SSf
United States '. The Office of Air Quality Planning and Standards, RTP, Nortt.i-Carolina, and
Environmental Protection The Environmental Criteria and Assessment Office. Cincinnati, Ohio
Agency October, 1986
Methodology for the Assessment of Health Risks
Associated with Multiple Pathway Exposure to
Municipal Waste Combustor Emissions
DEPOSITION
ON WATER
DEPOSTION;
GROUND
DEPOSITION
ON FOOD
AND FEED
RIGATION
i
\
EATING
VEGETABLES
DRINKING
INHALATION
DERMAL
ABSORPTION
- UPTAKE
A Staff Paper Submitted for Review to the Science Advisory Board
-------
METHODOLOGY FOR THE ASSESSMENT OF HEALTH RISKS
ASSOCIATED WITH MULTIPLE PATHWAY EXPOSURE TO
MUNICIPAL WASTE COMBUSTOR EMISSIONS
U.S. Environmental Protection Agency
The Office of Air Quality Planning and Standards, RTP, North Carolina
and
The Environmental Criteria and Assessment Office, Cincinnati, Ohio
October 1986
-------
DISCLAIMER
This document is an external draft for review purposes only and does not
constitute Agency policy. Mention of trade names or commercial products does
not constitute endorsement or recommendation for use.
ii
-------
CONTENTS
LIST OF TABLES [[[ vii
FIGURES [[[ x
PREFACE [[[ xi
EXECUTIVE SUMMARY .................................................. xi i
LIST OF ABBREVIATIONS .............................................. xix
AUTHORS, CONTRIBUTORS, AND REVIEWERS ............................... xxii
1. INTRODUCTION AND BACKGROUND
2. MUNICIPAL WASTE COMBUSTOR TECHNOLOGY AND EMISSIONS ............. 2-1
2. 1 DESCRIPTION OF MWC TECHNOLOGIES ........................... 2-1
2.1.1 Massburn Facilities ................................ 2-1
2.1.2 Small Modular Incinerators ......................... 2-3
2. 1. 3 Refuse-Derived Fuel ................................ 2-5
2.2 DISTRIBUTION OF EXISTING MUNICIPAL WASTE COMBUSTION
FACILITIES IN THE UNITED STATES ........................... 2-7
2.3 DISTRIBUTION OF PLANNED MUNICIPAL WASTE COMBUSTION
FACILITIES IN THE UNITED STATES ........................... 2-8
2.4 PROJECTED GROWTH OF MUNICIPAL WASTE COMBUSTION THROUGH
THE YEAR 2000 ............................................. 2-11
2. 5 MUNICIPAL WASTE COMBUSTOR EMISSIONS ....................... 2-15
3. EXPOSURE .MODELING OF MUNICIPAL WASTE COMBUSTOR EMISSIONS ....... 3-1
3.1 THE INDUSTRIAL SOURCE COMPLEX MODEL ....................... 3-6
3.2 THE HUMAN EXPOSURE MODEL (HEM) ............................ 3-15
3. 3 TERRESTRIAL FOOD CHAIN MODEL .............................. 3-25
3.3.1 General Considerations ............................. 3-26
3.3.1.1 Most-Exposed Individuals (MEIs) ........... 3-26
3.3.1.2 Soil Deposition Rate of Contaminants ...... 3-27
3.3.1.3 Soil Incorporation of Deposited Con-
tami nants ................................. 3-28
3.3.1.4 Contaminant Loss from Soils ............... 3-28
3.3.1.5 Contaminant Uptake Relationships
in Plants ................................. 3-29
3.3.1.6 Contaminant Uptake by Animal Tissues ...... 3-31
3.3.1.7 Human Diet ................................ 3-31
3.3.2 Deposition-Soil-Plant-Human Toxicity Exposure
Pathway ............................................ 3~32
3.3.2.1 Assumptions ............................... 3-32
3.3.2.2 Calculation Method ........................ 3-32
3.3.2.3 Input Parameter Requirements .............. 3-37
3.3.2.4 Example Calculations ...................... 3-37
3.3.3 Deposition-Human Toxicity ("Pica") Exposure
-------
CONTENTS (continued)
Page
3.3.3.3 Input Parameter Requirements 3-41
3.3.3.4 Example Calculations 3-43
3.3.4 Exposure Pathways for Herbivorous Animals for
Human Consumption 3~44
3.3.4.1 Assumptions 3-44
3.3.4.2 Calculation Method 3-44
3.3.4.3 Input Parameter Requirements 3-46
3.3.4.4 Example Calculations " 3-50
3.4 SURFACE RUNOFF 3-53
3.4.1 General Considerations 3-53
3.4.2 Assumptions 3-53
3.4.3 Calculations 3-55
3.4.4 Required Inputs 3-64
3.4.5 Example 3-64
3.4.5.1 Tier 1 3-66
3.4.5.2 Tier 2/3 3-68
3.5 GROUNDWATER INFILTRATION .' 3-76
3.5.1 General Considerations 3-76
3.5.2 Assumptions 3-77
3.5.3 Calculations 3-77
3.5.4 Requi red Inputs 3-86
3.5.5 Example 3-88
3.5.5.1 Tier 1 3-88
3.5.5.2 Tier 2/3 3-91
3.6 DERMAL EXPOSURE MODEL 3-95
3.6.1 General Considerations 3-95
3.6.1.1 Most-Exposed Individual (MEIs) 3-95
3.6.2 Deposition-Human ("Dermal") Toxicity Exposure
Pathway 3-95
3.6.2.1 Assumptions 3-95
3.6.2.2 Calculation Method 3-96
3.6.2.3 Input Parameter Requirements , 3-98
3.6.2.4 Example Calculations 3-101
4. ESTIMATING CARCINOGENIC AND NONCARCINOGENIC RISKS TO HUMANS BY
INDIRECT EXPOSURE 4-1
4.1 DETERMINATION OF THE ADJUSTED REFERENCE INTAKE (RIA) ' 4-2
4.1.1 Threshold-Acting Toxicants 4-2
4.1.1.1 Risk Reference Dose (RfD) 4-2
4.1.1.2 Human Body Weight (BW) 4-3
4.1.1.3 Relative Effectiveness of Exposure (RE) ... 4-3
4.1.1.4 Total Background Intake of Pollutant
(TBI) 4-4
4.1.2 Calculation of RIA for Cadmium 4-5
4.1.3 Non-Threshold Toxicants -- Carcinogens 4-6
iv
-------
CONTENTS (continued)
Page
4.1.3.1 Human Cancer Potency (qj) 4-8
4.1.3.2 Risk Level (RL) 4-8
4.1.3.3 Human Body Weight (BW) 4-8
4.1.3.4 Total Background Intake of Pollutant
(TBI) 4-8
4.1.3.5 Relative Effectiveness of Exposure (RE) ... 4-9
4.1.4 Calculation of RIA for Benzo(a)Pyrene 4-9
4.2 COMPARISON OF DAILY INTAKES (DI) AND DERMAL DAILY INTAKE
(DDI) WITH THE ADJUSTED REFERENCE INTAKE (RIA) 4-10
4.3 DETERMINATION OF THE REFERENCE WATER CONCENTRATION
(RWC) 4-10
4.3.1 . Threshold-Acting Toxicants 4-10
4.3.1.1 Risk Reference Dose (RfD) 4-11
4.3.1.2 Human Body Weight (BW) 4-11
4.3.1.3 Water Ingestion Rate (I ) 4-11
4.3.1.4 Relative Effectiveness Sf Exposure (RE) ... 4-12
4.3.1.5 Total Background Intake Pollutant (TBI) ... 4-12
4.3.1.6 Bioconcentration Factor (BCF) 4-12
4.3.1.7 Fish Consumption Rate (If) 4-15
4.3.2 Calculation of RWCw for Cadmium .! 4-18
4.3.3 Non-Threshold Toxicants — Carcinogens 4-18
