United States
Environmental Protection
Agency
Office.of Water
Washington. DC 20460
EPA-822-R-96-003
August 1996
EPA Technical Support
Document for the Round
Two Sewage Sludge
Pollutants
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Technical Support Document
for the Round Two Sewage Sludge Pollutants
Prepared by:
U.S. Environmental Protection Agency
Office of Water
Office of Science and Technology
Health and Ecological Criteria Division
With the assistance of:
Abt Associates Inc.
Hampden Square - Suite 600
4800 Montgomery Lane
Bethesda, MD 20814
Under Contract No. 68-C3-0332
August 1996
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ACKNOWLEDGMENTS
This study was prepared by Abt Associates Inc., under Contract Number 68-C3-0332,
for the U.S. Environmental Protection Agency's Health and Ecological Criteria Division of
the Office of Water. The following Abt Associates staff contributed to the analysis, writing,
editing, and production of this document:
Elizabeth Fechner Levy Environmental Scientist
Susan Egan Keane Environmental Scientist
Vicki Hutson Technical Reviewer
Dan McMartin Environmental Modeler .
Josh Kanner Research Assistant
Caryl Waggett Research Assistant
Rich Walking Research Assistant
Han Wang . Research Assistant
Michael Wise Research Assistant
Julie Wormser Research Assistant
Abt Associates staff would like to thank Yogi Patel and Barbara Corcoran for their
guidance and support as EPA Project Managers at different stages of this project. We would
also like to thank Robert Southworth, Maria Gomez-Taylor, Mark Morris, and Alan Hais of
the Office of Water for their useful comments and valuable insights on various aspects of this
study. In addition, we thank Chuck White of SAIC for his statistical analyses of the National
Sewage Sludge Survey data.
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EXECUTIVE SUMMARY
Section 405(d) of the Clean Water Act requires the U.S. Environmental Protection
Agency (EPA) to promulgate regulations, including numerical limitations for pollutants that
may be present in sewage sludge, to protect public health and the environment from any
reasonably anticipated adverse effects of the pollutants. The statute requires EPA to
promulgate regulations' hi two stages. When the Agency failed to meet the statutory deadlines
for promulgation of these regulations, several groups sued EPA. The Agency agreed to
promulgate the first stage of regulation by November 1992, and the second stage by
December 2001. Under the agreement, filed hi Federal district court, EPA also agreed to
identify pollutants that it was considering for regulation in the second stage by November
1995 (Gearhart et al. v. Browner, Civ. No. 89-6266-HO, D. Oregon).
Implementation of the Clean Water Act (CWA) has resulted in higher levels of
wastewater treatment, leading to both greater pollutant removal from the wastewater before
discharge and generation of larger amounts of sewage sludge. Publicly Owned Treatment
Works (POTWs) currently generate 5.3 million dry metric tons of sewage sludge per year,
or approximately 47 pounds per person per year.
A POTW has a number of alternative practices for the use or disposal of sewage
sludge. These practices include, but are not limited to, land application, surface disposal, and
incineration. Approximately 33 percent of the sewage sludge generated by POTWs is Used
to condition the soil or to fertilize crops grown in the soil. Approximately ten percent of the
sewage sludge is placed hi surface disposal units, such as surface impoundments (into which
liquid sewage sludge is placed), sludge-only landfills (monofills), and piles left on land
surfaces. Approximately 16 percent of the sewage sludge generated by POTWs is combusted
in sewage sludge incinerators. The requirements that have to be met for these practices are
described in the Standards for the Use or Disposal of Sewage Sludge (40 CFR Part 503),
which were published under the authority of section 405(d) of the CWA. These standards
are known as the "Round One sewage sludge regulation." Most of the remaining sewage
sludge generated by POTWs is co-disposed hi landfills, which have requirements established
by the Solid Waste Disposal Facility Criteria,(40 CFR Part 258 and Part 503,4).
In November 1992, EPA promulgated numerical limits for ten pollutants in sewage
sludge. At the same tune, the Agency promulgated an operational standard for total
hydrocarbons as a surrogate for limits on organic pollutants hi the exit gas from sewage
sludge incinerators. EPA is now required to make a determination as to whether it is
necessary to propose and promulgate a regulation covering a second set of pollutants that may
cause adverse effects to public health or the environment (Round Two). A candidate list of
pollutants for the second round of the sewage sludge regulations was provided to the District
Court in Oregon hi May 1993. The final list of pollutants was provided to the District Court
hi Oregon hi November 1995. The purpose of this Technical Support Document is to provide
information on how both the candidate list and the final list of pollutants for the Round Two
sewage sludge regulation were derived.
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To select the candidate pollutants for Round Two, a Preliminary Hazard Identification
study was conducted. First, pollutants that were detected frequently in sewage sludge were
determined by using data from the 1988 National Sewage Sludge Survey (NSSS). Analytical
results for the 411 pollutants for which samples were analyzed in the NSSS were reviewed
to determine the frequency of detection for each pollutant. Pollutants with a frequency of
detection of less than ten percent were deleted from further consideration for the Round Two
pollutant list. There were 254 pollutants that were not detected, and another 69 pollutants
that had a frequency of detection of greater than zero but less than ten percent. If a pollutant
had a frequency of detection of ten percent or greater, and had not already been regulated in
Round One, then scientific literature was reviewed to determine whether there were toxicity
data for a pollutant. If no human health or ecological data were found for a pollutant, no
further consideration was given to that pollutant for the Round Two list. Based on the results
of the Preliminary Hazard Identification, a list of 31 Round Two pollutant candidates was
submitted to the District Court hi Oregon in May, 1993.
The 31 pollutant candidates identified in the Preliminary Hazard Identification study
were then evaluated hi a Comprehensive Hazard Identification study to determine the final
list of pollutants for the Round Two sewage sludge regulation. In the Comprehensive Hazard
Identification study, a quantitative risk assessment, including dose-response evaluation,
exposure assessment, and risk characterization, was performed. The goal was to identify
pollutants that may potentially cause human health or ecological risk for a Highly Exposed
Individual (HEI). The risk to the HEI was estimated using a combination of high-end and
average assumptions designed to give a plausible estimate of the individual risk at the upper
end of the risk distribution (e.g., above the 90th percentile of the actual distribution). In
general, high-end assumptions were used to characterize sewage sludge concentrations and
certain exposure parameters, while average values were typically used to characterize
use/disposal practices and soil and meteorological characteristics. Sewage sludge
concentrations were based on the 95th percentile concentrations of pollutants obtained hi the
NSSS, with non-detects set equal to the minimum level (e.g., the minimum concentration of
pollutant that could be measured).
For land application, risks were estimated for 15 exposure pathways. For surface
disposal, risks were estimated for two exposure pathways, and for incineration, one exposure
pathway. If risk values were greater than certain thresholds for a given pollutant and
exposure pathway, that pathway was defined as "critical" for that pollutant. The threshold
for carcinogens was an individual risk of IxlO"4 or higher; for non*carcinogens, a ratio of
exposure to the Risk Reference Dose of one or greater; and for ecological risk, a Risk
Quotient of one or greater. Critical pathways were identified for 12 pollutants. These 12
pollutants were then evaluated further. As a result of this evaluation, EPA reported hi
November 1995 to the District Court hi Oregon that it only was considering proposing
regulation of two pollutants, dioxins/dibenzofurans and coplanar polychlorinated biphenyls,
in Round Two.
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TABLE OF CONTENTS
ACKNOWLEDGMENTS .
• • • • • .......... ..................... i
EXECUTIVE SUMMARY ...... ..... ...... .... „
TABLE OF CONTENTS ........................ . . •. iv
LIST OF EXHIBITS ............. ; ...I..........,...;.. vii
LIST OF MATHEMATICAL SYMBOLS
1. INTRODUCTION ...
1.1 BACKGROUND
1.2 PURPOSE ..... . ..... . . ........ ............ x_3
2. ROUND ONE SEWAGE SLUDGE REGULATION 21
2.1 DESCRIPTION OF A PART 503 STANDARD .....]... . . . .' " .' " 2-1
2.2 SEWAGE SLUDGE USE OR DISPOSAL PRACTICES . . ..... 2-2
2.2.1 Land Application ...... ....... . . . 2-2
2.2.2 Surface Disposal ..... ; ..... 2-2
2.2.3 Incineration .......... ....... .... 2-3
2.3 DATA GATHERING STUDIES . . . ____ . ..... 2-3
2.3.1 The 40 City Study . . . . ......... .. .[.... . . . .2-3
2.3.2 Environmental Profiles and Hazard Indices . ........ . 2-4
2.3.3 Sewage Sludge Incinerator Field Studies . ..... 2-5
2.3.4 National Sewage Sludge Survey ...... ..... 2-5
2.4 ROUND ONE POLLUTANTS . . ......... 2.6
3. CANDIDATE LIST OF ROUND TWO POLLUTANTS 31
3.1 SELECTION PROCESS ..... ........ ;.......'.'.'.'.'.'.'.''' 3-1
3.2 NATIONAL SEWAGE SLUDGE SURVEY POLLUTANTS ........ 3-1
3 . 3 RESULTS OF PRELIMINARY HAZARD IDENTIFICATION
....... ........ .......... ... ............ 3_10
3.3.1 Pollutants Removed from Further Consideration .......... 3-10
3.3.2 Individual Pollutants Combined Into Classes ........... 3-io
3.3.3 Frequency of Detection of Pollutants ......... ...... 3.10
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3.3.4 Available Human Health and Ecological Toxicity Data ...... 3-11
3.3.5 Pollutant Candidate List for Round Two Regulation ....... 3-12
4. FINAL LIST OF ROUND TWO POLLUTANTS ............ ..... 4-1
4.1 GENERAL APPROACH FOR THE COMPREHENSIVE HAZARD
IDENTIFICATION ......... . . ....... ............... 4-1
4.2 LAND APPLICATION PATHWAY EXPOSURE METHODOLOGIES . . 4-4
4.2.1 Pathway 1 - Ingestion of Crops Grown on Sewage Sludge-
Amended Soil . . ............................. 4-10
4.2.2 Pathway 2 - Ingestion of Crops Grown hi Sewage Sludge-
Amended Home Gardens ........................ 4-19
4.2.3 Pathway 3 - Direct Ingestion of Sewage Sludge by Children . . 4-25
4.2.4 Pathway 4 - Ingestion of Animal Products Produced From
Animals Consuming Forage/Pasture Grown on Sewage Sludge-
Amended Soil ... ...... .......... .......... . . 4-26
4.2.5 Pathway 5 - Consumption of Animal Products Produced From
Animals That Ingest Sewage Sludge ............... . . 4-33
4.2.6 Pathway 6 - Animal Toxicity From Plant Consumption ..... 4-36
4.2.7 Pathway 7 - Animal Toxicity From Direct Ingestion of Sewage
Sludge ............ . ...................... 4-38
4.2.8 Pathway 8 - Toxicity to Plants .................... 4-40
4.2.9 Pathway 9 - Toxicity to Soil-Dwelling Organisms ......... 4-40
4.2.10 Pathway 10 - Toxicity to Predators of Soil-Dwelling
Organisms . .......... ....... ...... ... ..... 4^1
4.2.11 Pathway 11 - Human Toxicity Through Inhalation of
Particulates Resuspended by Tilling Sewage Sludge ......... 4-45
4.2.12 Pathway 12 - Ingestion of Fish and Water from Surface Water
that Receives Eroded Soil ....................... 4-48
4.2.13 Pathway 13 - Inhalation of Pollutants Volatilized from Land- .
Applied Sewage Sludge ............ .......... ..... 4-82
4.2.14 Pathway 14 - Ingestion of Groundwater Containing Leached
Pollutants .......... ............ . .......... 4-90
4.2.15 Pathway 15 - Infant Exposure to Pollutants Through
Breastfeeding ........... ..... : . . . . .......... 4-96
4.3 SURFACE DISPOSAL EXPOSURE METHODOLOGIES ........ 4-103
4.3.1 Definitions of a Monofill and a Surface Impoundment ..... 4-103
4.3.2 Methods for the Monofill Prototype . . .............. 4-103
4.3.3 Methods for the Surface Impoundment Prototype ........ 4-116
4.3.4 Estimating Human Exposure ................... . . 4-129
4.3.5 Data Inputs .................. . . . ........... 4-130
4.3.6 Modeling of Surface Impoundments hi the Comprehensive
Hazard Identification ......... ... .............. 4-136
4.3.7 Example Exposure Calculations for Surface Disposal ...... 4-136
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4.4 INCINERATION EXPOSURE METHODOLOGIES . . 4-153
4.4.1 Estimating Emissions of Pollutants from Incinerators 4-153
4.4.2 Modeling the Dispersion of Pollutants in Air 4-154
4.4.3 Mapping Dispersion and Pollutant Concentrations Onto a
Unified Grid . 4-154
4.4.4 Estimating Human Exposure 4-156
4.4.5 Data Inputs 4-156
4.4.6 Example Calculations for Incineration . 4-157
4.5 RISK CALCULATIONS ;..... 4-158
4.5.1 Human Health Risk Calculations 4.153
4.5.2 Ecological Risk Calculations 4-165
4.5.3 Human Health and Ecological Risk Results . 4-169
5. FURTHER ANALYSES OF ROUND TWO POLLUTANTS 5 1
5.1 INTRODUCTION -.-..........,........!"!'''' 5-1
5.2 POLLUTANTS THAT WARRANT FURTHER CONSIDERATION . . . 5-1
6. LIST OF POLLUTANTS FOR THE ROUND TWO REGULATION • ' -
SUBMITTED TO THE COURT 6_j
7. REFERENCES 7_j
APPENDICES
Appendix A: Analysis of Pollutants Detected Less than Ten Percent of the Time
Appendix B: Statistical Analyses of the National Sewage Sludge Survey Data
Appendix C: Calculation of a "Square Wave" for the Groundwater Pathway
Appendix D: Evaluation of Candidate Pollutants for the Round Two Sewage Sludge
Regulation
Appendix Dl: List of 31 Candidate Pollutants for the Round Two Sewage Sludge
Regulation Submitted to the Court
Appendix D2: Final List of Pollutants for the Round Two Sewage Sludge Regulation
Submitted to the Court
Appendix D3: Responses to Requests for Data on the Round Two Candidate
Pollutants
VI
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LIST OF EXHIBITS
Exhibit 3-1: Pollutants Analyzed in the National Sewage Sludge Survey 3-4
Exhibit 3-2: Pollutants Regulated in Round One Sewage Sludge Regulation 3-13
Exhibit 3-3: Pollutants Combined Into Classes or Removed from Further
Consideration 3-14
Exhibit 3-4: Pollutants With a Frequency of Detection of Zero Percent in the
National Sewage Sludge Survey 3-15
Exhibit 3-5: Pollutants With A Frequency of Detection of Less Than Ten Percent in
the National Sewage Sludge Survey . 3-19
Exhibit 3-6: Pollutants With A Frequency of Detection of Ten Percent or Greater in
the National Sewage Sludge Survey 3-21
Exhibit 3-7: Available Human Toxicity Data . 3-23
Exhibit 3-8: Available Ecological. Toxicity Data . . 3-25
Exhibit 3-9: Pollutants With No Human Health or Ecological Toxicity Data ...... 3-25
Available 3-27
Exhibit 3-10: Round One Pollutants Included as Potential Candidates For Round Two . 3-28
Exhibit 3-11: 31 Pollutant Candidates For Round Two Regulation ; . 3-29
Exhibit 3-12: Rationale for the Number of Pollutants Selected as Candidates for the
Round Two Sewage Sludge Regulation ...,., 3-30
Exhibit 4-1: 95th Percentile Concentrations for Round Two Candidate Pollutants .... 4-3
Exhibit 4-2: Definitions of Exposure Pathways and Highly Exposed Individuals (HEIs) for
Land Application 4.5
Exhibit 4-3.: Average Values for Sewage Sludge Land Application Parameters ...... 4-9
Exhibit 4-4: Background Concentrations of Pollutants in Soil . . 4-10
Exhibit 4-5: Dietary Assumptions for Pathway 1 ...................... 4.13
Exhibit 4-6: Available Plant Uptake Slopes for Agricultural Pathway 1 . . 4-15
Exhibit 4-7: Available Plant Uptake Slopes for Non-Agricultural Pathway 1 4-17
Exhibit .4-8: Dietary Assumptions for Pathway 2 4-19
Exhibit 4-9: Available Plant Uptake Slopes, for Agricultural Pathway 2 ......... 4-21
Exhibit 4-10: Dietary Assumptions for Pathway 4 4-27
Exhibit 4-11: Forage/Pasture Uptake Slopes for Agricultural Pathway 4 4-29
Exhibit 4-12: Animal Uptake Slopes for Agricultural Pathway 4 . 4-31
Exhibit 4-13: Animal Uptake Slopes for Non-Agricultural Pathway 4 . : 4-32
Exhibit 4-14: Dietary Assumptions for Pathway 5 . 4.35
Exhibit 4-15: Bioaccumulation Factors for Soil-Dwelling Organisms 4-44
Exhibit 4-16: Non-Pollutant-Specific Parameters for Pathways 12, 13, and 14 4-67
Exhibit 4-17: Environmental Fate and Transport Parameters 4-68
Exhibit 4-18: Parameters Used to Calculate
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Exhibit 4-23: Environmental Fate and Transport Parameters 4-132
Exhibit 4-24: Human Health Toxicity Numbers . 4-159
Exhibit 4-25: Threshold Limit Values for Pollutants ....." 4-161
Exhibit 4-26: Toxicological Reference Values for Mammals 4-166
Exhibit 4-27: Toxicological 'Reference Values for Soil-Dwelling Organisms 4-169
Exhibit 4-28: Risk Results for Highly Exposed Individual for Pathway 1 ........ 4-170
Exhibit 4-29: Risk Results for Highly Exposed Individual for Pathway 2 4-170
Exhibit 4-30: Risk Results for Highly Exposed Individual for Pathway 3 4-171
Exhibit 4-31: Risk Results for Highly Exposed Individual for Pathway 4 4-173
Exhibit 4-32: Risk Results for Highly Exposed Individual for Pathway 5 4-173
Exhibit 4-33: Risk Results for Highly Exposed Individual for Pathway 6 4-174
Exhibit 4-34: Risk Results for Highly Exposed Individual for Pathway 7 4-175
Exhibit 4-35: Risk Results for Highly Exposed Individual for Pathway 9 4-176
Exhibit 4-36: Risk Results for Highly Exposed Individual for Pathway 10 4-177
Exhibit 4-37: Risk Results for Highly Exposed Individual for Pathway 11 4-178
Exhibit 4-38: Risk Results for Highly Exposed Individual for Pathway 12 4-179
Exhibit 4-39: Risk Results for Highly Exposed Individual for Pathway 13 4-180
Exhibit 4-40: Risk Results for Highly Exposed Individual for Pathway 14 4-181
Exhibit 4-41: Risk Results for Highly Exposed Individual for Pathway 15 4-183
Exhibit 4-42: Risk Results for Highly Exposed Individual for the Surface Disposal
Pathways 4-183
Exhibit 4-43: Risk Results for Highly Exposed Individual for Incineration Pathway 4-184
• • -
Exhibit 5-1: Pollutants with Critical Land Application Pathways 5-2
Exhibit 5-2: Pollutants with Critical Surface Disposal Pathways 5-3
Exhibit 5-3: Summary of Critical Pathways and HEIs for Inorganic Pollutants 5-4
Exhibit 5-4: Summary of Conservative Assumptions hi Critical Pathways 5-5
Exhibit 5-5: Measurement Endpoints for Toxicological Reference Values for
Inorganic Pollutants [ 5-6
Exhibit A-l: Available Human Toxicity Data for 72 Chemicals Detected Less Than 10
Percent of the Time A-2
Exhibit A-2: Agricultural Pathway 3 Analysis for 43 Chemicals with Oral RfD or QJ*
Values A-4
Exhibit A-3: Pollutant-Specific Data Required for Pathways 12 and 13 A-5
Exhibit A-4: Individual Cancer Risks for Aldrin/Dieldrin from Pathway 12 A-6
Exhibit A-5: Individual Cancer Risks for Aldrin/Dieldrin from Pathway 13 A-6
vin
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LIST OF MATHEMATICAL SYMBOLS
= intermediate diffusivity variable defined in Eq. 4-31 (nrVsec),
= change in total mass of pollutant hi soil (kg),
Ar = one year,
fj.a ~ viscosity of air (g/cm-sec),
pa = density of air (g air/cm3 air),
Psi ~- particle density of sewage sludge (kg sewage sludge/m3 sewage sludge),
. Pss = particle density of sewage sludge-soil mixture (kg/m3),
Pw = density of water (kg water/m3 water),
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C2 = total concentration of pollutant in solids layer (kg pollutant/m3 solids layer),
Cg = concentration of gaseous pollutant in air-filled pore space (kg pollutant/m3
air), .
Ca,r>J = average concentration of pollutant j in ambient air at the downwind edge
of the site (/*g pollutant/m3 ah-),
Caoti.] ~ dry weight concentration of pollutant j in eroded soil entering the stream
(mg pollutant/kg eroded soil),
CffJ = concentration of pollutant j in fish fillets (mg pollutant/kg fish fillet),
Q = concentration of pollutant hi inflow to the impoundment (kg pollutant/m3
sewage sludge),
Cj = concentration of pollutant j in sewage sludge (mg pollutant/kg sewage
sludge),
C^ ~ concentration of pollutant hi water leaching from sewage sludge-amended
soil (kg pollutant/m3 porewater),
Cieej = average concentration of pollutant j in water leaching from the sewage
sludge-amended soil (mg pollutant/L water),
cmiiifaij = concentration of pollutant j in maternal milk (mg pollutant/kg milk fat),
Ct — concentration of sorbed pollutant on sewage sludge-amended soil particles
(kg pollutant/kg soil),
Cstp.j — average concentration of pollutant/' hi water seeping through the bottom
of the impoundment (mg pollutant/L water),
csaf, j ^ concentration of pollutant/ hi sewage sludge-amended soil eroded from the
land application site (mg pollutant/kg sewage sludge-amended soil),
CSH.J = concentration of pollutant/ in surface water (mg pollutant/L water),
Ct = total concentration of pollutant hi bulk sewage sludge-amended soil (kg
pollutant/m3 total bulk soil volume),
C* = concentration of dissolved pollutant hi water-filled pore space (kg
pollutant/m3 porewater),
Cwet. ~ concentration of pollutant/ hi well water (mg pollutant/L water),
= concentration of pollutant j in annual product k (mg pollutant/kg animal
tissue),
— tissue concentration (dry weight) of pollutant j in forage/pasture (mg
pollutant/kg forage/pasture),
= tissue concentration (dry weight) of pollutant/ hi crop / (mg pollutant/kg
crop tissue),
= tissue concentration (dry weight) of pollutant/ hi forage/pasture (mg
pollutant/kg forage/pasture),
= incremental cancer risk from pollutant/ for exposed individual (incremental
risk of developing cancer per lifetime of exposure),
= concentration of pollutant/ hi sewage sludge-amended soil (mg pollutant/kg
•sewage sludge-amended soil),
d • = depth of incorporation of sewage sludge (cm),
d, = depth of liquid layer (m), • ' •
d2 = depth of solids layer (m),
4> = depth of aquifer (m),
dc = depth of soil cover (m),
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'de = average rate of soil loss due to erosion from sewage sludge-amended land
each year (m/yr),
dm/ ~ depth of a monoflll active sewage sludge unit (m),
dsi = total depth of impoundment (m),
Df = anti-dilution factor (dimensionless),
D™g = geometric mean of the maximum dispersion ratios for 172 incinerators (/*g
pollutant/m3 air per g pollutant/sec),
Dca - molecular diffusivity of pollutant in air (cmVsec),
DC* = diffusivity of pollutant in water (cm2/sec),
Dei •= intermediate diffusivity variable defined in Eq. 4-30 (m2/sec),
Dah = diffusivity of diethyl ether in water (cnrVsec),
Df .. = dilution factor (dimensionless),
D'p = dispersion ratio for grid cell z impacted by incinerator p (fig pollutant/m3
air per g pollutant/sec),
DAk = daily dietary consumption of animal product k (g dry weight animal
product/day),
DCt = daay dietary consumption of crop / (g crop tissue/day),
de = effective diameter (or fetch) of surface impoundment (m),
DE , = exposure duration adjustment (number of years of exposure divided by
expected lifetime of 70 years),
DV = rate of change in the volume of the liquid layer (m3 liquid layer/sec),
Eavg.j = average emission rate for pollutant/'(g pollutant/sec),
EJP = emission rate for pollutant j at incinerator p (g pollutant/sec),
ED = exposure duration (yr),
EXPfJ = exposure to pollutant j through ingestion of fish (mg pollutant/kg body
weight-day),
EXPy = exposure to pollutant./ for individuals.living in grid cell i (mg pollutant/kg
body weight-day),
EXP; = exposure to pollutant/ (mg pollutant/kg body weight-day),
EXP*.J = exposure to pollutant j through direct ingestion of surface water (mg
pollutant/kg body weight-day),
EXPAj = exposure of animal to pollutant/ (mg pollutant/kg diet),
EXPIj = infant's average daily exposure to pollutant/ (mg pollutant/kg body weight-
day),
EXPOj = exposure of soil-dwelling organisms to pollutant/ (mg pollutant/kg sewage
sludge-amended soil),
EXPTj =• exposure of tractor operator to pollutant/ (mg pollutant/m3 air),
f*-J = proportion of ingested pollutant/ that isi stored in fat (dimensionless),
f* = proportion of mother's weight that is fat (kg maternal fat/kg total body
weight),.
fs = proportion of fat in breast milk (dimensionless),
&• J- • = proportion of ingested pollutant / that is absorbed (dimensionless),
ft = fraction of total pollutant lost during monofill's active lifetime
(dimensionless),
faa = fraction of each year's loading of pollutant lost during each year of the
surface impoundment's active phase (dimensionless),
XI
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fcg = fraction of monofill's active lifetime that typical active sewage sludge unit
contains sewage sludge with temporary soil cover (dimensionless),
fdi = fraction of total pollutant in liquid layer that is dissolved (dimensionless),
jfo = fraction of total pollutant in solids layer that is dissolved (dimensionless),
fte = fraction of "total pollutant loss during monofill's active operation
attributable to degradation (dimensionless),
ftteg — fraction of total pollutant loss caused by degradation (dimensionless),
fjqj = fraction of total pollutant lost from liquid layer that is lost to degradation
(dimensionless), .
fdeg2 = fraction of pollutant reaching the solids layer that is lost to degradation
(dimensionless),
faeti = , fraction of total pollutant lost from the liquid layer as a result of the
diminishing volume of the liquid layer (dimensionless),
fdta = fraction of pollutant reaching the solids layer that is stored in the
accumulating depth of this layer (dimensionless),
fen = fraction of total pollutant loss caused by erosion (dimensionless),
fia = fraction of total pollutant loss during monoflU's active operation
attributable to leaching (dimensionless),
ftfc . = fraction of total pollutant loss caused by leaching (dimensionless),
" fig == fraction of total cumulative loading lost in individual's lifetime to all four
loss processes (dimensionless),
fa. = fraction of organic carbon (dimensionless),
fout = fraction of total pollutant lost from the impoundment through outflow
(dimensionless),
foiai = fraction of total pollutant lost from liquid layer that is lost in outflow from
the impoundment (dimensionless),
fsep = fraction of total pollutant lost from the impoundment through seepage
(dimensionless),
ftepi ~. fraction of total pollutant lost from liquid layer that is lost to seepage
(dimensionless),
fsepz = fraction of pollutant reaching the solids layer that is lost to seepage
(dimensionless),
fst = fraction of monofill's volume containing sewage sludge (dimensionless),
fsoi = fraction of solids in sewage sludge (kg solids/kg sewage sludge),
fun = fraction of monofill's active lifetime that a typical active sewage sludge
unit contains sewage sludge without soil cover (dimensionless),
fva = fraction of total pollutant loss during monofill's active operation
attributable to volatilization (dimensionless),
Jw = fraction of total pollutant loss from inactive monofill attributable to
volatilization (dimensionless),
/vir = fraction of pollutant mass that volatilizes over a human lifetime
(dimensionless),
fvoi — fraction of total pollutant loss caused by volatilization (dimensionless),
frit =• fraction of total pollutant lost from liquid layer that is lost to volatilization
(dimensionless),
xu
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F* . = the volume of fluid passing through a vertical cross-section of the aquifer
oriented perpendicular to the direction of flow (m3/sec),
FAk = fraction of dietary consumption of animal product k produced on sewage
sludge-amended soil (dimensionless), ~
FAUc = annual average flux of pollutant leached from sewage sludge-amended soil
(kg pollutant/ha-yr),
FAUcJ = annual average flux of pollutant./ leaching (kg pollutant/ha-yr),
FAvolJ = annual average flux of pollutant J volatilizing from the sewage sludge-
amended soil (kg pollutant/ha-yr),
Fq = fraction of dietary consumption of crop /grown in sewage sludge-amended
soil (dimensionless),
FD = fraction of diet considered to be soil organisms (dimensionless, kg soil
organisms/kg diet),
FM = pollutant-specific food chain multiplier (dimensionless),
F& - ratio of fetch to depth (dimensionless),
FS = fraction of animal's diet that is sewage sludge (dimensionless, kg sewage
sludge/kg diet),
H = Henry's Law constant (dimensionless),
H = Henry's Law constant (atm-m3/mol),
hj ' = half-life of pollutant j in adults (days),
IA • . = - . inhalation rate (m3 air/day),
IF = daily consumption of fish fillets (kg fish fillets/day),
IM •= higestion rate of breast milk (kg milk/day), ,
IS = sewage sludge higestion rate (g sewage sludge/day),
Tw = volume of water ingested daily (L water/day),
Ki = soil-water partition coefficient (L water/kg soil),
K** = loss rate due to abiotic or microbial degradation of the pollutant on-sewaee
sludge-amended land (yr1),
Kdegi = anaerobic rate of pollutant degradation in liquid layer (sec'1),
JW - anaerobic rate of pollutant degradation hi solids layer (sec-1)'
m ~ Ioss1 rate due to erosion of the pollutant from sewage sludge-amended land
Kg = mass transfer coefficient for the gas layer (m/sec),
A"/ = mass transfer coefficient for the liquid layer (m/sec),
Kte ~ l°ss rate due to leaching of the pollutant (yr1),
KOC = organic carbon-water partition coefficient (mLwater/g organic carbon)
A*. - octanol-water partition coefficient (dunensionless, mg pollutant/L octan'ol
per mg pollutant/L water),
Kta = total loss rate of pollutant due to leaching, volatilization, and degradation
during monofilFs active operation (yr1),
Kti = total loss rate of pollutant from inactive monofill^yr1),
Km, = • total loss rate for the pollutant from sewage sludge-amended lancl (yr1)
A-., = coefficient for the total rate at which pollutant is lost from the liquid layer
of a surface impoundment (mVsec), '
coefficient for the total rate at which pollutant is lost from or stored in the
solids layer of a surface impooundment (mVsec),
*-totl
Xlll
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= loss rate of pollutant due to volatilization during monofill' s active operation
(yr1),
= loss rate of pollutant due to volatilization from inactive monofill (yr1),
= loss rate due to volatilization of the pollutant from sewage sludge-amended
lander1),
KVOII — rate of pollutant volatilization from liquid layer (m/sec),
LF = active lifetime of monofill (yr),
LS = average human lifetime (yr),
mbce*growui.f = maternal intake of pollutant/ from sources other than sewage sludge (mg
pollutant/kg body weight-day), . ' ,
Mmg = geometric mean of. sewage sludge feed rates for 172 incinerators (kg
sewage sludge/yr),
MO, = mass of gaseous pollutant (kg),
Ma = mass of sorbed pollutant (kg), .
Ma = total mass of pollutant hi soil (kg),
MC,, = mass of dissolved pollutant (kg),
MEsile = rate of soil loss for land treated with sewage sludge (kg sewage sludge-
amended soil/ha-yr),
MEm = estimated rate of soil loss (erosion) for the watershed (kg soil/ha-yr),
Mff = mass of pollutant in sewage sludge/soil at end of monofill's active lifetime
(kg pollutant/ha),
MIS = mass °f pollutant hi soil at end of a period equal to an individual lifetime
(mg pollutant/ha);
Mp = mass of sewage sludge incinerated at incinerator p each year (kg sewage
sludge/yr),
Ms = mass of soil (kg),
MS = mass of soil in mixing zone of one hectare of land (Mg soil/ha land),
msiudge.j = maternal intake of pollutant j from relevant sewage sludge exposure
pathways (mg pollutant/kg body weightrday),
MW = molecular weight of pollutant (g/mol),
n = number of incinerators modeled,
, N = total number of years sewage sludge is applied to land (yr),
Nsaf = site life (yr),
No. = total average emissions from the soil surface over tune interval te (kg
pollutant/m2 soil),
Nas = emissions from the soil surface in first second (kg pollutant/m2),
Nay =• total average emissions from the soil surface hi first year (kg pollutant/m2),
NR = annual recharge to groundwater (m3 recharge/m2 area-yr, or m
recharge/yr),
PI • = fraction of solids (by mass) in liquid layer (kg solids/kg liquid layer),
P2 = fraction of solids (by mass) in solids layer (kg solids/kg solids layer),
Pf — ratio of pollutant concentration in fillet to whole fish (dimensionless),
= intake level of pollutant/ hi insectivorous mammal's diet (mg pollutant/kg
diet),
= tune-weighted average pollutant flux from typical monofill unit over the
active lifetime of the monofill (kg pollutant/m2 unit-sec),
xiv
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% - pollutant flux from treated soil for covered period (kg pollutarit/m2-sec)
qJ = human cancer potency of pollutanty (mg pollutant/kg body weight-day)-'
?«« = pollutant flux from sewage sludge/soil for uncovered period (kg
pollutant/m2 unit-sec),
Qi = rate at which sewage sludge enters the impoundment (m3 sewage
sludge/sec), &
Qo = rate at which outflow leaves the impoundment (m3 sewage sludge/sec),
QW = rate of seepage beneath the impoundment (m/sec),
r' = distance from center of the land application site to the downwind edge (m)
R = gas constant (L-atm/mol-K),
RJ = combined removal efficiency for pollutanty of furnace and control devices
expressed as fraction of original pollutant mass retained by the furnace or
pollution control devices (dimensionless),
RJP = combined removal efficiency for pollutanty of furnace and control devices
for incinerator p expressed as fraction of original pollutant mass retained
by the furnace or pollution control devices (dimensionless),
Risk Reference Dose for pollutanty (mg pollutant/kg body'weight-day)
ratio of the exposure to the RiD for pollutanty (dimensionless),
tun
ecological risk quotient for pollutanty (dimensionless),
RTCj = ration of the exposure to the TLV-TWA for pollutant y (dimensionless),
^i . '.' — concentration of solids in liquid layer (kg/m3),
"V = concentration of solids hi solids layer (kg/m3)' ~
ScG = the Schmidt number on the gas side (dimensio'nless),
SSM = sediment delivery ratio for the land application site (dimensionless)
^ ~ sediment delivery ratio for the watershed (dimensionless),
SC = estimated mass of sewage sludge contained in one hectare of completed
. monofill (kg/ha), F
SRR = source-receptor ratio (sec/m),
t = time (yr),
= duration of emissions (sec),
time that a typical active sewage sludge unit contains uncovered sewage
sludge (yr), *
= temperature (Kelvin),
= total exposure of tractor operator to soil dust (mg soil dust/m3 air)
TF - estimated active lifetime of surface impoundment (sec)
TLV-TWAj = Threshold Limit Value-Time Weighted Average for pollutant / (mg
pollutant/m3 air),
TP = duration of "square wave" for approximating the loading of pollutant into
the unsaturated soil zone (yr),
TPN = total mass of pollutant available at a site after the final year of application
(mg pollutant/ha), .
TRVj = toxicological reference value for pollutanty for an animal (mg pollutant/kg,
diet),
TSS = concentration of total suspended solids in the stream (mg solids/L water)
u = average wind speed (m/sec), '
u"> = average wind speed 10 m above surface (m/sec),
xv
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forage, j = rate °^ uptake of pollutant j into tissue of forage/pasture (mg pollutant/kg
dry weight forage/pasture per mg pollutant/kg soil),
= rate of uptake of pollutant y into tissue of crop / (jig pollutant/g dry weight
crop tissue per /zg pollutant/g sewage sludge-amended soil),
Ujf. = rate of uptake of pollutant j into animal product k (mg pollutant/kg dry
weight animal tissue per mg pollutant/kg dry weight diet),
v = vertical term for dispersion of pollutant hi air (dimensionless),
vh = regional velocity of horizontal groundwater flow (m/sec),
vt = superimposed radial velocity from water, seeping from impoundment
(m/sec),
vv = vertical velocity due to the source (m/sec),
Va = volume of air in soil (m3),
Vs = volume of solids in soil (m3),
V, = total bulk volume of soil (m3),
Vw = .volume of water in soil (m3),
x = distance from the center of the land application site to the downwind edge
(km),
•xy = lateral virtual distance to land application site (m).
XVI
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1. INTRODUCTION
Under section 405(d) of the Clean Water Act, the U.S. Environmental Protection
Agency (EPA) is required to promulgate regulations to protect public health and the
environment from any reasonably anticipated adverse effects of pollutants in sewage sludge
In November 1992, limits were promulgated for ten pollutants in sewage sludge. At the same
time, an operational standard was promulgated for total hydrocarbons as a surrogate for
organic pollutants in the exit gas from sewage sludge incinerators. EPA is now required to
propose and promulgate a regulation covering a second set of pollutants that may cause
adverse effects to public health or the environment.
1.1 BACKGROUND
Congress adopted the Clean Water Act (CWA) to "restore and maintain the
chemical, physical, and biological integrity of the Nation's waters" (section 101(a) 33 U S C
1251(a)). To achieve this goal, the CWA prohibits the discharge of pollutants into navigable
waters except in compliance with the statute. The CWA directs EPA to promulgate
regulations that establish limits on the types and amounts of pollutants discharged from
various industrial, commercial, and public sources of wastewaters. In addition the CWA
requires EPA to promulgate regulations limiting pollutant discharges to sewers'flowine to
Publicly Owned Treatment Works (POTWs). . _
POTWs provide treatment to domestic sewage and other wastewaters- the effluent
is generally discharged to surface waters and the residual material (i.e., sewage sludge) may
be used as a fertilizer or soil nutrient or disposed by a number .of practices. Sewag? sludge
is mostly water (approximately 90 percent), but also contains solids and dissolved substances
The chemical and biological constituents in sewage sludge depend on the composition of the
wastewater .that enters the treatment works and the processes used to treat the wastewater
Typically, constituents of sewage sludge include organic chemicals, organic solids nutrients'
inorganic chemicals, and disease-causing organisms or pathogens Ce.gi, bacteria, viruses and
neimmth ova).
Implementation of the CWA has resulted in higher levels of wastewater treatment
leading to both greater pollutant removal from the wastewater before discharge and generation
of larger amounts of sewage sludge. In the United States, the amount of sewage sludge
generated has almost doubled since enactment of the Water Pollution Control Act
Amendments in 1972. POTWs currently generate 5.3 million dry metric tons of sewage
sludge per year, or approximately 47 pounds per person per year. Proper management of the
use or disposal of the sewage sludge is important.
A POTW has a number of alternative practices for the use or disposal of sewage
sludge. These practices include, but are not limited to, land application, surface disposal
placement in a municipal solid waste landfill, and incineration: The requirements that have
to be met for all of these practices are described in the Standards for the Use or Disposal of
Sewage Sludge (40 CFR Part 503), which were published under the authority of section
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405(d) of the CWA. These standards are known as the "Round One sewage sludge
regulation."
The Federal Water Pollution Control Act Amendments of 1972 regulated sewage
sludge use or disposal in only one limited circumstance: when the use or disposal poses a
threat to navigable waters. Section 405(a) of the CWA prohibited the disposal of sewage
sludge if it would result in any pollutant from the sewage sludge entering navigable wasters
unless hi accordance with a permit issued by the EPA Administrator. In 1977, Congress
amended section 405 to add a new subsection 405(d) that required EPA to develop regulations
for the use or disposal of sewage sludge. The regulations had to establish guidelines that:
(1) identify uses for sewage sludge, including disposal; (2) specify factors to be taken into
account hi determinhig the methods and practices applicable to each use or disposal practice;
and (3) identify concentrations of pollutants that interfere with each use or disposal practice.
In 1987, Congress amended section 405 again and for the first tune set forth a
comprehensive program to reduce the potential public health and environmental risks from
sewage sludge and to maximize the beneficial use of sewage sludge. Amended section 405(d)
established a timetable for the development of the sewage sludge use or disposal guidelines.
The basis for the program Congress mandated to protect public health and the environment
is the development of technical requirements or standards for sewage sludge use or disposal
and the implementation of the standards, in part through a permit program.
Under the current section 405(d), EPA must first identify, based on available
information, toxic pollutants that may be present in sewage sludge in concentrations that may
adversely affect public health and the environment. Next, for each identified use or disposal
practice,- EPA must promulgate regulations that specify acceptable management practices and
numerical limits for sewage sludge that contains these pollutants. The management practices
and numerical limits must be "adequate to protect public health and the environment from any
reasonably anticipated adverse effects of each pollutant. " Section 405(d) requires that EPA
promulgate the sewage sludge regulations in two rounds and review the regulations
periodically to identify additional toxic pollutants for regulation.
On February 19, 1993, EPA published the Round One sewage sludge regulation
in the Federal Register (58 FR 9248). Subsequently, the regulation was amended on
February 24, 1994 (59 FR 9095) and on October 25, 1995 (60 FR 54764).
A candidate- list of pollutants for the second round of the sewage sludge
regulations (i.e., Round Two) was developed hi May 1993 in accordance with an agreement
filed with the District Court hi Oregon (Gearhart et al. v. Browner, Civ. No. 89-6266-HO,
D. Oregon). The final list of pollutants was provided to the District Court in Oregon in
November 1995, and the Round Two regulation is scheduled for proposal .hi December 1999
and for publication hi December 2001 . The Round Two regulation will be an amendment to
the Round One regulation.
€>
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1.2 PURPOSE
The purpose of this document is to provide information on the list of pollutants for
the Round Two sewage sludge regulation. It discusses the Round One regulation (Chapter
2), describes the methodology used to evaluate pollutants for inclusion on the candidate list
of Round Two pollutants and presents the candidate list of Round Two pollutants (Chapter
3), describes the methodology used to develop the final list of Round Two pollutants
(Chapters 4 and 5), and presents the final list (Chapter 6).
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2. ROUND ONE SEWAGE SLUDGE REGULATION
2.1 DESCRIPTION OF A PART 503 STANDARD
The Round One sewage sludge regulation established standards for the use or
disposal of sewage sludge when the sewage sludge is applied to the land (including sewage
sludge sold or given-away in a bag or other container for application to the land), placed In
a surface disposal site, placed in a municipal solid waste landfill, or incinerated. In general.
the rule does not apply to the processing of sewage sludge before use or final disposal.
A Part 503 standard contains seven elements: general requirements, pollutant
.limits, management practices, operational standards, and frequency of monitoring.
recordkeeping, and reporting requirements. These elements are designed to protect public
health and the environment (only public health was evaluated in the case of incineration) from
the reasonably anticipated adverse effects of pollutants in sewage sludge.
The general requirements hi a Part 503 standard contain what are often called
"administrative requirements." For example, the general requirements for land application
require the preparer of the sewage sludge to notify the permitting authority about the
interstate transfer of sewage sludge that is land applied. Without the general requirements.
a Part 503 standard is incomplete. Thus, it is important that the regulated community
understand and follow the Part 503 general requirements. Note that the general requirements
are different for land application, surface disposal, and incineration.
Specific pollutant limits are established separately for each regulated use or
disposal practice. EPA developed these limits based on the results of a risk assessment. For
example, in the case of land application, up to 14 exposure pathways were evaluated in the
risk assessment for each pollutant. The most stringent value for all of the exposure pathways
for a pollutant was used to determine the limit for that pollutant in the Part 503 land
application subpart.
Management practices contain requirements that have to be met at a use or
disposal site, or, in the case of incineration, requirements for the incinerator. For example,
one of the surface disposal management practices indicates that to protect public health, food
crops cannot be grown on a surface disposal site unless otherwise authorized by the
permitting authority.
Operational standards are technology-based requirements that in the judgement of
the EPA Administrator protect public health and the environment. Section 405(dX3) of the
CWA indicates that when risk-based pollutant limits and management practices cannot be
developed, requirements such as an operational standard can be utilized. For land application
and surface disposal, the pathogen and vector attraction -reduction requirements are
operational standards. For incineration, the allowable concentration of total hydrocarbons in
the exit gas from the incinerator stack is an operational standard.
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The final three elements of a Part 503 standard help make the Round One regulation
self-implementing. This means that the Part 503 requirements have to be met even in the
absence of a federal permit. The frequency of monitoring requirements establish the
frequency at which sewage sludge samples have to collected and analyzed, while the
recordkeeping requirements indicate who has to keep what records and for how long.
Finally, the reporting requirements indicate who has to report information to the permitting
authority annually.
2.2 SEWAGE SLUDGE USE OR DISPOSAL PRACTICES
This section discusses the use or disposal practices for which requirements are
included in the Round One sewage sludge regulation. These practices include land
application, surface disposal, and incineration. These same use or disposal practices were
evaluated during the development of the final list of pollutants for the Round Two regulation,
as discussed in more detail in Chapter 4. Note that the placement of sewage sludge in a
municipal solid waste landfill is not considered under Round Two, because the requirements
have already been established by the Solid Waste Disposal Facility Criteria (40 CFR Part 258
and Part 503.4).
2.2.1 Land Application . ,
Approximately 33 percent of the sewage sludge generated by POTWs is used to
condition the soil or to fertilize crops grown in the soil. Approximately 22 percent of the
sewage sludge generated by POTWs is applied to agricultural land (including pasture and
range land). Approximately 11 percent of the sewage sludge generated by POTWs is applied
to forests, public contact sites (e.g., parks, ball fields, and highway median strips), and
reclamation sites, or is sold or given away in a bag or other container for home gardens. The
method ofapply ing sewage sludge to the land depends on the physical characteristics of the
sewage sludge and the soil, and on the types of crops or vegetation grown. Liquid sewage
sludge may be applied to the land surface or may be injected below the land surface.
Dewatered sewage sludge is usually applied to the soil surface and either is left on the soil
surface or incorporated into the soil by plowing or disking.
Ten or more States have undertaken sewage sludge application to forests, at least on
an experimental field-scale level. The most extensive experience with this practice is in the
Pacific Northwest. Usually, partially dewatered sewage sludge is sprayed onto established
forest stands using mobile equipment.
Sewage sludge that is sold or given away in a bag or other container for application
to the land often is composted with another material or heat-dried. The sewage sludge most
often is used to fertilize lawns and home gardens.
.2.2.2 Surface Disposal
Surface disposal is the term used to describe the placement of sewage sludge on land
for final disposal. Surface disposal is used to describe a number of different practices,
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including disposal in surface impoundments (into which liquid sewage sludge is placed).
disposal in sludge-only landfills (monofills), and disposal in piles left on land surfaces. These
practices account for the disposal of approximately ten percent of the sewage sludge generated
by POTWs. ......
2.2.3 Incineration
Incineration is a practice that reduces the volume of sewage sludge through
combustion. A sewage sludge incinerator is an enclosed device in which sewage "sludge and
auxiliary fuel are fired at high temperatures. During the incineration process, emissions are
discharged to the atmosphere. The residual ash that is-generated is not sewage sludge, by
definition.
Approximately 16 percent of the sewage sludge generated by POTWs (i.e.,
865,000 dry metric tons) is fired in a sewage sludge incinerator. Most of the sewage sludge
incinerators employ the multiple hearth technology. Other types of sewage sludge
incinerators include multiple hearth incinerators with a secondary combustion chamber,
fluidized bed incinerators, and electric incinerators.
2.3 DATA GATHERING STUDIES
The results of several EPA-sponsored data gathering studies were used to support
development of the Round One sewage sludge regulation. These include the "40 City Study"
(U.S. EPA, 1982), Environmental Profiles and Hazard Indices (U.S. EPA, 1985), sewage
sludge incinerator field studies, and the National Sewage Sludge Survey (U.S. EPA, 1990a).
These studies are described below.
2.3.1 The 40 City Study
During 1979 and 1980, EPA conducted a study, known as the "40 City Study"
(U.S. EPA, 1982), to determine the fate of priority toxic pollutants in POTWs. Wastewater
samples were collected at 40 POTWs throughout the United States and analyzed for priority
toxic pollutants. Analytical data from the study were used to determine whether priority
pollutants were in the influent to POTWs and, if they were, what happened to them as they
moved through the POTW.
At several of the POTWs where wastewater samples were collected, sewage
sludge samples also were collected and analyzed for priority toxic pollutants. Analytical data
for the sewage sludge samples were used to determine the concentrations of 40 pollutants (12
metals, six base neutral organic compounds, six volatile organic compounds, nine pesticides,
and seven polychlorinated biphenyls) in sewage sludge. These concentrations were used to
develop the Environmental Profiles and Hazard Indices discussed below.
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2.3.2 Environmental Profiles and Hazard Indices
In the spring of 1984, EPA enlisted the assistance of Federal, State, academic, and
private sector experts to determine which pollutants likely to be found in sewage sludge
should be examined closely as possible candidates for Part 503 pollutant limits. These
experts screened a list of 185 pollutants in sewage sludge that, if used or disposed
improperly, could cause adverse public health or environmental effects. The experts were
then requested to revise the list, adding or deleting pollutants, based on the potential risk to
public health and the environment when sewage sludge containing a particular pollutant was
applied to the land, placed in a surface disposal site, or incinerated. The Agency also
requested that the experts identify the most likely route that a pollutant would travel to reach
target organisms, whether human, plant, or animal. The experts recommended that the
Agency gather additional information on approximately 50 pollutants. During 1984 and 1985,
the Agency collected data and information from published scientific reports on the toxiciry,
persistence, means of transport, and environmental fate of the 50 pollutants.
EPA made an assessment of the likelihood that each pollutant would adversely affect
public health or the environment by using data from the'40 City Study on the frequency of
detection and concentration of pollutants hi sewage sludge and preliminary information on
pollutants' toxicity and persistence, the pathways by which the pollutants travel through the
environment to a receptor organism (human, animal, or plant), the mechanisms that transport
or bind the pollutants in the pathway, and the effects of pollutants on the target organism.
For these analyses, EPA relied on simple screening models and calculations to predict the
exposure to each .pollutant that would occur if the pollutant were present hi surface or ground
water, soil, air, or food. EPA then developed a hazard index for each pollutant by comparing
the predicted exposure to an Agency human health criterion, such as a drinking water
standard promulgated under the Safe Drinking Water Act, or an animal or plant toxicity
number, to determine whether the pollutant could be expected to have an adverse effect on
public health or the environment. For purposes of this initial screening, EPA assumed: (1)
conditions that would maximize the exposure of an individual human, animal, or plant to the
pollutant and (2) the worst possible toxicity effects of the pollutant.
Using the hazard indices, EPA scored each pollutant and ranked them for more
rigorous analysis. Two categories of pollutants were excluded from further evaluation. First,
EPA excluded pollutants that had a hazard index of less than one (i.e., presented no risk to
public health or the environment at the highest concentration that the Agency found in the "40
City Study" or hi other available databases). Second, EPA deferred consideration of
pollutants for which EPA lacked human health criteria or sufficient data. An environmental
profile for each pollutant was then developed by combining environmental fate, transport, and
toxicity data on each pollutant, the results of the simple screening models and calculations,
and the hazard indices for that pollutant (U.S. EPA, 1985). The environmental profiles were
used to determine the pollutants for which limits would be developed through a detailed risk
assessment of exposure pathways.
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2.3.3 Sewage Sludge .Incinerator Field Studies
In 1987, the Agency initiated a series of field studies at ten sewage sludge incinerators
to support the Part 503 rulemaking effort. The purposes of the on-site tests were to obtain:
(1) information about the percentage of hexavalent chromium in the total chromium in the exit
gas from a sewage sludge incinerator, (2) information on the percentage of nickel subsulfide
in the total nickel in the exit gas from a sewage sludge incinerator, (3) total hydrocarbon
(THC) emissions data for the sewage sludge incinerators, and (4) information about organic
compounds hi the exit gas from a sewage sludge incinerator.
Data from the .studies on the exit gas THC concentrations were used as the basis for
the THC operational standard hi the Round One sewage sludge regulation. This standard is
technology-based in that it is based on performance data from sewage sludge incinerators
Test data indicated that there is a significant correlation between THC and organic compound
concentrations hi the exit gas, which is important because sampling and analysis techniques
are not available to identify or quantify all potential organic compounds emitted from sewage
sludge incinerators, nor are toxicity data available for all compounds.
2.3.4 National Sewage Sludge Survey
In 1988, EPA conducted the National Sewage Sludge Survey (NSSS) to collect
information and data necessary to produce national estimates of: (1) concentrations of toxic
pollutants in sewage sludge, (2) sewage sludge generation and treatment processes, (3) sewage
sludge use or disposal practices, and (4) sewage sludge treatment and use or disposal costs.
The NSSS consisted of a questionnaire survey and an analytical survey The sample
for the questionnaire survey was selected from the 11,407 POTWs in the United States
Puerto Rico, and the District of Columbia, identified in the 1986 Needs Survey as having at
least secondary treatment. Secondary treatment was defined as primary clarification followed
by biological treatment and secondary clarification. The sample for the analytical survey was
a subset of the sample for the questionnaire survey.
. POTWs m me two samples were selected using a stratified probability design based
on 24 mutually exclusive groups. Membership in these groups was based on four categories
of wastewater flow rates and six use or disposal practices. The flow rates and use or disposal
categories were as follows: v
\
POTW average daily flow rate categories: ,
• Flow rate less than or equal to one million gallons per day (MOD)
• Flow rate greater than one MGD, but less than or equal to 10 MOD
• Flow rate greater than 10 MGD, but less than or equal to 100 MGD
• Flow rate greater than 100 MGD
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POTW sewage sludge use or disposal practice groups:
• Land application
• Distribution and marketing (now called sale or give-away in a bag or
other container for application to the land)
• Incineration
• Monofill (sewage sludge-only landfill)
• Co-disposal landfill
• Ocean disposal
For the questionnaire survey, a 50 page questionnaire was mailed to 479 POTWs.
Information collected through the questionnaire included service area, POTW operating data,
sewage sludge use or disposal practices, pretreatment activities, wastewater and sewage
sludge testing frequencies, and POTW financial information. In addition to the six use or
disposal practices used for the stratified sampling, information was collected on co-
incineration and surface disposal.
POTWs in the analytical survey were restricted to POTWs in the contiguous States
and the District of Columbia. As mentioned above, the POTWs were a subset of the
POTWs in the questionnaire sample. Sewage sludge samples were collected at 208 POTWs
from the four flow rate categories and analyzed for 411 analytes. (Note that in 58 FR 9269,
it is stated that samples were analyzed for a total of 412 analytes. In Appendix C-List of
NSSS Analytes hi the NSSS Data Element Dictionary for the Questionnaire and Analytical
Data Bases, (U.S. EPA, 1990a), a total of 413 analytes are listed. However, the pollutants
crotoxyphos and phosphorus were listed twice in Appendix C (crotoxyphos under both
semivolatile organics and pesticides/herbicides; phosphorus under both metals and classicals).
Therefore samples were analyzed for a total of 411 analytes.) Results of the NSSS are
discussed in Chapter 3.
2.4 ROUND ONE POLLUTANTS
The Round One sewage sludge regulation included limits (or equations for calculating
these limits) for 11 pollutants for one or more use or disposal practices. Not every pollutant
is regulated for each practice. When the rule was proposed, limits were developed for 28
pollutants for one or more use or disposal practices. The final Round One regulation contains
limits for the following pollutants:
Land application: arsenic, cadmium, chromium (deleted from the land application
subpart on October 25, 1995), copper, lead, mercury,
molybdenum, nickel, selenium, and zinc
Surface disposal: arsenic, chromium, and nickel
Incineration: arsenic, cadmium, chromium, lead, nickel, and total
hydrocarbons. Beryllium and mercury are regulated through a
National Emission Standard
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3.CANDIDATE LIST OF ROUND TWO POLLUTANTS
3.1 SELECTION PROCESS
The Clean Water Act requires EPA to: (1) identify pollutants that may be present
in sewage sludge in concentrations that may adversely affect public health and the
environment, and (2) promulgate standards, including numerical limits for the identified
pollutants. To select the pollutants on the candidate list of Round Two pollutants a
Preliminary Hazard Identification study was conducted. The first step in the study was to
determine the pollutants not regulated in the Round One sewage sludge regulation that are
detected frequently in sewage sludge. This was done using data from the 1988 NSSS
Analytical results for the 411 pollutants for which samples were analyzed in the NSSS were
reviewed to determine the frequency of detection for each pollutant. Pollutants with a
frequency of detection of less than ten percent were deleted from further consideration for
the Round Two pollutant list, if the results of a toxicity analysis did not indicate potential for
adverse public health effects (see Appendix A). If a pollutant had a frequency of detection
of ten percent or greater, the second step in the Preliminary Hazard Identification study was
conducted. . . - . . .
In the second step, scientific literature was reviewed to determine whether there
were toxicity data for a pollutant. If no human health or ecological data were found for a
pollutant, no further consideration was given to that pollutant for the Round Two list.
Pollutants for which sewage sludge samples were analyzed in the NSSS are
presented in Section 3.2. Section 3.3 presents the results of the Preliminary Hazard
Identification study for pollutants with a frequency of detection of ten percent or greater and
for which toxicity data were available. .is«*uerdna
3.2 NATIONAL SEWAGE SLUDGE SURVEY POLLUTANTS
Sewage sludge samples collected during the NSSS were analyzed for 411 analvtes
as shown in Exhibit 3-1 at the end of this section. These analytes included inorganics a^
well as eveiy organic, pesticide, dibenzofuran, dioxin, and polychlorinated biphenyl (PCB)
for which EPA had gas chromatography and mass spectrometry (GC/MS) standards The
pollutants also were selected in consideration of the CWA priority toxic pollutants'" toxic
pollutants identified in the Domestic Sewage Study (U.S, EPA? 1986a>! and Resource
Conservation and Recovery Act (RCRA, Pub. L. 94-580) appendix VIE pollutants.
Three categories of POTWs were excluded from calculations of national estimates
of pollutant concentrations in sewage sludge: Primary Samples Only, No Sludge Sampled
SSs?? flBusmeAss/Ineligible- ,In addition, data for one POTW were not entered into the
NSSS database. As a result, data from only 176 of the 208 POTWs sampled were used to
estimate pollutant concentrations. The rationale for excluding the three categories is
described below:
3-1
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Primary Samples Onty - Samples of sewage sludge generated during less than
secondary treatment were collected from some POTWs; for two POTWs, they
were the only type of sample collected. These data were not suitable to use in
determining the national estimates for pollutant concentrations in sewage sludge
generated during secondary or more-advanced treatment.
No Sludge Sampled - No sewage sludge samples were collected at 24 of the
selected POTWs. For 18 of those POTWs, pre-sampling phone contacts indicated
that sewage sludge was not being used or disposed at these POTWs during the
time of the NSSS. These 18 POTWs claimed to treat their wastewater in
stabilization ponds; sewage sludge generated in a stabilization pond remains in the
pond until it is removed for final use or disposal through some other practice.
Out of Business/Ineligible - Five of the selected POTWs were either closed or
could not be contacted by telephone. Therefore, no data were available for those
POTWs because no samples could be collected.
Sewage sludge sampling, preservation, and analytical protocols were developed
specifically for the NSSS. Analytical methods 1624 and 1625 were developed for volatile and
semi-volatile organics, respectively, hi the sewage sludge matrix; these methods use gel
permeation chromatography sample clean-up followed by isotope dilution GC/MS analyte
identification and quantification. Pesticides and PCBs, and dibenzofurans/dioxins, were
analyzed using analytical methods 1618 and 1613, respectively. Metals, other inorganics, and
classicals were analyzed using standard FJPA methods. The analytical methods were
developed or adapted specifically for the sewage sludge matrix to give the most reliable,
accurate, and precise measurements of the 411 analytes.
Each NSSS sample was tested by EPA contract laboratories for 411 pollutants.
A minimum level, one land of "detection limit" used by the Agency, was identified for each
pollutant in the protocol of the analytical method. That minimum level, as applied to the
determination of pollutants by GC/MS, is defined by the EPA?s Engineering and Analysis
Division'in the Office_ of Water as the level at which the entire analytical system shall give
recognizable mass spectra and acceptable calibration points. In the NSSS, the minimum level .
is equivalent to the minimum concentration of pollutant that could be measured.
Pollutant concentrations arid minimum levels were reported in dry weight units
due to differences hi the percent solids in sewage sludge samples; percent solids ranged from
less than one percent to 100 percent in NSSS samples. The use of dry weight units allows
all sewage sludges to be evaluated on an equivalent basis. Implicit in this form of reporting
is that pollutants are associated with the solid phase of sewage sludge.
For any given pollutant, the minimum levels varied due to varying volumes or
masses of samples tested, dilution of the sample or extract, and matrix effects or interference.
All analytical protocols specified the volume or amount of sewage sludge to be tested. When
matrix interferences prevented accurate determination of pollutant concentrations, however,
samples were diluted with reagent water and then reanalyzed. The minimum level for a
diluted sample was then raised by the dilution factor. For example, if a sample was diluted
3-2
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by a factor of 10, then the minimum level was raised by a factor of 10. Analytical protocols
provided explicit guidance as to the limits of dilution.
For each pollutant, EPA statisticians calculated the frequency of occurrence as
weir as the median and other percentile concentrations. Non-detected values were treated in
two ways to capture the full range of possible concentrations: nondetects were set to both the
minimum level and to zero. These national pollutant concentration estimates reflect the
distribution of pollutant concentrations in dry weight sewage sludge that is generated by
secondary or more advanced wastewater treatment at POTWs. Appendix B provides details
on the statistical analyses performed on the NSSS data.
3-3
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EXHIBIT 3-1
Pollutants Analyzed in the National Sewage Sludge Survey
Pollutant1
Acenaphthene
Acenaphthylene
Acetic acid, 2,4-dichlorophenoxy
Acetophenone
Acrylonitrile
Aldrin
Aluminum
Aminobiphenyl, 4-
Aniline
Aniline, 2,4,5-trimethyl-
Anisidine, o-
Anthracene
Antimony
Aramite
Arsenic
Azinphos ethyl
Azinphos methyl
Barium
Benzanthrone
Benzene
Benzenethiol
Benzidine
Benzofluorene, 2,3-
Benzoic acid
Benzonitrile, 3,5-dibromo-4-hydroxy-
Benzoquinone, 2,6-di-tert-butyl-p-
Benzothiazole, 2-(methylthio)
Benzo(a)pyrene
Benzo(b)fluoranthene
Benzo(ghi)perylene
Benzo(k)fluoranthene
Benzyl alcohol
Benz(a)anthracene
Beryllium
BHC, alpha-
BHC, beta-
BHC, delta-
Biphenyl
Biphenyl, 4-nitro
Bismuth
Bis(2-chloroethoxy)methane
Bis(2-chloroethyl)ether
Bis(2-chloroisopropyl)ether
Bis(2-ethylhexyl)phthalate
Boron - ,
Bromodichloromethane
Bromomethane
Butadiene, 2-chloro-l,3-
Butanone, 2-
Butene, trans-l,4-dichloro-2-
Butyl benzyl phthalate
Cadmium
Calcium
Captafol
I Captan
J Carbazole
Carbon disulfide
Carbophenothion (Trithion)
Cerium
Chlordane
Chloroacetonitrile
Chloroaniluie, p-
Chlorobenzene
Chlorobenzene, l-bromo-2-
Chlorobenzene, l-bromo-3-
Chlorobenzilate
Chloroethane
Chloroethylvinyl ether, 2-
Chlorofenvinphos
Chloroform
Chloromethane
Chloronaphthalene, 2-
Chlorophenol, 2-
Chloropropene, 3-
Chlorpyrifos
Chromium
3-4
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Pollutant1
Pollutant1
cnrysene «
Ciodrin
Cobalt
Copper
Coumaphos
Cresol, o-
Cresol, p-
Crotonaldehyde
Crotoxyphos
Cyanides (soluble salts and complexes)
Cymene, p-'
DDD
DDE
DDT
Decane, n-
Demeton
Diallate
Diazinon
Dibenzofuran
Dibenzothiophene
Dibenz(a,h)anthracene
Dibromochloromethane
Dibromoethane, 1,2-
Dibromomethane
DicMoroaniline, 2,3-
Dichlorobenzene, 1,2-
Dichlorobenzene, 1,3-
Dichlorobenzene, 1,4-
Dichlorobenzidine, 3,3'-
Dichloroethane, 1,1-
Dichloroethane, 1,2-
Dichloroethene, 1,1-
Dichloroethene, trans-1,2-
Dichloronitrobenzene, 2,3-
Dichlorophenol, 2,4-
Dichlorophenol, 2,6-
Dichloropropane, 1,2-
Dichloropropane, 1,3-
Dichloropropene, cis-1,3-
Dichloropropene, trans-1,3-
Dichlorvos
Dicrotophos (Bidrin)
Dieldrin
Diepoxybutane, 1,2:3,4-
Diethyl ether
Diethyl phtbalate
Dimethoate
Dimethoxybenzidine, 3,3'-
Dimethyl phtnalate
Dimethyl sulfone
Dimethylaminoazobenzene, p-
Dimethylbenz(a)anthracene, 7,12-
Dimethylphenanthrene, 3,6-
Dimethylphenol, 2,4-
Dinitrobenzene, 1,4-
Dinitrophenol, 2,4-
Dmitrophenol, 2Tsec-butyl-4,6-
(Dinoseb)
Dinitrotoluene, 2,4-
Dmitrotoluene, 2,6-
Dioxane, 1,4-.
Dioxathion
Diphenyl ether
Diphenylamine
Diphenyldisulfide
Diphenylhydrazuie, 1,2-
Disulfoton
Di-n-butyl phtnalate
Di-n-octyl phthalate
Di-n-propyhiitrosamme
Docosane, n-
Dodecane, n-
Dysprosium
Eicosane, n-
Endosulfan sulfate
Endosulfan-I
Endosulfah-n
3-5
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EXHIBIT 3-1
Pollutants Analyzed in the National Sewage Sludge Survey (cont'd.)
Pollutant1
Endnn
Endrin aldehyde
Endrin ketone
Erbium
Ether, 4-bromophenylphenyl
Ether, 4-chlorophenylphenyl
Ethion
Ethyl cyanide '
Ethyl methacrylate
Ethyl methanesulfonate
Ethylbenzene
Ethylenethiourea
Europium
Famphur
Fensulfothion
Fenthion
Fluoranthene
'Fluorene
Fluoride
Gadolinium
Gallium
Germanium
Gold
Hafnium ~ • •>
Heptachlor
Heptachlor epoxide
Heptachlorodibenzofuran, 1,2,3,4,6,7,8-
Heptachlorodibenzofuran, 1,2,3,4,7,8,9-
Heptachlorodibenzo-p-dioxhi, 1,2,3,4,6,7,8-
Hexachlorobenzene
Hexachlorobutadiene
Hexachlorocyclopentadiene
Hexachlorodibenzofuran, 1,2,3,4,7,8-
Hexachlorodibenzofuran, 1,2,3,6,7,8-
Hexachlorodibenzofuran, 1,2,3,7,8,9-
Hexachlorodibenzofuran, 2,3,4,6,7,8-
Hexachlorodibenzo-p-dioxin, 1,2,3,4,7,8-
Hexachlorodibenzo-p-dioxin, 1,2,3,6,'7,'s-
PoUutant1
lexachlorodibenzo-i
Hexachloroethane
Hexachlorbpropene
Hexacosane, n-
Hexadecane, n-
Hexanoic acid
Hexanone, 2-
Holmium
Indeno(l ,2,3-cd)pyrene
Indium
Iodine
lodomethane
Indium
Iron
Isobutyl alcohol
Isodrin
Isophorone
Isopropylnaphthalene, 2-
Isosafrole
Lanthanum
Lead
Leptophos
Lindane
Lithium
Longifolene
Lutetium
Magnesium
Malachite green
Malathion
Manganese
Mercury
Mestranol
Methapyrilene
Methoxychlor
Methyl methacrylate
Methyl methanesulfonate
Methyl parathion
Methylbenzothioazole, 2-
lioxin,
3-6
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EXHIBIT 3-1
Pollutants Analyzed in the National Sewage Sludge Survey (cont'd.)
Pollutant1
Pollutant1
Methylcholanthrene, 3-
Methylene chloride
Methylene phenanthrene, 4,5-
Methylene bis(2-chloroaniline), 4,4'-
Methylfluorene, 1-
Methyhiaphthalene, 2-
Methylphenanthrene, 1-
Methylphenol, 4-chloro-3-
Mevinphos (Phosdrin)
Mirex
Molybdenum
Monocrotophos
N,N-dimethylfonnamide
Naled (Dibrom)
Naphthalene
Naphthalenediamine, 1,5-
Naphthoquinone, 1,4-
Naphthoquinone, 2,3-dichloro-l,4-
Naphthylamine, 1-
Naphthylamine, beta-
Neodymium
Nickel
Niobium
Nitrate
Nitrite
Nitroaniline, 2,6-dichloro-4-
Nitroanilhie, 2-
Nittoaniluie, 3-
Nitroaniline, 4-chloro-2-
Nitroaniline, p-
Nitrobenzene
Nitrobenzene, l-chloro-3-
Nittofen (TOK)
Nitrophenol, 2-
Nitrophenol, 4-
N-nitrosodiethylamine
N-nitrosodimethylamine
N-nitrosodiphenylamine
N-nitrosodi-N-butylamine
N-nitrosomethylethylamine
N-nitrosomethylphenylamine
N-nitrosomorpholine
I N-nitrosopiperidine
I Octachlorodibenzofuran
Octachlorodibenzo-p-dioxin
Octacosane, n-
Octadecane, n-
Osmium
Palladium
Parathion
PCB-1016
PCB-1221
PCB-1232
PCB-1242
PCB-1248
PCB-1254
PCB-1260
Pentachlorobenzene
Pentachlorodibenzofuran, 1,2,3,7,8-
Pentachlorodibenzofuran, 2,3,4,7,8-
Pentachlorodibenzo-p-dioxin, 1,2,3,7,8-
Pentachloroethane
Pentachloronitrobenzene
Pentachlorophenol
Pentamethylbenzene
Pentanone, 4-methyl-2-
Perylene
Phenacetin
Phenanthrene
Phenol
Phenol, 2-methyl-4,6-dinitro-
Phenothiazine
Phenymaphthalene, 1-
Phenymaphthalene, 2-
Phorate
Phosmet
3-7
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EXHIBIT 3-1
Pollutants Analyzed in the National Sewage Sludge Survey (cont'd.)
Pollutant1
Pollutant1
Phosphamidon
Phosphoric acid, trimethyl ester
Phosphoric acid, tri-o-tolyl ester
Phosphoric triamide, hexamethyl-
Phosphorus
Picoline, 2-
Platinum
Potassium
Praseodymium
I Pronamide
Propane, 1,2-dibromq-3-chloro-
Propanol, l,3-dichloro-2-
Propanone, 2-
Propenal, 2-
Propenenitrile, 2-methyl-2-
Propen-1-ol, 2-
Propionic acid, 2-(2,4,5-trichlorophenoxy)
(Silvex)
Pyrene
Pyridine
Resorcindl
Rhenium
Rhodium .
Ruthenium
Safrole
Samarium
Santox (EPN)
Scandium
Selenium
Silicon
Silver
Sodium
Squalene
Strontium
Styrene
Sulfiir
Tantalum
Tellurium
Terbium
Terbufos ~
Terpineol, alpha-
Tetrachlorobenzene, 1,2,4,5-
Tetrachlorodibenzofuran, 2,3,7,8-
Tetrachlorodibenzo-p-dioxin, 2,3,7,8-
Tetrachloroethane, 1,1,1,2-
Tetrachloroethane, 1,1,2,2-
Tetrachloroethene
Tetrachloromethane
Tetrachlorophenol, 2,3,4,6-
Tetrachlorvniphos
Tetracosane, n-
Tetradecane, n-
Tetraethyldithiopyrophosphate
Tetraethylpyrophosphate
Thallium
Thianaphthene
Thioacetamide
Thioxanthe-9-one
Thorium
Thulium
Tin
Titanium
Toluene
Toluene, 2,4-diamino-
Toluidine, 5-chloro-o-
Toluidme, 5-nitro-o-
Toluidine, o-
Total heptachlorodibenzofurans
Total heptachlorodibenzo-p-dioxms
Total hexachlorodibenzofurans
Total hexachlorodibenzo-p-dioxins
Total kjeldahl nitrogen
Total pentacMorodibenzoiurans
Total pentachlorodibenzo-p-dioxins
Total residue
Total tetrachlorodibenzofurans
Total tetrachlorodibenzo-p-dioxms
3-8
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EXHIBIT 3-1
Pollutants Analyzed in the National Sewage Sludge Survey (cont'd.)
