and the relationship for determining
particulate concentration in air can be described as follows:
W < 10 m/sec: C. = 1 x 10"7 Cen<1 (21)
CFair = I x 10'7 Kg/m3
W > 10 m/sec: C = (1 x 10"7) W2 (W - 9) C (22)
a1r - soil
CFai> » (1 x 10'7) W2 (W - 9) Kg/m3
where Cai-r = dioxin concentration in air above soil (ug/m )
C .-, = dioxin concentration in soil at the point of interest (ug/Kg)
Because W is an average wind speed over time for chronic exposure, the second
relationship will not be required unless acute exposure calculations are
desired.
Mechanical disturbances can increase dust emissions significantly. For
dioxin levels that would occur during episodes of mechanical disturbance, the
relevant values in Figure 5 should be employed to increase CF .
air
proportionally to the ratio of the relative resuspension factors for the
Q 1
disturbed state and the calm state (RF = 10 m" ), respectively. Hence, for
-20-
-------
tractor use (RF = 10'7 to 10"6 nf1), CF . would be 1 to 2 orders of
<11 i
magnitude higher, or CFair = 1 x 10"5 to 1 x 10"6 Kg/m3.
This approach is based on calculation of atmospheric levels as a
function of the soil at the point of interest and does not account for
dispersion downwind. Thus, for downwind areas, C ••• is not the original
concentration at the source but the concentration at the point of exposure.
If a measured value is not available, an estimated value is employed using
the methods described in Section 4.0. This approach was taken because
atmospheric levels are believed to arise primarily from resuspension of local
soils (Sehmel, 1980). This assumption further assumes that activity levels
at the point of exposure.are greater than or equal to activity levels at the
source. More complex methods for rapid assessment of particulate emissions
have been summarized by MRI (1984) and are recommended if time and resources
allow.
Sites downwind from the original source will reflect dioxin levels in
soil resulting from all transport mechanisms, runoff and wind. Further,
atmospheric levels for chronic exposure will be a function of average wind
conditions, i.e., wind speeds of about 5.5 m/sec (12 mi/hr) (Versar, 1983).
Most wind-transported dioxin will have arrived from discrete storm events,
therefore, the conversion factor CFa. for W > 10 m/sec is required to
air ~
determine downwind soil levels as described in Section 4.0. These
atmospheric deposition contributions are combined with runoff input, also
described in Section 4.0, to yield an estimate of total average soil
concentration at the downwind point of exposure. The conversion factor CF .
a i r
is then applied to calculate atmospheric levels, as follows:
-21-
-------
Cair ' ("air) ("=„„)
where C,,-,, = dioxin concentration in air above soil (ug/m )
a 1 1
Csoj-| = dioxin concentration in soil at the point of interest (ug/Kg)
3.3 Gaussian Dispersion
The conversion factor derivation described in Section 3.2 applies only
to the air directly above a point of interest. Atmospheric concentrations
downwind from a source can be estimated without knowing soil concentrations
if the airborne particulate plume is considered 'to be subject to Gaussian
dispersion. In this case, atmospheric concentrations, C . , at point x can
be calculated using the following formula (Turner, 1970):
Q2
air(x,o,o) = TTO a u ^
y zw
where Cai-r = concentration in air (g/m3)
Q2 =• emission rate (g/sec)
r)
ayaz = ProdlJCt °^ tne Gaussian dispersion coefficients (m ),
from Figure 4
W = wind speed (m/sec)
Therefore, the ratio of concentrations at two points can be determined by the
ratio:
Cairl _ (gyqz)2 ....
Cair2" l
-22-
-------
If Ca-jr2 is set at the boundary of the source, it can be equated to
Therefore,
CF (aya2
where x = distance from source boundary to the point of interest
F = fetch or downwind dimension of the source
and CX+F = CF yz F (27)
\y z'x+F
The approach described by equations 24, 25, 26, and 27 assumes
conservation of mass.
Values for a a downwind from a source area are shown in Figure 4 for
the following six stability classifications:
A = Extremely Unstable
B = Moderately Unstable
C = Slightly Unstable
D = Neutral, considered to be representative of average, long-term
conditions
E = Slightly Stable
F = Moderately Stable
The combined conversion factors that describe particulate dioxin
concentrations downwind from a source are as follows:
-23-
-------
If W < 10 m/sec:
CFair = CFparticulate = l x 10
If W > 10 m/sec:
-7
'x+F
Kg/trT
(8)
"7
CFair ' ^articulate = l * 10" w
-------
1984; Thibodeaiix, 1983), with ultimate photolytic decomposition (Nash and
Beall, 1980).
In his study at the Vertac site, Thibodeaux (1983) calculated that
vaporization of dioxin from the soil surface was the major route of dioxin
loss from the site. Mass flux calculations, based on estimated values for
pertinent environmental and chemical properties, predicted that vaporization
losses from the site were much greater than losses from entrainment of soil
particles.
Nash and Beall (1980) reported that dioxin volatilized from soil in
microecosystem chambers and from field plots. Significant quantities of
dioxin in the air from both experiments appeared to be dechlorinated. The
researchers concluded that" atmospheric photodegradation was occurring. The
rates of both volatilization and degradation depended on the dioxin
application formulation and the temperature of the systems.
Freeman and Schroy (1984) used vaporization processes to model dioxin
movement in a soil column. However, the researchers suggest that
photodegradation at the soil surface will dominate vaporization losses during
daylight hours. Thus dioxin losses to the atmosphere should occur primarily
at night, with rapid photodegradation the next morning.
No research has been conducted to date on atmospheric degradation of
dioxin. Research results suggest that dioxin is lost from the soil, but the
loss mechanism and environmental fate are only poorly understood. Although
volatilization may be an important loss mechanism, potential photodegradation
may reduce any environmental transport. If the degradation process is
-25-
-------
occurring, as postulated by Nash and Seal! (1980) or Freeman and Schroy
(1984), the potential off-site exposure to vapor will be very low.
Due to the uncertainties in volatilization and photodegradation, it is
not possible to derive a conversion factor for dioxin vapor air
concentration. As more research is conducted, derivation of a conversion
factor for vapor (CFvapor) may be possible. This CFy r should be added to
the CFa-r presented in this report to estimate total atmospheric dioxin
concentrations at potential exposure points.
-26-
-------
4.0 SOIL CONCENTRATIONS
4.1 Introduction
Overland transport of contaminated soil via runoff is an important
mechanism which contributes to the potential for human exposure to dioxin.
Human activities such as farming, gardening, excavating, and recreation can
result in dermal absorption of contaminants or ingestion of contaminated
soil, particularly by. children. In order to estimate potential exposures
downflow from a source, an approximation of the soil loss from a source and
the redeposit.ion of contaminated soil away from a source must be calculated.
4.2 Dioxin Behavior in Soil
Soil at a source becomes contaminated by adsorption of dioxin. No data
were found to quantitatively describe dioxin concentration versus soil
characteristics, such as particle size or organic content. However, it is
assumed that due to its high KQW, dioxin will be adsorbed primarily on the
organic fraction of the soil. It is this high affinity for organics in soil
and low solubility in water that are believed to account for the vertical
immobility of dioxin (Kearney, Woolson, and Ellington, 1972; Matsumura and
Benezet, 1973). Because small particles have a higher surface-to-volume
ratio than large particles, it is also assumed that the small particle-size
fraction of the soil would have a higher contaminant concentration than a
bulk soil sample. Walling (1983) summarizes the relationship of particle
-27-
-------
size and organic content characteristics of eroded soil to those of the
original soil in five test-plot studies. These data suggest that
contaminants such as nutrients or pesticides may be enriched up to 1.5 times
on clay-sized particles, and more than 2 times on the organic fraction. Lack
of quantitative data concerning these phenomena, however, precludes
incorporating them into the derivation of conversion factors. It has been
assumed, therefore, that all transported soil has the same dioxin
concentration as determined for bulk soil samples from the site.