4.3.4 Calculation of RWC for Benzo(a)Pyrene 4-19
4.4 COMPARISON OF CONTAMINANTWCONCENTRATIONS (Ci and Xi) WITH
REFERENCE WATER CONCENTRATIONS (RWC) 4-20
5. ECOLOGICAL EFFECTS FROM MWC EMISSIONS 5-1
5.1 TERRESTRIAL FOOD CHAIN MODEL 5-1
5.1.1 General Considerations 5-1
5.1.1.1 Most-Exposed Individuals (MEIs) 5-1
5.1.1.2 Soil Deposition Rate of Contaminants 5-2
5.1.1.3 Soil Incorporation of Deposited
Contami nants 5-2
5.1.1.4 Contaminant Loss from Soils 5-2
5.1.1.5 Contminant Uptake Relationship 5-2
5.1.1.6 Toxicity Thresholds for Nonhuman
Organi sms 5-2
5.1.2 Exposure Pathways for Toxicity to Herbivorous
Animals 5-3
5.1.2.1 Assumptions 5-3
5.1.2.2 Calculation Method 5-3
5.1.2.3 Input Parameter Requirements 5-4
5.1.2.4 Example Calculations 5-4
5.1.3 Deposition-Soil-Plant Toxicity Exposure Pathway 5-5
-------
CONTENTS (continued)
5.1.3.1 Assumptions 5-5
5.1.3.2 Calculation Method 5-5
5.1.3.3 Input Parameter Requirements 5-5
5.1.3.4 Example Calculations 5-5
5.1.4 Exposure Pathways for Toxicity to Soil Biota and
Their Predators 5-5
5.1.4.1 Deposition-Soil-Soil Biota Toxicity
Exposure Pathway 5-5
5.2.4.2 Deposition-Soil-Soil Biota Predator
Toxicity Exposure Pathway 5-6
5.2 SURFACE RUNOFF AND GROUNDWATER MODELS 5-10
5.2.1 General Considerations 5-10
5.2.1.1 Aquatic Life Protection 5-10
5.2.1.2 Wildlife Protection 5-11
REFERENCES 6-1
APPENDIX A A-l
APPENDIX B B-l
VI
-------
LIST OF TABLES
Number
2-1
2-2
2-3
2-4
2-5
2-6
2-7
3-1
3-2
3-3
3-4
3-5
3-6
3-7
3-8
3-9
3-10
Planned MSW Combustion Facilities for States with the
Greatest Growth
Percentage by Region of the Forecast Waste to Energy
Throughout 1985-2000
Pollutants Quantified in Municipal Combustion Emissions ...
Summary Matrix of Emissions Data Gathered for MSW
Incinerators
Emissions Data for Massburn MWC Facilities
Uncontrolled Metals Emission Factors for MWCs
Controlled Inorganic Emissions from the Hampton MWC
Modeling Parameters for the Model Plant
Model ing Parameters for Hampton, VA
Summary of Scavenging Coefficients Expressed per Second
of Time, Used in Computing Wet Surface Deposition
Particle Size Distribution Determined in Particulate
Matter Emissions at the Braintree MWC
Typical Particle Size Distribution Determined in
Particulate Emissions at the Wurzburg Massburn MWC
Ratio of Metal Emissions as a Function of Particle
Diameter
Unit Cancer Risk Estimates for Inhalation Exposure to
Specific Chemicals
Ambient Air Concentrations of B(a)P as Predicted by the
ISC Model (with ESP control )
Ambient Air Concentrations of B(a)P as Predicted by the
ISC Model (with Dry Scrubber-Fabric Filters)
HEM Output as to the Distribution of Risk by Population
Resulting from B(a)P Emissions at the Model Plant (with
ESP control )
Page
2-13
2-14
2-16
2-16
2-20
2-22
2-22
3-5
3-5
3-10
3-12
3-12
3-13
3-20
3-21
3-22
3-23
vn
-------
LIST OF TABLES (continued)
Number
3-11
3-12
3-13
3-14
3-15
3-16
3-17
3-18
3-19
3-20
3-21
3-22
3-23
3-24
3-25
4-1
4-2
HEM Output as to the Distribution of Risk by '.Population
Resulting from B(a)P Emissions at the Model Plant (with
Dry Scrubber-Fabric Filter Control)
Average Daily Dry-Weight Consumption of Food Groups, Based
on a Reanalysis of the FDA Revised Total Diet Food List ...
Assumptions for Terrestrial Food Chain
Assumptions for Deposition-Soil-Plant-Human Toxicity
Exposure Pathway
Average Percent of Annual Consumption which is Homegrown
for Various Foods, Rural Farm Households
Assumptions for Deposition-Human Toxicity ("Pica")
Exposure Pathway
Assumptions for Pathways Dealing with Herbivorous
Animal s
Surface Runoff Methodology Assumptions
Input Parameters for the Runoff Pathway Methodology
Input Parameters for the Example Calculations
Assumptions for the Groundwater Pathway Methodology
Metal Contaminants Simulated in the Geochemical Portion
of the Groundwater Pathway
Input Parameters for Groundwater Pathway
Input Parameters for the Example Calculations
Assumptions and Uncertainties for Dermal Exposure
Model
Illustrative Categorization of Carcinogenic Evidence
Based on Animal and Human Data
Water Ingestion and Body Weight by Age-Sex Group in
the United States
rage
3/\*\
-23
3-33
3-34
3-36
3-36
3-43
3-46
3-54
3-65
3-67
3-78
3-82
3-89
3-90
3-97
4-7
4-12
vm
-------
LIST OF TABLES (continued)
Page
United States Annual Per Capita Consumption of
Commercial Fish and Shellfish, 1960-1984 4-16
5-1 Assumptions for Ecological Effects for Terrestrial
Food Chain 5-7
5-2 Assumptions for Deposition-Soil-Soil Biota-Predator
Toxicity Exposure Pathway 5-8
ix
-------
FIGURES
Number
1-1 Potential Exposure Pathways to Pollutants Emitted from
the Stacks of Municipal Waste Combustors 1-5
2-1 Massburn MSW Incinerator with Overfeed Stoker Grates 2-2
2-2 Smal1 Modular Incinerator with Heat Recovery 2-4
2-3 RDF Processing Facility with On-Site Boiler 2-6
2-4 Distribution of Existing Installed Incinerator Capacity ,
by Design Type • • • 2-8
2-5 Distribution of Municipal Waste Combustors in the U.S.
by Region 2-9
2-6 Distribution of Installed MSW Incinerator Capacity by
Design Type 2-11
2-7 Projected Growth of Municipal Waste Combustion Between
1986 and 1990 2-12
3-1 Surface Runoff Pathway Methodology 3-56
3-2 Logic Flow for Groundwater Pathway Evaluation 3-80
3-3 Example MINTEQ Speciation Results for Entry of a Contami-
nant into the Standard Zone for Condition of pH = 7.0 and
Eh-1.50 mv 3-87
3-4 Groundwater Cadmium Speciation 3-94
-------
PREFACE
Parts of this document have been taken directly from Development of Risk
Assessment Methodology for Land Application and Distribution and Marketing
of Municipal Sludge (U.S. EPA, 1986h) and Development of Risk Assessment
Methodology for Municipal Sludge Landfilling (U.S. EPA, 1986i). These
documents are undergoing review by the Environmental Engineering Committee
of the Science Advisory Board.