Pollutant1
Ioxapnene
Triacontane, n-
Tribromomethane
Trichlorobenzene, 1,2,3-
•Trichlorobenzene, 1,2,4-
Trichloroethane, 1,1,1-
Trichloroethane,. 1,1,2-
Trichloroethene
Trichlorofluoromethane
Trichlorofon
Trichlorophenol, 2,3,6-
Trichlorophenol, 2,4,5-
Trichlorophenol, 2,4,6-
Trichlorophenoxyacetic acid, 2,4,5-
Trichloropropane, 1,2,3-
Trifluralin (Treflan)
Pollutant1
inmetnoxybenzene, 1,2,3-
Triphenylene
Triprppyleneglycol methyl ether
Trithiane, 1,3,5-
Tungsten
Uranium
Vanadium
Vinyl acetate
Vinyl chloride
Xylene, m-
Xylene, o- and p-
Ytterbium
Yttrium
Zuic
Zirconium
1 This exhibit presents-the 411 specific analytes in the NSSS. In 58 ^ 9269, it is stated that samples were analyzed
for a total of 412 analytes. However, in Appendix C-List of NSSS Analytes in U.S. EPA, 1990a, there are 411
unique analytes out of 413 listed (crotoxyphos and phosphorus are listed twice).
3-9
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3.3 RESULTS OF PRELIMINARY HAZARD IDENTIFICATION STUDY
The purpose of the Preliminary Hazard Identification study was to identify a list
of Round Two pollutant candidates that would be further evaluated in the Comprehensive
Hazard Identification study. Pollutants analyzed in the NSSS formed the initial set of
candidates for Round Two regulation. This section discusses the methods used to reduce the
initial set of candidates to the list of Round Two pollutant candidates submitted to the Court
in May, 1993. These methods included removing pollutants already regulated and pollutants
with analytical problems; combining pollutant congeners into classes; removing pollutants
based on frequency of detection data; and ascertaining the availability of human and
ecological toxicity data.
3.3.1 Pollutants Removed from Further Consideration
Several NSSS pollutants were removed from further consideration for Round Two
Total residue was deleted because it is the inert material left after all the pollutants have been
extracted from sewage sludge. Total kjeldahl nitrogen, the sum of the concentrations of
ammonia (NH3) and organic nitrogen compounds in the trinegative oxidation state (N(-HI))
was deleted because the concentrations of the individual compounds cannot be discerned and
because nitrogen is already controlled through management practices hi Part 503 The
herbicide Dinoseb (2-sec-butyl-4,6-dinitrophenol) could not be analyzed in the NSSS
Because all U.S. uses of this herbicide have been cancelled, Dinoseb was eliminated from"
further consideration under Round Two. Lastly, the ten individual pollutants regulated under
Round One, shown in Exhibit 3-2, were not considered to be pollutant candidates for Round
Two because they were already regulated. This does not preclude possible re-evaluation of
the Round One pollutants hi the Round Two regulation, however.
3.3.2 Individual Pollutants Combined Into Classes
Two classes of pollutants, polychlorinated biphenyls (PCBs) and
dioxms/dibenzofurans, were formed by aggregating individual pollutants (Exhibit 3-3) For
PCBs, seven Aroclor mixtures were combined into one category. For dioxins and
dibenzofurans, 25 chlorinated dioxins and dibenzofurans were combined by multiplying each.
congener's concentration by its corresponding toxicity equivalency factor (U.S. EPA 1989b)
and adding the resultant values over all congeners for each POTW.
3.3.3 Frequency of Detection of Pollutants
The Agency removed from further consideration all pollutants that had a
frequency of detection of zero percent in sewage sludge sampled in the NSSS. A total of 254
pollutants, shown in Exhibit 3-4, fell into that category.
The remaining 114 pollutants hi the NSSS were then analyzed further The
Agency used non-zero frequency of occurrence data to evaluate which of the remaining 114
pollutants should be candidates for Round Two regulation. For each pollutant in the NSSS
except dioxms/dibenzofurans, there was only one frequency of occurrence value For
dioxms/dibenzofurans, however, frequency of occurrence was calculated in two ways: (1)
3-10
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dioxins/dibenzofurans were considered to be detected if at least one of the individual
congeners was detected above the minimum level; (2) dioxins/dibenzofurans were.considered
to be detected if all of the individual congeners were detected above the minimum level.
Because all of the individual congeners were not detected above the minimum level at any
given POTW, the second method resulted in zero percent detection. At the other extreme,
the first method resulted in 100 percent detection.
Of the 114 pollutants, the Agency deleted from further consideration those pollutants
that were detected less than ten percent of the time if the results of a toxicity analysis (see
Appendix A) indicated that a pollutant did not have the potential to affect public health
adversely. EPA concluded that if a pollutant was not present in more than ten percent of the
sewage sludge samples and is not highly toxic, the potential to adversely affect public health
and the environment is low. Of the 69 pollutants deleted for this reason, 35 had a frequency
of detection of two percent or less; 22 had a frequency of detection greater than two percent,
but equal to or less than five percent; and 12 had a frequency of detection greater than five
percent, but less than ten percent. After deleting those 69 pollutants, 45 pollutants remained
as candidates for the Round Two regulation, as shown in Exhibit 3-6.
3.3.4 Available Human Health and Ecological Toxicity Data
For the 45 candidate pollutants that had frequencies of occurrence of ten percent or
greater, EPA obtained available human and ecological toxicity data. For human toxicity data,
EPA predominantly relied on IRIS (Integrated Risk Information System) and HEAST (Health
Effects Assessment Summary Tables). EPA also attempted to obtain human toxicity data
from the following sources in descending order of preference:
Reportable Quantities/Potency Factors,
California Environmental Protection Agency data,
HSDB (Hazardous Substances Data Base),
RTECS (Registry of Toxic Effects of Chemical Substances), and
Office of Pesticide Programs data sources.
For carcinogenic effects, EPA obtained both the slope potency factor, qr*, hi units of
(mg/kg-day)'1 for the oral and inhalation pathways, and the weight of evidence factors as
classified by the Cancer Risk Assessment Guidelines (U.S. EPA, 1986b). For non-
carcinogenic effects, the Risk Reference Doses (RfDs) for the oral pathway and the Risk
Reference Concentrations (RfCs) for the inhalation pathway, in units of (mg/kg-day) and
(mg/m3), respectively, were obtained. RfCs were converted to units of (mg/kg-day) by
assuming a person weighs 70 kg and breathes 20m3 of air per day. For those chemicals with
only noncarcinogenic effects, it was also noted whether any positive mutagenicity studies had
been reported. These human health toxicity data are shown hi Exhibit 3-7.
EPA's Health and Ecological Criteria Division provided several sources of ecological
toxicity data. These sources included work hi support of the Great Lakes Water Quality
Initiative (U.S. EPA, 1993b), "Implementation of a Chemical Ranking System" (U.S. EPA,
1990b), Oneliner in PIRANHA (U.S. EPA, 1991b), "Screening Study for Wildlife Criteria
Development" (U.S. EPA, 1989f), Devillers and Exbrayat (1992), Harfenist et al. (1989),
3-11
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and 40 CFR Part 131 (Federal Register, Part H, December 22, 1992). These ecological
toxicity data are shown in Exhibit 3-8. .
For 14 pollutants, shown in Exhibit 3-9, no human or ecological toxicity data were
available; these pollutants were not considered further. In addition, there were no human
toxicity data from IRIS or HEAST for the pollutant n-decane; the only available data were
from the least preferable source. Therefore n-decane also was not considered further.
Although human or ecological toxicity data were riot available for aluminum, some evidence
exists that aluminum may be a phytotoxin. Florida farmers have reported that high
background concentrations and high loadings of aluminum have led to phytotoxicity and
decreased crop yields. Therefore, aluminum was kept and evaluated in the Comprehensive
Hazard Identification study.
3.3.5 Pollutant Candidate List for Round Two Regulation
Of the remaining 30 pollutants, seven pollutants were evaluated for some use or
disposal practices under Round One; they are being re-evaluated in Round Two because they
were not evaluated for all use or disposal practices (Exhibit 3-10). Beryllium was initially
modeled only under the incineration pathway for the Round One risk assessment due to a lack
of data; it is considered as a pollutant candidate for Round Two for the land application and
surface disposal pathways. Bis(2-ethylhexyl)phthalate was not modeled under the land
application pathways hi Round One due to a lack of data, and therefore is considered as a
pollutant candidate in Round Two. Although the PCB Aroclor mixtures were evaluated for
Round One, coplanar congeners of PCBs were not evaluated; coplanar PCBs are pollutant
candidates for Round Two because they have characteristics similar to those of dioxins. Four
compounds considered hi the Environmental Profiles ranking system (discussed in Section
2.3.2), cyanide, fluoride, methylene chloride, and phenol, are considered again hi Round
Two, because previously they were not fully evaluated for all use or disposal practices. -
Even though it is not known if asbestos is present hi sewage sludge, asbestos was
raised as a possible pollutant candidate after the public comment period for the Round One
rule-was closed. NSSS sewage sludge samples were not analyzed for asbestos; friable
asbestos, however, is reported as being released under the Toxics Release Inventory. If
asbestos were present hi sewage sludge, it could potentially be released and become airborne
as the sewage sludge matrix breaks down. It could then undergo secondary deposition on soil
and water, as well as on food and feed crops. The secondary deposition pathway was never
evaluated during Round One. After including asbestos as a potential candidate, a total of 31
pollutant candidates resulted, as shown hi Exhibit 3-11.. Exhibit 3-12 summarizes the
rationale for the number of pollutants selected as candidate pollutants for the Round Two
sewage sludge regulation.
3-12
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EXHIBIT 3-2
Pollutants Regulated in Round One Sewage Sludge Regulation
Pollutant
Arsenic
Beryllium1
Cadmium
Chromium
Copper
Lead
Mercury1 •
Molybdenum
Nickel
Selenium
Zinc
Total Hydrocarbons2
1 Regulated for incineration through a National Emission Standard, which is referenced in Part 503.
2 Note that "Total Hydrocarbons" is not a pollutant explicitly listed in the 411 analytes in the NSSS. Total
Hydrocarbons encompass all organic compounds in the stack gas exiting an incinerator.
3-13
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EXHIBIT 3-3
Pollutants Combined Into Classes or Removed from Further Consideration
Number of Pollutants
Deleted
Explanation
24
7 Aroclor mixtures combined into one composite
called "PCBs".
25 chlorinated dioxins and furans combined using
Toxic Equivalency Factors into one surrogate
called "dioxins".
Total residue was removed from further
consideration because it is not a pollutant; it is the
inert material left after all the pollutants have been
extracted from sewage sludge.
Total Kjeldahl Nitrogen, the sum of the
concentrations of ammonia (NH3) and organic
nitrogen compounds hi the trinegative oxidation
state (N(-H[)), was removed from further
consideration because the concentrations of the
individual compounds cannot be discerned and
because nitrogen is controlled through
management practices hi Part 503.
Dinitrophenol, 2-sec-butyM,6- (also called
Dinoseb) was dropped from the NSSS due to
analytical problems.
33
Total number of pollutants from the NSSS
analytical survey removed from further
consideration.
3-14
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.EXHIBIT 3-4
Pollutants With a Frequency of Detection of Zero Percent in the National Sewage
Sludge Survey
Pollutant
Acenapntnene
Acenaphthylene
Acrylonitrile
Aminobiphenyl, 4-
Aniline
Aniline, 2,4,5-trimethyl-
Anisidine, o-
Aramite
Azinphos ethyl
Benzanthrone
Benzene
Benzenethiol
Benzidine
Benzo(ghi)perylene
Benzofluorene, 2,3-
Benzonitrile, 3,5-dibromo-4-hydroxy-
Benzoquinone, 2,6-di-tert-butyl-p-
Benzothiazole, 2-(methylthio)
Biphenyl, 4-nitro
Bis(2-chloroethoxy) methane
Bis(2-chloroethyl) ether
Bis(2-chloroisopropyl) ether
Bismuth
Bromodichloromethane
Bromomethane
Butadiene, 2-chloro-l,3-
Butene, trans-l,4-dichloro-2-
Captafol
Captan
Carbazole
Carbophenothion (Trithion)
Cerium
Chlordane
Cfiloroacetdnitrile
Chlorobenzene, l-bromo-2-
Chlorobenzene, l-bromo-3-
Pollutant
unioroetnane
Chloroethylvinyl ether, 2-
Chlorofenvinphos
Chloromethane
Chlorophenol, 2-
Chloropropene, 3-
Ciodrin
Coumaphos
Crotonaldehyde
Grotoxyphos
DDD
Demeton
Di-n-propyMtrosamine
Diallate
Dibenz(a,h)anthracene
Dibenzothiophene
Dibromochloromethane
Dibromoethane, 1,2-
Dibromomethane
Dichloroaniline, 2,3-
Dichlorobenzene, 1,2-
Dichlorobenzene, 1,3-
Dichlorobenzidme, 3,3'-
Dichloroethane, 1,1-
Dichloroethane, 1,2,-
Dichloroethene, 1,1-
Dichloronitrobenzene, 2,3-
Dichlorophenol, 2,4-
Dichlorophenol, 2,6-
Dichloropropane, 1,2-
Dichloropropane, 1,3-
Dichloropropene, cis-1,3-
Dichloropropene, trans-1,3-
Dichlorvos
Dicrotophos (Bidrin)
Diethyl ether
3-15
-------
EXHIBIT 3-4
Pollutants With a Frequency of Detection of Zero Percent in the
National Sewage Sludge Survey (cont'd.)
Pollutant
Diethyl phthalate ~~
Dimethoxybenzidine, 3,3'-
Dimethyl sulfone
Dimethylaminoazobenzene, p-
Dimethylbenz(a)anthracene, 7,12-
Dimethylphenanthrene, 3,6-
Dimethylphenol, 2,4-
Dinitrobenzene, 1,4-
Dinitrophenol, 2,4-
Dinitrotoluene, 2,4-
Dinitrotoluene, 2,6-
Dioxathion
Diphenyl ether
Diphenylamine
Diphenyldisulfide
Diphenyihydrazine, l,2r
Disulfoton
Dysprosium
Endosulfan sulfate
Endrin aldehyde
Endrin ketone
Erbium
Ether, 4-bromophenylphenyl
Ether, 4-chlorophenylphenyl
Ethion
Ethyl cyanide
Ethyl methacrylate
Ethyl methanesulfonate
Ethylenethiourea
Europium
Famphur
Fensulfothion
Fenthion
Fhiorene
Gadolmium
Gallium
Germanium
Pollutant
Gold ^ —
Hafnium
Heptachlor
Hexacnlorobenzene
Hexachlorobutadiene
Hexachlorocyclopentadiene
Hexachloroethane
Hexachloropropehe
Hohnium
Indeno(l ,2,3-cd)pyrene
Indium
Iodine
lodomethane
Indium
Isodrin
Isophorone
Isopropyuiaphthalene, 2-
Isosafrole
Lanthanum
Leptophos
Lindane
Lithium
Longifolene
Lutetium
Malachite green
Malathion
Mestranol
Methapyrilene
Methoxychlor
Methyl methacrylate
Methyl methanesulfonate
Methyl parathion
Methylbenzothioazole, 2-
Methylcholanthrene, 3-
Methylene bis(2-chloroaniluie), 4,4'-
Methylene phenanthrene, 4,5-
Methylfluorene, 1-
3-16
-------
EXHIBIT 3-4
Pollutants With a Frequency of Detection of Zero Percent in the
National Sewage Sludge Survey (cont'd.)
Pollutant
Pollutant
iviernyipnenanrnreryft( i-
Methylphenol, 4-chloro-3-
Mevinphos (Phosdrin)
Mirex
Monocrotophos
N-nitrosodi-n-butylamine
N-nitrosodiethylamine
N-nitrosodimethylamine
N-nitrosomethylethylamine
N-nitrosomethylphenylamine
N-nitrosopiperidine
N-nitrosomorpholine
Naphthalenediamine, 1,5-
Naphthoquinone, 1,4-
Naphthoquinone, 2,3-dichloro-l,4-
Naphthylamine, 1- .
Naphthylamine, beta-
Neodymium
Niobium
Nitroaniline, 2-
Nitroaniline, 2,6-dichloro-4-
Nitroaniline, 3-
Nitroaniline, 4-chloro-2-
Nitroaniline, p-
Nitrobenzene
Nitrobenzene, l-chloro-3-
Nitrophenol, 2-
Nitrophenol, 4-
N,N-dimethylfonnamide
Osmium
Palladium
Parathion
Pentachlorobenzene
Pentachloroethane
Pentachlorophenol
Pentamethylbenzene
Perylene
Fnenacetin
Phenol, 2-methyl-4,6-dmitro-
Phenothiazine
Phenylnaphthalene, 1-
Phenyhiaphthalene, 2-
Phorate
Phosmet
Phosphoric acid, trimethyl ester
Phosphoric triamide, hexamethyl-
Phosphorus
Platinum
Potassium
Praseodymium
Pronamide
Propane, l,2-dibromo-3-chloro-
Propanol, l,3-dichloro-2-
Propen-1-ol, 2-
Propenal, 2-
Propenenitrile, 2-methyl-2-
Pyridhie
Resorcinol
Rhenium
Rhodium
Ruthenium
Safrole
Samarium
Scandium
Silicon
Squalene
Strontium
Sulfur
Tantalum
Tellurium
Terbium
Terbufos
Tetrachlorobenzene, 1,2,4,5-
Tetrachloroethane, 1,1,1,2-
3-17
-------
EXHIBIT 3-4
Pollutants With a Frequency of Detection of Zero Percent in the
National Sewage Sludge Survey (cont'd.)
Pollutant
Tetrachloroethane, 1,1,2,2T
Tetrachlorophenol, 2,3,4,6-
Tetrachlorvinphos
Tetraethyldithiopyrophosphate
Thianaphthene
Thioacetamide
Thioxandie-9-one
Thorium
Thulium
Toluene, 2,4-diamino-
Toluidine, 5-chloro-o-
Toluidine, 5-nitro-o-
Toluidine, o-
Toxaphene
Tribromomethane
Trichlorobenzene, 1,2,3-
Trichlorobenzene, 1,2,4-
PoUutant
richloroethane, 1,1,1-
Trichloroethane, 1,1,2-
Trichlorofon
Trichlorophenol, 2,3,6-
Trichlorophenol, 2,4,5-
Trichlorophenol, 2,4,6-
Trichloropropane, 1,2,3-
Trimethoxybenzene, 1,2,3-
Triphenylene
Tripropyleneglycol methyl ether
Trithiane, 1,3,5-
Tungsten
Uranium
Vinyl acetate
Vinyl chloride
Ytterbium
Zirconium
3-18
-------
EXHIBIT 3-5
Pollutants With A Frequency of Detection of Less Than Ten Percent in the National
Sewage Sludge Survey
Pollutant
Frequency of Detection (%)
Acetophenone
Aldrin
Anthracene
Azinphos methyl
Benz(a)anthracene
Benzo(a)pyrene
Benzo(b)fluoranthene
Benzo(k)fluoranthene
Benzole acid
Benzyl alcohol
BHC, alpha-
BHC, beta-
BHC, delta-
Biphenyl
Butyl benzyl phthalate
Chloroaniline, p-
Chlorobenzene
Chlorobenzilate
Chloroform
Chloronaphthalene, 2-
Chlorpyrifos
Chrysene
Cobalt
Cresol, o-
Cymene, p-
DDE
DDT
Di-n-butyl phthalate
Di-n-octyl phthalate
Diazinon
Dibenzofuran
Dichlorobenzene, 1,4-
Dichloroethene, trans-1,2-
Dieldrin
Diepoxybutane, 1,2:3,4-
2
3
2
2
4
3
7
5
6
1
2
6
2
1
9
5
2
7
1
1
3
5
9
6
7
1
2
5
1
2
1
2
1
5
2
3-19
-------
EXHIBIT 3-5
Pollutants With A Frequency of Detection of Less Than Ten Percent
in the National Sewage Sludge Survey (cont'd.)
Pollutant
Frequency of Detection (%)
Dimethoate
Dimethyl phthalate
Dioxane, 1,4-
" Docosane, n-
Endosulfan-I
Endrin
Ethylbenzene
Fluoranthene
Heptachlor epoxide
Hexanone, 2-
Isobutyl alcohol
Methylnaphthalene, 2-
N-nitrosodiphenylamine
Naled (Dibrom)
Naphthalene
Nitrofen (TOK)
Octadecane, n-
Pentanone, 4-methyl-2-
Phenanthrene
Phosphamidon
Phosphoric acid, tri-o-tolyl ester
Picolhie, 2-
Pyrene
Santox (EPN)
Styrene
Terpineol, alpha-
Tetrachloroethene
Tetrachloromethane
Tetraethylpyrophosphate
Trichloroethene
Trichlorofiuoromethane
Trifluralhi (Treflan)
. Xylene, m-
Xylene, o- and p-
3-20
-------
EXHIBIT 3-6
Pollutants With A Frequency of Detection of Ten Percent or Greater in the National
Sewage Sludge Survey
Pollutant
Frequency of Detection (%)
Acetic acid (2,4-dichlorophenoxy)
Aluminum '
Antimony
Barium
Beryllium
Bis(2-ethylhexyl)phthalate
Boron
Butanone, 2-
Calcium
Carbon disulfide
Crespl, p-
Cyanides (Soluble salts and complexes)
Decane, n-
Dioxins
Dodecane, ri-
Eicosane, n- •
Endosulfan-n
Fluoride
Hexacosane, n-
Hexadecane, n-
Hexanoic acid
Iron
Magnesium
Manganese
Methylene chloride
Nitrate
Nitrite
Octacosane, n-
PCBs
Pentachloronitrobenzene
Phenol
Propanone, 2-
Propionic acid, 2-(2,4,5-trichlorophenoxy)
Silver
Sodium.
Tetracosane, n-
Tetradecane, n-
16
100
38
100
22
62
48
34
100
10
43
37
10
100
14
13
12
63
15
12
40
100
100
100
42
95
83
13
19
10
34
58
15
84
100
15
14
3-21
-------
EXHIBIT 3-6
Pollutants With A Frequency of Detection of Ten Percent or Greater in the
National Sewage Sludge Survey (cont'd.)
Pollutant
Frequency of Detection (%)
Thallium
Tin
Titanium
Toluene
Triacontane, n- s
Trichlorophenoxyacetic acid, 2,4,5-
Vanadium
Yttrium
15
84
98
61
14
29
62
61
3-22
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3-26
-------
EXHIBIT 3-9
Pollutants With No Human Health or Ecological Toxicity Data Available
Pollutant
Calcium
Decane, n-1
Dodecane, n-
Eicosane, n-
Hexacosane, n-
Hexadecane, n-
Hexanoic acid
Iron
Magnesium
Octacosane, n-
Sodium
Tetracosane, n-
Tetradecane, n-
Triacontane, n-
Yttrium
1 Available data from the Office of Pesticide Programs RfD Tracking Report (27 January 1992) were not
considered to be appropriate for the assessment of the Round Two list of pollutants.
3-27
-------
EXHIBIT 3-10
Round One Pollutants Included as Potential Candidates For Round Two
Pollutant
Rationale
PCBs
Beryllium
Phenol.
Methylene chloride
Bis(2-ethylhexyl)phthalate
Cyanide
Fluoride
Aroclor mixtures of PCBs were evaluated in Round One.
Coplanar PCBs are being considered for Round Two
because of their chemical and lexicological similarities to
dioxins.
Not evaluated for all exposure pathways
practices in Round One.
Not evaluated for all exposure pathways
practices in Round One.
Not evaluated for all exposure pathways
practices in Round One.
Not evaluated for all exposure pathways
practices in Round One.
Not evaluated for all exposure pathways
practices in Round One.
Not evaluated for all exposure pathways
practices hi Round One.
and use/disposal
and use/disposal
and use/disposal
and use/disposal
and use/disposal
and use/disposal
3-28
-------
EXHIBIT 3-11
31 Pollutant Candidates For Round Two Regulation1
Pollutant
Acetic acid (2,4-dichiorophenoxy)
Aluminum2
Antimony
Asbestos3
Barium
Beryllium
Bis(2-emylhexyl)phthalate -
Boron
Butanone, 2-
Carbon disulfide
Cresol, p-
Cyanides (soluble salts and complexes)
Dioxins and dibenzofurans
Endosulfan-E
Fluoride
Manganese
Methylene chloride
Nitrate -
Nitrite
Pentachloronitrobenzene
Phenol .
Polychlorinated biphenyls - coplanar
Propanone, 2-
Propionic acid, 2-(2,4,5-trichlorophenoxy)
Silver
Thallium
Tin
Titanium
Toluene
Trichlorophenoxyacetic acid, 2,4,5-
Vanadium
' Pollutants detected at a frequency of at least ten percent with human health and/or ecological toxicity data available
Aluminum does not have human health or ecological toxicity data available, but is included due to its potential for
phytotoxicity.
3 Asbestos was not tested in the NSSS, but is toxic, persistent, and may be in sewage sludge.
3-29
-------
EXHIBIT 3-12
Rationale for the Number of Pollutants Selected as Candidates for the Round Two
Sewage Sludge Regulation
Rationale
Number of
Pollutants
Pollutants for which NSSS sewage sludge samples
were analyzed (Exhibit 3-1).
Pollutants regulated in the Round One sewage
sludge regulation (Exhibit 3-2).
Pollutants combined into classes of congeners or
removed from further consideration (Exhibit 3-3).
Pollutants with a frequency of detection of zero
percent (Exhibit 3-4).
Pollutants deleted because tfiey did not have a non-
zero frequency of detection of ten percent or
greater (Exhibit 3-5). ^
Pollutants detected at a frequency of ten percent or
greater (Exhibit 3-6).
Pollutants detected at a frequency of ten percent or
greater, but with insufficient human health or
ecological toxicity data available (Exhibit 3-9).
Pollutants detected at a frequency of ten percent or
greater, with human health and/or ecological
toxicity data available.
Asbestos added.
Number of candidate pollutants for Round Two
sewage sludge regulation (Exhibit 3-11).
411
10
33
254 .
69
45
15
30
1
31
3-30
-------
-------
4. FINAL LIST OF ROUND TWO POLLUTANTS
This section presents the methods and the results of the Comprehensive Hazard
Identification study, in which a quantitative risk assessment for the Highly Exposed
Individual, including dose-response evaluation, exposure assessment, and risk
characterization, is performed. The 31 pollutant candidates for the Round Two sewage
sludge regulation, identified in the Preliminary Hazard'Identification, are evaluated in this
Comprehensive Hazard Identification. Note that in the Technical Support Documents for
Round One, the calculations begin with an acceptable level of risk, and work backwards to
determine what pollutant concentrations in the sewage sludge are acceptable for that
use/disposal practice, thereby calculating pollutant limits. In this study, levels of risk that
might be associated with a given pollutant under a given use/disposal practice are estimated
based on sewage sludge pollutant concentrations from the NSSS. Those pollutants/practices
with high risk estimates are candidates for the final list of Round Two pollutants. The results
of this Comprehensive Hazard Identification indicate that only a subset of the 31 pollutants
should be considered for regulation in Round .Two.
4.1 GENERAL APPROACH FOR THE COMPREHENSIVE HAZARD
IDENTIFICATION ™^AKU
The purpose of the Comprehensive Hazard Identification study is to identify pollutants
that warrant further consideration for the final list of Round Two pollutants. Analyses are
performed to identify pollutants that may potentially cause human health or ecological risk
for a Highly Exposed Individual (HEI). Consistent with the EPA Guidelines for Exposure
Assessment (57 FR 22888, May 29, 1992), the risk to the HEI is estimated using a
combination of high-end and average assumptions designed to give a plausible estimate of the
individual risk at the upper end of the risk distribution (e.g., above the 90th percentile of the
actual distribution).
In general for mis study, high-end assumptions are used to characterize sewage sludge
concentrations and certain exposure parameters, while average values are typically used to
characterize use/disposal practices and soil and meteorological characteristics. Specifically
sewage sludge concentrations are based on the 95th percentile concentrations of pollutants
obtained in the NSSS, with non-detects set equal to the minimum level (e.g., the minimum
concentration of pollutant that could be measured) (see Exhibit 4-1). For each sewage sludge
use/disposal practice, the HEI is defined as "an individual who remains for an extended
period of time at or adjacent to the site where the maximum exposure occurs" (U.S. EPA,
1992a). Numerous exposure assumptions are specific to each exposure pathway and are given
in the subsequent sections.1
This chapter describes the methods used for each sewage sludge use or disposal
practice and exposure pathway (Section 4.2 (land application), 4.3 (surface disposal), and 4 4
(incineration)). Next, methods for estimating human health and ecological risk are presented
More detailed explanations for the derivations of the specific values used in this study are provided in the
Technical Support Documents for Round One (e.g., U.S. EPA, 1992a).
4-1
-------
(Section 4.5). For those pollutant-exposure pathway combinations for which an exposure can
be estimated, the calculated risk associated with that exposure is then presented. Based on
these risk estimates, the pollutants that warrant further consideration for inclusion on the final
list of Round Two pollutants are presented in Chapter 5.
4-2
-------
EXHIBIT 4-1
95th Percentile Concentrations for Round Two Candidate Pollutants
Pollutant
95th Percentile Sewage Sludge
Concentrations1 (mg/kg dry weight)
Acetic acid (2,4-dichlorophenoxy)
0.030
Aluminum
36,400
Antimony
24
Asbestos2
Not Available
Barium
1,730
Beryllium
8
Bis(2-ethylhexyl)phthalate
191
Boron
182
Butanone, 2-
69.3
Carbon disulfide
3.13
Cresol,
306
Cyanides (soluble salts and complexes)
130
Dfoxins and dibenzofurans
3.11 x
Endosulfan-n
0.0667
Fluoride
411
Manganese
1,620
Methylene cWoride
31.3
Nitrate
5,020
Nitrite
462
Pentachloronitrobenzene
0.0793
Phenol
Polychlorinated biphenyls — coplanar3
Propanone, 2-
Propionic acid, 2-(2,4,5-trichlorophenoxy)
Silver
57.5
5.4
116
I^^^BH^^HK
0.040
128
4-3
-------
Pollutant
Thallium
Tin
Titanium
Toluene
Trichlorophenoxyacetic acid. 2,4,5-
Vanadium
95th Percentile Sewage Sludge .
Concentrations1 (mg/kg dry weight)
10.6
138
363
238
0.0505 ,
64.1
1 Non-deiccts set equal to the Minimum Level. Concentrations from the NSSS.
* Asbestos was not tested in the NSSS, but is toxic, persistent, and may be in sewage sludge.
3 Sewage sludge samples were not analyzed for coplanar PCBs in the NSSS. A composite PCB concentration was
estimated by combining the concentrations of the seven Aroclor mixtures measured in the NSSS.
4.2 LAND APPLICATION PATHWAY EXPOSURE METHODOLOGIES
Methods were developed in Round One to evaluate risk from sewage sludge that is land-
applied to agricultural and non-agricultural sites. These methods include evaluating both
human health and ecological risks associated with exposure to sewage sludge through 14
different pathways. In Round Two, an additional exposure pathway, breastfeeding, is also
considered. These 15 exposure pathways and the corresponding HEIs are summarized hi
Exhibit 4-2."
To estimate exposure to pollutants in sewage sludge that is land-applied, several non-pollutant-
specific types of data are required, including information on application practices and soil
characteristics. The average values used for these parameters are shown in Exhibit 4-3. The
way hi which these data were combined with pollutant-specific data to estimate exposure is
described in Sections 4.2.1 through 4.2.15. "
For both agricultural and non-agricultural land application sites, it is necessary to
estimate the number of years that a site is used and the rate at which sewage sludge is applied
to that site. An estimate of the depth to which sewage sludge is incorporated at the site also
must be made. The values used for these parameters in this analysis are presented in Exhibit
4-3. Note that for the forest.and public contact sites, depth of incorporation is assumed to
be zero, and therefore the "soil concentration" of a pollutant is the same as the sewage sludge
concentration.
For the soil at agricultural land and reclaimed sites, an average bulk density of 1600
kg/m3 is assumed (U.S. EPA, 1992a). For the bulk density of sewage sludge-amended soil
an average value of 1400 kg/m3 is assumed (U.S. EPA, 1992a). For natural "background"
soil concentrations of inorganics, 90th percentile concentrations of inorganics in soil
throughout the United States were used (U.S. Geological Survey, 1992). To ascertain
4r4
-------
whether sewage sludge pollutant concentrations were greater than natural soil concentrations
the 90th percentile soil concentrations were compared to the 95th percentile pollutant
concentrations in sewage sludge. As shown in Exhibit 4-4, for three inorganics the 95th
percentile sewage sludge concentrations are less than (aluminum and titanium) or
approximately equal to (vanadium) the 90th percentile natural soil concentrations This
suggests that natural concentrations of these inorganics contribute significantly to the overall
exposure of an HEI. For antimony and tin, sewage sludge concentrations are much greater
than natural soil concentrations, indicating that sewage sludge loadings are the major
determinant of an HEI's exposure to those inorganic pollutants. In this analysis, background
concentrations of nitrate and nitrite are not applicable due to the cycling of nitrogen in soil
systems. The background concentration of asbestos also is not applicable because asbestos
would not be taken up by crops. Cyanide was. the only applicable inorganic pollutant for
which background soil concentration data were not available. For the organic pollutants
background concentration data are not required because it is assumed that there are not
natural concentrations of organic pollutants in soil.
4-5
-------
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4-6
-------
4-7
-------
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SB & S S g
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4-8
-------
EXHIBIT 4-3
Average Values for Sewage Sludge Land Application
Parameters
Parameter
Definition
Value
Notes
N, agricultural
land
Number of years of application to
agricultural land
20
assumed applied
once even' year for
20 year site life
(U.S...EPA, 1993a)
N, forest
Number of years of application to forest
land
assumed applied
once every three
years for 20 year
site life (U.S. EPA,
1984)
N, reclamation
sites
Number of years of application to
reclamation sites
1
assumed one-time
application
, public contact
Number of years of application to public
contact sues
10. assumed applied
once every two
years for 20 year
site life (U.S. EPA,
1984)
NJW, agricultural
land, forest,
public contact
Site life, i.e., period of time during
which sewage sludge may be applied
20yr
3OT, reclamation
site
Site life, i.e., period of time during
which sewage sludge may be applied
lyr
d, agricultural
land
Depth of incorporation on agricultural
land
15 cm
U.S. EPA, 1992a
d, forest
Depth of incorporation on forest land
Ocm
no incorporation
assumed
d, reclamation
sites
Depth of incorporation on reclamation
sites
10cm
U.S. EPA, 1984
d, public contact
Depth of incorporation on public contact
sites
Ocm
no incorporation
assumed
AR, agricultural
land
Annual whole sludge application rate for
agricultural land
U.S.. EPA; 1992a
dry Mg/ha-yr
Annual whole sludge application rate for
forest land
26
dry Mg/ha-yr1
U.S. EPA, 1992a
AR, reclamation Annual whole sludge application rate for
74
dry Mg/ha-yr1
U.S. EPA, 1992a
reclamation sites
Annual whole sludge application rate for
public contact sites
18
dry Mg/ha-yr'
U.S. EPA, 1992a
Bulk density of soil at agricultural land
and reclamation sites
U.S. EPA/ 1992a
Bulk density of sewage sludge-amended
soil
U.S. EPA, 1992a
Note that 1 Mg = 1 megagram = 1 metric ton = 1000 kg.
4-9
-------
EXHIBIT 4-4
Background Concentrations of Pollutants in Soil
Pollutant
Aluminum
Antimony
Barium
Beryllium
Boron
Cyanide (soluble salts and
complexes)
Fluoride
Manganese
Silver
Thallium
Tin
Titanium
Vanadium
90th Percentile
Background
Concentration in Soil1
(mg/kg dry weight)
68,000
0.51
452
0.65
27
NA
220
342
O2
O3
0.94
3400
60
Ratio of Sewage Sludge
(95th) to Soil
Concentrations
(dimensionless)
. , 0.54
47
3.8
12
6.7
NA
1.9
4.7
NA
NA
150
0.11
1.1
NA means Not Available.
1 Concentration obtained by calculating 90th percentile value, based on geometric means and standard deviations
obtained from U.S. Geological Survey, 1992.
2 Silver was measured too infrequently for a reliable mean concentration to be calculated, as discussed in U S
Geological Survey, 1992.
Thallium was analyzed for in all samples but was never found, as discussed in U.S. Geological Survey, 1992.
4.2.1. Pathway 1 - Ingestion of Crops Grown on Sewage Sludge-Amended Soil
Pathway 1 evaluates human ingestion of plants that have taken up pollutants from
sewage sludge-amended agricultural and non-agricultural lands. Non-agricultural, lands
include forests, reclaimed land, and public contact sites.
4-10
-------
Mass Balance Equations
to be conservative in these analyses, the entire mass of a pollutant applied hi sewage
sludge over the life of a land application site is assumed to be available for plant uptake.
Unlike Pathway 1 in the Technical Support Document for Land Application of Sewage
Sludge, developed for the Round One regulation (U.S. EPA, 1992a), this analysis does not
assume that organic pollutants either degrade or volatilize each year between sewage sludge
applications. This conservative approach is used for this level of assessment because: (a)
plant uptake of most pollutants is low and (b) the dissipation of many of the pollutants being
considered in Round Two is slow. If this pathway yielded high risk for a particular pollutant
the assumptions would be refined.
In calculating total pollutant concentration hi the soil (CZ}) for the agricultural and
ed land scenarios, the following equation is used for both inorganic and organic
reclaimed land scenarios,
pollutants:
1 (N -AR) + MS
where:
4
= concentration of pollutant j in sewage sludge-amended soil (mg
pollutant/kg sewage sludge-amended soil),
= background concentration (dry weight) of pollutant j hi soil (mg
pollutant/kg soil), •
MS = mass of soil hi mixing zone of one hectare of land (Mg soil/ha land),
AT = total number of years sewage sludge is applied to land (yr),
Cj = concentration of pollutant/ hi sewage sludge (mg pollutant/kg sewage
sludge), and
AR = annual whole sludge application rate of sewage sludge to land (dry Mg
sewage sludge/ha-yr).
The mass of soil in the mixing zone of one hectare of land is calculated as:
MS = BD^ • d • lO'1 , . (4-2)
where:
BDsoil - bulk density of soil (kg soil/m3 soil),
d = depth of incorporation (cm), and .
10'1 = constant to convert (kg -cm/in3) to (Mg/ha).
For forests and public contact sites, it is assumed that there is no incorporation of
land-applied sewage sludge. Therefore, the concentration of each pollutant in the "soil" is
set equal to its concentration in the sewage sludge. •
4-11
-------
The concentrations of pollutants in crops grown on sewage sludge-amended soil are
calculated as; . • ' .
UC
(4-3)
where:
CD,.
UC, =
tissue concentration (dry weight) of pollutant j in crop / (mg
pollutant/kg crop tissue), and
rate of uptake of pollutant j into tissue of crop i (jig pollutaut/g dry
weight crop tissue per fig poUutant/g sewage sludge-amended soil).
Exposure Equation
Once the various concentrations of a pollutant in crop tissues are estimated, they are
combined with data on the fraction of crops grown on sewage sludge-amended soil and the
daily dietary consumption of crops to estimate human exposure:
1Q-3
BW
icr3 CT.
CDfJ FCf DCt
BW
1 £ UC, PC, DC
where:
10-3 =
BW =
DC; =
Data Inputs
exposure to pollutant./ from crops produced on sewage sludge-amended
soil (mg pollutant/kg body weight-day),
constant to convert units from (g) to (kg),
body weight (kg), assumed to be 70 kg,
fraction of dietary consumption of crop / grown in sewage sludge-
amended soil (dimensionless), and
daily dietary consumption of crop / (g crop tissue/day).
There are three types of data inputs specific to this pathway: daily dietary
consumption of various crops, fraction of consumption derived from sewage sludge-amended
soil, and pollutant-specific plant uptake rates. Values for the daily dietary consumption of
crops and the fraction of consumption derived from sewage sludge-amended soil are presented
in Exhibit 4-5.
4-12
-------
EXHIBIT 4-5
Dietary Assumptions for Pathway 1
Daily Dietary Consumption
of Crop (g dry
weight/day)1
Fraction of Consumption
Derived from Sewage Sludge-
Amended Soil2
Crop
=
Garden Fruits
Grains and Cereals
Leafy Vegetables
Root Vegetables
Mushrooms
f°°d ***" f°r "» Cr°PS< in **.***. -less otherwise
Fractions represent reasonable estimates, unless otherwise noted. U S EPA 1992a
Fractions represent reasonable worst-case assumptions. U.S. EPA, 1992a.
_ Plant uptake slopes are needed for the seven agricultural crop categories that represent
whr. TS ^f ^ gafden *"* (e'g" t0mat0e<* «"*» ™« steals ^g. bSy
wheat); leafy vegetables (e.g., swiss chard, cabbage, lettuce); dry and fresh legumis (e g
eaS);HP^anUtS; F0tat°eS; 3nd' r°0t Vegetables
-------
• peanut uptake slopes and legume uptake slopes were considered
interchangeable; and
• any vegetative or leafy growth uptake slopes.identified hi a study (e.g.,
soybean leaves) were used for leafy vegetable uptake slopes if no leafy
vegetable studies could be identified.
When multiple data points were available for a particular pollutant and crop from a
variety of studies, the average of the data from the most appropriate studies was used. The
appropriateness of a given study was determined from the study hierarchy established in
Round One: data from sludge-amended field studies were preferred over data from sludge-
amended pot studies, which hi turn were preferred over data from metal-salt-amended field
or pot studies. If. uptake slope data existed for a particular pollutant hi a particular crop
category from more than one study of the same hierarchical level, they were averaged.
For the agricultural pathway, if uptake slope data were not available or could not be
estimated using the above extrapolations for all seven crops for a particular pollutant, then
exposure to that pollutant was not estimated. Available plant uptake slopes for agricultural
land crops are presented hi Exhibit 4-6. Note that only 14 pollutant candidates had available
data on plant uptake slopes for at least one crop; only three pollutant candidates had uptake
slope data available for all seven crops. Therefore only three pollutants could be evaluated
for this exposure pathway. -
The non-agricultural Pathway 1 models human consumption of plants grown hi forests,
on reclaimed lands, or on public contact sites that have been amended with sewage sludge!
Humans are assumed to be potentially exposed to wild berries or mushrooms. For this
pathway, garden fruits were used as a surrogate for berries if actual data did not exist for
organic and inorganic pollutant uptake. For uptake of organic and inorganic pollutants .into
mushrooms, data were not available. In the Technical Support Document for Land
Application of Sewage Sludge (U.S. EPA, 1992a), it was noted that mushrooms have
potential to bioaccumulate both mercury and cadmium. As shown hi Exhibit 4-7, information
is not available on whether mushrooms can bioaccumulate any of the Round Two pollutant
candidates. Note that only four pollutant candidates, those with available wild berry uptake
data, are included hi the exhibit. No pollutants could be evaluated for the non-agricultural
pathway due to the lack of uptake data for mushrooms.
4-14
-------
4-15
-------
u
IS
vo
3
S
a
09
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ca
j>
ca
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vo-
4-16
-------
EXHIBIT 4-7
Available Plant Uptake Slopes for Non-Agricultural Pathway 1
Pollutant
Dioxinsand Dibenzofurans
Fluoride
Plant Uptake Slopes
Qtg/g crop tissue per >g/g soil)
0.0000362
Key to Study Type:
A: Studies conducted in the field with sewage sludge.
B: All other studies conducted with sewage sludge.
C: All other studies.
Footnotes:
1 Soon and Bates, 1985 (com kernels: 2.29): A.
2 U.S. EPA, 1992b. Estimated from model. Assumed dry weight basis
Doss et al., 1977 (tomatoes: 0.35): C.
4 Tonkpnozhenko and Khlyupina, 1974 (corn kernels: 0.136): C.
Example Exposure Calculation for Pathway 1
' - ' The following example presents the calculations for estimating human exnosure to
using £0^-2* 6
°f
MS' fa 6Stimated f°r •&***** land by
where:
1600 =
15 -
10 =
MS.- Jggfe .-!<,-,'.
_ 2400Afg
kg • cm/m3 ha
BDsoit (bulk density of soil) from Exhibit 4-3,
d (depth of incorporation for agricultural land) from Exhibit 4-3 and
constant to convert (kg •cm/m3) to (Mg/ha).
Then, the concentration of fluoride in agricultural soil is calculated using Eq. 4-1:
4-17
-------
CT
fluoride
= 230mg/kg
where:
220 = BSj (background concentration of fluoride in soil) from Exhibit 4-4,
2400 = MS (mass of soil hi agricultural mixing zone), estimated above,
20 = N (total number of years sewage sludge is applied to agricultural land)
from Exhibit 4-3,
411 = Cj (fluoride concentration hi sewage sludge) from Exhibit 4-1, and
7 « ^(application rate of sewage sludge to agricultural land) from Exhibit
4-3.
Total dietary exposure is then determined using Eq. 4-4:
] . [(0.35*/g - 0.025 - 4.15g/day)
+ (0.44gfg • 0.025 • 90.7g/day) + (L9g/g • 0.025 - L91g/day)
+ (2.2g/g • 0.025 • %.15glday) + (2.2g/g • 0.025 • 2.2Sg/day)
+ (0.25^/g • 0.025 • 15.6gfday) + (0.25g/g - 0.025 • L6g/day)]
= 0.0061mg/kg-day
where:
10" = constant to convert units from (g) to (kg),
230 = CZ} (concentration of fluoride hi agricultural soil),
estimated above,
70 - 5W(body weight), assumed to be 70 kg,
0.35-0.025-4.15 = contribution to dietary exposure from garden fruits,
0.44-0.025-90.7 = contribution to dietary exposure from grains and cereals,
1.9 • 0.025 • 1.97 = contribution to dietary exposure from leafy vegetables,'
2.2 • 0.025 • 8.75 = contribution to dietary exposure from legumes,
2.2 • 0.025 • 2.25 = contribution to dietary exposure from peanuts,
0.25 • 0.025 • 15.6 = contribution to dietary exposure from potatoes, and
0.25-0.025-1.6 = contribution to dietary exposure from root vegetables.
4-18
-------
Contribution to dietary exposure is derived as the product of UCy (uptake slope of fluoride
into crop) from Exhibit 4-6, FCt (fraction of dietary consumption of crop grown in sewage
sludge-amended soil) from Exhibit 4-5, and DC, (daily dietary consumption of crop) from
Exhibit 4-5. * .
4.2.2 Pathway 2 - Ingestion of Crops Grown in Sewage Sludge-Amended Home Gardens
Pathway 2 evaluates human ingestion of plants that have taken up pollutants from
sewage sludge-amended home gardens. The mass balance and exposure equations used are
identical to those for Pathway 1 (Section 4.2.1).
Data Inputs
There are three types of data inputs specific to this pathway: the daily dietary
consumption of specific crops, the fraction of that daily consumption that comes from sewage
sludge-amended home gardens, and pollutant-specific plant uptake slopes. Values for
daily dietary consumption of crops and the fraction of consumption derived from sewage sludg
amended soil are presented in Exhibit 4-8.
Garden Fruits
Grams and Cereals
•~""™-~^™— •— ™— ™™— ^«_
Leafy Vegetables
"
Fresh Legumes
Potatoes
Root Vegetables
Sweet Corn
EXHIBIT 4-8
Dietary Assumptions for Pathway 2
Consumption of Crop
(g/day)1
==
4.15
•^•^^H^V^
89.1
•— ™»™
1.97
. 3.22
15.6
1.60
TOp
=====
Fractio
Derived f
A
=====
-^
Amended Soil2
==
0.58
0.0043
0.58
0.58
^MM^H^^H
0.37
0.58
Values represent-the estimated lifetime average daily food intakes for the crops US EPA 1992a
Values are for a Highly Exposed Individual. U.S. EPA, 1992a.
7 UP *? ** Deeded f°r ** Seven Cr°Ps ** ******* major human dietary
aer^r/r?0^ fTi^ *** ^^^ garden ^^
-------
legumes (e.g., beans, peas); potatoes; root vegetables (e.g., carrots, beets); and sweet corn
(U.S. EPA, 1992a). Note that there are three differences between these crops and the
agricultural crops used to evaluate exposure in Pathway 1. First, only fresh legumes are
considered for Pathway 2, not both fresh and dry, because home gardeners do not usually
grow the dried legumes they consume. Second, peanuts are not considered in Pathway 2,
also because home gardeners do not usually grow the peanuts they consume. Third, sweet
corn is separated out as a food group for home .gardeners because so many gardeners grow
sweet corn. In Pathway 1, sweet corn is included in the category of cereals and grains In
Pathway 2, the percent of sweet corn that is homegrown differs from the percent of grains
and cereals that are homegrown, and thus the two food categories are separated, the non-
agricultural crops of Pathway 1, berries and mushrooms, which are grown on forest land
reclaimed sites, and public contact sites, are not considered relevant for home gardens.
Data on plant uptake slopes did not exist for all pollutant candidates and for all crops
When no data were available for a particular crop, the following extrapolations were made
between crops for a given pollutant:
• grain and cereal uptake slopes and forage/pasture uptake slopes (used
in animal exposure pathways) were considered interchangeable;
• potato uptake slopes and root vegetable uptake slopes were considered
interchangeable; ' • ~
• peanut uptake slopes and legume uptake slopes were considered ^^
interchangeable; and
• any vegetative or leafy growth uptake slopes identified in a study (e.g
soybean leaves) were used for leafy vegetable uptake slopes if no leafy
vegetable studies could be identified.
When multiple data points were available for a particular pollutant and crop from a
variety of studies, the average of the data from the most appropriate studies was used The
appropriateness of a given study was determined from the study hierarchy established in
Round One: data from sludge-amended field studies were preferred over data from sludge-
amended pot studies, which in turn were preferred over data from metal-salt-amended field
or pot studies. If uptake slope data existed for a particular pollutant in a particular crop from
more than one study of the same hierarchical level, they were averaged. -
Even after all of the above extrapolations were made, many pollutants still had data
gaps.. If uptake slope data were not available or could not be estimated using the above
extrapolations for all seven crops for a particular pollutant, then exposure to that pollutant
was not estimated. Available plant uptake slopes are presented in Exhibit 4-9. Note mat the
exhibit includes only the 14 pollutant candidates for which there were uptake data for at least
^nr°P' , ?y **$ P°Uutant candidates CPOM te ev^ted for this exposure pathway
because only.three pollutant candidates had uptake slope data available for all seven crops
4-20
-------
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4-22
-------
Example Exposure Calculation for Pathway 2
The following example presents the calculations for estimating human exposure to
fluoride in sewage sludge applied to a home garden. The example uses the equations and
input data presented above for Pathway 2.
JFirst, the mass of soil in the mixing zone, MS, is estimated for a home garden by
using Eq. 4-2:
MS = 1600*g - 15cm . W^Mg/ha = 2400Mg
m3 . kg - cm/m3 ha
where:
1600 = BDsoil (bulk density of soil) from Exhibit 4-3,
fill = d (4epth °f mc°IP°ration for agricultural land) from Exhibit 4-3 and
10- - constant to convert (kg - cm/m3) to (Mg/ha).
Then, the concentration of fluoride in soil is calculated using Eq. 4-1:
(220mg . 2400Af£J + /; .4llmg 7Mg
CT^ = Li_ ha } ( kg ha.,.
ha-yr j ha
= 230mg/kg
where:
(back8round concentration of fluoride in soil) from Exhibit 4-4
oo Sv (maSS °f S0il " h°me garden mbdnS zone>' estimated above, '
20 - N (total number of years sewage sludge is applied to home garden)
from Exhibit 4-3,
411 = q.(fluoride concentration in sewage sludge) from Exhibit 4-1 and
7 - ^(application rate of sewage sludge to home garden) from Exhibit 4-
4-23
-------
Total dietary exposure is then determined using Eq. 4-4:
0.58
• 0.0043 • S9.1g/day) + (l<9g/g • 0.58 • 1.97g/day)
0.58 • 3.22g/day) + (0.25g/g • 0.37
- (0.25g/g - 0.58 • 1.6g/day) + (0.35^ • 0.58
= Q.03lmg/kg-day
where:
10~3 = constant to convert units from (g) to (kg),
,
230 = C3} (concentration of fluoride in agricultural soil),
estimated above,
70 = BW (body weight), assumed to be 70 kg,
0.35 • 0.58 -4.15 = contribution to dietary exposure from garden fruits,
0.44-0.0043-89.1 = contribution to dietary exposure from grains and
cereals,
1.9 • 0.58 • 1.97 . = contribution to dietary exposure from leafy vegetables,
n'?c °n5? " 3"22 = contributi°n to dietary exposure from fresh legumes, '
0.25 • 0.37 • 15.6 = contribution to dietary exposure from potatoes,
0.25 • 0.58 - 1.6 = contribution to dietary exposure from root vegetables,
and • '
0.35-0.58-1.6 = contribution to dietary exposure from sweet com.
Contribution to dietary exposure is derived as the product of UCfj (uptake slope of fluoride
into crop) from Exhibit 4-9, FCt (fraction of dietary consumption of crop grown in sewage
gidge-amended soil) from Exhibit 4-8, and DC, (daily dietary consumption of crop) from
4-24
-------
4.2.3 Pathway 3 - Direct Ingestion of Sewage Sludge by Children
Pathway 3 evaluates children's exposure to pollutants from direct ingestion of sewaee
sludge applied to land. The agricultural and non-agricultural scenarios are different in their
assumptions regarding the age of children who ingest sewage sludge. At agricultural and public
contact sites, children ages 1 to 6 are assumed to be exposed, whereas forforest and reclaimed
sites, only older children ages 4 to 6 are assumed to have the opportunity for exposure For all
scenarios, children are assumed to be exposed directly to sewage sludge from storage piles or
from the soil surface, not to a sewage sludge/soil mixture.
Data Inputs and Exposure Equation
As in Round One, children (ages 1 to 6) exposed to agricultural land and public contact
sites are assumed to ingest 0.2 g soil (dry weight) per day, and weigh 16 kg. Older children
(ages 4 to 6) exposed to forest and reclaimed sites are assumed to ingest 0.2 g soil per day and
Wp?r A HM", ™mf ?^ me °f °'2 g SOU P£r day is a high-end value, but does not represent
a PICA child. The body weights are average values. Exposure is calculated as:
= —1. (4-5)
where:
EXPj = exposure to pollutant y in sewage sludge (mg pollutant/kg body weight-
day), &
IS ^ = sewage sludge ingestion rate (g sewage sludge/day),
10- = constant to convert units from (g/day) to (kg/day)
q = concentration of pollutant j in sewage sludge (mg pollutant/kg sewage
sludge), and "
BW = body weight (kg).
Example Exposure Calculation for Pathway 3
The exposure of a child to fluoride from directly ingesting sewage sludge applied to
agricultural land can be estimated from Eq. 4-5:
-------
10~* = constant to convert (g) to (kg),
411 = Cj (concentration of fluoride in sewage sludge) from Exhibit 4-1, and
16 = BW (body weight of child assumed to be exposed to agricultural land),
from-Methods section above.
4.2.4 Pathway 4 - Ingestion of Animal Products Produced From Animals Consuming
Forage/Pasture Grown on Sewage Sludge-Amended Soil
i .
Pathway 4 calculates human exposure to pollutants through consumption of animals that
ingest forage/pasture grown on sewage sludge-amended soils. In the agricultural Pathway 4,
animals ingest forage and pasture produced oh sewage sludge-amended soil. Humans then ingest
animal products, such as beef, pork, lamb, poultry, dairy products, and eggs. The non-
agricultural Pathway 4 examines human exposure to pollutants through consumption of deer and
elk that forage on sewage sludge-amended forest land and reclaimed land.
Methods
As in Pathways 1 and 2, to be conservative, the entire mass of pollutant applied in
sewage sludge over the life of a land application site is assumed to be available for plant uptake
(see Section 4.2.1). In calculating total pollutant concentration hi the soil for the agricultural
and reclaimed land scenarios, Eq. 4-1 from Section 4.2.1 is used for inorganics and organics.
The expected concentrations of pollutants hi forage/pasture grown on sewage sludge-amended
soil are then estimated using Eq. 4-3 from Section 4.2.1.
To estimate concentrations of pollutants hi animal tissues, the forage/pasture pollutant
concentrations are combined with animal uptake rates. In this calculation it is assumed that
forage is 100 percent of the animal's diet:
Ujk (4-6)
where:
' = concentration of pollutant j in animal product k (mg pollutant/kg
animal tissue),
- tissue concentration (dry weight) of pollutant/ in forage/pasture
(mg pollutant/kg forage/pasture), and
ujt = rate of uptake of pollutant y into animal product k (mg pollutant/
kg dry weight animal tissue per mg pollutant/kg dry weight diet).
Once the concentrations of pollutants hi animal tissues have been estimated, they are
combined with data on the daily dietary consumption of animal products and on the fraction ,of
these animal products that are produced on sewage sludge-amended soil to estimate human
exposure: ^^
4-26
-------
where:
EXPj
lO'3 =
BW =
FAk =
DA, =
Data Inputs
IP
-3
(4-7)
ID'3 CD.
forage, j
BW
£
k
UjkFAkDAk
exposure to pollutant j from animal products produced on sewage sludge-
amended soil (mg pollutant/kg "body weight-day),
constant to convert units from (g) to (kg),
body weight (kg), assumed to be 70 kg,
fraction of dietary consumption of animal product k produced on sewage
sludge-amended soil (dimensionless), and
daily dietary consumption of animal product k (g dry weight animal
product/day).
This pathway requires data on pollutant uptake rates into forage/pasture and animal
products as well as data on daily dietary consumption of specific animal products and the
fraction of that daily consumption that comes from animals feeding on forage/pasture produced
on sewage sludge-amended soil. Values used for the latter two parameters are shown in Exhibit
4-10. •
EXHIBIT 4-10
Dietary Assumptions for Pathway 4
Animal Product
Beef (lean)
Beef Fat
Beef Liver (lean and fat)
Dairy (non-fat)
Dairy Fat
Eggs
Lamb (lean)
Lamb Fat
Daily Dietary
Consumption of Animal
Product (g/day)1
19.3
15.5
1.1.
28.9
18.1
8.3
0.20
0.21
Fraction of Consumption
Derived from Sewage
Sludge-Amended Soil2
9.7xlO'2
9.7xlO'2
9.7xlO'2
S.lxlO'2
S.lxlO"2
7.9xlO'2
9.7xlO'2
9.7xlO-2
4-27
-------
EXHIBIT 4-10
Dietary Assumptions for Pathway 4 (cont'd.)
Animal Product
Poultry (lean)
Poultry Fat
Pork (lean)
Pork Fat
Deer (lean)
Deer Fat
Deer Liver (total)
Elk (lean)
Elk Fat
Elk Liver (total)
Daily Dietary
Consumption of Animal
Product (g/day)1
6.7
' 1.3
. 9.0
12.7
15.33
5.13
0.383
30.63
10.23
0.763
Fraction of Consumption
Derived from Sewage
Sludge-Amended Soil2
l.lxlO'1
1.1x10-'
9.7xlO'2
9.7xlO-2
1
1 .
1
0.5
0.5
0.5
' Values represent the estimated lifetime average daily food intakes for the animal products unless otherwise noted U S
EPA. 1992a.
: Fractions represent reasonable estimates. U.S. EPA, 1992a.
' It was assumed that total consumption of deer and elk meat and fat constitutes 50 percent of the HEI's consumption
of agricultural animal products. U.S. EPA. 1992a.
If no data were available on pollutant uptake slopes into forage/pasture, it was assumed
that grain and cereal uptake slopes and forage/pasture uptake slopes were interchangeable.
When multiple data points were available from a variety of studies for the uptake of a particular
pollutant into forage/pasture crops, the average of the data from the most appropriate studies was
used. The appropriateness of a given study was determined from the study hierarchy established
in Round One: data from sewage sludge-amended field studies were preferred over data from
sewage sludge-amended pot studies, that in turn were preferred over data from metal-salt-
amended field or pot studies. If uptake slope data existed for a particular pollutant from more
than one study of the same hierarchical level, they were, averaged.
Available plant uptake data are presented in Exhibit 4-11. Note that if forage/pasture
uptake data were not available for a particular pollutant, that pollutant is not included in the
exhibit.
4-28
-------
EXHIBIT 4-11
Forage/Pasture Uptake Slopes for Agricultural Pathway 4
Pollutant
Aluminum
Beryllium
Boron
Dioxihs and Dibenzofurans
Fluoride
Manganese
Polychlorinated biphenyls (coplanar)
Silver
Forage/Pasture Uptake Slopes
Qtg/g plant per ftg/g sou)
Key to Study Type:
A: Based on sewage sludge field study.
B: Based on non-field sewage sludge study.
C: Based on non-sewage sludge study.
Footnotes:
1 Muchovej et al., 1986 (ryegrass: 3.13): C.
2 Bonn and Seekamp, 1979 (oats: 0.31): C.
3 Soon and Bates, 1985 (bromegrass: 1.64): A.
* U.S. EPA, 1992d. Estimated from model
DO* =,
): c:
3.131
0.312
3,9s
0.0244
0.67755
8.226
^^Hm
2.07
0.0218
« 0.33):
8 Romney et al., 1977 (barley: 0.021): C
» Tonkonozhenko and Kblyupina, 1974 (winterwheat leaves+stem: 0.07, clover leaves + stems: 0.063, rice: 0.03): C.
For non-agncultural am
products, however, tissue values for sheep, goats dee
HVCr ^ -*«ta.^S
- For coplffiar PCBi- """^ vataes were
4-29
-------
Exhibit 4-12 presents the available animal uptake slopes for the agricultural animal
products of interest; Exhibit 4-13 presents the available uptake slopes for non-agricultural
animals. As before, note that only those pollutant candidates with data for at least one animal
uptake slope are included hi the exhibits.
4-30
-------
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4-31
-------
EXfflBIT4-13
Animal Uptake Slopes for Non-Agricultural Pathway 4
Pollutant
Aluminum
Barium
Boron
Manganese
• Polychlorinated
biphenyls (coplanar)
Animal Uptake Slopes Qtg/g animal per /tg/g feed) -
Deer
(Lean)
0.021 l
0.0883
0.0911
0.009254
NA
Deer
Fat
NA
NA
NA
NA
4.2152
Deer
Liver
0.0071
0.05353
0.09151
0.009254
6.6642 '
Elk
(Lean)
0.021 '
0.0883
0.091 l
0.009254
NA . ,
Elk
Fat
NA
NA
NA
NA
4.2152
Elk
Liver
0.0071
0.05353
0.09151
0.009254
6.6642
NA means Not Applicable.
1 Bray et al., 1985 (goat).
2 U.S. EPA, 1992a.
•'Whelan, 1993 (sheep:muscle).
4 Bray et al., 1985 (goat); Voigt et al., 1988 (beef).
Example Exposure Calculation for Pathway 4
The following example presents the calculations for estimating human exposure to boron
from consumption of deer and elk which have eaten forage plants on forest lands amended with
sewage sludge. Because it is assumed that there is no incorporation of sewage sludge into forest
soils, the concentration of boron hi sewage sludge is used as the relevant concentration for
uptake into the forage. The concentration of boron in forage is calculated using Eq. 4-3:
CD
^"
boron
= 182
18Z ~~
V-gg
= 710 mg/kg
where:
182 =
3.9
j (concentration of boron in forest soil, equivalent to the concentration
of boron in sewage sludge) from Exhibit 4-1, and
y (uptake slope of boron into forage/pasture) from Exhibit 4-11.
4-32
-------
Total dietary exposure to boron from wild animals eating plants in sewage sludge-
amended forests is then determined using Eq. 4-7:
1-3
+ (0.092g/g • 1 • 0.3Sg/day) + (Q.09lg/g ' 0.5 - 30.6g/day)
• + (0.092g/g • 0.5 - Q.76g/day)]
= Q.Q29mg/kg-day
where: "
10"3 = constant to convert units from (g) to (kg),
710 - CTj (concentration of boron in forest soil), estimated above,
70 = BW (body weight), assumed to be 70 kg,
0.091 -1-15.3 = contribution to dietary exposure from lean deer,
0.092 • 1 • 038 = contribution to dietary exposure from deer liver,
0.091 • 0.5 • 30.6 = contribution to dietary exposure from lean elk, and
0.092 • 0.5 • 0.76 = contribution to dietary exposure from elk liver.
Contribution to dietary exposure is derived as the product of UCik (uptake slope of boron
into animal product) from Exhibit 4-13, FAk (fraction of dietary consumption of animal product
derived from sewage sludge-amended soil) from Exhibit 4-10, and DAt (daily dietarv
consumption of animal product) from Exhibit 4-10. .
4.2.5 Pathway 5 - Consumption of Animal Products Produced From Animals That Insest
Sewage Sludge „ e
Pathway 5 involves the application of sewage sludge to land, the direct ingestion of
sewage sludge by animals, and finally, the consumption of animal products by humans
Agricultural Pathway 5 considers only the direct ingestion of sewage sludge by livestock
following the surface application of sewage sludge to pasture crops. Non-agricultural Pathwav
5 considers the direct ingestion of sewage sludge by livestock that graze on grasses growing on
forest land or reclaimed land; the animals are then ingested by humans. The pathway does not
consider the grazing of livestock on public contact sites because it is assumed such grazing
would .be controlled. The pathway also does not consider wild animals in forest land because
deer do not graze on plants close to the ground and would not inadvertently ingest sewage
sludge. Furthermore, other wild herbivorous game animals are assumed to grazl over large
territories. °
Methods
To estimate concentrations of pollutants in animal tissues, sewage sludge pollutant
concentrations are combined with percent of animal ingestion that is sewage sludge and tissue
uptake rates. In this calculation the sewage sludge pollutant concentration multiplied by the
- 4-33
-------
animal's percentage sewage sludge consumption and pollutant uptake rate is assumed to equal
the resulting animal tissue concentration of the pollutant:
where:
CAU
FS
CAjk = C. • FS - U.k
(4-8)
concentration of pollutant j in animal product k (mg pollutant/kg
animal tissue),
concentration of pollutant j in sewage sludge (mg pollutant/kg
sewage sludge),
fraction of animal's diet that is sewage sludge (unitiess, kg sewage
sludge/kg diet), and
rate of uptake of pollutant/ into animal product k (mg pollutant/
kg dry weight animal tissue per mg pollutant/kg dry weight diet).
Once the concentrations of pollutants in animal tissues have been estimated, they are
combined with data on the daily dietary consumption of animal products and on the fraction of
those animal products that are produced on sewage sludge-amended soil to estimate human
exposure:
"3
= 4L 2: CA*FA* DAt
where:
ID'3
BW
FAk
DAt
10
-3
CJ
FS
BW
UjkFAtDAt
(4-9)
exposure to pollutant/ from animal products produced oh sewage sludge-
amended soil (mg pollutant/kg body weight-day),
constant to convert units from (g) to (kg),
body weight (kg), assumed to be 70 kg,
fraction of dietary consumption of animal product k produced on sewage
sludge-amended soil (dimensionless), and
daily dietary consumption of animal product k (g dry weight annual
product/day).
Data Inputs
For this pathway, there are four data inputs required to calculate human exposure from
consumption of animal products produced from animals that ingest sewage sludge: the percentage
of animal diet that consists of sewage sludge, the daily dietary consumption of specific animal
4-34
-------
products, the fraction of that daily consumption .which comes from animals that ingest sewage
sludge, and animal uptake rates of pollutants.
The percentage of a grazing cattle's diet that is sewage sludge, averaged over a season,
is estimated to be 2.5 percent (Chaney et al., 1987; Bertrand et al., 1981). However, given that
in any one year, the maximum percentage of a farm treated with sewage sludge is approximately
33 percent, and assuming that livestock are rotated among several pasture fields, the actual
percentage of the diet that is sewage sludge is assumed to be lower than 2.5 percent. The diet
for cattle grazing on land treated with sewage sludge that was applied the previous growing
season has been shown to be approximately 1.0 percent sewage sludge (Decker et al., 1980)
When a weighted average is calculated from these two values of sewage sludge ingestion, the
long-term average percentage of diet that is sewage sludge is 1.5 percent: V3(2.5%) + %fl 0%)
= 1.5% (Chaney etaL, 1991). . . .
Values for the daily dietary human consumption of specific animal products and the
fraction of consumption from sewage sludge-amended soil are shown in Exhibit 4-14 For
animal uptake slopes for beef, dairy, and lamb products, see Exhibit 4-12 in Section 4.2.4.
EXHIBIT 4-14
Dietary Assumptions for Pathway 5
Beef (Lean)
Beef Fat
Beef Liver
Dairy (Non-Fat)
Dairy Fat
Lamb (Lean)
Lamb Fat
roduct
=========
)
Daily Dietary
Consumption of A
Product (g/day
19.3
15.5
1.1
28,9
18.1
0.20
0.21
Fraction of Consumption
Derived from Sewage
Sludge-Amended Soil2
=====
9.7xlO'2
9.7xlO-2
S.lxlO'2
S.lxlO-2
9.7xlO'2
^S^Sj^^SSS^SS^^SE^^S^^^^^^^s:^^^^^^^^^^^^^^^~^~^^^^^m^^
: daily food intakes for the animal products unless otherwise noted. U.S.
: Fractions represent reasonable estimates. U.S. EPA, 1992a.
Example Exposure Calculation for Pathway 5
The following example presents the calculation for estimating human exposure to
manganese from consumption of livestock that ingest sewage sludge on sewage sludge-amended
forest land. Using Eq. 4-9: .
4-35
-------
EXP.
manganese
io-3** . 1620-22 • 0.015
70kg
[fo.00925^ • 9.7xlO~2 • 1.1-S-} +
\ . 8 day)
fo.0005^ - 9.7x20-z • 19.3-3-} + fo.00925^ • 9.7x10-2 - 0.20-M
\ 8 day) ( g day)
V
g
day
8.9xlO-7—™8__
kg-day
where:
io-3
1620
0.015
70
(0.00925 -9.7xlO-2 -1.1)
(0.0005 • 9.7 x IO-2 • 19.3)
(0.00925 - 9.7 x 10'2 • 0.20)
(0.0005 -3.1 xlO'2 -28.9)
constant to convert units from (g) to (kg),
Cj (concentration of manganese in sewage
sludge) from Exhibit 4-1,
FS (fraction of animal's diet that is sewage
sludge), discussed above,
BW(]body weight), assumed to be 70 kg,
contribution to dietary exposure from beef
liver,
contribution to dietary exposure from lean
beef,
contribution to dietary exposure, from lean
lamb, and
contribution to dietary exposure from non-
fat dairy.
Contnbution to dietary exposure is derived as the product of Uf (uptake of manganese into
animal product) from Exhibit 4-12, FAk (fraction of dietary consumption of animal product
derived from sewage sludge-amended land), from Exhibit 4-14, and DAk (daily dietary
consumption of animal product) from Exhibit 4-14.
4.2.6 Pathway 6 - Animal Toxicity From Plant Consumption
Pathway 6 calculates herbivorous animal toxicity caused by the consumption of plants that
are grown on sewage sludge-amended soil on both agricultural and non-agricultural lands Non-
agncultural lands include forests, reclaimed land, and public contact sites. In the agricultural
pathway, the HEI is the most sensitive herbivorous livestock that consumes plants grown on
sewage sludge-amended soil. For the non-agricultural forest land Pathway 6, two exposure
scenarios are possible. In one, the HEI is a small herbivorous mammal that spends its entire
life in a sewage sludge-amended area feeding on seeds and small plants close to the sewage
sludge/soil layer. In the second scenario, the HEI is an herbivorous livestock that grazes on the
. 4-36
-------
grasses growing on sewage sludge-amended forest land: The HEI for reclaimed land is
livestock; the HEI for public contact sites is a small herbivorous mammal.
Methods
For the agricultural Pathway 6, the animals of interest are herbivorous livestock. For
the non-agricultural Pathway 6, the .animals of interest are small herbivorous mammals as well
as herbivorous livestock. Eq. 4-1 from Section 4.2.1 is used to calculate total pollutant
concentration in the soil (C7}) for the agricultural and reclaimed land scenarios. Exposure is
reported in terms of milligrams pollutant per kilogram of diet, assuming that the herbivore
receives its total diet from forage/pasture grown on sewage sludge-amended land Therefore
the dietary exposure for the HEI can be expressed as: '
(4-10)
where:
EXPAJ = exposure of animal to pollutant./ (mg pollutant/kg diet),
CDJ ~ ti85116 concentration (dry weight) of pollutant j in forage/pasture
(mg pollutant/kg forage/pasture),
01 i - concentration of pollutant j in sewage sludge-amended soil (mg
pollutant/kg sewage sludge-amended soil), and
UCforageJ = rate of uptake of pollutant j into tissue of forage/pasture (mg
pollutant/kg dry weight forage/pasture per mg pollutant/kg soil).
Data Inputs
For this pathway, data are needed on pollutant uptake rates into forage/pasture For
available uptake rates into forage/pasture, see Exhibit 4-11 in Section 4.2,4.
Example Exposure Calculation for Pathway 6
live^n 5"°^ ^"^^ ^^ ** ^^^^ for estimating exposure of herbivorous
livestock to fluoride in sewage sludge applied to agricultural land.
E 4_2FirSt' ^ maSS °f S0il fa ** mbcmg zone' MS> is estimated for agricultural land by using
MS = g • i5cm . W~Mg/ha = 2400Mg
m3 kg ' cm/m3 ha
where:
1600 = BDSOU (bulk density of soil) from Exhibit 4-3,
15 = d (depth of incorporation for agricultural land) from Exhibit 4-3 and
iu - constant to convert (kg -crn/m3) to (Mg/ha).
4-37
-------
Then, the concentration of fluoride in agricultural soil is calculated using Eq. 4-1:
(22Qmg 24QQMg\ /.,.