4.3 Photodegradation
Photodegradation is another process which may affect the amount of
dioxin available for transport from a site and the amount to which humans may
be exposed. Ultraviolet wavelengths have been shown to be effective in
photodegrading dioxins. Photolysis apparently removes one or more chlorine
atoms from the dioxin molecule, thereby making it less toxic but not
destroying the basic dioxin nucleus (Crosby et al., 1971).
Esposito et al., (1980) provide a comprehensive review of numerous
Photodegradation studies and the inconsistent results. Crosby et al., (1971)
applied dioxin to several matrices. Although decomposition was rapid in
alcohol solution, there was negligible loss from aqueous suspension and wet
or dry soil after 96 hr of irradiation. However, the researchers suggested
that in the natural environment, waxy leaf cuticles, surface slicks on water,
and spray oils or solvents commonly incorporated in pesticide formulations
may serve as the organic hydrogen donors necessary for Photodegradation.
Dioxin applied to soil and exposed to artificial sunlight (sunlamp) for
96 hr showed no degradation, as reported by Crosby, Moilanen, and Wong
-28-
-------
(1973). In other studies (Crosby and Wong, 1977), dioxin-contaminated
Herbicide Orange was applied to plant leaves and soil and exposed to
sunlight. After 6 hr of exposure, 0-30% of the dioxin remained on the plant
leaves, with 30% remaining on soil which had received the lowest applied
n O
concentration (1.3 ng/cnr). At the application rate of 10 mg/cm
approximately 90% of the dioxin remained after 6 hr of exposure. The
researchers believed that surface soil particles shaded the underlying
particles, thereby preventing photodecomposition at depth. It was concluded
from the 1977 study that the three requirements for dioxin
decomposition/dechlorination are: 1) dissolution in light-transmitting film;
2) the presence of an organic hydrogen donor, such as solvent or pesticide;
and 3) ultraviolet light. All three conditions should be present in the
application or accidental loss of materials commonly contaminated with dioxin
f
(2,4,5-T, trichlorophenol, PCB road oils). Crosby and Wong conclude that
dioxin is not stable as a contaminant in thin herbicide films exposed to
outdoor light.
In response to the work by Crosby and co-workers, photodegradation was
evaluated as a decontamination technique in Seveso, Italy (Liberti et al.,
1978). Exposure of dioxin-contaminated soils to artificial ultraviolet light
and natural sunlight in the presence of a hydrogen donor resulted in
degradation at the surface and to a certain extent, degradation beneath the
soil surface. The degradation rate in soil from natural sunlight would be
affected by sunlight intensity, nature of the contaminated medium,
temperature, and the amount of vegetative cover at a site.
-29-
-------
Although photodegradation may have a significant effect on environmental
dioxin concentrations, it was not included in deriving soil conversion
factors. The amount of degradation appears to depend on site-specific
factors (sunlight intensity, temperature, substrate, cq-contaminants) that
.are not amenable to a generic approach. If photodegradation is occurring,
the method presented in this report will overestimate the dioxin
concentrations at the point of exposure.
4.4 Volatilization
As discussed in Chapter 3.0, volatilization from a site may be a
significant loss mechanism. Nash and Beall (1980) and fhibodeaux (1983)
report that volatilization may be a major pathway. Matsumura and Ward (1976)
indicate that the water content of soil may mediate the evaporation rate.
The effects of volatilization on concentrations at a site and at points of
exposure have not been considered in deriving CFsoii, so overestimation of
soil exposure concentrations may result when using this method.
4.5 Approach
Empirical and deterministic methods were evaluated for applicability in
deriving conversion factors. The empirical approach involved collecting.
monitoring data from specific sites and trying to correlate observed
distributions with characteristic site parameters. Because it was expected
that few comprehensive sets of dioxin data would be available, monitoring
data for other persistent contaminants such as polychlorinated biphenyls
(PCBs), polybrominated biphenyls (PBBs), heavy metals, and radionuclides were
also sought.
-30-
-------
Only two sites were found where sampling had been conducted at and away
from a site. Roberts, Cherry, and Schwartz (1982) studied the distribution
and surface translocation of a serious PCB spill at a transformer
manufacturing plant in Regina, Saskatchewan, Canada. The researchers found
that particle transport in runoff from eroding areas was an . important
migration mechanism. However, PCB distributions were "extremely
heterogeneous," with "no definable trends in concentrations."
Dioxin contamination in Seveso, Italy, was caused by wind-influenced
atmospheric deposition from the 1976 explosion at the Givaudan-LaRouche
ICMESA plant. Sampling was conducted within 110 hectares southeast of the
site for over 3 years (DiOomenico et al., 1980). Dioxin concentrations at
locations 100 m apart varied by as much as a factor of 100, and this highly
irregular distribution changed very little during the three-year study
period. Although the mechanism by which dioxin was initially distributed
differs from that characteristic of uncontrolled disposal sites, the method
of transport from the originally contaminated area is similar to the problem
addressed in this report. Of particular interest are the following:
• Areas of high contamination showed little dioxin contamination reduction
over three years.
• Slightly contaminated or uncontaminated areas downwind and within runoff
routes showed no statistically-significant increase in dioxin
concentrations over three years.
Assessment activities at dioxin-contaminated sites in Missouri and
Arkansas did not include systematic sampling/analysis at the sites and at
intervals away from the sites. Sampling was not conducted at known high
-31-
-------
concentration sources scheduled for remedial action at the Vertac site (JRB,
Inc., 1983). Sampling was conducted at numerous sites in Missouri, but only
to locate areas of high concentration. Few samples were acquired from each
site and no descriptions of site or pathway characteristics were provided
(U.S. EPA, 1982a; U.S. EPA, 1982b).
Based on available site data, it was not possible to derive correlations
between dioxin concentrations at sources and at points of exposure. The
Regina and Seveso data provide qualitative indications that contaminant
concentration distributions will probably be irregular and thus difficult to
predict. Belli et al., (1983) report that the statistical analysis of data
from regions of low contamination at Seveso was most strongly affected by the
sensitivity and precision of analytical instrumentation.
Deterministic approaches involve mathematical modeling of the physical
transport process from a site to a selected exposure point. Onishi, Whelan,
and Skaggs (1982) present a review of overland soil and sediment transport
models and divide them into three groups based on their degree of complexity
and the extent to which they represent physical processes.
The simplest models require the least amount of site-specific data and
use an empirical formula to estimate average soil losses from an area. A
modified version of the Wischmeier and Smith (1978) USLE requires limited
data on watershed characteristics.
The second group of models requires considerable amounts of detailed
hydrologic, meteorologic, and site-specific physical characteristics to
simulate soil erosion and transport. If the required data are available,
these models are generally more accurate than the simplest, empirical models.
-32-
-------
Both the first and second model groups account for chemical distributions via
loading factors.
The final group of models simulates environmental chemical behavior,
such as adsorption-desorption and decay, as well as runoff and erosion. In
addition to the data required for Group 2 models, the most complex models
require chemical characteristics and distributions on the land surface.
The last two groups of models are useful only to those who have access
to a digital computer, and are therefore of no interest to those wishing to
calculate simple conversion factors. In addition, these models require
detailed site and chemical data that are generally not available without
extensive field investigations.