XI
-------
EXECUTIVE SUMMARY
Each year the collective social and commercial activity in the United
States produces >150 million tons of discarded waste. Commonly termed munici-
pal solid waste (MSW), this discarded material must somehow be managed to avoid
undesirable adverse consequences on human life, and the vitality of terrestrial
and aquatic life.
The age-old solution to the problem of managing MSW has been to dispose of
the waste in the ground in land areas dedicated to that purpose. Currently
about 80% of the.MSW is disposed of by land burial in ~10,000 landfills nation-
wide. If not properly sited, designed, and managed, these landfills can cause
serious damage to the environment. For example, gases can escape the landfill
and travel to residential areas potentially impacting human health, or contami-
nated leachate can migrate off-site into sources of potable drinking water and
into sensitive natural ecosystems. Because of the possible adverse environmen-
tal impact posed by landfills, many States have imposed strict siting require-
ments, landfill cover requirements, leachate collection and treatment require-
ments, landfill gas capture and treatment requirements, and groundwater monitor-
ing requirements to the design and operation of landfills. These requirements
have significantly increased the cost of disposing MSW in landfills, and have
limited land areas suitable for landfill sites.
Meanwhile the amount of MSW needing disposal continues to increase with
the increase in the U.S. population. By the year 2000 U.S. society may be faced
with managing the disposal of >250 million tons of MSW each year. Methods of
waste management are limited by available technology. Communities can continue
to only landfill MSW, or they can utilize technologies that will substantially
reduce the volume of waste that is ultimately landfilled, e.g., recycling of
waste and incineration of waste. While recycling strategies are being encour-
aged and fostered, many communities are turning to municipal waste combustion
(MWCs) in order to incinerate and reduce the volume of waste by 70-90%. Current
MWC technology is a distinct improvement in the design, combustion efficiency,
and pollution control over combustors planned and constructed a decade ago.
xn
-------
They not only reduce the volume of waste, but have the added advantage of ther-
mally recovering energy from combustion in the form of steam or hotwater that
can be used in industrial cogeneration, used to generate electricity, and used
to heat and cool residential and commercial properties.
The U.S. EPA predicts a substantial growth in MWC will occur over the next
10-20 years. Today 99 MWCs nationwide incinerate about 4% of the annual vol-
ume of MSW, whereas it is conceivable that by the year 2000 one-third of the
MSW will be incinerated in >300 MWCs. There is a definite trend moving toward
incineration of MSW, and away from exclusively landfilling the waste.
The U.S. EPA has a limited opportunity to prospectively evaluate the poten-
tial environmental and health impact that may result from a sudden proliferation
of municipal waste combustion. In this regard the agency has developed a
methodology for the evaluation of emissions of pollutants into the atmos-
phere from the stacks of MWCs during incineration. The methodology consists
of a series of environmental fate and transport models that utilize the known
physical and chemical properties of specific pollutants to predict the atmos-
pheric dispersion from stack emissions, the potential for surface deposition and
accumulation; the movement of the settled pollutants through and into various
environmental media; the potential bioaccumulation of pollutants into trophic
systems; the potential for adverse effects on the vitality of natural ecosy-
stems; and the potential for adverse effects on human health. With regard to
evaluating potential human health effects, the methodology will estimate health
risks resulting from inhalation of predicted ambient air concentrations of pol-
lutants; ingestion of pollutants deposited on the ground an bioaccumulated into
the food chain; ingestion of potable water or aquatic organisms contaminated
by the surface runoff and the leaching and percolation of settled pollutants
into water supplies; and ingestion of soil particles contaminated by deposited
incinerator emissions.
The utility of the present methodology is limited by a number of gaps in
the available technical data and significant uncertainties in many of the major
analytical parameters. There is little question that the methodology can be
improved by further research. One major limitation is that the methodology
focuses only on pollutants emitted from the stacks of MWCs. Ideally the total
pollutant loading resulting from the incineration process should be evaluated,
xiii
-------
e.g., ash residues, aqueous residues, and stack emissions. The evaluation of
stack emissions is further limited by the relatively small number of organic
and inorganic pollutants that have been measured in MWC emissions. A final con-
straint on the methodology is the limited amount of data regarding the physi-
cal and chemical behavior of specific pollutants in the natural environment,
and the adverse impact these pollutants may have on human health.
In the evaluation of the potential environmental impact of combustion
sources, the U.S. EPA has traditionally focused on air emissions from the
source, and on the human health risks from direct inhalation of predicted
ambient air concentrations of pollutants. The present methodology represents
an expansion of the analytical scope-to include consideration of multiple
pollutants, multiple exposure pathways, carcinogenic and noncarcinogenic risks
posed to humans, and potential adverse effects to the natural environment.
Human exposure to incinerator emissions results from direct inhalation of
ambient air concentrations of the pollutants and indirectly from skin contact
of the pollutants, and ingestion of contaminated soil particles, water and
food. Detailed experimental evaluation of the environmental fate and transport
of MWC emissions have not been conducted under actual conditions. Therefore,
mathematical models of fate and transport are currently the most feasible
alternative to the assessment of exposure to MWC emissions. In addition to
estimating concentrations that will be inhaled, these models can also be used
to estimate the potential accumulation in soils of pollutants adverse to the
promotion of human, animal and plant life, and accumulation of pollutants into
the human and ecological food chain. The models specifically used in this
analysis of MWC emissions are: the Human Exposure Model (HEM), the Industrial
Source Complex Short-Term Air Dispersion Model; the Terrestrial Food Chain
Model, the Surface Runoff Model, the Groundwater Contaminant Model, and the
Dermal Exposure Model.
Given the complexities of predicting the environmental fate and transport
of specific chemicals emitted, as well as predicting multiple routes of human
exposure to specific chemicals, it is not currently feasible nor practical to
apply the models to every existing or planned MWC. Therefore, the methodology
employs a simplified modeling approach by using a hypothetical plant (in
xiv
-------
western Florida) to characterize the potential adverse impacts of emissions
from technologies typical of MWCs currently being planned or considered, and
the Hampton, Virginia MWC to represent a reasonable worst case of the potential
adverse impacts on air pollutant emissions from existing MWC technology.
The Industrial Source Complex Model (ISC)
The industrial source complex model is used to predict the dispersion of
smokestack emissions from the hypothetical plant and the Hampton facility
through the atmosphere, as well as to predict both wet and dry deposition of
pollutants onto the surface. Assessments of potential risk from air emissions
have primarily been concerned with health risks resulting from direct inhala-
tion of ambient air concentrations of pollutants. The ISC assists in extending
the risk evaluation to a consideration of their routes of population exposure
to environmental pollutants and allows the U.S. EPA to predict the rate of
deposition, over time, of pollutants believed to be adsorbed onto particulate
matter in the smokestack exhaust gas, and attempt to calculate the spatial and
temporal accumulation of these pollutants on the soil, surface water, ground-
water and terrestrial food chain.
For purposes of exposure analysis from MWC emissions, the ISC Short-Term
(ISCST) model program is utilized. The program makes mathematical calculations
of dispersion and dry deposition and produces a printout of these values. How-
ever, the ISCST model as originally developed had no provision for calculating
wet deposition of the emissions. Because this deposition pathway is considered
to be of potential significance, the present methodology included an algorithm
to estimate the effect of precipitation events on the rate of surface deposi-
tion.