„ • _(-*-—•ri'i**
' fhtoridf
* 20yr • -2ML + MM**
V ha-yr) ha
- 230mgfkg
where:
220 = 55y (background concentration of fluoride hi soil) from Exhibit 4-4,
2400 = JWS (mass of soil in agricultural mixing zone), estimated above,
20 = N (total number of years sewage sludge is applied to agricultural land)
from Exhibit 4-3,
411 = Cj (fluoride concentration in sewage sludge) from Exhibit 4-1, and
7 = AR (application rate of sewage sludge to agricultural land) from Exhibit
4-3.
The expected exposure of livestock to fluoride hi forage/pasture is calculated usins
Eq. 4-10: . '
0
EXP = 230mS . 0.68g
kg g
= 160 mg/kg
where: "...
230 = CTj (concentration of fluoride in agricultural soil), estimated above, and
0.68 =
ucforage.j (uptake slope of fluoride into forage/pasture) from Exhibit 4-11.
4.2.7 Pathway 7 - Animal Toxicity From Direct Ingestion of Sewage Sludge
This pathway calculates herbivorous animal toxicity caused by the direct ingestion of
sewage sludge on both agricultural and non-agricultural lands. Non-agricultural lands include
forests and reclaimed land. For both the agricultural and non-agricultural pathways, the HEI
is the most sensitive or most exposed "herbivorous livestock that directly ingests sewage sludge
from sewage sludge-amended soil. This pathway does not consider exposure to sewage sludge
on public contact sites because livestock usually are not grazed there.
Methods
For both the agricultural and non-agricultural Pathway 7, the HEI is herbivorous
livestock. Exposure is reported in terms of milligrams pollutant per kilogram of diet, assuming
4-38
-------
that the herbivore receives its total diet from forage/pasture grown on sewage sludge-amended
land. Therefore, the dietary exposure for the HEI can be expressed as:
EXPAj = Cj -FS (4-11)
where:
EXPAj = exposure of animal to pollutant j (mg pollutant/kg diet),
Cj = concentration of pollutant j in sewage sludge (mg pollutant/kg sewage
sludge), and .
FS = fraction of animal's diet that is sewage sludge (unitless, kg sewage
sludge/kg diet).
Data Inputs
For this pathway, data are required on the percentage of animal diet that consists of
sewage sludge. As described for Pathway 5, the percentage of a grazing cattle's diet that is
sewage sludge, averaged over a season, is estimated to be 2.5 percent (Chancy et al 1987-
Bertrandetal., 1981). However, given that in any one year, the maximum percentage of a fann
treated with sewage sludge is approximately 33 percent, and assuming that livestock are rotated
among several pasture fields, the actual percentage of the diet that is sewage sludge is assumed
to be lower than 2.5 percent. The diet for cattle grazing on land treated with sewage sludge that
was applied the previous growing season has been shown to be approximately 1 0 percent
sewage sludge (Decker et al., 1980). When a weighted average is calculated from these two
values of sewage sludge ingestion, the long-term average percentage of diet that is sewage sludge
is 1.5 percent: V3(2.5%) + %(1.0%) = 1.5% (Chaney et al., 1991). *
Example Exposure Calculation for Pathway 7
Uveaock " mwsansse *- dircct tegesti°n °f
0.015
is
kg
where: -
, and
FS (fraction of animal's diet that is sewage sludge), discussed above.
4-39
-------
4.2.8 Pathway 8 - Toxicity to Plants
This pathway could not be evaluated due to a lack of data on the phytotoxicity effects of
the candidate Round Two pollutants.
4.2.9 Pathway 9 - Toxicity to Soil-Dwelling Organisms
Pathway 9 evaluates toxicity to soil-dwelling organisms due to the presence of pollutants
•in sewage sludge that is land-applied to agricultural and rion-agricultural lands. Non-agricultural
lands include forests', reclaimed land, and public contact sites. The soil-dwelling organisms
considered are earthworms. There is no evidence that earthworms are the most sensitive
species; however, because of a lack of data for other soil-dwelling species, earthworms are
considered the HEI for this pathway.
Data Inputs and Exposure Equation
For each pollutant and type of land application site, the concentration of pollutant hi the
soil had to be calculated. For agricultural and reclaimed land, the pollutant concentration was
calculated using Eq. 4-1 in Section 4.2.1. For forests and public contact sites, it was assumed
that there would be no incorporation of land-applied sewage sludge. This implies that the soil
layer to which the soil-dwelling organisms are exposed is pure sewage sludge. Therefore, the
concentration of each pollutant hi the exposure layer ("soil") was set equal to its concentration
hi the sewage sludge. For this pathway, exposure to pollutants in soil by earthworms is
measured by the concentration of th6 pollutants hi the sewage sludge/soil:
EXPO; = CTj (4-12)
where: .
EXPO; = exposure of soil-dwelling organisms to pollutant/' (mg pollutant/kg
sewage sludge-amended soil), and
£2} = concentration of pollutant /" hi sewage sludge-amended soil (mg
pollutant/kg sewage sludge-amended soil).
Example Exposure Calculation for Pathway 9
The following example estimates exposure of earthworms to manganese hi agricultural
sou.
For agricultural land, the mass of soil in the mixing zone, MS, must first be estimated
by using Eq. 4-2:
4-40
-------
MS = i°9°*£ . 15cm . W-lMg/ha = 240QMg
m3 kg • cmjm3 ha
where:
1600 . = BDsoil (bulk density of soil) from Exhibit 4-3,
15 = d (depth of incorporation for agricultural land) from Exhibit 4-3, and
10 - constant to convert (kg • cm/m3) to (Mg/ha).
Then, the concentration of manganese in the soil, and thus the earthworm's exposure to
manganese, is calculated using Eqs. 4-1 and 4-10:
?42jng • 24gOJKg + 1620mg ^^. •
CT~*_ " EXP0mangmese = US. *" I I kg ha-yrl
(2Qyr-™8\+~
V ha-yr)
= 410mg/kg
where:
^ 342 - ^(background concentration of manganese in soil) from Exhibit 4-4
•v 'macc nf c~'l m agricultural mixing zone), estimated above,
of years sewage sludge .is applied to agricultural land)
1620 = q (manganese concentration in sewage sludge) from Exhibit 4-1 and
- AR (application rate of sewage sludge to agricultural land) fron/Exhibit
4.2.10 Pathway 10 - Toxicity to Predators of Soil-Dwelling Organisms
Pathway 10 evaluates toxicity to animals feeding on soil-dwelling
Sge *#""«*«.•** on ^th agricultural and nln^cul^
lands include forests, reclaimed land, and public contact sites. For both the
non-agncultural Pathway 10, the HEI is a small insectivorous mammal
organisms-
Methods
in Section
incorporation
4-41
-------
of land-applied sewage sludge. This implies that the soil layer to which the soil-dwelling
organisms are exposed is pure sewage sludge. Therefore, the concentration of each pollutant
in the exposure layer ("soil") is set equal to its concentration in the sewage sludge.
Exposure for Pathway 10 is reported in terms of milligrams pollutant per kilogram of diet
for the insectivorous mammal, assuming that the only source of the pollutant hi the mammal's
diet is from ingestion of contaminated soil-dwelling organisms (eaVthworms). This dietary
concentration is referred to as the "pollutant intake level" and is represented by P/Ly hi the
following equation:
EXPAj = PILj. = CTj' BACCj • FD (4-13)
= exposure of insectivorous mammal to pollutant j (mg pollutant/kg
diet>>
= intake level of pollutant j in insectivorous mammal's diet
(mg pollutant/kg diet),
= concentration. of pollutant j in sewage sludge-amended soil
(mg pollutant/kg sewage sludge-amended soil),
BACCj = bioaceumulation factor for pollutant j (mg pollutant/kg soil
organisms per mg pollutant/kg sewage sludge-amended soil), and
FD = fraction of diet considered to be soil organisms (unitless, kg soil
organisms/kg diet).
First, the concentration of the pollutant in the soil-dwelling prey is calculated by
multiplying the pollutant concentration hi the soil (CTJ) by a bioaceumulation factor (BACC).
To adjust the pollutant concentration hi soil-dwelling organisms to the analogous concentration
in the entire diet of the insectivorous mammal, the concentration hi the soil-dwelling organisms
is then multiplied by the fraction of the insectivorous mammal's diet that consists of soil-
dwelling organisms. " . •
Data Inputs
There are two data inputs required for this pathway: fraction of diet considered to be
soil organisms and bioaceumulation factors. As in. Round One, it is assumed that the fraction
of the HEI's diet that is composed of' soil-dwelling organisms is one-third, based on a
consideration of maximum chronic consumption of earthworms by wildlife (U.S. EPA, 1992a).
Because earthworms are generally the most conspicuous prey item of the soil biota and
are considered a potential vector for the transfer of sewage sludge pollutants up the food chain,
a number of studies have measured bioaceumulation hi earthworms. Most of the studies,'
however, have been focused on a select set of metals (e.g., cadmium, lead, mercury, selenium!
zinc, copper, nickel, and chromium) or persistent chlorinated hydrocarbons, such as DDT (see
Beyer, 1990; Gfflett, 1994). Very few data were found on bioaceumulation of Round Two
candidate pollutants by earthworms from soil.
4-42
-------
Bioaccumulation of Organic Pollutants. No empirical data were found on
bioaccumulation in soil organisms for any of the candidate organic pollutants for Round Two
except dioxins and dibenzofurans. A predictive equation that describes bioaccumulation of all
organic compounds hi earthworms inhabiting contaminated soils was found (Menzie et al
1992). This equation is based on a relationship between the fraction of organic carbon in the
soil and the lipid content of earthworms: .
• BACC = - -
0-66 •/«
where: .
BACC = bioaccumulation factor for earthworms (dimensionless),
YI = Hpid content of earthworms (fraction),
0.66 - constant derived by Menzie et al., 1992, and
fac = the fraction of organic carbon in the mixing zone of the sewage sludge-
amended soil.
The lipid content of earthworms can be assumed to be two percent, as reported in Menzie et al
(1992) for the earthworm Lumbricus terrestris. Assuming the fraction of organic carbon in the
mixing zone of the sewage sludge-amended soil to be 0.01 (U.S. EPA 1992a) the
bioaccumulation factor in earthworms for all organic pollutants would be 3.0. This value was
not used hi Round Two, however, because it is not pollutant-specific.
Bioaccumulation of Inorganic Pollutants. Very few studies were found from which
bioaccumulation factors for the inorganic Round Two candidate pollutants could be determined
Walton (1987) investigated sodium fluoride accumulation in earthworms (primarily Lumbricus
terrestris). From the results of one of the experiments, the bioaccumulation factor for worms
(including gut contents) in soil with a high level of fluoride was calculated to be 0 670 This
value was used for the BACC for fluoride.
Helmke et al. (1979) investigated effects of land-applied sewage sludge on the
concentration of many different elements in earthworms (Aporrectodea tuberculata) Four of
the elements measured were candidate Round Two pollutants: antimony, barium, manganese
and thallium. By taking the ratio of the reported concentrations of these metals in earthworms
living in the control soil to the concentrations of the metals in the control soil, BACCs were
calculated for antimony, barium, manganese, and thallium. The authors noted that the
earthworms may not have truly accumulated these metals into their tissues. Instead the
concentrations measured were probably due to the metals in the casts (soil in the gut of the
worms). However, because predators of earthworms eat entire earthworms, including casts
Aese bioaccumulation values were considered appropriate for this analysis. Exhibit 4-15
displays the bioaccumulation factors used for the candidate Round Two pollutants. Note that
it pollutants do not have data on bioaccumulation, they are not included in the exhibit
4-43
-------
EXHIBIT 4-15
Bioaccuinulation Factors for Soil-Dwelling Organisms
Pollutant
J=— - •- • • - ----- - - - - "
Antimony
Barium
Dioxins and Dibenzofurans
Fluoride
Manganese
Thallium
-• • - ~ -^•^=^r=^=
Bioaccumulation Factor
0.13
0.062
10
0.67
0.073
0.062
,!.
Reference
Helmke et al. (1979)
Helmke et al. (1979)
U.S. EPA (1994a)
Walton (1987)
Helmke et al. (1979)
Helmke et al. (1979) .
Example Exposure Calculation for Pathway 10
The following example estimates exposure of predators of soil-dwelling organisms to
manganese in agricultural soil. . • -
Eq. 4-2:
First, the mass of soil in the mixing zone, MS, is estimated for agricultural land by using
where:
MS = 1600*g • I5cm . W^Mg/ha _ 2400Mg
J"3 kg ' cm/m3 ha
1600 = BDsoil (bulk density of soil) from Exhibit 4-3,
1 *\ -— j /-a-»_-.i_ _^ ? .• . '
15
10:1
-,
d (depth of incorporation for agricultural land) from Exhibit 4-3 and
constant to convert (kg • cm/m3) to (Mg/ha).
The concentration of manganese hi the soil is then calculated using Eq. 4-1:
4-44
-------
manganese
f342mg . 2400Afg\ + f < 1620wg , 7Mg
ha-yr) . ha
= 410mg/kg
where:
.
342 . = flSy (background concentration of manganese in soil) from Exhibit 4-4,
2400 = MS (mass of soil hi agricultural mixing zone), estimated above,
20 = N (total number of years sewage sludge is applied to agricultural land)
from Exhibit 4-3,
1620 = Cj (manganese concentration hi sewage sludge) from Exhibit 4-1, and
7 = AR (application rate of sewage sludge to agricultural land) fron/Exhibit
Hie "pollutant intake level" (PZL), and thus exposure of the HEI, is then calculated using
"
. *T~ 1 3 I
• 0.073 '
kg 3
= 10 mg/kg
where: -
n1*™ I ^ (concentration of manganese in agricultural soil), estimated above,
0.073 - &4CCy(bioaccumulation factor for manganese for soil-dwelling organisms)
from Exhibit 4-15, and
1/3 = FD (fraction of diet that is soil-dwelling organisms), discussed above.
4.2.11 Pathway 11 - Human Toxicity Through Inhalation of Particulates Resuspended by
Tilling Sewage Sludge
Pathway 11 evaluates human (tractor operator) exposure to particles that have been
resuspended by the tilling of dewatered sewage sludge into the soil. Because this type of
exposure to -"* sludge ^^ to
Methods
To ^Iculate uie total poUutant concentration in the soil for both organics and inorganics
' T° ^
4-45
-------
EXPTj = CTj • TDA - IV* ' (4-14)
where:
j = exposure of tractor operator to pollutant j (mg pollutant/m3 air),
,
concentration of pollutant j in sewage sludge-amended soil
(mg pollutant/kg sewage sludge-amended soil),
TDA = total exposure of tractor operator to soil dust (mg soil dust/m3 air) and
10"* = constant to convert (kg) to (mg).
Data Inputs
As in Round One, the total dust exposure of the tractor operator is assumed to be the
total dust standard of 10 mg/m3 established by the American Conference of Governmental
Industrial Hygienists (ACGIH). ' wvcnamanai
Example Exposure Calculation for Pathway 11
The following example estimates exposure of a tractor operator to manganese.
maSS °f S0il m ^ mixillS zone' M5' is estimated for agricultural land by using
MS = 1GOOkg - 15cin . W~lMg/ha = 2400Mg
m3 kg • cm/m3 ha
where:
1600 = BDsoil (bulk density of soil) from Exhibit 4-3,
15 =
d (depth of incorporation for agricultural land) from Exhibit 4-3 and
= constant to convert (kg • cm/m3) to (Mg/ha).
The concentration of manganese in the soil is then calculated using Eq. 4-1:
(342^ . 2400Mg\ + / . 1620mg . 7Mg
Ljfe *« I (^ kg ha-yr
where:
[2Qyr - J!&-} + 24QQ^
{ ha-yr) ha
410mg/kg
concentration of manganese in soil) from Exhibit 4-4
MS (mass of soil in agricultural mixing zone), estimated above,
4-46
-------
20 = N (total number of years sewage sludge is applied to agricultural land)-
from Exhibit 4-3,
1620 = Cj (manganese concentration in sewage sludge) from Exhibit 4-1, and
7 = AR (application rate of sewage sludge to agricultural land) from Exhibit 4-
3. • • •
Using Eq. 4-14, exposure of the tractor operator to manganese is then estimated as:
• 10 • !()•«-
g m3 mg
= 4.b:10-3 mg/m3
where: •
410 = CTj (concentration of manganese in agricultural soil), estimated above.
10 = TDA (total exposure of tractor operator to soil dust), discussed above, and
10'6 = constant to convert (kg) to (mg).
4-47
-------
4.2.12 Pathway 12 - Ingestion of Fish and Water from Surface Water that Receives
Eroded Soil
Pathway 12 evaluates human ingestion of fish and water from surface water that receives
eroded soil from sewage sludge-amended agricultural and non-agricultural lands. Non-
agricultural lands include forests, reclamation sites, and public contact sites.
To estimate exposure for this pathway, a mass balance analysis is required. This mass
balance analysis accounts for the partitioning of pollutants into different soil phases (solids, air,
and water) and the subsequent losses of pollutants from the land application site. Pollutants are
lost from a land application site by: erosion of soil particles, which releases sorbed pollutants
into surface waters; volatilization of pollutants into air; leaching of pollutants into groundwater;
and degradation. A mass balance for a pollutant must be maintained, given these four competing
loss processes of erosion, volatilization, leaching, and degradation. Once mass balances for
pollutants have been established, exposures to pollutants that have eroded, volatilized, or leached
are calculated under three separate pathways: surface water (Pathway 12), air (Pathway 13), and
groundwater (Pathway 14). It is assumed that if pollutants degrade, they degrade into chemicals
that do not pose unacceptable risks to public health or the environment.
The methods for performing the mass balance calculation for Pathways 12, 13, and 14
are discussed below. Subsequent to this discussion, the equations particular to Pathway 12, for
estimating the maximum amount of pollutant available for erosion at a site and the transport of
that pollutant mass to a surface water stream, are presented.
Method for Mass Balance (Pathways 12, 13, and 14)
There are two major steps involved in the mass balance calculation. First, a pollutant
is partitioned among the three phases present hi a soil: solids, air, and water. Second, the rates
at which the four loss processes occur (erosion, volatilization, leaching, and degradation) are
estimated. These .two steps are described in the next two subsections.
Step 1: Partitioning of Pollutant Among Solids, Air, and Water in Sewage Sludge-
Amended Soil. This section describes the methods for partitioning a pollutant in the sewage
sludge-amended soil at a land application site, assuming that equilibrium is maintained among
the pollutant concentration sorbed onto soil particles (which erodes into surface water), the
pollutant concentration in the air-filled pore space (which volatilizes), and the pollutant
concentration in the porewater (which leaches to groundwater).
Equilibrium partitioning between sorbed and dissolved phases is described by soil-water
partition coefficients; partitioning between dissolved and gaseous phases is described by Henry's
Law constants. From these assumptions and the definitions of concentrations in different phases
presented below, equations are derived to describe pollutant partitioning among all of the phases.
4-48
-------
Mathematically, pollutant concentrations in different phases can be expressed as:
C =
Ms
s
and:
C{ = ~f = V 1 V^r ^ (4"16)
where: .
C, = concentration of sorbed pollutant .on sewage sludge-amended soil particles
(kg pollutant/kg soil),
Mcs = mass of sorbed pollutant (kg),
M-s = mass of soil (kg),
Vs = volume of solids in soil (m3),
Ca = concentration of gaseous pollutant in air-filled pore space of sewage
sludge-amended soil (kg pollutant/m3 air),
Mia = mass of gaseous, pollutant (kg),
Va = volume of air in soil (m3),
Cu. = concentration of dissolved pollutant in water-filled pore space of sewage
sludge-amended soil (kg pollutant/m3 porewater),
Mm. = mass of dissolved pollutant (kg),
K = volume of water in soil (m3),
C, = total concentration of pollutant in bulk sewage sludge-amended soil (kg
pollutant/m3 total bulk soil volume),
Ma = total mass of pollutant in soil (kg), and
V, = total bulk volume of soil (m3).
The definitions of equilibrium partition coefficients and soil characteristics, such as bulk
density and porosity, are used in conjunction with Eqs. 4-15 and 4-16 to estimate the pollutant
concentrations in each soil phase (solids, air, and water). The equilibrium partition or
distribution coefficient (Kd), describing the partitioning of a pollutant between pollutant sorbed
on solids and pollutant dissolved in porewater, can be defined as:
(4-17)
where:
Kd = soil-water partition coefficient (L water/kg soil), and
103 = constant to convert (m3) to (L).
4-49
-------
The dimensionless Henry's Law constant, which describes the partitioning of a pollutant
between gaseous and dissolved phases, is defined as:
= ,. _. (4-18)
l*"cw/ r *J Me* Va
where:
H = Henry's Law constant (dimensionless).
The bulk density of soil is defined as:
= MJ Vt (4-19)
where: • .
#Anir = bulk density of sewage sludge-amended soil (kg soil/m3 total bulk soil
volume).
The air-filled porosity of soil is defined as: . • -
0C = va I Vt (4-20)
where: .
^a = air-filled porosity (dimensionless).
The water-filled porosity is defined as:
ew = rw/*V (4-21)
where:
0W = water-fmed porosity (dimensionless).
And, the total porosity of soil is defined as:
* * S t O W \ ™<«'^' t
where: . .
^t — total soil porosity (dimensionless).
By combining Eqs. 4-15 through 4-22, equations that describe the pollutant concentrations in the
air and water phases in terms of the total pollutant concentration can be derivS:
4-50
-------
c . s • •
a Kd BDmiI _3 6M. (4-23)
~~5 x +1+ 6a
and:
Cw = - - - - ^— - — (4-24)
'
Kd • ID
Equations 4-23 and 4-24 are used to estimate the first-order rate constants for
volatilization and leaching in the second step of. the mass balance calculation.
It is important to note that in these derivations, Cr, the total pollutant concentration in
sewage sludge-amended soil, is expressed as mass per volume. Recall that in Pathways 1
through 1 1 , the total pollutant concentration is expressed as mass per mass of soil (CTJ). These
two quantities can be related through the following equation:
Ct = CTj • BD^ • 1
-------
Ku>< = *m + *„, + KUC + K*8 (4-26)
where:
KM = total loss rate for.the pollutant from sewage sludge-amended land (yr1),
Kcro *= loss rate due to erosion of the pollutant from sewage sludge-amended land
K^i - loss rate due to volatilization of the pollutant from sewage sludge-amended
land (yr1), .
K& - loss rate due to leaching of the pollutant from sewage sludge-amended
land (yr1), and .
K
-------
where:
dt = average rate of soil loss due to erosion from sewage sludge-amended land
each year (m/yr),
d ~ dePtih of incorporation of sewage sludge (cm), and
10'2 = constant to convert (cm) to (m).
Note that this assumes that the loss to erosion includes pollutant mass in all three phases
(solids, air, and water); therefore, the first-order loss coefficient is not pollutant-specific That
is, given the assumption of even incorporation, if one-tenth of the sewage sludge-soil mixture
is removed by erosion, one-tenth of the mass of pollutant is also removed.
First-Order Loss Rate for Volatilization. Estimates of volatile emissions from
uncovered soil are used in conjunction with estimates of pollutant concentrations in the air-filled
pore space of soil to estimate the loss rate coefficient for volatilization, Kml.
,i«o^ ?timates of volatile emissions are based on equations provided by Hwang and Falco
(1986) for contaminated soil with no cover:
2te6eD C
Na —-I * (4-29)
where: . '
Na = total average emissions from the soil surface over time interval t
(kg pollutant/m2 soil), f
le — duration of emissions (sec),
6e = effective porosity of soil (dimensionless),
Dei = intermediate diffusivity variable (defined in Eq. 4-30) (nrVsec),
Q = concentration of gaseous pollutant in air-filled pore space of sewage
sludge-amended soil (kg pollutant/m3 ah-), and
a-, = intermediate diffusivity variable (defined in Eq. 4-31) (m2/sec).
In Hwang and Falco (1986), C9 is estimated from the concentration of sorbed pollutant (C)
However this analysis requires the relationship between the total concentration of pollutant *in
sewage sludge-amended soil (in sorbed, gaseous, and dissolved phases) and the concentration
m gaseous phase within the soil's air-filled pore space. Therefore, Eq. 4-23 is used ^e*
foil™-™6 1interm;;dia* diffusivity variables required in Eq. 4-29 are obtained from the
following relationships (Hwang and Falco, 1986):
4-53
-------
and:
Dei =
(4-30)
and:
««=
Kd'io-3/H
(4-31)
where:
(4-32)
and where:
H
T
lO'3
H
R • r • 10
-3
(4-33)
the molecular diffusivity of pollutant in air (cm2/sec),
constant to convert units fronv(cm2) to (m2),
effective porosity of soil (dimensionless),
particle density of sewage sludge-soil mixture (kg/m3),
soil-water partition coefficient (L water/kg soil),
Henry's Law constant (dimensionless),
bulk density of sewage sludge-amended soil (kg soil/m3 total bulk soil
volume),
total soil porosity (dimensionless),
Henry's Law constant (atm-m3/mol),
gas constant (L-atm/mol-K),
temperature (Kelvin), and
constant to convert. (L) to (m3).
Equation 4-29 provides an estimate of total average emissions from an uncovered layer
of soil as a function of both time and the initial concentration of pollutant. For consistency with
methods used to estimate losses for other pathways, Eq. 4-29 is evaluated for te equal to 1 year
(r^3.2x!07 sec), and results are used to estimate an annual loss coefficient. Losses predicted
for the first year (Afap are divided by the total mass of pollutant hi soil to estimate the
approximate fraction of available pollutant lost per unit of time. For a unit concentration (1
kg/m) of the pollutant hi soil (i.e., Q, the mass of pollutant beneath one square meter of soil
surface (in kg/m2) is equal to the volume of treated soil beneath a square meter of surface (m3
4-54
-------
per nr), that is, equal to the depth of incorporation (m). The estimated loss rate (in kg/nr-yr)
is approximated as a comparable first-order loss, coefficient (in yr1) as:
. / Na \
K^ * -In 1- Jl_ , . (4-34)
I d • 10-2J .
where:
K^ = loss rate due to volatilization (yr1),
Nay = total average emissions from the soil surface in first year
(kg pollutant/m2), estimated using Eq. 4-29,
d = depth of incorporation of sewage sludge (cm. When converted to (m),
equivalent to kg/m2 for a unit concentration of pollutant hi sewage sludge-
amended soil), and
10'2 = constant to convert (cm) to (m).
Because Eq. 4-29 was derived by assuming the column of soil is of infinite depth it can predict
greater than 100 percent loss within a year for a relatively shallow layer of treated soil and a
relatively volatile pollutant. For such cases, Eq. 4-34 cannot be evaluated and the loss rate
coefficient is instead estimated from predicted emissions in the first second (t = 1 sec) The loss
rate coefficient estimated from the first second of emissions is then converted to an annual loss
rate coefficient:
K^ * -3.2xl07lnfl- Na'. } (4-35)
( d-W2) .
where: .
. Na* = emissions from the soil surface in first second (kg pollutant/m2)
estimated using Eq. 4-29,
3.2xl07 = constant to convert (sec'1) to (yr1), and
10~2 - constant to convert (cm) to (m).
' Equation 4-34 was used to evaluate carbon disulfide, dioxins and dibenzofurans
endosulfan, pentachloronitrobenzene, PCBs, and 2-(2,4,5-trichloroPhenoxy) propionic acid
However, for the other Round Two pollutants that volatilize, Eq. 4-35 was used.
First-Order Loss Rate for Leaching. To estimate pollutant loss to leaching the
fo owing Aquation which computes a first-order loss rate for a pollutant leaching from treated
soil (US. EPA, 1989f), was modified to take into account that leaching is only one of four
competing pollutant loss processes:
4-55
-------
NR
Kd
where:
K& == loss rate due to leaching of the pollutant from sewage sludge-amended
land (yr1), . .
NR = annual recharge to groundwater beneath the treated soil (m3 recharge/m2
area-yr, or m recharge/yr),
BDma ' = bulk density of sewage sludge-amended soil (kg ,soil/m3 total bulk soil
volume),
Kd = soil-water partition coefficient (L water/kg soil),
d = depth of incorporation of sewage sludge (cm), .
10'3 = constant to convert (L) to (m3), and
10"2 = constant to convert (cm) to (m).
To derive a coefficient for first-order loss to leaching while maintaining a mass balance,
the mass of pollutant expected to be lost each year is estimated and divided by the available mass
of pollutant. The mass of pollutant that will be lost to leaching in any interval of time per unit
area (i.e., the flux of pollutant) can be described by the volume of water percolating through the
treated soil multiplied by the average concentration of pollutant in that water: • -
. FAUc = NR • CUc • 10,000 (4-37)
where: ' ,
FA,ec = annual average flux of pollutant leached from sewage sludge-
amended soil (kg pollutant/ha-yr),
Ore = concentration of pollutant in water leaching from sewage sludge-
amended soil (kg pollutant/m3 porewater), and
10,000 = constant to convert units from (kg/m2-yr) to (kg/ha-yr).
Assuming that all the porewater forms leachate, Eq. 4-24 can be used to estimate the
pollutant concentration in the leachate: .
C"^°>>^-ie.^ej
where:
CH. = concentration of dissolved pollutant in water-filled pore space of sewage
sludge-amended soil (kg pollutant/m3 porewater),
' Q = total concentration of pollutant in bulk sewage sludge-amended soil
(kg pollutant/m3 total bulk soil volume),
0W = water-filled porosity (dimensionless),
4-56
-------
H = Henry's Law constant (dimensionless),
. 0a = air-filled porosity (dimensionless), and
10'3 = constant to convert (L) to (m3).
Given that first-order loss rates are being assumed, the total concentration of pollutant in soil
decreases due to leaching according to the following equation:
dC
~dt
1 = ~KUcCt (4-39)
KUc can be estimated with the discrete approximation:
».; — —— — • ' = > I i~r~*T\))
f f \g
ui <-, Az_ r
where:
t — time (yr).
The change in the total pollutant mass in the soil with time can be expressed in terms of the
pollutant flux leaching to the groundwater:
AM"
^* ^ ^y A A « f\'*& ' / A A •< \
--£- -^fa'^ -10-4 (4-41)
where: .
A34/A/ = change in total mass of poUutant in soil over time interval of one
year(kg/yr),
A — • area of land application site (m2), and
10^ = constant to convert units from (m2) to (ha).
Combining Eqs. 4-40 and 4-41:
. • A - ID"4
-r; ' (4-42)
4-57
-------
Given that the total bulk volume of sewage sludge-amended soil equals the area of the land
application site multiplied by the depth of incorporation:
Vt = A • d • 1(T2
where:
Vt = total bulk volume of soil (m3),
and rearranging Eq. 4-16:
the following equation can be written:
Using Eqs. 4-37, 4-38, and 4-43, Eq. 4-42 can be rewritten as:
NRCu< _ NR
(A-AA\
( '
Vmix'Kd* QW + & Oj d • ID'2
Equation 4-44 is used to predict the rate of pollutant loss to leaching. -
First-Order Loss Rate for Degradation. Values of K^ are obtained from the literature
for each pollutant. For K^, rates of abiotic reactions such as hydrolysis are preferentially used
because they can be more reliably measured; for pollutants only degraded by microbes, the
lowest aerobic rates measured under environmental conditions are used. For inorganics' and
some persistent organic pollutants such as dioxins and dibenzofurans, K^. is zero.
Methods Specific to Pathway 12
Overview of Methods Specific to Pathway 12. To calculate the average concentration
of a pollutant in eroding soil, estimates of the maximum mass of pollutant available for erosion
from a land application site are combined with estimates of the rate at which the pollutant is lost
from the site due to the four loss processes.over a human lifetime. The average concentration
of pollutant in eroded, sewage sludge-amended soil is then "diluted" by the erosion of clean soil
from the remainder of the watershed. Estimates of the concentration of pollutants in the stream
receiving the eroded soil are then made, derived from the mass of pollutant on eroded soil and
4-58
-------
on soil- water partition coefficients. Pollutants then partition into fish inhabiting the stream.
Humans are exposed to pollutants through both direct ingestion of surface water and ingestion
of fish. For this analysis, organics as well as inorganics are allowed to build up in the soil over
the active life of the land application site.
Maximum Pollutant Mass Available for Erosion. The maximum mass of pollutant
available for erosion at a land application site occurs after the final application of sewage sludge
to the site. This maximum mass is estimated as:
TPN = AR • Cj • (\+e~c'Ka*+e~*c.'K**+...+e~b' *'***) • 1000 • 1 (4-45)
where:
TPN = total mass of pollutant available at a site after the final year of application
(mg pollutant/ha), .
AR = annual whole sludge application rate of sewage sludge to land (dry Mg
sewage sludge/ha-yr),
Cj = concentration of pollutant j in sewage sludge (mg pollutant/kg sewage
sludge),
c = application interval (number of years between applications),
Ktot = total loss rate for the pollutant from sewage sludge-amended land (yr1),
b = model parameter that is equal to the integer part of ((^te-l)/c), where N ~
is the site life (yr), .. • «*
1000 = constant to convert (Mg) to (kg), and
1 =• interval of time over which pollutant loss is evaluated (yr).
Losses to Erosion Averaged Over a Human Lifetime. Because human exposure is
assumed to continue for the duration of an individual's lifetime, concentrations of pollutants in
surface water are calculated based on losses of pollutant through surface erosion for a period
equal to the human life expectancy. Therefore, loss to erosion during the period between the
end of land application and the end of an individual's lifetime must be estimated. To do so the
mass of pollutant left at the end of an individual lifetime is first calculated. After the' last
application of sewage sludge to a site, a pollutant continues to be depleted according to the
following equation:
MLS = TPN •e~*-V*~ir"''> (4-46)
where:
MU = mass of pollutant in soil at end of a period equal to an individual lifetime
(mg pollutant/ha),
LS = average human lifetime (yr), and
Nsiu = site life (yr).
4-59
-------
The fraction of total, cumulative loading lost to all four loss processes in the human
lifetime (over both the application and post-application periods) can then be calculated as:
frs =
where:
' AR ' 100°) "
N • C, • AR '• 1000
(4-47)
JLS
fraction of total cumulative loading lost in individual's lifetime to all four
loss processes (dimensionless). .
This fraction is used'to estimate the average pollutant concentration in eroded soil for
both organic and inorganic pollutants. The estimated total loading of pollutant is multiplied by
the fraction expected to be lost to erosion in the human lifetime (f^ -/^), and divided by the
total mass of eroded soil lost during that period to calculate the expected average concentration
of each pollutant in eroded soil:
where:
•"me. j
ME,.
"site, j
ILS
(4-48)
the concentration of pollutant j in sewage sludge-amended soil
eroded from the land application site (mg pollutant/kg sewage
sludge-amended soil),
rate of soil loss for land treated with sewage sludge (kg sewage
sludge-amended soil/ha-yr), and
where:
BD.
10,000
'10,000
(4-49)
average rate of soil loss due to erosion from sewage sludge-
amended land each year (m/yr),
bulk density of sewage sludge-amended soil (kg soil/m3 total bulk
soil volume), and
constant to convert (m2) to (ha).
Note that Eq. 4-48 assumes that the same mass of pollutant leaves the site by erosion every year.
Dilution of Eroded, Sewage Sludge-Amended Soil with Non-Sewage Sludge-Amended
Soil. The extent to which eroded soil from the land application site is "diluted" by soil from
the untreated remainder of the watershed also needs to be estimated. A "dilution factor"
4-60
-------
describes the fraction of the total eroded soil in the watershed originating in the land application
site:
where:
Df = dilution factor (dimensionless),
Asite = area of land application site (ha),
Aws = area of the watershed (ha),
Ssite = sediment delivery ratio for the land application site (dimensionless),
Sws = sediment delivery ratio for the watershed (dimensionless), and
ME** - estimated rate of soil loss (erosion) for the watershed (kg soil/ha-yr).
If the rates of soil erosion from the land application site and the remainder of the watershed are
assumed to be the same, ME!ite and MEWS cancel from Eq. 4-50, and the dilution factor can be
calculated by:
Df = :
The sediment delivery ratios for the land application site arid the watershed are calculated
with the following empirical relationship (Vanoni, 1975):
S = 0.872 A-*135
Thus, the sediment delivery ratio for the site and for the watershed are:
S^ = 0.872 [A^-*-125 (4-52)
and
= 0.872 HJ-0'125 (4-53)
4-61
-------
If all of a pollutant entering the stream on eroded soil is assumed to originate from the •»
land application site, the dilution factor, Df, also describes the ratio between the average ^^
concentration of pollutant hi soil entering the stream and the average concentration in soil
eroding from the land application site:
C. = D • C" (4,-^£\
esoil,j ^f site,j (,t*-Jtt)
where:
weight concentration of pollutant/ in eroded soil entering the stream
(mg pollutant/kg eroded soil).
Pollutant Concentration in the Stream. The estimated concentration of pollutant in the
eroded soil is used as an input to calculate the expected concentration of pollutant in the stream.
Once the eroded soil enters the stream, the pollutant is assumed to partition between the solid
and liquid phases of the stream according to equilibrium conditions. The total amount of
pollutant available.to partition depends on the amount of eroded soil in the stream. Assuming
that all of the total suspended solids in the stream are from eroded soil particles, Eq. 4-55
partitions the total mass of pollutant between dissolved and sorbed phases. The lefthand term
represents the total mass of pollutant (per liter of water) entering the stream hi eroded soil; the
middle term represents the dissolved pollutant; and, the right-most term represents the mass of
sorbed pollutant (per liter of water):
10"6 =
-------
*<-*<*;/«< (4'57)
where:
Koc = organic carbon-water partition coefficient (mL water/g organic carbon),
and v .
f^ = fraction of organic carbon in suspended solids (dimensionless, g organic
carbon/g suspended solids).
*
A value of 0.01 is used for the .4 of suspended solids, to correspond to the .4 of -the mixing
zone from which the suspended solids are assumed to have eroded (U.S. EPA, 1993a)
To estimate an organic pollutant's Kx value, an empirical regression equation presented
by Hasset et al. (1983) is used:
loSio(*oc) ".0.088 ' + 0.909 log^^) (4-58)
where:
K^ = octanol-water partition coefficient (dimensionless, mg pollutant/L octanol
per mg pollutant/L water).
Partition Coefficients for Inorganic Pollutants. For inorganic pollutants, it is much
more difficult to predict a "typical" Kd value. Metals can sorb onto soils through the processes
of ion exchange, specific adsorption, co-precipitation with hydrous oxides, and incorporation into
canonic lattice sites in crystalline sediments (Bodek et al., 1988). Therefore, clay minerals
organic matter, and manganese and iron oxides are .all important sorbents of metals in soil
(Bodek et al. 1988). The pH of the system also affects metal sorption, with most metals
tending to sorb more at higher pHs. Therefore, measured Kd values reported in the literature
tor conditions similar to those being modeled are used.
Estimating Human Exposure
Humans can be exposed to surface water contaminated by sewage sludge through two
pathways: ingestion of water and ingestion of fish. Potential exposure through ingestion of .
contaminated surface water is calculated as: s^uuu "i
(4-59)
on
where:
exposure to pollutant j through direct ingestion of surface water
(mg pollutant/kg body weight-day),
4-63
-------
Cwj = concentration of pollutant j in surface water
(mg pollutant/L water),
IW = quantity of water ingested daily (L water/day), and
BW = body weight (kg), assumed to be 70 kg.
Exposure through ingestion of fish is calculated based on estimates of the bioaccumulation
of a pollutant in fish and the assumed rate of fish ingestion. Bioaccumulation is the process by
which aquatic organisms accumulate pollutants, from both water and food, at concentrations
higher than the ambient concentration. The process by which a pollutant is absorbed from water
through gill membranes or other external body surfaces is called bioconcentration, and the
measure of a pollutant's tendency to bioconcentrate is described by the bioconcentration factor.
For organic pollutants, a regression equation based on logCK^) values is used to estimate
bioconcentration factors (BCFs). The equation was developed for a three percent limd content
of fish (U.S. EPA, 1990): H
log10(£CP) = 0.79 log^*^ - 0.80 (4-60)
where:
BCF — pollutant-specific bioconcentration factor (L water/kg fish).
For inorganic pollutants, available literature values were used for BCFs.
Biomagnification denotes the process by which the concentration of a pollutant increases
hi different organisms occupying successive trophic levels. The combined accumulation from
bioconcentration and biomagnification is represented by the bioaccumulation factor, which
calculated as the product of the bioconcentration factor and a food chain multiplier: '
is
BAF = BCF-FM (4-61)
r
where:
BAF = pollutant-specific bioaccumulation factor (L water/kg fish), and
FM = pollutant-specific food chain multiplier (dimensionless).
Assuming that humans only ingest fish fillets, and not the whole fish, the pollutant
concentration in fish fillets can be expressed as:
CJKl * c~j ' *** ' Pf (4-62)
where: '
CKJ = concentration of pollutant/ in fish fillets (mg pollutant/kg fish fillet), and
4-64
-------
•Pf = ratio of pollutant concentration in fillet to whole fish (dimensionless).
Thus, human exposure through ingestion of contaminated fish can be expressed as:
EXPf . = ffJ'IF (4-63)
fj BW
where:
' EXPf.j ~ exposure to pollutant/ through ingestion of fish (mg pollutant/kg body
weight-day), and
IF = daily consumption of fish fillets (kg fish fillets/day).
For this analysis of the surface water pathway, exposures through drinking water
consumption and fish ingestion are combined:
EXPJ = EXPWJ + EXPfJ . . . (4-64)
where': . . _
EXPj = exposure to pollutant/ through consumption of both surface water and fish
combined (mg pollutant/kg body weight-day).
Data Inputs
Both non-pollutant-specific and pollutant-specific inputs are required for the exposure
equations described above. Values for non-pollutant-specific inputs, such as the area of a
watershed, its hydrogeological characteristics, and the daily consumption of fish and drinking
water, are presented hi Exhibit 4-16. Note that all parameters necessary for Pathways 12 13
and 14 are presented in this Exhibit.
There are several pollutant-specific fate and transport parameters required to maintain the
mass balance of a pollutant among the four loss processes and to estimate the rates at that those
four loss processes occur. In Exhibit 4-17, all of the fate and transport parameters are
presented.
To obtain estimates of inorganic Kd values for six Round Two pollutants studies of
adsorption described in Gerritse et al. (1982) were used. Gerritse et al. present a range of Kd
values for various inorganics in two soil types: sand and sandy loam. In the sandy soil there
was 0.035 g/g organic matter, 0 g/g clay, 0.22 meq/g cation exchange capacity (CEC), and the
porewater had a pH of 5. In the sandy loam soil, there was 0.025 g/g organic matter, 0.2 g/g
clay, 0. 16 meq/g CEC, and the porewater had a pH of 8. For this analysis, the Kd values from
sand, which were lower than those in sandy loam, were used. In addition, the lowest Kd value
from the range available for each of the six Round Two inorganics tested was used.
4-65
•d
-------
For aluminum and fluoride, available data on Langmuir isotherm parameters were used
to estimate Kd values (Bodek et al., 1988). For aluminum, data were for silica, at a pH of 5.
For fluoride, data corresponded to clay loam, containing 10.4 percent clay, 0.94 percent organic
carbon, and 825 /zg/g aluminum, with a pH of 5.9. For boron, thallium, tin, and titanium, Kd
values were not available. ' • '
In the absence of pollutant-specific data for the ratio of pollutant concentration in fillet
to the concentration in whole fish, it is assumed that these concentrations are the same (Pf = 1)
for all pollutants except PCBs and dioxins. PCBs are assumed to behave similarly to dioxins,
for which a ratio of 0.5 has been estimated (Branson et al., 1985).
For BCF values for inorganic pollutants, the Ambient Aquatic Life Water Quality Criteria
documents were reviewed. Only three Round Two inorganics had such documents available:
aluminum, antimony, and silver. For aluminum, bibconcentration factors for young brook trout
were reported to range from 50 to 231 (U.S. EPA, 1988a). The geometric mean (107) was used
in this analysis. For antimony, one study on bioconcentration in bluegill found no significant
accumulation above controls (U.S. EPA, 1988b). AQUIRE (Aquatic Toxicity Information
Retrieval) was then searched for BCF data on antimony. The AQUIRE run turned up values for
one saltwater fish and one fish that may or may not be saltwater. The BCFs for the shanny
(Blennius pholis) and for the two-spot goby (Gobiusculus flavescens) are 0.40 and. 0.15,
respectively. The geometric mean of these two values (0.24) was used in this analysis. For
silver, the Ambient Aquatic Life Water Quality Criteria for Silver document (U.S. EPA, 1987)
had BCF information for two freshwater fish species, bass and bluegill. The geometric mean
of the bass BCFs (11 and 19) is 14; the geometric mean of the bluegill BCFs (15 and 150) is 47.
The geometric mean of the two species' BCFs is 26; this value was used in this analysis.
AQUIRE was then searched for all the remaining inorganics (Ba, Be, B, F, Mn, Th, Sn
Ti, and V). No data on BCF values for freshwater fish were found.
For FM values, if an organic pollutant had a log(Kow) value less than or equal to five, a
value of one was used; otherwise a value of ten was used for FM (U.S. EPA, 1990). This
relationship is applicable to a species on a trophic level of three. For inorganic'pollutants an
FM value of one was used.
4-66
-------
EXHIBIT 4-16
Non-Pollutant-Specific Parameters for Pathways 12, 13, and 14
Parameter
LJ :
r<
k
ra
K
•^site
II -^ws
jjc
U
1
" '
\\TSS
I
11
1 T
UF
\\IA
r-
1**
r
Definition
1 total soil porosity
effective soil porosity
air-filled porosity
water-filled porosity
area of land treated with
sewage sludge
area of watershed
application interval
yearly depth of soil eroded
annual recharge to
groundwater
density of water
total suspended solids in
surface water
angle subtended by the
land application site's
width
average wind velocity
vertical dispersion of
pollutant in air
average air temperature
daily consumption of fish
fillets
daily inhalation rate
daily ingestion of water
site life
]
gas constant <
|]
Value
— — ^—
0.4 (dimensionless)
0.4 (dimensionless)
0.2 (dimensionless)
0.2 (dimensionless)
1,074 ha
440,300 ha
varies
6 x 10-4 m/yr
0.5 m/yr.
1 kg/L
16 mg/L
22.5°
4.5 m/sec
1 (dimensionless)
288 Kelvin
0.04 kg/day
20m3/day
2 L/day
ZOyrag., forest,
sub., 1 yr reel.
3.082 L-atm/mol- •
Kelvin
1
Reference
4 — 1 =
U.S. EPA, 1992a
U.S. EPA. 1992a 1
U.S. EPA, 1992a 1
U.S. EPA, 1992a |
1 U.S. EPA, 1992a
U.S. EPA, 1992a
See # in Exhibit 4-3 ||
USDA, 1987 I
U.S. EPA, 1992a
approximate density of -
water under
environmental conditions
U.S. EPA, 1992a
U.S. EPA, 1992a
U.S. EPA, 1992a ~]|
U.S. EPA, 1992a 1
U.S. EPA, 1992a
U.S. EPA, 1992a
1
U.S. EPA, 1992a
U.S. EPA, 1992a T
" : 1
constant
4-67
-------
4-68
-------
4-69
-------
•• C
o
S
c-
t- —
5 I
H S
II
65
C
4-70
-------
I
C
u
00 •
o
T3
n
«
_0
5
I
£
2
ts
s
o
§N
£
-a
c
ts
c
o
I
1
CO
S
ON
W
I
o
ON
ON
1
O
s.
.$>
g
o 3
tu a.
S s
5
b
si
s
oo
ON
•-- «n .
2 2! R
& 52 s
*i^5ff§
llill;-!^
«•!
g s
If
u
•S S
m CQ
« S
Jf 3
S 5 5
4-71
-------
Example Exposure Calculations for Pathway 12
The following example calculates exposure of humans to dioxins and dibenzofurans
through ingestion of water and fish from surface water receiving eroded sewage sludge-amended
soil from agricultural land.
Step 1: Partitioning of Pollutant
In Step 1 of the mass balance calculation, relationships among pollutant concentrations
in the bulk sewage sludge-amended soil (Q, in the air-filled pore space (Q, and in the water-
filled pore space (CJ are derived. In this example calculation, these relationships are used to
estimate Kwl and KlK in Step 2. •
Step 2: Estimation of Km
Equation 4-28 can be used to calculate the loss rate coefficient for erosion (Kero):
j, 6xlO"4 mlyr ,.,«-•? i
Kero = —LZ. = 4x10 3 yr'1
15 cm • 10~2 m/cm
where:
6x1 Q-4 = de (average rate of soil loss due to erosion from sewage sludee-amended
land each year) from Exhibit 4-16,
15 = d (depth of incorporation for sewage sludge on agricultural land) from
- Exhibit 4-3, and
10': = constant to convert (cm) to (m).
Step 2: Estimation of Knl
Several equations are used to calculate the loss rate coefficient for volatilization (Kml).
First, the intermediate diffusivity variable £>„ is calculated from Eq. 4-30:
Dei = 4.4x10~2 cm2/sec • 10'4 m/cm2 • (0.4)1/3 = 3.2x10~6 m2/sec
where: . •
4.4xlO'2 = bm (diffusivity of pollutant in air) from Exhibit 4-17,
10"4 = factor to convert (cm2) to (m2), and
°-4 = Oe (effective porosity of soil) from Exhibit 4-16.
4-72
-------
4-32:
Second, the particle density of the sewage sludge-soil mixture GO is calculated using Eq
where:
1400 =
0.4 =
* (bulk density of sewage sludge-amended soil) from Exhibit 4-3 and
6, (total porosity of soil) from Exhibit 4-16.
Third, the dimensionless Henry's Law constant (H) is calculated using Eq. 4-33:
H =
6.SxlQ-5atm-m3Imol
(0.082 L-atm/mol-K) - 288* • 0.001 m?IL
= 2.9xlO~3
where:
6.8xlO'5
0.082
288
0.001
H (Henry's Law constant) from Exhibit 4-17,
R (universal gas constant) from Exhibit 4-16,'
T (average air temperature in Kelvin) from Exhibit 4-16 and
factor to convert (L) to (m3).
Fourth, the intermediate diffusivity variable a, is calculated from Eq. 4-31:
a. = 3.2x70•* m2/sec • 0.4
0.4
2-3xl°3
(1-Q.4) • 13.000 Lfke
= 2.0x10 -13m2/sec
1000 L/m
where:
3,2x10-*
0.4
2.3X103
13,000
2.9xlO'3
1000
Dj (calculated above),
6e (effective porosity of soil) from Exhibit 4-16
pss (calculated above),
Kd (partition coefficient for dioxins between water and soil) from
Exhibit 4-17, ~ "
A (calculated above), and
constant to convert (m3) to (L).
4-73
-------
Next, to calculate Ca, Eq. 4-23 is used with a unit concentration of lkg/m3 for CT:
kg/m:
[13,000 L/kg • 1400 kg/m3 • 0.001 m3/L\ + 0.2
2.9x70 -3 2.9x70 '3
Q
where:
1
13000
1400
0.001
2.9xlO'3
0.2
0.2
C, (unit concentration of pollutant in bulk sewage sludge-amended
soil),
Kd (soil-water partition coefficient) from Exhibit 4-17,
BD^ (bulk density of sewage sludge-amended soil) from Exhibit
4-3,
constant to convert (L) to (m3),
H (diniensionless Henry's Law constant) calculated above,
0W (water-filled porosity) from Exhibit 4-16, and
0a (air-filled porosity) from Exhibit 4-16.
4-29:
Total average emissions from the soil surface hi one year are then calculated using Eq.
Na = 2 • 31,536,000 sec • 0.4 • 3.2x70 "6 /K2/sec • 1.6x2Q-7 kgfm3
2.9x10
V/7t • 2.0x70 ~13 m2/sec • 31,536,000 sec
where:
31,536,000 =
0.4
3.2x10-*
1.6xlO'7
2.0xlO:'3
te (duration of emissions), corresponding to one year,
0e (effective porosity of soil) from Exhibit 4-16,
Da (intermediate diffusivity variable) calculated above,
Ca (concentration of dioxins in air-filled pore space) calculated
above, and
a, (intermediate diffusivity variable) calculated above.
4-74
-------
Finally, K^ can be calculated using Eq. 4-34:
2.9x10-*
0.15 kg/m2.
= 1.9x70
where:
2.9xlO'3
0. 15
Nay (total average emissions from the soil surface in the first year),
calculated above, and
d (depth of incorporation of sewage sludge) from Exhibit 4-3; see
text for explanation of why this may be expressed as a mass per
area.
Step 2: Estimation of
Equation 4-44 can be used to approximate the loss rate coefficient for leaching
KUc «
0.5 m/yr
_
[1400 kg/m3 • 13,000 L/kg • 0.001 m3/L + 0.2 + 2.9xlO~3 • 0.2] • 0.15
= 1.8x70-* yr~\
m
where:
°-5
1400
13,000
0-°°1
°-2
°-2
2.9xlO'3
°- 15
Step 2: Estimation of K
M
NR (annual recharge to ground water) from Exhibit 4-16,
BD^ (bulk density of sewage sludge-amended soil) from Exhibit
4-3, -
Kd (soil-water partition coefficient) from Exhibit 4-17,
constant to convert (L) to (m3),
0* (water-filled porosity) from Exhibit 4-16,
^? (air-filled porosity) from Exhibit 4-16,
H (dimensionless Henry's law constant) calculated above, and
d (dePth of hicorporation of sewage sludge) from Exhibit 4-3.
The total loss rate for dioxins from the sewage sludge-amended agricultural land is
calculated from Eq. 4-26:
4-75
-------
Ktot = 4*10
~l + 1.8x10 ~* yr~l
= 2.3x70 ~2 yr
~2 '1
where:
4x10°
1.9xlO-2
1.8x10"*
0
Kero (loss coefficient for erosion) calculated above,
Kwl (loss coefficient for volatilization) calculated above,
KlK (loss coefficient for leaching) calculated above, and
Kdeg (loss coefficient for degradation) from Exhibit 4-17.
Maximum Pollutant Mass Available for Erosion
Once Ktot is calculated, it is then used in Eq. 4-45 to calculate the maximum mass of
pollutant onsite:
2J*S_ .
35 mg/ha
kg
kg_ .
Mg
where:
S.llxlO"4
1
2.3xlO'2
19
AR (application rate) from Exhibit 4-3,
Cj (concentration of dioxins in sewage sludge) from Exhibit 4-1,
c (application interval) .from Exhibit 4-3,
Km (total loss rate for dioxins) calculated above,
b (equal to integer part of (AT-l)/c, where TV is the site life), from
Exhibit 4-3,
constant to convert (Mg) to (kg), and
interval of time over which pollution loss is evaluated.
To calculate the mass of pollutant left in the sewage sludge-amended soil at the end of
an individual's lifetime, Eq. 4-46 is used:
where:
35
= 35 mg/ha
= 11 mg/ha
TPN (total mass of dioxins available at a site after the final year of
application) calculated above,
4-76
-------
2.3xlO'2 = Km (total loss rate for dioxins) calculated above,
70 =' LS (average human lifetime), assumed to be 70 yr and
20 = ATV (site life) from Exhibit 4-3.
Then, f^ can be calculated using Eq. 4-47:
- 20 yr •
Mg) ha
20 yr • 3.11x10-*' 28-' • 7 Mg • 1000 -**'
kg ha-yr Mg
= 0.75
where:
20 = ^(number of years sewage sludge is applied to land) from Exhibit
4-3,
S.llxlQ-4 = q , (concentration of dioxins in sewage sludge) from Exhibit 4-1
7 - AR (application rate) from Exhibit 4-3,
1000 = constant to convert (Mg) to (kg), and ' •' • -
11 = Mu (mass of dioxms in soU at end of period equal to an
individual's lifetime) calculated above.
Pollutant Concentration on Eroded Soil
site (C^) 'reoukes ^^"T"**1 °f P°Uutant on soil erodinS from the land application
Sfh» ? / q * ? additional parameters to be calculated. First, the fraction of total
pollutant loss caused by erosion is calculated by Eq. 4-27: "*-u«n or total
where:
f _ yr n^
fero = - T^ - = 0.17
2.3x10-* r~l
5* (1°SS coefficient for erosion) calculated above, and
(total loss rate for dioxins) calculated above.
4-49
(- Second, the calculated rate of soil loss for a land application site is calculated using Eq.
4-77
-------
= 6x10 ^ m/yr • 1400 kg/m3 - 10,000 m2/ha
- 8400 kg/ha-yr
where:
6x10^
1400
10,000
de (average rate of soil loss due to erosion from sewage sludge-
amended land each year) from Exhibit 4-16,
BDmu (bulk density of sewage sludge-amended soil) from Exhibit
4-3, and
constant to convert (ha) to (m2).
Then, Csitetj can be calculated using, Eq. 4-48:
20 yr • S.llxlO-4 ^ - 7 -^- - 1000 -ZL - 0.17 • 0.75
r = *g ha-yr Mg
siu,dioxaa ~~r ——
8400 —%— - 70 yr
ha-yr
= 9.5x10-* 2£
kg
where:
20
3.11x10^
7
1000
0.17
0.75
8400
70
N (number of years sewage sludge is applied to land) from Exhibit
4-3, .
Cj (concentration of dioxins in sewage sludge) from Exhibit 4-1,
AR (application rate) from Exhibit 4-3,
constant to convert (Mg) to (kg),
fm (fraction of total pollutant loss caused by erosion) calculated
above,
fus (fraction of total cumulative loading lost hi individual's lifetime
to all four loss processes) calculated above,
MEsi!e (rate of soil loss) calculated above, and
LS (lifetime of an individual), assumed to be 70 years.
A dilution factor, to represent the extent to which eroded soil from the land application
site is diluted" by soil from the untreated remainder of the watershed, is calculated using Eq
4-51. First, the -sediment delivery ratios for the land application site and watershed are
calculated using Eqs. 4-52 and 4-53: .
4-78
-------
= 0.872 • (1074 to)"0-125 = 0.36
= 0.872 • (440,300 fez)"0'125 = 0.17
where:
1074
440,300
Asite (area of land application site treated with sewage sludge) from
Exhibit 4-16, and
(area of the watershed) from Exhibit 4-16.
Then the dilution factor is calculated using Eq. 4-51: .
D = _ 1074 ha -0.36 _
f (1074 ha • 0.36) + [(440,300 ha - 1074 ha) • 0.17]
= 5.2xlO-3 . .
where:
°-36
°-17
$te (sediment delivery ratio for land treated with sludge) calculated above
and '
Swt (sediment delivery ratio for the watershed) calculated above.
Eq 4-5?e
C°nCentration of P°llutant m erod«i soil (Cao^ can be calculated using
= 5.2x10-* • 9.5X10-62$.
where:
5.2x10^
9.5x10
*
kg
^(dilution factor) calculated above, and
c^j (concentration of dioxins in sewage sludge-amended soil
eroded from the land application site) calculated above.
t0 CalCUlate ** concentratio° of Pollutant dissolved in the
4-79
-------
4.9x70 ~8
16
10 kgfmg
6.5XJ0'13 mgIL
where:
4.9xlO'8
16
13,000
j (dry weight concentration of dioxins in eroded soil)
calculated above,
= TSS (concentration of total suspended solids in the surface water)
from Exhibit 4-16,
= Kd (soil-water partition coefficient for dioxins in the stream) from
, ' Exhibit 4-17, and
= constant to convert (mg) to (kg).
Exposure Calculations
Potential human exposure to dioxins through direct ingestion of surface water is
calculated using Eq. 4-59:
_ 6.5x20 713 mg/L • 2 L/day
= -
= l.SxlO -14 mg/kg-day
where:
6.5xlO'13
2
70
Cswj (concentration of dioxins in surface water) calculated above,
IW (ingestion rate of water) from Exhibit 4-16, and
BW (body weight), assumed to be 70 kg.
Potential human exposure to dioxins through consumption of contaminated fish is
calculated using Eqs. 4-60 through 4-63. First, the BCF for dioxins is calculated using Eq. 4-
60:
log.JBCF) = 0.79 - 6.64 - 0.80 = 4.4
where:
6.64 = log10(/$roHI) from Exhibit 4-17.
4-80
-------
The BAF is calculated using Eq. 4-61:
BAF = 28,000 • 10 = 2.8xl05
where:
28,000 = BCF (bioconcentration factor for dioxins) calculated above by
exponentiating log10(5CF), and
10 = FM (food chain multiplier for dioxins) from Exhibit 4-17.
/ .
The concentration of dioxins in fish fillets is then calculated from Eq.. 4-62:
CffJ = 6.5x10 ,'13 mg/L • 2.Sxl05 L/kg • 0.5
= 9.0x70 ~8 mg/kg
where:
6.5xlO"13 = C^j (concentration of dioxins in surface water) calculated above,
2.8X105 = BAF (bioaccumulation factor for dioxins) calculated above, and -'
°-5 = pf (ratio of pollutant concentration in fillet to whole fish) from
Data Inputs.
Human exposure through ingestion of fish fillets is then calculated using Eq. 4-63:
EXp = 9.0x20 -* mg/kg •• 0.04 kg/day
. /•*—'" : 70kg
= 5.2x20 -11 mg/kg-day
where: .
- 9.0xlO'8 = CffJ (concentration of dioxins in fish fillets) calculated above,
°-04 = IF (daily ingestion of fish) from Exhibit 4-16, and
70 = BW (body weight), assumed to be 70 kg.
Total exposure to dioxins in surface water is the sum of the exposures to dioxins in water
and fish, as shown in Eq. 4-64:
4-81
-------
= '1.8x70-14 mg/kg-day + 52x20-nmg/kg-day
= 5.2x10'" mg/kg-day
where:
l.Sxlfr14 = EXPw,j (exposure to dioxins through ingestion of surface water),
calculated above, and '
5.2x10-" = EXPf.j (exposure to dioxins through ingestion of fish) calculated
above.
4.2.13 Pathway 13 - Inhalation of Pollutants Volatilized from Land-Applied Sewage
Sludge
Pathway 13 evaluates human exposure to pollutants volatilizing from both agricultural
and non-agricultural lands to which sewage sludge has been applied. Non-agricultural lands
include forests, reclamation sites, and public contact sites.
ir
To estimate exposure for this pathway, a mass balance analysis is required This mass
balance analysis accounts for the partitioning of pollutants into different soil phases (solids air-
and water) and the subsequent losses of pollutants from the land application site. Pollutants are
lost from a land application site by: erosion of contaminated soil particles, which releases
pollutants into surface waters; volatilization of pollutants into air; leaching of pollutants into
groundwater; and degradation. A mass balance for a pollutant must be maintained, given these
four competing loss processes of erosion, volatilization, leaching, and degradation Once mass
balances for pollutants have been established, exposures to pollutants that have eroded
volatilized, or leached are calculated under three separate pathways: surface water (Pathway 12)'
air (Pathway 13), and groundwater (Pathway 14). Pollutants which have degraded are assumed
to have degraded into chemicals that do not pose unacceptable risks to public health or the
environment.
The methods for performing the mass balance calculation for Pathways 12 13 and 14
are discussed in Section 4.2.12. In this section, the equations particular to Pathway 13 for
estimating the pollutant mass expected to volatilize and its transport to the downwind edge of
the land application site, are presented. Note that volatilization is assumed to occur within a
one-year period; any contribution to volatilization from sewage sludge applied in prior years is
considered negligible. J
4-82
-------
Methods Specific to Pathway 13
There are two major steps required to estimate the concentration of a volatilized pollutant
in air at the downwind edge of the land application site:
1) Using the mass balance calculations presented hi Section 4.2.12, the mass
of pollutant expected to volatilize from the land application site within a
period equivalent to a human lifespan is estimated.
2) Using a simplified version of the Industrial Source Complex Long Term
Model (ISCLT), the transport and dispersion of pollutant in ambient an-
al the downwind edge of the land application site are modeled.
In the first step, the rate at which a pollutant volatilizes from the site is estimated, based
on the assumption that equilibrium has been achieved between annual pollutant loadings and total
losses: -
^j--AR-Cj-f^ . (4-65)
where:
FA^j = annual average flux of pollutant j volatilizing from the-sewage sludge-
amended soil (kg pollutant/ha-yr),
0.001 = constant to convert units from (Mg-mg/kg) to (kg),
AR = annual whole sludge application rate of sewage sludge to land (dry Mg
sewage sludge/ha-yr),
Cj = concentration of pollutant j in sewage sludge (mg pollutant/kg sewage
sludge), and
. /«,„ - . fraction of total pollutant loss caused by volatilization (dimensionless).
The fraction of total pollutant loss caused by volatilization is obtained from the mass balance
calculation presented hi Section 4.2.12.
In the second step, pollutant concentrations in ambient air at the downwind edge of the
land application site are estimated, based on pollutant fluxes from the site. The model used to
simulate^transport of pollutants from treated land is described by U.S. EPA (1986d) and is based
on equations provided by Environmental.Science and Engineering (1985). These equations are
simplifications of equations used hi ISCLT.
• The exposed individual is assumed to live within 1 km of the land application site and
to be exposed to concentrations present at the downwind edge of the land application site A
source-receptor .ratio is calculated to relate the concentration of pollutant in ambient air at'that
mdividual s location (g/m3) to.the rate at which that pollutant is emitted from the treated soil
(e/m-sec):
(g/m2-sec):
4-83
-------
- 2fW9 A* ' V ' 10>QO°
- 2.032 — - - . — (4-66)
(r + *P ' u • az
where:
SRR = source-receptor ratio (sec/m),
2.032 = empirical constant, ' .
Asue — ' area of land application site (ha),
v = vertical term for dispersion of pollutant in air (dimensioriless),
10,000 = constant to convert (ha) to (m2),
r> = distance from center of the land application site to the downwind
edge(m),
xy ~ ^teral virtual distance to land application site (m),
u . = average wind speed (m/sec), and
°i = standard deviation of the vertical distribution of pollutant
concentration hi air (m).
rr ^ vertical term (v) is a function of source height, the mixing layer height and a
Under stable conditions the mixing layer height is assumed infinite, and for a pollutant releas'e"
height of zero, v=l. The lateral virtual distance is the distance from a virtual point source to
the land application site, such that the angle 6 subtended by the site's width is 22 5° This
distance is calculated as: ' '
_ A^-10,000 e
x - \ cot — (4-67)
N •' it 2
The distance from the center of the land application site to the downwind edge is
calculated assuming a square land application site:
• 10.000
at™ J?' T^ f ,°f **•VertiCal distribution of concentration (az) is defined by an
atmosphenc stability class and the distance from the center of the site to the downwind edge
Exhibit 4-18 provides values for two parameters, a and b, for a range of distances under stable
4-84
-------
atmospheric conditions. Based on values from this table, an appropriate value of a is calculated
.as: z
az = a or*
(4-69)
where r
x =.
(4-70)
and:
x - distance from the center of the land application site to the downwind edge
(km), and
10'3 = constant to convert (m) to (km).
EXHIBIT 4-18
Parameters Used to Calculate
-------
Once the source-receptor ratio has been estimated, it is combined with the estimated
average flux of pollutant to predict the average concentration of pollutant in ambient air at the
downwind edge of the site:
airj
SRR • 0.00317
(4-71)
where:
0.00317
average concentration of pollutant y in ambient air at the downwind
edge of the site (pg pollutant/m3 air), and
constant to convert (kg/ha-yr) to (/*g/m2-sec).
Estimating Human Exposure
The estimated concentrations of pollutants in air are converted to estimates of human
exposure based on assumptions about the rate at which the average individual inhales air:
EXP. =
3
10"3 • C
BW
(4-72)
where:
EXPj -
lO"3 -
IA'* -
BW -
exposure to pollutant./ in sewage sludge (mg pollutant/kg body weight-
day),
constant to convert (/*g) to (mg),
average concentration of pollutant j in ambient air (jig pollutant/m3 air),
inhalation rate (m3 air/day), and
body weight (kg).
Data Inputs
All the non-pollutant-specific data inputs required for this pathway are presented in
Exhibit 4-16 and all the pollutant-specific parameters are presented in Exhibit 4-17, both hi
Section 4.2.12. As shown in Exhibit 4-16, the daily inhalation rate for humans is assumed to
be 20 m /day, the average wind velocity is assumed to be 4.5 m/sec, and the average air
temperature is assumed to be 288 K.
4-86
-------
Example Exposure Calculations for Pathway 13
The following example calculates exposure of humans to dioxins and dibenzofurans
through inhalation of dioxins volatilized from sewage sludge-amended soil on agricultural land.
The mass balance portion of the calculation, which is the same for Pathways 12, 13, and 14, is
presented in the section "Example Exposure Calculations for Pathway 12". In the mass balance
calculation, /w/ is estimated from Eq. 4-27: .
2.3*10-2/yr
where:
1.9 x 10"2
2.3 x 10'2
Kw[ (loss rate due to volatilization of dioxins from sewage sludge-
amended land) calculated in Pathway 12, and
Kut (total loss rate for dioxins from sewage sludge-amended land)
calculated hi Pathway 12.
Equation 4-65 is used to estimate the annual average flux of dioxins volatilizing from
sewage sludge-amended land: . -
= 0-001 • IMglha-yr -3.1 1x10 ^mg/kg • 0.82
where:
3.11x10^
0.82
constant to convert (Mg-mg/kg) to (kg),
AR (annual whole sludge application rate of sewage sludge to land)
from Exhibit 4-3,
Cj (concentration of dioxins in sewage sludge) from Exhibit 4-1
and
fwl (fraction of total pollutant loss caused by volatilization)
calculated above.
The source-receptor ratio is then calculated by Eq. 4-66. Three variables, x r' and a
^6e7stimated- ^ lateral virtual distance to the land application site, Xy, £ calculated
4-87
-------
' x 1074/m • lQ,OOOm2/ha m ^22.5
= 9295 m
where: •
1074 = Asite (area of land application site treated with sewage sludge) from
Exhibit 4-16,
10,000 = constant to convert (ha) to (m2), and
22.5 = 6 (the angle subtended by the site's width) from Exhibit 4-16.
The standard deviation of the vertical distribution of concentration (o^) is then calculated
using Eqs. 4-68 through 4-70. The distance from the center of the land application site to the
downwind edge is calculated using Eq. 4-68:
, , V1074/M • 10,000 =
r ,
where:
1074 - Asitf (area of land application site treated with sewage sludge) from Exhibit
4-16, and ' . <
10000 = constant to convert (ha) to (m2).
The distance from the center of the land application site to the downwind edge is then converted
to kilometers using Eq. 4-70:
x = 10'3 km/m • 1639m = 1.6 km
where:
10'3 = constant to convert (m) to (km), and
1639 = r' (distance from center of the land application site to the downwind edge)
. calculated above.
4-88
-------
Then the standard deviation of the vertical distribution is calculated using Eq. 4-69:
= 13.953 • 1.6a63227 = 19m
where:
13.953
0.63227
a (corresponding to x = 1.6 km) from Exhibit 4-18, and
b (corresponding to x = 1,6 km) from Exhibit 4-18.
The source-receptor ratio can then be calculated using Eq. 4-66:
where:
2.032
1074
1
10000
1639
9295
4.5
19
SRR = 2.032 •
= 23sec/m
1074te- 1 • 10.000
(1639m + 9295m) "• 4.5m/sec • 19m
empirical constant,
Asiu (area of land application site treated with sewage sludge) from Exhibit
4-16, - .
v (vertical dispersion term) from Exhibit 4-16,
constant to convert (ha) to (m2),
r' (distance from center of the land application site to the downwind edee)
calculated above, 6 '
Xy (lateral virtual distance to land application site) calculated above
u (wind speed) from Exhibit 4-16, and
a (standard deviation of the vertical distribution of concentration in air)
calculated above.
of
c
air,
-------
Potential human exposure to dioxins through inhalation of dioxins volatilizing from
sewage sludge-amended land is calculated from Eq. 4-72:
- 10"3 • 1.3x70-7u.g/m3 • 20m3/day
~ WJT-.
= 3.8xlO-nmg/kg-day
where: • ,
10"3 = constant to convert (/zg) to (kg),
l.SxlO"7 = C&j (average concentration of dioxins hi ambient air) calculated
above, *
20 = IA (daily inhalation volume) from Exhibit 4-16, and
70 = BW (body weight), assumed to be 70 kg.
4.2.14 Pathway 14 - Ingestion of Groundwater Containing Leached Pollutants
Pathway 14 evaluates human exposure to pollutants through ingestion of groundwater that
receives leachate from agricultural and non-agricultural lands to which sewage sludge has been
applied. Non-agricultural lands include forests, reclamation sites, and public contact sites.
To estimate exposure for this pathway, a mass balance analysis is required. This mass
balance analysis accounts for the partitioning of pollutants into different soil phases (solids air
and water) and the subsequent losses of pollutants from the land application site. Pollutants are
lost from a land application site by: erosion of contaminated soil particles, which releases
pollutants into surface waters; volatilization of pollutants into ah-; leaching of pollutants into
groundwater; and degradation. A mass balance for a pollutant must be maintained, given these
four competing loss processes of erosion, volatilization, leaching, and degradation Once mass
balances for pollutants have been established, exposures to pollutants that have eroded
volatilized, or leached are calculated under three separate pathways: surface water (Pathway 12)'
air (Pathway 13), and groundwater (Pathway 14). Pollutants which have degraded are assumed
to have degraded into chemicals that do n6t pose unacceptable risks .to public health or the
environment.
The methods for performing the mass balance calculation for Pathways 12 13 and 14
are discussed hi Section 4.2.12. In this section, the equations particular to Pathway 14 for
estimating the concentration of pollutant in leachate from a site and modeling the transport of
that pollutant to the groundwater, are presented.