The model used to derive conversion factors from average soil loss and
deposition is consistent with the first group of models and does not require
•>
a computer or cumbersome amounts of site-specific data. Because chronic
exposure is of primary interest, average loss and deposition are appropriate
for assessing lifetime exposure, rather than for cyclic or acute events. The
»
approach utilizes the USLE.
4.6 Average Soil Losses
Average annual soil losses from a contaminated site can be approximated
by using the USLE, .an empirical formula which was developed for agricultural
land using data from numerous field test soil plots (Wischmeier and Smith,
1978). The equation input factors have been modified slightly for use in
areas other thjan cropland. The USLE provides an approximation of sheet and
rill erosion losses, in tons per acre per year, due to the interaction of six
physical factors which can be expressed numerically as site characteristics.
-33-
-------
Tables and maps are provided for use in selecting site-specific values for
these factors.
The USLE defines loss as:
L=RxKxLSxCxP (28)
where L = computed average annual soil loss (tons/acre/yr)
R = rainfall erosion index (yr"1)
K = soil credibility factor (tons/acre per unit of rainfall factor,
. R) . .
LS = topographic factor (dimensionless)
C = cover and management factor (dimensionless)
P = support practice factor (dimensionless)
Useful procedures for the estimation of USLE parameters for both agricultural
and non-agricultural conditions can be found in Mills et al., (1982).
The average annual soil loss per unit area, L-, represents an average
annual value and is obtained by multiplying the rainfall erosion index (which
provides estimated soil losses due to rainfall and runoff for a geographic
area) by a series of ratios. These ratios represent the relationship of
actual parameters to those observed in test soils and standardized
agricultural plots.
The rainfall erosion index, R, expresses erosion potential for average
annual rainfall at a location. A map of average R values for the U.S., based
on over 30 yrs of measurements, is provided in Figure 6. Interpolation
between contour lines is necessary for many areas of the country.
-34-
-------
e. 5
°
O 3
I
CO
W.H. WlICIWMiw. SEA. 1976
,t>
Figure 6. Average Annual Values of the •Rainfall Erosion
Index (yr~l) (from Wischmeier and Smith, 1978)
-------
Values for K, the soil erodibility factor, have been experimentally
determined for a number of benchmark soils at erosion research stations in
the U.S. Average values of K, based on a range of soil types, are provided
in Tables 2 and 3. The soil erodibility for a particular site can be
approximated by using the K value corresponding to the predominant soil type.
Average values for basic soil types are provided in Table 2. Assuming soil
organic content is known or can be estimated, more specific values for K are
available in Table 3.
The topographic factor, LS, combines the effect of slope length and
steepness. Values for the area under consideration can be determined using
the average percent slope and slope length, measured in ft. A.listing of LS
values for slopes of varying gradients and lengths is provided in Table 4.
Interpolation between listed values may be necessary.
The cover and management factor, C, is most significant for agricultural
land where it is a" function of vegetative cover, crop sequence, crop
rotation, and tilling practices. Wischmeier and Smith (1978) provide
guidelines for determining C values for construction sites, pasture, range,
idle land, and forested areas. In order to simplify site characterization,
two C values have been selected. A C value of 1.0 represents a worst-case
scenario and should be used when vegetation is completely absent. Examples
of this type of site would be horse arenas, unpaved roads, and unvegetated
landfills. A C value of 0.5 should be used for any other type of site.
Because 0.5 represents a high value for permanent pasture, range, wooded, and
idle land, a worst-case scenario for vegetated land has been assumed. For
wooded areas with highly erodible soil and no surface vegetative cover, the C
-36-
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Table 2. Average Values for the Soil Erodibility Factor,
K, for Soils on Erosion Research Stations
(After Wischmeier and Smith, 1978)
Average K Value
Soil Type (tons/acre)
Silt Loam
Loam 0.4
Sandy Clay Loam
Silty Clay Loam
Clay 0.3
Clay Loam
Fine Sandy Loam 0.2
Loamy Sand
Flaggy Silt Loam 0.1
Gravelly Loam <0.1
-37r
-------
Table 3. General Magnitude of the Soil/Erodibility Factor,
K*, when Organic Content Data are Available
(Carsel et al., 1984)
Organic Matter Content
Texture Class
Sand
Fine Sand
Very Fine Sand
Loamy Sand
Loamy Fine Sand
Loamy Very Fine Sand
Sandy Loam
Fine Sandy Loam
Very Fine Sandy Loam
Loam
Silt Loam
Silt .
Sandy Clay Loam
Clay Loam
Silty Clay Loam
Sandy Clay
Silty Clay
Clay
<0.5%
0.05
.16
.42
.12
.24
.44
.27
.35
.47
.38
.48
.60
.27
.28
.37
.14
.25
2%
0.03
.14
.36
.10
.20
.38
.24
.30
.41
.34
.42
.52
.25
.25
.32
.13
.23
0.13-0.29
4%
0.02
.10
.28
' .08
.16
.30
.19
.24
.33
.29
.33
.42
.21
.21
.26
.12
.19
*The values" shown are estimated averages of broad ranges of specifiw-soil
values. When a texture is near the borderline of two texture classes, use
the average of the two K values. For specific soils, Soil Conservation
Service K-value tables will provide much greater accuracy.
-38-
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Table 4. Values of the Topographic Factor, LS, for Specific
Combinations of Slope Length and Steepness (From
Wischmeier and Smith, 1978)
Nrctn'
t-apt
0.2 ...
• 0,5
0.8
2
3 ....
4
3
6
8 ...
10
12
14
16
18
20
Slop* lingth ;f»»t:
23
. . 0.060
.073
.086
133
190
230
.268
. .336
496
685
903
.... 1.15
. ... 1.42
... 1 .72
. . . . 2.04
50
0.069
.083
.098
.163
.233
.303
.379
.476
.701
.968
1.28
I.i2
2.01
2.43
2 38
73
0.075
.090
.107
.185
.264
.357
.464
.583
859
119
1.56
1.99
2.46
2.97
353
100
0.380
.096.
.113
^201
.287
.400
.536
.673
.992
1.37
1.80
2.30
2.84
343
4.08
130
0.086
104
.123
.227
.325
'.471
.656
.824
1.21
1 68
2.21
2.81
3*8
4.21
5.00
200
0.092
.110
.130
.248
.354
.528
.758
.952
1.41
1.94
2.S5
3.25
4.01
3.86
5.77
300
0.099
.119
.141
.280
.400
.621
.928
1.17
1.72
2.37
3.13
3.98
4.92
5.95
7.07
400
0.105
.126
.149
.305
.437
.697
1.07
1.35
1.98
2.74
3.61
4.59
5.68
687
8.16
300
0.110
.132
.156
.326
.466
.762
1.20
1 50
2.22
3.06
4.04
5.13
6.35
7.68
9.12
600
0.114
.137
.162
.344
.492
.820
1.31
1.65
2.43
3.36
4.42
5.62
6.95
8.41
10.0
MO
O.!21
.143
.171
.376
.536
.920
1.52
1.90
2.81
3.37
5.11
6.49
8.03
9.71
11.5
1.000
0.126
.152
179
.402
.573
1 .01
1.69
2.13
3.14
4.33
5.71
7.26
8.98
10.9
12.9
: LS = ;V 72.6 ™ '65.41 »in: 6 4- 4.56 tin 8 - 0.065) whtr* X = slop* l«ngrh in f**t: m = 02 far
grod.tnn < 1 percent, 0.3 (or 1 to 3 p«rc«nt ilopn, 0.4 for 3.5 to 4.5 p*r:tnl tlopci, 0.5 for 5 pxrcint
ilopti and it»«p«r, and 6 = ongl* of llop«. (For oth«r combinationi of length and gradient, inttrpolatt
b«fw««n adjactnt valu« or M< fig. 4.1
Reproduced from jjpyt
best available copy.