Human Exposure Model (HEM)
The ISCST output is a concentration array for a total of 160 receptors, or
10 receptors along each of 16 wind directions, specified in concentric radial
distances from each facility of 0.2, 0.5, 1, 2, 5, 10, 20, 30, 40 and 50 kilo-
meters computed every 22.5° on a radius-polar grid pattern. This output is a
suitable format for utilization with HEM. HEM is a general model that has
been routinely used with the EPA's air regulatory program to estimate the
carcinogenic risk to the population exposed by inhalation to predicted ambient
xv
-------
air concentrations of specified pollutants. The HEM also is capable of air
dispersion modeling, and is often used in nationwide analysis of source cate-
gories.
Terrestrial Food Chain Model (TFC)
Contaminants associated with emissions from MWC are subject to deposition
on surfaces downwind form the MWC. The fallout may be deposited on soil and/or
vegetation.
Humans in the vicinity of the MWC have the potential to ingest contaminated
soil directly or consume vegetation and animal tissues containing the contami-
nants. The TFC model has separate components for examining each potential expo-
sure pathway. These components describe methods for using empirical data on
contaminant uptake by plant or animal tissues to estimate tissue concentrations,
and for integrating these estimates to give a picture of potential human dietary
exposure. Potential exposure of children resulting from soil ingestion ("pica")
is also estimated.
Surface Runoff Model
Contaminants associated with particulates emitted by MWCs are subject to
deposition on surfaces downwind from the MWC at rates determined by meteorology,
terrain, and particle physics. This fallout is subsequently subject to dissolu-
tion and/or suspension on runoff after precipitation events. Runoff moves over
the surface of the earth to a surface water body where it mixes with other
waters. As a consequence, humans utilizing water from the surface water body
or aquatic life living therein may be exposed to runoff transported contami-
nants.
The methodology is formulated in three successive tiers that begin with
simple but very conservative estimates, and proceed to more detailed analyses
if the first tiers predict unacceptable risks. Both acute events and chronic
exposure are evaluated, using standard approaches to calculate runoff volume
and associated runoff potential. The methodology was originally developed to
evaluate impacts from the application of municipal waste-waters sludge to land.
xvi
-------
Groundwater Infiltration Model
Contaminants associated with particulate emitted from MWCs are subject to
deposition on surfaces downward from the facility. This fallout is subsequently
subject to dissolution in rain or meltwater from precipitation events. The dis-
solved portion can follow one of two pathways: either move over the surface
as runoff to a surface water body or infiltrate into the ground and recharge
the groundwater. As a consequence, persons using the groundwater may be exposed
to groundwater transported contaminants. Aquatic life inhibiting surface water
bodies fed by the contaminated aquifer could be exposed as well.
The methodology derived to calculate risks from the groundwater pathway
was originally developed to evaluate impacts from the landfilling of municipal
sludge. As for surface runoff, this methodology is formulated in three succes-
sive tiers. Only chronic exposure is evaluated using standard approaches to
calculate leachate generation and associated groundwaters transport in the un-
saturated zone.
Dermal Exposure Model
The dermal exposure model refers to human skin contact with contaminants
from emissions of MWC deposited on the soil. The tissue of dermal absorption
of deposited contaminants is very complex. There is a fundamental lack of data
for percutaneous absorption of chemicals in human skin from soil. Other factors
important for estimation of human exposure to contaminants by the dermal route
also have many uncertainties. The model described in this document is offered
as a possible approach for the estimation of human exposure and risk associated
with a dermal exposure, but it is recognized that in most, if not all cases,
the available data will not provide a satisfactory basis for risk calculations.
Systemic toxic thresholds or carcinogenic potencies of chemicals by a dermal
route of exposure have not been delineated by the U.S. EPA at the present time.
Ecological Effects from MWC Emissions
Methods to assess risk to terrestrial organisms represent a follow-up to
the Terrestrial Food Chain (TFC) model. Components for assessing effects of
deposited pollutants on herbivores, soil biota, predators of soil biota, and
xvn
-------
plants are included. Methods to assess risk to aquatic organisms and wildlife
preying on aquatic organisms follow from the surface runoff and groundwater
infiltration models. Surface water concentrations predicted by these models
are used to predict adverse effects on aquatic organisms or wildlife.
Example Calculations
Two chemicals have been selected to provide example calculations of risk:
benzo(a)pyrene [B(a)P] and cadmium (Cd). Both chemicals are used only as
examples for each methodology. The .examples are shown for the Human Exposure,
Terrestrial Food Chain, Surface Runoff, Groundwater Infiltration and Dermal
Exposure models. The purpose of the examples are to assist the reader in the
functional operation of the calculations for each methodology.
xviii
-------
LIST OF ABBREVIATIONS
AD Annual deposition rate of pollutant (g/m2 [year]"1)
ADI Acceptable daily intake (mg/kg/day)
AF Absorption fraction (%/day)
AFC Animal feed concentration (ug/day)
AQDM Air Quality Display Model
B(a)P Benzo(a)pyrene
BCF Bioconcentration factor (Jfc/kg)
B6/ED Block group/enumeration district
BI Background intake of pollutant (mg/kg)
BW Body weight (kg)
CA Contact amount (mg/cm2)
Cd Cadmium
CD Cumulative soil deposition of pollutant (kg/ha)
CDM Climato"logical Dispersion Model
Ci Concentration of contaminant i in receiving water (mg/£)
cm Centimeter
CRSTER Single Source Model
CT Contact time (hours/day)
DA Daily dietary consumption of animal tissue food group
(g DW/day)
DC Daily dietary consumption of crop food group (g DW/day)
DDI Human dermal daily intake (ug/day)
DI Human daily intake (ug/day)
DW Dry weight
EDA Exposure duration adjustment (unitless)
ESP Electrostatic precipitator
FA Fraction of food group (unitless)
FC Fraction of crop (unitless)
FF Fabric filters
FS Fraction of animal diet adhering to soil (unitless)
xix
-------
LIST OF ABBREVIATIONS (continued)
ha Hectare
HEM Human exposure model
1^ Human consumption of fish
Ig Soil ingestion rate (g DW/day)
I Total water ingestion rate (£/day)
W
ISC Industrial Source Complex
ISCLT ISC long term
ISCST ISC short term
k Loss rate constant (years *)
LC Maximal soil concentration (ug/g)
MEI Most-exposed individual
Mg Megagram
MS 2.7 x 103 Mg/ha = Assumed mass of upper soil layer (20 cm)
mt Metric tons
MWC Municipal waste combustor
PCDD Polychlorodibenzodioxins
PCDF Polychlorodibenzofurans
PFC Predator feed concentration (ug/g)
PM Particulate matter
PVC Polyvinyl chloride
q* Human cancer potency (mg/kg/day) 1
RDF Refuse-derived fuel
R£ Relative effectiveness of ingestion exposure (unitless)
Risk reference dose (mg/kg/day)
Adjusted reference intake (ug/day)
RI_ Risk level (unitless)
Reference water concentration (mg/£)
Exposed skin surface area (cm2)
Small modular incinerators
Total period of deposition (years)
xx
-------
LIST OF ABBREVIATIONS (continued)
Threshold feed concentration (ug/g DW)
Total background intake (mg/day)
2,3,7,8-tetrachlorodi benzodi oxi n
TPD Tons per day
UA Uptake response slope in animal tissue (ug/g feed DW
[ug/g]'1)
UB Uptake response slope in soil biota (ug/g [kg/ha]"1)
uc Uptake response slope in crops (ug/g [kg/ha]"1)
xi Concentration of leachate entering aquifer (mg/£)
xxi
-------
AUTHORS
0. Cleverly
Pollutant Assessment Branch
Office of Air Quality Planning and Standards
U.S. Environmental Protection Agency
Office of Air and Radiation
Research Triangle Park, NC 27711
L. Fradkin
Environmental Criteria and Assessement Office
Office of Health and Environmental Assessment
Office of Research and Development
Cincinnati, OH 45268
R. J. F. Bruins
Environmental Criteria and Assessment Office
Office of Health and Environmental Assessment
Office of Research and Development
Cincinnati, OH 45268
P. M. McGinnis
Center for Chemical Hazard Assessment
Syracuse Research Corporation
Syracuse, NY 13210
G. W. Dawson
ICF Northwest
Rich!and, WA 99352
R. Bond
ICF'Northwest
Rich!and, WA 99352
CONTRIBUTORS, AND REVIEWERS
N. C. Possiel
Air Dispersion Modeling
Source Receptor Analysis Branch
Office of Air Quality Planning and Standards
Office of Air and Radiation
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
G. J. Schewe
Air Dispersion Modeling
PEI Associates, Inc.