*
4-90
-------
Methods Specific to Pathway 14
There are two steps required to estimate the concentration of each pollutant in
groundwater near the land application site:
1) Determine the concentration of pollutant in water leaching through the
treated soil.
2) Use mathematical models for the transport of pollutant through the
unsaturated and saturated soil zones to estimate expected concentrations
of pollutant in groundwater.
The maximum mass of pollutant available for leaching from a site is estimated first For
• all organic pollutants, except dioxins and dibenzofurans and coplanar PCBs, it is assumed that
pollutant concentrations gradually increase in the soil until the rates of annual loss equal the rates
of annual loading, and steady-state is achieved. At steady-state, the rate at which the organic
pollutant leaches from the site can be determined from the annual loading (which equals total
annual losses) and the fraction of total losses attributable to leaching:
FA*cj = AR • Cj 'fuc ' °-001 (4-73)
where: ,
FAleCf j = annual average flux of pollutant j leaching from sewage sludge-amended
soil (kg pollutant/ha-yr),
AR = annual whole sludge application rate of sewage sludge to land (dry Me
sewage sludge/ha-yr),
Cj = concentration of pollutant j in sewage sludge (mg pollutant/kg sewaee
, ' sludge), . * 5
•fbm = fraction of total pollutant loss caused by leaching (dimensionless), and
0.001 - constant for converting units from (Mg-mg/kg) to (kg).
For inorganic pollutants, dioxins and dibenzofurans, and coplanar PCBs sewage sludge
is assumed to be applied over a 20 year period, followed by an inactive period. During the
inactive period, pollutant is depleted from the treated soil by leaching and erosion, and for the
two classes of organic pollutants, volatilization and degradation, as well. To simulate potential
contamination of groundwater, the loading of pollutant into the unsaturated zone is "linearized"
into a pulse of constant magnitude to represent the maximum annual loss of pollutant (in ke/ha-
yr) occurring over the 300-year simulation period modeled. The duration of that pulse is
calculated so that pollutant mass is conserved. For land application sites, the maximum rate of
loss is expected in the year immediately following the last application of sewage sludge because
the concentration of pollutant at the site reaches its peak at that time. As explained in Appendix
C, this peak loss rate could be maintained for a maximum length of time described by
4-91
-------
N '
TP = :—=— (4-74)
'
where:
TP = duration of "square wave" for approximating the loading of pollutant into
the unsaturated soil zone (yr), and
N = total number of years sewage sludge is applied to land (yr).
This result is combined with an estimate of the fraction of total pollutant loss to leaching
for a conservative estimate of the average flux of pollutant leaching from the land application
site: *
(4-75)
'7*2)
The fraction of total pollutant loss caused by leaching is -obtained from the mass balance
calculation presented hi Section 4.2. 12.
For both organic and inorganic pollutants, the estimated flux from either Eq. 4-73 or Eq.
4-75 can 'be combined with the assumed rate of net recharge -to groundwater at the land
application site to derive an estimate of the average concentration of pollutant in the leachate:
(4-76)
NR
where:
Q«,y = average concentration of pollutant j in water leaching from the sewage
sludge-amended soil (mg pollutant/L water),
0.1 = constant to convert, units from (kg/ha-m) to (mg/L), and
NR = annual recharge to groundwater beneath the treated soil (m recharge/yr).
Next, the leachate concentration is used to estimate the concentration of pollutant in
drinking water wells near the site. Two mathematical models are combined to calculate an
expected ratio between these two concentrations. The Vadose Zone Flow and Transport finite
element module (VADOFT) from the RUSTIC model (U.S. EPA, 1989d,g) is used to estimate
flow and transport through the unsaturated zone, and the AT123D analytical model (Yeh, 1981)
is used to estimate pollutant transport through the saturated zone.
4-92
-------
VADOFT allows consideration of multiple soil layers, each with homogeneous soil
characteristics. Within the unsaturated zone, the attenuation of organic pollutants is predicted
based on longitudinal dispersion, an estimated retardation coefficient derived from an equilibrium
partition coefficient, and a first-order rate of pollutant degradation. The input requirements for
the unsaturated zone module include various site-specific and geologic parameters and the rate
of gfoundwater recharge in the area of the site. It is assumed that the flux of pollutant mass into
the unsaturated zone beneath a land application site can be represented by results from the mass
balance calculations described above. Results from analysis of the unsaturated zone give the
flow velocity and concentration profiles for each pollutant of interest. These velocities and
concentrations are evaluated at the water table, converted to a mass flux, and used as input to
the AT123D saturated zone module.
The flow system hi the vertical column is solved with VADOFT, that is based on an
overlapping representation of the unsaturated and saturated zones. The water flux into the
unsaturated zone is specified for the bottom of the zone of incorporation for sewage sludge In
addition, a constant pressure-head boundary condition is specified for the bottom of the
unsaturated zone beneath the land application site. This pressure-head is chosen to be consistent
with the expected pressure head at the bottom of the saturated zone. Transport in the
unsaturated zone is determined using the Darcy velocity and saturation profiles from the flow
simulation: From these, the transport velocity profile can be determined.
» - •
Although limited to one-dimensional flow and transport, the use of a rigorous finite--
element model in the unsaturated zone allows consideration of depth-variant physical and
chemical processes that would influence the mass flux entering the saturated zone Among the
more important of these processes are advection (that is a function of the Darcy velocity
saturation and porosity), mass dispersion, adsorption of the leachate onto the solid phase and
both chemical and biological degradation. . •'
To represent the variably saturated soil column beneath the application site the model
discretizes the column into a finite-element grid consisting of a series of one-dimensional
elements connected at nodal points. Elements can be assigned different properties for the
simulation of flow in a heterogenous system. The model generates the grid from user-defined
zones; the user defines the homogeneous properties of each zone, the zone thickness and the
number of elements per zone, and the code automatically divides each zone into a series of '
elements of equal length. The governing equation is approximated using the Galerkin finite
element method and then solved iteratively for the dependent variable (pressure-head) subject
to the chosen initial and boundary conditions. Solution of the series of nonlinear simultaneous
equations generated by the Galerkin scheme is accomplished by either Picard iteration a
Newton-Raphson algorithm or a modified Newton-Raphson algorithm. Once the finite^element
calculation converges, the model yields estimated values for all the variables at each of the
discrete nodal points. A detailed description of the solution scheme is found in U.S. EPA
\ • o/*
One-dimensional, advective-dispersive transport is estimated with VADOFT based on the
estimated mass flux of pollutant into the top of the soil column, and a zero concentration
vT^Sr f °n at ** b0tt°m °f ** saturated zone- ^ resuWng mass flux from the
VADOFT simulation is used as input for the AT123D model, that simulates pollutant transport
4-93
-------
through the saturated zone. It is represented as a- mass flux boundary condition applied over a
rectangular area representative of the land application site. The transient nature of the flux into
the saturated zone is represented by tune-dependent levels interpolated from the results generated
by the VADOFT simulation.
As in calculations for the unsaturated'zone, degradation of organic pollutants is assumed
to be first-order during transport through the aquifer. Speciation and complexation reactions are
ignored for metals, leading to the possible over- or underestimation of expected concentrations
of metals in groundwater at the location of a receptor well. Detailed descriptions of the
AT123D model are provided by U.S. EPA (1986d) and by Yeh (1981) and will not be repeated
here. In general, the model provides an analytical solution to the basic advective-dispersive
transport equation. One advantage of AT123D is its flexibility: the model allows the user up
to 450 options and is capable of simulating a wide variety of configurations of source release
and boundary conditions. For the current application, AT123D uses the source term from
VADOFT and other input parameters to predict concentrations of pollutant within 300 years in
a receptor well at the downgradient edge of the land application site.
Estimating Human Exposure
Once pollutant concentrations in groundwater are estimated, estimates of human exposure
are made based on assumptions about the rate that the average individual consumes drinking
water. Potential exposure through ingestion of contaminated groundwater is estimated as:
C • TW
EXP. = welj (4-77)
' BW
where:
EXPj = exposure, to pollutant j in sewage sludge (mg pollutant/ke body weight-
day),
CUt/ = concentration of pollutant j in well water (mg pollutant/L water),
TW = volume of water ingested daily (L water/day), and
BW = body weight (kg).
Data Inputs
All the data inputs required for this pathway are presented in Section 4.2.12.
Example Exposure Calculations for Pathway 14
The following example calculates exposure of humans to dioxins and dibenzofurans
through ingestion of groundwater that has received leachate from sewage sludge-amended soil
on agricultural land. The mass balance portion of the calculation, which is the same for
Pathways 12, 13, and 14, is presented in the section "Example Exposure Calculations for
Pathway 12". In the mass balance calculation, f^ is estimated from Eq. 4-27:
4-94
-------
_ 7R 3
~ 7.0x10
2.3xlQ-2/yr
where:
1.8 x 10^ . = Kkc (loss rate due to leaching of dioxins from sewage sludge-
amended land) calculated in Pathway 12, and
2.3 x 10' = Kut (total loss rate for dioxins from sewage sludge-amended land)
calculated in Pathway 12.
The duration of the square wave for approximating the loading of dioxins into the
unsaturated soil zone is calculated using Eq. 4-74:
£1 _ g
where:
20 • = •' N
-------
Eq. 4-76 is then used to calculate the average concentration of pollutant in the. leachate:
= 0.1 • 6AxlQ-9kglha-yr
0.5m/yr
= 1.3xlQ-9mg/L
where:
0.1
6.4xlO-9
0.5
constant to convert units from (kg/ha-m) to (mg/L),
FA^ j (annual flux of dioxins leaching from the site) calculated
above, and
NR (net recharge to groundwater in treated area) from Exhibit 4-
16.
The concentration of dioxins hi leachate is then converted into a well concentration
through the use of VADOFT and AT123D. The concentration of dioxins hi the well is estimated
to be 0 mg/L; i.e., dioxins are not transported through the unsaturated and saturated zones to
the well hi appreciable concentrations. Using Eq. 4-77, human exposure to dioxins hi
groundwater is thus estimated to be:
where:
0
2
70
Cweij (concentration of dioxins in well water) obtained through VADOFT
and AT123D modeling,
IW (quantity of water ingested daily) from Exhibit 4-16, and
BW (body weight), assumed to be 70 kg.
4.2.15 Pathway 15 - Infant Exposure to Pollutants Through Breastfeeding
Pathway 15 evaluates exposures of infants to pollutants hi breast milk. Only highly
lipophilic pollutants are evaluated because these are expected to . concentrate in milkfat.
Concentrations of pollutants hi breast milk result in part from the mother's exposure to pollutants
in sewage sludge through several human exposure pathways, including pathways 1, 2, 4, 5, 12,
13 and 14, as described in previous sections. In addition, the mother is assumed to be exposed
to background concentrations of pollutants from sources other than sewage sludge. This analysis
presents two exposure scenarios that differ hi exposure duration and/or percent of an infant's
lifetime over which dose is averaged.
4-96
-------
Methods
The method used to estimate infant exposure is taken from Estimating Exposures to
Dioxin-Like Compounds (U.S. EPA, 1994b), and is based on an approach developed by Smith
(1987). The method assumes that the concentration of pollutant in breast milk fat is the same
as the concentration in maternal fat. The following calculation is used:
+ mS!Udge.J>' hj '/I. /
0.693
(4-78)
where:
».y
^background, j
"^sludge, j
- concentration of pollutant/ in maternal milk (mg pollutant/kg milk
fat), •
maternal intake of pollutant / from sources other than sewage
sludge (mg pollutant/kg body weight-day),
= maternal intake of pollutant / from relevant sewage sludge
exposure pathways (mg pollutant/kg body weight-day),
= half-life of pollutant j in adults (days),
= proportion of ingested pollutant / that is stored in fat
(dimensionless), ' '.'" ~
= ln(0.5), to convert the half-life of pollutant / into a rate
(dimensionless), and
•^ • • ' ' = proportion of mother's weight that is fat (kg maternal fat/kg total
body weight).
This steady-state model assumes that the pollutant levels in maternal fat remain constant i e
changes m maternal levels of pollutant during breastfeeding do not occur.
The concentration of pollutant in mother's milk is then used in the following equation to
estimate the daily dose to the infant: .
0.693
AT
(4-79)
where:
IM
ED
infant's average daily exposure to pollutant/ (mg pollutant/kg body
weight-day),
proportion of fat hi breast milk (dimensionless),
proportion of ingested pollutant/ that is absorbed (dimensionless),
ingestion rate of breast milk (kg milk/day),
exposure duration (yr),
4-97
-------
= average body weight of the infant during the exposure period (kg),
and
AT = averaging time (yr).
Daily doses to the infant are determined for each of the two following scenarios, which
are distinguished by exposure duration and averaging time:
Scenario 1: An infant breastfeeds for one year and the daily dose is averaged over this
•one year exposure period. Infant body weight at one-half exposure
duration (six months) is used to represent the average weight-over the
exposure duration in this scenario.
Scenario 2: Under the second scenario, the infant breastfeeds for two years. The dose
is averaged over 70 years.
The data used in these scenarios are discussed hi the next section.
Data Inputs '
The exposure scenarios were constructed using a mix of central tendency and high-end
values. The central tendency values were used for the following parameters: background
maternal intake, proportion of pollutant stored in fat, fat content of breast milk, and absorption
rate for ingested pollutants. The remainder of the parameters were set to then: high-end values.
Sources for the input values are discussed below.
General Inputs. Exhibit 4-19 shows the general (i.e., not pollutant-specific) inputs used
to calculate exposure to infants through breastfeeding. The values for many of the input
parameters were taken from Smith (1987) and U.S. EPA (1994b). The fat content of milk
(0.04) and the ingestion rate of milk come from information reported in Smith (1987). A study
of British children found mean intakes hi the first 7 to 8 months that ranged from 677 to 922
ml per day (Whitehead and Paul, 1981 as cited in Smith, 1987). A value of 0.9 L is used in
this assessment; assuming the density of breast milk is approximately that of water, 0.9 L is
equivalent to 0.9 kg. Smith (1987) presents two studies that estimate the percent of maternal
body weight that is fat. One study estimated a fat content of about 33% (Timson and Coffman,
1984, as cited hi Smith, 1987), and a second study found a reduction from a mean of 28% to
26.3% during four months of lactation (Butte et al., 1984, as cited in Smith, 1987). Based on
these data, the current assessment uses 0.3.as the proportion of maternal weight that is fat.
•The exposure duration and infant body weight over the exposure period differ between
the two scenarios. This analysis uses exposure durations of one or two years. Because the body
weight of an infant differs for these two exposure durations, the current analysis uses different
infant body weights for each exposure duration. For a one-year exposure duration, an infant
body weight of 9.1 kg (an average for babies 6 to 11 months old) is used to represent an average
body weight during the first year. For the two year exposure duration, a body weight of 11.3
kg (the average for 1 year old babies) is used as the average body weight over the first two years
4-98
-------
of life. Both values are taken from the National Center for Health Statistics (1987) (cited in
U.S. EPA, 1994b).
The appropriate choice of averaging time for these less-than-lifetime exposures depends
on the health endpoint assessed. In this analysis, different averaging times are used to estimate
the daily dose. First, the averaging time is set equal to the exposure duration (Scenario 1) to
obtain an estimate of daily dose for the period during which exposure occurs. Such an estimate
may be appropriate to evaluate health effects (such as developmental effects) that can occur from
short-term exposures (U.S. EPA, 1994b). (At this time, corresponding health risk values are
not available. Therefore, this calculation is only carried to the point of estimating exposure )
Second, this analysis also uses an averaging time of 70 years (Scenario 2) to calculate a lifetime
average daily dose (LADD) to correspond with exposure duration assumed for cancer potency
estimates.
EXHIBIT 4-19
General Input Parameters to Estimate Exposure to Pollutants through Breastfeeding
Parameter values common to all three scenarios:
fz: proportion of mother's weight that is fat (dimensionless)
0.3
•f3: proportion of fat hi breast milk (dimensionless)
0.04
IM: ingestion rate of breast milk (kg/day)
==========^=^=^====^=^==—
Parameter values that differ among scenarios:
ED: exposure duration (yr)
scenario 1
scenario 2
0.9
1
2
'- infant's average body weight (kg) scenario 1
scenario 2
9.1
11.3
AT: averaging time (yr)
scenario 1
scenario 2
1
70
Pollutant-Specific Inputs. For this pathway, only pollutants with a log Kow value greater
than five were evaluated. Other pollutants were not considered sufficiently lipophilic to warrant
further analysis. Of the Round Two candidate pollutants, only three-dioxins and dibenzofurans
coplanar PCBs and bis (2-ethylhexyl) phthalate (BEHP)-have log K^ values of five or greater
(see Exhibit 4-17 in Section 4.2.12). After further research, infant exposure to BEHP through
breastfeeding was not evaluated because the biological half-life of BEHP in humans is less than
one day. Schmid and Schlatter (1985) estimated the urinary elimination half-life to be 12 hours
and concluded that accumulation of BEHP in the human body was unlikely Sjoberg et al
4-99
-------
(1985) determined that BEHP levels in the human body decline with a half-life of 10 hours.
Based on these data, infant exposure to BEHP through breastfeeding did not require further
consideration. Therefore, only dioxins and furans and coplanar PCBs are considered for this
pathway.
There are five pollutant-specific parameters used to estimate exposure through this
pathway: maternal background exposure from other sources; maternal exposure from the relevant
sewage sludge pathways; percentage of pollutant ingested by the mother that is stored in her fat;
the half-life of the pollutant in the adult human body; and the percentage of pollutant ingested
by the infant that is absorbed. The pollutant-specific inputs for these pollutants are described
below; Exhibit 4-17 summarizes the pollutant-specific input values and the references for these
values.
Maternal exposure from other sources. For dioxins, a background exposure of
119 pg/day is assumed. This value is derived from environmental concentration
data collected hi rural, pristine, and urban areas not thought to be affected by
local sources (U.S. EPA, 1994b). Dividing this value by the standard adult body
weight of 70 kg yields 1.7 pg/kg-day. A similar value for coplanar PCBs was
not found; therefore, only incremental risk from exposure above background can
be assessed for coplanar PCBs.
Maternal exposure from sewage sludge pathways. For the estimate of maternal
exposure from the relevant sewage sludge pathways, the results of the exposure
assessments for pathways 1, .2, 4, 5, 12, 13 and 14 (described hi previous
sections) were considered simultaneously as the source of maternal body burden.
Percentage of pollutant ingested by the mother stored in her fat. The proportion
of ingested dioxins stored hi fat (0.9) is taken from Smith (1987) (as cited in U.S.
EPA, 1994b). Coplanar PCBs are assumed to behave in a similar manner.
Distribution studies of PCBs demonstrate that the adipose/plasma partition ratio
for PCBs ranged between 185/1 and 210/1 depending on the PCB involved
(Brown and Lawton, 1984). Based on these data (and assuming some PQBs are
• stored in other body tissues besides adipose tissues and plasma), it is assumed that
90 percent of the ingested coplanar PCBs are stored in fat.
Biological half-lives. U.S. EPA (1994b) presents different half-lives for 2,3,7,8-
TCDD in humans, from 5.8 years to 7 years. For this analysis, a half-life'of 7
years is used as a conservative estimate.
For PCBs, Yakushiji et al. (1978) reported an approximate half-life of 8 months.
In an analysis of exposure to fish, contaminated by PCBs hi the Great Lakes,
Anderson and Amrhein (1993) assumed a half-life of one year, based on their
review of the literature, although it should be noted that longer half-lives were
estimated for particular .PCB mixtures (up to a suggested half-life of 10 years for
congener #153). For this assessment, a half-life of one year was used for
coplanar PCBs.
4-100
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Percentage of pollutant ingested by the infant that is absorbed. Ninety percent
of ingested dioxins were assumed to be. absorbed by the infant (Smith 1987 as
cited in U.S. EPA, 1994b). For PCBs, studies of monkeys (Allen et al 1974)
and ferrets (Bleavins et al., 1984) demonstrated .90% and 85.4% absorption
efficiencies, respectively. Based on the data for monkeys, 90% absorption
efficiency in humans was assumed for.coplanar PCBs.
EXHIBIT 4-20
Pollutant-Specific Input Parameters to Estimate Exposure to Pollutants through
Breastfeeding
Pollutant
Maternal
Background
• Intake
(mg/kg-day)
Proportion
of
Pollutant
Stored in
Proportion
of Ingested
Pollutant
Absorbed
Half-Life of
Pollutant in
Adults (hj)
Dioxins and Dibenzofurans
1.7xlO-9(I>
0.9®
0.9®
7 years(1)
Polychlorinated biphenyls (coplanar)
'U.S. EPA (1994b).
2Smith (1987).
3Brown and Lawton (1984).
"Allen et al. (1974).
5Anderson and Amrhein (1993).
0.9(3>
0.9(4>
Example Exposure Calculations for Pathway 15
The following example calculates the exposure of infants to dioxins and dibenzofurans
in breast milk First, the maternal intake of dioxins in sewage sludge from agricultural pathways
i, 2, 4, 5, 12, 13, and 14 (msludge diolij!s) is calculated using the methods discussed hi the
corresponding sections. The value of msludgt_ ^ is 1.4 x 10'9 mg/kg-day. Equation 4-78 is then
used to calculate the concentration of dioxins hi milk fat:
4-101
-------
C -
mittfat, dioxins
(1.7xlQ-9mg/kg-day
2555days -0.9
0.693 • 0.3
= 3.5x20-5mg/kg
where:
1.4xlO-9
2555
0.9
0.693
0.3
, / (maternal intake of dioxins from sources other than
sewage sludge) from Exhibit 4-20,
msiadge.j (maternal intake of dioxins from sewage sludge exposure
pathways 1, 2, 4, 5, 12, 13, and 14),
hj (half-life of dioxins in adults) from Exhibit 4-20,
' /7i j (proportion of ingested dioxins that is stored in fat) from
Exhibit 4-20, .
ln(0.5), to convert the half-life of dioxins into a rate, and
/2 (proportion of mother's weight that is fat) from Exhibit 4-19.
Exposure to the infant under Scenario 2 can then be calculated using Eq. 4-79:
diorins
= 3.5xlQ-5mg/kg • 0.04 • 0.9 • 0.9kg/day • 2yr
11.3kg • 70yr
= 2.SxlO -9mg/kg -day
•where:
3.5xlO'5
0.04
0.9
0.9
2
11.3
70
(concentration of dioxins in maternal milk) calculated
above,
f3 (proportion of fat hi breast milk) from Exhibit 4-19,
ft.j (proportion of ingested dioxins that is absorbed) from Exhibit
4-20,
IM (ingestion rate of breast milk) from Exhibit 4-19,
ED (exposure duration) from Exhibit 4-19 for Scenario 2,
BWinfant (average body weight of the infant during the exposure
period) from Exhibit 4-19 for Scenario 2, and
AT (averaging time) from Exhibit 4-19 for Scenario 2.
4-102
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4.3 SURFACE DISPOSAL EXPOSURE METHODOLOGIES
Under the surface disposal practice for sewage sludge, humans can be exposed to
pollutants volatilizing or leaching from sewage sludge placed in either a monofill or a surface
impoundment. Humans can either inhale air or ingest groundwater containing pollutants. In
this section, the modeling prototypes for monbfills and surface impoundments are defined.
The methods used to model volatilization and leaching from the prototypes are then presented.
4.3.1 Definitions of a MonofiU and a Surface Impoundment
The monofill prototype is a sewage sludge-only trench fill. Only de-watered sewage
sludges with solids content greater than or equal to 20 percent are assumed to be suitable for
placement in a monofill, and the sewage sludge often is mixed with a bulking agent (e.e
soil) to increase solids content. Operating procedures assumed include daily" cover which
reduces odors and provides vector control, and a final cover placed on the monofill after
closure.
The surface impoundment prototype receives a continuous inflow of sewaee sludge
with a low solids content (between two and five percent). A vertical outflow pipe maintains
the surface liquid level at a constant height, and liquid is assumed to leave the impoundment
both in the outflow (possibly for return to the treatment works) and in seepage through the
floor of the impoundment. Over time, paniculate settling occurs and a denser layer of "solids
accumulates on the floor of the impoundment. Eventually, this layer of solids reaches the top
of the impoundment and no further inflow is possible. Upon closure, the sewage sludge is
left permanently in place and remains uncovered.
One key difference between the surface impoundment and monofill prototypes is that
the active surface impoundment is assumed to contain significantly more liquid than the active
monofill. Seepage through the floor of the surface impoundment is expected to be greater
than seepage from the monofill, and may be sufficient to sustain a local moundiiuTof the
underlying water table. The surface layer of the impoundment also is assumed to"be in a
liquid state over the active lifetime of the impoundment. The volatilization of organic
pollutants from this liquid layer is expected to differ from that predicted for a monofill which
is assumed to contain a higher percentage of solids and to receive a daily, and eventually a
permanent, cover, •
4.3.2 Methods for the Monofill Prototype
Method for Mass Balance
Pollutant mass is assumed to enter the monofill through daily deposits of sewage
sludge and to be removed through degradation, leaching, and volatilization Rates of
pollutant loss are assumed to be first-order (that is, proportional to the residual concentration
ot pollutant in the monofill); mass balance calculations begin by estimating first-order loss
coefficients for each competing loss process.
4-103
-------
Pollutant Losses Through Leaching. A coefficient for the rate of pollutant loss to
leaching is calculated by assuming that pollutant mass hi a filled monofill active sewage
sludge unit is partitioned at equilibrium between dissolved and adsorbed phases. Based on
mathematical relationships presented hi Section 4.2.14, the concentration of pollutant
dissolved in water within the monofill can be estimated from the total concentration of
pollutant within the monofill:
where:
Cw - concentration of dissolved pollutant hi water-filled pore space of
sewage sludge/soil hi monofill (kg pollutant/m3 porewater),
C, = total concentration of pollutant hi bulk sewage sludge/soil in monofill
(kg pollutant/m3 total bulk soil volume),
BDma = bulk density of sewage sludge/soil (kg soil/m3 total bulk soil volume),
Kd = soil-water partition coefficient (L water/kg. soil), .
#u- = water-filled porosity (dimensionless),
H = Henry's Law constant (dimensionless),
6* = air-filled porosity (dimensionless), and
10-3 = constant to convert (L) to (m3).
The dimensionless Henry's Law constant can be calculated by:
H = - - (4_81)
R • T • ICT3
where: .
H = Henry's Law constant (atm-m3/mol),
R = gas constant (L-atm/mol-K),
T = temperature (Kelvin), and
10'3 = constant to convert (L) to (m3).
For an arbitrary unit concentration of pollutant in the sewage sludge/soil
(1 kg pollutant/m3 sewage sludge/soil), a flux of pollutant mass leaching from the monofill
can be calculated as the product of net recharge (NR) and the expected concentration of
pollutant in leachate. Moreover, with a unit concentration of pollutant, the mass of pollutant
beneath one square meter of surface is equal to the volume of sewage sludge/soil beneath that
area. This volume can be expressed as kg pollutant/m2 area/m depth of the monofill. As
discussed hi Section 4.2.14, the estimated flux of leaching pollutant is divided by mis mass
to derive a first-order loss coefficient for leaching:,
4-104
-------
7VR
where:
KIK = loss rate due to leaching of the pollutant from monofill (yr1),
NR = annual recharge to groundwater beneath the monofill (m recharge/yr),
and
dmf = depth of a monofill active sewage sludge unit (m).
Pollutant Losses Through Volatilization. Rates of volatilization from a filled active
sewage sludge unit in a monofill will vary according to whether a cover layer of soil has been
applied. Each active sewage sludge unit in the monofill is assumed to contain uncovered
sewage sludge for a few hours on each of the days it receives sewage sludge. Following each
placement, a temporary cover layer of soil is applied. Once the monofill's capacity is
exhausted, a thicker permanent cover of soil is applied to the entire monofill (U.S. EPA,
1986d). A time-weighted average of emission rates with and without cover is used, therefore.
to describe the average rate of volatile emissions for an individual active sewage sludge unit
in the monofill. The fraction of the monofill's active lifetime that a typical active sewage
sludge unit will be uncovered is calculated as:
• : ' ' ' /. - 7=; . . ' (4-83)
where:
fm = fraction of monofill's active lifetime that a typical active sewage sludge
unit contains sewage sludge without soil cover (dimensionless),
tm = tune that a typical active sewage sludge unit contains uncovered sewage
sludge (yr), and "
LF - active lifetime of monofill (yr).
Some monofill active sewage sludge units will be filled early in the monofill's
operation, others nearer to the monofill's closure. The average monofill active sewage sludge
unit is assumed to contain sewage sludge for half the active lifetime of the monofill The
fraction of the monofill's active lifetime that such a unit will contain sewaee sludge that is
rr\\rf*rt*ri ic- •• ° °
covered is:
4-105
-------
fco 2 fun ^
where:
/„ = fraction of monofiU's active lifetime that typical active sewage sludge
unit contains sewage sludge with temporary soil cover (dimensionless"),
and
1A = fraction of monofill's active lifetime that typical active sewage sludge
unit contains sewage sludge (dimensionless).
A time-weighted average flux of pollutant emissions from a typical monofill active
sewage sludge unit is calculated from equations describing emissions from a unit with and
without soil cover. According to Environmental Science and Engineering (1985) as discussed
in U.S. EPA (1986d), emissions from an uncovered landfill cell can be described by:
0.17 u 0.994(r-293) Ca.
where: ~
-------
Equations 4-85 and 4-86 require an estimate of the concentration of pollutant in air-
filled pore space within the active sewage sludge unit. As discussed in Section 4.2.14, this
concentration can be related to the total concentration of pollutant hi sewage sludge/soil as:
10-3 8M (4-87)
6_
H H
where: . . .
Ca = concentration of gaseous pollutant hi air-filled pore space of monofill
(kg pollutant/m3 ah:),
C, = total concentration of pollutant hi bulk sewage sludge/soil in monofill
(kg pollutant/m3 total bulk soil volume),
KA .- ' soil-water partition coefficient (L water/kg soil),
B.Dmix -- bulk density of sewage sludge/soil (kg soil/m3 total bulk soil volume),
. H = Henry's Law constant (dimensionless),
#w — water-filled porosity (dunensionless),
0a = air-filled porosity (dimensionless), and
10'3 = constant to convert (L) to (m3).
Estimated pollutant fluxes from an uncovered and temporarily covered monofill active
sewage sludge unit are combined to derive a tune-weighted average pollutant flux from a
monofill unit during the monofilFs active lifetime:
' . £« = 4unfun + Icofco . (4'88)
where:
qac = time-weighted average pollutant flux from typical monofill unit over the
active lifetime of the monofill (kg pollutant/m2 unit-sec).
For a unit concentration (C,=l kg pollutant/m3 sewage sludge/soil) of pollutant in
sewage sludge/soil, the mass of pollutant beneath one square meter of monofill surface
.(kg/m2) is equal to the depth of the monofill (m). Therefore, converting the estimated loss
rate (kg/nr-sec) into a first-order loss coefficient (yr1) requires division by depth and
adjustment of units from seconds to years:
4-107
-------
qnr 3.16xl07
K = ?£. - (4-89)
"
where:
KM • - loss rate of pollutant due to volatilization during monofill's
active operation (yr1), and
3.16xl07 = constant to convert units from (sec"1) to (yr1).
Fraction of Pollutant Loss to Each Pathway. Estimated coefficients for losses to
volatilization and leaching are combined with assumed rates of degradation to yield a
"lumped" coefficient describing pollutant loss through all three pathways during the monofill's
active lifetime:
X =KUC + Zva+KtUg (4-90)
where: .
K,a = total loss rate of pollutant due to leaching, volatilization, and
degradation during monofiU's active operation (yr1).
The fraction of pollutant loss attributable to each individual process during the
monofill's active lifetime is:
4-= /«»= /*, = (4-91)
A«z *M ' Kv
where: .
/to = fraction of total pollutant loss during monofill's active operation
attributable to leaching (dimensionless),
fw = fraction of total pollutant loss during monofill's active operation
attributable to volatilization (dimensionless), and
ftu, = fraction of total pollutant loss during monofill's active operation
attributable to degradation (dimensionless).
The fraction of total loading lost within the monofill's active lifetime is calculated
numerically from the lumped rate of pollutant loss, assuming a time step of one year and a
unit pollutant loading of one kg/ha-yr:
4-108
-------
Mt = 0 (r=0)
(4-92)
The fraction of total pollutant lost during a monofill's active lifetime can then be calculated
as:
' --- (4-93)
.
ac 1 - LF
where: • " . • .
/„,. = fraction of total pollutant lost during monofill's active lifetime
(dimensionless),
Mff = mass of pollutant in sewage sludge/soil at end of monofill's active
lifetime (kg pollutant/ha), and
1 = annual unit loading of pollutant (kg/ha-yr).
Once the monofill's capacity is exhausted, a permanent cover layer of soil is applied
to its surface. This permanent cover reduces the rate of volatilization, changing both the total
rate of pollutant loss and the relative fraction of that loss attributable to volatilization,
leaching, or degradation. Based on the increased thickness of cover, an estimated rate of
volatilization from the -inactive monofill (£„-) is calculated with Eqs. 4-86 through 4-88 by
setting /„„ to zero. Rate coefficients for loss to leaching and degradation are assumed to be
unaffected by soil cover, so the lumped rate of loss for the inactive monofill is described by:
K« ' Kuc + *v, + K^ (4-94)
where:
Kti = total loss rate of pollutant from inactive monofill (yr1), and
Kvi = loss rate of pollutant due to volatilization from inactive monofill (yr1).
The fraction of loss attributable to volatilization is calculated as:
j£-
fvi = -/ (4-95)
&ri
where:
/" = fraction of total pollutant loss from inactive monofill attributable to
volatilization (dimensionless).
As will be discussed below, these fractions and the lumped rate coefficients for pollutant loss
are used to estimate pollutant concentrations in air and groundwater near the site.
4-109
-------
Method for Groundwater Pathway
Upon completion of the mass balance calculations described above, two additional
steps are used to calculate the concentration of each pollutant hi groundwater:
1) Determine the concentration of pollutant hi leachate from the bottom of the
monofill.
2) Use mathematical models for the transport of pollutant through the unsarurated
and saturated soil zones to estimate expected concentrations of pollutants in
groundwater. .
With the mass balance calculations, the total rate at which a pollutant is lost from the
monofill, and the fraction of that, loss attributable to leaching, are estimated. The amount of
time that would be required to deplete the entire mass of pollutant placed in a monofill at the
maximum predicted rate of loss for that pollutant is estimated. This approach is conservative
because using higher estimates for pollutant flux leaving the monofill will yield a higher
estimate of pollutant concentrations at the well.
For monofills, the rate of maximum total pollutant loss (hi kg/yr) will occur in the
year immediately following the last placement of sewage sludge, because the total mass of
pollutant at the site reaches its peak at that tune. As explained in Appendix C, this peak rate
of loss could be maintained for a maximum length of. tune described by :
where:
TP = duration of "square wave" for approximating the loading of pollutant
into me unsaturated soil zone (yr).
This result is combined with the estimate of the fraction of totaf pollutant loss through
leaching for a conservative estimate of the average flux of pollutant leaching from the
monofill:
(4-97)
where: , . .
j = annual average flux of pollutant j leaching from the monofill
(kg pollutant/ha-yr),
Cj = concentration of pollutant j in sewage sludge (mg pollutant/kg
sewage sludge),
sc = estimated mass of sewage sludge contained in one hectare of
completed monofill (kg/ha), and
4-110
-------
icr
constant for convening units from (mg/ha-yr) to (kg/ha-yr).
Because sewage sludge is often combined with soil to increase solids content when
placed in a monofill, the volume (and mass) of sewage sludge in the monofill is only a
fraction of the total volume of the monofill. Therefore, the dry mass of sewage sludge
contained in one hectare of filled monofill is calculated by multiplying the monofiH's depth
by the fraction of its volume containing sewage sludge and by the mass of solids in one cubic
meter of sewage.sludge:
where:
= dmffsl BDsludge
(4-98)
BD
and:
BD
sludge
L
Psl
PH-
Jsol
io*.
sludge
(4-99)
bulk density of sewage sludge (kg sewage sludge/m3 sewage
sludge),
fraction of monofill's volume containing sewage sludge
(dimensionless), "-''.'' ~
particle density of sewage sludge (kg sewage sludge/m3 sewaae
sludge),
density of water (kg water/m3 water),
fraction of solids in sewage sludge , (kg solids/kg sewaee
sludge), and
constant for converting units from (kg/m2) to (kg/ha).
Next, dividing this estimated flux by the assumed net recharge and adjusting units
yields the estimated average concentration of pollutant in leachate:
0.1 FA
'lecj
lecj
NR
(4-100)
where:
c
0.1
average concentration of pollutant/ in water leaching from the monofill
site (mg pollutant/L water), and
constant to convert from (kg/ha-m) to (mg/L).
The next step is to relate the leachate concentration to the expected concentration of
pollutant in drinking water wells near the site. Two mathematical models are combined to
calculate an expected ratio between these two concentrations. The Vadose Zone Flow and
Transport finite element module (VADOFT) from the RUSTIC model (U.S. EPA, 1989d,g)
is used to estimate flow and transport through the unsaturated zone, and the AT123D
4-111
-------
ana
zone.
tlytical model (Yeh, 1981) is used to estimate pollutant transport through the saturated
le.
VADOFT allows consideration of multiple soil layers, each with homogeneous soil
characteristics. Within the unsaturated zone, the attenuation of organic pollutants is predicted
based on longitudinal dispersion, an estimated retardation coefficient derived from an
equilibrium soil-water partition coefficient, and a first-order rate of pollutant degradation.
The input requirements for the unsaturated zone module include various site-specific and
geologic parameters and the leakage rate from the bottom of the monofill. It is assumed that
the flux of pollutant mass into the top of the unsaturated zone beneath a monofill can be
represented by results from the mass-balance calculations described above. Results from
analysis of the unsaturated zone give the flow velocity and concentration profiles for each
pollutant. These velocities and concentrations are evaluated at the water table, converted to
a mass flux, and used as input to the AT123D saturated zone module.
The flow system in the vertical column is solved with VADOFT, which is based on
an overlapping representation of the unsaturated and saturated zones. The water flux at the
soil/liquid interface is specified for the bottpm of the monofill, which defines the top of the
unsaturated zone in the model. In addition, a constant pressure-head boundary condition is
specified for the bottom of the unsaturated zone beneath the monofill. This pressure-head is
chosen to be consistent with the expected pressure head at the bottom of the saturated zone
without consideration of the added flux seeping from sewage sludge in the monofilf
Transport in the unsaturated zone is determined using the Darcy velocity and saturation
profiles from the flow simulation. From these, the transport velocity profile can be
determined.
Although limited to one-dimensional flow and transport, the use of a rigorous finite-
element model in the unsaturated zone allows consideration of depth-variant physical and
chemical processes that would influence the mass flux entering the saturated zone. Among
the more important of these processes are advection (which is a function of the Darcy
velocity, saturation, and porosity), mass dispersion, adsorption of the leachate onto the solids
phase, and both chemical and biological degradation.
To represent the variably saturated soil column beneath the floor of the monofill the
model discretizes the column into a finite-element grid consisting of a series . of one-
dimensional elements connected at nodal points. Elements can be assigned different
properties for the simulation of flow in a heterogenous system. The model generates the grid
from user-defined zones; the user defines the homogeneous properties of each zone, the zone
thickness, and the number of elements per zone, and the code automatically divides each zone
into a series of elements of equal length. The governing equation is approximated using the
Galerkm finite element method and then solved iteratively for the dependent variable
(pressure-head), subject to the chosen initial and boundary conditions. Solution of the series
of nonlinear simultaneous equations generated by the Galerkin scheme is accomplished by
either Picard iteration, a Newton-Raphson algorithm or a modified Newton-Raphson
algorithm. Once the finite-element calculation converges, the model yields estimated values
for all the variables at each of the discrete nodal points. A detailed description of the solution
scheme is found in U.S. EPA (1989g).
4-112
-------
One-dimensional, advective-dispersive transport is estimated with VADOFT based on
the estimated mass flux of pollutant into the top of the soil column, and a zero concentration
boundary condition at the bottom of the saturated zone. As discussed earlier, sewage sludge
is assumed to be deposited in the monofill for 20 years, followed by an inactive period in
which pollutant is depleted from the monofill by leaching, volatilization, and degradation.
To simulate potential contamination of groundwater, the loading of pollutant into the
unsaturated zone beneath the monofill is "linearized" into a pulse of constant magnitude (TP)
to represent the maximum annual loss of pollutant (hi kg/ha-yr) occurring over the 300-year
simulation period modeled. The duration of that pulse is calculated so .that pollutant mass is
conserved.
As in calculations for the unsaturated zone, degradation of organic pollutants is
assumed to be first-order during transport through the aquifer. Speciation and complexation
reactions are ignored for metals, leading to the possible over- or under-estimation of expected
concentrations of metals hi groundwater at the location of a receptor well. Detailed
descriptions of the AT123D model are provided by U.S. EPA (1986d) and by Yeh (1981) and
will not be repeated here. In general, the model provides an analytical solution to the basic
advective-dispersive transport equation. One advantage of AT123D is its flexibility: the
model allows the user up to 450 options and is capable of simulating a "wide variety of
configurations of source release and boundary conditions. For the current application,
AT123D uses the source term and other input parameters to predict concentrations of
pollutant (Cwel) within 300 years in a receptor well at the downgradient edge of the site~'s
property boundary.
Method for Volatilization Pathway
Two steps provide an estimate of the concentration of a volatilized pollutant in air near
the monofill: ••„.•'
1) Use the mass balance calculations summarized above to determine the mass of
pollutant expected to volatilize from the monofill within a period equivalent to '
a human lifespan, and .
2) Use a simplified version of the Industrial Source Complex Long Term Model
(ISCLT) to model the transport and dispersion of pollutant in ambient air near
the monofill.
Results from the mass balance calculations are used to estimate the fraction of total
pollutant mass expected to volatilize from the monofill within an expected human lifetime
(assumed to be 70 years), which spans both the active and inactive phases of the monofill's
operation:
4-113
-------
-"-^ (4-101)
where:
jta = fraction of pollutant mass that volatilizes over a human lifetime
(dimensionless), and
LS = average human lifetime (yr).
Next, this fraction is multiplied by the total mass of pollutant placed in the monofill,
and divided by the time of release to calculate "an average flux:
• . L*J
where: •
j = annual average flux of pollutant./ volatilizing from the monofill
(kg pollutant/ha-yr),
= constant to convert units from (mg/ha-yr) to (kg/ha-yr), and
Cj , = concentration of pollutant j in sewage sludge (mg pollutant/kg
sewage sludge). . . -
The next step is to relate releases of volatilized pollutant to the expected concentrations
in ambient air. The model used to simulate transport of pollutant from a monofill site is
described by U.S. EPA (1986d) and is based on equations provided by Environmental Science
and Engineering (1985). These equations are simplifications of equations used in ISCLT
The exposed individual is assumed to live at the downwind property boundary of the monofili
site. A source-receptor ratio is calculated to relate the. concentration of pollutant in ambient
air at that individual's location (g pollutant/m3 air) to the rate at which that pollutant is
emitted from the monofill (g pollutant/m2 monofill area-sec):
= 2.032 — (4-103)
(r'+xy)uoz
where: •
SRR = source-receptor ratio (sec/m),
2.032 = empirical constant,
Amono = area of monofill (m2),
V, = vfrtical term for dispersion of pollutant in air (dimensionless),
r' =B distance from the monofill's center to the downwind edge (m),
xy - lateral virtual distance to the monofill (m),
u - average wind speed (m/sec), and
or. = standard deviation of the vertical distribution of pollutant concentration
in air (m).
4-114
-------
The vertical term (v) is a function of source height, the mixing layer height, and oz.
Under stable conditions the mixing layer height is assumed to be infinite, and for a pollutant
release height of zero, v=l. The lateral virtual distance is the distance from a virtual point
source to the monofill, such that the angle 6 subtended by the monofill width is 22.5°. This
distance is calculated as:
^
Amono cot - (4-104)
The distance from the center of the land application site to the downwind edee is
calculated assuming a square monofill: -
(4-105)
The standard deviation of the vertical distribution of concentration (a.) is defined by
an atmospheric stability class and the distance from the center of the monofill to the
downwind edge. Exhibit 4-18 provides values for two parameters, a and fe, for a range of
distances under stable atmospheric conditions. Based on values from this table, an
appropriate value of a. is calculated as:
oz = a xb (4-106)
where:
* = 10-3 -r1 ' (4-107) .
and: •
x = distance from the center of the monofill to the downwind edge
(km), and
10"3 = constant to convert (m) to (km).
This result is combined with the estimated average flux of pollutant to predict the
average concentration of pollutant in ambient air over this period:
0.00317 , (4-108)
where: ' .
C<*r.j — average concentration of pollutant j hi ambient air at the
downwind edge of site (fig pollutant/m3 air), and
0.00317 = constant to convert (kg/ha-yr) to G*g/m2-sec).
4-115
-------
4.3.3 Methods for the Surface Impoundment Prototype
The methods for estimating exposure for surface impoundments are similar to those
described in Section 4.3.2 for monofills. As with monofills, a mass balance of pollutant
losses from the surface impoundment is calculated first.
Method for Mass Balance
Pollutants hi sewage sludge are assumed to enter the surface impoundment through
continuous inflow, and to be removed through four general processes:
1) degradation within the surface impoundment (e.g., photolysis,
hydrolysis, or microbial decay);
2) transportation out of the surface impoundment by seepage through the
floor of the impoundment;
3) outflow (possibly for return to the treatment works); and
4) volatilization from the liquid surface of the impoundment.
The model for describing these four processes in this analysis is adapted from a two-
layer model suggested by Thomann and Mueller (1987) for modeling toxic substances in a
lake. For the water column of a lake, those authors consider the inflow and outflow of
pollutant, diffusive exchange between the solids layer and the water column, degradation
volatilization, the settling of paniculate toxicant from the water column to the solid and the
re-suspension of particulate from the solids layer to the water column. For the solids layer
they consider diffusive exchange with the water column, decay processes, particulate settling
from the overlying water column, re-suspension from the solids to the water column, and loss
of toxicant from the solids due to net sedimentation or burial.
A similar, two-layer model is used for surface impoundments. The "liquid" layer
begins at the surface and has the same average solids content as inflow to the surface
impoundment; the "solid" layer beneath has a higher solids content. Although a gradient of
solids concentrations is likely to form in an actual impoundment, each layer is idealized as
homogeneous for both solids and pollutant concentrations.
Thomann and Mueller provide explicit equations for predicting settling velocities for
particulates and rates of diffusive exchange between the two layers, but the present
methodology derives simpler equations by assuming the solids layer will eventually reach the
surface of the impoundment and outflow contains negligible concentrations of suspended
solids. All loss processes are approximated as proportional to pollutant concentration; i.e.,
loss rates at any time are proportional to the current concentration of pollutant in the
impoundment. .
4-116
-------
Two additional simplifying assumptions are made:
1) Concentrations of pollutant within each layer are assumed to be at
steady-state and to be partitioned at equilibrium between adsorbed and
dissolved phases..
2) Rates of pollutant transfer and loss when the impoundment is half-filled
. with solids are assumed to be typical of the surface impoundment both
. before and after it fills with sewage sludge.
If rates of loss to effluent, volatilization, seepage, and degradation are all proportional
to pollutant concentration, the maximum total rate of loss will occur if equilibrium
concentrations are attained. Moreover, after the continuous placement of sewage sludge in
the surface impoundment is terminated, the rates at which pollutant is lost to seepage and
volatilization should decline. By assuming that equilibrium conditions represent the entire
(active and inactive) lifetime of the surface impoundment, this methodology, probably
overestimates rates of pollutant loss through seepage and volatilization, leading to
conservative estimates of risks.
Liquid Layer. The concentration of pollutant in the inflow of the impoundment (Q
and in the liquid layer (C,) are assumed to remain constant throughout the surface
impoundment's active lifetime. The partitioning of pollutant in the liquid layer is described
as:
+ DVC, (4-109)
where: . '. •
Qi =• rate at which sewage sludge enters the impoundment (m3 sewage
sludge/sec),
C, = concentration of pollutant in inflow to the impoundment
(kg pollutant/m3 sewage sludge),
Qo - rate at which outflow leaves the impoundment, possibly for further
treatment (m3 sewage sludge/sec),
fa = fraction of total pollutant in liquid layer that is dissolved
(dimensionless),
C, = total concentration of pollutant (adsorbed and dissolved) in liquid layer
of impoundment (kg pollutant/m3 liquid layer),
Kdegi = anaerobic rate of pollutant degradation in liquid layer (sec'1),
A = area of surface impoundment (m2),
di '- depth of liquid layer (m),
KmU = rate of pollutant volatilization from liquid layer (m/sec),
Qsep = rate of seepage beneath the impoundment (m/sec), and
DV = rate of change in the volume of the liquid layer (m3 liquid layer/sec).
4-117
-------
Because the total depth of the impoundment (including both liquid and solids layers)
is assumed constant, the depth of the liquid layer is reduced as more sewage sludge
accumulates in the solids layer. If the rate at which the solids accumulates is constant over
the active lifetime of the surface impoundment, the rate of accumulation can be determined
by dividing the total depth of the impoundment by its expected active lifetime: '.
DV = (4-110)
TF
where: ' .
dsi = total depth of surface impoundment (m), and
TF = estimated active lifetime of surface impoundment (sec).
The active lifetime of the surface impoundment is calculated as:
(4-111)
where: . -
Sj = concentration of solids in liquid layer (kg/m3), and
5, = concentration of solids in solids layer (kg/m3).
For the first term on the right of Eq. 4-109, the volume of outflow from the surface
impoundment (Q0) is calculated to be consistent with assumptions about rates of inflow,
seepage, and accumulation of the solids layer:
S*\ ( 5, \
1 I f\ A T*Tlt 1 2 I (4-H2)
P-l P*
The concentration of solids in the liquid and solids layers is calculated from parameters
describing the fraction of solids (by mass) in each layer:
(4-113)
• P.f,io-(i-p1)pj, P.P.ioa-jyp,
where:
PI = fraction of solids (by mass) in liquid layer (kg solids/kg liquid layer),
P2 = fraction of solids (by mass) in solids layer (kg solids/kg solids layer),
Pw = density of water (kg water/L water),
Pst = particle density of sewage sludge (kg sewage sludge/m3 sewage sludge),
'
4-118
-------
103 = constant to convert (L) to (m3).
In both the liquid and solids layers, a pollutant is partitioned between adsorbed and
dissolved phases. The partitioning depends on both the pollutant-specific partition coefficient
and the concentration of solids in the layer:
fdl "
'5,
where: '
•fdj = fraction of total pollutant hi liquid layer that is dissolved
(dimensionless), .
Kd = soil-water partition coefficient (L water/kg soil), and
10'3 = constant to convert (L) to (m3).
The second term on the right side of Eq. 4-109 describes degradation of the pollutant
through photolysis, hydrolysis, microbial decay, and other processes. Values for Kd , are
taken from studies of hydrolysis and microbial degradation, and are applied to pollutant in
both dissolved and adsorbed phases.
The third term. oh the right side of Eq. 4-109 describes pollutant loss through
volatilization, and is the only term directly linked with human exposure. The overall mass
transfer coefficient for volatilization (Kwa) is calculated with a two-film resistance model
(Tnomann and Mueller, 1987) in which the overall resistance equals the sum of the liquid and
gas phase resistances:
1 1 RT • 10'3
- j— = -? + —-77jr~ • (
^voii Ki "Kg
where:
Kwtl = rate of pollutant volatilization from liquid layer (m/sec),
KI . = • mass transfer coefficient for the liquid layer (m/sec),
Kg = mass transfer coefficient for the gas layer (m/sec),
R . = gas constant (L-atm/mol-K),
.. T = average air temperature (Kelvin),
H = - Henry's Law constant for pollutant (atm-m3/mol), and
10" = constant to convert (L) to (m3).
Numerous methods for calculating Kt and Kg for water surfaces have been proposed
(see for example: Hwang, 1985; Mackay and Leinonen, 1975; Mackay and Yeun 1983-
Shen, 1982; Springer etal., 1984; U.S. EPA, 1987b; U.S. EPA, 1989e). This methodology
follows an approach described in U.S. EPA (1987b, 1989e) for estimating volatilization from
4-119
-------
surface impoundments. The selection of appropriate equations for, calculating mass transfer
coefficients depends on two characteristics of the site: (1), the ratio of the impoundment's
effective diameter (or "fetch") to its depth and (2) the local average wind speed. Effective
diameter (hi meters) is defined as the diameter of a circle with area equal to that of the
impoundment. Depth is defined as that of the liquid layer, which for the purpose of this
calcuiation is assumed to average half of the impoundment's total depth. The ratio of fetch
to depth is therefore calculated as:
de =
& k (4-H6)
FR = —
where:
2 = factor to convert radius to diameter,
A — area of surface -impoundment (m2),
de = effective diameter (or fetch) of surface impoundment (m),
FR = ratio of fetch to depth (dimensionless), and
d, — depth of liquid layer (m).
For surface impoundments where the average wind speed 1Q m above the liquid surface is
greater than 3.25 m/s and FR ^ 51.2 (as in the scenario used for the surface impoundment
prototype):
K, = 2.61 IxlO'7.u{
2/3
(4-117) '
where:
2.6llxlO~7 = empirical constant, /
u,0 = ' average wind speed 10 m above surface (m/sec),
•Dot- = diffusivity of pollutant in water (cm2/sec), and
Dai, = diffusivity of diethyl ether hi water (8.5 x 10'6 cnr/sec).
Calculation of the mass transfer coefficient for the gas phase is based on Hwang
(1982). For all values of FD and uw, the mass transfer coefficient for the. gas layer is
calculated from:
Kg = 1.8xlO-Vo78 Sc™ de™ (4-118)
where: •
Kg = mass transfer coefficient for the gas layer (m/sec),
LSxlO'3 = empirical constant, and
&G — the Schmidt number on the gas side (dimensionless), defined
below.
4-120
-------
Scc = —j- , • (4-119)
where:
fia = viscosity of air (g/cm-sec),
Pa = density of ah- (g air/cm3 air), and
Dca = molecular diffusivity of pollutant in air (cm2/sec).
Equations 4-117 and 4-118 are sufficient to estimate Kwll, the overall mass transfer coefficient
for the dissolved fraction of the pollutant.
The fourth term on the right side of Eq. 4-109 describes losses of dissolved pollutant
from the liquid layer as a result of the seepage through the solids layer and the floor of the
impoundment. The rate of seepage (Qsep) is based on measured values from sewage sludse
lagoons. Only dissolved pollutant concentrations are included in this term; adsorbed pollutant
concentrations are included in the fifth term of the equation, which describes loss of pollutant
from the liquid layer as a result of the diminishing volume of that layer.
All terms on the right side of Eq. 4-109 are proportional to the concentration of
pollutant in the liquid layer. A coefficient for the total rate at which pollutant mass is lost
from the liquid layer (Ktotl, in m3/sec) can be defined as: -
Kua= QJdi + K^gidiA + KvoiifdiA + QSfpfdlA + DV (4-120)
so that:
Q;C; = K, ,C, (4-Pii
^' I tOll 1 VT J.4-J./
Because all estimated rates of pollutant loss are proportional to the concentration of pollutant
m toe liquid layer, total losses can be partitioned among competing loss processes according
to fixed ratios. Of the total mass of pollutant lost from the liquid layer, the fraction lost to
each process is:
/.
Jf •'VOll jp
(4-122)
where:
fraction of total pollutant lost from liquid layer that is lost in outflow
from the impoundment (dimensionless),
4-121
-------
fdegi — fraction of total pollutant lost from liquid layer that is lost to
degradation (dimensionless),
/«>/; = fraction of total pollutant lost from liquid layer that is lost to
volatilization (dimensionless),
fscpi = fraction of total pollutant lost from liquid layer that is lost to seepage
(dimensionless), and
fdeii = fraction of total pollutant lost from the liquid layer as a result of the
diminishing volume of the liquid layer (dimensionless).
Solids Layer. Pollutant mass accumulates in the solids layer as the depth of mis layer
increases and eventually reaches the surface of the impoundment. If the only source of
pollutant mass for the solids layer is the loss estimated for the liquid layer, then:
t = Kdeg2d2AC2 + Q^faAC2 + DVC2 (4-123)
where: -
faz = fraction of total pollutant in solids layer that is dissolved
(dimensionless),
d2 = depth of solids layer (m),
Kdeg2 = anaerobic rate of pollutant degradation in solids layer (sec^1), and
C, = . total concentration of pollutant hi solids layer (kg pollutant/m3 solids
layer).
Similar to the liquid layer, the partitioning of a pollutant in the solids layer can be
expressed as:
' <° ' i * r, .w* • s, (4-124>
where: . .
Kd = soil-water partition coefficient (L water/kg soil),
S2 = concentration of solids in solids layer (kg/m3), and
1CT3 = constant to convert (L) to (m3).
A coefficient for the total loss or storage of pollutant in the solids layer (Ktol2, in m3/sec) can
be defined as:
+ DV (4-125)
As with the liquid layer, this coefficient can be partitioned into its individual components:
4-122
-------
_ DV
./&£> -jp— (4-126)
wr2 »r2 ^Wj2
where:
/^2 = fraction of pollutant reaching the solids layer that is lost to degradation
(dimensionless),
fsep2 = fraction of pollutant reaching the solids layer that is lost to seepage
(dimensionless), and
fata - fraction of pollutant reaching the solids layer that is stored in the
accumulating depth of this layer (dimensionless).
If concentrations of pollutant in the liquid and solids layers can be approximated as
steady-state for the duration of the impoundment's active lifetime, and if the partitioning of
pollutant among competing loss processes halfway: through the impoundment's active lifetime
is assumed typical of its entire active phase, then the fraction of each year's loading of
pollutant lost during each year of the surface, impoundment's active phase can be calculated
as:
fact = /«,/; + f*gi + /«| + (4p/ •*/*!/) (krf +4,2) (4-127)
where: .
• faa = fraction of each year's loading of pollutant lost during each year of the
surface impoundment's active phase (dimensionless).
Finally, if all pollutant is eventually lost from the impoundment and the partitioning
of pollutant mass halfway through the surface impoundment's lifetime is generalized for the
entire mass of pollutant, the fraction of pollutant mass lost through each pathway can be
calculated as:
4-123
-------
f = **epl +^del1'
J SfD
'scp /.
Jo
act
Jvol
fdeg
f
fact
_ Joutl
«p, */«)/«
fact
(4-128)
out
Jact
where: •
fsep , = fraction of total pollutant lost from the impoundment through seepage
(dimensionless),
fwl = fraction of total pollutant lost from the impoundment through
volatilization (dimensionless), •
fdtg = fraction of total pollutant lost from the impoundment through
degradation (dimensionless), and
/„„ * - fraction of total pollutant lost from the impoundment through outflow
(dimensionless).
These results are used to calculate concentrations of pollutant in groundwater and air
near the surface impoundment.
Method for Groundwater Pathway
The methods for estimating concentrations of pollutants in groundwater near a surface
impoundment are almost identical to those discussed above for monofills. First the
concentration of pollutant in sewage sludge is used to estimate the expected flux of pollutant
info the top of the unsaturated zone. To simplify the calculations, this pollutant flux is
represented as a pulse of constant magnitude or "square wave, " with its duration calculated
so that the entire mass of pollutant will be depleted at the equilibrium rates calculated for the
active impoundment:
TP _ TF
= f^ • 31,536,000 , (4"129)
, •*
where:
- duration of "square wave" for approximating the loading of
pollutant into the unsaturated soil zone (yr),
TF = estimated active lifetime of surface impoundment (sec), and
31,536,000 = constant to convert (sec) to (yr).
4-124
-------
This result is combined with another result from the mass balance calculations to derive a
conservative estimate of the average flux of pollutant to the unsaturated zone beneath the site:
where:
PAiK.j = annual average flux of pollutant j leaching through the floor of
the surface impoundment (kg pollutant/ha-yr),
°-01 . = constant to convert (mg/m2) to (kg/ha), and
ci = concentration of pollutant/ in sewage sludge (mg pollutant/kg
sewage sludge).
Next the average flux is used to estimate the average concentration of pollutant in
seepage: '
Qsep ' 31,536,000
where: .
0.1 = constant to convert (kg/ha-m) to (mg/L), and
CsepJ = average concentration of pollutant j in water seeping through .the
bottom of the ^impoundment (mg pollutant/L water).
As discussed in Section 4.3.2 for the monofill prototype, two mathematical models are
combined for this purpose. The VADOFT component of the RUSTIC model (U S EPA
1989d,g) estimates flow and transport through the unsaturated zone, and the AT123D model
(Yeh, 1981) estimates pollutant transport through the saturated zone.
Minor adjustments have been made to the linked models to represent a phenomenon
unique to the surface impoundment prototype: seepage from a surface impoundment can cause
local elevation of the water table if rates of seepage from the impoundment exceed natural
rates of aquifer recharge in the surrounding area. Such elevation of the water table or
mounding, has two implications for the expected concentrations of sewage sludge pollutants
at a receptor well. The first is that the reduced vertical distance between the impoundment
and the local water table will result in decreased time of travel for water moving between the
impoundment and the saturated zone. The second is that an increased hydraulic gradient will
form in the aquifer between the surface impoundment and the downgradient receptor well
This change in the gradient will increase the expected rate of horizontal transport of the
pollutant through the saturated zone.
PTT*T J° ac?°,mmod^these ^o effects in the model calculations, an approach used in the
RUSTIC model is modified. The first component (VADOFT) of the modified linked model
performs calculations for a vertical column containing both unsaturated and saturated zones,
4-125
-------
and predicts the extent to which the elevation of the water table will be increased by the flux
of water from the impoundment. Once the vertical column problem has been solved for mass
and water fluxes at the water table elevation, the second model component (AT123D)
simulates the movement of pollutants through the saturated zone, with adjustments to
represent increased elevation of the water table. Unlike RUSTIC, however, the present
methodology does not allow for partial feedback between the unsaturated and saturated zone
components of the model; the saturated zone is represented separately by an analytical
transport model.
A
Saturated Zone. The AT123D model accepts as input the flux of pure pollutant mass
entering the top of the saturated zone, and does not consider the extent of the pollutant's
dilution by water from the source area, or the impact of that water on groundwater flow
within the saturated zone. When the vertical movement of pollutant through the unsaturated
zone is due only to infiltration throughout the area, the gradient within the aquifer is a
function of the water entering the saturated zone, and neglect of the diluted state of the source
term may be valid. For the.case of a surface impoundment, however, neglect of the extent
of the pollutant's original dilution could result hi non-trivial overestimation of the source
concentration, leading to an overestimation of pollutant concentrations at the receptor well.
Furthermore, neglect of mounding effects could lead to incorrect assumptions about the
velocity of groundwater flow near the site.
These concerns are addressed with three simple adjustments to the execution of the
AT123D model. First, to correct for AT123D's potential overestimation of the original
concentration of pollutant at the aquifer's boundary, the mass flux estimated from VADOFT
results is adjusted by a dilution factor (Df) as follows:
Df= AQ \p . • (4-132)
where:
. Fa = the volume of fluid passing through a vertical cross-section of the
aquifer oriented perpendicular to the direction of flow, and having a
width equal to the source width and a depth equal to the saturated
thickness of the aquifer (m3/sec).
The excess water released by seepage from a surface impoundment also can result in
a superimposed radial velocity field on the background or regional velocity field of
groundwater flow. In other words, the horizontal velocity of water within the aquifer can be
decreased upgradient of the surface impoundment, and increased downgradient of the surface
impoundment. This change hi the velocity field might result in reduced time of travel for
pollutants moving to receptor wells downgradient of the impoundment, which could in turn
lead to reductions hi pollutant degradation prior to human exposure. Accurate accounting of
the influence of mixing and degradation would require a fully three-dimensional flow and
transport model; this methodology uses a simpler approach to estimate a conservative limit
to pollutant decay within the system. The limit is estimated by increasing the estimated
velocity of groundwater flow to account for the maximum downgradient increase in velocity
4-126
-------
due to the source. The velocity increase can be approximated by idealizing the surface
impoundment as a circular source, so that the rite at which seepage passes outward through
a cylinder beneath the perimeter of the impoundment is:
(4-133)
••"a
where:
v,. = - superimposed radial velocity from water seeping from impoundment
(m/sec), and . .
da .= depth of aquifer (m).
In addition to increasing the expected velocity of pollutant transport through the
aquifer, this superimposed velocity also would have the effect of increasing AT123D's
estimate of pollutant dilution within the aquifer. This additional dilution effect must be
subtracted back out of the model calculations, because the true dilution is explicitly included
in the factor introduced by Eq. 4-132. The model performs this calculation automatically
based on the following equation for the anti-dilution factor-
(4-134)
vh
where: ,
Daf = anti-dilution factor (dimensionless),
v,. = vertical velocity due to the source (m/sec), and
v* = regional velocity of horizontal groundwater flow (m/sec).
the v^inrm, h. »K * JJ ' methodology is conservative, because it overestimates
the velocity beneath the source and does not allow for decreases in the superimposed velocity
beyond the source. As a result, the methodology is more conservative than a three-
dimensional model. In comparison with a two-dimensional cross-sectional flow and transport
model, the model is more conservative beneath the source, but less conservative beyond the
oUUlwC. - ^
combining the VADOFT model with AT123D, and by adjusting calculations in
mn, t , 9sfaaa^to ^ dil*ion and superimposed velocity-described above,
concentrations of a pollutant in groundwater at a receptor well can be predicted as a function
of the liquid concentration of pollutant near the floor of the impoundment, the rate of seepage
from the surface impoundment, and hydrogeological characteristics of the area. It should be
npted tfiat.all of the calculations described above are linear with respect to pollutant
concentrations in liquid seeping from the impoundment
4-127
-------
Method for Volatilization Pathway
Estimates of exposure for the vapor pathway are based on the highest average
concentrations ,of pollutant to be encountered over an expected human lifetime. At the
maximum rate at which pollutant is lost during the surface impoundment's active operation,
the fraction that would be lost to all processes over a period equivalent to the life expectancy
is:
fvls = fact ' fvoll • (4-135)
where: -
fiis - fraction of pollutant mass that volatilizes over a human lifetime
(dimensionless),
faa = fraction of each year's loading of pollutant lost during each year
of the surface impoundment's active phase (dimensionless), and
fwu = fraction of total pollutant lost from liquid layer that is lost to
volatilization (dimensionless).
This fraction can be converted to an average flux of pollutant volatilizing from the site
as: ..'••-
where:
FAwij = annual average flux of pollutant /( volatilizing from the surface
impoundment (kg/ha-yr), •
Cj - concentration of pollutant/ in sewage sludge (mg pollutant/kg sewage
sludge),
/* = fraction of pollutant mass that volatilizes over a human lifetime
(dimensionless),
0.01 =3 constant to convert (mg/m2) to (kg/ha),
$> = concentration of solids in solids layer (kg/m3),
dfi = total depth of impoundment (m), and
LS = average human lifetime (yr). . , > •
. The next step is to relate releases of volatilized pollutant from the site to the expected
concentration in ambient air. As .before, the simplified version of ISCLT described above
to calculate a source-receptor ratio (SRR) is used. Multiplying the SRR by the average
volatilized flux and adjusting units yields a conservative estimate of the expected average
concentration of pollutant hi ambient ah- near the site-
'
4-128
-------
C = —*fe/ _ (4-137)
mj 315.36
where: ' - . '
c<*.j = average concentration of pollutant j in ambient air at the
downwind edge of the site (jig pollutant/m3 air),
SRR = source-receptor ratio (sec/m), and
315.36 = constant to convert (kg/ha-yr) to (jig/nr-sec).
4.3.4 Estimating Human Exposure
To estimate human exposure, the methods discussed in Sections 4.3.2 and 4.3 3 are
used to estimate the concentrations of each pollutant in air and groundwater near each type
of surface disposal site prototype. Estimated concentrations in environmental media are
converted to estimates of human exposure based on assumptions about the rate at which the
average individual inhales air and ingests drinking water. For air, human exposure is
calculated as: •
icr3 • c . . • IA
EXP. = ^ (4-138)
' BW , '. ~
where:
EXPj = exposure to pollutant/ in sewage sludge (mg pollutant/kg body weight-
. day),
10'3 = constant to convert units from (fig) to (mg),
CairJ = average concentration of pollutant j in ambient air at the downwind
edge of the site (jtg pollutant/m3 air),
I A = inhalation rate (m3 air/day), and
BW = body weight (kg).
Potential exposure to pollutants through ingestion of groundwater is calculated as:
C TW
EXP. = . welj (4-139)
BW
where:
C ij = concentration-of pollutant j in well water (mg pollutant/L water), and
™ = volume of water ingested daily (L water/day).
4-129
-------
4.3.5 Data Inputs
Exhibits 4-21 and 4-22 present the non-pollutant-specific data for the monofill
prototype and the surface impoundment prototype.
EXHIBIT 4-21
Site and Sewage Sludge Parameters for Monofill Prototype'
j Parameter
| Area of Monofill (m2)
I Depth of Monofill (m)
|] Distance to Well (m)
1 Thickness of Daily Cover (m)
J Thickness of Permanent Cover (m)
II Time Each Unit Uncovered (hr)
[J Tune Average Unit Contains Sewage"
Sludge (hr) - -
II Sewage Sludge as Fraction of Total
Volume (m3/m3)
| Active Monofill Life (yr)
I! Average Wind Speed (m/sec)
| Average Air Temperature (Kelvin)
I! Vertical Term for Pollutant Dispersion in
Air (dimensionless)
[I Net Recharge (m/yr)
IJ Solids Content of Sewage Sludge (kg/kg)
Value
10,000
3.46
150
0.3
1
12
87,660
0.63
20
4.5
288
1'
0.5
0.20
Reference
U.S. EPA, 1992c
U.S. EPA, 1992c
U.S. EPA, 1992c
U.S. EPA, 1992c
U.S. EPA, 1992c
U.S. EPA, 1992c
U.S. EPA, 1992c
U.S. EPA, 1992c
U.S. EPA, 1992c
U.S. EPA, 1992c
U.S. EPA, 1992c.
U.S. EPA, 1992c
U.S. EPA, 1992c
U.S. EPA, 1992c
4-130
-------
EXHIBIT 4-22
Site and Sewage Sludge Parameters for Surface
Parameter
Area of Surface Impoundment (m2)
Value
—
20,236
Impoundment Prototype
Reference
=====
U.S. EPA, 1992c
Total Depth of Surface Impoundment (m)
U.S. EPA, 1992c
Distance to Well (m)
150
U.S. EPA, 1992c
Rate of Inflow (m3/sec)
0.0022
U.S. EPA, 1992c
I Average Wind Speed (m/sec)
4.5
U.S. EPA, 1992c
Average Air Temperature (Kelvin)
288
U.S. EPA, 1992c
Solids Content of Inflow (kg/kg)
0.03
U.S. EPA, 1992c
Solids Content of Liquid Layer (kg/kg)
0.03
U.S. EPA, 1992c
0.175
U.S. EPA, 1992c
1200
U.S. EPA, 1992e
I Net Seepage Beneath Impoundment (m/yr)
2.5
U.S. EPA, 1992c
Depth of Solids Layer (m)
Assumed half-filled
I Depth of Liquid Layer (m)
Vertical Term for Pollutant Dispersion in
Air (dimensionless)
Assumed half-filled
U.S. EPA, 1992c
| Density of Air at 15 °C (g/cm3)
1.226xlO-3
Weast, 1990
Viscosity of Air at 15 °C (g/cm-sec)
1.79x10^
nterpolated from Henry and
Heinke, 1989
There are several pollutant-specific fate and transport parameters required to maintain
the mass balance of a pollutant among the three loss processes and to estimate the rates at
which those three loss processes occur. In Exhibit 4-23the fate and transport parameters for
both the monofill and surface impoundment prototypes are presented. For degradation rates
estimates of both aerobic and anaerobic degradation rates are presented. As in Round One'
for the degradation rate of-a given pollutant in the unsaturated zone, ten percent of the
aerobic biodegradation rate is used if an hydrolysis rate was unavailable. For the degradation
rate for a given pollutant in the saturated zone, the arithmetic-average of the unsaturated zone
degradation rate and the anaerobic degradation rate for that pollutant is used.
To obtain estimates of inorganic Kd values for six Round Two pollutants, studies of
adsorption described in Gerritse et al. (1982) were used. Gerritse et al. present a range of
^.values for various inorganics in two soil types: sand and sandy loam. In the sandy soil
there was 0.035 g/g organic matter, 0 g/g clay, 0.22 meq/g cation exchange capacity (CEC)'
and the porewater had a PH of 5. In the sandy loam soil, there was 0.025 g/g organic
matter, 0.2 g/g clay, 0.16 meq/g CEC, and the porewater had a pH of 8. For this analysis
4-131
-------
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4-135
-------
\ .
, the Kd values from sand, which were lower than those hi sandy loam, were used. The
median Kd value from the range available for each of the six Round Two inorganics tested
was used.
For aluminum and fluoride, available data on Langmuir isotherm parameters were
used to estimate Kd values (Bodek et al., 1988). For aluminum, data were for silica, at a pH
of 5. For fluoride, data corresponded to clay loam, containing 10.4 percent clay, 0.94
percent organic carbon, and 825 /zg/g aluminum, with a pH of 5.9. For boron, thallium, tin,
and titanium, Kd values were not available. •
4.3.6 Modeling of Surface Impoundments in the Comprehensive Hazard Identification
For this risk assessment, only surface impoundments were modeled for the surface
disposal option. Exposure to pollutants in surface impoundments is greater than exposure to
pollutants in monofills for both the air and groundwater exposure pathways. Pollutants
volatilize more readily from surface impoundments than from monofills because there is no
soil cover. Pollutants also leach more readily from surface impoundments because the
recharge rate to groundwater is higher due to the liquid sewage sludge being placed in the
impoundment. Therefore, if a pollutant hi a surface impoundment did not result in high risk
levels, then its presence hi a monofill also would not be of concern.