-39-
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value would approach 1.0. Estimation of soil losses using the C values
discussed here would probably result in higher rates than would be actually
observed in the environment, but are deemed to be acceptable approximations.
The support practice factor, P, is also dependent on agricultural
techniques and is a function of such practices as contouring and terracing.
Because there is no counterpart to P on natural land or construction sites,
the value of P has been set at 1.
Users of the USLE must be aware of its limitations (Wischmeier, 1976;
Walters, 1983). Soil losses from a source or area can be determined by using
the USLE; however, the USLE provides only an estimate of the amount of soil
eroded from a specific area and does not indicate the amount of sediment
actually delivered to streams. The sediment yield is the total amount of
soil loss from the area less the amount of deposition which occurs.
R values obtained from Figure 6 (Wischmeier and Smith, 1978) are
applicable only for long-term erosion averages. Values for K, the soil-
erodibility factor {Tables 2 and 3), are averages for soil types, but the
actual amount of soil loss for any soil type can vary widely as a function of
antecedent soil moisture conditions. The amount of runoff will be
significantly different for saturated and unsaturated soils.
The USLE was developed primarily from data obtained east of the Rocky
Mountains, so its applicability to the arid western states may be somewhat
limited. Use of the USLE may result in significant errors due to the
predominance of high intensity, short duration rainfall in the West, and the
greater effect of other physical conditions such as wind, humidity, and heat.
-40-
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Because the USLE was developed using data from small field plots, it
predicts sediment yields of particles of 1-mm diameter (coarse sand) and
finer sediments. The USLE is not applicable to coarser sand, gravel, and
larger particles.
4.7 Soil Deposition
A Modified Universal Soil Loss Equation (MUSLE) (Simons, Li, and
Associates, 1982; Walters, 1983) has been developed for determining single-
.storm event sediment yields from drainage basins. The substitution of a
runoff factor for the rainfall factor, R, in the USLE makes the MUSLE better
suited for use in areas west of the Rocky Mountains. Use of the MUSLE,
however, requires calculating site-specific coefficients, which preclude its
general use for determining sediment yields. The MUSLE has been further
modified for computing annual sediment yield, but this calculation is also
site-specific based on weighted storm yields for selected return periods.
To determine sediment yield to a stream, the sediment delivery ratio is
used (Piest and Miller, 1975):
D = Y/L (29)
where D = sediment delivery ratio, the change per unit area of sediment
delivery downstream (dimensionless)
Y = sediment yield at measuring point (tons/acre/yr)
L = total amount of sediment eroded from drainage area .-"stream
of measuring point, estimated using the USLE (tons/acre/yr).
-41-
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To date there have been no comprehensive studies defining sediment
delivery relationships for the U.S. on a regional basis. It is impossible to
define relationships which would hold true for all geographic areas. Walling
(1983) states that the processes of sediment delivery are very complex and
are dependent on a variety of factors, including topography of the source
area, stream channel characteristics, drainage patterns, vegetative cover,
land use, soil properties, and the distribution of sediment sources.
Interrelationships between these factors are difficult to define, and errors
can be introduced because measured sediment yields are compared to total
erosion from a source estimated with a generalized soil loss equation.
As summarized by Walling (1983), there is evidence that only a small
percentage of the drainage basin area provides storm runoff in humid regions,
and the actual runoff area for the same delivery location varies in extent
and location depending on antecedent moisture conditions. This evidence
suggests that the sediment delivery ratio is dependent on only the
characteristics Of that portion of the drainage basin which produces storm
runoff and would change with time as the area changes.
Some relationships have been characterized sufficiently to show general
trends between the size of the drainage basin and the sediment delivery
ratio. Piest and Miller (1975) present a summary of this relationship in
Figure 7. Walling (1983) provides a curve showing the relationship between
the sediment delivery ratio and drainage basin area for the central and
eastern U.S., as developed by the U.S. Department of Agriculture Soil
Conservation Service. He also gives a summary of 10 relationships from
selected drainage basins in the U.S. and other countries as well as a summary
-42-
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£E
Cr .
r
r M
T
, HEOmLLS PM»S*OWWHC ARtA TElAb AWO OKLAHOMA
• WSSOvOi 8ASM LOESS r«.L5 • IO«A AX) FCBAASKA
» eLACKLANOMAIRCS. TEXAS
^ SANL- CtAV NLLS -MSSi^^Pi3!
"•>OUTH":.AI "•""•""i- " r'"'v "
i SMMOFiLDPLAn.LuNOlSINOT USED KOt TEHMK*TON OF O«v£)-
I I
Figure 7. Generalized Relationship Between Size of
Drainage Basin and Sediment Delivery Ratio
(from Piest and Miller, 1975)
Reproduced from
best available copy.
-43-
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of 13 sediment yield/drainage basin area relationships which are directly
analogous to the delivery ratio/drainage basin area relationship (Piest and
Miller, 1975; Walling, 1983). Some of these relationships show marked
similarities to the relationship in Figure 7.
The relationship shown in Figure 7 is generally applicable to the
central and eastern U.S. No comparable data could be found for the western
U.S. The sediment delivery ratios shown in Figure 7 vary widely for a given
drainage basin area. The values for basins often vary by a factor of 2, and
sometimes by an order of magnitude.
It is assumed that sediment delivery ratio is inversely proportional to
drainage basin size because of greater redeposition that will occur as
sediment travels over greater distances before reaching the point of
interest. It can be concluded by analogy that redeposition of contaminated
soil will become smaller as locations of interest are more distant from the
source site.
The concentration of dioxin in soil at a point of interest, x, is:
mass of dioxin delivered to x
mass of soil delivered to x
(mass of dioxin lost from source)(fraction delivered to x) /,,N
mass of soil delivered to x
The fraction delivered from the source to point x is the sediment
delivery ratio for an assumed watershed which begins at the source and
encompasses the natural drainage area between the source and the point of
interest. The mass of soil lost from the source can be estimated using the
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USLE and the source area. The total mass of soil delivered to point x can be
estimated using, the USLE, the watershed area, and the sediment delivery ratio
for the entire watershed area. Frequently, USLE estimates will be based on
the average slope and slope length for the watershed. The expression for
soil concentration, thus, becomes:
r source x Lsource x Asource x Psource ,,0x
Soil = [ T~A 75 ( '
Lbasin x Abasin x ubasin
where Ccnny,ra = dioxin concentration in soil at the source (ug/Kg)
SOUiCc
Lsource 3 estimated soil loss from the USLE for the source'
(tons/acre/yr)
Asource = source area (acres)
"source = sediment delivery ratio for the area between the source
and the point of interest (dimensionless)
Lbasin = estimated soil 1°ss ^rom tne USLE for the watershed
upstream of the point of interest (tons/acre/yr)
Abasin = watershed area upstream of the point of interest (acres)
Dbasin = sec'''ment delivery ratio for the watershed area upstream of
the point of interest (dimensionless)
From Figure 7, it appears that the sediment delivery ratio for a
specific basin area can range from about 0.1 or 0.2 to 1.0, depending on
factors other than basin area. On the other hand, the delivery ratio as a
function of basin size only varies over about the same range. Therefore, in
general, the sediment delivery ratio for the source and watershed basin areas
are not considered to be significantly different. If sediment delivery
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ratios at the point of interest are not known, C •-, can nevertheless be
estimated using a simplified form of equation 45:
P - P
soil " source
Lsource x source
basin x Abasin
(9)
In areas where the downgradient points of exposure lie in the downwind
direction from the site, it will be necessary to consider the effect of
atmospheric concentrations, as well as soil concentrations. Atmospheric
concentrations and the corresponding conversion factor were discussed in
Chapter 3.0.