Cincinnati, OH 45246
xxii
-------
H. V. Geary
Air Dispersion Modeling
H. E. Cramer Company, Inc.
Salt Lake City, UT 84115
Technical assistance was provided by the Environmental Criteria and
Assessment Office and Northrop Services, Inc., Research Triangle Park, NC.
xxi n
-------
1. INTRODUCTION AND BACKGROUND
The Environmental Protection Agency (EPA) estimates that more than 150
million tons of municipal solid waste (MSW) are generated each year in the
United States (U.S. EPA, 1986a). Only about 5.8 million tons, or 3.7 percent,
of the MSW presently generated are incinerated in approximately 99 municipal
waste combustors (MWCs). By comparison about 137 million tons of MSW are
buried each year in about 10,000 municipal, county, and privately operated
sanitary landfills. Over 80 percent of MSW is disposed through landfill ing the
waste. Currently, EPA estimates that more MSW is recycled and recovered as raw
materials for manufacturing (12.7 million tons/year), than is incinerated.
Many states have implemented strict requirements governing the siting of
sanitary landfills, and the daily operation of landfills. States have imple-
mented rules requiring extensive geologic and hydro!ogic study of proposed
sites for new sanitary landfills. In many cases potential and existing sources
of groundwater must be identified, and extensive soil and geologic analysis
must be made to assess the potential for leachate contamination migrating off
the proposed site. In many states, existing landfills must conform to regula-
tions specifying daily operational management practices at landfills including
covering the disposed MSW, prevention of percolation of rain water through the
landfill, leachate collection and treatment processes, and the placement of
groundwater monitoring wells. Stricter environmental regulations are limiting
the land areas suitable for landfilling, as well as significantly increasing
the costs of siting, operating, and closing sanitary landfills.
Faced with increasing problems of land disposal, many communities are
actively considering or pursuing incineration as a disposal alternative. These
systems are designed for the recovery of heat from refuse combustion in the
form of steam or hot water to be used as an energy source to generate electric-
ity, to supplement energy demands of industry, and for use in district heating
and cooling of residential and commercial properties. In addition to heat
recovery, MWCs reduce the volume of waste requiring landfill ing by 70 to 90
October 1986 1-1 DRAFT-DO NOT QUOTE OR CITE
-------
percent, and therefore extend the operational life of existing landfills.
Without a reduction in the volume of waste that ultimately is landfilled some
urban areas will soon reach the design capacity of existing landfills and will
be compelled to select some method of MSW disposal or face a possible waste
disposal crisis.
The EPA estimates that significant growth in the population of MWCs will
occur in the United States between 1985 and the year 2000. In 1985 approxi-
mately 47,000 tons per day (TPD) of MSW was incinerated in 99 facilities. By
EPA estimates, this may rise to a daily volume of 156,000 TPD in 223 facili-
ties by the year 1990 (U.S. EPA, 1986a). In the year 2000 about 310 additional
MWCs may be incinerating about 252,000 tons of MSW per day. If it is assumed
that very few of the existing incinerators are permanently shut down over the
next 14 years, then it is possible that 33 percent of the projected MSW through-
put in the year 2000 may be incinerated in about 400 facilities nationwide.
All of the future population of MWCs are expected to be heat recovery systems,
whereas about 66 percent of existing MWCs have heat recovery. Data on plants
in the planning or construction stage suggest that MWCs with a total capacity
of over 1000 tons per day will make up over 50 percent of the new facilities
built by 1990.
The preferred solid waste management practice seems to be shifting away
from land disposal of MSW to incineration, and a large growth in MWCs is
expected to occur over a relatively short time period. A fundamental issue
attendant to this sudden growth is the total environmental impact that may
result from the emissions of pollutants from the incinerators. Ideally incin-
erator emissions should be evaluated according to the way in which they parti-
tion into various environmental media, i.e., air, water, and soil. Emissions
should be thought of as the the total pollution loading originating from the
incineration process, e.g., solid residues, aqueous residues, and stack gas
emissions. Such emissions should be studied in a manner that determines the
ultimate biological, physical and chemical disposition of the pollutants in the
human and natural environment, and in a manner that elucidates the ultimate
adverse effect to human health, and stresses to the natural environment.
Ideally the aggregate environmental assessment should be treated as an
audit and rely upon actual field measurements and observations of both
ecological and human health effects that can be statistically associated with
pollution loadings from MWCs. The environmental audit, however, should
encompass various technologies, various physical settings, and MWCs with
October 1986 1-2 DRAFT—DO NOT QUOTE OR CITE
-------
various degrees of pollution control. The audit would require extensive
monitoring in the land area that could be potentially impacted from emissions,
and would index and trace the movement of all pollutants through the environ-
ment that arise from the incineration process. The indexed pollutants would be
traced for accumulation through certain trophic systems, including the human
food web. Stressing of, and toxicity to, interconnected terrestrial and aquatic
ecosystems would be observed. Ambient air measurements of pollutants would be
measured. The biological and chemical kinetics of adsorption, absorption,
metabolism, transformation, and degradation would be recorded. The persist-
ence, decay, transformation, or magnification of pollutants transported through
the atmosphere, soil, water bodies, and biota would be measured. The human
population would be extensively observed, over a long period of time, to see if
pollutants given off during MSW incineration have any observable acute, sub-
chronic, or chronic effects on their health. To date an evaluation of this
magnitude and completeness has not been done on any operating municipal waste
combustor.
There exists a need to somehow evaluate the potential adverse environmen-
tal effects from MWC emissions. Lacking information from a complete environ-
mental assessment of representative MWC technologies, the EPA can turn to a
different set of tools, tools that involve computer simulation of the fate and
transport of chemicals in the environment. These tools, or models, attempt to
mimic actual circumstances, and use field and laboratory measurements of water
solubility, vapor pressure, octanol-water partitioning, bioconcentration
factors, soil adsorption constants, degradation rates, and half-lives, soil
characteristics, meteorology and climate as a means of simulating the environ-
mental movement, accumulation, partitioning into certain compartments, and
potential human exposure. With regard to the latter, there is a model that
can match the best estimates of the distribution of the human population to
predicted ambient air concentrations of containments. For the most part these
simulation models are emerging tools that can potentially enlarge the analytical
horizon and predict adverse effects before they can be observed in the field.
To be sure, there is a need for additional research as well as field studies to
refine these models.