4.3.7 Example Exposure Calculations for Surface Disposal
The following example calculates exposure of humans to methylene chloride present
in sewage sludge placed hi a surface impoundment. Exposure occurs through ingestion of
groundwater that has received leachate from a surface impoundment and through inhalation
of"methylene chloride that has volatilized from a surface impoundment.
Groundwater Exposure Pathway for Surface Impoundment
To estimate human exposure through the ground water pathway, computations must
be made for both the liquid layer of the surface impoundment and the solids layer.
Calculations for Liquid Layer of Surface Impoundment. Several steps are
necessary to estimate the rate at which methylene chloride is lost from the liquid layer of a
surface impoundment. • First, the mass transfer coefficients for the liquid layer of the surface
impoundment and the air layer immediately above it are calculated. Equation 4-117 is used
to calculate the mass transfer coefficient for the liquid layer:
€>
4-136
-------
2/3
5T, = 2.611x20-7 • (4.5m/sec)2 • I 9-2*™'6<™W
8.5x70-6cm2/secj
= 5.6xlO-6m/sec
where:
2.611xlO'7 = empirical constant,
4-5 = u!0 (average wind speed 10 m above surface) from Exhibit 4-22,
9.2xlO-6 = Dw (difftisivity of. methylene chloride in water) from
Exhibit 4-23, and '
S.SxlQ-6 = Dah (diffusivity of diethyl ether in water) from text.
To calculate the mass transfer coefficient for the gas layer immediately above the
surface impoundment, the Schmidt number and the effective diameter of the surface
impoundment must be calculated. The Schmidt number is calculated using Eq. 4-119:
1.79x10 -tg/cm -sec
ScG = —
1.23x20-3g/cm3 • 1.0xlO-\cm2/sec
. =1.5
where:
1.79XK)-4 = ^ (viscosity of air) from Exhibit 4-22,
1.23xlO'3 = Pa (density of air) from Exhibit 4-22, and
1. Oxl O'1 = Dca (diffusivity of methylene chloride in air) from Exhibit 4-23.
Equation 4-116 is used to calculate the effective diameter of the surface impoundment:
~ 20,236m2
= 2. — i - =
t „
160m
where:
^ = factor to convert radius to diameter, and
20,236 = A (area of surface impoundment) from Exhibit 4-22.
The mass transfer coefficient for the gas layer is then calculated using Eq. 4-118:
4-137
-------
K = 1.8x70 -3 • 4.5°'78 • 1.5-°'67 • 16Q-0-11
o -
2.6x70 -3--
sec
where:
1.8xlO~3
4-5
1-5
160
empirical constant,
uIO (average wind speed 10 m above surface) from Exhibit 4-22 ,
Scc (Schmidt number on the gas side) calculated above, and
^(efTectivediameterofsurfaceimpoundment)calculatedabove.
Both Kg and K, are then used in Equation 4-115 to calculate the overall mass transfer
coefficient for volatilization:
-5.6x.70-6ro/sec
1.8xl05secfm
0.082(L-atm)f(mol-K) • 288K • lQ-3m3fL
2.QxlO-3(m3-atm)l(mot) • 2.6x10-3m/sec
where:
5.6X10"6
0.082
288
2.0xlO'3
2.6xlO'3
KI (mass transfer coefficient for the liquid layer) calculated
above,
R (gas constant) from Exhibit 4-16,
T (average air temperature) from Exhibit 4-22,
H (Henry's Law constant for methylene chloride) from
Exhibit 4-23, and .
Kg (mass transfer coefficient for the gas layer) calculated above.
Taking the inverse of
= 5.4x70 -6m/sec
Second, the active lifetime of the surface impoundment and the rate of change in the
volume of the liquid layer are calculated. The concentrations of solids in the liquid and solids
layers must first be calculated by using Eq. 4-113:
4-138
-------
s = 1200kg/m3 - Ikg/L • 0.03 • 1000L/m3
• 0.03 • lOOOL/m3 + (1-0.03) • 1200kg/m
and: -
= 30kg/m
Ikg/L • 0.175 • 1000iy/n3 3
0.03 • lOOOL/m3 + (1-0.175) • 1200£g/m3
where:
1200 = p;, (particle density of sewage sludge) 'from Exhibit 4-22,
1 = pH. (density of water) from Exhibit 4-16,
0.03 = P, (fraction of solids (by mass) .in liquid layer) from Exhibit 4-22,
0. 175 = P2 (fraction of solids (by mass) in solids layer) from Exhibit 4-22 and
1000 = constant to convert (L) to (m3).
The active lifetime of the surface impoundment is then calculated using Eq. 4-111:
_ 4m • 20.236/B2 • 206kg/m3
0.0022w3/sec - 3Qkg/m3 •
= 2.5x20ssec
where:
^ , ~ dsi (total depth of surface impoundment) from Exhibit 4-22,
20,236 = A (area of surface impoundment) from Exhibit 4-22,
206 = ^2 (concentration of solids in solids layer) calculated above,
0.0022 = 0, (rate at which sewage sludge enters the surface
impoundment) from Exhibit 4-22, and
30 = si (concentration of solids in liquid layer) calculated above.
4-139
-------
The rate of change in the volume of the liquid layer is calculated using Eq. 4-110:
= 4m • 20,236m2
2.5xlO*sec
= 3.2x20 "
where:
4
20,236
2.5xl08
= d^ (total depth of surface impoundment) from Exhibit 4-22,
=. A (area of surface impoundment) from Exhibit 4-22, and
= TF (active lifetime of the surface impoundment) calculated
above.
Third, the volume of outflow from the surface impoundment is calculated using
Eq. 4-112: &
Q0 = 0.0022-
sec
1 -
30-^
m3
1200-&
m
- 2.5— -20,236m2 -
13L 32xlO-4^-
31,536,000sec ' sec
1 -
206-^
-3
1200-
-3
= 2.1x10-*—
sec
where:
0.0022
30
1200
2.5
3.2X10-4
206
Qi (rate at which sewage sludge enters the surface
impoundment) from Exhibit 4-22,
Sj (concentration of solids in liquid layer) calculated above,
ft a (particle density of sewage sludge) from Exhibit 4-22,
Qsep (rate of seepage beneath the impoundment) from
Exhibit 4-22,
£>V(rate of change in the volume of the liquid layer) calculated
above, and
S2 (concentration of solids in solids layer) calculated above.
4-140
-------
Fourth, Eq. 4-114 is used to calculate the fraction of methylene chloride dissolved in
the liquid layer:
fa =
1
1+0.19^- • 10-*— -30-^-
kg L _3
= 0.99
where:
0.19 =
10'3
30
Kd (soil-water partition coefficient for methylene chloride) from Exhibit
4-23,
constant to convert (L) to (m3), and
Sj (concentration of solids in liquid layer) calculated above.
Fifth, the total rate at which methylene chloride is lost from the liquid layer of a
surface impoundment can be calculated using Eq. 4-120:
Ktotl = (2.7x70 -4m3/sec • 0.99) + (12/yr
lyr
31,536,000sec
(5.4x70-6/n/sec • 0.99 • 20,236m2) +
lyr
2m • 20,236m2)
(2.5m/yr •
= 0.13m3/sec
31,536,000sec
0.99 • 20,236m2) + (3.
where:
2.7x10^
0,99
12
2
20.236
5.4X10-6
2.5
3.2x10^
Q0 (rate at which outflow leaves the impoundment) calculated
above,
fa (fraction of total methylene chloride in liquid layer that is
dissolved) calculated above,
Kdegl (anaerobic degradatipn rate for methylene chloride) from
Exhibit 4-23,
dj (depth of liquid layer) from Exhibit 4-22,
A (area of surface impoundment) from Exhibit 4-22,
K^j (rate of volatilization of methylene chloride from liquid
layer) calculated above,
Qsep (rate of seepage beneath the impoundment) from
Exhibit 4-22, and
DV(rate of change in the volume of the liquid layer) calculated
above. -
4-141
-------
The last calculations pertaining to the liquid layer estimate the fractions of total
methylene chloride lost from the liquid layer to seepage and as a result of the diminishing
volume of the liquid layer as the surface impoundment is filled. Eq. 4-122 is used to
calculate the fraction of total methylene chloride lost from the liquid layer to seepage:
. 0.13m3/sec
= L3xlO~2
where:
2-5 = Qsep (rate of seepage beneath the impoundment) from
Exhibit 4-22,
°-99 = fa (fraction of total methylene chloride in liquid layer that is
dissolved) calculated above,
20,236 • = A (area of surface impoundment) from Exhibit 4-22, and
°-13 = K,oa (coefficient for the total rate at which methylene chloride
is lost from the liquid layer) calculated above.
Eq. 4-122 is also used to calculate the fraction of total methylene chloride lost from the liquid
layer due to the diminishing volume of the liquid layer:
f =
Jdell ^~ I
0.13m3/sec
where: .
3.2X1Q-4 = DV (rate of change in the volume of the liquid layer) calculated
above, and
°-13 = KMI (coefficient for the total rate at which methylene chloride
is lost from the liquid layer) calculated above.
Calculations for Solids Layer of Surface Impoundment. As for the liquid layer of
the surface impoundment, several computational steps are required for the solids layer. First,
the fraction of total methylene chloride that is dissolved hi the solids layer is calculated using
Eq. 4-124:
4-142
-------
fdz = - = 0.96
1 + 0.19- • KT3— -206-^
kg L m*
where: .
0, 19 = Kd (soil-water panition coefficient for methylene chloride) from Exhibit
Q-2.3 ,
10'3 = constant to convert (L) to (m3), and
206 = 5, (concentration of solids in solids layer) calculated above.
Second, the coefficient for the total loss or storage of methylene chloride in the solids
layer is calculated using Eq. 4-125,:
31.536.
31.536.msec
where:
12 = K*g2 (anaerobic degradation rate for methylene chloride) from
Exhibit 4-23,
2 ~ d2 (depth of solids layer) from Exhibit 4-22,
20 236 = A (area of surface impoundment) from Exhibit 4-22,
Q«P (rate of seepage beneath the impoundment) from
Exhibit 4-22,
°-96 fa (fraction of total methylene chloride in solids layer that is
dissolved) calculated above, and
3.2X10-4 = D V (rate of change in the volume of the liquid layer) calculated
above.
4-143
-------
Third, the fraction of methylene chloride reaching the solids layer that is lost to
seepage is calculated using Eq. -4-126:
2.5m/yr - —IE-- • 0.96 • 20,236m*
Jsep2 •— -
= 8.9x10 ~2
where: .
2-5 = Qup (rate of seepage beneath the impoundment) from
Exhibit 4-22,
°-96 = • /^'(fraction of total methylene chloride in solids layer that is
dissolved) calculated above,
20.236 = A (area of surface impoundment) from Exhibit 4-22, and
1.7xlO'2 = Ktot2 (coefficient for the total rate at which methylene chloride
is lost from or stored in the solids layer) calculated above.
Fourth, the fraction of each year's loading of methylene chloride lost during each year
of the surface impoundment's active phase must be calculated. This calculation requires
seven fractions, as shown in Eq. 4-127. From Eq. 4-122, the fraction of total methylene
chloride lost from the liquid layer to volatilization is calculated:
, _ 5.4x10 "6m/sec • 0.99 • 20,236m2
' Jvoll
0.13m 3/sec
= 0.86
where:
5.4x10"6 = Krn/f (rate of volatilization of methylene chloride from liquid
layer) calculated above,
°-99 = fdi (fraction of methylene chloride -dissolved in liquid layer)
calculated above,
20,236 = A (area of surface impoundment) from Exhibit 4-22. and
°-13 = K-MI (coefficient for the total rate at which methylene chloride
is lost from the liquid layer) calculated above.
Also from Eq. 4-122, the fraction of total methylene chloride lost from the liquid layer to
degradation is calculated:
4-144
-------
31,S36,OOOsec
• 2m -20,236m2
0.13—
= 0.12
where: •
12 = ^dqti (anaerobic degradation rate for methylene chloride) from
Exhibit 4-23, .
2 = d, (depth of liquid layer) from Exhibit 4-22,
20.236 . - = A (area of surface impoundment) from Exhibit 4-22.
°'13 = 5"" (coefficient for the total rate at which methylene chloride
is lost from the liquid layer) calculated above.
Also from Eq. 4-122, the fraction of total methylene chloride lost from the liquid layer to
outflow is calculated:
f = 2.7x10 ^m^/sec • 0.99
Joutl ~ ~
0.13m3/sec
where:
2.7X10-4 = Qa (rate at which outflow leaves the impoundment) calculated
above,
°'99 = fa (fraction of methylene chloride dissolved in liquid layer)
calculated above, and , -
°'13 = K>»» (coefficient for the total rate at which methylene chloride
is lost from the liquid layer) calculated above.
From Eq. 4-126, the fraction of methylene chloride reaching the solids layer that is lost to
degradation is calculated:
f urannn-. '20,236m2
Jdeg2
= 0.89
where:
12 = ^W' (anaerobic degradation rate for methylene chloride) from
Exhibit 4-23,
4-145
-------
2 = J, (depth of solids layer) from Exhibit 4-22,
20,236 = A (area of surface impoundment) from Exhibit 4-22. and
1.7xlO~2 = Klol2 (coefficient for the total rate at which methylene chloride
is lost from or stored in the solids layer) calculated above.
The fraction of each year's loading of methylene chloride lost during each year of the surface
impoundment's active phase then can be calculated from Eq. 4-127:
fact = 0.86 •+ 0.12 + 2.1x10'* + (1.3x10~2 + 2.5*10~3) • (0.89 + 8.9*70'2) = 1.0
where: • . -
0.86 = • fml1 (fraction of total methylene chloride lost from liquid layer
that is lost to volatilization) calculated above.
0.12 = fdefl (fraction of total methylene chloride lost from liquid layer
that is lost to degradation) calculated above,
2.1xlO'J = foul, (fraction of total methylene chloride lost from liquid layer
that is lost in outflow from the impoundment) calculated above.
l.SxlO"2 = " fsepl (fraction of total methylene chloride lost from liquid layer
that is lost to seepage) calculated above,
2.5x1 0"J = fM, (fraction of total methylene chloride- lost from the liquid
layer as a result of the diminishing volume of the liquid layer)
calculated above,
°-89 = && (fraction of methylene chloride reaching the solids layer that
is lost to degradation) calculated above, and
8.9x10'- = f^ (fraction of methylene chloride reaching the solids layer that
is lost to seepage) calculated above.
Fifth, the duration of the square wave is calculated by using Eq. 4-129:
TP = - =
1.0 • 31,536,000sec/yr
where: .
2.5x10s = TF (estimated active lifetime of surface impoundment)
calculated above,
i-0 = faa (fraction of each year's loading of methylene chloride lost
during each year of the surface impoundment's active phase)
calculated above, and
31,536,000 = constant to convert (sec) to (yr).
4-146
-------
Sixth, the fraction of total pollutant lost from the impoundment through seepage is
calculated using Eq. 4-128:
_ (1.3X10 + .. ,„*„ ! y.7*.±v _ 1AxJQ-3
where:
1.3x10'- = fsep] (fraction of total methylene chloride lost from liquid layer
that is lost to seepage) calculated above,
2.5x10° = fMJ (fraction of total methylene chloride lost from the liquid
layer as a result of the diminishing volume of the liquid layer)
calculated above,
8.9x 1O'2 = . f^ (fraction of methylene chloride reaching the solids layer that
is lost to seepage) calculated above, and
L0 = fac, (fraction of each year's loading of methylene chloride lost
during each.year of the surface impoundment's active phase).
Seventh, the average flux of methylene chloride to the unsaturated zone beneath the
surface impoundment is calculated using Eq. 4-130:
0.01-^- • 1.4*70-3 • 206-^ • 4m • .31.32*
fA ' — ' (mS/m ) m3 kg
* lee, methylene chloride ' •
S.Oyr
= 4.4x10-2 kg
ha-yr
where: ' . . ' . ' ' • • •
°-01 = constant to convert (mg/m2) to (kg/ha),
1.4x10° = /.c/, (fraction of total methylene chloride lost from impoundment
through seepage) calculated above,
206 = s: (concentration of solids in solids layer) calculated above,
• . , . !j " d» (total dePm of surface impoundment) from Exhibit 4-22,
-31--3 = , Cf (concentration of methylene chloride in sewage sludge) from
Exhibit 4-1, and
?'° = JP (duration of "square wave" for approximating the loading of
methylene chloride into the unsaturated soil zone) calculated
above.
Eighth, Eq. 4-131 is used to calculate the average concentration of methylene chloride
in seepage from the surface impoundment:
4-147
-------
x-. _ (kg/ha -m)
sep, methylene chloride
ha-yr
2.5-
= 1.7xlQ-3^
L
where:
0.1
4.4x1 0'2
2.5
constant to convert (kg/ha-m) to (mg/L),
FA,CC j (annual average flux of methyiene chloride leaching
through the floor of the surface impoundment) calculated above,
and
Oxef> (rate of seepage beneath the impoundment) from
Exhibit 4-22.
.chioncie ls then used jn the linked unsatuTated zone and saturated zone models
to estimate the concentration of methylene chloride at the well. C^., for methylene chloride
is modeled to be lAxW6 mg/L. This concentration is then used in Eq. 4-139 to estimate
human exposure:
1-4*10
• 2Llday
- ~ =
where:
1.4xlO'6
2
70
Cml j (concentration of methylene chloride in well water),
IW (volume of water ingested daily) from Exhibit 4-16, and
BW (body weight), assumed to be 70 kg.
Volatilization Exposure Pathway for Surface Impoundment
The volatilization of methylene chloride from a surface impoundment is calculated
below. Many of the parameters required for the volatilization exposure route were calculated
above for the groundwater exposure pathway, and therefore are not repeated below.
4-148
-------
Equation 4-13 5 is used to calculate the. fraction of total methylene chloride volatilizing
during a human lifetime:
fvh = 1.0 -0.86 =0.86
where: . '
L0 = .4, (fraction of each year's loading of pollutant lost during each
year of the surface impoundment's active phase), and
°-86 = fvoti (fraction of total pollutant lost from liquid layer that is lost
to volatilization) calculated above.
Equation 4-136 is then used to calculate the average flux of methylene chloride
volatilizing from the surface impoundment:
FA , = 31.3 mg/% • 0.86 • 0.01 • 2Q6kfflm3 • 4m
vol,methylenechloride ~ - • - 2i— -
70yr
= 3.2kgfha-yr -
where: '
31.3 = Cj (concentration of methylene chloride in sewage sludge) from
Exhibit 4-1, . . ..
°-86 = /•/, (fraction of total methylene chloride volatilizing during human
lifetime) calculated above,
0.01- = factor to convert (mg/m2) to (kg/ha),
206 = , 5", (concentration of solids in solids layer) calculated above, -
4 = 4,,- (total depth of surface impoundment) from Exhibit 4-22.' and
70 = LS (life expectancy), assumed to be 70 yr.
The source-receptor ratio (SRR) must be calculated next. First, the lateral virtual
distance to the surface impoundment is calculated using Eq. 4-104:
4-149
-------
20,236 m
22.5
where:
20.236
22.5°
= 403m
A (area of surface impoundment) from Exhibit 4-22. and
0 (the angle subtended by the surface impoundment's width)
from text.
The-standard deviation of the vertical distribution of concentration is calculated using
Eqs. 4-105 through 4-107:
where:
20.236
A (area of surface impoundment) from Exhibit 4-22.
FromEq. 107:
x = lQ-3km/m • 71m = 0.071 km
where:'
71
constant to convert (m) to (km), and
r'.(distance from the surface impoundment's center to the receptor)
calculated above.
From Eq. 106:
o. = (15.209 • 0.0710-81558) = 1.8m
where:
15.209
0.81558
0.071
a (corresponding to x = 0.071 km) from Exhibit 4-18,
b (corresponding to x = 0.071 km) from Exhibit 4-18, and
x (distance from the surface impoundment's center to the
receptor) calculated above.
4-150
-------
Equation 4-103 is then used to estimate the source-receptor ratio:
SRR = 2.032 • 20,236m2 • 1
(71m + 403m) • 4.5m/sec • 1.8m
= llsec/m
where:
2.032 • = empirical constant,
20.236 = A (area of surface impoundment) from Exhibit 4-22,.
1 = v (vertical term for dispersion of methylene chloride in air) from
Exhibit 4-22,
71 = *' (distance from me surface impoundment's center to the
receptor) calculated above,
403 = xy (lateral virtual distance to the surface impoundment)
calculated above.
4-5 = «/»(average wind speed) from Exhibit 4-22, and
L8 =
-------
EXP = 10~3mg/\ig • O.llng/m3 • 20m3/day
me thy lene chloride
= 3.1x1 0~smg/kg -day
where: .
10° = constant to convert (fig) to (mg).
0- ! ! " = . CairJ (average concentration of methylene chloride in ambient air
at the receptor location) calculated above, . •
20 = IA (hihalation rate) from Exhibit 4-16, and
70 = BW (body weight of an adult), assumed to be 70 kg.
4-152
-------
4.4 INCINERATION EXPOSURE METHODOLOGIES
This section evaluates human exposure to inorganic pollutants from the incineration of
sewage sludge. For Round Two, the main concern is the emission of inorganic pollutants,
because organic pollutant emissions were regulated as "Total Hydrocarbons" (THC) in Round
One. THC was used to take into account the fact that organic pollutants are both destroyed and
• created in the incineration process. Although a pollutant-specific limit may be developed for
dioxins and dibenzofurans in Round Two, in this Section only inorganic pollutant emissions are
discussed.
The analysis uses four steps to estimate risks from incineration of sewage sludge:.
1) estimate the rate at which pollutants are emitted from incinerator stacks;
2) estimate the transport and dispersion of pollutants in ambient air near
incinerators, .and determine the extent to which pollutant plumes overlap;
3) map expected, ground-level concentrations of pollutants onto human
populations; and
4) determine the extent of human exposure to emitted pollutants and the
resulting health risks. -
4.4.1 Estimating Emissions of Pollutants from Incinerators
The first step in estimating human exposure to pollutants through incineration is to
determine the rate at which pollutants are emitted from the stacks of sewage sludge incinerators.
The rate at which an inorganic pollutant is emitted is based on the mass of pollutant entering the
incinerator, the removal efficiency of the furnace, and any operating pollution control devices:
(4.140)
,,000 • 10
where: .
EJP = emission rate for pollutant j at incinerator/? (g pollutant/sec),
Cj — concentration of pollutant j in sewage sludge (mg pollutant/kg
sewage sludge),
MP = mass of sewage sludge incinerated at incinerator/? each year (kg
sewage sludge/yr),
RJP • - combined removal efficiency for pollutant/' of furnace and control
devices for incinerator p expressed as fraction of original pollutant
mass retained by the furnace or pollution control devices
(dimensionless),
31,536,000 = constant to convert (yr) to (sec), and
103 = constant to convert (mg) to (g).
4-153
-------
The rate at which a pollutant enters the incinerator is based on the feed rate for sewaee
sludge (Mp) and the concentration.of pollutant hi the sewage sludge (Cy). For a given mass of
an inorganic pollutant entering the incinerator, some fraction will remain in the bottom ash of
the furnace. Of the remainder, some is trapped by pollution control devices and the rest is
emitted from the stack. To estimate the fraction of pollutant released to the atmosphere, the
mass entering the incinerator (per unit time) is adjusted for the removal efficiency of the furnace
and controls (Rjp). The resulting estimates for emissions from individual incinerators represent
stack emissions hi units of grams per second for each pollutant (Ejp).
4.4.2 Modeling the Dispersion of Pollutants in Ak-
in Round One, dispersion of pollutants hi air was simulated with the Industrial Source
Complex Long Term (ISCLT) model (Bowers et.al., 1980; U.S. EPA, 1986) as implemented
in the Graphical Exposure Modeling System for personal computers, or PC-GEMS (U.S. EPA,
1989a). The model described the dispersion of pollutants as steady-state Gaussian plumes, and
allowed the user several modeling options. '
In Round One, all incinerator stacks were modeled as point sources. Depending on the
velocity and temperature of exit gases, plume rise was modeled as either momentum- or
buoyancy-induced; the appropriate option was selected automatically by the program. Both the
downwash and plume-rise-by-distance options were used, but (for lack of sufficient data) the
effects of surrounding terrain were ignored. For computational efficiency, the dispersion of ^
pollutants near each incinerator was modeled only once, using a unit rate of emissions (i e one MB
g/sec of pollutant emitted per kg/sec of sewage sludge incinerated). Resulting dispersion
estimates were converted to ground-level concentrations at individual locations, scaled by
expected emissions of each pollutant from each individual incinerator.
4.4.3 Mapping Dispersion and Pollutant Concentrations Onto a Unified,Grid
In Round One, results from the ISCLT model were reported as dispersion ratios in units
of /zg/m of pollutant concentration in ambient air per g/sec of pollutant emissions from
incinerator stacks. Separate coefficients were provided for selected locations in the area
surrounding an individual incinerator. The model allowed the user to choose between a
rectangular or polar grid for specifying these locations. In Round One, the rectangular grid was
selected and coordinates specified in such a way that results from the modeling of individual
incinerators could be integrated into a unified mapping of dispersion ratios for the U.S. as a
whole. Explicit details of this analysis are provided in the Risk Assessment Document (U S
EPA, 1993a).
Within each cell of the grid system, expected pollutant concentrations were calculated by
combining emission estimates from each incinerator with results from ISCLT. When a cell was
4-154
-------
'impacted by more than one incinerator, pollutant concentrations were summed to calculate total
expected concentrations for that cell:
where:
AA,j = estimated ambient air concentration of pollutant y in grid cell / due to
sewage sludge incineration (fig pollutant/m3 air),
Dip = dispersion ratio for grid cell i impacted by incinerator p (fig pollutant/m3
- air per g pollutant/sec), and
n = number of incinerators modeled.
For Round Two, rather than model the entire U.S. for this Comprehensive Hazard
Identification exercise, modeling results from the analyses performed for Round One were used.
For each of the 172 sewage sludge incinerators modeled in Round One, the maximum dispersion
ratio for any cell impacted by the incinerator was identified. The geometric mean of these
maximum dispersion ratios was then calculated. (An arithmetic average was not calculated
because the maximum dispersion ratios appeared to be log-normally distributed.) The geometric
mean of the incinerator's sewage sludge feed rates, was also determined; again, the feed rates
appeared to be log-normally distributed. Using an average removal efficiency for any given
pollutant, an "average" ambient air concentration was estimated:
or
. I 31,536,000 • 103
where:
~ average ambient air concentration of pollutant j due to sewage
sludge incineration 0*g pollutant/m3 air),
= geometric mean of the maximum dispersion ratios for 172
incinerators (fig pollutant/m3 air per g pollutant/sec),
~ average emission rate for pollutant j (g pollutant/sec),
= concentration of pollutant / in sewage sludge
(mg pollutant/kg sewage sludge),
Mavg = geometric mean of sewage sludge feed rates for 172 incinerators
(kg sewage sludge/yr),
Rj = combined removal efficiency for pollutant/ of furnace and control
devices expressed as fraction of original pollutant mass retained by
the furnace or pollution control devices (dimensionless),
31,536,000 = constant to convert (yr) to (sec), and.
1Q3 = constant to convert (mg) to (g).
4-155
-------
4.4.4 Estimating Human Exposure
Once average ambient air concentrations of each inorganic pollutant are estimated, an
estimate of human exposure is made by combining the concentrations with assumptions about
daily inhalation volume and body weight. Individual exposure to each pollutant is calculated as:
AA.IA 10~3
EXP. = — i - (4-1.43)
1 BW
where:
EXPj = . exposure to pollutant j (mg pollutant/kg body weight-day),
IA = inhalation rate (m3 air/day),
10'3 = constant to c'onvert 0*g) to (mg), and
BW = body weight (kg).
As can be seen from the equation, the conservative assumption that each person inhales
air at the estimated (outdoor) concentration for 24 hours per day for his or her entire lifetime
is made. It is also assumed that all of the inhaled pollutant is absorbed into the body, and thus
exposure is effectively equivalent to dose.
4.4.5 Data Inputs
\
To estimate human exposure through this pathway, several types of incinerator data are-
required: sewage sludge feed rates, dispersion ratios, and removal efficiencies. As mentioned
in Section 4.4.3, average values for these parameters were calculated. For sewage sludre feed
rates, a geometric mean of the feed rates for the 172 sewage sludge incinerators was calculated
to be 1.04 x 10 kg/yr. For dispersion ratios, after the maximum dispersion ratio had been
identified for each incinerator, the geometric mean was calculated to be 3,36 ^g/m3 per g/sec.
The removal efficiencies in two types of incinerators, multiple hearth with wet scrubber
and fluidized bed with wet scrubber,' were needed for inorganic pollutants. Unfortunately
removal efficiency data were not available for any of the inorganics other than beryllium which
was already evaluated in Round One. Therefore, exposures were calculated for two removal
efficiencies: 50- percent and 90. percent. Fifty percent was chosen as a very conservative
number; 90 percent was chosen as a more reasonable number, based on the removal efficiencies
of inorganic pollutants evaluated in Round One.
4-156
-------
4.4.6 Example Calculations for Incineration
To estimate human exposure to manganese from an incinerator with 50 percent removal
efficiency, Eq. 4-142 is used first to estimate the average ambient air concentration of
manganese:
AA = 3'36 ^S/m3} I (g/sec). • 1620 mg/ks - 1.04xl06 kg/yr • (1-0.5)
manganese ~~ - ' - : — — - 2-=- - - - -
31,536,000 sec/yr • 103 mgfg
= 0.090 \igfm3
where:
3;36 = Davg (geometric mean of maximum dispersion ratios) from text,
1620 • = Cj (manganese concentration hi sewage sludge) from Exhibit 4-1
1.04x10 = Mavg (geometric mean of sewage sludge feed rates) from text,
°-5 - Rj (combined removal efficiency) assumed to be 50 percent, '
31,536,000 = constant to convert (yr) to (sec), and
JO3 .= constant to convert (mg) to (g).
To then calculate the exposure to manganese, Eq. 4-143 is used:
EXP - a09° VSt™2 ' 20 *"3/
-------
4.5 RISK CALCULATIONS
In Sections 4.2, 4.3, and 4.4, estimates were made of exposure to pollutants from sewage
sludge that is land applied, placed in surface disposal sites, or incinerated. In this section, the
exposure estimates are combined with human and ecological toxicity values to obtain estimates
of risk by pollutant and exposure pathway. If risk values are greater than certain thresholds for
a given exposure pathway, that pathway is defined as "critical" for that pollutant.
4.5.1 Human Health Risk Calculations
For land application Pathways 1, 2, 4, .5, 12, 13, 14, and 15, and for surface disposal
and incineration, human exposure estimates to carcinogenic pollutants are combined with cancer
potency slopes to estimate individual risk:
. C/y. = EXPj • q/ (4-144)
where:
CIj = incremental cancer risk from pollutant j for exposed individual
(incremental risk of developing cancer per lifetime of exposure),
EXPj. = exposure to pollutant j (mg pollutant/kg body weight-day), and
q~ = human cancer potency of pollutant j (mg pollutant/kg body weight-day)'1 ~
For land application Pathway 3, the exposure must be modified to account for the
duration of exposure relative to lifetime:
j • DE • q}" (4-145)
where:
DE = .exposure duration adjustment (number of years of exposure divided by
expected lifetime of 70 years.).
For Pathway 3, an exposure duration adjustment of (5/70) was used.
In this analysis, if the individual risk for a given pollutant and exposure pathway
exceeded a value of 10"4 (one in 10,000), then the pathway was considered critical.
For non-carcinogenic pollutants for land application Pathways 1 through 5 and 12 through
15, as well as for surface disposal and incineration, estimated exposure was compared to the
Risk Reference Dose (RfD):
4-158
-------
RNCj =
RfD.
(4-146)
where:
RJDj =
ratio of the exposure to the RJD. for pollutant j (dimensionless), and
risk reference dose for pollutant y (mg pollutant/kg body weight-day).
If the ratio was equal to or greater than one for a given pollutant and exposure pathway then
the pathway was considered critical.
Since May, 1993, when human toxicity data were first obtained for the 31 -pollutant
candidates, some pollutant toxicity numbers have been changed, and others withdrawn Exhibit
4-24 presents the q, and RfD values used in the Comprehensive Hazard Identification for both
oral and inhalation exposure routes and indicates those numbers which have changed from those
EXHIBIT 4-24
Human Health Toxicity Numbers1
Pollutant
Acetic acid (2.4-
j.dichlorophenoxy)
Aluminum
I Antimony
Asbestos3
Barium
[Beryllium
I Bis(2-ethylhexyl-)phthalate
[j Boron-
[JBuianone. 2-
Carbon Disulfide
Cresol, p-
Cyanides (soluble salts and
complexes)
Dioxins and dibenzofurans
_
Endosulfan-II^
Cancer Potency Slopes
(risk/mg/kg-day)
1 Inhalation
....
8.4
3.0 x 10W4>
Oral
=====
1.9.x- lO'2
" — I.
Reference Doses (mg/kg-day)
Inhalation
J
:
4.3
1,4 x 10'2
1.4x 10^2)
5.7 x lO'3
8.6 x lO'2
2.9xlO"3
3.0 x 106*4' |
•
Oral
===^
l.OxlO'2 j
4.0 xlO-4
7.0 x ID'2
5.0 x lO'3
2.0 x lO'2
9.0 x 10-2
6.0xlO-"2)
— " —
IxlO'1
(3)
2.0 x 1C'2
5.0 xlO'5
4-159
-------
EXHIBIT 4-24
Human Health Toxicity Numbers (cont'd)1
Pollutant
Fluoride
Manganese
Methylene chloride
Nitrate
Nitrite
Pentachloronitrobenzene
Phenol
Polychlorinated biphenyls —
coplanar
Propanone, 2-
Propionic acid, 2-(2,4,5-
trichlorophenoxy)
Silver
Thallium
Tin
Titanium
Toluene
Trichlorophenoxyacetic acid,
0 4 5-
*-,*T,J
Vanadium
Cancer Potency Slopes
• (risk/mg/kg-day)
Tnha'?tini
1.5 x lO'3
-
Oral
7.5 x ID"3
2.6xlO-1
7.7
Reference Doses (mg/kg-dlay)
Inhalation
5 x ID'5®
9.0 x 10-'
4 x 10'1(2)
Oral
6.0 x 10-2
5.0 x 10-3(2>- (5)
6.0 x 1Q-2
1.6
1.0 x 10-'
S.OxlO'3
6.0 x 10"'
l.OxlO'1
8.0 x lO'3 .
5.0 x 10'3
8.0 x 10-5(2)
6.0 x 10-'
(3)
2.0 x 10-'
1.0 x 1012
7.0 x 10'3
Notes:
1 See Section 3.3.4 for a complete description of sources reviewed; see Exhibit 3-7 for references for individual toxicity
numbers unless marked with a (2).
1 Toxiciry data from IRIS (March 3, 1995).
* Toxiciry number withdrawn from IRIS for further consideration (March 3, 1995). No toxicity number available in
HEAST (March. 1994 tables).
4 U.S. EPA (1994c). • ' '
5 This RfD is for water intake, assuming the necessary amount of the trace nutrient has already been ingested with food.
4-160
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For land application Pathway 11, exposure was compared to an occupational Threshold
Limit Value-Time Weighted Average (TLV-TWA). The TLV-TWA is the time-weighted
average concentration of a pollutant to which nearly all workers can be repeatedly exposed over
an 8-hour workday and 40-hour work week, without adverse effect. The ratio of exposure to
the TLV-TWA was taken:
RTC. = -
1 TLV-TWA.
(4-147)
where:
RTC,
TLV-TWAj
ratio of the exposure to the TLV-TWA for pollutant j
(dimensionless), and
Threshold Limit Value-Time Weighted Average for pollutanty (mg
pollutant/m3 air).
Exhibit 4-25 presents the Threshold Limit Values used in this analysis. As shown in the
Exhibit, values were not available for three organic compounds.
EXHIBIT 4-25
Threshold Limit Values for Pollutants
Threshold Limit Values-
Tune Weighted Average
(mg/m3)1
Acetic acid (2,4-dichlorophenoxy)
0.2 fiber/cm3 3
Bis(2-ethylhexyl) phthalate
Butanone, 2-
Carbon disulfide
Cyanides (soluble salts and complexes)
4-161
-------
Pollutant
Dioxins and Dibenzofurans
1 Endosulfan-H
(I Fluoride
|j Manganese
1 Methylene chloride
1 Nitrate
1 Nitrite
|] Pentachloronitrobenzene
1 Phenol ,
1 Polychlorinated biphenyls (coplanar)
1 Propanone, 2-
Propionic acid, 2-(2'.4,5-trichlorophenoxy)
| Silver
1 Thallium
Tin
Titanium
Toluene
j Trichlorophenoxy acetic acid, 2,4,5-
Vanadium
1
Threshold Limit Values-
Tune Weighted Average
(mg/m3)1
0.1s
1.6
• r-
174
NA
NA
0.5s
19s
0.5s
1780
0.018
• o.i5
2
JO9
188
0.05
Notes:
NA means Not Applicable.
1 American Conference of Governmental Industrial Hygienists (1994).
2 Soluble salts.
* Crocidolite.
J Anhydrous sodium tetraborate.
"Skin.
* Hydrogen cyanide.
7 Fume. . . •
* Soluble compounds.
* Titanium dioxide.
4-162
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Example Calculations for Human Risk
For carcinogens, this example estimates the risk associated with the application of sewage
sludge containing dioxins and dibenzofurans to agricultural land. The exposure pathway is
Pathway 12, ingestion of fish and water from surface water that receives eroded soil. In the
example calculation. for exposure presented in Section 4,2.12, the sum of the exposures to
dioxins in water and fish was estimated to be 5xlO'n mg/kg-day. To estimate the individual risk
from this exposure to these carcinogens, Eq. 4-144 is used:
.
kg-day mgjkg-day
= 2xlO~4
where:
5.2xlO'n - = £XPy (exposure to dioxins) calculated in Section 4.2. 12, and
. S.OxlO6 = q*j (human cancer potency of dioxins) from Exhibit 4-24.
Given that the individual risk exceeds 10^, this pathway is critical for dioxins and dibenzofurans.
To demonstrate the use of an exposure duration adjustment for carcinogens, exposure to
a carcinogen must be calculated for Pathway 3. Using Eq. 4-5 for beryllium in sewage -sludge
used on agricultural land:
0.2g .
EXP - 8
kg-day
where:
0.2 = 75 (sewage sludge ingestion rate for agricultural land), from Section 4.2.3,
10'3 = constant to convert (g) to (kg),
8 = ^(concentration of beryllium in sewage sludge) from Exhibit 4-1, and
16 = BW (body weight of .child assumed to be exposed to agricultural land)
from Section 4.2.3.
4-163
-------
Then using Eq. 4-145:
" l.Qr/0 g- • • 4.3
/
70 (mg/kg-day)
where:
1.0x10^ "= EXPy (exposure to beryllium) calculated above,
5/70 = £>£ (exposure duration adjustment) from text above, and
4-3 = tf* (human cancer potency of beryllium) from Exhibit 4-24.
Given that the individual risk does not exceed 10^, this pathway is not critical for beryllium.
For non-carcinogens, boron in sewage sludge applied to forest lands in Pathway 4
provides an example. Total dietary exposure to boron from wild .animals was calculated in
Section 4.2.4 to be 0.029 mg/kg-day. Using Eq. 4-146:
0.029
mg
0.090—^-
kg-day
where: • . .
0.029 = EXPj (exposure to boron) calculated in section 4.2.4, and
0.090 = ##} (risk reference dose for boron) from Exhibit 4-24.
This ratio of exposure to RJD is less than one, and thus this pathway is not critical for boron.
For Pathway 11, a different type of risk to human health is calculated. Using the
exposure to manganese calculated in Section 4.2.11 and using Eq. 4-147:
— - 4. 1x20
m3
where:
4.1xlO'3 • = EXPj (exposure to manganese) calculated in section 4.2.11, and
4-164
-------
1 == TLV-TWAj (Threshold Limit Value-Time Weighted Average for
manganese) from Exhibit 4-25.
This ratio of exposure to TLV-TWA is less than one, and thus this pathway is not critical for
manganese.
4.5.2 Ecological Risk Calculations
For Pathways 6, 7, 9, and 10, to estimate risk to an herbivorous or an insectivorous
mammal, or an earthworm, an ecological risk quotient was calculated. The ecological risk
quotient is the ratio of the predicted exposure to an appropriate lexicological reference value:
.where:
RQj = ecological risk quotienl for pollutant j (dimensionless),
EXPAj = exposure of animal to pollutant j (mg pollutant/kg diet), and
TRVj = toxicological reference value for pollutant/ for an animal'(mg pollutant/ke
. diet). . . -
Ideally, for toxicological reference values (77?Vs), data for livestock, earthworms and
small mammals such as shrews and moles would be available; however, toxicity data are
generally not available for all of .these species. Instead, toxicity tests are most often performed
on a select number of "laboratory species," such as rats, mice, and dogs. Results from tests on
these species are assumed to be represeniative of the sensitivity of species experiencing exposure
in the field. • ;
Toxicological reference values for livestock and small mammals were taken from three
wrro f™011,1?6?; f JS^R (Agencv for Toxic Substances and Disease Registry) documents;
WHO (World Health Organization) documents, and data provided in HSDB (Hazardous
Substances Data Bank). Each of these sources summarized results of toxicity studies For this
analysis, none of the original studies were obtained. The documents reported either NOAELs
(no observed adverse effect levels), LOAELs (lowest observed adverse effect levels) or
information, necessary lo calculate NOAELs or LOAELs. When only a LOAEL was provided
the LOAEL was divided by a factor of ten to make it more comparable with the NOAELs For
a given pollutant, the lowest NOAEL (or adjusted LOAEL value) reported in any of the sources
was chosen to be the toxicological reference value for that species. This conservative practice
was deemed appropriate for this effort.
As shown in Eq, 4-8 in Section 4.2.6, animal exposure is calculated in terms of
concentration of pollutants in the food items (mg/kg); therefore, the toxicological reference
values also should be in concentration units (mg/kg). While HSDB provides more details on the
protocol of the toxicity studies, the ATSDR and WHO documents only report the toxicity test
4-165
-------
results in terms of mg/kg-day, regardless of the exposure metric actually employed in the
toxicity tests. When the reference values came from sources where the exposures were given
as mg pollutant/kg body weight-day doses, the values were convened to a mg/kg food
concentration. In many of the toxicity studies, exposure was probably originally reported as
food concentrations. However, because the original studies were not consulted, standard values
were used for the data needed to make the conversions from the mg pollutant/kg body weisht-
day values reported in the ATSDR and WHO documents and HSDB to food concentration
equivalents, as described below.
To make the conversions from mg pollutant/kg body weight-day to food concentration,
data on body weights, food consumption rates, and, sometimes, water consumption rates were
needed. Body weights for the various species in the tests were taken from the table of reference
body weights in EPA's report Recommendations for and Documentation of Biological Values for
Use in Risk Assessment (Table 1-2, U.S. EPA, 1988). Allometric equations for daily food and
water consumption were taken from Table 1-3 of the same source. Daily doses were convened
to food concentrations by multiplying the dose by the body weight and dividing by the daily food
consumption rate. . .
When the exposure was given in terms of pollutant concentration in drinking water, the
water concentration was first convened to a mg pollutant/kg body weight-day dose by
multiplying by the drinking water rate and dividing by the body weight. The dose was then
convened to a food concentration in the same manner as described in the previous paragraph
In essence, the water concentration was multiplied by the ratio of the water consumption rate
to the food consumption rate.
Exhibit 4-26 presents the TRVs used for each pollutant for Pathways 6, 7 and 10 It
also shows which.species was used to derive the TRY, whether exposure conversions were
necessary, and whether the TRY was based on a NOAEL or a LOAEL. Toxicological reference
values could not be obtained for endosulfan-II or 2-propanone.
EXHIBIT 4-26
Toxicological Reference Values for Mammals
Pollutant
—
Acetic acid (2.4-dichlorophenoxy)'
Toxicological
Reference Value
(mg pollutant/kg diet)
• ^^-«.^_
180
Reference
(Species)
—
HSDB (mammals)N
Aluminum1
1400
Domingo et al., 1987 (rat)N
Antimony1
0.34
Schroeder et al., 1970 (rat)L
Barium1
0.70
Perry et al., 1983, 1985, 1989
(rat)N
Beryllium1
9.1
Schroeder and Mitchener, 1975
(rat)N
4-166
-------
EXHIBIT 4-26
Toxicological Reference Values for Mammals (cont.'d)
Toxicological
Reference Value
(mg pollutant/kg diet)
Reference
(Species)
Bis(2-ethylhexyl) phthalate
Canning et al., 1991 (rat)N
Weir and Fisher, 1972 (doe)N
Butanone. 2-
Ralston et al., 1985 (rat)N
Carbon disulfide
Jones-Price et al., 1984 (rabbirt1-
Hornshaw et al., 1986 (mink)N
Cyanides (soluble salts and
complexes)1
Gerhart, 1987 (rat)N
Dioxins and Dibenzofurans
Kociba et al., 1978 (rat)N;
Murray et al., 1979 (rat)L
HSDB (mice)*1
Laskey et al., 1982 (rat)L
Methylene chloride
Scrota et al., 1986 (rat)N
Pentachloronitrobenzene3
HSDB fdoe)L
NCI, 1980 (rat)N
Polychlorihated biphenyls
(coplanar)1
Barsotti and Van Miller, 1984
(monkey)1* .
Propionic acid, 2-(2,4,5-
trichlorophenoxy)
HSDB (dog)N
Rungby and Danscher. 1984
(mouse)L
Downs et al., 1960 (rat)N
Schroeder et al., 1968 (rat)L
Schroeder et al.. 1964 (mouse)N
NTP, 1990 (rat)N
Tnchlorophenoxy acetic acid,
2,4,5-'
HSDB (rat)N
Domingo et al., 1985 (rat)N
Notes:
1 Secondary source reported exposure
in units of mg pollutant/kg body weight/day.
4-167
-------
" Secondary source reported exposure in units of mg pollutant/kg drinking water.
1 Secondary source reported exposure in units of mg pollutant/kg food.
N Toxicological Reference Value was based on a NOAEL.
L Toxicological Reference Value was based on a LOAEL.
For soil-dwelling organisms in Pathway 9, TRVs also were needed. The soil-dwelling
biota includes a taxonomically very diverse array of organisms. There are very few
lexicological,data available, however, for most groups of soil-dwelling organisms. In ecological
risk assessments, one or a few species of earthworms are generally chosen to represent the soil-
dwelling niche. Therefore, searches focused on toxicity data for earthworms for the Round Two
candidate pollutants. . .
For Pathway 9, exposure to pollutants in soil by earthworms is measured by the
concentration of the pollutants in the sewage sludge/soil; therefore, the TRVs for earthworms
should be in units of soil concentration. While there is a considerable amount of toxicity data
for earthworms, the variability in test quality and designs makes results difficult to compare-
thus, it is very difficult to assess potential environmental hazards of pollutants to earthworms
(Roberts and Dorough, 1985; Edwards and Bohlen, 1992). There have been recent attempts to
standardize earthworm toxicity testing protocols (Roberts and Dorough, 1984; Callahan et al
1994). However, while standard laboratory testing protocols should help in comparing toxicities
of different pollutants or sensitivities of different species of earthworms, many of thelaboratory
testing protocols produce results mat are almost impossible to interpret in terms of field exposure
(Edwards and Bohlen, 1992). For instance, while placing earthworms in contact with pollutants
on filter papers in Pern dishes for two days may provide measurable LC50s, it is not clear how
these results can be used to determine risk from exposure to pollutants in the field In short
there is a dearth of toxicity data for earthworms that can be used in risk assessments.
The toxicity literature was searched through computerized databases (e g BIOSIS and
HSDB) and recent review articles were examined to find toxicity information on earthworms or
other soil biota for the Round Two candidate pollutants (e.g., Callahan et al 1994- JEdwards
and Bohlen, 1992; Beyer, 1990; Roberts and Dorough, 1985). Usable data were found only for
two of the pollutants. For phenol, Neuhauser and Callahan (1990) determined a NOAEC (no
observed adverse effect concentration) of 5900 mg pollutant/kg soil for mortality in the
earthworm Eisema fetida. The earthworms were exposed for eight weeks.to phenol in a
combination of sand and horse manure in Petri dishes. In a separate study, Hartenstein et al
(1981) investigated the effects of cations (including heavy metals) and anions added to activated
sewage sludge on the growth of E. fetida. The only Round Two metal investigated was
manganese. It was determined that manganese at the highest level tested (22 000 mg
pollutant/kg sludge) was innocuous to this species of earthworms. Therefore, a conservative
NOAEC estimate of 22,000 mg pollutant/kg sludge is used for manganese. Exhibit 4-27
summarizes these available data.
4-168 ,
-------
EXHIBIT 4-27
Toxicological Reference Values for Sou-Dwelling Organisms
Pollutant
Manganese
Toxicological Reference
Value (NOAEC in mg
pollutant/kg sewage sludge)
~" • ' I,....
22,000
Reference
(Species)
=^===
Hartensteinet al., 1981
(Eisenia fetida)
Phenol
5,900
Neuhauser and Callahan, 1990
(Eisenia fetida)
Example Calculation for Ecological Risk .
; ,^o T° estflate^isk to animals from sewage sludge-amended land, the same equation Eq
4-148, is used for Pathways 6, 7, 9, and 10. As an example, the risk for predators of soil
dwellmg organisms (Pathway 10) through exposure to manganese in sewage sludge-amended
agricultural soil can be estimated using Eq. 4-148:
manganese
llmg/kg
where:
10
17
EXPAj (exposure to manganese) calculated in Section 4.2.10, and
TRVj (toxicological reference value for manganese) from Exhibit 4-26.
10 is «*
4.5.3 Human Health and Ecological Risk Results
Presented in this section are estimates of risk for those pollutant-exposure pathway
combinations for which all pollutant-specific data are available. For these estimates 95th
percentile pollutant concentrations, with non-detects set equal to the minimum detection 'level '
as determined in the 1988 National Sewage Sludge Survey were used. Risks for each of the land
application exposure pathways are presented first, followed by risks for surface disposal and
mcmeration Note that in the following exhibits, a blank entry means that a risk estimate could
not be calculated for that pollutant-pathway combination, either because not all pollutant-specific
data were available or, in the case of human endpoint pathways, because neither q* or RfD
values were available. - • y
4-169
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4-173 .
-------
EXHIBIT 4-33
Risk Results for Highly Exposed Individual for Pathway 6
Pollutant
Aluminum
Beryllium
Boron
Dioxins arid
Dibenzofiirans
Fluoride
Manganese
Polychlorinated
biphenyls (coplanar)
Silver
Titanium
Risk Quotient1
Agricultural
Land
100 ,
0.04
0.7
0.03
10
200
5'
\
0.02
7
Forest
80
0.3
4
0.6
30
800
80
0.3
0.7
Reclamation
Site
100
0.03
0.7
0.03
10
200
4
0.01
7
Public
Contact Sites
. 80
0.3
4
0.6
30
800
80
0".3 :
0.7
1 Ratio of exposure to ecological toxicological reference value.
4-174
-------
EXHIBIT 4-34
Risk Results for Highly Exposed Individual for Pathway 7
I Pollutant
Acetic acid (2,4-
| dichlorophenoxy) '
1 Aluminum
•jj Antimony
1 Barium
-j| Beryllium
Bis(2-ethylhexyl) phthalate
jj Boron
|j Butanone, 2-
j| Carbon disulfide
I Cresol, p-
Cyanides (soluble salts and
jj complexes)
[Dioxins and Dibenzofurans
;
Fluoride
jj Manganese
1 Methylene chloride
Pentachloronitrobenzene
Phenol
Polychlorinated biphenyls
(coplanar)
Propionic acid, 2-(2,4,5-
j trichlorophenoxy)
1 Silver
[ Thallium
Tin
1 _
Agricultural
Land
<0,01.
0.4
1
40.
0.01
. 0.02
0.01 .
<0.01
<0.01
0.01
0.2
0.4
0.6
1
<0.01
<0.01
<0.01
0.6
<0.01
0.2
0.03
2
Risk Quotient1
Forest
'
<0.01
0.4
1
40
0.01
0.02
0.01
<0.01
<0.01
0.01
0.2
0.4
0.6
1 .
<0.01
<0.01
<0.01 .
0.6
<0.01
0.2
0.03
2
Reclamation
Site
<0.01
0-4'
1
"40
^
0.02
0.01 I
<0.01 "I
<0.01 ||
0.01 . 1
0.2 1
:
0.4
0.6 1
.''I.-'
<0.01~
<0:01 1
<0.01 ]
0.6 1
<0.01 • 1
0.2
0.03
2 I
4-175
-------
EXHIBIT 4-34
Risk Results for Highly Exposed Individual for Pathway 7 (cont'd.)
Pollutant
Titanium
Toluene
Trichlorophenoxy acetic acid,
2,4,5-
Vanadium
Risk Quotient1
Agricultural
Land
0.2
<0.01
<0.01
0.2
Forest
0.2 .
<0.01
<0.01
0.2
Reclamation
Site
0.2
<0.01
<0.01
0.2
1 Ratio of exposure to ecological toxicological reference value.
EXHIBIT 4-35
Risk Results for Highly Exposed Individual for Pathway 9
Pollutant
Manganese
Phenol
Risk Quotient l
Agricultural
Land
0.02 .
<0.01
Forest
0.07
<0.01
Reclamation
Site
0.02
<0.01
Public
Contact Sites
0.07
<0.01
1 Ratio of exposure to ecological toxicological reference value.
4-176
-------
EXHIBIT 4-36
Risk Results for Highly Exposed Individual for Pathway 10
Pollutant
Antimony
Barium
Dioxins and
Dibenzofiirans
Fluoride
Manganese
Thallium
Risk Quotient1
Agricultural
Land
10
0.6
<0.01
50
80
0.04
Ratio of exposure to ecological toxicological reference value.
Reclamation
Site
10
0.6
<0.01
Public
Contact Sites
50
80
JL
2
0.04
4-177
-------
EXHIBIT 4-37
Risk Results for Highly Exposed Individual for Pathway 11
Pollutant
Aluminum
Antimony
Barium
Beryllium
Bis(2^ethylhexyl) phthalate
Boron
Butanone, 2-
Cresol, p-
Cyanides (soluble salts and complexes)
Endosulfan-H
Fluoride
Manganese
Methylene chloride
Phenol
Polychlorinated biphenyls (coplanar)
Propanone, 2-
Silver
Thallium
Tin
Titanium
Toluene
Vanadium
Exposure
as a Fraction of TLV-TWA
0.1
<0.01
<0.01
<0.01
<0.01
<0.01
<0.01
<0.01
<0.01
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EXHIBIT 4-43
Risk Results for Highly Exposed Individual for Incineration Pathway
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4-184
-------
-------
5. FURTHER ANALYSES OF ROUND TWO POLLUTANTS
5.1 INTRODUCTION
Chapter 4 presented the risk assessments used in the Comprehensive Hazard Identification
to evaluate the list of candidate pollutants for the Part 503 Round Two regulation. In that
Chapter, results are presented only for those pollutant-exposure pathway combinations for which
all pollutant-specific data are available. Examples of pollutant-specific data are plant-uptake
slopes for different crops; animal uptake slopes for livestock, poultry, etc.; distribution
coefficients (K^); and human and ecological toxicity values. In this chapter, the candidate
pollutants that warrant further consideration for the final list are presented. For each pollutant,
the critical pathways, defined as exposure pathways for which the carcinogenic risk is 1 x 10~*
or higher, the ratio of exposure to the Risk Reference Dose (RfD) is one or greater, or the
ecological risk quotient (RQ) is one or greater, are summarized.
5.2 POLLUTANTS THAT WARRANT FURTHER CONSIDERATION
Based on the results of the risk assessments of the Comprehensive Hazard Identification,
12 pollutant candidates have critical pathways for land application and five pollutant candidates
have critical pathways for surface disposal. These pollutant candidates and their critical
pathways are summarized below in Exhibits 5-1 and 5-2, respectively. None of the inorganic
pollutants evaluated had a critical pathway for incineration.
5-1
-------
> EXHIBIT 5-1
Pollutants with Critical Land Application Pathways
I Pollutant
|| Aluminum
Antimony
Barium
J] Beryllium
|| Boron
| Dioxins and
Dibenzofurans
|| Fluoride
Manganese
PCBs-coplanar
Thallium
1 Tin
Titanium
' '..
Critical Agricultural Pathways
6
7, 14
7, 10, 14
14
2, 3, 10, 12, 13, 15
6, 10
3, 6, 7, 14
3, 4, 5, 6, 15
3
7
6
:^
..
Critical Non-Agricultural Pathways
6 (for., rec., pub.)
7 (for., rec.); 10 (for., pub.); 14 (for., rec.,
pub.)
7 (for., rec.); 10 (for., rec., pub.); 14 (for., I
rec., pub.) . j
14 (for., rec., pub.) j
6 (for., pub.) j
3 (for., rec., pub.); 10 (for., rec., pub.); 12 1
(for., rec., pub.); 13 (for., rec., pub.);
15 (for., rec., pub.)
6 (for., rec., pub.); 10 (for., rec., pub.)
3 (for., rec., pub.); 4 (for., rec.); 6 (for.,
rec., pub.); 7 (for., rec.); 10 (for., pub.); 14
(for., rec., pub.)
3 (for., rec., pub.); 4 (for., rec.); 5 (for.,
rec.); 6 (for., rec., pub.); 13 (for., rec.);
15 (for., rec., pub.)
" .
3 (for., rec., pub.)
7 (for., rec.)
6 (rec.) ||
Notes:
Pathway 2 = residential .home gardener
Pathway 3 = child ingesting sewage sludge
Pathway 4 = human ingesting animal products
Pathway 6 = livestock ingesting forage/pasture
Pathway 7 = livestock ingesting sewage sludge
Pathway 10 = soil organism predators ingesting soil organisms
Pathway 12 = humans ingesting surface water and fish
Pathway 13 = humans inhaling volatilized pollutants
Pathway 14 = humans ingesting groundwater
Pathway 15 = breastfeeding infant
.for. = forest land
rec. = reclamation site
pub. = public contact site .
5-2
-------
EXHIBIT 5-2
Pollutants with Critical Surface Disposal Pathways
Pollutant
Antimony
Barium
Beryllium
Dioxins and
Dibenzofurans
Manganese
Surface Impoundments
Groundwater
Groundwater
Groundwater
Air
Groundwater
From Exhibits 5-1 and 5-2, it is evident that the organic candidate pollutants dioxins and
dibenzofurans as well as coplanar PCBs have more critical pathways than the inorganic candidate
pollutants, except for manganese, which has the same number of critical pathways. These two
organic pollutant candidates are recommended to be included on the list of pollutants for the
Round Two regulation. The Agency has decided not to recommend including any of the
inorganic pollutants on the list for the Round Two regulation, however. The justifications for
that decision are presented in Appendix D on a pollutant by pollutant basis.
5-3
-------
-------
6. LIST OF POLLUTANTS FOR THE ROUND TWO REGULATION
SUBMITTED TO THE COURT
In May, 1993, the Agency submitted a list of 31 pollutant candidates for the Part 503
Round Two regulation to the District Court in Oregon. A copy of the court notice is presented
in Appendix Dl. On November 30, 1995, EPA submitted the final list of pollutants for the Part
503 Round Two regulation to the court. A copy of that court notice is presented in Appendix
D2. '
After considering the results of the Comprehensive Hazard Identification, the analysis of
pollutants that warranted further consideration, and information received from others, EPA
concluded that two pollutants should be on the list for each use or disposal practice. Tliey are:
dioxins/dibenzofurans (all monochloro to octachloro congeners) and polychlorinated biphenyls
(coplanar). The court notice indicates that EPA may, hi the exercise of its discretion, determine
to add or delete other pollutants to or from this list at the tune the Round Two regulation is
proposed. .
In addition to the list of pollutants submitted to the court, EPA may change a limit for
any of the pollutants in the Round One regulation during development of the Round Two
regulation. For this reason, the Round One pollutants also are considered pollutants for the
Round Two regulation.
Including the pollutants from Round One regulation, the list of pollutants for the Part 503
Round Two regulation by use or disposal practice is:
Land application
arsenic, cadmium, copper, lead, mercury, molybdenum, nickel, selenium, zinc,
dioxins/dibenzofurans, and coplanar polychlorinated biphenyls
Surface disposal .
arsenic, chromium, nickel, dioxins/dibenzofurans, and coplanar polychlorinated
biphenyls
Sewage sludge incineration
arsenic, beryllium, cadmium, chromium, lead, mercury, nickel, dioxins/
dibenzofurans, and coplanar polychlorinated biphenyls
Dioxins/dibenzofurans and coplanar polychlorinated biphenyls were included on the list
of pollutants for sewage sludge incineration even though they were regulated under the Total
Hydrocarbons operational standard hi Round One. EPA currently is conducting a reassessment
of dioxins/dibenzofurans. Because the results of this assessment are unknown,
dioxins/dibenzofurans were included on the Round Two list of pollutants for all use or disposal
practices. At the completion of the dioxin reassessment, EPA may decide not to regulate
6-1
-------
dioxins/dibenzofurans for a particular use or disposal practice of may. decide to regulate
dioxins/dibenzofurans on an accelerated schedule.
6-2
-------
7. REFERENCES
w
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7-1
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v
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7-2
-------
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U.S. EPA. 1993b Comparison and Rank of Proposed Human Health Bioaccumulation
Factors for the Great Lakes Initiative. EPA-822-R-93-010. Office of Water. August.
U'S- EP£™94t Revision of Assessment of Risks to Terrestrial Wildlife from TCDD and
Ism oncSt a* PSf S1U??f • Prepared by Abt Associates **. under contract no. '
68-CO-0093 for the Office of Pollution Prevention and Toxics. December.
U.S. EPA. 1994b. Estimating Exposure to Dioxin-Like Compounds Volume II-
Properties, Sources, Occurrence and Background Exposures. Office of Health and
Environmental Assessment. June. EPA/600/6-88/005Cb. External Review Draft.
U"S' ^r!^40^1 * Assessment Document for 2,3,7,8 - Tetrachlorodibenzo-p-Dioxin
(TCDD) and Related Compounds. Volume II. EPA/600/BP-92/001b. External Review
Draft. June.
U.S. Geological Survey. 1992. Element Concentrations in Soils and Other Surficial
Materials of the Conterminous United States. H.T. Shacklette and J.G Boerngen
U.S. Geological Survey Professional Paper 1270. Second printing.
Vanoni Vita A. (ed.). 1975. Sedimentation Engineering. Prepared by the ASCE Task
Committee for the Preparation of the Manual on Sedimentation of the Sedimentation
Committee of the Hydraulics Division, New York, NY.
7-16
-------
Verschueren, K. 1983. Handbook of Environmental Data on Organic Chemicals Van
Nostrand Reinhold Co., New York, NY. 2nd ed. [Cited in ATSDR, 1992f, 1993b.]
Voight, G., K. Henrichs, G. Prohl, and H.G. Paretzke. 1988. Measurements of Transfer
Coefficients for 137Cs, 60 Co, 54 Mn, 22 Na, 1311 and 95mTc from Feed into Milk
and Beef. Radiation and Environmental Biophysics. 27:153-164.
Walton, K.C. 1987. Effects of Treatment with Sodium Fluoride and Subsequent Starvation
on Fluoride Content of Earthworms. Bulletin of Environmental Contamination and
Toxicology. 38:163-170,
Wang, C.H. and F.E. Broadbent. 1972, Kinetics of Losses of PCNB and DCNA in Three
California Soils. Soil Sci. Soc. Amer. Proc. 36:742-745.
Weast, R.C. (ed.). 1990. CRC Handbook of Chemistry and Physics 70th ed CRC Press
Inc., Boca Raton, FL. - ,
Webber, M.D H.D. Monteith, and D.G.M. Corneau. 1983. Assessment of Heavy Metals
and PCBs at Sludge Application Sites. Journal of the Water Poll. Control Fed.
Weir, R.J., Jr. and R.S.S Fisher. 1972. Toxicologic Studies on Borax, and Boric Acid
Toxicol. Appl. Pharmacol. 23:351-364. [Cited in ATSDR, 1992d.]
Whelan, B.R. 1993 Effect of Barium Selenate Fertilizer on the Concentration of Barium
in Pasture and Sheep Tissues. J. Agric. Food Chem. 41:768-770.
'9 w A«- U, -L Infant GTOV/th and Human Milk Requirements.
Lancet. 2:161-163. [Cited in Smith, 1987.]
WHO. 1982. World Health Organization. Environmental Health Criteria 24- Titanium
Geneva. ' .
Wilson J.T., J.F. McNabb, D.L. Balkwill, and W.C. Ghiorse. 1983. Enumeration and
ria Indigen°US tO a Shallow Water-Table Aquifer. Ground
Yakushiji T LWatanabe K.Kuwabura,etal. 1978. Long-Term Studies of the Excretion
of PolychJonnated Biphenyls (PCBs) Through the Mother's Milk of an Occupational
ArCh' EnVir°n' C°ntam' T°XiCOL 7:493-504-
1993] rC' nVr°n' °ntam' °XiCOL 7:493-504- tcited in ATSDR,
fT Transport One-, Two-, and Three-Dimensional
of Waste Transport in the Aquifer System. Oak Ridge National
Laboratory, Environmental Sciences Division. Publication No. 1439. March.
7-17
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-------
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APPENDIX A
ANALYSIS OF POLLUTANTS DETECTED
LESS THAN TEN PERCENT OF THE TIME
-------
-------
Introduction
EPA conducted two screening analyses to determine if any of the 69 pollutants
detected less than ten percent of the time in sewage sludge might still pose an unacceptable
risk to human health. For the first screening analysis, EPA used the algorithms from
agricultural Pathway 3. This pathway tends to result in high risk because small children are
directly ingesting sewage sludge, without any of the mitigating influences of degradation
dilution, etc. found in other pathways. For the second screening analysis, EPA Devaluated
other pathways for pollutants with relatively large cancer potency slopes, or q,* values.
To conduct these analyses, human toxicity data were needed. Exhibit A-l presents
the available human toxicity data for the 69 pollutants as well as each pollutant's frequency
of detection, as measured in the 1988 National Sewage Sludge Survey (U.S. EPA, 1989a).
Screening Analysis Based on Pathway 3
To calculate exposure from agricultural Pathway 3, the only pollutant-specific data
required is the pollutant's concentration in sewage sludge, as described in Section 423
EPA chose to use 98th percentile pollutant concentrations with non-detects set equal to the
minimuni detection level. The Agency did not use 99th percentile concentrations because
such estimates are not as statistically meaningful when pollutants are only detected a few
percent of the time. For the non-pollutant-specific data required for this analysis, a sewage
sludge mgestion rate of 0.2 g/day, a body weight of 16 kg, and an exposure duration (for
cancer) of 5/70 were used.
„ A u rlSk' £ither ™ °ral Risk Reference Dose (RfD) or an oral q, * value was
needed. .Of .the 69 pollutants detected less than ten percent of the time, 49 had at least one
of these estimates of toxicity. Six of these pollutants had already been evaluated for Pathway
3 in Round One, and so were not considered further: aldrin, dieldrin. benzo(a)pyrene, DDT
DDE, and trichloroethene. For the remaining 43 pollutants, EPA estimated risk. For those
pollutants with an oral RfD value, the ratio of exposure to RfD was calculated For those
" value' ""• risk of cancer was '
As shown in Exhibit A-2, for all but one of the 43 pollutants analyzed, the ratio of
exposure to RfD was. below one and the cancer risk was below one in 100 000 For 2-
picoline the ratio of exposure to RfD was five. EPA chose to not evaluate 2-picoline
further, however, because it was only detected one percent of the time in the 1988 NSSS
A-l
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-------
Screening Analysis Based on Cancer Potency Slopes
The second screening analysis EPA conducted consisted of identifying those pollutants
with relatively high cancer potency slopes. As shown in Exhibit A-l, four pollutants, aldrin
dieldrin, heptachlor epoxide, arid benzo(a)pyrene, have relatively large q,' values. These
pollutants were evaluated further.
Although aldrin and dieldrin are both insecticides, they are often evaluated together
as aldrm/dieldnn, because dieldrin is an environmental degradation product of aldrin In
addition, aldrin and dieldrin have the same human health toxicity values. In Round One
aldrm/dieldrm were evaluated for Pathways 1 through 11,. but not 12, 13, or 14.
Given the log(Kow) value for dieldrin is greater than five, aldrin/dieldrin misht pose
an unacceptable risk by sorbing to. panicles that subsequently erode and enter a^stream
Aldnn/dieldnn is not expected to leach significantly to groundwater, given the high log(AT )
value However, aldrin/dieldrin might also pose an unacceptable risk through volatilization
Therefore, EPA evaluated risks from Pathway 12 and Pathway 13. for aldrin/dieldrin using
the assumptions and equations presented in Sections 4.2.12 and 4.2.13, respectively.
To correspond to the methods used 'in the Comprehensive Hazard Identification
exercise, the 95th percentile pollutant concentrations with the non-detect values set equal to
the minimum detection level were used. The pollutant-specific data for both pathways are
presented in Exhibit A-3. .**..-.
EXHIBIT A-3
Pollutant-Specific Data Required for Pathways 12 and 13
95th percentile concentration (mg/kg)
Kd(L/kg)
Henry's Law constant (atm-m3/mol)
(yr1)
Diffusivity in Air (cnr/sec)
BCF (L/kg)
FM (dimensiohless)
5.482
11733
l.lxlO-5(2)
O4
4xl0-2(3)
34003
^ Composite aldrin/dieWrin concentration from 1988 NSSS
' Schwarzenbach et al., 1993.
. Calculated using equations in Section 4.2.12.
4 Howard, 1991.
A-5
-------
Results of the analysis are presented in Exhibit A-4 for Pathway 12 and Exhibit A-5
for Pathway 13. For Pathway 12, the individual cancer risks range from 7xlO'9 for
reclamation sites to 2X10"8 for other land application sites. For Pathway 13, individual cancer
risks range from 9xlO'8 for agricultural land to IxlO"6 for reclamation sites.
EXHIBIT A-4
Individual Cancer Risks.for Aldrin/Dieldrin from Pathway 12
! Agricultural Land
2xlO'8
Forest
2xlO-8
Reclamation Site
7xlO-9
Public Contact
Site
2xlO'8
EXHIBIT A-5
Individual Cancer Risks for Aldrin/Dieldrin from Pathway 13
1 Agricultural Land
9xlO-8
Forest
4xlO-7
Reclamation Site
IxlO-6
Public Contact
Site
2xlO'7
For heptachlor epoxide, the individual risk for a child directly ingesting sewage sludge
(Pathway 3) was calculated above to be 9 x 10'7 (Exhibit A-2). Given the low magnitude of
the risk, this pollutant was not evaluated further.
Benzo(a)pyrene" was fully evaluated for all land application pathways in Round One
except Pathway 11 (tractor driver). Benzo(a)pyrene cannot be considered further in Round
Two for Pathway 11, however, because there is not a Threshold Limit Value for this pollutant
to be evaluated under Pathway 11.
A-6
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APPENDIX B
STATISTICAL ANALYSES
OF THE NATIONAL SEWAGE SLUDGE SURVEY DATA
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Final Report:
Percentile Estimates Used to Develop the List of Pollutants
for Round Two of the Part 503 Regulation
Submitted to:
Environmental Protection Agency
Office of Science and Technology
Engineering and Analysis Division
401 M Street, SW. (4303)
Washington, DC 20460
Submitted by:
Health and Environment Studies and Systems Division
Science Applications International Corporation
1710 Goodridge Drive
McLean, VA 22102
Contract No. 68-C4-0046
SAIC Project No. 01-0813-07-5046-010
An-Employee-Owned Company
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I. INTRODUCTION
In Februaiy, 1993, the Environmental Protection Agency (EPA) promulgated limits for nine toxic
pollutants in sewage sludge. These limits which were issued by EPA under the authority of
section 405(d) Clean Water Act, as amended, are referred to as the "Round One" sewage sludee
regulation. In May, 1993, the EPA submitted to the court a list of 31 candidate pollutants for
"Round Two" regulations. This report presents percentile estimates used to develop the list of
pollutants for Round Two of the Part 503 Regulation. All elements, compounds, or solids
physically measured will be referred to in this report as pollutants. The term pollutant is used
here to mean only that a substance, in certain quantities, could cause harm to human health or the
environment; not that it adll cause harm to human health or the environment.
\
Data analyzed to produce these pollutant concentration percentile estimates are from the EPA's
1988 National Sewage Sludge Survey (NSSS). Section H briefly describes the NSSS Data
conventions are presented in Section m. Section TV provides the statistical methodology
employed to produce me percentile estimates. And finally, Section V presents tabulated percentile
H. EPA's 1988 NATIONAL SEWAGE SLUDGE SURVEY
To support Round One and Two regulatory development efforts, the EPA's 1988 NSSS collected
sewage sludge quality and pollutant occurrence data from a national probability sample of Publicly
^ P°TWS iaC6
S± T ^ ^ (P°TWS) PiaC6^S 3t ^ "*"d«y "-—l °f wastewW
OperationaUy, secondary treatment was defined as a primary clarifier process followed by
biological treatment and secondary clarification. In 1988, 11,407 POTWs in the 50 States, Puerto
Rico, and the District of Columbia met this criteria.