4.8 Wind Deposition
The constructs for calculating Cair presented in Chapter 3.0 incorporate
the assumption that the soil concentration of dioxin (Csoil) is known at the
location of interest. If observed data are not available, a means of
calculating downwind soil concentrations arising from deposition during
previous storm events will be necessary.
For the purposes of this report, it is assumed that dioxin-contaminated
soils in downwind areas arise from windblown particulates. Although some
vapor transport may occur, this process was neglected due to the
uncertainties discussed in Section 3.4. If the transported particles of
interest are in the range £ 20 micron, they are subject to dilution as
predicted by a Gaussian distribution for the plume. Because respirable
particles are £ 10 micron, this assumption incorporates all particulates of
interest. Larger particles containing dioxin will settle more rapidly and
therefore reduce atmospheric dioxin levels. Hence, this assumption may lead
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to overprediction. In this case, atmospheric concentrations, Cai>, at point
x can be calculated using the following formula (Turner, 1970):
Qz
Cair(x,o,o) *~^-^ (24)
yV
where Qz = emission rate (gm/sec)
o" a = Product of the Gaussian dispersion coefficients (m ),
from Figure 4
W = wind speed (m/sec)
Therefore, the ratio of concentrations at two points can be determined by the
ratio:
^M = lVil2 • (25)
Lair2 lCTyVl
If the phenomena that relate atmospheric levels to soil concentrations
at a given location are essentially the same for all points in the downwind
direction (i.e., CFair = Ca1r/Cso1l), it holds that
Csoill (avaz)2
c = / V (33)
Lsoil2 (yz)i
Values for a az downwind from a source are given in Figure 4 for the six
stability classifications. Assuming that most particulate transport arises
from major storm events, the a az values for Stability Class A are most
appropriate for predicting downwind soil concentrations. Therefore, downwind
soil dioxin levels at distance x can be calculated using the following
relationship:
-47-
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CF
..
5011 Source «VzW
where F = fetch or downwind dimension of the source (m)
x = distance from source boundary to the point of interest (m)
p
ova, = product of the Gaussian dispersion coefficients (m ),
A • £ • .
from Figure 4.
When runoff patterns and prevailing wind direction are coincident, soil
concentrations should be based on the summation of the two contributions, as
follows:
Csoil _ Lsource x Asource
5011 " Csource " Lbasin x Abasin
This surface concentration value relates to the top centimeter of soil
and should be used for subsequent calculation of downwind dioxin-particulate
levels in the atmosphere, assuming that no soil mixing occurs prior to
resuspension.
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5.0 SEDIMENT CONCENTRATIONS
5.1 Introduction
Human health risks may arise when individuals are exposed to dioxin
through the consumption of dioxin-contaminated fish. Fish accumulate dioxin
from an aquatic environment in two ways (Isensee and Jones, 1975). Bottom
feeding species, in particular, may ingest contaminated sediment along with
their food. Any fish species can accumulate dioxin directly from the water.
In these cases, dioxin is desorbed from contaminated sediments or absorbed
from stream water. To quantify human exposure, it is necessary to develop a
means of approximating the concentrations of dioxin in stream sediments in
the vicinity of a contaminated site. The conversion factor described in this
report can be used in the algorithm developed by Schaum (1984) to estimate
the bioconcentration of dioxin in various fish species and the subsequent
human exposure.
5.2 Dioxin Behavior in Water
Some of the processes which can affect dioxin when it is exposed to air
(as discussed in Chapters 3.0 and 4.0) are expected to have minimal effects
on dioxin in an aquatic environment. Crosby et al. (1971) report that
volatilization does not appear to be of major importance in water. Other
researchers report that evaporation from or with water may be a major cause
of the disappearance of dioxin in. a model aquatic ecosystem (Ward and
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Matsumura, 1978), but the experimental results and their application to the
natural environment are as yet inconclusive. Photodegradation is thought to
occur so slowly as to be negligible in water (Crosby, Moilanen, and Wong,
1973; Crosby et al., 1971; Isensee and Jones, 1975; Matsumura and Ward,
1976). Some evidence of microbial degradation under experimental conditions
has been documented; however, dioxin in water is generally thought to resist
microbial effects (Ward and Matsumura, 1978; Matsumura and Ward, 1976). The
half-life of dioxin was found to be on the order of 600 days in a model
aquatic ecosystem (Ward and Matsumura, 1978).
When present at very low concentrations on sediment, dioxin is generally
not expected to desorb due to its low solubility. At concentrations as low
as 0.1 ppb, however, dioxin can desorb. Isensee and Jones (1975) report that
under experimental conditions the concentrations of dioxin in water and
sediment reached equilibrium in 4 to 15 days. The temporal variation was
attributed to the difference in adsorption capacities of the two soils used.
5.3 Sediment Transport in Streams
Contaminants can be transported in streams by three processes: 1) as
dissolved compounds in stream water, 2) as compounds adsorbed onto sediments
and transported as suspended load, and 3.) as compounds adsorbed onto
sediments and transported as bed load. The low solubility and high affinity
of dioxin for soils, particularly soils high in organic content (Isensee and
Jones, 1975; Kearney, Woolson, and Ellington, 1972), suggest that dioxin
would be transported primarily in ;? adsorbed phase on stream sediments. No
data could be found on preferential adsorption of dioxin to any particular
sediment particle size, so it was assumed that dioxin would be adsorbed
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equally on all available particle size fractions. Pritchard (1984) reports
that for polynuclear aromatic hydrocarbons, partition coefficients and
natural transport of sediments adequately accounted for the observed
distribution of the contaminants in an aquatic environment.
The source of stream sediment contamination in the vicinity of a
contaminated site is surface soil on which dioxin is adsorbed. Such
contaminated soil may reach a stream by direct stream erosion of the soil or,
more commonly, by overland sediment transport. The latter process is known
as sheet and rill erosion and occurs during runoff of precipitation. This
type of sediment transport was discussed in Chapter 4.0. Of the soil eroded
and transported by overland processes, some can be expected to reach both
major and minor streams within a drainage system.
Finer soil particles, such as clay, silt, and fine sand, that reach the
stream are usually transported as part of the suspended sediment load.
Coarser particles, such as coarse sand and gravel, are usually transported as
bed sediment load or are deposited in the stream bed.
Most streams normally flow at less than their capacity. This normal
flow is called the mean annual discharge and corresponds to a water depth of
only about one-third of the stream capacity or bank-full depth. Mean annual
discharge is equaled or exceeded on an average of 25% of the time (Leopold,
Wolman, and Miller, 1964).
For streams in general, the amount of suspended sediment varies
logarithmically with respect to stream discharge. As discharge increases,
streams can also transport larger sediments. Correspondingly, the majority
of sediment transport in a given stream occurs during high flow conditions
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and seasonal floods, rather than during very low or even normal flow
conditions. During the higher flow events, particles previously transported
as bed sediment load and sediments deposited in the stream bed may be added
to the suspended sediment load. Extremely high flow events may transport
very large quantities of sediments, but are so infrequent and of such short
duration that their effect on the average sediment discharge is minimal
(Longwell, Flint, and Sanders, 1969).