In the past the EPA has focused only on the potential public health risks
posed by chronic stack or fugitive emissions of pollutants from combustion
sources. These risks were calculated as a result of a lifetime inhalation of
October 1986 1-3 DRAFT—DO NOT QUOTE OR CITE
-------
concentrations of pollutants in the ambient air. The Office of Air Quality
Planning and Standards (OAQPS) of the Office of Air and Radiation (OAR), and
the Environmental Criteria and Assessment Office (ECAO-Cincinnati) of the
Office of Research and Development (ORD) have collaborated in an effort to
develop a methodology to expand these analytical horizons. The methodology
permits the assessment of the potential human health risks posed by direct and
indirect exposure pathways resulting from the wet and dry surface deposition
and ambient air concentrations of pollutants emitted from the stacks of MWCs.
The pathways of exposure are depicted in Figure 1-1. The pollutants are
available for direct human inhalation when concentrations are predicted by
the models to be in the ambient air. The pollutants are available for dermal
absorption and ingestion when settled to the top layers of the soil, migrate to
underground aquifers, runoff into surface waters, or bioaccumulate into the
human food web following prolonged surface deposition. Finally, to the extent
practicable and feasible, and within the constraints of the environmental fate
and transport models, the methodology allows rough estimates to be made of the
possible adverse and toxicological effects on plant and animal life as a result
of exposure to deposited emissions from MWCs. The estimated effects include
estimates of soil biota toxicity, soil biota predator toxicity, phytotoxicity,
effects on herbivorous animals, and aquatic toxicity.
The major limitation to these analyses is that only the impact of stack
emissions is being considered. There are two other activities within EPA
addressing the potential adverse health effects arising from the land disposal
of solid residues from municipal waste combustion, e.g., fly ash and bottom
ash. The Office of Solid Waste is currently evaluating the potential leaching
of chemical constituents sorbed to the flyash matrix, and is undertaking health
risk assessment under the authority of the Resource, Conservation and Recovery
Act. The Office of Policy, Planning and Evaluation is conducting a comparative
risk assessment between land disposal of ash, land disposal of MSW., and MSW
incineration. The results of these analyses will be folded into the analysis
of MWC emissions when the projects have been completed in 1987.
The purpose of this report is to describe the multipollutant and multiple
exposure pathway risk assessment methodology that has been developed to evalu-
ate pollutant emissions from municipal waste combustors. The staff paper is
organized by descriptions of municipal waste combustor technology and quanti-
fied stack emissions (Section 2), methods for modeling of human exposure
October 1986 1-4 DRAFT—DO NOT QUOTE OR CITE
-------
DEPOSITION
ON WATER
DEPOSITION
ON GROUND
RUNOFF
PERCOLATION *^ V
DEPOSITION
ON FOOD
AND FEED
IRRIGATION
L
EATING
VEGETABLES
DRINKING
MILK
SOIL ,
INGESTION
EATING;
FISH
INHALATION
DRINKIN
WATER
DERMAL
ABSORPTION
UPTAKE
BY BIOTA
Figure 1.1. Potential Exposure Pathways to Pollutants Emitted from the
Stacks of Municipal Waste Combustors.
1-5
-------
pathways to emissions (Section 3), estimating carcinogenic and noncarcinogenic
risks to humans by indirect exposure to MWC emissions (Section 4), and methods
for assessing ecological effects (Section 5). The EPA recognizes there are
uncertainties inherent in the methodology. For example, the methodology can
be applied only to the finite number of chemicals that have actually been spe-
ciated in incinerator emissions. There are potentially hundreds of pollutants
that may be emitted but have not been looked for or measured. Emissions for
the most part were measured during good operations of the incinerator. There-
fore, emissions resulting from perturbations, mechanical failures, excessively
wet garbage feed, shutdown and cold start of the plant are not reflected in
the methodology. In terms of population risk assessment the effects of multiple
pollutant exposure is assumed to be additive. Synergism, antagonism, and po-
tentiation are not considered given the lack of biological information on the
effects of exposure to mixtures. There are other data limitations that ul-
timately constrain the methodology. The methodology therefore is not a compre-
hensive environmental audit, but is best regarded as an evolving and emerging
process that moves EPA beyond the analysis of potential effects on only one
medium (air), one exposure pathway (inhalation), and one animal species (hu-
mans), and proceeds to the consideration of other exposure pathways and on
other terrestrial and aquatic life.
October 1986 1-6 DRAFT—DO NOT QUOTE OR CITE
-------
2. MUNICIPAL WASTE COMBUSTOR TECHNOLOGY AND EMISSIONS
The EPA estimates there are currently 99 MWCs operating in the United
States with a total design capacity of about 47,000 tons per day of MSW input.
About 66 percent of these facilities are equipped for the recovery of energy in
the form of steam or hot water* during the combustion of refuse. Technologies
for heat recovery operations may be divided by three principle types of design:
massburn, small modular incinerators, and facilities that produce and combust
refuse-derived fuel (RDF). These designs are briefly and generically described
in the following subsection.
2.1 DESCRIPTION OF MWC TECHNOLOGIES
2.1.1 Massburn Facilities
Massburn is the predominant method of incinerating municipal solid waste.
The term massburn means that the raw MSW is combusted as received at the
facility without any preprocessing other than removing bulky items (stoves,
telephone poles, etc.) and mixing to produce a more homogenous fuel.
In a typical massburn incinerator, an overhead crane loads the waste from
the storage pit to the feed chute. The feed chute deposits solid waste on the
first, or dry-out, grate. Ignition starts at the bottom of the dryout grate
and is concentrated on the second, or combustion, grate. The third grate, a
burn-out grate, provides final combustion of the waste before the resulting ash
falls into the flooded ash pit and is sent to a landfill. In some cases,
ferrous metals are removed from the ash by magnetic separation. Figure 2-1 is
a schematic of a typical massburn overfeed stoker incinerator. These types of
incinerators are typically field erected. Individual incinerators can range in
size from 50 up to 1000 tons per day of capacity.
There are several types of grate systems in use with massburn incinera-
tors. All of these grate designs are similar in that they are designed to move
the waste through the incinerator and promote complete combustion. The grates
October 1986 2-1 DRAFT—DO NOT QUOTE OR CITE
-------
ro
i
ro
sniw
Figure 2-1. Mass Burn HSM incinerator with overfeed stoker grates.
-------
use either a traveling, rocking, or reciprocating motion to accomplish this.
In addition, one recently introduced design uses a rotary combustor rather than
a mechanical grate, followed by an inclined grate for final burnout.
The incinerator shown in Figure 2-1 has a waterwall furnace to recover the
energy from waste combustion in the form of steam. All new massburn incinera-
tors are expected to have waterwall furnaces for energy recovery. Many older
facilities have refractor lined walls rather than waterwalls. However, the
basic grate design is the same.
2.1.2 Small Modular Incinerators (SMI)
Combustion of MSW in SMI was introduced in the late 1960s. Small modular
incinerators are shop fabricated on a package basis. Individual modules were
originally limited to approximately 5 to ,50 tons per day (TPD) of MSW in size.
However, several manufacturers are currently marketing single modules which can
combust up to 100 TPD (Franklin et a!., 1982). The required plant capacity is
achieved by using multiple modules. The modular system allows the plant
capacity to easily be expanded as refuse generation increases.
Figure 2-2 presents the basic components of a SMI. The primary chamber is
fed using a hopper and ram feed system and ignited using a gas or oil burner.
Air is supplied to the primary chamber at substoichiometric levels. This
results in a lower air velocity in the primary combustion chamber than if
excess air was used and minimizes entrainment of fuel particles and ash in the
flue gas. The incomplete combustion products, primarily carbon monoxide and
low molecular weight hydrocarbons, pass into the secondary combustion chamber.