A statistical probability sample of 208 POTWs in the contiguous states and the District of
Columbia comprised the analytical component of the 1988 NSSS. These POTWs were randomly
drawn from secondary or higher treatment POTWs which were categorized into one of four stZ
based on their average daily flow rate. These strata are defined as follows:
1) Flow greater than 100 million gallons per day (MGD)
2) Flow more than 10 MGD but less than or equal to 100 MGD
3) How more than 1 MGD but less than or equal to 10 MGD
4) Flow less than or equal to 1 MGD.
EPA contract personnel collected sewage sludge samples from 180 POTWs in the analvtical
"
l - ae
All sample collection and preservation was conducted according toprotocol Contract
-i^W'VBWV riUdge *** for 412 «** ipAPadapS'aS
methods 1624 and 1625 to allow volatile and semi-volatile organic analytes to bTcfuantified
the .sewage sludge matrix. Pesticides and polychlorinated biphenyls (PCBs) were
1
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according to EPA method 1618; method 1613 measured dibenzofiirans and dioxins; metals, other
inorganics, and classical were quantified according to standard EPA methods.' All chemical
analysis methods were either developed, chosen, or adapted to allow for the most reliable and
accurate measurement of the 412 analytes in the sewage sludge matrix.
A more detailed discussion the NSSS sampling plan, POTWs, and data is included in a November
1992 final report entitled "Statistical Support Documentation for the 40 CFR, Part 503 Final
Standards for the Use or Disposal of Sewage Sludge."
ffl. DATA CONVENTIONS
A total of 208 POTWs were selected for sampling as part of the analytical component of the 1988
NSSS. However, 32 POTWs were excluded from the statistical analyses because sewage sludge
samples were not obtained after the completion of secondary treatment of wastewater. POTWs that
were selected for the NSSS but excluded from the statistical analyses are listed on Table 1 The
EPISODE number listed on Table 1 designates the POTWs identification number in the analytical
survey. An episode number of "0" indicates that the POTW was selected for sampling as part of
the analytical-probability sample but samples of sewage sludge were not collected.
The reported national pollutant concentration estimates were calculated from a sample of 176
POTWs. These estimates apply to a population of 7,750 POTWs that practiced at least secondary
treatment of wastewater during 1988. Pesticides were not quantified for Surveylb 35-38-348
^SSSSf ° J116"*016' «*»*» f°r pesticides reported on the tables result from a sample
of 175 POTWs and are projected to a population of 7,720 POTWs in the Nation. Sewage sludge
samples from SurveylDs 23-07-036 (Episode =1554) and 35-05-012 (Episode=1561) were not
analyzed for the dioxin/furans. Therefore the dioxin estimates, generated from a sample of 174
^S>, "5 { ? a P0?"1^011 of 7'714 POTWs- Adjusted stratum weights for each sample size
are tabulated below. •
ADJUSTED WEIGHTS for STRATA (wj) by Sample Size
Sample size = 174
»i_-_»_^_i
27/7,714
301/7,714
1,838/7.714
5,548/7,714
Sample size = 175
——^____
27/7,720
307/7,720
1,838/7,720
Sample size = 176
-5=5S=__S
27/7,750
307/7,750
1,868/7,750
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In the NSSS, if a pollutant was measured above the Minimum Level, as adjusted for interferences
then the measure is considered a detection. In the August, 1989 document titled "Analytical
Methods for the National Sewage Sludge Survey," the EPA's Industrial Technology Division
defines a Minimum Level for pollutants quantified by gas chromatography combined with mass
spectrometry (GCMS) as the level at which " the entire analytical system shall give recognizable
mass spectra and acceptable calibration points." For elemental pollutants, the Minimum Level
is defined as "the minimum concentration of substance that can be measured and reported in 99%
confidence that the value is above zero." The final report for Round One Part 503 regulations
refers to the Minimum Level as "roughly equivalent to the minimum concentration or amount of
pollutant that could be measured."
If a pollutant was not measured above the Minimum Level, then estimates were generated using
two substitution methods. One set of estimate were produced using the value of the Minimum
Level for those samples for which the pollutant was considered to be an non-detect The second
set of estimates substituted zero for pollutant concentration value for those samples from which
a pollutant was not quantified above the Minimum Level. Tabulated results identify the
substitution method employed for the reported set of estimates.
Prior to calculating the estimates, pollutant concentrations were aggregated on a POTW basis to
form one concentration value per POTW for each pollutant. Field duplicate samples were
averaged together. For POTWs with multiple treatment trains, sample measuremen^pSlu^
concentrations were averaged together, using a weighted average based on the dry weight of
f ^ t? T** by ** treatment *** aSSOdated *** «* -Me, Primary sarnp7« f were
» a
secondary treatment Because the percent solids in sampled sewage sludge ranged from kn ten
raf^toofT^^^
as a function of the sample s percent solids: This transformation allows a standardized basis for
The dioxins and furans are reported individually and in aggregate Agereeates were
-
first convention the composite dioxin was considered a detect if all of the individual conveners
a7^S"1rr
SL™ TV- ^ ^ °f determining a detec*°n for the composite dioxin, the
compete dioxin was considered a detect if at least one of the individual congeners was detiS
T111"1 T^651^ » ^ignated «dioxinb." TEF adjusted estimates of the
congeners appear in Section V.
-------
PCBs were also mathematically aggregated. These aggregates were generated as described above
with the exception that the individual PCB's were hot multiplied by a toxicity equivalence factor.
IV. STATISTICAL METHODS
Percentile estimates were calculated using the nonparametric, weighted cumulative distribution
function (CDF) technique. Denote the dry weight concentration of a given pollutant in the
sampled sewage sludge from the j* POTW in the i* survey flow stratum as X,. The vahjes of the
variable X,. were then sorted in order of increasing concentration. The values of the adjusted
sur^w«ghts (Wi) associated with the ordered values of X are then summed until the first
If Xp is defined as the concentration of the p* percentile then,
4
X = F(XUp where F(X) .= £ V.F. (X)
•^~ J. 2. „
with
7=1
and KXg 5 x) = 1 if X^ £x for x aO
= 0 otherwise.
To determine the pollutant concentration associated with the p* percentile, an inverse function was
applied to the^ cumulative distribution function. Define the D* oercentile as
P !, a
PJPtog .p/100 Tneinverseof this function F'(p), is the smallest value of x satisfy lx),p
where p is the desired percentile point (P) divided by 100.
Because the cumulative distribution created by application of the formula in the previous section
^empirical, integer valued percentile points are not always realized in the data. The conation
** ^f11^011 aSS°daled ^ ** P* "«*« P6^^6 fi«» *e empiric^
was to determine the smaUest concentration value x such that FUx) >p Tto
^^0^ «-—*»• ^ next smallest conc^S^f^m
nH f^ wth me (q-Dst ordered concentration was then defined. The
concentetion value for the p* percentile was obtained using linear interpolation between the a*
ana (q-lj values. , . n
Nonparametric estimates of pollutant concentration means and standard deviations are also
reported in the tables. Retaining the definition of X, as the dry weight concentration of a. given
-------
pollutant in the sampled sewage sludge from the j* POTW in the i* survey stratum and w- as the
adjusted survey weight for the i* stratum, then the mean pollutant concentration was estimated as
listed on the next page.
£
v
n..
The poUutant concentration standard deviation was estimated as the square root of the method of
moments estimator of the variance. That is:
1/2
•* "it y *•
V(X)i/2 =
V. POLLUTANT CONCENTRATION PERCENTILE ESTIMATES
Tables 3 and 4 present pollutant concentration percentile estimates for pollutants from the 1988
National Sewage Sludge Survey (NSSS.) Taking into account the individual dioxin and furan
congeners and the PCS aroclors, Tables 3 and 4 present concentration estimates for 353
pollutants. The listing of pollutants is ordered by percent detection. The ordering is from highest
to lowest detection rates in the nation. Excluded from this listing are the metals regulated in
Round One, and the 42 semiquantitative metals listed on Table 5. Of the 42 semiqukntitative
metals, 36 had no quantitative measurements recorded in the NSSS database. Of the remainine
six, potassium and iodine had one recorded measure while silicon, strontium, and sulfur had
TrTZ6"!5 re«)rded/oi;itwo ^P165- AU other samples were missing measurements. This
precluded estimation of poUutant concentrations. Estimates of phosphorus concentrations were
generated from data collected using colorimetric method 365.2 as reported in EPA's August 1989
Analytical Method for the National Sewage Sludge Survey."
For each poUutant, the tables report the foUowing: poUutant type, unit of measure, sample size
t^*?*^*S* Pf^detect' mean> ^dard deviation, the observed maximL, and
me w , 9S" 9S», 90" and median percentiles estimated from empirical national, cumulative
rf?S?^fj?° M«^ncen^°f ""*'C0lumn labeled "Sample""records *e number
of POTWs in the NSSS from which data were used to generate the reported estimates.
Table 3 is subtitled "Nonparametric Substitution Method Estimation Procedure - Nondetects Set
to the Minimum Level." The nonparametric estimation procedure is that described in Section IV.
-------
The substitution of nondetects set to Minimum Levels indicates that Minimum Level of a pollutant
was used in the estimation procedure for those samples that were not quantified above the
pollutant's Minimum Level of detection. Estimates on Table 4 were generated using the value
zero for samples from which a pollutant was not quantified above the Minimum Level.
Tables 3 and 4 indicate that there.are 45 tested pollutants detected at an estimated national rate of
ten percent or higher from sewage sludge resulting from secondary or higher treatment of
wastewater in 1988. EPA used this list of pollutants in conjunction with human health and
ecological toxicity data to select the 31 candidate pollutants for Round Two regulation
-------
TABLE 1.
LISTING OF POTWS EXCLUDED FROM PERCENTILE ESTIMATION
FLOW
STRATUM
SURVEYED
12-49-455
^—"•^•••—™
21-25-234
REASON
——»••—••••.
Ineligible/Out of Busing
Not sampled
^••^•^^^M^hvM^H
Ineligible/Out of Business
Only primary sludge sampled
^^^^^^^^^""^""""^•^•^^•^^"•^^•^M
Data not entered into database
Only orimarv sludge samoled
25-50-472
•^—••
31-18-140
31-23-206
41-24-215
41-36-312
Not sampled
•••^^••^^^"—••^^MIM
Not sampled
Wastewater Stabilization pond
_._. *~
45-13-083
"^^^^•"«««"™"l^^^^
45-13-089
—•"™«—^«»^_
45-14-092
"••'' —
45-15-112
45-16-130
45-17-131
45-19-154
••^•MHHH.
45-23-208
45-24-220
45-25-229
^~^"^^"«™^^^-™
45-25-231
45-26-237
Ineligible/Out of Busin^c
-------
45-28-246
45-29-248
45-30-253
45-37-339
45-42-387
45-42-392-
45-45-415
45-45-423
45-50-463
45-50-474
0
0
0
0
0
- 1488
0
0
0
0
WWSP
WWSP
WWSP
Not sampled
Ineligible/Out of Business
Ineligible/Out of Business
WWSP
Not sampled
Not sampled
WWSP
4
4
4
4
4
4
4
4
4
4
8
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TABLES
SEMIQUANTTrATIVE METALS in the NATIONAL SEWAGE SLUDGE SURVEY
| CAS NUMBER | CHEMICAL NAME |
I 7440699 1 . BISMUTH j
I 7440451 | CERIUM |j
1 74299.16
1 7440520
' 7440531
7440542
7440553
• 7440564
7440575
7440064
7440097
7440100
7440155
7440166
7440188
7440199
7440202
7440213
7440586
7440600
7440746
DYSPROSIUM II
ERBIUM 1
EUROPIUM 1
GADOLINIUM |
GALLIUM III
GERMANIUM III
GOLD |
PLATINUM I
POTASSIUM I
PRASEODYMIUM |||
— ill
RHENIUM |
RHODIUM III
RUTHENIUM |||
.SAMARIUM |
SCANDIUM If
SILICON I
— "' lil
HAFNIUM I
HOLMIUM |
INDIUM If
| CAS NUMBER 1 CHEMICAL NAME
1 7440246 j STRONTIUM
7704349
7440257'
13494809
7440279
7440291
7440304
7440337
7440611
7440031
, 7440644
7440042
7440677
7440053
7723140
7553562
7439885
7439910
7439932
7439943
7440008
SULFUR
TANTALUM
TELLURIUM
TERBIUM
THORIUM
THULIUM
TUNGSTEN
URANIUM
NIOBIUM
YTTERBIUM
OSMIUM
ZIRCONIUM:
PALLADIUM:
PHOSPHORUS
IODINE
IRIDIUM
LANTHANUM
LITHIUM
LUTETIUM
NEODYMIUM
46
-------
APPENDIX C
CALCULATION OF A "SQUARE WAVE"
FOR THE GROUNDWATER PATHWAY
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Potential human exposure and risk through the groundwater pathway are estimated for
VAnn^T apP'1C,ati°n ai?d surfacfe disP°saI of sewa^ sludge. To prepare input for the
VADOFT model of pollu£am ^p^ ^^ ^ unsaturated 2one k ig conservativej
assumed for both land application and surface disposal that the pollutant is consistently loaded
into the top of the unsaturated zone at the maximum rate estimated by mass balance
calculations. The.duration of this constant pulse, or "square wave", is constrained so fl«£
total mass of pollutant leaching or seeping from the site is conserved. Althoush the general
approach is the same for both land application and surface disposal, details differ aSne
to which management practice is being considered. This append provides a brief dTscuSon
of Ac methods for estimating the magnitude and duration of the "square wave" of poHumm
loading for land application and both prototype facilities for surface disposal
Land Application
Both inorganic and organic pollutants can accumulate in soil with repeated applicatio
of sewage sludge. As described in Chapter 4, it is assumed that all compel?poSumnt
processes for sewage sludge-amended soil can be approximated as firsSfrder and
coefficients describing the rate of loss to each process can be summed to
lumped" coefficient for first-order loss. Losses at any time , can then te
dMt
where:
M, - mass of pollumnt in sewage sludge-amended soil at time t (kg) and
K,0, - total loss rate for the pollutant from sewage sludge-amended soil (yr>).
Mt = fpA e~K^ dx =
o K.
where:
PA
total annual loading of pollutant to site (kg/yr).
Aw approaches inflnity, M, therefore app^ches (PA,/Km and yearly loss approaches yearly
C-l
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soil. Estimates of risks from organic pollutants on land application sites are derived for this
steady-state condition. The amplitude of the square wave pulse for the groundwater pathway
model is therefore equal to the annual loading of pollutant multiplied by the fraction of annual
loss attributable to leaching, the length of the square wave is equal to the length of the
simulation (300 years).
For inorganic pollutants, this condition of steady-state is not necessarily reached. The
leaching of inorganic pollutants from sewage sludge to groundwater depends not only on the
cumulative loading of inorganic pollutants, but also on the period of time in which this
cumulative loading takes place. It is assumed that, after 20 years, applications are
discontinued. To capture the risks associated with the peak rate at which inorganic pollutants
leave the soil layer, the peak loss rate (calculated for the 20th year of application) is used for
the calculations. The length of the square wave is calculated by dividing the total
(cumulative) loading of pollutant by this maximum rate of loss: • •"
TP = N PA = N
PA (l-e~K^) (1-
where:
TP = duration of "square wave" for approximating the loading of pollutant
into the unsaturated soil zone (yr).
'
---------- c --- ________ „ *.«wtJFt
The modeling of the groundwater pathway for the monofill prototype of surface
disposal is similar to that for land application. For both cases, it is assumed that the site
receives repeated loadings of pollutant for the duration of its active lifetime. By analogy with
the above discussion for land application, this maximum rate of loss from the facility can be
described as a function of its yearly loading, yearly loss, and number of years of active
operation:
*„**„* PA tl-e-*-")
where:
LF = active lifetime of monofill (yr),
to,* = .mass of pollutant in sewage sludge/soil at end of monofill's active
lifetime (kg), and
PA = total annual loading of pollutant to monofill (kg/yr).
•
C-2
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The length of time this maximum rate of loss could be maintained is then:
TP =-
LF PA LF
PA'a-e'*-1*) 1-e-*-^
Surface Disposal: Surface Impoundment Prototype
For the surface impoundment prototype of surface disposal, calculations are based on
the conservative assumption that steady-state is maintained for concentrations of pollutants
within the liquidI and sediment layers of the impoundment. It is also assumed that the flux
of pollutant leaching from the impoundment is constant with respect to time, at least until the
tota mass of pollutant deposited in the impoundment has been depleted For this orototvne
the length of the square wave used for execution of the VADOFT model is therefore eaual
to the total mass of pollutant entering the impoundment each year, multiplied by the expected
lifetime of the facility and divided by the amount lost each year- expected
PA
31,536,000 -PA •/„ 31,536,000 -/^
where:
total annual loading of pollutant into the surface impoundment
\*v§' j*/»
estimated active lifetime of surface impoundment (sec),
constant to convert (sec) to (yr), and
fraction of each year's loading of pollutant lost during each year
of the surface impoundment's active phase (dimensionless)
C-3
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APPENDIX D
EVALUATION OF CANDIDATE POLLUTANTS
FOR THE ROUND TWO SEWAGE SLUDGE REGULATION
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EVALUATION OF CANDIDATE POLLUTANTS FOR THE
ROUND TWO SEWAGE SLUDGE REGULATION
by
U.S. ENVIRONMENTAL PROTECTION AGENCY
401 M STREET, S.W.
WASHINGTON, D.C. 20460
AUGUST 1996
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TABLE OF CONTENTS
Page
TABLE OF CONTENTS . . . ;
1. INTRODUCTION . j
1.1 Background - j
1.2 Purpose . 2
. 1.3 Policy Decisions 2
1.4 Additional Information 3
2. POLLUTANT EVALUATIONS 4
2.1 Candidate Pollutants That Warrant Consideration 4
2.2 Information Used to Develop Rationales to Exclude
Inorganic Pollutants From Further Consideration 6
2.2.1 Land Application 7
2.2.2 Surface Disposal ...... -12
2.3 Rationales for Excluding Inorganic Pollutants From
Further Consideration 12
2.3.1 Land Application . . . . . ; « - . j^
2.3.2 Surface Disposal 28
2.3.3 Incineration 29
2.4 Pollutants Recommended by Others for the Round Two
List of Pollutants
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TABLE OF CONTENTS (cont'd)
Page
3. LIST OF POLLUTANTS FOR THE ROUND TWO SEWAGE SLUDGE
REGULATION 32
4. REFERENCES . 34
APPENDIX Dl: List of 31 Candidate Pollutants for the Round Two Sewage Sludge
Regulation Submitted to the District Court in Oregon
APPENDIX D2: Final List of Pollutants for the Round Two Sewage Sludge Regulation
" Submitted to the District Court in Oregon
APPENDIX D3: Responses to Requests for Data on the Round Two Candidate Pollutants
11
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1. INTRODUCTION
1.1 BACKGROUND
In 1987, Congress amended section 405 of the Clean Water Act (CWA) to require a
comprehensive program to reduce the potential public health and environmental risks from the
use or disposal of sewage sludge, which is solid, semi-solid, or liquid residue generated during
the treatment of domestic sewage in a.treatment works. Amended section 405(d) established a
timetable for the development of the sewage sludge use or disposal regulations. The basis for
the program Congress mandated to protect public health and the environment is the development
of technical requirements or standards for sewage sludge use or disposal, and the implementation
of the standards through a permit program.
Under the current section 405(d), EPA first had to identify toxic pollutants that may be
present in sewage sludge in concentrations that may affect public health and the environment.
Next, for each identified use or disposal practice, EPA had to publish regulations that specify
management practices for sewage sludge that contains the toxic pollutants and establish numerical
limits for the toxic pollutants. The management practices and numerical limits must be "adequate
to protect public health and the environment from any reasonably anticipated adverse effect of
each pollutant." Section 405(d) requires that EPA publish the sewage sludge regulations in two
rounds and then review the regulations periodically to identify additional pollutants for regulation.
On February 19, 1993, EPA .published the Round One sewage sludge regulation (i.e., the
Standards for the Use or Disposal of Sewage Sludge - 40 CFR Part 503)in the Federal Register
(58 FR 9248). It was amended subsequently on February 24, 1994 (59 FR 9095), and on
October 25, 1995 (60 FR 54164). .
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A candidate list of pollutants for the second round of the sewage sludge regulations (i.e.,
Round Two) was provided to the District Court in Oregon in May 1993 (see Appendix Dl). The
final list of pollutants was submitted to the District .Court in Oregon in November 1995 (see
Appendix D2). The Round Two sewage sludge regulation is scheduled for proposal in December
1999 and for publication in December 2001.
To develop the final list of pollutants for the Round Two sewage sludge regulation, a
Comprehensive Hazard Identification study was conducted by use or disposal practice for the 31
pollutants on the candidate list. Results of that study were used to determine the candidate
pollutants that warrant further consideration for the Round Two list of pollutants.
1.2 PURPOSE
This paper reviews the candidate pollutants from the Comprehensive Hazard Identification ^^
study that warrant further consideration for the Round Two list of pollutants and presents the
rationales for not including some of the pollutants on the final list. It also presents the pollutants
on the final list of pollutants for the Round Two sewage sludge regulation.
1.3 POLICY DECISIONS
For the review of the candidate pollutants from the Comprehensive Hazard Identification
study that warrant further consideration for the Round Two list, EPA made several policy
decisions. They are:
• Uptake rates from non-sewage sludge studies (i.e., crops for which the uptake rates
were obtained were not grown in sewage sludge-amended soil) are not appropriate
for crops grown in sewage sludge-amended soils because sewage sludge is
expected to "bind" pollutants and makes them less available for plant uptake
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(Corey et al., 1987). . '
Potential population effects are of greater concern than are individual effects for
exposure pathways in which the Highly Exposed Individual (HEI) is a
nonendangered animal.
The route through which a pollutant is administered (e.g., in drinking water or
food) hi a toxicity study should be considered when determining the applicability
. of the study to an exposure pathway.
A soil type for all land application sites and surface disposal sites of either sandy
loam, shrinking clay, or sand is reasonable.
A margin of safety that is smaller than the total uncertainty factor used for the
Reference Dose (RfD) is reasonable in .certain cases.
1.4 ADDITIONAL INFORMATION
Questions about the information in this paper should be addressed to:
Yogendra M. Patel or Robert M. Southworth
U.S. Environmental Protection Agency (4304)
401 .M Street, S.W.
Washington, D. C. 20460
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2. POLLUTANT EVALUATIONS
2.1 CANDIDATE POLLUTANTS THAT WARRANT CONSIDERATION
During the Comprehensive Hazard Identification study (U.S. EPA, 1996), 15 exposure
pathways were evaluated for land-applied sewage sludge and two pathways were evaluated for
sewage sludge placed on a surface disposal site. A pathway was considered "critical" for a
pollutant if the risk level for a carcinogenic pollutant was W4 or higher; the ratio of exposure
for a noncarcinogeriic pollutant to its Reference Dose (RfD) was equal to or greater than one; or
the risk quotient (RQ) for a pollutant for the ecological pathways was equal to or greater than
one.
Based on the results of the Comprehensive Hazard Identification study several of the
candidate pollutants had critical pathways for land application and for surface disposal. The
candidate pollutants and their critical pathways are presented in Table 2.1 for land application and
Table 2.2 for surface disposal. The exposure pathway for incineration (i.e., inhalation) was not
critical for any of the candidate inorganic pollutants. That pathway was not evaluated for the
organic pollutants because organic pollutants are controlled by the allowable concentration of total
hydrocarbons in the exit gas from a sewage sludge incinerator in the Part 503 regulation.
As indicated on Tables 2.1 and 2.2, dioxins, dibenzofurans, and coplanar polychlorinated
biphenyls (PCBs) have several critical pathways. For this reason and because dioxins,
dibenzofurans, and coplanar PCBs are bioaccumulative pollutants (i.e., they accumulate in human
and animal tissues) with reproductive effects, EPA concluded that those pollutants should be on
the final Round Two list of pollutants for land application and surface disposal.
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TABLE 2.1 - POLLUTANTS WITH CRITICAL LAND APPLICATION PATHWAYS
Pollutant
Aluminum
Antimony
Barium
Beryllium
Boron
Dioxins/furans
Fluoride
Manganese
PCBs - coplanar
Thallium
Tin
Titanium
Critical Ag Pathway
6
7,14 .
7,10,14
14
2,3,10,12,13,15
6,10
3,6,7,14
3,4,5,6,15
3
7
6
Critical Non-Ag Pathway
6(f,r,p)
7(f,r); 10(f,p); 14(f,r,p,)
7(f,r); 10(f,r,p); 14(f,r,p)
H(f,r,p)
6(f,p)
3(f,r,p); 10(f,r,p); 12(f,r,p);
13(f,r,p); 15(f,r,p)
6(f,r,p); 10(f,r,p)
3(f,r,p); 4(f,r); 6(f,r,p); 7(f,r);
10(f,p); 14(f,r,p)
3(f,r,p); 4(f,r); 5(f,r); 6(f,r,p);
13(f,r); 15(f,r,p)
3(f,r,p)
7(f,r)
6(r).
Pathway 2 - residential home gardener
Pathway 3 - child ingesting sewage sludge
Pathway 4 - human ingesting animal products (foraging animals)
Pathway 5 - human ingesting animal products (grazing animals)
Pathway 6 - livestock ingesting forage/pasture
Pathway 7 - livestock ingesting sewage sludge
Pathway 10 - soil organism predators ingesting soil organisms
Pathway 12 - humans ingesting surface water and fish
Pathway 13 - humans inhaling volatilized pollutants
Pathway 14 - humans ingesting groundwater
Pathway 15 - breast-feeding infant
f - forest; r - reclamation site; p - public contact site; ag - agricultural land; non-ag - non-
agricultural land
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TABLE 2.2 - POLLUTANTS WITH CRITICAL SURFACE DISPOSAL PATHWAYS
Pollutants
Monofills
Surface Impoundments
Antimony
Ground water
Barium
Ground water
Beryllium
Ground water
Dioxins/furans
Air
Manganese
Ground water
EPA also concluded that the inorganic pollutants with critical pathways for land
application and surface disposal should not be on the final list of pollutants for the Round.Two
regulation. The rationales for excluding those pollutants from the list are presented below.
2.2 INFORMATION USED TO DEVELOP RATIONALES TO EXCLUDE INORGANIC
POLLUTANTS FROM FURTHER CONSIDERATION
The Comprehensive Hazard Identification study used to evaluate the candidate inorganic
pollutants was, by design, conservative. After the critical pathways were identified for each
pollutant, a detailed examination of each pathway was conducted by EPA to confirm that the
pathway results supported inclusion of the pollutant on the final Round Two list of pollutants.
As part of the detailed examination for each critical pathway for a pollutant, three reviews
were conducted. First, the assumptions made in conducting the pathway exposure assessment
were reviewed. Next, the relevance of available toxicity data for a pathway to the Highly
Exposed Individual (HEI) for the pathway was reviewed. Finally, the magnitude of the ratio of
estimated exposure to the RfD for a noncarcinogenic pollutant in the non-ecological pathways
or
6
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the magnitude of the ratio of the estimated exposure to the toxicological reference value (TRV)
for a pollutant in the ecological pathways was reviewed.
2.2.1 Land Application • .
The information in Tables 2.3, 2.4, 2.5, and 2.6 was used in the detailed examination of
the critical land application pathways. Table 2.3 contains a summary of conservative assumptions
for several of the critical pathways. Table 2.4 contains the Highly Exposed Individual (HEI) for
each of the critical pathways, and Table 2.5 contains the measurement endpoint for each pollutant
by critical pathway and the species used to develop the endpoint. Table 2.6 contains the results
of the Comprehensive Hazard Identification study for each of the critical pathways.
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TABLE 2.3 - SUMMARY OF CONSERVATIVE ASSUMPTIONS
Pathway
.3
4
6
. 7
10
14
Conservative Assumption
One hundred percent of the material that the child ingests is sewage sludge,
not a mixture of soil and sewage sludge.
Results from non-sewage-sludge studies can be used to develop pollutant
uptake slopes into forage/pasture.
Herbivorous livestock or small herbivorous animals forage only on land on
which sewage sludge has been applied; results from non-sewage-sludge
studies can be used to develop pollutant uptake slopes, into forage/pasture.
Herbivorous livestock graze only on land on which sewage sludge has been
applied.
All of the soil organisms ingested by small mammals are exposed to sewage
sludge-amended soil and, therefore, bioconcentrate pollutants.
The soil-water partition coefficient used is the lowest soil-water partition
coefficient for sandy soil with a porewater pH of 5.
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TABLE 2.4 - HIGHLY EXPOSED INDIVIDUALS FOR CRITICAL PATHWAYS
Pathway Number
3-agricultural
3-non-agricultural
4-non-agriculturaI
6-agricultural
6-non-agricultural
7-agricultural
7-non-agricultural
10-agricultural
1 0-non-agricultural
14-agricultural
1 4-non-agricultural
Highly Exposed Individual (HEI)
Child ingesting sewage sludge
Child ingesting, sewage sludge
Human uigesting deer and elk
Herbivorous livestock
Herbivorous livestock (forest, reclamation site); small
herbivorous mammal (forest, public contact site)
Herbivorous livestock
Herbivorous livestock
Small insectivorous mammal ingesting soil organisms
Small insectivorous mammal ingesting soil organisms
Human ingesting ground water
Human ingesting ground water
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TABLE 2.5 - MEASUREMENT ENDPOINTS FOR CRITICAL PATHWAYS
Pollutant
Aluminum
Antimony
Barium
Beryllium
Boron
Fluoride
Manganese
Thallium
Tin
Titanium
Pathway Number
6
7
10
. 14
7
10
14 •
3
14
.6
6
10
3
4
6
7
10
14
3
7
6
Endpoint/Species '
TRV/rat
TRV/rat
TRV/rat
RfD/rat
TRV/rat
TRV/rat -
RfD/human
CRL
CRL
. TRV/dog
TRV/mice
TRV/mice
RfD/human
RfD/human
TRV/rat
TRV/rat
TRV/rat
RfD/human
RfD/rat
TRV/rat
TRV/mice
1 CRL - cancer risk level
RfD - risk reference dose
TRV - toxicological reference value
10
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TABLE 2.6 - RESULTS OF RISK ASSESSMENT FOR CRITICAL PATHWAYS
Pollutant
Aluminum
Antimony
Barium
Beryllium1
Boron
Fluoride
Manganese
Thallium
Tin
Titanium
Pathway Number
6(ag,i>,p)
7(ag,f,r)
10(f,p)
14(ag,f,r,p)
7(ag,f,r)
10(ag,f,r,p)
14(ag,f,r,p)
14(ag,f,r,p)
6(f,p)
6(ag,f,r,p)
10(ag,f,r,p)
3(ag,f,r,p)
4(f,r) .
6(ag,f,r,p)
7(ag,f,r)
10(f,p)
14(ag,f,r,p)
t
3(ag,i>,p)
7(ag,f,r)
6(ag,r)
RfD Ratio1
'
20(ag),40(f),3(r),60(p)
9(ag),20(f),l(r), 20(p)
7xlO-4(ag),9xlO-4(f)
SxlO-VXlxlO-fa)
'
-
4(ag,p),3(f,r) '
10(f), 40(r)
700(ag),1000(f),
30(r),2000(p)
2(ag,p),l(f,r)
-
•
RQ2
80(f,p),
100(ag,r)
l(ag,f,r)
3(f,p)
40(ag,f,r)
10(ag,r), 50(f,p)
_
4(f,p)
10(ag,r),30(f,p)
5(ag,r),8(f,p)
200(ag,r),800(f,p)
l(ag,f,r)
2(f,p)
-
2(ag,f,r)
7(ag,r)
'Ratio of estimated exposure to Reference Dose (RfD). For beryllium, the value is a carcinogenic
risk level.
2Risk Quotient - ratio of estimated exposure to Toxicological Reference Value (TRY).
ag - agricultural land; f - forest land; r - reclamation site; p - public contact site
11
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2.2.2 Surface Disposal
During the Comprehensive Hazard Identification study for surface disposal, soil-water
partition coefficients for sand with a porewater pH of 5 were used for the ground-water pathway.
This is the same conservative assumption that was used in the groundwater pathway analyses for
land application.
Results of the Comprehensive Hazard Identification study for the critical surface disposal
pathways are presented in Table 2.7.
TABLE 2.7 - RESULTS FOR CRITICAL SURFACE DISPOSAL PATHWAYS
Pollutant .
Antimony
Barium
Beryllium
Manganese
Pathway
Ground water
Ground water
Ground water
Ground water
Cancer Risk Level
-
,
2 x 10-4
.
•RfD Ratio1
• 4
1
-
90
*
'.Ratio of estimated exposure to the Reference Dose (RfD).
2.3 RATIONALES FOR EXCLUDING INORGANIC POLLUTANTS FROM FURTHER
CONSIDERATION
The rationales for excluding inorganic pollutants from the list of pollutants for the Round
Two sewage sludge regulation for land application and ^surface disposal are presented below.
12
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2.3.1 Land Application ^^
Aluminum
The critical pathway for aluminum for land application is Pathway 6 (animal foraging)
for both agricultural land and non-agricultural land (forest, reclamation sites, and public contact
sites). As indicated in Table 2.3, the uptake slopes used in the Pathway 6 analyses were obtained
from.non-sewage sludge studies (i.e., crops from which the uptake slopes were obtained-were not
grown in sewage sludge-amended soil). EPA concluded it is not appropriate to use those uptake
slopes to estimate the uptake of aluminum into forage grown in sewage sludge-amended soils (see
Policy Decision on page 2). No other information was available on uptake slopes for aluminum.
Because aluminum is not a bioaccumulative pollutant (i;e., does not accumulate in human
or animal tissue); because Pathway 6 was the only critical pathway for aluminum from the
Comprehensive Hazard Identification study; and because after the detailed review of Pathway 6,
it could not be evaluated using available information, EPA concluded that aluminum should not
be on the list of pollutants for the Round Two regulation for land application.
Antimony
One of the critical pathways for antimony for land application is Pathway 7 (grazing
animal that ingests sewage sludge directly). As indicated in Tables 2.3 and 2.5, the measurement
endpoint (i.e., the lexicological reference value (TRY)) for this pathway for both agricultural and
non-agricultural land is based on results of studies using laboratory animals (i.e., rats). This
endpoint was extrapolated to the appropriate HEI (i.e., herbivorous animals) for the land
application risk assessments.
13
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The lowest observed adverse effect level (LOAEL) for antimony is 0.262 mg/kg-body
weight/day, which is based on the results of a study in which antimony was fed to rats in water
(Schroeder et al., 1970). This value was converted to a dietary value (i.e., 3.4 mg/kg-food) using
a standard body weight of 0.4 kilograms for a rat and allometric equations (U.S. EPA, 1988).
A decrease in survival and longevity for male and female rats was observed at this dose
equivalent. The dietary value was divided by 10 to obtain the TRY for antimony.
There are two reasons why it is not appropriate to use the TRV for laboratory animals as
the TRV for the HEI in the Pathway 7 exposure analyses for agricultural land, forests and
reclamation sites. First, the study on which the LOAEL for antimony was based (Schroeder,
1970) indicates that the effect from exposure to antimony (a decrease.in survival and longevity)
occurs later in the life of a rat and growth was not affected. Thus, the potential for antimony to
interfere with growth and reproduction (i.e., population effects) is unclear, Also, results of
another study (Schroeder et al., 1968a) indicate a decrease in survival and longevity due to
exposure to antimony was not observed in mice.
Second, the LOAEL on which the TRV is based was obtained from a study in which
.antimony was fed to rats in water. Gastrointestinal absorption of antimony in food is expected
to be lower than the gastrointestinal absorption of antimony in drinking water. For example,
results of other rat studies (Sunagawa, 1981; Smyth and Thompson, 1945) in which antimony was
administered in food indicate that the no observed adverse effect level (NOAEL) for antimony
can be as high as 200 mg/kg-day and not cause specific systemic effects (e.g., changes in blood
pressure). This value, which did not result in population effects, is over two orders of magnitude
higher than the LOAEL used to develop the TRV for antimony.
14
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Because there is uncertainty in the animal studies about whether exposure to antimony
causes population effects and because the LOAEL used to develop the TRY is based on the
results of a study in which rats were fed antimony in water, EPA concluded that it is not
appropriate to use the TRY in the Comprehensive Hazard Identification study in the Pathway 7
analyses. In those analyses, the HEI ingests sewage sludge while gracing on sewage sludge-
amended soil. Other TRY values would likely be much higher based on other toxicity data. If
the TRY is based on a NOAEL of 200 mg/kg-day (i.e., the NOAEL. from rat studies in which
rats were fed antimony in food), the risk quotient for the Pathway 7 analyses would be less than
one. For these reasons, EPA concluded that antimony should not be on the Round Two list of
pollutants based on exposure through Pathway 7.
Pathway 10 (predator of soil organism) also was critical for antimony for land application.
EPA concluded that the TRY used in the Comprehensive Hazard Identification study is not
appropriate for this pathway for the same reasons the TRY for Pathway 7 is not appropriate.
1 Given that the RQ was 3 and that other TRY values would likely be much higher based on other
toxicity data, EPA concluded antimony should not be included on the final Round Two list of
pollutants for land application based on exposure through Pathway 10:
Pathway 14 (i.e., ground water) in the. land application Comprehensive Hazard
Identification study for agricultural land and non-agricultural land also was critical for antimony.
One way to evaluate the RfD ratio for this pathway (i.e., the highest'ratio is 60 for public contact
sites) is to consider the uncertainty factor for the RfD with respect to .the RfD ratio and the effect
upon which the RfD is based.
- The antimony RfD is based on an uncertainty factor of 1000 (IRIS, 1996). The highest
15
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RfD ratio for this pathway is 60 for public contact sites. In this case, the margin of safety (that
is, the ratio between the uncertainty factor and the RfD ratio) is approximately 17 (i.e., 1000/60).
EPA concluded that a margin of safety of 17 is sufficiently protective for the HEI (i.e., human)
in this case because .the effect upon which the RfD is based (i.e., changes in cholesterol and
glucose blood levels) is not severe and is likely reversible. EPA also concluded that the margins
of safety for the other types of land application sites (i.e., 50 for agricultural land, 25 for forest,
and 333 for reclamation sites) are protective for the HEIs for those types of land application sites.
The above information indicates that the critical pathways from the Comprehensive Hazard
Identification study should not be used as the basis for including antimony on the Hst of
pollutants for the Round Two sewage sludge regulation. For this reason, antimony was not
included on the list for land application.
Barium
One of the critical pathways for barium for land application was Pathway 7 (grazing
animal that ingests sewage sludge directly). The TRY for this pathway for agricultural land,
forest, and reclamation sites is based on results of studies using laboratory animals (i.e., rats).
This endpoint was extrapolated to the appropriate HEI (i.e., herbivorous animals) for the land
application risk assessments.'
Study results reported in the Agency for toxic Substances and Disease Registry (ATSDR,
1992a) were used as the basis for the TRY for barium. In those studies (Perry et al., 1983, 1985,
1989), barium was fed to rats in drinking water. The NOAEL for barium was 0.056 mg/kg- body
weight/day, which corresponds to a concentration in drinking water of 1 ppm. The dietary
16
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equivalent is 0.7 mg/kg-food when the NOAEL is converted using allometric equations (U!S.
EPA, 1988).
The RQ for barium for Pathway 7 was 40. .Even if the LOAEL were used as the basis
for the TRY, instead of the NOAEL, the RQ would be 4. This means that the estimated exposure
for Pathway 7 could cause the LOAEL to be exceeded.
The effect for the LOAEL for barium is an increase in systolic blood pressure. This effect
was not seen, however, until the eight month of a 16 month rat study. No other toxic effects
were observed in the study, and growth was not impaired. The impact of slight increases hi
systolic blood pressure for cattle, other grazing animals, and small mammals is unclear, and
population effects (i.e., growth, reproductive, and mortality) for those animals cannot be evaluated
using the results of the rat study.
A 1975 study found reduced life span in male mice given 5 ppm barium in drinking _
water (Schroeder and Mitchener, 1975). The calculated LOAEL for this study was 0.95 mg/kg- ^^
body weight/day, which has a dietary equivalent of 4.8 mg/kg-food when converted using an
allometric equation (U.S. EPA, 1988). During the study, longevity only was reduced slightly.
Other studies in which cardiovascular and other systemic effects from exposure to barium were
evaluated found NOAELs at an order of magnitude higher than in the NOAEL based on the
results of the Perry et al. studies.
EPA concluded that it is not appropriate to use the above TRY as the TRY for the HEI
in the Pathway 7 exposure analyses because the observed effects from exposure to barium, which
is a non-bioaccumulative pollutant, were not population effects. In addition, the effects tirat were
observed (i.e., increase hi systolic blood pressure) occurred as a result of exposure to barium in
17
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drinking water. The absorption of barium in drinking water is likely to be higher than absorption
of barium in food or in sewage sludge. For these reasons, EPA concluded that barium should
not be on the Round Two list of pollutants for land application based on exposure through
Pathway 7.
Pathway 10 (predator of soil organism) also was critical for barium hi the Comprehensive
Hazard Identification study for agricultural land and non-agricultural land. EPA concluded that
the TRY used in that study for Pathway 10 is not appropriate for the same reasons the TRY for
Pathway 7 is not appropriate. Therefore, EPA concluded barium should not be on the Round
Two list of pollutants based on exposure through Pathway 10.
Pathway 14 (i.e., ground water) also was critical for barium for agricultural and non-
agricultural land application. Two conservative assumptions were made for this pathway in the
Comprehensive Hazard Identification study. One was the type of soil at the land application sites
and the other was the value for the soil-water partition coefficient (KJ.
The type of soil affects the ability of a pollutant to move vertically to an aquifer and
laterally to a nearby well. Soil types in the unsaturated zone beneath a land application site in
order of increasing pollution potential are: (1) nonshrinking clay, (2) clay loam, (3) silty loam,
(4) loam, (5) sandy loam, (6) shrinking clay, (7) sand, (8) gravel, and (9) thin or absent soil (U.S.
EPA, 1992). EPA concluded that it is reasonable to assume a soil type of either sandy loam,
shrinking clay, or sand as the soil type for all land application sites. In the case of barium, the
assumed soil type for the land application sites was sand.
The Kd value for sand with a porewater pH of 5 varies from 6 liters per kilogram to 174
liters per kilogram (Gerritse et al., 1982). In the Comprehensive Hazard Identification study,
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Pathway 14 was critical for barium because the lower end of the Kj range (i.e., 6) was used to
estimate exposure from barium. If the upper end of the K,, range (i.e., 174) is used, Pathway 14
is not critical (i.e., the RfD ratio is less than one) for barium.
EPA concluded that because there is an acceptable range of partition coefficients, it is
appropriate to use the upper end of the range, particularly when the soil type for all land
application sites is assumed to be sand. Because Pathway 14 is not critical when the upper end
of the partition coefficient range is used, EPA concluded that barium should not be on the Round
Two list of pollutants for land application based on exposure through Pathway 14.
The above information indicates that after the detailed examination of the critical pathways
for barium (i.e., 7, 10, and 14) in the Comprehensive Hazard Identification study, none of the
pathways are critical for both agricultural land and non-agricultural land. For this reason, barium
was not included on the final list of pollutants for the Round Two regulation for land application.
Beryllium .
Pathway 14 was critical for beryllium for both agricultural and non-agricultural land
(forest, reclamation sites, and publication sites). As mentioned previously, the assumed soil type
and the partition coefficient are important for this pathway.
In the case of beryllium, the assumed soil type for all land application, sites is sand. This
is a reasonable assumption, particularly for agricultural land. Loam soils (sandy loam, silty loam,
silty clay loam) are predominant on agricultural land throughout the United States (sand and
sandy loams predominate in the southeast). Of the loam soils, sandy loam has the highest
pollution potential (U.S. EPA, 1992). ,
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During the Comprehensive Hazard Identification study, the partition coefficient at the
lower end of the range of partition coefficients for sand with a porewater pH of 5 was used.
EPA concluded that because a reasonable soil type was used, it is appropriate to use any of the
partition coefficients in the range of partition coefficients.
When the median partition coefficient value for sand with a porewater pH of 5 is used.
Pathway 14 is not critical for beryllium (i.e., the cancer risk level is lower than 10"4). For this
reason, beryllium was not included on the final list of pollutants for the Round Two sewage
sludge regulation for land application.
Boron
The critical pathway for boron for land application is Pathway 6 (animal foraging) for
forest and reclamation sites. None of the pathways for agricultural land were critical for boron".
The uptake slopes used hi the Pathway 6 analyses were obtained from non-sewage-sludge
studies (i.e., crops for which the uptake slopes were obtained were not grown in sewage sludge-
amended soil). EPA concluded that it is not appropriate to use those uptake slopes to estimate
risks from boron in crops grown in sewage sludge-amended soils (see Policy .Decision on page
2).
No other information is available on uptake slopes for boron. Because Pathway 6 was
the only 'critical pathway for boron and because this pathway could not be evaluated using
available information after the detailed examination of the critical pathways, EPA concluded that
boron should not be on the list of pollutants for the Round Two regulation for land application.
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Fluoride
Pathways 6 and 10 were critical for fluoride for both agricultural land and non-agricultural
land (i.e., forest, reclamation sites, and public contact sites) in the Comprehensive Hazard
Identification study. For Pathway 6 (animal foraging), "the. uptake slopes used in the analyses
were obtained from non-sewage-sludge studies (i.e., crops from which the uptake slopes were
obtained were not grown on sewage sludge-amended soils). EPA concluded it is not appropriate
to use those uptake slopes to estimate risks from fluoride in forage grown in sewage sludge-
amended soils (see Policy Decision on page 2).
No other information is available on uptake slopes for fluoride. Because Pathway 6 could
not be evaluated using existing information after completion of the detailed examination of the
critical pathways, EPA concluded that Pathway 6 is not critical. For this reason, fluoride v/as not
included on the list of pollutants for the Round Two regulation for land application based on
exposure through Pathway 6.
Pathway 10 (predator of soil organism) also was critical for fluoride for agricultural land
and non-agricultural land. The TRY for this pathway was based on a NOAEL of 10 mg/L in
drinking water administered to mice (Kanisawa and Schroeder, 1969). This was converted to a
dietary equivalent value of 11 mg/kg-food using allometric equations (U.S. EPA, 1988). Results
of other studies indicate that a dietary equivalent value for fluoride of 52 mg/kg-food resulted in
changes in teeth and liver, and structural and functional changes in the kidney (Jankauskas, 1974;
Lim et al., 1975; Roman et aL, 1977, as cited in IARC, 1982).
The HEI for Pathway 10 is the predator of a soil organism (e.g., a shrew). The effect
from the exposure in Pathway 10 is mild systemic changes (e.g., changes in teeth and liver).
21
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Population effects from this exposure are unknown.
Because the effect for which the TRY is protective is mild systemic changes and not
population effects, and because there is some evidence that fluoride is necessary for fertility in
mice (Messer et al. 1973, as cited in IARC, 1982), EPA concluded that the TRV used in the
Comprehensive Hazard Identification study was not appropriate. For this reason, and because no
other relevant toxicological information on small mammals was available for Pathway 10, EPA
concluded that Pathway 10 could not be evaluated for fluoride. Thus, Pathway 10 is not critical
for fluoride.
The above information indicates that the critical pathways from the Comprehensive Hazard
Identification study should not be the basis for including fluoride on the Round Two list of
pollutants. For this reason, fluoride was not placed on the list of pollutants for the Round Two
sewage sludge regulation for land application. -
Manganese .
Pathways 3, 6, 7, and 14 were critical for manganese for agricultural land. Pathways 3,
4, 6, 7, 10, and 14 were critical for manganese for non-agricultural land.
Pathway 3 is the child ingestion pathway. For agricultural land and public contact sites,
a child between the ages of 1 and 6 is assumed to ingest 0.2 gram of sewage sludge (not the
sewage sludge-soil mixture) daily. For forest and reclamation sites, a child between the ages of
4 and 6 is assumed to ingest 0.2 grams of sewage sludge daily.
The Reference Dose (RfD) for the Pathway 3 analyses in the Comprehensive Hazard
Identification study was 0.005 milligrams of manganese per kilogram of body weight per day.
j
22
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On May 1, 1996, the RfD for manganese in EPA's Integrated Risk Information System (IRIS)
was changed. The current RfD for manganese in IRIS is 0.14 milligrams of manganese per
kilogram of body weight per day for dietary exposure. As indicated in the Uncertainty and
Modifying Factors section in IRIS, when assessing exposure to manganese from food, a
modifying factor of one is used. When assessing exposure to manganese from drinking water
or soil, a modifying factor of three is used. Because the HEI ingests sewage sludge, which is
similar to soil, an uncertainty factor of three was applied to the RfD. The RfD for the Pathway
3 analyses should be 0.14 divided by 3, resulting in 0.05 milligrams of manganese per kilogram
of body weight per day.
Using the current RfD for manganese, the RfD ratio for Pathway 3 is 0.4 for agricultural
land and public contact sites, and 0.3 for forest and reclamation sites. Because these values are
less than one, Pathway 3 is not critical for manganese. For this reason, EPA concluded that
manganese should not be on the list of pollutants for the Round Two sewage sludge regulation
based on exposure through Pathway 3.
The uptake slopes in Pathway 4, which was criticaTfor forest and reclamation sites, were
obtained using non-sewage-sludge studies (i.e., crops from which the uptake slopes were obtained
were not grown in sewage sludge-amended soils). EPA concluded that it is not appropriate to
use those uptake slopes for crops grown in sewage sludge-amended soils (see Policy Decision on
page 2). Because there is no other information on manganese uptake slopes, manganese was not
included on the Round Two list of pollutants based on exposure through Pathway 4.
Pathway 6 also was critical for manganese for agricultural land, forests, reclamation sites,
and public contact sites. EPA concluded that manganese should not be included on the Round
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Two list of pollutants based on exposure through this pathway because the uptake slopes used in
the analyses were obtained from non-sewage-sludge studies. It is not appropriate to use those
uptake slopes for crops grown in sewage sludge-amended soils (see Policy Decision on page 2).
*. . •
The TRVs for Pathway 7, which was critical for manganese for agricultural land, forest,
and reclamation sites, and for Pathway 10, which was critical for forest and public contact sites,
are based on the results of studies using laboratory animals (i.e., rats). After reviewing the results
in the original study (Laskey et al., 1982) used to develop the TRY, an error was found in the
dietary value. The dietary value used to develop the TRY in the Comprehensive Hazard
Identification study was 170 mg/kg-food. This value was divided by 10 to determine the TRY.
The dietary value in the Laskey study was 350 mg/kg-food. Thus, the TRY should have
been 35 mg/kg-food instead of 17 mg/kg-food. When the revised TRY was used to calculate the
RQs. for Pathways 7 and 10, the RQ for Pathway 7 was 0.7 and the RQ for Pathway 10 was just
1. Therefore EPA concluded that manganese should not be included on the Round Two list of
pollutants, because the RQ became less than one for one pathway, and just met the level of
concern for the other pathway.
The final pathway that was critical for manganese is Pathway 14 (i,e., ground water).
This pathway was critical for both agricultural land and non-agricultural land (i.e., forest,
reclamation sites, and public contact sites).
Two of the important variables for^this pathway are soil type and partition coefficient.
As mentioned previously, EPA concluded that assuming a soil type of either sandy loam,
shrinking clay, or sand is conservative. During the detailed examination of the critical pathways,
the assumed soil type for Pathway 14 for manganese was sandy loam, not sand.
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The partition coefficient for sandy loam with a porewater pH of 8 ranges from 8418 liter
per kilogram to 15,774 liter per kilogram (Gerritse et al., 1982). Using any partition coefficient
within that range is conservative. For the Pathway 14 analysis for manganese, when a value in
the middle of the above range is used, Pathway 14 is not critical for manganese.
The above information indicates that after completion of the detailed examination of the
critical pathways for manganese from the Comprehensive Hazard Identification study, none of
the pathways are considered to be critical for agricultural land and non-agricultural land. For this
reason, manganese was not included on the final list of pollutants for the Round Two sewage
sludge regulation for land application.,
Thallium
The critical pathway for thallium for agricultural land, forest, reclamation sites, and public
contact sites was Pathway 3 - child ingestion of sewage sludge. In the Comprehensive Hazard
Identification study, the ratio of exposure from Pathway 3 to the RfD for thallium was two.
The thallium RfD is based on the results of a 90-day study during which rats ingested
soluble thallium salts in drinking water (IRIS, 1996). The uncertainty factor in the RfD is 3,000.
In the case of the Pathway 3 analysis, the margin of safety is 1,500 (i.e., 3,000. divided by an
RfD ratio of 2). • '
The absorption of metals like thallium in sewage sludge in the gastrointestinal tract, after
the sewage sludge is ingested by a child is expected to be lower than the absorption of soluble
salts of thallium. For this reason and because the margin of safety for the Pathway 3 analysis
is 1,5.00, EPA concluded that Pathway 3 was not critical for thallium. Thus, thallium was not
25
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included on the final list of Round Two pollutants' for land application for either agricultural land
or non-agricultural land.
*
Tin
The critical pathway for tin for agricultural land, forest, and public contact sites was
Pathway 7 (i.e., grazing animal that ingests sewage sludge directly). The TRY for tin was based
on the results of a study in which female rats were fed 5 ppm tin hi drinking water (Schroeder
et.al., 1968b). The observed effect in this study was decreased longevity.
The LOAEL reported in ATSDR (1992b) was 0.7 mg/kg/day, which is equivalent to a
dietary value of 9 mg/kg-food. This value was divided by 10 to obtain a TRY for Pathway 7 of
0.9 mg/kg-food. When reviewing the original study on which the TRY is based, an error was
found. The TRY should be 0.45 mg/kg-food, which means the RQ for tin for agricultural land,
forest, and reclamation sites should have been four instead of two.
' Studies other than the Schroeder et al. study (1968b) failed to find any effects in mice
administered 5 ppm tin in drinking water (Schroeder and Balassa, 1967). In addition, other
studies examining systemic effects in rats and mice found NOAELs an order of magnitude or
• more higher than the LOAEL from the Schroeder et al. study (1968b). Effects observed in these
studies are not clear with respect to population effects from exposure to tin.
Because the LOAEL used to calculate the TRY for tin is from a study in which rats were
administered tin in drinking water (absorption of tin in food or sewage- sludge is likely to be
lower than absorption of tin in drinking water); because results of other studies indicate that the
NOAEL for tin is higher than the LOAEL from the Schroeder et al. study (1968b); and because
26
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the population effects from exposure to tin are not known, EPA concluded that the TRY from
the rat study should not be used as the TRY for the HEI hi Pathway 7.
Because there is no other more appropriate information on the TRY for tin, Pathway 7
could not be evaluated for tin after completion of the detailed examination of the critical
pathways. For this reason, Pathway 7 is not critical for tin, and tin was not included on the
Round Two list of pollutants based on exposure through Pathway 7.
Titanium
The critical pathway for titanium for agricultural land arid reclamation sites was Pathway
6 (i.e., animal foraging on sewage sludge-amended soils). The uptake slopes used in the Pathway
6 analyses were obtained from non-sewage-sludge studies (i.e., crops from which the uptake
slopes were obtained were not grown in sewage sludge-amended soils). EPA concluded that it
is not appropriate to use uptake slopes from non-sewage-sludge studies for forage grown in
sewage sludge-amended soils (see Policy Decision on page 2). -
No other information is available on uptake slopes for titanium. Because Pathway 6 could
not be evaluated using available information, EPA concluded that Pathway 6 is not critical and
that titanium should not be on the list of pollutants for the Round Two sewage sludge regulation
for land application based on exposure through Pathway 6.
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2.3.2 Surface Disposal
Antimony and Barium
The critical pathway for antimony and barium for surface disposal is the groundwater
pathway. As mentioned in the above discussion of antimony for land application, one way to
evaluate the RfD ratio (i.e., four for antimony and one for barium) for this pathway is to consider
the uncertainty factor for the RfD with respect to the RfD ratio and the effect for which the RfD
is protective. ,
The antimony and barium RfDs are based on an uncertainty factor of 1000. The margin
of safety for a surface disposal site (i.e., surface impoundment) would be 250 (i.e., 1000 divided.
by 4) for antimony and 1000 (i.e., 1000 divided by one) for barium. EPA concluded that for
antimony a margin of safety of 250 is sufficiently protective for the HEI (i.e^, human) in this case"
because the effect upon which the RfD is based (i.e., changes in cholesterol and glucose blood
levels) is not severe and is likely reversible. EPA also concluded that barium just met the critical
pathway criteria. For these reasons, EPA concluded after completion of the detailed examination
of the critical pathways that antimony and barium should not be on the Round Two list of
pollutants for surface disposal based on exposure through the groundwater pathway.
Beryllium and Manganese i
The groundwater pathway also was the critical pathway for beryllium and manganese for
, surface disposal. As mentioned previously during the discussion of the groundwater pathway for
land application, two important parameters for the groundwater pathway are soil type and soil-
water partition coefficient. .
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During the Comprehensive Hazard Identification study, the soil type for all surface
disposal sites was assumed to be sand. EPA concluded that using a soil type of eithe:r sandy
loam, shrinking clay, or sand is conservative.
A soil-water partition coefficient for sandy soil with a porewater pH of 5 was used in the
Comprehensive Hazard Identification study for surface disposal. However, if the median partition
coefficient for sandy loam with a porewater pH of 8 is used in the analysis, the groundwater
pathway is no longer critical for beryllium and manganese for. surface disposal.
EPA concluded that it is reasonable to use the sandy loam soil type in the surface disposal
groundwater analysis. It is also reasonable to use the median value for partition coefficient in
the range of partition coefficients for sandy loam soil in the analysis. When this value is used,
the groundwater pathway is not critical for beryllium and manganese for surface disposal. For
this reason, EPA concluded that those pollutants should not be on the, final list of pollutants for
the Round Two regulation for surface disposal based on exposure through the groundwater
pathway. •
2.3.3 Incineration
Results of the Comprehensive Hazard Identification study indicate that no pollutants
warrant consideration for the "list of pollutants for the Part 503 Round Two regulation for
incineration. Dioxins/furans will be re-evaluated for the Part 503 use or disposal practices,
including incineration, at the completion of EPA's dioxin reassessment.
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2.4 POLLUTANTS RECOMMENDED BY OTHERS FOR THE ROUND TWO LIST
OF POLLUTANTS
Prior to conducting the Comprehensive Hazard Identification study for the 31 candidate
pollutants for the Round Two list of pollutants, EPA programs and experts from outside of EPA
were contacted to obtain data (e.g., plant and animal uptake data) on the 31 candidate pollutants.
Comments were received from Dr. George O'Connor from the University of Florida and Dr.
Rufus Chaney from the U. S. Department of Agriculture (see Appendix D3).
Dr. O'Connor provided references on plant bioavailability for some of the candidate
organic pollutants. Information from those references was used in the Comprehensive Hazard
\ •
Identification study, where applicable. •
Dr. Chaney also provided information on several of the candidate pollutants. He
recommended that beryllium, boron, dioxins/furans, coplanar polychlorinated biphenyls, cobalt,
fluoride, and iron be on the Round Two list of pollutants for land application.
With the exception of cobalt and iron, the pollutants that Dr. Chaney recommended for
the Round Two list of pollutants for land application were evaluated in the Comprehensive
Hazard Identification study. The results of the detailed examination of the critical pathways.for
those pollutants are presented in other sections of the Technical Support Document (U.S. EPA,
1996).
Both cobalt and iron were evaluated for the list of pollutants for the Part 503 Round One
regulation for land application. Neither pollutant was include on the Round One list of
pollutants.
Cobalt was not included on the Round One list of pollutants because the hazard index
(estimated exposure divided by the reference dose) was less than one. Dr. Chaney stated that '
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results of cobalt feeding trials indicate that a cobalt concentration between 5 and .10 milligrams
per gram of diet may be injurious to sheep and cattle. Cobalt was detected in nine percent of the
samples from the National Sewage Sludge Survey. Using the 98th percentile concentration for
cobalt from the NSSS (i.e., 104 mg/kg with non-detected values set equal to the minimum level)
and the fraction of the animal's diet that is sewage sludge used in the Round One risk
assessments (i.e., 1.5 percent), the 5-10 milligram per kilogram diet concentration for cobalt is
not expected to be reached in an animal's diet from ingestion of sewage sludge. In addition,
none of the updated information submitted by Dr. Chancy suggests that the original hazard index
for cobalt would change. For these reasons, EPA concluded that cobalt should not be on the list
of pollutants for the Part 503 Round Two regulation for land application.
Iron was not included on the Round One list of pollutants even though the hazard index
for grazing animals that ingest the sewage sludge/soil mixture (i.e., Pathway 7) was 2.1. TKe
rationale for not including iron on the Round One list was that the gracing animal index was
based on a worst worst-case sewage sludge iron concentration and the assumption that five
percent of the animal's diet is Sewage sludge. If sewage sludge with a "typical" iron
concentration (i.e., 28,000 mg/kg (U.S. EPA, 1985)) is used in the analysis, the hazard index for
grazing animals is less than one. The hazard index for iron also is expected to be less man one
if the fraction of the animal's diet from the risk assessment for the Round One regulation (i.e.,
1.5 percent) and the 90th percentile concentration for iron from the NSSS (i.e., 41,800 mg/kg)
are used to develop the index. For these reasons, EPA concluded that iron should not be on the
list of pollutants for the Part 503 Round Two regulation for land application.
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3. LIST OF POLLUTANTS FOR THE ROUND TWO
SEWAGE SLUDGE REGULATION
On November 30, 1995, EPA submitted the list of pollutants for the Round Two sewage
sludge regulation to the District Court in Oregon. The court notice is presented hi Appendix D2.
After considering information from the Comprehensive Hazard Identification study; the
rationales for deleting inorganic pollutants from the list of pollutants that warranted further
consideration; and information received from others, EPA concluded that two pollutants should
be on the list for each use or disposal practice. They are: dioxins/furans (all monochloro to
octachloro congeners) and polychlorinated biphenyls (coplanar). The court notice indicates that
EPA may, in the exercise of its discretion, determine to add or delete other pollutants to or from
this list at the time the Round Two regulation is proposed.
In addition to the list of pollutants submitted to the court, EPA may change a limit for
the pollutants in the Round One regulation during development of the Round Two regulation.
For this reason, the Round One pollutants also are considered pollutants for the Round Two
sewage sludge regulation. . :
Including the pollutants from the Round One regulation, the list of pollutants for the
Round Two sewage sludge regulation by use or disposal practice is:
Land application
arsenic, cadmium, copper, lead, mercury, molybdenum, nickel, selenium, zinc,
dioxins/furans, and coplanar polychlorinated biphenyls.
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Surface disposal
arsenic, chromium, nickel, dioxins/furans, and coplanar polychlorinated biphenyls
Sewage sludge incineration
arsenic, beryllium, cadmium, chromium, lead, mercury, nickel, total hydrocarbons (or
carbon monoxide), dioxins/furans, and coplanar polychlorinated biphenyls
• T
Dioxins/furans were included on the list of pollutants for sewage sludge incineration even
though results of the screening risk assessments indicate that no pollutant warrants consideration
for the Round Two list of pollutants for incineration. EPA currently is conducting a reassessment
of dioxins/furans. Because the results of this assessment are unknown, dioxins/furans were
included on the Round Two list of pollutants for all use or disposal practices. At the completion
of the dioxin reassessment, EPA may decide not to regulate dioxins/furans for a particular .use
or disposal practice or may decide to regulate dioxins/furans on an accelerated schedule.
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4. REFERENCES
Agency for Toxic Substances and Disease Registry. 1992a. lexicological Profile for Barium
and Compounds. Prepared by Clement International Corporation under contract no. 205-
88-0608. U:S. Public Health Service. ATSDR/TP-91/03.
Agency for Toxic Substances and Disease Registry. 1992b. Toxicological Profile for Tin and
Compounds. Prepared by Life Systems under subcontract to Clement International
Corporation under contract no. 205-88-0608. U.S. Public Health Service. ATSDR/TP-
91/27.
Corey, R.B., L.D. King, C. Lue-Hing, D.S. Fanning, J.J. Street, and J.M. Walker. 1987. Effects
of Sludge Properties on Accumulation of Trace Elements by Crops. In: Land Application
of Sludge. A.L. Page, T.J. Logan, and J.A. Ryan. Lewis Publishers, Inc. Chelsea, MI.
Gerritse, R.G., R. Vriesema, J.W. Dalenberg, and H.P. De Roos. 1982. Effect of Sewage Sludge
on Trace Element Mobility in Soils. Journal of Environmental Quality. ll(3):359-364.
IARC (International Agency for Research on Cancer). 1982. IARC Monographs on the
Evaluation of the Carcinogenic Risk of Chemicals to Humans. Some Aromatic Amines,
Anthraquinones and Nitroso Compounds, and Inorganic Fluorides Used in Drinking-water
and Dental Preparations. Vol. 27.
IRIS. 1996. Integrated Risk Information System. June.
Jankauskas, J. 1974. Effects of Fluoride on the Kidney (A Review). Fluoride. 7-93-105 [As
cited in IARC, 1982].
Kanisawa, M. and H.A. Schroeder. 1969. Life Term Studies on the Effect of Trace Elements
on Spontaneous Tumors in Mice and Rats. Cancer Res. 29:892-895.
Laskey^ J.W., G.L. Rehnberg, J.F. Hein, and S.D. Carter. 1982. Effects of Chronic Manganese
(Mn3O4) Exposure on Selected Reproductive Parameters in Rats. J. Toxicol. Environ
Health. 9:677-687.
Lim. J.K.J., G.K. Jensen, and O.K. King, Jr. 1975. Some Toxicological Aspects of Stannous
Fluoride After Ingestion as a Clear, Precipitate Free Solution Compared to Sodium
Fluoride. J. Dent. Res. 54:615-625. [As cited in I ARC, 1982].
Messer, H.H., W.D. Armstrong, and L. Singer. 1973., Influence of Fluoride Intake on
Reproduction hi Mice. J. Nutr. 103:1319-1327. [As cited in IARC, 1982].
Perry, H.M., Jr., SJ. Kopp, M.W. Erlanger, and E.F. Perry. 1983. Cardiovascular Effects of
'34
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Chronic Barium Ingestion. Trace Subst. Environ. Health. 17:155-164.
Perry, H.M., Jr., E.F. Perry, M.W. Erlanger, .and SJ. Kopp. 1985. Barium-Induced
Hypertension. In: Advances in Modern Environmental Toxicology, Vol. IX, Inorganics
in Drinking Water and Cardiovascular Disease. EJ. Calabrese, R.W. TuthilL, and L.
Condie, eds. Princeton Scientific Publishing Co., Inc. Princeton, N.J. pp. 221-229.
Perry, H.M., Jr., SJ. Kopp, E.F. Perry, and M.W. Erlanger. 1989. Hypertension and Associated
Cardiovascular Abnormalities Induced by Chronic Barium Feeding. J. Toxicol Environ
Health. 28:373-388.
Roman, R.J., J.R. Carter, W.C. North, and M.L. Kauker. 1977. Renal Tubular Site of Action
of Fluoride in Fischer 344 Rats. Anesthesiology. 46:260-264. [As cited in I ARC, 1982].
Schroeder, H.A. and J.J. Balassa. 1967. Arsenic, Germanium, Tin and Vanadium in Mice:
. Effects on Growth, Survival and Tissue Levels. J. Nutr. 92:245-252.
Schroeder, H.A., M. Mitchener, J.J. Balassa, M. Kanisawa, and A.P. Nason. 1968a. Zirconium,
Niobium, Antimony and Fluorine in Mice: Effects on Growth, Survival'and Tissue Levels'
'J. Nutr. 95:95-101. -
Schroeder, H.A., M. Kanisawa, D.V. Frost, and M. Mitchener. 1968b. Germanium, Tin and
Arsenic in Rats: Effects on Growth, Survival, Pathological Lesions and Life Span J
Nutr. 96:37-45.
Schroeder, H.A., M. Mitchener, and A.P. Nason. 1970. Zirconium, Niobium, Antimony,
Vanadium and Lead in,Rats: Life Term Studies. J. Nutr. 100:59-68.
Schroeder, H.A. and M. Mitchener. 1975. Life-term Effects of Mercury, Methyl Mercury, and
Nine Other Trace Metals on Mice. J. Nutr. 105:452-458.
•Smyth, H.F., Jr. and W.L. Thompson,. 1945. The Single Dose and Subacute-Toxiciry of
Antimony Oxide (Sb2O3). Melon Institute of Industrial Research, University of Pittsburgh
OTS 206062. [As cited in ATSDR, 1992bj.
Sunagawa, S. 1981. Experimental Studies on Antimony Poisoning. Igaku kenkyu 51-129-142
[As cited in ATSDR, 1992b].
U.S. EPA. 1985. Environmental Profiles and Hazard Indices for Constituents of Municipal
Sludge: iron. Office of Water, Regulations and Standards. June.
U.S. EPA. 4988:. Recommendations for and Documentation of Biological Values for Use in
. Risk Assessment. Environmental Criteria and Assessment Office, Office of Health and
Environmental Assessment, Office of Research and Development. EPA/600/6-87/008.
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February. . •
U.S. EPA. 1992. Technical Support Document for Land Application of Sewage Sludge.
Appendix J. Office of Water, Office of Science and Technology. EPA 822/R-93-001a.
November.
U.S. EPA. 1996. Technical Support Document for the Round Two Sewage Sludge Pollutants.
Health and Ecological Criteria Division, Office of Science and Technology Office of
Water. EPA-822-R-96-003. August.
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APPENDIX Dl
LIST OF 31 CANDIDATE POLLUTANTS FOR THE
ROUND TWO SEWAGE SLUDGE REGULATION
SUBMITTED TO THE DISTRICT COURT IN OREGON
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v . IN THE UNITED STATES DISTRICT COURT
FOR THE DISTRICT 0? OREGON
FRANK GEARHART,, CITIZENS INTERESTED )' "'" -V ": ^T
IN BULL. RUN, INC., An Oregon )
Corporation, KATHY WILLIAMS, AND )• ; ' ..
PRANCES PRICE COOK, (•:.••'-.
I '•'••. :i
Plaintiffs, ) '
NATURAL RESOURCES DEFENSE COUNCIL, )
Intervenor Plaintiffs, )
ASSOCIATION OF METROPOLITAN SEWERAGE I ClVl1 Np« 89-fi266-HO
AGENCIES, j
Intervenor Plaintiffs, )
v.
CAROL M. BROWNER ]
Administrator, United States \
Environmental Protection Agency, )
Defendant. j • .
NOTICE OF POLLUTANTS
Pursuant to paragraph 2. of the Consent Decree entered in this
proceeding on September 5, 1990, as modified by this Court's
September 14, 1993 order, the U.S. Environmental Protection Agency
("EPA") hereby gives notice that, based on available information
reviewed to date, EPA presently intends to propose for regulation
under section 4OS(d)<2J{B)(i) of the Clean Water Act, 33 U.S.C. §
1345(d)(2)(B)(i), the following pollutants:1
Acetic acid (2, 4, -dichlorophenoxy), aluminum, antimony,
asbestos, barium, beryllium, boron, butanone (2-), carbon
^? information available at the time of proposal, EPA
« + discretion to either add or delete pollutant^ from the
of those that it currently intends to propose for regulation^
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cyanides
fluoride,.
phenol.
biphenyls
, tin, titani™, tolu«n«,
tarichlorophenoxyacetic iciii fa 4 s-\
C2' 4' 5 }'
acid (C2 - C2/4/ Sr)J/
^
o '
Respectfully
E. FLINT
Acting Assistant Attorney General
v
—r—*- »*«vitt. of Justice
„ :-.- Pennsylvania Ave., N.W
Washington, O.c. 20530
(202) S14-3785
/v,
RJCHARO T. WITT, Attorney
ri0 Counsel (LE-132W1
401 M Street, s.W.
Washington, D.C. 20460
(202) 260-7715
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JACK C. WONG - Bar No. 67138
United States Attorney
RAY - Bar HO. 72319~~
itant United States Attorney
District of Oregon
701 High Street .
Eugene, Oregon 97401
(503) 465-6771
OF COUNSEL:
GERALD H. .YAKADA
Acting General Counsel
DAVID M. GRAVALLESE
Assistant General Counsel
U.S. Environmental Protection Agency
Dated: May 21, IS93
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APPENDIX D2
FINAL LIST OF POLLUTANTS FOR THE
ROUND TWO SEWAGE SLUDGE REGULATION
SUBMITTED TO THE DISTRICT COURT IN OREGON
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LOIS J. SCHIFFER
Assistant Attorney General
Environment and Natural Resources
Division
MARK A. NITCZYNSKI, Attorney
Environment and Natural Resources
Division
RICHARD T. WITT, Attorney
Office of General Counsel (LE-132W)
U.S. Environmental Protection Agency
JACK C. WONG - Bar No. 67138
United States Attorney.