Studies of PCB concentrations and transport in the Hudson River in New
York State (Turk, 1980) show a constant transport rate of PCBs during
moderate non-flood discharges and increased transport during floods. PCB
concentrations were found to increase as discharge fell below an intermediate
value. At low discharge, resuspension of bottom material was minimal, but
less dilution of contaminants occurred. At higher discharges, increased
concentrations were due to resuspension of contaminated bottom sediments.
During intermediate discharges, PCB concentrations were found to be a
function of both sediment resuspension and dilution. However, the net effect
of these opposing influences produces concentrations less than those achieved
during either low or high stream discharges.
Deterministic and empirical approaches were examined for their
usefulness in deriving exposure factors for stream transported sediments. In
general, deterministic approaches involve the use of numerical models to
approximate natural physical processes. Onishi, Whelan, and Skaggs (1982)
provide a comprehensive review of a number of models which could be used to
simulate sediment transport by streams. These models are divided into three
groups, as described in Chapter 4.O.. The least complex models require the
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least amount of site-specific data. Stream transport models in this group
involve dilution of contaminant concentrations with increases in stream
discharge and distance downstream from the source. No allowance is made in
these models for adsorption/desorption factors.
The two remaining groups of models require much more site-specific data,
including detailed stream channel and flow characteristics, as well as
adsorption/desorption and contaminant degradation factors. These models are
best applied when such specific information is available and are unsuited for
characterization of transport processes operating over a wide range of
geographic areas where site generalization must be made.
The unsuitability of the last- two groups of models for deriving simple
conversion factors stems from the extensive, site-specific data requirements
and the complexity of the computations. However, even the first group of
simple models requires flow rates and sediment size distributions for each
stream-transport scenario being considered. The sediment transport rate is
derived using stream-specific characteristics and empirically derived
constants that must be estimated for each sediment size range. Thus, none of
the models described by.Onishi, Whelan, and Skaggs (1982) are applicable for
estimating non-stream-specific sediment transport.
Procedures utilizing USLE losses and sediment delivery ratios to
estimate the sediment yield to streams from a source area have been discussed
in Chapter 4.0. In order to determine the concentration of dioxin on
sediments delivered to a stream, or at any point downstream of the source
area, it is necessary to estimate what "fraction of the total sediments at
that point were derived from the source area. Consequently, the size of the
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source area and the concentration of dioxin on the soil at the source must be
known. The size of the watershed must also be known or estimated from
topographic maps. Mechanisms of sediment supply, transport, and deposition
within the drainage basin are assumed to be in equilibrium.
The sediment delivery ratio for a given drainage basin decreases with
distance downstream of a contaminated source ; therefore, with distance
downstream, a decreasing portion of the total sediment yield reaches the
streams in the drainage area. The concentration of dioxin in stream
sediments at the point of exposure is a function of the downstream decrease
in sediment yield, due to deposition of contaminated sediments along the path
of sediment transport between the source area and the point of exposure. The
relationship between sediment delivery ratio and drainage area can and should
be regionalized when applied to a given site due to the effect of features
such as dams. The required data for regionalization are available for some
.watersheds.
Because the processes described in Section 4.7 are the same as those
affecting sediment transport, the dioxin concentration relationship for
sediments in a drainage system is:
rp sediment ^source x source no,
ur sediment = ~7~ = T 71 (U)
source basin basin
where C - dioxin concentration in soil at the source (ug/Kg)
^sediment = dioxin concentration in sediment at point of interest
(ug/Kg)
-54-
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^source = estimatecl soil 1°ss ^rom the USI-E for tne source
(tons/acre/yr)
Asource = source area (acres)
Lbasin = estiniated soil loss from the USLE for the watershed
upstream of the point of interest (tons/acres/yr)
Abasin = watershed area upstream of the point of interest (acres)
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6.0 MISSOURI BEEF DISTRIBUTION PATTERNS
Human exposure to dioxin through the consumption of beef products may
result if the livestock ingested and accumulated dioxin as a result of
contact with contaminated soils. Although dioxin exposure may occur from
consumption of dairy products from contaminated cattle, only meat consumption
is considered in this discussion. In areas where beef consumption involves
locally grown and fed cattle, this pathway can be additive to those stemming
from consumption of local fish, inhalation of dusts and vapor, and contact
with (and/or ingestion of) soil. In order to quantify this potential
pathway, an understanding of the pattern of beef production and meat
processing in the area of interest is necessary. The area of interest for
this report is the State of Missouri.
The beef industry in Missouri focuses largely on cow-calf production,
i.e., grazing herds which are utilized to produce calves to a point of
weaning. Backgrounding (preparing calves for feedlots) occurs in Missouri to
a lesser extent, and feeding comprises only a small segment of the state beef
industry. The small feeding segment is due, in part, to the fact that
Missouri is a grain-deficit state and does not produce a sufficient excess of
grain to economically support feeding operations. As a consequence, a large
percentage of the Missouri calf crop is shipped to Nebraska and Kansas for
backgrounding and feeding until it reaches" marketable size.
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The Missouri Crop Reporting Service reported that the January, 1984 herd
consisted of 2,376,000 head of cows (Sestak, 1984). Approximately 25% of the
herd is turned over through replacement, i.e., heifer calves are held back to
replace death losses, dry and barren cows, or older animals which are
slaughtered for low grade meat. The low grade meat is later distributed
nationally for hamburger and could be shipped anywhere in the United States.
Of the 25% of the herd turned over through replacement, roughly 60%, or 15%
of the total herd, represent cows sent to. slaughter for a variety of reasons.
.Total national input to. this pool averages 7,000,000 head annually (USDA,
1984), so, the 594,000 replacement figure in Missouri constitutes 5% of the
national inventory of cows destined for slaughter [(0.60)(594,000)
7,000,000].
At the replacement rate of 25%, 1,782,000 head of Missouri cows are
available for transport to feeding. Of that number, roughly 150,000 are fed
in state for commercial slaughter (Sestak, 1984). A second group of cows is
held for home slaughter. Because there are 107,000 cattle ranches and beef-
raising farms in Missouri (Grimes, 1984), with an estimated household size of
3.8 people, and assuming an average annual beef consumption level 77.4
Ibs/capita (Berglund, 1984), home slaughter could account for 107,000 x 3.8 x
77.4 = 31,470,840 Ibs/yr. This estimate is conservative because not all
farms slaughter their own beef for personal consumption. If the average
yield per head for home feeding is 550 Ibs, 31,470,840 Ibs/yr equate to
57,219 head. Hence, roughly 200,000 head of cattle are raised, fed,
slaughtered, and consumed in Missouri each year. The remaining 1,582,000
head of calves are shipped out of state for feeding, the bulk of which are
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sent to Nebraska and Kansas. In 1983, Nebraska marketed 4,580,000 head of
fed cattle and Kansas marketed 3,410,000 head, for a total of 7,990,000 head
(Gustafson, 1984). If it is assumed that the 1,582,000 head of Missouri-
raised calves are uniformly mixed into this pool, fed, slaughtered, and
distributed throughout the area as retail beef, approximately 20% of the beef
imported into Missouri would have been calved in Missouri. This estimate is
conservative because some beef may also be imported from Illinois and other
neighboring states.
The fraction of cattle that leave Missouri and then return as wholesale
and retail beef will have been subjected to "clean" feed during their
confinement. The feed period often lasts as long as 6 months, or 24 weeks.