In the secondary chamber, excess air is added and combustion is completed. The
auxiliary burner shown in Figure 2-2 is an integral part of SMI and is fired
whenever the secondary chamber temperature falls below the set point
(Frounfelker, 1979). The resulting hot gases can be passed through a heat
recovery boiler for energy recovery. Although several existing SMI do not
have heat recovery, almost all new SMI projected for startup in the next 15
years will have heat recovery boilers.
The SMI previously described is typically called a controlled air or
starved air incinerator. There is another type of incinerator which also is a
modular design but operates differently. This incinerator is also ram fed, but
excess air is used in the primary chamber, and no air is added in the secondary
chamber. In this case, the secondary chamber simply provides additional
residence time to complete combustion.
October 1986 2-3 DRAFT-DO NOT QUOTE OR CITE
-------
H«t Recovery
Stack
—
*
I
— •
1
srsrs
t
1
ss
J
I IJ
I"~" By-pass Stack
.- — Damper
COMIUSTIOM
Figure 2-2. Small modular incinerator with heat recovery.
2-4
-------
2.1.3 Refuse-Derived Fuel (RDF")
One alternative to direct incineration of MSW is to process the waste to
produce RDF. The RDF produced may then be combusted on-site in a boiler
specifically dedicated to that fuel, or sold for firing as a fuel off-site.
Figure 2-3 presents a schematic of a representative RDF facility with an
on-site boiler. The advantages of producing RDF are that RDF is a more homoge-
nous fuel with a greater heat value per pound and requires less costly grate
designs to combust. Also, producing RDF allows a greater portion of the raw
MSW to be recovered for recycling.
The designs of dedicated boilers used to combust RDF are basically the
same as those for coal-fired boilers, and can include suspension, stoker, and
fluidized bed designs. If the RDF is sold as a fuel, it may be cofired with a
fossil fuel (usually coal). Existing dedicated RDF-fired boilers range in size
up to 1000 TPD (based on the raw MSW input into the RDF processing plant)
(Anonymous, 1985). Briefly, the various types of RDF which can be produced are
as follows.
Fluff RDF
Fluff RDF is prepared by mechanical shredding of MSW, followed by air
classification, magnetic separation, or trommeling to reduce the noncombustible
content of the waste stream. Coarse fluff RDF utilizes a single shredding
stage and can be fired in stoker type boilers, either alone or with coal. If
multiple shredding stages are used, fine RDF is produced. Fine RDF may be
fired in suspension boilers.
Densified RDF
Densified RDF (d-RDF) is produced by extruding fine RDF in a pellet mill.
It has the advantages of being much easier to store and transport than fluff
RDF. This makes d-RDF marketable as a fuel for existing stoker type coal-fired
boilers.
Powdered RDF
The production of powdered RDF requires mechanical, thermal, and chemical
processing of shredded MSW which has undergone screening and magnetic separa-
tion. Powdered RDF is fine enough to be cofired with fuel oil. It should also
be suitable for suspension type boilers, such as pulverized coal boilers.
October 1986 2-5 DRAFT—DO NOT QUOTE OR CITE
-------
MSW Receiving
and Storage Area
Primary
Shredder
Air
Qassifler
rx
/\ RefuseDerived
/ \ Fuel Storage
Aluminum
Magnet
Magnets
0«nsifl«rs
Residue Aluminum
Heavy
Fenoua
Ugnt
Ferrous
Stack
o
o
Boiler
W*'W*
r Metering
si/ -
Aan
System
Figure 2-3. RDF processing facility with on-site boiler.
2-6
-------
Wet-Pulped RDF
In the wet pulping process, the pulper is fed MSW which has been sluiced
with water. The pulper acts much like a kitchen blender to reduce particle
size. Noncombustibles are removed in a liquid cyclone. The RDF is then
mechanically dewatered to a moisture content of 50 percent. Wet-pulped RDF is
intended for use in a dedicated boiler on site.
2.2 DISTRIBUTION OF EXISTING MUNICIPAL WASTE COMBUSTION FACILITIES IN THE
UNITED STATES
The EPA has recently characterized the MWC industry currently operating in
the United States (U.S. EPA, 1986a). Approximately two-thirds of the 99
facilities that could be identified as currently operating are designed with
heat recovery boilers. Massburn facilities constitute 68.0 percent of current
installed incineration capacity. Twenty-five massburn facilities have heat
recovery, and 16 facilities do not recover heat during MSW combustion. Twelve
massburn MWCs have a capacity equal to or greater than 1000 tons per day,
15 facilities have a capacity in the range of 250-1000 tons per day, and 14
massburn facilities burn less than 250 tons MSW per day.
The RDF facilities represent 23.6 percent of current incinerating capaci-
ty. There are nine operating facilities in the U.S., all of which have energy
recovery units. Four of the nine RDF units are designed to process more than
1000 tons per day of MSW. One unit burns about 120 tons MSW per day, and four
facilities have a capacity in a range of 250-1000 tons per day.
The small modular incinerators (SMI) represent about 8.4 percent of U.S.
incineration capacity. There are about 49 SMIs currently operating of which
35 recover energy and 16 do not recover energy. Forty-eight SMIs have a design
capacity equal to or less than 250 tons MSW per day, and one SMI has a capacity
of 250-500 tons MSW per day.
Figure 2-4 summarizes the percent distribution by technology of existing
MWC facilities. Figure 2-5 shows the distribution of MSW facilities regionally
throughout the United States. The majority of MWC facilities currently operat-
ing in the U.S., (35 of the 99 existing facilities) are located in the New
England and Mid-Atlantic states. The pattern of distribution seems to follow
population density with those areas of the U.S. with the highest population
densities having the most numbers of MWCs. The Mountain and Pacific states
October 1986 2-7 DRAFT-DO NOT QUOTE OR CITE
-------
THE PERCENT OF TOTAL DESIGN CAPACITY
DISTRIBUTED AMONG EXISTING FACILITIES
SMALL MODULAR INCINERATORS
MASS BURN (68.0%)
(8.4%)
(23.6%)
Figure 2-4 Distribution of existing Installed MWC facility
capacity by design type.
2-8
-------
3-
NEW ENGLAND SOUTH
AND MID-ATLANTIC CENTRAL
SOUTH
ATLANTIC
REGION
NORTH MOUNTAIN AND
CENTRAL PACIFIC
Figure 2-5 Distribution of existing Installed MWC facility
capacity by region.
2-9
-------
have the least numbers of MWCs of any region with a total of eight existing MWC
incinerators. Public and regulatory concerns have impeded the development of
the MWC industry in California. The EPA will continuously revise the statis-
tics on existing MWCs as additional information is developed.
2.3 DISTRIBUTION OF PLANNED MUNICIPAL WASTE COMBUSTION FACILITIES IN THE
UNITED STATES
The EPA estimates there are 185 MWC facilities in the United States that
are either in some conceptual planning stage or that are under construction
(U.S. EPA, 1986a). All of these facilities are heat recovery systems. Mass-
burn technology is expected to predominate with more than 58 percent of the
total design capacity (106,256 tons per day) accomplished with 107 facilities.
Forty-nine massburn facilities (46%) will have a design capacity equal to or
greater than 1000 tons MSW per day, 25 facilities will have a capacity of
between 500-1000 tons per day, 18 facilities will have a capacity of 250-500
tons per day, and 15 facilities are expected to have a design capacity of less
than 250 tons per day.
RDF units are expected to be about 23.7 percent of the total design
capacity with 30 facilities either planned or under construction. Seventeen
facilities (57%) will have a design capacity equal to or greater than 1000 tons
per day, nine facilities (30%) will have a capacity of 500-1000 tons per day,
one RDF facility will have a capacity of 250-500 tons per day, and three
facilities will have a design capacity less than 250 tons per day.