JOHN C. RAY - Bar No. 72319
Assistant United States Attorney
District of Oregon
701 High Street
Eugene, Oregon 97401 ~
(503) 465-6771
95HQV29
• ERK. U.S. DISTRICT COURT
DISTRICT OF OREGON
EUGENE. OREGON
IN THE UNITED STATES DISTRICT COURT
FOR THE DISTRICT OP OREGON
FRANK GEARHART, CITIZENS INTERESTED
IN BULL RUN, INC., An Oregon
Corporation, KATHY WILLIAMS, AND
FRANCES PRICE COOK,
Plaintiffs,
NATURAL RESOURCES DEFENSE COUNCIL,
INC.,
Intervenor Plaintiffs,
ASSOCIATION OF METROPOLITAN SEWERAGE
AGENCIES, • ,
Intervenor Plaintiffs,
v.
CAROL M. BROWNER
Administrator, united States
Environmental Protection Agency,
Defendant.
Civil No. 89-6266-HO
REVISED NOTICE OF
POLLUTANTS
Revised Notice of Pollutants -
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On May 24, 1993, pursuant to Paragraph 2 of the Consent
Decree entered in -this proceeding on September 5, 19 9O, as
subsequently.modified by this Court's orders, the U.s.
Environmental Protection Agency ("EPA") submitted a Notice of
Pollutants ("Notice"). The Notice stated that the Agency was
considering proposing 31 pollutants for regulation under section
405(d)(2)(B)(i) of the Clean Water Act, 33 U.S.C. §
1345(d) (2) (B) (i)'v Paragraph 9d of the Consent Decree provides
that the Agency may revise this list of pollutants if it
concludes that regulations are not needed for some or all -of the
31 pollutants. Based on current information, EPA has concluded
that 29 of the listed pollutants need not be regulatedj
acetic acid (2, 4, -dichlorophenoxy), aluminum, antimony,
asbestos, barium, beryllium, boron, butanone (2-), carbon
disulfide, cresol (p-), cyanides (soluble salts and complexes),
endsulfan-II, fluoride, manganese, methylene chloride, nitrate,
nitrite, pentachloronitrobenzene, phenol, phthalate (bis-2-
ethylhexyi), propanone (2-), silver, thallium, tin, titanium,
toluene, trichlorophenoxyacetic acid (2, 4, 5-),
trichlorophenoxypropionic acid ([2 - (2,4, 5-)], and vanadium.
Thus, EPA has concluded that only two of the listed
pollutants warrant further consideration for regulation:
dioxins/dibenaofurans (all monochloro to octochloro congeners)
and polychlorinated biphenyls (co-planar). EPA may, in the
exercise of its discretion, determine to add or dele-te other
pollutants from this list at the time of proposal.
Revised Notice of Pollutants - 2
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Dated: November 28, 1995
Respectfully submitted,
LOIS J. SCHIFPER
Assistant Attorney General
Environment and Natural Resources
Division
MARK A. NITCZYkSKI, Attorney
Environment and Natural Resources
. Division
U.S. Department of Justice
10th & Pennsylvania Ave. , N,w.
Washington/ D.C. 20530
(202) 514-3785
\
J~
RICHARD T. WITT, Attorney
Office of General Counsel (LE-132W)
U.S. Environmental Protection
Agency
401 M Street, S.W.
Washington, D.C. 20460
(202) 260-7715
JACK C. WONG - Bar No. 6713 S
United States Attorney
JOHN C, RAY - Bar No. 72319
Assistant United States Attorney
District of Oregon
701 High Street
Eugene, Oregon 97401
(503) 465-6771
Revised Notice of Pollutants - 3
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CERTIFICATE OP SERVICE
1 hereby certify that pn this November 28, 19? 5 I caused a
copy of the foregoing Revised Notice of Pollutants to be served
by first class mail, postage prepaid, on the following counsel:
WILLIAM CARPENTER
474 Willamette
Suite 303
Eugene, OR 97401
Counsel for Plaintiffs
JESSICA IANDMAN • .
Natural Resources Defense Council, Inc.
1350 New York Ave. , N.W.
Suite 300
Washington , DC .20005
Counsel for Natural Resources Defense Council, Inc. -
LEE WHITE '
122S I Street, N.W. , Suite 300
Washington, DC 20005
counsel for Association of Metropolitan Sewerage Agencies
&
Annette Bucco
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APPENDIX D3
RESPONSES TO REQUESTS FOR DATA
ON THE ROUND TWO CANDIDATE POLLUTANTS
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United Sratej
Department of
Aoricuiture
Agricultural.
Raeaarch
Seivica
Beitsviiie Area
Baftsvilfa. Agricultural
Rasiarch Canter
BeltSvUla. Maryland
20705
May 10, 1995
SUBJECT: Round 2 contaminants.
TO: Alan B. Hals, Chief, Multimedia Risk Assessment Branch.
Yogi Pstel, Multimedia Risk Assessment Branch.
FROM: R.L. Chaney, USDA-ARS, Environmental Chemistry Lab,
I am responding to your letter of April 18, 1995 requesting information on
plant uptake of these compounds or metals. 1 have written about the risks of
most of these metals, and some of the organics over the last 10 years. I
have huge amounts of literature on these elements, and several you appear
to left our of consideration. Where uptake by plants is known to occur to
any significant level from sludge-amended soils, these lesser-studied
elements have often been examined by pot and field studies of Dr. Don Usk
and his collaborator* (including me); they examined the Sludges, soils, plants,
and animal tissues using neutron activation (and atomic absorption or ICP) to
analyze over 40 elements in numerous experiments.
I would hope that demonstrated Iron toxlclty to cattle and horses from high
Fe sludges would put Fe on the list. Similarly, Co is a significant possibility
based on food-chain Injury to cattle and sheep. Fluoride Is also a
demonstrated risk from sludges, although mostly in the livestock grazing on
surface-applied sludges. 1 fiad brought up these omissions in Round 1, so I
am a little surprised that Fe and Co were not on the list. Even more
surprised when Tl, Sn, and some of the others on your preliminary list were
being considered when papers I have given EPA clearly show the lack of risk
under any route of exposure to sludges. It would seem to me that your list
partially came from the Water people, and they base their concern on toxlclty
of water soluble salts in distilled water, or even on injected water soluble
salts (e.g., Ag, TI, Sn. etc.).
If there is a message to this letter, it is my concern about the need to have
iron and cobalt on the thorough evaluation list. Comments below will
provide a summary of the literature related to Pathway Analysis of Risk, and
useful references.
If you want to reach me regarding these comments on the Round 2 List of
Contaminants, I will bo at my lab (301-504-8324) May 10 and 11, leaving
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for England In mid-afternoon. I will return the evening "of May 18, and be in
the iab on May 19.
Aluminum: AI is severely phytotoxic to plants when soils remain at pH lower
than about 5.2 for a number of years. Clays dissolve and AJ3+ enters the
cation exchange complex in soil. The water soluble AI3+ injures root
Initials, reduces root grown and reduces yield. Toxic AI in subsoils
prevents plants from using water stored in subsoils. AI phytotoxicity Is a
common problem on agricultural and forest land. Addition of inorganic AI
salts would allow development of AI phytotoxiclty soon after acidification
since precipitated AI(OH)3 !s present when the soil pH is over 5.2-5.5.
Little AI Is absorbed and translocated into plant shoots, and even less
into fruits and grains. Most plant AI is soil contamination from "wind-
blown dust in the field. Soil AI has lower bioavailability than do soluble
salts of AI. Other than phytotoxicity, we know of no Pathway in which
sludge-borne AJ in soils will cause risk compared to unsludged soils. AI
should be deleted from the list.
Antimony: In the 1970s and 1980s. Dr. Lisk and his collaborators used
neutral activation to measure many elements in plants, sludges, and
soils.. In pot and field experiments. There were some limitations in these
studies. However, the results with antimony were useful to your need.
The normal chemical form of antimony (Sb3*) in soils is quite Insoluble at
normal soil pH levels. Plant leaves, fruits, or grains had unchanged Sb
concentration even when soil Sb was significantly increased by applied
sludges; and animals did not accumulate Sb from sludge grown crops of
Chaney et al. (1978). Sb has little toxicity to animals or plants. It is
used some medications. 1 believe Sb should be deleted from your list.
Barium: in normal soils, which have adequate amounts of Ca and Mg even
when sludges are utilized on land, Ba is an exchangeable cation which is
pretty insoluble when sulfate is at the levels in soil required to produce
high yielding crops. Plant shoots have little response to added sludge
Ba, again from the data of Lisk et al. (including the Chaney et al., 1978b)
paper on chard fed to Guinea pigs) show no risk of injury or residue
transfer to livestock or wildlife. Barium occurs at unusual levels in a few
crop species, including Brazil nut, but Lisk and other researchers have
not shown significant increase In crop Ba on sludge-amended soils.
Beryllium: Added to soils as a soluble salt. Be has low phytoavailability. Lisk
found little evidence that sludge Be moved Into plants. And no evidence
that Be accumulated in, animal tissues when sludge grown crops were
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fed to test animals. Be may require full evaluation because of known
possible uptake and important industrial toxicology information.
However, only Lisk may have measured Be in sludge research studies,
and I'm not sure even he did. My comments are based on basic studies
in which Be salts were added to soils for plant studies, and the NRC
(1980) book on livestock.
Boron: Boron is important in agriculture and the environment because It is
phytotoxlc. High water soluble B in soils is accumulated by most plants,
and they suffer phytotoxlclty at foliar B levels which are not high enough
to be toxic to livestock chronically fed the crops suffering B toxicity.
There Is reasonably good evidence that B is required by animals, and that
dietary B is generally low. I can perceive no risk except phytotoxlclty
from sludge B; Lisk et al. provided good evidence of lack of B toxicity or
. food-chain accumulation of boron.
Only a few studies of sludge or effluent use on cropland or forests has
shown B phytotoxicity. In one, a sensitive crop received spray-applied
effluent with over 1 mg B/L. in a sludge study, a sensitive crop suffered
B phytotoxicity when a sludge containing glass fiber wastes was land
applied. Slow dissolution of B from the glass fibers caused excessive B
uptake. More B tolerant crops would not have been expected to suffer
any effects of biosoiids-applied B in that study. I summarized sludge and
compost B data in the Chaney and Ryan (1993) paper from the Ohio
Composting Conference (see at end of reference section). The
appropriate analysis of sludge boron risk will require extraction of "hot
water soluble" boron. Based on substantial animal tolerance of B {NRC,
1980), only the phytotoxicity pathway will require risk assessment.
Fluoride: A few sludges contain very high levels of F, resulting from
computer chip manufacturing wastewaters {HF is used to leach Si from
. marked surfaces of the chip}, and from aluminum smelting processes.
One sludge containing about 5% F was studied by Davis, 1980. Ha
found this sludge could induce F phytotoxicity in ryegrass from soil
applied high-fluoride sludge. Generally, foliar exposure of plants to HF
causes high accumulation of F in the plants, which In turn poisons
livestock. It is widely shown that animals are at much greater fluoride
risk from sludge of soil ingestion than from plant uptake.
In the Denver sludge feeding studies (Klenhoiz et al. and Baxter et al.),
CaF (the solid phase F compound in sludges) could be dissolved in the
digestive system of cattle, and it could cause bones to become brittle
and teeth to break. Analysis of sludges, using some selected
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concentration below which no harm is expected to plants or livestock,
will provide the protection needed for humans, livestock, and wildlife.
Only highly contaminated soils will have phytoavailabte F.
Manganese: Few sludges contain high levels of Mn (> 1500 ppmDW), In
fact, the principle problem regarding sludges is the induction of Min
deficiency when lime-treated sludges are used on coastal plain soils
(historically depleted of total soil Mn, so they are more susceptible to '
limo-induced-Mn deficiency}. I reviewed Mn in the Chaney and Ryan
(1993a) paper at the Ohio Composting Conference.
We have been testing use of Mn amendments to sludges to prevent
induced-Mn deficiency from lime-treated sludges, and have found no
evidence of plant toxicity when limed sludge was enriched In Mn by
about 6,000 ppm. Al Rubin heard our seminar on May 3 at the Maryland
Department of the Environment. .
When high Mn soils are strongly acidified (pH £ 5.4), Mn24 accumulates
among the exchangeable cations, and can cause phytotoxicity to
sensitive crops. However, except for rare Mn hyperaccumulator species,
plants suffer phytotoxicity and leaves remain low in Mn such that they
do not comprise chronic toxicity risk to livestock or wildlife. Farmers are
forced to add limestone to raise soil pH to prevent Mn phytotoxicity in
strongly acidic high Mn soils. I believe that the added risk from
sludge-borne Mn is trivial.
Silver: Silver is toxic to animals when injected, but not when ingested with a
complete diet; AgCi precipitate is formed in the gut, and Ag is not toxic.
•When Ag is added to soils, it is strongly precipitated and adsorbed by the
soils. Plants accumulate only traces of Ag, and no evidence of plant
uptake which might comprise a chronic ingestion risk has been found.
Most environmental concern about Ag results from toxicologists testing
soluble Ag salts in purified waters. Never from sludge. Even when
sludge was fed to livestock, sludge Ag was not toxic nor accumulated.
Silver should be deleted from the list.
Thallium: Although Tl appears to comprise a risk to plants or the food-chain
from deposition of aerosols on plants, there is little evidence that sludge-
applied Tl is moved into edible plant tissues. Again, the studies of Lisk
et al. using neutron activation provide adequate evidence that sludge Tl
has not been found to comprise risk. Tl can be emitted from
incinerators, and cement manufacturers commonly emit Tl and cause
local enrichment of soils.
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Tin: Sn is normally Sn * in the soil environment, and very insoluble. Like TX
and Cr, Sn is a good label for non-absorbed soil in the diet. Sludges
seldom have really high levels of Sn> and no evidence of plant uptake of
Sn from sludge-amended soil has been reported. Lisk included Sn in hi*
studies by neutron activation. Actually, sludge Sn is not a risk to
livestock which ingest sludge, in strong contrast with sludge Fe and F.
Tin should be deleted from the list.
Titanium: Ti is usually Ti4* in soils, and is very insoluble as TIO2. Soil Tl is
not found Inside plants, only as soil or dust contamination on the plants.
Soil/Sludge Ti is so insoluble that it does not comprise risk even when
Ingested by livestock. Titanium should be deleted from the list.
Vanadium: In nutrient solutions, certain unstable V salts can be accumulated
by plants, and vanadate interferes with ion uptake by ATPase enzymes in
the roots. Little V is translocated to edible crop tissues. The Lisk work
usually showed that-V'was not accumulated by crops, nor in animal
tissues. Vanadium should likely be deleted from the list.
Iron: I am a little concerned that no one in your team chose to enter Fe
{iron} or cobalt (Co) into the Round 2 review. In 1976-1979, a
cooperative study in Maryland allowed us to characterize Fe toxicity to
cattle fed high Fe U1%) and low Cu sludges on pastures. When a
sludge or compost with only about 4% Fe was surface-applied on
pastures or added to feeds in a feeding study with cattle, they did not
cause the Fe toxicity, but some accumulation of Fe in the spleen, liver,
and duodenum was observed. Several other controlled feeding studies in
the US did not find evidence of Fe toxicity from ingested sludges with 1-
2% Fe, and seldom found Fe accumulation is tissues. The usual action
of excessive Fe Intake is to induce chronic Cu deficiency which causes
joint disease. Because Fe has poisoned livestock in several sludge
experiments, and if high Fe sludges are found by monitoring, the sludge
can be required to be injected or incorporated rather than left on the
surface, avoidance of sludge Fe risk is comparatively easy. When the
ferrous Fe in anaerobic sludges becomes oxidized in the soil, or during
composting, the ferric Fe has much lower solubility or toxicity to cattle.
So the method of sludge processing and the concentration of Fe In the
final product are important in prediction of animal risk. Humans seldom
ingest sludges which are freshly anaerobic, and no evidence of human
risk from sludge Fe has been identified.
In the Oklahoma miniature horses case, the horses were alleged to have
suffered Fe toxicity, but the soil appears to have been the major source
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of the Fe exposure. Historic observations of induced Cu deficiency on
lateritic or other high Fe soils has been reported in cattle from many
locations. Another ease in Virginia may have comprised Fe poisoning,
but the details of the source of excess soluble Fe remain unclear. Ona
common symptom of Fe toxicity is red coloration of the duodenum from
ferritln accumulation. Tissues (liver; kidney, spleen, blood/serum) have
increased Fe concentrations when higher Fe sludges are Ingested by
livestock.
Cobalt: Because sludges normally do not contain high Co concentrations
without unusual industrial discharge, no Co problems have been
observed En sludge research. However, my analysis of the "Soil-Plant
Barrier" indicated that plants could tolerate higher Co concentrations
than can be tolerated by ruminant livestock. Apparently vitamin B14 is
formed in the rumen, and this form of Co causes toxicity in the livestock.
Co feeding trials (see NRC, 1980} have shown that 5-10 ppm Co In diets
injures sheep and cattle. I have done a substantial risk assessment on
Co for a compost to be made from wastewater treatment blosolids at a
manufacturing plant of DuPont, and this could be made available to you
upon request. Thus, although no adverse effects of sludge-applied Co
have been reported to date, it is at least possible to poison ruminants by
Co in forage plants. Analysis can identify the very few high Co sludges
and require practices to prevent adverse effects.
So, of all the elements you have listed, Fe and F are the only ones with
sludge research showing a toxic environmental effect from sludges utilized
on land. Please add Fe and Co to the list now. And delete Ai, Sb, Ba, Mn,
Ag, Sn, Ti, and V. '.
Organics with substantial vapor pressure Uoluene; 2-butanona; methylene
chloride; phenol; 2-propanone; toluene} are expected to be volatilized or
biodegraded during activated sludge treatment of the wastewater, and trace
residues will collect in the sludge. Each of these compounds is readily
metabolized by soils, with short half-lives. These should be deleted because
Round 1 consideration of other volatile compounds showed that no residue
reached humans or livestock.
The 2,4-D, 2,4,5-T, and 2-(2,4,5)-TP are residues of pesticides which have
lower use today because of their adverse effects in Agent Orange which was
contaminated with dioxins produced as byproducts. These compounds are
usually sprayed on the plant, and metabolized fairly rapidly by tolerant plants,
but slowly by sensitive plants. These reactions are well reported in pesticide
applications at EPA. Because these are not very lipophilic, they are usually
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blodegraded rather than persistent in soils.
Cyanides can accumulate in sludges after precipitation of ferricyanide by
other metals. Soluble cyanide is present only at very low concentrations.
Sludges have low levels of total CM, and essentially all sludge/soil cyanide is
found to be bound to Fe.
"*!
I know little about CSa. But these is little evidence it would survive aerobic
treatment of the water.
Co-planar PCBs, PCNB, Dioxins, Dibenzofurans, and endosulfan are
persistent halogenated hydrocarbons. Detailed evaluation will be required for
these compounds. But the toxicity endpoints for the halogenated
hydrocarbons are seldom reached from these compounds in land-applied
sludges. The Madison, Wl, studies showed that no significant transfer of
sludge-applied PCBs was observed in above ground plant biomass =
forages. Direct ingestion of sludges allows digestion of these compounds
from the sludge. Accumulation of dioxins in earthworm-food-webs is
expected, but not yet shown to induce toxicity to animals.
Nitrate accumulates in fields with aerobic soils after sludge has been
incorporated. Some plants accumulate excessive levels of plant nitrate
{spinach, beet), and comprise nitrate-poisoning risk to infants. Further,
excessive nitrate accumulation in some forage crops can poison livestock.
Nitrite seldom accumulates unless some toxic factor inhibits nitrification of
the nitrite. Because sludge application rate is limited to the fertilizer
requirement of the crop, nitrate and nitrite so not require regulation.
1 heard a story about tungsten toxicity in a field study in the UK, but no
papers were prepared from the thesis and report to the funding agency. I
hope to visit the University of Sheffield and obtain a copy of the thesis on
May 12. Dr. Steven McGrath hypothesized that tungstate Interfered with
use of moiybdate in plants by competition as a co-factor for an enzyme
involved in N-fixation or nitrate reduction by the plants.
Thus, several elements on the list are of potential importance because of
their phytotoxicity rather than food-chain-transfer. These include Al, B and
F. Some comprise food-chain risk to livestock which graze the fields (F; and
possibly Be, Ba, and Be). Some are not dangerous to livestock even when
ingested (Ti, Sn, Sb, and probably Sb). As noted above, Fe and Co also
comprise risk until sludge analysts provides the management Information
needed to prevent risk.
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As shown by the Round 1 contaminants/ Se and Mo are accumulated by
forage crops such that they comprise risk to the livestock rather than to
humans consuming garden crops in Pathway 3. Mo needs to be finalized.
Several scientists have been conducting studies on Mo uptake by crops on
sludge-amended soils. High sludge Fe reduces Mo phytoavailability as noted
in my 1991/1992 comments on 503 Mo limits in which I clarified .errors in
the database on Mo uptake from sludges. The potential role of sludge Fe
(and Al) in binding sludge F can sharply reduce fluoride risk assessment for
sludges.
Based on widely accepted data about the trace elements on this list, I .believe
that the following should be deleted from Round 2 now (AI, Sb, Ba, Mn, Ag,
Tl, Sn, TI, and V, and the volatile organics). Others are only a risk in sludge
is ingested (Fe, F}, and some are sufficiently phytotoxic (based on field
studies with sludge) that they might be regulated to avoid phytotoxicity): Al,
B, Mn. And Fe should be added to include a well characterized sludge risk
from anaerobic treatment conditions. Cobalt is theoretically.toxic to
ruminant livestock after it is accumulated in forage plants.
Please feel free to call or write me for further information if needed. I
enclose several references which cover the Lisk/Furr papers, and have
several databases on the sludge-trace element literature in WordPerfect 5.1
which contain references on these rarer elements in sludges.
PLEASE CONFIRM RECEIPT OF THIS MEMORANDUM.
References cited in letter to Hais:
Boyer, K.W., J.W. Jones, D. Linscott, S.K. Wright, W. Stroube and W.
Cunningham. 1381. Trace element levels in tissues from cattle fed a sewage
sludge-amended diet. J. Toxicol. Environ. Health. 8:281-295.
VREF-VER/Copy [Sewage Sludge—CO: "Baxter et al.] "The levels of 20 elements (Al. Ca, Cd,
Cl. Co. Cu. Fe. K, Mo. Mn. Mo. Na, Mi. P. Pb, Rb. Sb. Se. V, and Zn ere reported for kidney, liver.
musda, spleen, and brain tissues taken from two groups of 6 steers per group In a feeding study
conducted at Colorado State University. The control group was fed a normal feedlot cattlo ration
and the test group was fed the sama ration amended with 12% (by weight) air-dried municipal
sewage sludge, elemental levels ere also reported for the contra! and test diets, control and test
faces, arid sewage sludge added to the diet. All samples were analyzed by 3CP-plasma emission
cpectroccopy and neutron activation analysis. Brief descriptions of the analytical methods are
included, the levels of all metals determined were elevated In the test diet (as much as 19>fold for
Cd) compared with the control diet. The levels of Pb and Cd in kidney and of Pb. Cd, and Cu ki
Ever In tha test animals were high enough to causa concern from a lexicological standpoint If those
tissues were consumed regularly by humans. None of the levels of any of the other elements in the
control and test animals tissues were high enough to cause similar concern with respect to human
consumption."
Samples from the 2nd study, with Ft. Collins sludge when It was still high in Cd and Cu. Wat
ashed samples. For higher concn metals, ran on ICP directly. For lower conen metals, adjusted to
pH near 5 and used chelex resin to collect metals from a larger aliquot, and than add stripped the
8
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metals into small volume for analysis. Co was by NAA. Co In kidney of (Control/Test) were
0.020/0.041 f/gig FW; liver Co: 0.047/0.077; Muscle Co: 0.07/0.01?; Spleen Co: 0.02/0.02; Brain:
0.009/6.019. Diet contained 0.09/0.43 ppm Co. sludge, 2.5 ppm; and f eces:0.43/1.60 ppm DW.
/RLC-fl
Capar, S.G., J.T. Tanner, M.H. Friedman and K.W. Boyer. 1978. Multielement
analysis of animal feed, animal wastes, and sewage sludge. Environ. Sci. Technol.
12:785-790.
VREF-VER/Copy (Sewage Sludge-CO: Baxter at aJ.J "Animal excreta and sewage sludge are
currently being used as animal feed ingredients on an experimental basis. The levels of 30 elements
are reported for a typical cattle feedlot diet, two dried cattle manures, a commercial cattle waste
product, tow dried poultry manures, and a metropolitan sewage sludge. The analyses arc
conducted using neutron activation analysis. Induction coupled plasma spectroscppy. atomic
absorption spectroscopy. and anodic stripping voltammetry. The levels of most Inorganic elements
are considerably higher In animal wastes and sewage sludge than in traditional animal feeds. For
most «iemem» tb« levels determined by several techniques are in good agreement. Problems of io«a
of lead with precipitate formation, accurate quantisation of elements present in high levels, and
obtaining homogeneous samples for analysis are discussed.* ,
Worried about clement contamination of sludge and manure if these are used as feed
Ingredients, thus analyzed many elements using newer techniques (at that time). The feedlot diet
contained 0.10 ppm Co, while manures contained 1.1-2.2 ppm Co, and Denver sludge, 7.1 ppm Co.
Also analyzed As. Ba. Be. Br. Cd, Cr. Cu. Eu. Hg. La. Mn. Mo. Pb. Rb. Sb. Sc. Se. Sn. Ti. V. Zn. Al,
Ca. CI, Fe. K, Mg. Na. and P. Found considerable contamination of samples with residues of e
homogenlzer (for Co, Cr, and Ni from stainless steel). Nete need for studies of risk and health of
animals which consume these contaminated materials. /KLC-Q
Chaney, R.L. and J.A. Ryan. 1993. Heavy metals and toxic organic pollutants in
MSW-composts: Research results on phytoavailabllity, bioavailability, etc. pp.
451-506. In H.A.J. Hoitink and H.M. Keener (eds.j. Science and Engineering of
Composting: Design, Environmental, Microbiological and Utilization Aspects. Ohio
State University, Columbus, OH.
Chaney, R.L., G.S. Stoewsand, A.K. Furr, C.A. Bacne and D.J. Usk. 1978b.
Elemental content of tissues of guinea pigs fed Swiss chard grown on municipal
sewage sludge-amended soil. J. Agr. Food Chem. 26:994-997.
V (Sewage Sludge-USDA: Chaney et al.-FEEDING] VREF-VER/Copy HCo In Soil/Plant: Misc.
Auth.) Sewage Sludge-USDA: Cheney et al.-BioavaiIabffityJ Because we used neutron activation
to analyze Co. data are available. "Swiss chard was grown on soil amended with municipal sewage
sludges from Baltimore and Washington, DC. The harvested crops were fed at 20 or 28% of diet to
guinea pigs for 80 days. Samples of soil, sludges, plant, and animal tissues were analyzed for up to
43 elements. The elements Br, Ca. Co. Eu. Fe. NI. and Sr were found at higher concentrations In
tissues of animals fed the chard cultured on sludge-amended soil than In control animals.
Composting sludge prior to amending the soil appeared to render certain elements such ac Cd, Cu,
NI. and Zn less available to Swiss chard subsequently grown.'
COBALT SUMMARY: Chard was grown on plots of Woodstown silt loam amended with 56
Mg/ha of Baltimore digested sludge. 112 Mg/ha ef Blue Plains digested sludge, and 224 Mg/ha of
composted digested Blue Plains sludge, and on control. Because the BP compost included some
serpentine rock chips, compost and chard were higher In Co than the other sludges: Son * 9.1
ppm; Balto - 9.4 ppm; BP Dig - 8.0 ppm and BP Compost =15 ppm DW. The chard (harvested
at maturity, washed, rinsed, freeze»dried and ground]: Control - 0.4; Balto - 0.8; BP Dig = 2.2;
and BP Comp = 1.1 mg Co/kg DW. These results follow the pH of the plots rather than the Co
content of the "sludge" or the amended soils. pH at harvest was 6.6, 5.0. 5.7, and 6.7 indicating
that compost acted as a liming agent in contrast with sludge. Kidney of one of the 4 replicate
-------
gulnaa pigs was analyzed for many elements, and all kidney, and fiver samples wora analyzed for Nl,
Pb. and Cd. Kidney Co was: Control •• 0.6: Balto » 0.7; BP Dig « 1.0; BP Cemp * not reported.
No significance test was possible on (ha Co data. Ni was increasad in Baltimore chard and
kJdney/llver of the guinea pigs; Although ail sludge grown chard was higher in Cd than tlh» control,
no Increase was found In kidney, or liverf Attribute this to presence of 2n in same tissue. The
guinea pigs did equally weQ on ail sources of chard, growing 450 g in the 80 days.
Davis, R.D. 1980. Uptake of fluoride by ryegrass grown in soil treated with
sewage sludge. Environ. Pollut. 81:277-284.
Decker, A.M., R.L Chaney, J.P. Davidson, T.S. Rumsay, S.B. Mohanty arid R.C.
Hammond. 1980. Animal performance on pastures topdressed with liquid sawage
sludge and sludge compost, pp 37-41. In, Proc. Nat. Conf. Municipal and Industrial
Sludge Utilization and Disposal. Information Transfer, Inc., Silver Spring, IMD.
* RLC.JQ
Francois, L.E. 1986. Effect of excess boron on broccoli, cauliflower, and radish.
J. Am. Soc. Hort. Scl. 11.1:494-498.
Francois, L.E. and R.A. Clark. 1979. Boron tolerance of twenty-five ornamental
shrub species. J. Am. Soc.. Hort. Sci. 104:319*322.
Furr, A,K., W.C. Kelly, C.A. Bache, W.H. Gutenmann, and D.J. Lisk. 197(5.
Multi-element absorption by crops grown on Ithaca sludge-amended soil. Bull.
Environ. Contam. Toxicol. 16:756-763.
V RLC-Q .
Furr, A.K., T.F. Parkinson, D.C. Elfving et al. 1981. Element content of vegetable
and apple trees grown on Syracuse sludge-amended soils. J. Agrie. Food Cham.
29:156-160.
V RLC.Q
Hogue, D.E., J.J. Parrish, R.H. Foote, J.R. Stouffer, J.L. Anderson, G.S.
Stoewsand, J.N. Telford, C.A. Bache, W.H, Gutenmann and D.J. Lisk. 1984.
Toxicologic studies with male sheep grazing on municipal sludge-amended soil. J.
Toxicol. Environ. Health 14:153-161.
VREF-VER/Copy [Heavy Metals in Soil/Plants: Lisk et al.— SLUDGE] "Growing sheep were
grazed for 152 days on grass-legume forage growing on soil that had been amended with municipal
sewage sludge from Syracuse, NY, at 224 metric tons/ha. Cd was higher, but not significantly IP
> 0.05), In tissues of sheep fed the sludge-grown forage as compared to controls. No significant
differences between the sludge or control treatments were found In weight ef the complete or
•cauda epididymls or in % progressive motility of cauda epididymal sperm. The sludge-treatment
group had significantly larger testes (P< 0.025) when expressed as a percentage of body weight,
end higher blood uric add values IP < 0,05). There were no observable changes In tissue
ultrastructure of Ever, kidney, muscle, or testes as examined by electron microscopy in either of the
treatment groups. There were no significant differences for rate of animal weight gain, carcass
weight, dressing percentage, or quality or yield grade of the carcass between tha treatment
groups."
Syracuse sludge. April 1980. applied weathered (1 yr) sludge to subsoil of Chanango gravelly
loam, pH 7.1. Amended soil was pH 6.7. Collected grass-legume hay for feeding studiai; in 1980
and" 1981. In 1982, used for grazing study. Had been planted with alfalfa, blrdsfoot trefoil,
10
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timothy, and brornegrass. 3 month old 'Mortem' sheep used to graze the pastures for 152 days.
Each animal was also fed 250 g feed concentrate dally, and ad Ub water. Composite sott from fi«id
Sludge contained S3 ppm Cd; forage 0.09 vs. 1.14 pprnDW Cd. Feed concentrate contained
0.21 ppm Cd. Kidney contained 0,55 ±0.14. ppmDW Cd vs. 0.83 ±0.17 NS; liver contained
0.22 ±0.04 vs. 0.40 ±0.08 NS; musde contained 0.03*0.01 vs. 0.09 ±0.04 ppmDW Cd NS.
Rate of gain was higher for sludge than control animals NS.
Kienhoiz, E.W;, G.M. Ward, D.E. Johnton, J. Baxter, G. Braude and G. Stern.
1979. Metropolitan Denver sewage sludge fed to feedlot steers, J. Anim. Sci.
48:735-741.
VREF-VER/Copy [Sewage Sludga-CO: Baxter, Kienhoiz. et a!.] 'Feadlot steers received 0, 4,
or 12% Metropolitan Denver sewage sludge on a dry weight Intake basis for a 94-day finishing
. period. Th« sludge was anaerobically digested primary sludge that had been treated with
poly electrolyte to aid in dewstering during vacuum filtering. K was then dried to 35% water prior to
mixing Into the pelleted diet given the steers. Cattle (6 on each treatment) were slaughtered and
kidney, liver, musde, bone, brain, blood. lung, spleen, and fat were analyzed for As. Cd, Cu,.Hg,
Mo. Ni, Pb, Se, and Zn. . .
. "Growth of the sludge animals was less than controls (P < 0.02S) because sludge, apparently,
provided no energy. Sludge ingestion caused no pathology. All 10 inorganic elements except M
were increased In one or more body tissues following the 94 day sludge Ingestion. Percentage
whole carcass retentions of ingested minerals were estimated as follows: 0-2% As. 0.04% Cd,
0.3% Cu. 0.07% Hfl, 0.2% Mo, < 0.006% NI. 0.6% Pb, 1.3% Se. 0.2% Zn. and 32% f. Steers
retained low amounts of the toxic heavy metals from sludge Ingestion."
Sludge containad (ppmOW): 1.3 As. 21 Cd(die« 0.025. 0.65, and 1.9 ppm), 710 Cu(diats 3.2,
31. and 86 ppm), 11 Kg, 40 Mo, 125 NI, 780 Pbfdiets 0.6, 26, end 77 ppm). 5.4 Se. 1500 Zn, and
200 F. Diet was pelletted corn + cottonseed mean + molasses + limestone + NaCI. corn silage ad fib.
Bone samples were taken from the proximal half of the tarsal bone. Samples digested with low
metal acids. . For many elements (not kidney or liver), sample metals were extracted by APDC.
crystals collected, and filtered; Taken Into small volume for analysis. Carbon rod used for some
samples. Good QA/QC program. At 12% sludge. As was increased in fiver, Cd in fiver and kidney,
Cu Increased in liver, Hg increased in liver, kidney and muscle. Mo Increased in bone and liver, Pb
Increased in Ever, kidney, bone, and blood: Sa increased In blood: Zn increased only in Uver. At
both rates. F increased In bone. NI did not Increase in any tissues.
Pb in tissues: Liver 0.2a 3.3b, 4.6c ppmDW for 0/4/12% sludge: kidney: 0.9a 12.2 b 15.8 b;
Muscle: 0.2 . 0.2: bone:1a, 4b, lie: blood: 0.12a. ., 0.82b; fat:0.16, ., 0.16. Cd In tissues: iiver:
0.2a 0.5b 0.4b: kidney: l.la 2.5b 2.4b; muscle: <0.01, ., <0.01. Hg: Uver:0.01a 0.06b 0.14e;
kidney 0.1 a, 0.45b, 0.9c. Cu: liver: 124a, 260b 240b.
Neary, D.G., G. Schneider, and D.P. White. 1975. Boron toxicity in red pine
following municipal wastewater irrigation. Soil Sci. Soc. am. Proc. 39:981-982.
NRC (National Research Council). 1980. Mineral Tolerance of Domestic Animals.
National Academy of Sciences, Washington, D.C. 577pp.
• RLC-Q
Rea, R.E. 1979. A rapid method for the determination of fluoride in sewage
sludges. Water Pollut. Contr. 78:139-142.
Sanson, D.W., D.M. Hallford and G.S. Smith. 1984. Effects of long-term
consumption of sewage solids on blood, milk and tissue elemental composition of
breeding ewes. J. Anim. Sci. 59:416-424.
VREF-VER/Copy ISewaga Sludge-NM: Smith et a!.] "Fine-wool ewes received for 2 yr a
complete palleted diet (11% protein) or the basal diet fortified with 3.5% cottonseed meal (CSM,
11
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12* protein} or gamma-Irradiated (1 megarad) dried solids (SS. 12% protein) from primary
(undigested), sewage (Us Graces, NM municipal sewage). Five awe* fad each diet were sampled to
determine Ag, Ca, Cd, Co. Cr. Cu, Fa, K, Mg. Mn, Na. M. P, Pb, and Zn In blood, milk and tissues.
Tissues and blood ware sampled at slaughter 40 days after weaning of lambs. Mean whole blood
mineral concentrations were similar (P > 0.05) among treatments 3 d postpartum; however, at 42
days after lambing both basal and sewage fed awes had elevated blood Ca compared with awes fed
CSM. No biologically Important differences were detected In the concentrations of elements in milk.
Ewes fed SS had lower (P<0.05) flood Fe than animals in the other groups. Sewage-fed ewe« alee
had higher {P<0.05} liver Fe (1092 ppmDw) than basal-fed ewes (626 ppm) whereas Fe In CSM-fed
awes (873 ppm) was similar to both. Basal-fed animals had 1.1-1.3 times more (P<0.05> liver Mg
and 2-to-3-foid higher liver Na than CSM or SS. Uvars from SS-fed ewes had higher concentrations
0.051. element concentrations In
whole blood at weaning, after SB days of the feeding trial and at slaughter did not differ (F'>0.05)
between dietary groups. Serum chemistry determinations showed no biologically meaningful
patterns related to diets. Lambs fed SS had higher (P<0.05) canon, of Cu in livers (51.1 vs. 34.3
//g/g) and Pb In kidneys (4.0 vs. 2.2 ±0.3 j/g/g and lower Mg in kidneys. None of the elements in
spleen and muscle tissue differed (P>0.05) between diet groups. Lambs fed SS had elevated
(P<0.05) bone Co, Cu. Fe. K, and Na compared with those of CD. Lead concn. In bone were
Increased (P<0..05) by Ss over CD (30.5 vs. 26.3). but Cd and Zn did not differ. A feedlot diet with
7% SS did not appear to adversely affect growth or carcass characteristics of lambs. Serum
clinical profiles and chemical elements in blood and tissues were affected negligibly by SS as 7% of
the diet."
Sludge composition averaged: 3 Cd, 470 Cu, 9233 Fe, 110 Mn, 9 diet consumption, and
lower gain rates. Blood Cu not affected by sludge Ingestion. Uver contained: Cd < 0.07/0.07
ppmDW; Cu 34.3/51.1; Fe 179/190; Pb 2.5/3.5; Kidney: Cd <0.07/<0.07; Pb 2.2a/4.0b; bone:
Pb 26.3a/30.Bb.
Smith, G.S., D.M. Hallford and J.B. Watklns, III. 1985. lexicological effects of
gamma-Irradiated sewage solids fed as seven percent of diet to sheep for four
years. J. Anim. Sci. 61:931-941.
12
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VREF-VER/Copy [Sewage SIudge-NM: Smith at at.] 'Breeding awes ki dryiot ware fed pelleted
complete diets with 3% cottonseed meal {CSMJ or 7% dried, sarnma-lrradiated sewage so&ds
(DGSS) for 4 yr. Cytochrome P-450 content and enzyme activities for xenobiotic*
biotrarisformatlons ware assayed in livers after 3 yr and In liven, kidneys and Heat tissue after 4 yr.
Dietary DGSS caused no Increase in P-450 and few changes In activities of oxldativa, hydrolative,
and conjugates blotransformatlonaJ enzymes. Consumption of DGSS for 4 yr caused slight
enlargement of spleens (1.1-fold) and ovaries (1.3-fold, P<0.10). but no change In size of fiver*.
kidneys, hearts, adrenals and thyroids {R>0.10). nor Pver vitamin A levels (P>0.10). Of 22
refractory lipophilic residues assayed in abdominal adipose tissue, few were detected and of those
detected DGSS caused none to exceed normal levels. Dietary DGSS increased IP < 0.01) Fa In Ivan
1.5-fold and In spleens 5.6-fold, and Increased Cu in Overs 1.3-fold (P<0.01) and in kidneys 1,2-
foid. Dietary DGSS increased Cd level* in fivers but not in kidneys or spleens {P>0.10); yet all Cd
levels were within ranges for livestock fed conventional feed. Dietary DGSS caused no increase
(P>0.10) In levels of Ag. Caf Cr, Hg. K. Mg. Mn, Na, M. P. pfa. or Zn in livers, kidneys or spleens.
There were no histopathological lesions of toxicosis except mild hemosiderasis of spleens.
Consumption of a diet with 7% DGSS throughout 4 yr caused no hazardous accumulation of toxic
elements and little, if any, evidence of toxiclty."
Undigested sewage solids (primary and activated) from Las Crucas. NM. Dried and irradiated.
.Contained: 0.58% Fa; 606 ppm Zn; 405 ppm Cu; 361 ppm Cr: 150 ppm Pb; 99 ppm Pb; 11 ppm
Mi; <5 ppm Hg: <1 ppm So. 41.5% ash. Liver Fe was Increased. 849±387 (SD) vs. 1303 ±291
ppm DW. Uver Cu was raised: 597±308 vs. 761 ±259 ppmDW. liver Cd [<0.03 vs. 1.47*0.30
ppmDW] was raised, but kidney was not [2.8±0.3 vs. 3.6±0.6 ppmDWJ. Pb was unchanged and
at very-low levels In Over, kidney, and spleen «0.10 ppm DW). p.p'DDE was increased in fat, but
PCS and other chlorinated hydrocarbons were not increased. The animals were mature, tine-wool
ewes of Rambouillet breading. /BLC«Q
Vimmerstedt, J,P. and T.N. Glover. 1984. Boron toxicity to sycamore on mihesoil
mixed with sewage sludge containing glass fibers. Soil Sci. Soc. Am. J. 48:383-
393.
13
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TABLE 5. Maximum tolerable levels of dietary minerals for domestic livestock in
comparison with levels in forages.
'——~~ Eiement "Soil- Level in
Plant-Foliage* MaxirhumJLevels Chronically Tolerated* Plant
Barrier" Normal ghyto.toxic Cjsittfi Sheep
Swine Chicken
-mg/kg dry foliage ——mg/kg dry diet --
As. Inorg. yes 0.01*1
B
Cd'
Cr3*
Co
Cu
F
Fe
Mn
Mo
NI
Pb'
SQ
V
Zn
yes 7-75
Fails O.i-l
yes O.I-l
Fail? O.ON0.3
yes 3-20
yes? 1-5
yes 30-300
? 15-150
Fails 0.1-3.0
yes 0.1-5
yes 2-5
Fails 0.1-2
yes? O.I-l
yes 15-150
3-10 50. 50. 50. 50.
75 150. (150.) (150.) (150.)
5-700 0.5 0.5 0.5 0.5
20 (3000.) (3000.) (3000.) 3000.
25-100 10. 10. 10. 10.
25-40 100. 25. 250. 300. .
40. 60. 150. 200.
1000. 500. 3000. 1000.
400-2000 1000. 1000. 400. 2000.
100 10. 10. 20. 100.
50-100 50. (50.) (100.) (300.)
30. 30. 30. 30.
100 (2.) (2.) 2. 2.
10 50. 50. (10.) 10.
500-1500 500. 300. 1000. 1000.
£/ Based on literature summarized in Chaney et al. (1982).
&/ Basad on NRC (1980). Continuous long-term feeding of minerals at the
maximum tolerable levels may cause adverse effects. Levels in parentheses were
estimated (by NRC) by extrapolating between animal species.
sJ Maximum levels tolerated were based on Cd or Pb in liver, kidney, and bone in
foods for humans rather than simple tolerance by the animals.
From: Chaney and Ryan, 1993.
Boron Phytotoxicity: In contrast with municipal sewage sludge, MSW-compost
contains substantial levels of soluble boron (B). B toxicity from sewage sludge
application was reported only for an unusual case of a sensitive tree species
growing In soils amended with a sludge containing lots of glass fibers (Vimmerstedt
and Glover, 1984; see also Neary et al., 1975, regarding high B levels in
phosphate-free detergents). The glass fibers contained borosiiicate and release of
B caused phytotoxicity. Research has shown that much of the soluble 8 in MSW- _.
compost comes from glues (Voik, 1976). It has long been known that plant IM
samples placed in paper bags can become contaminated from B from glue used to
14
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hold the bag together. E! Bassarn and Thorman (1979) and Gray and Biddlestone
(1980) noted that the B level in MSW-composts was quite variable as might be
expected if composts are not well mixed.
In general, B phytotoxicity has occurred when high application rates were used,
and B-sensttlve crops were grown. However, when MSW-compost is used at
fertilizer rates In normal fields, the B might be important as a fertilizer rather than
as a potential phytotoxicity problem.
Boric acid and most berates are quite water soluble, although B can be adsorbed
on clays and by organic matter. Low soil pH facilitates B uptake by plants because
the H3BO3 molecule (predominant form at lower soil pH) is absorbed by roots rather
than anionic borates (Oertli and Grgurevic, 1975). Although most B toxicity has
been reported on alkaline soils, this is due to the lack of leaching for most of these
soils. Excess applications of soluble B are much more phytotoxlc in acidic soils,
and liming can correct B phytotoxicity. The usual liming action of compost should
help prevent this problem. .
There are large differences among crop species In tolerance of excessive soil B.
Some crops are very sensitive, and these are the species which have suffered
phytotoxicity from compost-applied B (bean, wheat, and mum). Francois has
summarized the significant differences among several groups of crops (Francois
and Clark, 1979; Gupta, 1979; Francois, 1986). Ornamental horticultural species
have been examined to some extent (information on Individual species can be
found by literature searching), but many horticultural crops have not been studied.
This is one research need related to practical microelement phytotoxicity from
compost.
Perhaps the first report on B toxicity from MSW-compost is that of Purves
(1972) who noted B phytptoxicity to beans on field plots which received high rates
of MSW-compost. The full description of the compost experiment is reported in :'
Purves and Mackenzie (1973). and a careful examination to prove B phytotoxicity
was reported by Purves and Mackenzie (1974). Bean (but not potato or other
species examined) suffered severe yield reduction at high compost rates; this yield
reduction was proportional to rate of compost application. Bean is known to be
especially sensitive to B phytotoxicity. Gray and Biddiestone (1980) also found B
phytotoxicity In sensitive species grown in field plots with high rates of MSW-
compost.
Gogue and Sanderson (1975) reported B phytotoxicity to chrysanthemums in
potting media containing MSW-compost. Foliar analysis clearly supported the
conclusion that B was toxic and that Mn, Cu, Zn, and other, elements were not at
toxic levels. They conducted a calibration experiment to determine the sensitivity
of chrysanthemums (Gogue and Sanderson, 1973), and the levels found in the
mums grown on the test media were In the phytotoxic range, in their research,
they adjusted the pH of the media to 6 using sulfur, rather than allowing the MSW-
compost to raise the pH of the media. This probably contributed to the severity of
B phytotoxicity observed. Some other horticultural species also suffered B
phytotoxicity in compost-containing media (GUUam and Watson, 1981). Sanderson
(1980) reviewed B toxicity in compost amended potting media. In contrast to
MSW-compost, sewage sludge composts with wood chips have not been found to
cause B phytotoxicity (Chaney, Munns, and Cathey, 1980). Only a few acid-loving
15
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species require acidification of media to do well on neutral compost-amended
media.
Interestingly, because the B which causes phytotoxiclty is water soluble, the B
phytotoxlcity problem from MSW-compost Is short-lived. Purves and Mackenzie
11373} noted that pre-leaehing MSW-compost prevented B phytotoxicity. Other
studies noted that the B-phytotoxlcity occurred only during the year of application,
and that soluble B was leached out of the root zone over winter (Volk, 1976) or by
leaching potting media with normal horticultural watering practices. Sanderson
(1980) noted that perllte also adds B to potting media, and that use of both may
cause B toxiclty when either periite or MSW-compost alone might not have don*
so. Lumis and Johnson (1982} studied leaching of B in relation to toxicity of salts
and B to Forsythia and Thuja. They reported that a simple leaching treatment
removed excess soluble salts, but was unable to remove enough B to prevent
phytotoxlcity (the compost they studied contained 225 mg B/kg, higher than most
reports}. Nogaies at at. (1987} also found compost-applied B leached quickly such
that crop B was reduced in each successive ryegrass crop.
B phytotoxicity Is significantly more severe when plants are N-deficiant IGogua
and Sanderson, 1973; Nogaies et a!., 1987; Gupta et al., 1973). This makes the B
in MSW-compost which is not properly cured (to avoid N immobilization) potentially
more phytotoxic than in well cured composts. Further, B flows with the
transpiration stream and accumulates in older leaves, in environments with low
humidity, more transpiration occurs (e.g., greenhouses), and B toxicity is more
severs. B and salt toxicity are easily confused; both are first observed in leaf tips
or margins of older leaves. Diagnosis of B phytotoxicity requires a knowledge of
relative plant tolerance of B, or analysis of the leaves bearing symptoms.
Thus, in general use, compost application at a reasonable fertilizer rate would
simply add enough B to serve as a fertilizer for B-deficiency susceptible crops such
as alfalfa or cole crops. However, use of MSW-compost at high rates in soils or
potting media could cause phytotoxicity if high soluble B were present. The B
phytotoxicity would not be persistent because soluble B would leach from The root
zone with normal rainfall or .irrigation. Compost-applied B would be more
phytotoxic in N-deflcient soils, which might result from application of Improperly
cured compost. Water soluble B should be one chemical which is regularly
monitored in MSW-composts so that the need for warning about rates of
application and use with sensitive crops can be Identified. Deliberate use of MSW-
compost as a B fertilizer for high B-requiring crops such as the cole crops (cabbage
family) might become a regular agronomic practice. Sources of soluble B In
modern MSW-compost should be evaluated, and alternative to B use identified.
16
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Reprinted from the Journal of Environmental Quality
Vol 19, no. 3, July-September 1990, Copyright C 1990, ASA. CSSA, SSSA
677 Sooth Segoe Road. Mmdinon, WI 53711 USA
Plant Uptake of Pentachlorophenol from Sludge-Amended Soils
Cheryl A. Bellin and George A. O'Connor
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Plant Uptake of Pentachlorophenol from Sludge-Amended Soils
Cheryl A. Bellin and George A. O'Connor*
ABSTRACT
A greenhouse study was conducted to determine the effects of
sludge OB plant •ptmke of "C-prntarslorophtsMC (PCP). Plants fe-
clwied tall tacae(FextmemmTm*di*ma* Schreb.), kttace (£«•«• je-
ft'wi L.), csmt (Dometts carota L,), and chfle pepper (Comou* am-
m»m L.X Minimal intact PCP was detected ia the fescue and lettoee
by gas chroamtognphy/inasf spectranwtry (GC/MS) analysis. No
intact PCP was detected in the carrot tissue yrtrartt. Chfle pepper
was not analyzed for intact PCP became netfaykoe chloride extracts
contained minimal "C. The GC/MS analysis of caQ extracts at har-
vest suggests a half-life of PCP of abort 10 d independent of stodge
rate or PCP loading rate. Kapid degndatioa of PCP is die soil
apparently linked PCP anSabflity to the plant. Bioeonceotntfoa
factors (dry plant wt/initial soil PCP coocentratio*) based oa intact
PCP were <(UH for aO craps, iism ilhn tittle PCP •stake. That,
food-chain crop PCP intake in these alkaline sofls sbooJd not limit
land application of stodge.
LAND APPLICATION of sewage sludge is an inexpen-
sive and convenient method of sludge disposal
that provides plant nutrients and improves soil struc-
ture. The potential bioavailability of toxic organics
that can contaminate some sludges (e.g^ polychlori-
nated biphenyls) may limit sludge utilization in agri-
cultural settings, however. The large number of toxic
organics that can contaminate sludges, complex anal-
ysis, and unlimited environmental conditions makes
it virtually impossible to study each compound in
every setting. Therefore, a compound is selected based
on physical and chemical properties to represent a
group of compounds.
Chlorophenols are a group of ionizable organic com-
pounds of environmental concern. Chlorophenols are
not generally detected in sludges nationwide. How-
ever, .156 of 223 industrial and municipal sewage
sludges from Michigan contained pentachlorophenol
(PCP) at concentrations ranging from 0.2 to 8 490 mg
kg-' dry weight (Jacobs et al., 1987). The median con-
centration of the PCP contaminated sludges was 5 mg
kg"1-
The widespread use of PCP as a wood preservative
and general biocide has lead to contamination of air,
food, sediment (Bevenue and Beckman, 1967), water,
and municipal sewage sludge (Buhler et aL, 1973). Pen-
tachlorophenol is categorized as very toxic [oral LDso
= 146 mg kg-1, rat (Rattus norvegicus); Crosby, 1980]
and is mutagenic in MP-1 yeast strain (Fahrig et al.,
1978). The generally low PCP concentration in con-
taminated sludges (5 mg kg-1) would be reduced 50 to
100 times during land application of sludge at normal
agronomic rates (22.5-45 Mg ha-1). Accumulation of
Both authors, Dep. of Agronomy and Horticulture, New Mexico
State Un iv.. Las Cruces, NM 88003. Journal Article no. 1499. Agric.
Exp. Stn., New Mexico State Univ. Although financial support for
2?L!?fe *?! Provjded «» Pan by Cooperative Agreement CR-
8126874)2 with the USEPA, this report has not had USEPA's re-
quired peer and policy review and does not necessarily reflect the
views of the agency. Received 14 July 1989. "Corresponding author.
Published in J. Environ. Qual. 19:598-602 (1990).
PCP in food-chain crops would not likely adversely
affect humans at the resulting PCP soil-sludge con-
centrations.
Bioavailability of nonpolar compounds such as PCP
- depends on the extent of sorption by the soil organic
fraction and other transformation processes (Hamalcer
and Thompson, 1972). Sludge additions increase the
soil organic fraction and thus possibly alter PCP bioa-
vailabDity. The chemical characteristics also contrib-
ute to the complexity of PCP sorption in the soil-
sludge system. Pentachlorophenol is a.weak acid (pK,
- 4.74; USEPA, 1979) and sorption increases with
decreasing soil pH (Lagas, I988a; Baaerji et al., 1986)
possibly altering the bioavailability of PCP in soils.
Uptake of I4C by crops grown in "C-PCP treated
soils has been reported; however, actual PCP and me-
tabolites were not verified (Kloskowski et aL, 1981;
Topp et aL, 1986). Topp et aL (1986) reported signif-
icant >4C uptake for barley (Hordeum vulgare L.) after
1 wk in a soil with a pH of 6.4. However, I4C in the
plant was attributed to rapid degradation and min-
eralization of PCP followed by assimilation of I4CO2.
On the contrary, CasterHne etaL (1985) reported intact
PCP uptake by spinach plants (Spinacia oleracea L.)
and soybean plants [Glycine max (L.) Merr.} from an
acid soiL Methoxytetrachlorophenol, 2,3,4,6-tetrach-
lorophenoL pentachloroanisoL and 2,3,4,6-tetrachlo-
roanisol were the primary metabolites detected in the
plant tissue. The purpose of this experiment was to
determine the effects of sludge on PCP uptake by
plants grown in alkaline soils in the greenhouse.
MATERIALS AND METHODS
. The soils used in this study were a Glendate clay loam
ffme-silty, mixed (calcareousX thermic Typic Torrifluvent]
and a Biucpoint sandy loam (mixed, thermic Typic Torrip-
samment), with pH values (water paste) of 7.8 and 8.3, re-
spectively. The sofls were air-dried and sieved (6.25 mm)
prior to the experiment.
A municipal sewage sludge obtained from Albuquerque,
NM, was anaerobically digested, air-dried, and gamma ir-
radiated (II7Cs 10 kGy) to reduce pathogens. The sludge was
ground to <2 mm and amended to the soils at rates of 0,
22.5, and 45 Mg ha-' (0, 10, and 20 g kg-')- The sludge was
PCP-free, <1 mg kg-1 as analyzed by gas chromatography
with a flame ionization detector (USEPA method 625).
Reagent-grade PCP (Chemical Dynamics Corp., South
Plainfield, NJ) and "C-PCP (Sigma Chemical Co., St. Louis,
MO) (specific activity 455.1 MBq mmol-', uniformly ring- ,
labeled) were used to obtain desired PCP rates (0, 0.1, 0.6,
1.1, and 5.1 mg kg-').
Fescue (Festuca anatdinacea Schreb., 'Ky 31') and three
foodchain crops: lettuce (Latuca saliva L., 'Great Lakes'),
carrot (Dauaa carota L, 'Nantes Scariet'X and chile pepper
(Capsicum, annum L., 'Espanola Improved*) were grown in
a greenhouse. The foodchain crops were chosen to represent
a leafy crop, a root crop, and a fruit crop, respectively. Treat-
ments were duplicated.
Greenhouse Procedure
Soil, sludge, and fertilizer (phosphate fertilizer 920 kg P2O5
ha-1) totaling 4 kg were mixed in a twin shell blender (18 kg
598
-------
BEUJN & O'CONNOR: PLANT UPTAKE OF PENTACHLOROPHENOL
599
capacity) for S min. This soil-sludge mixture was placed in
two pots (2 kg nor1) and leached to remove excess salts.
After drying, soil-sludge mixtures were spiked with '•'C-PCP
solutions (20 mL of 9.1 .AS Nad containing 740 MBq I4Q.
The soil was mixed in a twin shell blender for 3 min and
weighed into 18-cm diam. pots (2 kg pof). Drainage holes
in the pots were covered with fiberglass mesh. A 100-g sam-
ple of the soil-sludge mixture was retained for verification
of "C-PCP application rate.
Seeds were placed on the soil surfaces and covered with
1 cm of soil previously removed from the pot The pots were
then watered by a drip irrigation system to 80% pot water
holding capacity and maintained at this moisture content
gravimetneally. A total of 16 h daylight was maintained (ad-
ditional light fr°m c&bt 400-W sodium lamps). Tempera-
tures ranged from IS to 40 *C Liquid calcium nitrate fer-
tilizer (equivalent to 360 kg N ha-') was added to the
Bluepoint soil to compensate for N available in the sludge
treatments. The experimental design was a randomized com-
plete block for the lettuce, carrot, and chile and a completely
random design for the fescue.
Plant and Son Sampling
Fescue was grown in the Bluepoint soil and the Glendale
soil for 34 and 42 d, respectively. Lettuce was grown for 58
d, carrot for 79 d, and chile pepper for 115 d.
Plants were cut 2 cm above the soil surface to avoid con-
tamination from the soil Chile fruits were cut from the fo-
liage and carrot roots were removed from the soiL pie carrot
roots were washed in an ultrasonic bath with deionized water
for 5 min.
Fresh weights of each fraction (combustion and extraction
fractions) were recorded for an plant parts. The fraction for
combustion was oven-dried at SO °C, reweighed, and ground
in a. Wiley mill with a 20-mesh screen. Extraction samples
were stored in Ziploc bags in a freezer at —10 °C
Soil samples, consisting of four 2-cm diam. cores to the
depth of the pot (100 g), were taken from each pot at each
harvest The soil samples were divided into two fractions:
one for combustion and one for extraction. The samples for
combustion were air-dried and ground with a mortar and
pestle. The extraction samples were stored in Ziploc bags in
a refrigerator at 5 *C .
Analytical Methods
Combustion. Ground, air-dried, 300-mg soil samples and
ground, oven-dried, 50-mg plant samples were replaced in
boats, covered with activated alumina/cupric oxide (5:1 w/
w), and combusted at 1000 *.C (Lindberg furnace) for 8 min
under a stream of Oj (150 mL min-1). Evolved "COj was
trapped in a solution of 8 mL ethbxyethanolethanolamine
(3:2 v/v) and 10 mL Ready Gel cocktail (Beckman). Samples
were counted by liquid scintillation (LS) with a Beckman
LSI800 counter. Blanks and standards were combusted and
counted to determine background counts (40 disintegrations
per minute, dpm), oxidation efficiency (about 0.95), and
counting efficiency (0.65). The mean radioactivity was 0.184
± 0.018 MBq kg-' soil-sludge mixture in all PCP-containing
treatments.
Extraction Procedure. Soil and plant samples containing
the highest '^C concentrations were extracted and analyzed
by LS counting and gas chromatography with mass spec-
trometer detector (GC/MS). Samples were extracted with a
procedure modified from Casteriine et aL (1985). The orig-
inal procedure extracted PCP, lower chlorinated phenols,
and anisols individually from plant and soil samples (93.2%
total plant extraction efficiency). This procedure was sim-
plified because individual FCP metabolite identification was
not attempted.
Fine roots were removed from sofl samples and 10 g of
soil and 10 mL concentrated HQ were combined in glass
centrifuge bottles, and heated m an oven at 60 *C overnight
After cooling, the acidified soils were extracted with 50 mL
methylene chloride. The bottles were stoppered, shaken for
2 min, and centrifuged for 5 min at 2000 rpm (750 X g),
The methylene chloride was removed and collected in a
beaker. The soil mixture was extracted three more times for
a total methylene chloride volume of 125 mL,
Frozen plant samples were homogenized in a Waring
blender with 100 mL, 0.2 M Hd for 5 iron. The mixture
was transferred into a 200-mL glass centrifujie bottle, 100
mL methylene chloride added, and the bottle stoppered. Af-
ter shaking for 2 inin, the inixture was centrifui^ for 4 rnin
at 1000 rpm (300 X g). The methylene chloride was removed
and collected into a beaker. The plant and acid mixtures
were extracted three more times with 50, 25, and 25 mL
methylene chloride. All extracts were combined.
Extract Cleanup. Soil and plant methylene chloride ex-
tracts were evaporated to 25 mL. A 2-mL fraction of the soil
extract was evaporated to dryness in a LS vial, dissolved
with scintillation cocktail, and counted by LS, A 2-mL frac-
tion of the plant extract was evaporated to dryness in a LS
viaL Bleach (0.75 mL of 5.25% sodium hypochlorite) was
added and the suspension heated at 50 °C for HIT. A mixture
of acetic acid and Ready Gel cocktail (18 mL of 3:400 v/V)
was then added to the bleached plant samples, and the re-
sulting mixture counted by LS.
• The remaining 23-mL fractions of methylene chloride
were placed into glass centrifuge tubes with 10 mL NaOH
(pH 9), shaken for 2 min, and centrifuged for 4 min at 1000
rpm (300Xg). The methylene chloride was discarded. The
NaOH was acidified with 2 M Hd to pH<2 and extracted
with 2 mL methylene chloride. The acid fractions were ex-
tracted two more times with 2 mL methylene chloride. All
extracts were combined in one vial and evaporated to dry-
ness with N2 gas.
Extract residues were derivatized for GC/MS (MS Model
HP5970, connected to a GC Model HP5890) analysis by
heating at 70 °C for 15 min with 20 nL N,O-
bis(trimethylsilyl)-trifluoroacetamide (BSTFA) (Poole,
1978). Aliquots were injected on to a 0.32 mm :i.d by 25 m
Ultra 2 column (Hewlett Packard, Boulder, CO) with a 0.52
Mm film (Cross-linked 5% Phenol Methyl Silicons) at 150 °C.
The oven temperature was ramped at 25 °C min-' to 280 °C,
with a final hold time of 4 min. The detector temperature
was 270 °C The helium carrier gas flow rate was 1.5 mL
min"'.
The mass spectra were recorded with the electron multi-
plier at 1800 V and the ionization energy was preset at 70
eV. Selective ion monitoring (SIM) at ion masses of 321,
323,336, and 338 at 100 ms dwell times was used to analyze
the extracts.
Calculations
Total 14C estimates of plant uptake of PCP were based on
I4C of dry combusted plant material. This I4C represents the
maximum amount of PCP possible in the plant .is I4C-PCP,
l4CMabeled metabolites, or "CO* Carbon-14 was detected
in the control plants grown in soil containing no "C-PCP.
The '^C in the control plants probably represented foliar
assimilation of the I4CO2 released from pots containing I4C-
treated soils. This contamination was assumed uniform
across all treatments. Thus, "C contents for crops were re-
corded as net I4C (gross I4C dpm g-' in each treatment minus
I4C dpm g-1 in the controls, mean — 300 dpm g- ')• Biocon-
centration factors (BCF) were calculated by dividing the net
I4C (dpm g-1) in the dry plant material by the initial "C
(dpm g-') in the dry soiL
-------
600
J. ENVIRON. QUAL, VOL 19, JULY-SEPTEMBER 1990
RESULTS AND DISCUSSION
Carbon-14 Concentration in Plants
Analysis of plant tissue by combustion (total
suggested uptake of "C-PCP (Beffin, 1989)
mation of mtact PCP in the plant tissue, S
necessary due to possible "C-PCP degradation
detection of '^-labeled metabolites.
14O
i rom m
the highest combustion "C contents, The«'C i
methylene chloride extract represents PCP
cWonnated phenols, and anisols (CasteSne el
—— —— —uwM^yicnc
, -, —i significantly less
PentacUorophenoI Concentration in Plants
ett»ct»' ^A *e high-
n, were analyzed bv GC/VfS «nwn%«
identify only intact PCP. TrWamounteofLgSpCP
were detected in the fescue and fettucT NoPCT^S
™1
&
m
1.00
6.75
aso
ground tissue. The ^centratio? of^PTP ?^T
Bioconcentration Factors
O.25
aso
0.75
1.00
Organic' Carbon Content
t™™^™^^,^*^^^^^^^^^^
-------
-------
BELLIN & O'CONNOR: PUNT UPTAKE OF PENTACHLOROPHENOL
601
the change in BCF values, calculated from 14C in meth-
ylene chloride extracts. This effect was even more ap-
parent when BCFs were calculated from actual PCP
concentration, as determined by GC/MS SIM analysis
of the methylene chloride extracts.
The BCFs for fescue, carrot, and chile (Table 1) were
calculated from total 14C determined by combustion,
14C in the-methylene chloride extracts, and GC/MS
measurements. The BCFs calculated from total 14C
determined by combustion were maximum BCF val-
ues representing l4Cas PCP, PCP-metabolites, and any
other compound containing I4C. The BCFs from I4C
in the methylene chloride extracts (representing PCP,
lower chlorinated phenols, and anisols) were substan-
tially lower. The BCFs based on GC/MS analysis (ac-
tual PCP) were less than 0.01 for all crops. The PCP
was not detected in the carrot peels or tops. The ex-
tracted radioactivity in chile samples was only 5% of
the total (combustion) radioactivity. Thus, due to al-
ready low BCFs based on |*C in the methylene chloride
extracts, chile (foliage and fruit) was not analyzed by
GC/MS.
The BCFs based on actual PCP suggest minim?!
plant uptake in any sludge or PCP rate treatment This
was likely caused by rapid degradation of PCP in the
soils. •
Topp et aL (1986) reported PCP concentration fac-
tors (fresh plant weight/air dry soil) for barley (Hor-
deum vulgare L.) of about 7 after 1 wk in a soil with
a pH 6.4. This concentration factor, however, based
on I4Q likely overestimated actual PCP accumulation.
Topp et al. (1986) attributed the large bioconcentra-
tion factors to rapid degradation and mineralization
of PCP, followed by plant uptake of I4CO2.
Casterline et al. (1985) reported BCF values (fresh
plant weight and air dry soil) of »1.0 for spinach
plants, soybean plants, and soybean roots, based on
analysis by gas chromatography with an electron cap-
ture detector. Thus, BCFs suggested significant PCP
uptake and much greater uptake than measured here.
Several factors may have contributed to the greater
bioconcentration factors. Soil analysis revealed a half-
life of PCP of about 25 d (whereas the PCP half-life
in our soils was about 10 d). The relative average plant
exposure increases with increasing half-life. Based on
a 25- and 1 OK! PCP half-life the-relative average ex-
posure would be 0.3 and 045, respectively, for a 100-
d growing season (Ryan et aL, 1988). Thus, greater
uptake would be expected with a longer PCP half-life
Casterline et al. (1985) sterilized their soil to inten-
tionally delay degradation. The low sou* pll (estimated
to be <5.5) may also have reduced the activity of
microorganism responsible for PCP degradation
DeLauhe et aL (1983) reported maximal degradation;
at pH 8, and decreasing degradation with increasing
or decreasing pH. Additionally, PCP sorptipn (Bellin
et aL, 1990) was greater in an acid soil than two al-
kaline soils. Adsorbed PCP may be less available for
degradation than solution PCP. Speitel et aL (1989)
reported slower degradation when PCP was adsorbed
to granular activated carbon than when PCP remained
in solution. Therefore, slower degradation in the add
soil due to increased sorption allowing longer PCP
availability and thus, greater opportunity for plant up-
SUMMARY AND CONCLUSIONS
Detection of intact PCP by GC/MS in fescue and
lettuce revealed minimal plant uptake of intact PCP
The BCF values (plant dry wL/initial soil concentra-
tions) were <0.01 for fescue and lettuce. On a fresh-
weight basis, BCF values were <0.001. Intact PCP was
not detected in the carrot (foliage, peel, or pulp) and
chile plants (foliage or pods).
The GC/MS analysis of extracts of soil samples
taken after each crop harvest suggested PCP degraded
so rapidly in these soils that minimal plant contami-
Table 1. Fescue, carrot. and chile bioconcentratjon factors based on initial gog concentration «nd dry plant weights.
Carott
Chile*
Sludge
rate
PCP
rate
Plant
Total}
Peels
MeOf GC/MS*
Total
Med
Total
Med
Plant
Total
Fruit
Total
Mg/ha
0
22.5
45
nig/kg
0.6
5.1
0.6
5.1
0.06
5.1
0.77
134
1.18
2.25
0.84
2.18
0.08
0.14
0.08
0.08
0.06
0.10
-tt
0.0072
0.0001
Bfaepoint toil
0.06
1.09
0.04
OJ3
0.03
0.28
Gleodafesoil
0.06
0.02
0.04
1.22
0.48
1.34
OJS
1-32
0.33
0.46
0.24
0.40
0.23
0.29
0.17
0.19
0.08
0.13
0.06
0.15
0.06
0.09
0
22.5
45
0.6
5.1
0.6
5.1
0.6
5.J
1.32
0.96
0.61
0.60
0-24
0.86
0.12
0.06
0.06
0.02
<0.00004
0.41
O22
0.45
0.10
0.20
0.13
0.03
0.01
0.01
1.71
0.27
2.09
0.37
1.63
0.42
0.13
0.08
0.16
0.07
0.09
0.09
0.15
0.16
0.09
0.07
0.17
0.27
0.05
0.00
0.03
0.03
'0.07
—.-»...v wimwuMrMUMUUH TXJUCft IUT OTZVI PCCIS WCTC < 0.000 I
Mea bioconcentration &aors for chife plant and {rat were
-------
602
J. ENVIRON. QUAU, VOL. 19, JULY-SEPTEMBER 1990
nation could occur in the field. Thus, given normal
application rates of sludge with normal (or even ab-
nprmally high) PCP concentrations, concerns about
food chain plant uptake of PCP should not limit land
application of sludge in these soils. This conclusion is
likely appropriate to other high (>6.5) pH soils. How-
ever, fUrther study is necessary to determine the bioa-
vailabllity of PCP in sludge-amended, low-pH soils,
particularly those with high organic C contents where
PCP half-lives are reportedly much longer (Bellin et
aL, 1990).
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phenol adsorption on soils and its potential for migration into
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disposal VpL 6. ASTM STP 933. p. 120-139. Am. Soc. for Testing
and Materials. Philadelphia, PA.
Bellin, GA. 1989. Plant uptake of pentachlorophenol in sludge
amended soils. MS. thesis. New Mexico State Univ., Las Graces.
NM.
Bellin, CJ^, G.A. O'Connor, and Y. Jin. 1990. Sorption and deg-
radation of pentacbloropbenol in sludge-amended soils. J. Envi-
ron. QuaL 19:603-608 (ibis issue).
Be venue. A, and H. P^ium, 1967. Pentachlorophenol: A discus-
sion of its properties and its occurrence as a residue in human
and animal tissues. Res. Rev. 19:83-134.
Buhler, D.R., M.E Rassmusson, and US. Nakaue. -1973. Occur-
rence of bexachlorobenzene and pentachlorophenol in sewage
sludge and water. Environ. Sti. Techno!. 7:929-934.
Casterfine. J.L., N.M. Barnett, and Y. Ku. 198S. Uptake, translo-
cau'on, and transformation of pentachlorophenol in soybean and
spinach plants. Environ. Res. 37:101-118.
Crosby, D.G. 1980. Environmental chemistry of pentachlorophenol:
A special report on pentachlorophenol in the environment p.
1052-1080. In Commission on pesticide chemistry. Dep. of En-
vironmental Toxicology, Univ. of California, Davis, CA.
DeLaune, R.D., R.P. Gambrell, and KS. Reddy. 1983. Bite of pen-
uchtorophenol in estuarine sediment Environ. PolluL Ser. B
6:297-308.
Fahrig, R., CA. Nilsson, and C Rappe. 1978. Genetic activ.
chlorophenols and chlorophenoi impurities, p. 325-338. In \
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Hamaker, J.W., and J.M. Thompson. 1972. Adsorption, p, 49-1
In CA.L Goring and J.W. Hamaker (ed.) Organic chemicals in
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Jacobs, L.W., GA. O'Connor, M.R. Overcaslv MJ. Zabek, and P.
Rygwiecz. 1987. Effect of trace organic* in sewage sludges on soil-
plant systems and assessing their risk to humans, p. 101-143. In
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Kloskowski, R., I. Schuenert, W. Klein, and F. Korte. 1981. Lab-
oratory screening of distribution, conversion and mineralization
of chemicals in the soil-plant system and comparison to outdoor
experimental data. Chemosphere 10:1089-1100.