Agricultural researchers have determined that the half-life of dioxin in beef
is 16.5 + 1.4 weeks (Jensen et a!., 1981). Therefore, cattle fed out of
state will have had a period of up to 1.5 half-lives to eliminate dioxin from
their bodies. This will lead to an overall reduction of
-(In 2)t
C = CQe \ (34)
-(In 2)24
- C = CQe 16
C - C^'-
C/C0 =0.35
where C = dioxin concentration at time of consumption
C = initial dioxin concentration.
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Hence, the beef produced from those Missouri-raised calves that were sent to
out-of-state feed operations and then returned to the state will carry 35% of
their original dioxin levels. At the same time, the animals will have
doubled in size; thus, 35% of the original mass of dioxin in these animals
will be distributed over twice the total volume of beef, resulting in a
dioxin concentration equal to 18% of the original contamination level CQ in
the beef (i.e., C/CQ = 0.18).
Assuming an average annual consumption of 77.4 Ibs/capita, the 5,000,000
people residing in Missouri will consume 390,000,000 Ibs of beef each year,
or the equivalent of 710,000 head. As noted above, 50,000 head will have
been home slaughtered and 150,000 head will have been raised and fed in
Missouri. Therefore, 510,000 head, or 72% of all beef consumed in Missouri
(77% of the beef consumed by non-cattle-raising inhabitants) will have been
imported from adjoining states.
Actually, none of the Missouri herd is known to be contaminated.
However, if contamination were detected, a factor of H/HQ» where H is the
size of the contaminated Missouri herd and H0 is the total Missouri herd size
of 2,376,000, could be used to calculate potential market dilution effects.
H should be estimated on the basis of animals on contaminated pasture. If
the actual herd size is not known, it can be estimated based on acres of
contaminated pasture and cattle density for the state (i.e., cows/acre
pasture normal use).
Based on the considerations discussed previously, the following
conclusions can be drawn with respect to dioxin exposure in Missouri from the
consumption of beef:
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No natural degradation or market dilution will occur for the home-
slaughter situation. Thus, any people consuming home-slaughtered beef
where contamination is found would be in a high risk population with
respect to beef consumption as a dioxin-exposure pathway.
The remaining Missouri inhabitants who purchase wholesale or retail beef
will consume beef consisting of 23% raised and fed in Missouri; 15%
calved in Missouri but fed out of state, and 62% calved, raised, and fed
out of state (Figure 8).
raised and fed
out of state
raised and fed
in Missouri
raised in Missouri
and fed out of state
Figure 8. Origin of Beef Consumed in Missouri
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On the average, dioxin levels (C) in beef consumed by the non-cattle-
raising inhabitants will be:
C = [(0.23)CQ + (0.15)(0.18)CQ + (0.62)(0)cJ H/HQ (35)
= (0.23 + 0.027 + 0)(C0)(H/HQ)
= 0.26(CQ)(H/H0)
where C = predicted level of dioxin for beef raised entirely in a
dioxin-contaminated area
H = size of the contaminated herd
H = total size of the state herd.
This relationship assumes that all out-of-state cattle are dioxin free.
The total herd size (HQ) in January, 1984 was 2,376,000, therefore:
C = (0.26)(CQ)(H/2,376,000) .
= 1.1 x 10'7 CQH • ' _
Approximately 5% of the U.S. slaughter cow inventory comes from
Missouri. Thus, subsequent meat products such as hamburger may contain
dioxin contamination at
C = 0.05 (H/HQ)C0 (36)
and for January, 1984 data
C = 2.0 x 10"8 HCQ
It should be noted that if an entire beef or half a beef is purchased,
or if a large amount of retail cuts are purchased at a single time, all of
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the meat will be contaminated or contamination free depending on the source.
When viewed in this manner, the market dilution concept appears inaccurate.
However, when chronic exposure is considered, the market dilution concept is
analogous to the purchase of a small percentage of dioxin-contaminated beef
within a larger volume of total beef purchased.
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REFERENCES
ASCE. 1975. Sedimentation Engineering. American Society of Civil
Engineers, Manuals and Reports on Engineering Practice, No. 54, V. A.
Vanoni, Editor, New York, NY.
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Fred C. Hart Associates, Inc. 1984. Assessment of Potential Health Impacts
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APPENDIX A
EXAMPLE SITE 1
A.I SITE DESCRIPTION
Consider the case of a property where dioxin-contaminated soil was used
as fill. The filled area (the source) is approximately 100 ft (30.5 m) long
and 100 ft (30.5 m) wide (0.23 acre). Sampling indicated that the average
dioxin concentration in the source was 150 ug/Kg. The source is currently
without vegetative cover.
The 'property is located in a valley through which a creek flows
(Figure A-l). The source is about 50 ft (15 m) from the creek. The slope of
the property is 1%. Soil type in the area is primarily clay. Average annual.
wind speed is 5 mph (2.24 m/sec), with the predominant wind direction down
the valley.
A.2 PROBLEM
Sampling was conducted only in the source area where human health risks
were considered to be highest. However, potential exposure to dioxin is also
of interest for areas downwind, downs lope, and downstream of the site.
Concentrations of dioxin at such points of exposure can be predicted using
the appropriate conversion factors for the various modes of environmental
transport.
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Figure A-l. Example Site 1
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In this example, there is concern for dioxin concentrations at the
following locations:
1) atmospheric concentrations at the source;
2) soil and atmospheric concentrations near the creek; and
3) sediment concentrations in the creek .1,0.00 ft (300 m) down the creek
from the source.
The drainage basin area at Point 2 is 0.05 sq mi (32 acres). The
average slope in the area is 1%, with an average slope length of 200 ft
(61 m).
The drainage basin area at Point 3 is 0.7 sq mi (448 acres), with an
average slope of 2% and an average slope length of 500 ft.
A.3 ATMOSPHERIC CONCENTRATION AT THE SOURCE
Because the average wind speed is less than 10 m/sec, Equation 21 is
used to calculate air concentrations.
C •
CFai> « -iHL = i x 10"7 Kg/m3 (21)
Csoil
Caif - (1 x 10'7 Kg/m3) Csoi] ug/Kg
= (1 x 10"7 Kg/m3)(150 ug/Kg)
- 0.000015 ug/m3
= 15 pg/m3
A.4 SOIL CONCENTRATION NEAR THE CREEK
The volume of soil lost from the site source is:
L=RxKxLSxCxP tons/acre/yr = 12 tons/acre/yr
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R = 215 yr'1 (interpolated from Figure 6)
K = 0.3 tons/acre (from Table 2)
LS = 0.186 (interpolated from Table 4)
C = 1
P = 1
Because the source area is 0.23 acre, the mass of soil from the site is 2.76
ton/yr (2,500 Kg/yr). At a soil dioxin concentration of 150 ug/Kg, 376 mg of
dioxin are transported from the source annually.
The volume of soil lost from the drainage area above Point 1 is:
L=RxKxLSxCxP= 7.35 tons/acre/yr
where R = 215 yr"1
K = 0.3 tons/acre
LS = 0.228
C = 0.5
p = 1
With a drainage basin area of 32 acres, the mass of soil lost from the basin
is 235 tons/yr (213,000 Kg/yr).
The dioxin concentration in soil at Point 2 is:
C - r
soil ~ source
L x A
source source
. Lbasin x Abasin .
(9)
= (150 ug/Kg)I"12 tons/acre/yr x 0.23 acre
[_7.35 tons/acre/yr x 32 acre
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150 ug/Kg ||Zi
=1.8 ug/Kg
Because the predominant wind direction is down the valley (at right angles to
overland flow direction), atmospheric deposition need not be considered when
calculating soil concentrations at Point 2.