Lagas. P. 1988a. Sorption of chlorophenols in the soiL Chemosphere
17205-216.
Lagas, P. 1988b. Behavior of chlorophenols in soil p. 264-266. In
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Poole, CF. 1978. Advances in silylation of organic compounds for
GC p. 152-200. In K. Blau and G.S. King (ed.) Handbook of
derivatives for chronutography. Heydon- and Son, Ltd,. London.
Ryan, JA, RJvt Bell, J.M. Davidson, and GA, O'Connor. 1988.
Plant uptake of non-ionic organic chemicali from soils. Chemo-
sphere 17:2299—2323.
Schafer. W., and VL Sandermann, Jr. 1988. Metabolism of pentach-
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Scheel, D., W. Schafer, and H. Sandermann. Ir. 1984. Metabolism
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Speitel, G.E. Jr., C Lu, M. Turakhia, and X. Zhu. 1989. Biodeg-
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Topp, E, L Scheunert, A. Attar, and F. Korte. 1986. Factors affectii
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-------
Reprinted from the Journal of Environmental Quality
Vol. 19. no. 3. July-September 1990. Copyright © 1990. ASA, CSSA. SSSA
677 South Segoe Road. Madison. WI 53711 USA
Adsorption, Degradation, and Plant Availability of 2,4-Dinitrophenol
in Sludge-Amended Calcareous Soils
G. A. O'Connor,* J. R. Lujan, and Yan Jin
ABSTRACT
2,4-Dinitrophenol (DNP) is a moderately weak acid that is ex-
pected to be highly labile (teachable and plant available) in high-
pH soils. The adsorption and degradation behavior of DNP in two
sludge-amended, calcareous soils was determined and used to ex-
plain DNP uptake by plants grown in the soils in the greenhouse.
The DNP adsorption was minor in both soils and was only slightly
affected by sludge. The DNP degradation was rapid in both soils
and was unaffected by sludge. Thus, despite limited soil adsorption,
plant uptake of DNP was minor in all crops and plant parts owing
to rapid soil DNP degradation. Even if a municipal sludge highly
contaminated with DNP was identified (an unlikely occurrence), con-
cerns over possible plant contamination should not limit sludge ap-
plications to calcareous soils at agronomic rates. Rapid degradation
will minimize opportunities for plant uptake of DNP from contam-
inated soils or leaching of DNP to groundwater, given careful water
management
2,4-DiNiTROPHENOL (DNP) is an active compound
of considerable phytotoxicity to both animals and
plants (Shea et al., 1983). It occurs as a waste contam-
inant originating from several industrial sources, and
may occur as a degradation product of other com-
pounds. Shea et al. (1983) reviewed the various bio-
logical activities of DNP in plant-soil systems, and
noted that DNP behavior is highly pH-dependent ow-
ing to its weak acid character (pKa = 4.09). Adsorpr
tion of DNP is favored by low soil pH, but degradation
is favored by high soil pH. They cautioned that careful
management is necessary if DNP is a predominant
component in land-applied waste materials.
The purpose of this study was to determine DNP
behavior in calcareous (high-pH) soils amended with
municipal sewage sludge and to determine the extent
to which DNP is available to food-chain crops grown
in such soils.
In high-pH, calcareous soils, DNP adsorption
should be minimal and DNP should be readily avail-
able to plants (Shea et al., 1983; Ryan et al., 1988).
Sludge additions to soils may, however, increase DNP
adsorption (Shea et al., 1983) and thereby reduce DNP
.plant availability. Studies of DNP uptake by plants
from sludge-amended soils have not been conducted.
This study was intended to supply such information
and to explain the uptake data in terms of DNP soil
behavior..
MATERIAL AND METHODS
Adsorption, degradation, and plant uptake studies were
conducted with two calcareous soils from New Mexico. The
All authors. Dep. of Agronomy and Horticulture, New Mexico State
Uni v Las Graces, NM 88003. Journal Article no. 1491, Agric. Exp.
Stn New Mexico State Univ. Although financial support for this
study was Provided m part by Cooperative Agreement CR-812687-
02 with the USEPA, this report has not had USEPA's required peer
and policy review and does not necessarily reflect the views ofthe
agency. Received 1 Aug. 1989. 'Corresponding author.
Published in J. Environ. Qual. 19:587-593 (1990).
Glendale clay loam [pH 8.0, 6.5 g organic C (OQ/kg, and
132 g CaCOj/kg] is classified as a fine-silty, mixed thermic
Typic Torrifluvent The Bluepoint sandy loam (pH 8.0, 1.2
g OC/kg and 30 g CaCOj/kg) is classified as a mixed, thermic
Typic Torripsamment Surface (0-15 cm) samples of both
were air-dried and sieved (<2 mm) before use.
Soils were amended with a secondary, anaerobically di-
gested sewage sludge from Albuquerque, NM. Air-dried
sludge was gamma-irradiated (i:57Cs, 10 kGy) to further re-
duce pathogens. Sludge was thoroughly mixed, sieved (<2
mm), and, stored in 30-L plastic containers for use in all
studies. A screen for priority pollutants (USEPA method
625), utilizing a gas chromatograph with a flame ionization
detector, revealed no detectable (1 mg/kg) DNP in the
sludge. Sludge rates included 0 and 45 Mg/ha equivalent (20
g/kg).
Adsorption Study
The batch equilibration technique was used at a soil/so-
lution ratio of 10:11 (w/v). Initial DNP soil concentrations
were 1.0, 10.0, 50.0, and 100.0 mg/kg. Each solution con-
sisted of 10 mL of nonradioactive DNP of appropriate con-
centration and 1 mL of I4C-DNP. The rates were chosen to
cover a wide range of concentrations, including rates much
higher than expected in normal municipal sludge-amended
soils. The median concentration of most organics in mu-
nicipal sludges is <10 mg/kg. Such sludges applied at ag-
ronomic rates would result in DNP concentrations in soil-
sludge mixtures of <0.1 mg/kg (Jacobs et al., 1987). Uni-
formly ring-labeled MC-DNP (specific activity 1.89 BqAg,
>99% purity, Sigma Chemical Company, St Louis, MO)
was present in each flask at 2.442 dpm/kg soil (disintegra-
tions per minute per kilogram). All DNP solutions were pre-
pared in 10-3 M NaCl (as background electrolyte) adjusted
to pH 8.0 with NaOH. This pH was chosen to match the
soil initial pH values to reduce pH variations of soil solu-
tions that might otherwise occur with the different DNP
(acid) concentrations.
The suspensions were shaken on a wrist-action shaker in
the laboratory (22 °C) for 18 h. Such vigorous agitation pro-
motes thorough mixing of the solute-containing solution
with soil, and normally promotes rapid attainment of ad-
sorption equilibrium. A preliminary time study suggested
that equilibrium was not attained in 18 h. We nevertheless
chose 18 h of shaking for convenience and to minimize deg-
radation effects (see below) on the disappearance of DNP
from solution. Extrapolation of a preliminary percent ad-
sorbed vs. time curve suggested adsorption equilibrium
would require about 36 h and would yield an additional 5%
adsorption.
After shaking, the suspension pH of the lowest (LO) and
highest (100.0 mg/kg) treatments was measured. Suspen-
sions were then centrifuged (900 X g) to separate superna-
tants. An aliquot (0.5 mL) ofthe clear supernatant was added
to scintillation cocktail and counted to 2 sigma percent error.
An external standard was used to correct for'quenching.
Counting efficiency was typically ~0.75.
Mass of DNP adsorbed was calculated as the difference in
mass of DNP originally added and that remaining after 18
h. Two soil-less blanks were included for each soil to detect
DNP losses from solution by mechanisms other than ad-
sorption. The soil-less blanks consisted of 100 mg DNP/L.
587
-------
588
J. ENVIRON. QUAU VOL. 19. JULY-SEPTEMBER 1990
Degradation Study
The DNP degradation was measured in a simple flow-
through incubation system. Air under vacuum was first bub-
bled through water to humidify the air and to minimize soil
drying. The humid air was then drawn through the air space
of a flask containing 60 g of soil amended with 0 or 45 Mg/
ha (0.02 g/g) sludge. The DNP was added at 3.7 mg/kg, and
consisted of reagent-grade DNP plus 18.5 kBq I4C-DNP/
flask. Carbon dioxide (including "COj) was trapped in a final
test tube containing 1 M NaOH. Soils were moistened to
water contents representative of moisture conditions-main-
tained in the greenhouse study of DNP plant uptake (see
below). The moisture contents (0.16 kg/kg for the Bluepoint
sandy loam and 0.38 kg/kg for the Glendale clay loam) were
slightly greater than field capacity. A total of 14 flasks for
each soil-sludge mixture allowed duplicate sampling at 0, 1,
2,4,8, 16, and 32 d. (Preliminary studies suggested most of
the DNP would degrade between 2 and 16 d, and that vol-
atilization losses of DNP were negligible.) Flasks were re-
moved from the train at appropriate times, and soils were
immediately extracted with a mixture of 30 mL methanol
and 30 mL 0.1 M Nad. The suspensions were shaken 3 h,
and then centrifuged (900 X g) for 30 min to separate equi-
librium solutions. The supernatants were transferred to glass
bottles, pne-milliliter aliquots were removed for I4C assay;
the remainder was refrigerated for subsequent high pressure
liquid chromatography (HPLQ. analysis. Approximately 1
mL of each supernatant was centrifuged (12-680 X g) for 2
min to further dear the samples. Aliquots (25 to 250 pL,
depending on DNP concentration) were injected for HPLC
analysis.
The HPLC conditions were: RP-18 column, 4.6 X 250
mm, mobile phase, methanol: 1 % acetic acid at pH 2.8 (50:50
v/v), flow rate 1 mL/min. Under these conditions, the re-
tention time was 8.4 to 8.8 min. Detector UV at 254 mm,
•limit of detection was equivalent to 0.0174 mg DNPAg soil
dry weight (~0.5% of initial).
Extracted soil was air-dried and ground (mortar and pes-
tle) for combustion. About 300 mg soil was weighed into
ceramic boats, covered with catalyst (activated alumina/cu-
pric oxide powder, 5:1 w/w), and combusted at 1000 °C for
8 min in an oxygen stream of 150 mL/min. Evolved "CO2
was trapped in a mixture of 10 mL ethanolamine-ethoxy-
ethanol (2:3, v/v) and counted by liquid scintillation. Com-
bustion efficiencies were typically ~95%. Samples corre-
sponding to Day 0 and 1 were combusted to determine mass
balance and extraction efficiency. Mass balance (I4C ex-
tracted 4- MC remaining in soil) averaged 101% across soils
for times Day 0 and 1. Extraction efficiencies averaged 65%
(SD - 3.9%) for the Glendale soil and 69% (SD - 7.7%) .
for the Bluepoint soil, with no sludge effect in either soil. A
preliminary incubation study utilized alcohol (95% MeOH)
as the extractant (Overcash et al., 1982), but yielded extrac-
tion efficiencies that varied from 91 to 52% depending on
soil and sludge treatment, even over short (I d) times. Ov-
ercash et al. (1982) also reported that the alcohol extraction
efficiency varied (80-120%) with DNP concentration. The
mixture of MeOH and NaCl used in this study seems pref-
erable to alcohol alone. The extraction efficiency was less
than ideal, but was reproducible, and similar for both soils.
Greenhouse Study
Sludge, soils (2 kg), and 2 g of fertilizer (18-48-0; 18-21-
0, N-P-K) were thoroughly mixed in a blender. Fertilizer
additions were intended to equalize fertility differences of
soils and sludge treatments.
Six DNP rates (0. 0.1, 0.5, 1.0, 5.0, and 10.0 mg/kg) were
imposed on the soil and sludge treatments. The rates were
chosen to avoid possible phytotoxic (>20 mg DNP/kg) ef-
fects on the crops (Overcash et al., 1982) and to cncompa
both reasonable and excessively high DNP rates expec
from additions of sludge containing priority pollutants
concentrations <10 mg/kg (Jacobs et al.. 1987).
The 24 treatments (six DNP rates, two sludge rates, two
soils) per crop were replicated twice. The decision to invest
' in several DNP rates rather than more replicates was made
to detect the nature of plant response trends; with DNP rate
more precisely than with fewer DNP rates and more repli-
cates. With the same total number of pots, the estimate of
variance is the same, but the precision of the estimate of the
mean response is better with more DNP rates than with more
replicates; Each DNP treatment, except the control, con-
sisted of nonradioactiye, reagent-grade DNP plus uniformly
ring-labeled I4C-DNP (specific activity 1.89 Bq/kg, >99%
purity; Sigma Chemical Company, St. Louis, MO). The 0.1
mg/kg rate consisted of only 14C-DNP. The amount of I4C
added to each pot was the same (11.100 dpm/kg, SD =- 10%).
Soil-sludge-fertilizer mixes were spread uniformly on a tray
covered with aluminum foil. The DNP solutions (labeled
and uniabeled) were uniformly applied to the soils with a
syringe. The soils were mixed for several minutes by hand
and then transferred to plastic pots. The mass of prepared
soil in each pot was 1.8 kg. The excess soil (200 g) was
retained for DNP analysis (see below) and for covering seeds.
Soils were seeded with fescue (Festuca arundinacea
Schreb., *KY 31'), carrot (Daucus carota L., 'Nantes'), lettuce
(Lactuca saliva L., 'Black Seeded Simpson'), and chile pep-
per (Capsicum annuum L., 'Espanola' improved). Seeds
were sprinkled uniformly on the soil surface, and covered
with a few millimeters of dry soil reserved from each pot.
After seeding, pots were watered to pot-holding capacity
(0.16 kg/kg for Bluepoint, 0.38 kg/kg for Glendale). Each
was then covered with newspaper to minimize evapora
during germination. Some leaching occurred from several
pots in the initial watering, but leachate was collected in
plastic saucers and was reapplied to the respective pots.
Plants were watered approximately every 2 d to return them
to initial pot-holding weights. No subsequent drainage oc-
curred. '
Fescue was grown for 32 d^ lettuce for 43 d, carrot for 70
d, and chile pepper for 90 d. Natural light was supplemented
as needed by eight 400-W sodium lamps to supply 16 h of
light Temperatures, in the greenhouse varied from 16 to
35 °C during the experiment.
Fescue, lettuce, and chile were harvested by cutting plants
about 3 cm above the soil surface. Carrot plants were re-
moved with the main tap root intact Tops and roots were
separated with a razor blade and were washed with distilled
water until no visible soil particles remained. The roots were
then peeled, weighed, and stored in paper bags. Peels were
also weighed and stored in bags. Plant fresh weight yields
were recorded immediately after harvest. All plant material
was then dried (50 °C) for a minimum of 12 h. Dried plant
material was weighed, ground, and stored in'plastic bags for
later analysis.'
Soils retained from the initial mixing and labeling with
"C-DNP and dried plant samples were assayed for MC by
combustion as in the degradation study. Approximately 50
mg plant tissue or 300 mg ground (mortar and pestle) soil
was combusted.
Analysis of plant material for intact (parent compound)
DNP was performed by an independent analytical firm using
approved USEPA extraction and clean up procedures. De-
tection was by gas chromatography with flame ionization
detection. Given the limited mass of plant tissue available
for extraction (~9 g, reps combined), the limit of detection
for DNP was 0.146 mg/kg.
Treatments were arranged in a random complete block
design. Analyses of variance were conducted for the varia-
b •;:: crop, plant parts, DNP rates, sludge treatment, and
-------
O'CONNOR ET AL.: 2.4-DINITROPHENOL IN SLUDGE-AMENDED SOILS
589
bioconcentration factors based-pn MC. An LSD test was per-
formed for variable means exhibiting significant differences
in the analysis of variance. -
Results and Discussion
Adsorption
Adsorption data for the sludge-amended calcareous
soils are summarized in Table 1 along with the pH
values of the equilibrium suspensions measured. The
pH values of intermediate DNP treatments are as-
sumed to be similar. The paste pH values of both soils
in distilled water are 8.0. The lower pH values of the
adsorption equilibrium suspensions are primarily
caused by the background salt (IQ-3 M NaCl). rather
than acidifying effect of the DNP. Sludge-amended soil
pH values were consistently lower than unamended
soil pH values, but the effect on adsorption was prob-
ably minor. The pKa of DNP (4.09) is at least 2.6 units
lower than the soil pH values, so <0.25% of the DNP
exists as undissociated acid in the most acidic treat-
ment (pH 6.7) and <0.14% at pH 7.0.
The DNP adsorption was minor in both soils owing
to their negative charges and the dominance of the
anionic form of DNP. The Glendale soil.exhibited
positive adsorption (Freundlich K = 0.67 and 0.35.
unamended and amended, respectively) consistent
with adsorption of other weak acid organics 2 4-D
and 2,4,5-T (O'Connor et al., 1981) and pentachoro-
phenol (Belliri et al., 1990) on this soil. Despite the
soil's .negative charge, adsorption of weak acid com-
pound occurs, and is primarily associated with the
organic fraction (O'Connor and Anderson, 1974)
Sludge additions slightly decreased DNP adsorption
(Table 1), but had no effect on phenoxy herbicide ad-
sorption by the Glendale soil (O'Connor et al, 1981)
The DNP was repelled (negatively adsorbed) from
the Bluepomt soil at all DNP concentrations. This soil
is extremely low m organic matter (1.2 g OC/kg) and
apparently offered no positive adsorption sites. Sludge
addition reduced DNP repulsion (Table 1), but DNP
was negatively adsorbed in almost all treatments.
Given the minimal adsorption of DNP in both soils
in the presence and absence of sludge, DNP mobility
is expected to be great. Careful water management
would be necessary in DNP-contaminated soils to
avoid groundwater pollution (Shea et al., 1983) The -
DNP activity toward plants would be maximal in both
soils and unaffected by sludge additions. Almost all of
the chemical would remain .in solution available for
plant uptake, or other removal processes e g. degra-
dation and leaching. •«>•.&,
Degradation
The DNP degradation in the sludge-amended cal-
careous soils (Fig. 1 and 2) is presented as percent DNP
remaining (as determined by HPLC) plotted as a func-
tion of time. The data have been corrected for ex-
traction efficiencies. Similar plots (not presented) of
percent MC remaining in methanol/NaCI extracts
matched closely the data in both figures. Thus I4C
extracted by methanol/NaCI could have served'as a
surrogate for intact DNP, contrary to results we have
obtained for pentachlorophenol (Bellin et al 1990)
A semilog plot of the data was used to identify first-
order degradation kinetics. Neither soil, however
demonstrated the single linear decrease in DNP re-
maining with time consistent with simple first-order
kinetics. An initial linear decrease, lasting a few days
was followed by another linear decrease, of much
greater slope, until only a few percent of DNP re-
?lai™,^ls-J and 2)- Half-J'ves could be estimated
for DNP in both soils (~5 d in Glendale and ~9 d
m-BIuepomtX but are misleading. Much less DNP re-
mains in the soils after 2 half-lives than the 25% ex-
pected from first-order kinetics. A more meaningful
description would be that DNP has almost completely
degraded in the Glendale soil in 8 d, and in the Blue-
pomt soil m 16 d. There was no significant effect of
sludge on DNP degradation in either soil.
The DNP degradation in these soils was more rapid
^,^noted by other investigators. Overcash et al
(1982) reported 62 to 66% DNP loss after 4 wk in the
acid Davidson clay loam [clayey, kaolinitic, thermic
(oxidic) Rhodic Paieudults]. The USEPA (1979) re-
P°rts a half-life of 50 d for DNP. Miller (1977) clas-
sified phenolic compounds as slowly degradable. The
DNP can be rather persistent in both soils and aquatic
systems, but decomposition by certain strains of bac-
tencide and by a fungus has been demonstrated (Shea
et al., 1983). The DNP is bactericide at high concen-
trations and low pH, thus, both these factors influ-
enced toxicity and metabolism. The optimum pH for
microbial decomposition (by reduction of the nitro
groups to amino groups, followed by oxidative deam-
ination, or by the release of a nitro group as nitrite)
is near neutrality (Shea et al., 1983). Decomposition
by release of NO2 by Corynebacterium simplex was
maximal at pH 8.0 (Gundersen and Jensen, 1956)
Treatment
DNP
mg/kg
1.0
10.0
50.0
100.0
Sludge
Mg/ha
0
45"
0
45
0
45
0
45
Ci
0.93
0.93
9.1
9.1
45.5
45.5 '
90.9
90.9
Glendale
Ce
-mg/L — —
0.54
0.70
6.9
7.4
39.1
41.1
82.6
87.5
mg/kg
0.43
0.26
2.5
2.0
7.2
4.6
9.5
3.9 .
PH
7.0 .
6.8
7.0
6.9
Ce
mg/L
1.01
0.93
9.7
9.3
50.5
47.3
99.2
95.9
Blueprint
x/m
fflg/kg
-0.09
0.00
-0.6
-0.2
' -5.6
-2.0
-3.2
-5.6
PH
7.1
6.9
7.2
6.7
t a - inin-a. concentration; Ce - equiHbrium concentration; */m - amoun« adsorbed per Uni, mass of soi!. AH valuei are ,he average of ,wo rep.ica.es.'
-------
590
J. ENVIRON. QUAL.. VOL. 19. JULY-SEPTEMBER 1990
• Bluepoint Soil
T
2 4 6 8 10 12 14 16 18 2O 22, 24 26 28 3O 32
Time
: 1
Fig, 2. The DNP degradation in Glendale soil (means ± I SD).
Photochemical hydrolysis of DNP has not been
demonstrated, but has been suggested as possible
(Shea et al., 1983). Flasks in our study, however, were
wrapped with aluminum foil to exclude light, and
there was no evidence of photolysis in the adsorption
study (exposed containers). Volatilization was also
considered unlikely based on previous work (Overcash
et al., 1982) and DNP's low Henry's constant (2.7 X
10-«, dimensionless). Lack of volatilization was con-
firmed in a preliminary study with soil-less blanks;
Thus, the rapid dissipation of DNP in Our soils was
8
Time(days)
1O
12
14
16
attributed to microbial activity, favored by the initial
low (3.7 mg/kg) DNP concentration and high pH.
The rapid degradation of DNP in high pH soils has
important environmental consequences. The essen-
tially complete degradation of DNP in 8 or 16 d (Glen-
dale and Bluepoint, respectively) means that little
chemical remains in the soils long enough for signif-
icant plant uptake. Also, DNP is weakly, or negati vely
adsorbed in high-pH soils and could threaten ground-
water if the soils are leached excessively. Careful water
management that ensures chemical residence times of
-------
O'CONNOR ET AL.: 2,4-DlNITROPHENOL IN SLUDGE-AMENDED SOILS
591
Table 2. Effect of DNP and sludge.on yieldsf of crops in high pH soils in the greenhonse.
Glendaie
Treatment
DNP
rag/kg
0
0.1
0.5
1
5
10
Sludge
Mg/ha
0
45
0
45
0
45
0
45
0
45
0
45
. Fescue :
Glendaie Bluepoint
2.60
3.80
3.45
4.70
4.50
4.00
4.60
4.20
4.35
4.30
3.15
3.65
•iiii ,
1.65
3.05
Z80
Z8S '
3.20
Z80
3.80
3.15
3.20
3.30
3.00
ZOO
Lettuce
4.55
4.00
5.50
5.90
. 5.65
4.25
S3S
6.00
3.55
5.55
5.15
4.65
Tops
, Carrot
Roots
g dr^ wt/pot —
4.30 2.45
4.15 1.80
3.85 1.15
4.10 0.85
4.85 2.60
4.40 1.20
4.80 2.25
4.95 1.10
5-50 1.70
6-15 1.55
5.60 2.00
4.15 MS
t Mean of two replicates. . " ' ~
Peels
1.30
0.80
0.75
0.60
1.30
0.75
1.75
1.10
1.55'
0.95
1.50
0.75
Pl*nt
. 6.25
8.15
6.70
6.95
6.35
6J5
8.05
8.40
5.30
630
6.05
6.90
Chile
Fruit
1.35
1.75
Z25
1.55
1.65
Z85
4.90
4.20
Z3S
3.05
3.00
4.55
1
Table 3. The DNP bioconcentrarion factorst (dry-wt. basis) based on "C and intact DNP analysis.
Fescue
Glendaie
Bluepoint
Lettuce
DNP Sludge
rag/kg Mg/ba
0 0
45
0.1 0
45
0.5 0
45
1.0 0
. 45
5.0 0
45
10. 0
45
"C
_
:_
0.432
0.295
0.781
0.472
0.840
0.838
1.46
0.542
1.07
1.05
DNP
NP
0.068
-------
592
J. ENVIRON. QUAL.. VOL. 19, JULY-SEPTEMBER 1990
•Table 4. Various estimates of DNP bioconcentration ftctors (fresh-
wt. basis) at the highest initial DNP soil concentration (10 mg/
kg).
Bioconcentration factor
Crop
Fescue (Gtendak)
Fescue (Bluepoinl)
Lettuce
Carrot tops
Peels
Roots
Chile foliage
Fruit
"C
0.21
0.64
0.034 '
0.01
0.003
0.000
0.012
0.001
DNP
<0.040
<0.045
1) bioaccu-
mulation, whereas the BCFs for the other crops imply
much less (passive, BCFs < I) bioaccumulation.
The BCFs calculated on the basis of I4C are common
in the literature, but are misleading as the I4C is as-
sumed to represent intact parent compound. If the I4C-
labeled compound degrades in soil or is metabolized
within the plant, I4C contents of plant tissue falsely
describe parent compound contents. Some 14C was de-
tected in the plant tissues from control treatments.
Because there was no MC in the control soils (and no
DNP volatilized in the degradation study), I4C in plant
tissue in the controls was regarded as I4CO2 released
from l4C-tagged soils. This contamination was ac-
counted for by subtracting the average I4C content
(~0.300 dpm/kg, effective BCF = 0.027) of all con-
trols from the |4C content of each treatment. Never-
theless, net I4C contents of plants may still represent
|4C species accumulated other than 14C-DNP. Degra-
dation of DNP was rapid in both soils, being almost
complete in 8 d in the Glendale soil and 16 d in the
Bluepoint soil. Thus, even before the plants germi-
nated (5-20 d after planting), significant reductions in
actual soil DNP concentrations occurred, especially in
the Glendale soil. Fescue germinated 5 d after seeding
and was harvested 32 d after seeding. The three food-
chain crops grew for longer times, but germinated
more slowly. Lettuce germinated 20 d after seeding
and was harvested on Day 54; carrot germinated on
Day 12 and was harvested on Day 70; chile germinated
on Day 20 and was harvested on Day 90. Given these
growth characteristics, one would expect greatest con-
tamination in fescue, less in lettuce and carrot, and
much less in chile. The MC-based BCFs in Table 3
generally reflect the expected trend.
Evidence exists that DNP is metabolized within
plants (Berlin et al., 1971; Klepper, 1979). Thus, even
if intact '"C-DNP were accumulated by plants, I4C
contents of harvested plant material would not be clear
evidence of DNP in tissue. That plant metabolism can
completely obviate meaningful interpretations of
BCFs based on |4C was clearly demonstrated in studies
similar to this utilizing diethylhexyl phthalate (DEHP)
as the target chemical (Aranda et al., 1989). Based on
all of the above discussion,'it is clear that BCFs based
on |4C represent very conservative, and probably er-
roneously high, values.
Attempts to improve on the I4C data by analyzing
for intact DNP were disappointing. Because of limited
plant tissue, the limit of detection (LOD) for DNP in
extracted plant tissue was only 0.146 mg/kg dry
weight Replicates were combined to yield plant tissue
for analysis, but only 8 to 9 g dry tissue was obtained
(Table 2) resulting in the stated LOD. When less plant
tissue was available, the LOD, (per gram tissue) was
higher. Thus, BCFs calculated on the basis of detected,
intact DNP were limited to LOD values and usually
exceeded the BCFs based on I4C (Table 3). Exceptions
were for the highest initial DNP treatments. (10 mg/
kg). Particularly for fescue grown in Bluepoint soil,
actual DNP analysis showed the uC-based BCFs to be
grossly in error. . .-.
, Various estimates of DNP bioconcentration factors
for each plant and plant part are given in Table 4. The
BCFs are expressed on a fresh-weight- basis, because
all of these plants would be consumed fresh. Data for
unamended and sludge-amended soils have been av-
eraged because there was no significant effect of sludge
on BCFs across all treatments. The BCFs are presented
for the highest DNP concentration (10 mg/kg) because
DNP accumulation is likely greatest at this concen-
tration, and because DNP determinations of intact
parent compound are useful only for this DNP rate.
The last column represents our estimate of the likely
maximum BCFs. The values are the smaller (but more
reasonably accurate) than the BCFs based on I4C or
actual DNP determinations. Even this estimate is con-
servative as no intact DNP was indicated in any GC
chromatogram. Further, our degradation data suggest
minimal DNP existing in the soils long enough for
plant uptake. Plant metabolisrn of even .the small
amounts of DNP accumulated could also occur.
All of the BCFs (Table 4) are low, suggesting min-
imal plant contamination with DNP. Contamination
is minor regardless" of sludge treatment, at DNP con-
centrations an order of magnitude (or more) higher
than expected under normal conditions. Calcareous
soils more highly polluted (>20 mg/kg) with DNP
would likely result in plant phytotoxicities (Shea et al.,
1983). If the soils are acid, phytotoxicities-occur- at
lower DNP concentrations (Simon, 1953).
Thus, even if a municipal sewage sludge highly con-
taminated with DNP was identified (an unlikely oc-
currence), concerns over possible plant contamination
should not limit sludge application to calcareous soils
at agronomic rates. Degradation of DNP in these soils
would minimize the amount of chemical, available for
plant uptake. Careful water management for the first
30 d or so following DNP additions to high pH soils
would also minimize the amount of chemical available
for leaching to groundwater.
ACKNOWLEDGMENTS
The assistance of J. Aranda, L. Tinguely, and.C. Bellin in
the greenhouse study, of Dr. M. Southward in statistical anal-
ysis, and of J Aranda and Dr. W. Mueller in performing the
HPLC analysis for DNP is gratefully acknowledged.
REFERENCES
Aranda. J.M., G.A. O'Connor, and G;A. Eicemari. 1989. Effects of
sewage sludge on diethylhexyl phthalate uptake by plants. J. En-
viron. Qual. 18:45-50."
Bellin. C.A.. G.A. O'Connor, and Jin Van. 1990. Sorption and deg-
radation'of pentachlorophenol in sludge-amended soils. J. Envi-
-------
O'CONNOR ET AI-: 2,4-DlNITROPHENOL IM SLUDGE-AMENDED SOILS
593
ron. QuaL 19:603-608 (this issue).
Berlin, J., W. Barz, H. Harms, and K. Harder. 1971. Degradation
of phenolic compounds in cell cultures. FEES Lett. 16:141-146.
Gundersen, K_ and HI. Jensen. 1956. A soil bacterium decom-
posing organic nitrocompounds. Acta Agric. Scand. 6:100-114.
Jacobs, Lw.~G.A. O'Connor, MR. Oyercash, MJ. Zabek, and P.
Rygwiecz. 19S7. Effect of trace organics in sewage Sludges on soil*
plant systems and assessing their risk to humans, p. 101-143. In
A.L. Page et aL (ed.) Land application of sludge. Lewis PubL,
Chelsea, MI.
Klepper, I~A., 1979. Effects of certain herbicides and their combi-
nation on nitrate and nitrite reduction. Plant PhysioL 64:273-
275.
Miller, R.H. 1977. The soil as a biological filter, p. 70. In S. Sopper
and L.T. Kardos (ed.) Recycling treated municipal wastewater and
; sludge through forest and cropland. The Penn. State Univ. Press,
University Park, PA.
O'Connor, G.A., and J.U. Anderson. 1974. Soil factors affecting the
adsorption of 2,4,5-T. Soil Sci. Soc. Am. Proc. 38:433-436.
O'Connor, G.A^ B.C. Fairbanks, and E.A. Doyle, 1981. Effects of
sewage sludge on phenoxy herbicide adsorption and degradation
, in soils. J. Environ. QuaL 10:510-515.
Overcash, MR., J.B. Weber, and ML. Miles. 1982, Behavior of
organic priority pollutants in the terrestrial system: di-N-butyl
phthalate ester, toluene, and 2,4-dinitrophenoL Rep. 171. Water
Resources Res. Inst, Raleigh, NC.
Ryan, M^ R.M. BelL J.M. Davidson, and GA. O'Connor. 1988.
Plaint uptake of non-ionic organic chemicals from soils. Chemo-
sphere 17:2299-2323. .
Shea, PJ., J. Weber, and MR. Overcash. 1983. Biological activities
of 2,4-dinitrophenol in plant-soil systems. Residue Rev. 87:2-41.
Simon, E.W. 1953.Mechanisrnsofdinitrophenoltoxicity.Biol. Rev.
28:453^479. »
VS. Environmental Protection Agency. 1979. Water related envi-
ronmentatfate of 129 priority pollutants. VoL 1 and 2. EPA-440/
4-79-0296. NTIS, Springfield, VA.
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Chemosphere, Vol.17, No.12, pn 2299-2323, 1983 0045-6535/88 $3.00 +-.OO
Printed in Great Britain .Pergamori Press nlq
PLANT UPTAKE OF NON-IONIC ORGANIC CHEKICALS 7ROM SOILS
J.A. Ryan1*, R.H. Bell2, J.M. Davidson3, and G.A. O'Connor*
1.UEL, USEPA. Cincinnati, Ohio. *S268.
Z.Environaantal Ad»i»ory Unit, Univaraity of Liv«rpool, U.K.
, 3.Univaraity of Florid*. Gainnvilla, Florid..
«.N«w Naxico Stata University, Laa Crjjeaa, Nw Naxico.
i *v*iaMfei
Mcthooblogias utiliiirw tiapl* propartia* of chaaricals - half-Ufa . log octanol-vatar partition coafficiant (log K ) and
H«f»ry'« Lav constant (He) - art oavalopad to scraan organic chaaricals fcVpotantial plant mtaka? °"
XNTRODUCTIOM
Early in 1983, the American Chemical Society's Chemical Abstract Service
registered its 6,000,000th chemical. The Toxic Substance Control Act Inventory
list 63,000 chemical substances whose manufacture, processing and ultimate use
for commercial purposes has occurred in the United States since January, 1975
(TSCA inventory, USEPA, 1985). Additionally, the number of synthetic organic
chemicals used and disposed of by society is increasing at a rate of about 1000
new chemicals per year, (Loehr and Malina, 1986) . Thi« endless supply of compounds
together with the variety of -reactions they can undergo in the environment makes
describing their environmental impact exceptionally challenging.
Of the possible locations for the disposal of wastes - surface waters,'
atmosphere or land the latter represents a common location for waste disposal as
well as an opportunity to manage wastes with minimal environmental impact. The
object of the land disposal practice is to degrade, immobilize, and/or transform
the wastes into beneficial, or at least non detrimental constituents. There are
over 200 industrial waste 3and treatment sites in the United States, and a larger
number of land treatment sites for municipal wastewater and sludge (Loehr and
Malina, 1986). Land disposal of wastes has increased during the past decade and
is projected to continue to increase in the future (Loehr and Malina, 1986).
The study of organic chemicals in the soil environment has been dominated by
agricultural chemicals (e.g., insecticides, nematicides and herbicides) and
specific compounds that persist in the soil (e.g., PCB's, PBB's etc.). This
narrow perspective probably occurred because of the prevalence of agricultural
chemicals in soil, complexity of reactions, large number of compounds, and cost
associated with organic analysis. Specific compound attention has been propagated
by the formation of lists of specific compounds, such as the organic priority
pollutant list o£ 1976. Even with this narrowing of focus, the cost associated
with a chemical by chemical investigation is prohibitive. The approach therefore
2299
-------
2300
has been to utilize physicochemical parameters, or to group compounds on the'basis
of their chemical or physical properties and study selected compounds from each
group. Clearly, we must insure that the grouping of compounds is correct and that
the factors used in the groupings predict the behavior and impact of compounds
not studied.
The following attempts to provide a frameworJc which uses physicochemical
parameters to evaluate potential plant uptake of neutral or weakly ionized organic
chemicals from soil. The procedure does hot predict plant concentration of
organics in a field situation, but provides a procedure for grouping chemicals
by their relative potential for plant uptake. As such, it should allow compound
screening for their likelihood for plant uptake and, therefore, justify
experimental evaluation as well as identify chemicals of low concern where testing
say be counterproductive. It should also reveal where information is needed to
confirm the screening model.
BEHAVIOR OF ORGANIC CHZKICAL*
Many processes impact organic chemicals in the soil environment. The sum of
these actions determine the compounds environmental impact (Figure 1). Factors
such as pH, CEC, OM content, clay content and soil water content all impact the
rate and extent of these processes (Goring and Hamaker, 1972). In a given
situation (soil and environmental conditions) however, the processes are dependant
upon the physical and chemical properties of the, chemical. The characteristics
of a chemical that determine its distribution between vapor, solid, liquid and"
adsorbed phases in the soil, and its degradation rate become the characteristics
that determine its environmental fate and impact upon plants. These processes
determine not only the fora of the compound that is present, but also the speed
at which the compound moves or spreads through the soil and atmosphere to achieve
its impact. The importance of .each of these .processes will be discussed
separately.
FIGURE 1
SOIL TRANSFORMATIONS
-------
2301
Degradation - -
• Plant uptake of most chemicals is • concentration dependent, therefore a
compound's persistence can alter its ultimate fate and environmental impact.
An assessment of the half-life of a particular compound is a.relatively simple
way of limiting the number of soil borne organic ..compounds that need to be
considered as likely to impact 'a plant grown in contaminated soil. The
concentration of synthetic organic compounds.in the 'soil decrease with time,
providing no further additions occur.' Processes contributing to the decrease with
tine are.biological and/or chemical degradation. These processes have been shown
to be dependent on -soil and environmental 'factors (ie., temperature, water
content, soil pH, and organic C), . (Hamaker, 1972). Withput the quantitative
information necessary .to describe the functional dependence of degradation on
these factors, it has been shown that degradation of a specific organic chemical
can be described by a first order rate constant, p, (Nash, 1980; Rao and Davidson,
1980; Jury et al., 1983; Gillett,1983). This parameter is usually measured by
determining the fraction of an applied chemical remaining after a time t according
to Equation 1 : . . '. .' '•
M(t).= M(0) exp <~/*t> C1]
where M(t) is the quantity of the compound remaining in the soil at time t. The
half-life, T1/2, of a compound is defined as-the time required for one half of the
concentration of the chemical at any point in time to be lost from the soil. This
is related to the rate constant (ju) by : .
T/ - Q-69? , .
1/2 ~ M , ' ' [2]
Half-lives of many chemicals have been published (USEPA, 1979; Jury et al.,
1983; .smith and Dragun, 1984). Unfortunately, reported values of n may vary.
enormously because measured half-lives of compounds in the soil do not always
reflect degradation. Often losses include other pathways (i.e., volatilization,
leaching, etc.). Additionally, water content, microbial population, and
temperature can significantly influence the rate of loss thus, a chemicals life
r.ay vary from soil to soil. Half-lives are reported in Table 1 from data in USEPA,
1979. compounds are distinguish from one another on the basis of half-life in the
soil: less than 10 days, (Class A); between 10 and 50 days, (Class B);.and greater
than 50 days,-(Class C) . Gillett considered compounds of T1/2 greater than 14 days
of sufficient stability to be of concern (Gillett, 1983). The impact of chemical
half-lives on concentration of a pollutant in the soil over time is shown in
Figure 2. Pollutants with half-lives of less than 10 days, for example, are
reduced to less than 0.10% of their original concentration after 100 days in the
soil, m contrast; pollutants with half-lives of greater than 50 days are still
present at >25% of their original concentrations after 100 days. Their impact,
and relative potential for plant uptake, are much more pronounced than that for
-------
2302
compounds with half lives of less than 10 days.
6-
o
HALF-TIME (days)
100
20
40 60
TIME (days)
80
100
FIGURE 2 EFFECT OF CHEMICAL ' HALF LIFE AKD TIHE Oil
FRACTION REMAIHIHG
The average concentration present during the plant growing period can be
calculated by integration of Equation 1 between the limits 0 and t (growth, period)
and dividing by t. Assuming a growth period (i.e. 50 or 100 days) the effect of
half-life on the average soil concentration as a fraction of the amount originally
applied illustrates that the limits for classification of compounds based on half-
lives are arbitrary (Figure 3). The length of exposure (i.e. plant growth period)
and relative average exposure must be specified before compounds can be classified
by their half-lives. For example, our use of 10 and 50 day half lives as'
classification end points was based on a 100 day growth period and relative
average exposures of 0.15 and 0.5. Using the same half-life end points but a 50
day growth period means relative average exposures of 0.3 and 0.7.
20
40 60
HALF-LIFE (DAYS)
80
100
FIGURE 3
AVERAGE SOIL CONCENTRATION. VS HALF LIFE
FOR 50 AND 100 DAYS OF GROWTH - '
-------
2303
TABLE l. Log K^; Half-life and He for the Priority Pollutant*
t**r« log
Caipaini
PESTICIDES
20.Acroltin
22.Chtordv«
24.006
26.0i«ldrin
28.Endrfn
JO.K«pt»ehtor tpexidt
32.lind*nt
34.TO*
MLTOUXIMATB «MOnS
36i.Araditer 1016
36c.Arodilor 1232
3&t.Arochlor 1248
36g.Aroehlor 1260
.!!!.?~.l!/?...!!!.
-0.09 t 2.8E-03 21.Aldrin
4.3 . C 3.9E-03 23.000
5.69 I 9.0E-04 25.00T
4.5 C 1.7E-05 29.ltaptacnlor'
4 ...04
,
3.72
3.lMp*Mron.
3.SS
5.9 * *.2£.02
3.8 t 3.0E.04
1.70 , «J nd
3.85 c 2.1E.01
*.3B C 8.61-01 36b.*roditor 1221
*-54 c 2.1E«00 36d.*rodiior 12*2
!1i r J'«"JJ »*•*««*"«• «*
6.11 C 2.9E-01 37.2-chlom0itlMlm
4.11
c 1.3E-02
C 5 SE-02
«•<* e i:«-oT
*.« c 1.J£-02
IALOKWTED AllPMTIC
38.Chlora»tlunt 0.91 e
40.TriehloroBtthant .1.9 |
42.Chlorb*thm \ .54 |
44.1,2-dtdiloriMtlww 1.48 •
46.1.1,2-trietilerwtiwnt 2.17 nd
48.Hcjuehlore*than» ' 4.62 nd
50.1,1-dieiileratlMnt 1 48 A
SZ.THehlerettiwnt 2.29 A
54.1.2-dieMofOpfep«» 2;28 nd
S6.N*iuehloretutadi«n» 3.74 e
58.«roBMtlunt 1.10 I
60.DibranchleraBtthan* 2.09 nd
62.0ichterodifluoraHthint 2.16 C
•AUJGEJUTES ETKIS •
64.1U(ehloroiithyl>«thw -0.38 A
66.li«<2-ehtoroiMpropyl)*tlwr 2.M nd
6S.*-chloroph«nyt ptMnytctlwr 4.08 nd
70.tt*(2-ehlore*thaxyXMthan» 1.26 C
1.6M1
1.2£-01
6.1t-01
3.8E-02
3.1E-01
39.Dthyl
10).0f-n-propyl nitreuBin*
103.3.3-dichtoretanzidim
105.AcrylonitrU«
Ml
'
0.06
1.31
fl'25
.«E-03 95d.U0itiuil«nt
nd
nd
A
nd
nd
SM.Ind»t123-aapxr«»
100.0lph.vl nitrourint
102.8«tidfn.
10*'1'2-<»"ll*»lh»*«1'»
4.07 C
3.3T e
5.33 C
s.6i e
6.84 C'
4.04 C
7.66 C
4.8E-03
2.0E-02
4.0E-04
4.1E-05
nd
4.9E-01
nd
2.57 nd nd
i.gi A „,
3.03 nd nd
-------
2304
Partition
^ considerable research data exists on the equilibrium between an organic sorbed
to the soil and that in the soil-water phase. For simplicity, this is often
expressed as a linear sorption isotherm (Karickoff, 1981):
where Cs is the sorbed concentration (g/kg soil), q is the solution concentration
(g/m soil solution) and Kd (m3/kg) is the slope of the sorption isother* or
distribution coefficient (Kay and Elrick, 1967). Equation 3 assumes complete
reversibility and equilibrium between the two pha.es, which may not strictly occur
for some chemicals. Di Toro and Horzempa (1982), reported that the sorptive
process of 2,4,5,2-,4•,5'- hexachlorobiphenyl con.i.ted of both reversible and
strongly bound components, such bound residues could not be extracted by normal
analytical techniques, but could be detected by radiolabelling. similar findings
have been reported by others working with herbicides and chlorobenzenes (Khan,
1982; and Scheunert, et al., 1985) and may require the above mathematical approach
for sorption be modified to account for bound residuals.
in soil, .and sediments, where the clay content is relatively low, pollufant
sorption occurs primarily on the organic fraction of the .oil, (Hamaker and
Thompson, 1972; Rao and Davidson, 1980). The degree of sorption of the non ionic
organic pollutant is then dependant upon the organic carbon content in the soil
or sediment. Variation between materials, which otherwise exhibit a wide range
of physicochemical properties, can then be reduced by defining an organic carbon
distribution coefficient (K^):
KKjj . ,* ,
__ " _ »
•foe [4]
where Kd is the slope of the sorption isotherm in ,Vkg, and foe is the organic
carbon fraction in the soil or sediment, (Means, et al., 1982). This assumes that
all organic matter has the same chemical structure.
K is defined as the ratio of the organic chemical concentration in octanol
to that in water, when an aqueous solution of the organic chemical is mixed with
n-octanol and then the organic chemical allowed to partition between the two
Phases (Dawson, et al., 1980). There have been many investigations into the
relationship between K,, and K^. Briggs (1973) for example'reported:
iogK^ - 0.524 logK,,,, + 0.62
from his work with 4 agricultural .oil. and 30 chemicals chosen for their wide
range of properties, similar relationship., see Equation'. 6, 7, 8, 9, and 10
have been reported ( Means, et al., 1982; Schwarzenbach and Westall, 1981; Rao'
«t al., 1982; Karickhoff, 1981; and Brown and Flagg, 1981 respectively).
logK^ - logK^, - 0.317
[6]
- 0.721 lQgKM.+ 0.49
-------
2305
1(?9Koc ' 1-029 logK^, - 0.18
logK,,. = 0.989 logK,,,, - 0.346 . . '
logK^ - 0.937 logK^ -0.006 ' '" [1Q]
The relationships are surprisingly similar to on. another considering they cover
over 100 chemicals, as well as a large number of soils and sediments (Figure 4)
Thus when the sorption value of a particular pollutant in a particular soil is
not available, advantage can be taken of the relationship between the organic
carbon distribution coefficient (K^) and the octanol water partition coefficient
(K^,) of the chemical. Recently, a nonempirical measurement (first-order molecular
connective indexes) calculated from the non-hydrogen part of the molecule ha. been
shown to predict the KO- of organic compounds with great success (Sabljlc 1987)
As these calculated values for various organic compound, become available it will
allow for their use in place of K^ or
log Kov
FIGURE
RELATIONSHIP BETVEEN log Koc AND log
To have greatest impact upon plant uptake, the organic compound must stay
witlun the vicinity of the plant root, and not be quickly Reached away by mass
flow. For example most residual .oil-acting herbicide, have Kd value, in the range
of 1-20 with value, up to -40 being satifactory for mo.t soil applications
(Graham-Bryce, 1984), compounds with *,-. of gr.at.r tbmn 1000 becom. inactivated
by sou sorptxon (Graham-Bryce, 1984). Ba.«J on Equation 4 and Equation 9 for a
son With ,„> 0.0125 (OM « at,- Kd's of l, 20, 40, and 1000 would represent log
K,, s of 2.3, 3.6, 3.9, and 5.3, respectively.
. Vapor phase partitioning of a compound in the ..oil influence, the spread of
the compound through the soil. Even for cheaical. with relatively low vapor
*hi" tranSPOrt rWt* h" ^en *"«> *° «- ^^ificant (Mayer,. t al.,
Those chemxcals that have a high vapor pressure may easily move from the
sou solution into the soil air phase, where they can move throughout the soil
-------
23O6
and across the soil surface. The vapor-phase say be taken up by the plant either
through roots or by above ground portions of the plant.
The compartmentalization of the compound between the soil solution and the air
spaces in the soil is frequently described by Henry's Lav (Jury et al., 1983) with
the extent of partitioning described by Henry's Constant (He) . This can be
calculated as:
Henry's Constant (He) - 1$^04P M tll]
where P « vapor pressure of pure solute in mm/Hg,
H - molecular weight of solute,
T « absolute temperature, and
S - solubility in water ag/L
(Thibodeaux, 1979) . Henry's Constant may be expressed in different units and vary
by several orders of magnitude depending upon the source of the original data.
For example, estimated values for vinyl chloride of 2.3 X 10"2 to 6.39 ata ms/mol
are reported by Mackay and Shiu (1981) and Goldstein (1982) , respectively.
Experimentally determined He values are considered more reliable than calculated
values. Henry's Constant, dimensionless, for the priority pollutants is provided
in Table 1.
Comprehensive studies have not been conducted to determine the He above which
volatilization plays an important role in the transport of a chemical in the
atmosphere. Thus, it is not possible to select a He above which transport in the
soil will occur primarily in the vapor phase. However, a partition between the
vapor and aqueous phases of greater than 10"* is normally sufficient for a
chemical to be a good preemergence herbicide (Graham-Bryce, 1984) . Jury et
al.,(1984) utilized three volatility categories with He values of 2.5 x 10"3, 2.5
x 10"s and 2,. 5 x 10*7. Gillett (1983) utilized values of 10"3 and 6 x 10'5 in his
classification. Thus, the value of 10"* may be a reasonable transition point for
determining when vapor diffusion becomes important. This would mean that vapor
diffusion would be important for all. PCB's and halogenated aliphatics and
unimportant for some of the monocyclic and polycyclic aromatics and many
pesticides. Soil sorption can significantly reduce chemical ' volatilization
(Fairbanks et al., 1987) thus, the arbitrary value of 10"* may overestimate the
importance of volatilization in high organic carbon soils. Jury et al., (1983)
used He and K^ to calculate volatilization flux from' soil.
PLANT CPTAJOt OT OROAHIC
Chemical uptake by plants is a complex process that may involve a compound
specific active processes, and/or .a passive process in which the chemical
accompanies the transpiration water through the plant. If the former case
dominates, a rigorous relationship, between plant uptake and the chemicals
-------
2307
physicochemical parameters may not exist, although some general guidelines may
be expected. If uptake into the plant is a passive process, rigorous relationships
should exist.
It is generally accepted that there are four main pathways by which a chemical
in the soil can enter a plant (Topp et al., 1986). These ar«:
.1. root uptake and subsequent translocation by the transpiration stream,
2. vegetative uptake of vapor from the surrounding air,
3. uptake by external contamination of shoots by soil and dust, followed
by retention in the cuticle or penetration through it, and
4. uptake and transport in oil cells which are found in oil containing
plants like carrots and cress.
The amount of an organic chemical found in a plant will be the sum total of"
each of these transport routes minus metabolic losses. Their respective importance
will depend upon the nature of the organic chemical, the nature of the soil, and
the environmental conditions under which plant exposure occurs. Pathways 364
are significant only in specific situations. Thus, for the purpose of describing
the general case of plant uptake, they can be discounted as major routes of plant
contamination. Most reported instances of plant uptake of soil-borne organic
-compounds make no attempt to discriminate between pathways 1 & 2. Therefore, the
relative importance of each pathway, under different environmental conditions,
has not been assessed at present.
Reet Ootake And Traaaloeatien '
Shone and Wood (1.972) investigated the absorption" and translocation of the
herbicide simazine by 6-day-old barley plants in solution cultures. The
experiments were either 24- or 48-hour experiments conducted under different
conditions of humidity, light intensity, temperature, and levels of metabolic
inhibitors. The relationship between simazine transport and water uptake was
described by a transpiration stream concentration factor (TSCF), defined as:
TSCF = t"? ?imagine in shoots per mL water transnired
ng simazane per mL of external solution
They found that water was taken up preferentially to simazine, because the TSCF
was always less than unity, i.e., the concentration of simazine in the plant
shoots per mL of water transpired never reached that in the external solution.
There was no evidence of loss of or breakdown of the parent compound during the
experiment. The concentration of simazine in the plant roots, on a fresh weight
basis, however, reached a value greater than unity as a result of physical
sorption of the herbicide to the root tissue.
Evaluation of other triazines led to the conclusion that plant uptake was, in
general, a passive process because TSCF was less than unity, {Shone et al.,1973).
-------
2308
Plant uptake of 6 herbicides and a fungicide showed.that TSCF was independent of
concentration and less than unity for all except 2,4-D at pH 4.0 (Shone and Wood,
1974). In the case of 2,4-D at pH 4.0, plant uptake vac metabolically influenced.
Briggs et al.,(1982) evaluated plant uptake of 18 chemicals and found that the
TSCF was less than unity for all chemicals studied. They related the TSCF to the
octanol/water partition coefficient (KM) for the chemicals and found a bell
shaped relationship between TSCF and K^,, with a broad maximum around a K^ of 1.8.
A Gaussin curve (Figure 5) was fitted to the data such that:
TSCF-O^e-"10^- l-™> /2-«] [12]
/The authors suggested that at K^, values below 1.8, translocation is limited by
the lipid membranes in the-root. At Revalues above 1.8, translocation is limited
by the rate of transport of the lipophilic chemical from the plant root to the
top of the plant. All the TSCF values were below unity, suggesting 'passive
chemical movement into the shoot with the transpiration stream.( There wa» no
evidence that chemicals were taken up against a concentration gradient.
1.0-1
FIGURE 5 RELATIONSHIP BETVEEH log Kov AKD TRAHSPIRATIOH
STREAK COHCEHTRATIOH FACTOR
Shone and Hood (1974) proposed that the uptake of a chemical into a plant root
could be described by a root concentration factor (RCF), defined as:
TJCP m concentration in root.fug/a fresh wt.)
concentration in external solution, (pg/mli)
Using radiolabelled herbicides in solution culture with barley seedlings, they
showed that the quantity of the herbicide transported to the plant stems (TSCF)
could not be inferred from the concentration in the plant roots (RCF). In
addition, although the RCF of some of the tested herbicides exceeded unity, uptake
was not affected by temperature. This, suggests the compounds were retained by
physical sorption rather than biochemically.
When barley seedlings were transferred from the herbicide amended solution
culture to a herbicide free solution, RCF decreased before TSCF was affected by
-------
2309
the change (Shone et al., 1974). Thus lipophilic herbicides appear to penetrate
the cortical cells of the root whereas the lipophobic herbicides are largely
confined to the free cell space in the root.
Briggs et al., (1982) found that RCF was related to K,,,. Starting with a value
of less than unity for polar compounds, RCF increased with increasing l^,.
Sorption of chemicals by macerated roots was very closely related to the RCF of
living roots, for the more lipophilic chemicals. In contrast, the RCF of macerated
roots continued to decrease as the lipophilicity decreased (Figure 6). There was
a linear relationship between the log concentration factor of the macerated roots
and log K,,,:
* root) -' 0.77. logK,, - 1.52 [13]
100
MACERATED ROOT
. -0 • -1 ' 2 3 ' 4 5
log Xov
FIGURE 6 EFFECT OF TISSUE' STATUS OH THE RELATIONSHIP
BETVEEH log Kov AND ROOT COHCEHTRAT10H FACTOR
*OB0t»a from eriggs «t Ol . 1982
Assuming that RCF of living roots could be explained by two processes: (1) a
partitioning of the organic chemical between the lipophilic root tissue and
external solution culture and (2) a fraction of root that is aqueous and equal
in concentration to external solution phase (constant for all compounds, 0.82).
Briggs et al., (19«3) suggested that sorption of chemicals by the root is a
partitioning described by:
log(RCF - 0.82) - 0.77 logK,,. - 1.52 [14]
They proposed an analogous stem concentration factor (SCF) :
SCF - concentration in stm iua/a fr.MH wt.i
concentration in external solution (ng/*L)
Macerated stems sorption of organic compounds was also related to the K,,, of the
compound: '
logSCF(-,crlt-
.t-)
0.95
- 2.05
[15]
Assuming that the contribution of the aqueous phase in the stem was similar to
that in roots (0.82), the partition between the stem and xylem stream is: •
ll-Mp) - 0.8,2) - 0.95 logK., - 2.05 [16}
-------
2310
The SCF 'is then given by the K(,t-/xyli-Mp) partition co*fficient multiplied by th«
partition of the external solution present in the xylem sap (TSCF) :
SCF - [10(0.951ogKa, - 2.05) +
~ 1.78)2/2.44)
[17]
For *5 chemicals (logK^ from -0.57 to 3.7), th. .xp.rim.ntal point, fit th.
predicted line quite well (Figure 7) . Th. shift, in log K,,, wh.re TSCF reaches a
maximum (1.8) to where SCF reach., a maximum (4.5) ari... b«caus« sorption of
the more lipophilic compounds by th. stra tissu. incr.as.s faster than th. TSCF
decreases. The predicted decline in see for compound, of log * > 4.5 wa» not
tested. ' • '
FIGURE 7
8
RELATIONSHIP BETVEEH log Kov AND PLANT
CONCENTRATION FACTOR
There have been other attempts to r.lat. plant upta*. and translation of an
organic chemical to either the physical or ch«ical prop.rti.s of th. chemical
Topp et al., (1986) reported the relationship:
logRCF - 0.63 logK,,, - 0.959 ,. '.
following their exposure of barley seedling, for 7-day, to various chemicals in
water culture.
The concentration factor (CF) concept is a useful way of describing the
relative concentration of an organic chemical in a particular plant part, it has
cZica? vT', TVer> ^ ari" "—»" *~ CmCmtration °£ *****
chemxcals, both within th. soil or nutri.nt solution and within th. plant part
mLT TTn C°nStant With tim*- Ch«aical« in «» «U. or in.nutri.nt solution,
may be depleted by plant uptaJc. or degradation; chemical, in a plant may also be
reduced with time by degradation within th. plant, or by incr.as.s in plant mass
effectively diluting the chemical, change, in upta*. a. measured by th. CF, have
been reported, Figure 8 (Topp et al., 1986). Different CF's aris. dep.nding upon
the timing of th. actual sampling. Further it seem, logical that the CF would
depend upon soil concentration, initial v. soil concentration at time of plant
-------
2311
sampling. Further research on this topic is needed to define the effect of time
of sampling (both plant and soil) on CF'« so different experiments can be
compared.
is -i
10 -
5 -
BARLEY
50
TIKE Cdays)
100
FIGURE 8 EFFECT OF PLAHT TYPE AID LEBGTH OF GROWTH
PEROID OH THE PLAHT COHCEHTRATION FACTOR
««Bt«0 from TODD « a I 19i6
The work of Shone, Briggs, and their co-worker, reported above was carried out
in nutrient solution cultures where sorption and desorption effects of soil
organic matter were absent. The application of their results to plant uptake from
field soils requires that soil sorption be considered. The effect of soil sorption
on son solution concentration can be mathematically described using the following
relationship: ' • _
5CS
ec,
[19]
where CT is the total organic chemical concentration in the soil (ng/g), S is the
soil bulk density (g/cm3), c$ is. the adsorbed chemical concentration (Mg/g), 6 is
the soil-water content by volume (mX/cm5), and C, i. the chemical concentration
in the soil-water phase (Ag/mL). Using the linear equilibrium relationship in
Equation 3 and 4 allows Equation 19 to be rewritten in terms of c such that-
_c,. = s
CT ' 5Kpe^oc + S . " • [20]
It is now possible to combine equations relating soil sorption and soil
solution concentration and calculate RCF, TSCF, and SCF for different chemicals
on a total soil concentration basis. Substituting Equation 20 into' Equation 17
where CL is the external solution and:
SCF
(SOIL)
gives:
concentration in
concentration in soil
SCF,
(SOIL)
5-2.44]
[21],
For nutrient solutions this equation reduces to Eq 117] when /„,. - o, e - l,
and S = i. Inclusion of soil sorption into the SCF from Briggs et al.,
-------
2312
alters the relationship between SCF and 109 K^ such that the log K^ where plant
adsorption is a maximum decreases from 4.5 for nutrient solution to 1 for soils.
(Figure 9). The decrease in SCF for chemicals with log K^, greater than 1 is
supported by the published literature on plant uptake in soil systems (Travin and
Anus, 1988).
bu
U
cn
FIGURE 9 EFFECT OF SOIL ON THE RELATIONSHIP BETWEEN
log Row AND STEM CONCENTRATION FACTOR
Equation 21 implies that plant uptake is related to soil organic natter content
(Figure 10).Differences in the plant uptake of an organic chemical .in soils with
different organic carbon contents has been shown experimentally. Lichtenstein et
al., (1967) for example, showed higher concentrations of the pesticide aldrin in
roots of peas when grown in aldrin- polluted quartz sand compared to a loam soil
containing approximately the same total concentration of the pollutant.
FIGURE 10 EFFECT OF SOIL ORGANIC HATTER OH THE
RELATIONSHIP- BETVEER log Kow AHD
STEH COHCEHTRATIOH FACTOR
•It is also apparent from Equation 21 that increases in soil water content
reduce SCF (Figure 11). However, for a soil with a fx of 0.0075 (1.25% organic
Batter), changes in soil water content over the range 0.1 to 0.5 mL/cm3 altered
SCF less than 10% for chemicals with a K^ greater than 2.5. The fraction in
solution, (ecL/cT)j increases as soil water content increases' even though the
organic chemical concentration (CL) in the soil 'solution-phase decreases.
-------
2313
Therefore, if plant transpiration were increased by increasing soil water content,
plant concentration could be increased. Walker, (1971) found that the
phytotoxicities of the pesticides atrazine, simazine, linuron, lenacil, and
aziprotryne were increased as the moisture content of the soil increased. He
related the effect to differences in the quantities of the pesticides that were
accumulated by the plants, with the degree of accumulation being directly
proportional to water uptake. :
VOLUMETRIC VATER
CONTEXT
-1
,6'
FIGURE 11 EFFECT OF SOIL VATER COHTEHT OS THE RELATIONSHIP
BETVEEH log Kov AHD STEK COKCEffTRATIOS FACTOR
In conclusion, assuming degradation of the organic chemical does not occur
within the plant, and plant root uptake and translation of organic chemicals
from the soil is passive, plant uptake can be described as a series of consecutive
partitions reactions. Partitioning occurs between soil solids and soil water soil
water and plant roots, plant roots and transpiration stream, and transpiration
stream and plant stem. This partitioning can be related to the K of organic
compounds such that pollutants with high log *„, values, (eg. TCDD (6.14), PCB-S
(4.12-6.11), some of the phthalate esters (above 5.2) and the polycyclic aromatic
hydrocarbons (4.07-7.66)) are most likely to be .orbed, by the soil and/or plant
root. Chemicals with lower KO- values are likely to be translocated within the
Plant and may reach significant concentrations/within the above ground portions
of the plant. . • •
Vapor Phase
For volatile compounds, diffusion in the vapor phase and subsequent uptake by
the root and/or shoot may be an important route of chemical entry into plants
(Parker, 1966, and Prendeville, 1968). Two processes precede the penetration of
chemicals in the soil into plant tissue via the air: 1) volatilization of the
chemical from the soil and 2) deposition from the air onto the plant surface. Soil
volatilization depends upon the vapor pressure of the compound which varies
according to ambient temperatures, water solubility of the compound, and sorption
capacity and physical properties of the soil.
-------
2314
Increasing the soil-water content of a coil will increase the potential for
volatilization loss of a chemical (Guenzi and Beard, 1970). Harris and
Lichtenstein (1961) showed that the rate of volatilization of aldrin from soil
increased with aldrin concentration, soil moisture, relative humidity, temperature
and the rate of air movement. Chemical concentration effects cease when the
concentration reaches that required to give a maximum saturation vapor density
equivalent to that of the pure compound. For dieldrin in a Gila silt loan soil
this concentration was 25 ppm (Farmer et al., 1972). These authors also rttport
that under similar environmental conditions the rate of volatilization was lindane
> dieldrin > DDT, which is the same order for increasing vapor pressures. Jury
et al., (1983 and 1984) developed a behavior assessment model that separates
compounds into volatilization categories based on Henry's constants.
There have been few investigations aimed at separating root uptake and
translocation of a chemical from vapor phase uptake into plant shoots. In an
experiment designed to discriminate these 'effects, Beall and Mash (1971) found
soybean shoots were contaminated by soil applied dieldrin, endrin and heptachlor
largely via root uptake and subsequent translocation. Vapor phase .foliar sorption
however dominated for DDT and was nearly 7 times greater than root sorption and
translocation. Foliar contamination from vapor sorption of residues from all four
insecticides was similar (about 6.5 ppm plant dry weight), whereas contamination
from root sorption and translocation varied from 38 ppm to 1 ppm depending upon
the compound. . •
Using similar experimental techniques. Fries and Narrow (1981) found that PCBs
reached the shoots of plants via the vapor phase rather than from root uptake,
although the importance of this route for PCS contamination of plants remains
inconclusive.
Topp et al., (1986) investigated the uptake of 16 organic chemicals by barley
seedlings. Foliar uptake was related to the amount of chemical volatilized from
the soil surface.. The relationship (Figure 12) after 7 days exposure was:
FU - 46.11 + 28.95 log VOL [22]
where FU was foliar uptake as percent of total UC uptake, and VOL was the organic
UC trapped from the air plus that sublimated on the walls of the exposure chamber-
as percent of the total UC applied (Note that in the original publication the
sign in front of log VOL is negative, this is assumed to be a typographical
error). Four compounds (benzene, pentachlorophenol, diethylhexylphthalate, and
the phenylenediamine pigment) did not fit the calculated line because they were
nonpersistent and taken up after mineralization to 1*CO2.
There are many difficulties in extrapolating vapor phase uptake in the
-------
2315.
laboratory to that in the field. Overall, volatilization rates are likely to be
higher in the laboratory than in the field. This is because laboratory soils are
normally kept moist to encourage plant growth, and this encourages
volatilization. In addition, the actual deposition of volatilized chemicals onto
a plant in the field is likely to be lower as atmospheric turbulence nay be
higher.
-2
-i • -o
log VOLATILIZATION
FIGURE 12 RELATIONSHIP BETVEEH VOLATILIZATION AND
FOLIAR UPTAKE MBCKM from T
-------
2316
phase, and can be related to the octanol vater. partition coefficient of the
compound. Subsequent translocation of the chemical from roots to shoots depends
on the K^ of the compound and the transpiration rate of the plant. Based on
available data, compounds with a log K,,, of approximately 4.5 are most likely to
accumulate in the stem and leaf tissue of plants.
In soil systems, -there is competition between the plant and soil solids
(organic fraction) for.the partitioning of organic* from solution. As the sorption
of the compound by the soil organic phase increases, the quantity available for
plant uptake decreases. Based upon these considerations compounds with log K^
of 1 -2 are most likely to have significant transport of the chemical to above
ground plant tissue produced in soil systems. If metabolism of the compound in
the roots is significant, even compounds with low log K^'s may not be
translocated (HcFarlen et al., 1987). Compounds with high log K^ > 5.0 would not
be expected to be present in above ground plant tissue if plant uptake is limited
by soil solution.
The potential for root or plant sorption of organic compounds from vapor is
dependent upon the vapor pressure of the compound. Very few experiments on this
route of plant contamination have been conducted. Based upon the movement of
herbicides in the soil, a Henry's constant of 10"* may be used as a transition
point between primary movement in solution and vapor phases. If it can be assumed
that vapor movement in the soil will result in vapor uptake by the plant, then
those compounds with He >10"* are potential candidates for vapor phase uptake.
,superimposed upon both of these processes is the half-life of the compound.
If it is short, i.e., less than 10 days, the chemical is likely lost from the
system before it can be taken up by the plant. Those compounds with long half
lives, i.e., greater than 6 months or greater than the growing season of the
plant, presist long enough to impact plants.
Applying these screening processes to the priority pollutants, listed in Table
1, reduces the number of chemicals likely taken.up by plants. For example, if
plant uptake and translocation without vaporization is the pathway of concern,
tb* li*t of 107 chemicals is reduced to 50- on the basis of half-life and K^,
(Table 2). If vaporization is of concern the list is reduced from 107 to 64 on
the basis of half-life and He, (Table 3).
-------
2317
TABLE 2 Log K^, Half-life and He for Priority Pollutant* which ar* subject to
plant uptake from soil
tcsr
'1/2
He
PESTICIDES
2Q.Acrel«in
27.Endo«uU«n
31 .Nwuchlorocyctohwim
33.1taphorent
-0.09
3.55
3.8
1.70
8
C
8
nd
2.8E-03
nd-
3.0E-04
nd
26.D'i«ldrin
SO.ftpucnlor (pOKldt
35.Teu*hinB
2.9
3.9
3.72
3.8S
e
c
c
c
3.0E-04
3.2E-OS
6.0E-04
2.1E-01
NLYC8LOKIMTE9 B1PK8ILS ' '
MUCEMTEB ALIMATIC ITDtOCNB
38.Chtoranthan*
40.Tridttor<»thint
42.Chlerotthm
44.1,2-dichlerMthm
46.1,1.2-tricnloratthm
. 56.N*uchlorotoutadf«n»
59.lrandichleroMthn
61.TribronMtnm»
63.TriehlorofluaraHth«tt
MLOGEMTB cms
6S.8U(2-c*ilero*thyl)tth*r
70.8f(C2-dilorMtnexyMitam
mocraic AKMATICS
74. 1 ,3-dienlonbKuww
79.M1 tretenzanv
82.2,6-dinitretolum
86.2,4,6-tridilora^Mnpt
89.4-ni traphtnot
91 .2.4-diatthylpiwnol
93.4,6-dinitre-e-ermol
UK
0.91
1.9
1.54
1.48
2.17
3.74
1.88
2.30
2.53
1.58
1.26
2.84
3.55
1.85
2.05
3.61
1.91
2.50
2.85
C
8
8
8
nd
nd
nd
nd
nd
C
nd
nd
c
. nd
nd
.8
nd
nd
1.6E*01
1.21-01
6.1E-01
3.8E-02
3.1E-01
4.3E-01
nd
2.4E-02
2.4t*OO
4.7E-05
1. It-OS
1.5E-01
1.1E-01
5.4E-04
1.3E-02
1.7E-04
2.«E-04
7.SE-04
nd
39.0
-------
2318
TABLE 3. Log K^, Half-life and He for the Priority Pollutants which are subject
to plant uptake via volatilization
Cc*x»d 1
PESTICIDES
20.Acrol«in
23.BOO
27.Endc«ulfan
33.1tophorent
35.Teuph«nt
HLTOaOtlMTED 8IPKUIU
36».Arochtor 1016
36c.Arochler 1232
56«.Anxhlor 1248
36e.*ra»ler 1260
UUCEMTO ALIPMTIC ITOCCCMtl
40.TriehlortMt)un*
42.CMerotthan*
44.1,2-dtchlorotthm
46.1.1.2-trfchlerocthww
54 . 1 ,2-dfcJi toropropm
60 .0 i branch 1 ortacthm
62.0icnlorcdiftuoraMith«nt
KAUXSUTH) Eras
6A.li*(2-cMoroiMeraprl»tlMr
69.4-broKptMnyl ph«nyl «ttnr
KMOCTCLIC MOJMTICS
T2.Cntorct»nxm
74.1,3-dlchlerebtnun*
77. Muefl 1 enbtnztm
82.2,6-dlnitratolum
mmUUATE ESTEtS
KLTCTO.IC UCMTIC mCOCU«CB!
95c.Fluor«n*
97D.8*nze Oil f I uoranthm
97d.Chryunt
9Sc .6 1 b«ruo [t] tnthrietnt
"itmtwrr^w airuMn
101.DI-n-prapy( nltraUBin*
«**,
-0.09
5.99
3.55
1.70
3.85
4.38
4.54
5.6
6.11
OB*
0.91
1.9
1.54
1.48
2.17
2.28
1.10
2.09
2.16
2.58
4.28
2.84
3.55
6.18
2.05
3.22
C
4.13
4.18
6.57
5.61
7.23
5.97
0.06
1.31
T1/2
8
C
C
nd
C
C
• e
e
c
c
8
8
8
nd
nd
I
nd
C
nd
nd
nd
nd
C
nd
8
C
C
nd
C
C
C
nd
nd
He
2.8E-03
0.5E*00
nd
nd
2.1E-01
8.6E-01
2.1E*00
1.1E-01
2.9E-01
1.6E«01
1.2E-01
6.1E-01
3.8C-02
3.1E-01
1.2C-01
4.4E*00
nd
6.3E*01
4.7E-02
nd
1.5E-01
1.1E-01
7.0E-02
1.3E-02
1.9E-03
1. 06-02
4.8E-03
nd
8.8E-02
nd
nd
nd
nd
Ccapound
22.CMordm . -
25.DOT
31.N«aehleracycloh«iant ,
34.TCDO
36d.*radilor 1242 ,
36f.Arachter 1254
37.2-dilenraphtlwlm
41 .TctraehleroMthin*
43.1.1-dicMon»thm
45.1,1.1*triehlerattMra
56.ltaxKhlor«butid!m
61.TritaranBthin*
63.Tr
-------
2319
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(Received in Germany 11 September 1988; accepted 4 October- 1988)
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