A.5 ATMOSPHERIC CONCENTRATION NEAR THE CREEK
Because the predominant wind direction is down the valley (at right
angles to overland flow direction), dioxin in the soil at Point 2 will be the
primary source of atmospheric contaminants. At an average wind velocity of
2.24 m/sec, atmospheric concentrations can be calculated using Equation 21,
as follows:
Cai> =(1 x 10"7 Kg/m3)Csoil ug/Kg (21)
. = (1 x 10"7 Kg/m3)(1.8 ug/Kg)
= 0.00000018 ug/m3
= 0.18 pg/m3
A.6 SEDIMENT CONCENTRATION DOWNSTREAM
Soil losses from the source are the same as those calculated in
Section A.4. Soil losses from the basin upstream of Point 3 are:
L=RxKxLSxCxP=10.5 tons/acre/yr
1
where R = 215 yr
K = 0.3 tons/acre
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LS = 0.326
C = 0.5
P = 1
From Equation 12, the dioxin concentration in sediment at Point 3 is:
C - r
sediment ^source
source x source
basin x Abasin
(12)
= 150 ug/Kg
= 150 ug/Kg
12 tons/acre/yr x 0.23 acre
10.5 tons/acre/yr x 448 acres
[2.76'
[4700
= 0.088 ug/Kg
Alternately, the basin size and estimated dioxin concentration at Point 2 can
be used to estimate the dioxin concentration in sediment at Point 3, as
follows:
sediment " CPoint 2
LPoint 2 x APoint 2.
Lbasin x Abasin
= 1.8 ug/Kg
=1.8 ug/Kg
7.35 tons/acre/yr x 32 acres
10.5 tons/acre/yr x 448 acres
235
4,700
= 0.09 ug/Kg
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APPENDIX B
EXAMPLE SITE 2
B.I SITE DESCRIPTION
Consider the case of a private, rural lane (the source) that was sprayed
several times with dioxin-contaminated oil to control dust. The lane is 0.25
mi (400 m) long and 15 ft (4.6 m) wide (0.45 acre). Sampling indicated that
the average dioxin concentration in the source was 90 ug/Kg.
The source is located on a hill with a reported average slope of 3%
(Figure B-l). A creek flows through the valley at the bottom of the hill.
Soil type in the region is primarily silt loam. Average annual wind speed is
12 mph (5.4 m/sec), with the predominant wind direction down the valley.
B.2 PROBLEM
Sampling was conducted only in the source area where human health risks
were considered to be highest. However, potential exposure to dioxin is also
of interest for areas downwind, downs lope, and downstream of the site.
Concentrations of dioxin at such points of exposure can be predicted using
the appropriate conversion factors for the various modes of environmental
transport.
In this example, there are concerns about dioxin concentrations at the
following locations:
1) atmospheric concentrations at the source;
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Wind Direction
Overland Flow
Figure B-l. Example Site 2
-75-
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2) atmospheric concentrations near the residence, which is 100 ft (30 m)
beyond the end of the lane in the downwind direction;
3) soil and atmospheric concentrations adjacent to the creek that is 500 ft
(150 m) downslope from the source; and
4) sediment concentrations in the creek at a point 4,000 ft (1,200 m)
downstream of Point 3.
The drainage basin at Point 3 has an average slope of 3%, an average
slope length of 1,000 ft (300 m), and an area of 0.25 sq mi (160 acres).
The drainage basin at Point 4 has an average slope of 3%, an average
slope length of 1,000 ft (300 m), and an area of 1 sq mi (640 acres).
Soil type and average annual wind speed at the points of interest are
the same as those at the source.
B.3 ATMOSPHERIC CONCENTRATION AT THE SOURCE
Because the average wind speed is less than 10 m/sec, Equation 21 is
used to calculate air concentrations.
C
CFa.r =-^= 1 x ID'7 Kg/m3 (21)
Soil
Cair = (1 x 10~ K9/IT|) Csoil ug/Kg
m
3
Cai> = (1 x 10'7 Kg/m3) (90 ug/Kg)
Cai> = 0.000009 ug/m
= 9 pg/m3
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B.4 ATMOSPHERIC CONCENTRATION AT THE RESIDENCE
Equation 21 is modified to include the dispersion factor for calculating
air concentrations downwind of the source.
CF
air
soil
• 1 x lO
'7
(ayazW
where ^y^F = 46° m'
= 540m
2
-1
k_ - 0.001 sec
x = 30 m
F = 400 m
W = 5.4 m/sec
Substituting these values into Equation 8:
460
540
'soil
(8)
Cair = 0.00000009 ug/m3
= 0.09 pg/m3
B.5 SOIL CONCENTRATION NEAR THE CREEK
The soil lost from the source is
L=RxKxLSxCxP tons/acre/yr = 20 tons/acre/yr
where R = 215 yr"1
K = 0.4 tons/acre
LS = 0.233
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C = 1
p * -1
Because the source area is 0.45 acre, the mass of soil from the site is
9 tons/yr (8,200 Kg/yr). At a soil dioxin concentration of 90 ug/Kg, 738 mg
of dioxin are transported from the source annually.
The soil lost from the drainage basin above Point 3 is:
L=RxKxLSxCxP tons/acre/yr = 25 tons/acre/yr
where R = 215 yr"1
K = 0.4 tons/acre
LS = 0.573
C = 0.5
P = 1
With a drainage basin of 160 acres, the mass of soil lost from .the basin is
4,000 tons/yr (3,629,000 Kg/yr).
From Equation 9, the dioxin concentration in soil at'Point 3 is:
C - r
soil " source
= 90 ug/Kg
source x source
Lbasin x Abasin
20 x 0.45
(9)
25 x 160
=0.2 ug/Kg
Because the predominant wind direction is down the valley (at right
angles to overland flow direction), atmospheric deposition need not be
considered in calculating soil concentrations at Point 3.
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B.6 ATMOSPHERIC CONCENTRATION NEAR THE CREEK
Because the predominant wind direction is down the valley (at right
angles to overland flow direction), dioxin in the soil at Point 3 will be the
primary source of atmospheric contaminants. At an average wind velocity of
5.4 m/sec, atmospheric concentrations will be due to vaporization, according
to Equation 21:
Ca.r -(1 x 10'7 Kg/m3)Csoil ug/Kg- (21)
« (1 x 10'7 Kg/m3)(0.2 ug/Kg)
= 0.00000002 ug/m3
= 0.02 pg/m3
B.7 SEDIMENT CONCENTRATION DOWNSTREAM
Soil losses from the source are the same as those calculated in
Section B.5. Soil losses from the basin upstream of Point 4 are:
L=RxKxLSxCxP=25 tons/acre/yr
where R = 215 yr"
K = 0.4 tons/acre
LS = 0.573
C = 0.5
P = 1
From Equation 12, the dioxin concentration at Point 4 is:
L x A"
C r source source
sediment ~ source i Y /\
basin. * "basin
-79-
-------
= 90 ug/Kg
20 x 0.45
25 x 640
= 0.05 ug/Kg
Alternately, the basin size and estimated dioxin concentration at Point 3 can
be used to estimate the dioxin concentration in sediment at Point 4, as
follows:
CSediment = cPoint 3
= 0.2 ug/Kg
- 0.2 ug/Kg
0.05 ug/Kg
LPoint 3 x APoint 3
Lbasin x Abasin
25 tons/acre/yr x 160 acres
25 tons/acre/yr x 640 acres
4,000
16,000
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