The small modular incinerators (SMI) will account for only 2.6 percent of
the total planned capacity. This will be accomplished with 22 new units. Nine
SMIs will have a design capacity of between 250 and 500 tons per day, and
thirteen units will be less than 250 tons per day in capacity.
Figure_2-6 displays the percent of total design capacity by technology
among planned MWC facilities. It should be noted that the planned total
incineration capacity (106,256 tons per day) is approximately two and one-half
times the current incineration capacity of existing MWC facilities. For 26
facilities in the planning stage, data on the type of incinerator design was
either unavailable, or had not yet been decided. Figure 2-7 compares the
regional distribution of existing MWC facilities in the United States with the
October 1986 2-10 DRAFT—DO NOT QUOTE OR CITE
-------
THE PERCENT OF TOTAL DESIGN CAPACITY
DISTRIBUTED AMONG PUNNED FACILITIES
MASS BURN (58.4X)
SMALL MODULAR INCINERATORS (2.6X)
RDF (23.7X)
UNDECIDED/NOT AVAILABLE (15.3X)
Figure 2-6 Distribution of planned MSW Incinerator
capacity by design type.
2-11
-------
Figure 2-7. Projected Growth of Municipal
Waste Combustion Between
1986 and 1990's.
ro
i
MUNICIPAL WASTE COMBUSTOR LOCATIONS
1986
MUNICIPAL WASTE COMBUSTOR LOCATIONS
1990's
-------
anticipated growth of MWCs by 1990. Sixty-two facilities (48%) are either
planned or under construction in the New England and mid-Atlantic states, and
thirty-seven (29%) of the planned facilities will be built in the mountain and
Pacific states. The remainder of the planned facilities (29) will be geograph-
ically distributed as follows: north Central states - 9 facilities; south
Atlantic states - 13 facilities, and south Central states -. 7 facilities.
Table 2-1 summarizes the anticipated growth of planned MWC facilities among
states with the greatest rate of growth.
TABLE 2-1. PLANNED MSW COMBUSTION FACILITIES FOR
STATES WITH THE GREATEST GROWTH
State
California
New York
Connecticut
New Jersey
Massachusetts
Florida
Virginia
Pennsylvania
Number
52
20
17
9
10
13
7
33
Planned Facilities
Capacity, TPD
13,417
18,306
12,030
10,635
9,975
18,410
13,361
24,852
2.4 PROJECTED GROWTH OF MUNICIPAL WASTE COMBUSTION THROUGH THE YEAR 2000
Subsection 2-3 reviewed the characterization of planned facilities, and
facilities currently under some phase of construction. Most of these facili-
ties will be constructed by 1990 if there are not serious delays in implementa-
tion. The EPA has made projections of the growth of the MWC source category
through the year 2000. These projections are largely based on market surveys
and projections in the increase in the generation of municipal solid waste
(MSW) (U.S. EPA, 1986a). The projected increase in MWC includes new facilities
expected to be built and actually begin operating by the end of 1986.
Market analysis from several sources indicates a spread in the projected
total number and capacity of facilities out to the year 2000 (Franklin et al.,
1982; U.S. EPA, 1986a). Estimates indicate that about 310 new facilities will
be on-line and operating by the year 2000. Perhaps as much as 252,000 tons per
day of MSW will be processed in these new facilities.
Based on data on planned facilities, massburn waste-to-energy facilities
will likely dominate, and will account for between 60 and 70 percent of the
October 1986 2-13 DRAFT—DO NOT QUOTE OR CITE
-------
total population of refuse incinerators. Facilities that incinerate RDF will
constitute between 20 and 30 percent of the projected facilities by the year
2000, and modular systems may account for as much as 20 percent of the market.
Data on plants currently being planned, designed or constructed suggests that
over 50 percent of incineration capacity will be handled in facilities in
excess of 1000 tons per day in size. However, facilities in the capacity range
of 500-1000 tons per day will also be popular because the number of cities
having a municipal solid waste stream sufficient to sustain a plant greater
than 1000 tons per day are limited.
The net growth of MWCs in the year 2000 is dependent on the growth in new
incinerators minus the closing of existing facilities. The EPA estimates that
the number of existing facilities to be retired or closed over the next 15
years will be small in number. This is because the majority of existing MWCs
were built since 1970, and therefore will only be 30 years in operation by the
year 2000. There is expected to be an economic incentive to rebuild some aging
facilities instead of replacing them entirely with more expensive, newer
technology. A MWC in Oceanside, New York, for example was commissioned in 1965
and extensively modernized in 1977. The Betts Avenue incinerator in New York
City was built in 1965, modified in 1980, and is being considered for further
renovation. Thus, the total capacity of MWC by the year 2000 may be as high as
295,000 tons MSW per day in about 400 facilities. If this occurs then about
one-third of the MSW expected to be generated in the U.S. by the year 2000 will
be disposed of through incineration. Table 2-2 summarizes the increase in
incineration capacity to the year 2000.
TABLE 2-2. PERCENTAGE BY REGION OF THE FORECAST WASTE
TO ENERGY THROUGHPUT 1985 to 20001
Region
New England and Mid-Atlantic
North Central
South Atlantic
South Central
Mountain and Pacific
Throughout,
1985
45
19
28
7
1
Percent of
1990
45
13
21
6
15
Total
2000
45
13
20
7
14
1U.S. Environmental Protection Agency (1986a).
October 1986
2-14
DRAFT—DO NOT QUOTE OR CITE
-------
2.5 MUNICIPAL WASTE COMBUSTOR EMISSIONS
The EPA has collected information on the stack emission of participate
matter, organic pollutants, metals, and acid gases from MWC (U.S. EPA, 1985a).
This subsection will summarize these data, and more detailed analysis is dis-
cussed elsewhere (U.S. EPA, 1985b). Surveys of the pertinent literature were
performed including, contacts with other Federal agencies, trade organizations,
incinerator manufacturers, and incinerator operators. Table 2-3 is a list of
pollutants that have been measured in the stack emissions of MWCs that current-
ly will be employed in the risk assessment methodology. Table 2-4 summarizes
the number of facilities producing information on actual emissions testing.
These MWCs were or are operating in the U.S. and Canada. It is likely that
different sampling and analytical methods were used to (quantify emissions at
various facilities, so the data may or may not be directly comparable. The EPA
is currently reviewing many test reports to critically evaluate the validity of
the data. For the most part emissions of particulate matter emitted from MWC
seem well characterized. Measurements have been made at a number of facilities
representing a wide range of incinerator technologies. The extent of testing
is a consequence of regulating particulate emissions from MWCs. The most
significant data gap in the inventory of MSW incinerator emissions is the
speciation of halogenated and nonhalogenated organic compounds. In addition,
data were not found on the potential emission of asbestos, pathogens or radio-
nuclides. There is a more complete data set on the emission of trace elements
from conventional massburn incinerators, but not from RDF or modular facili-
ties. The EPA's assessment of MWC emissions is severely constrained by the
available published data. The EPA has recently reported the results of stack
sampling organic and inorganic emissions from a massburn, RDF, and a modular
facility, and the utilization of these data will extend the current data base
(U.S. EPA, 1986b). The report an'd analysis were approved for final publication
in September, and EPA intends to use the data for risk assessment purposes.
In addition EPA will spend $1.4 million on testing representative MWC
technologies during 1987 to better characterize the emission profile from
incinerators. These data will also be utilized in the risk assessment
methodology when they are available.
The multiple pollutant, multiple exposure pathway risk methodology will
first be applied to an assessment of the impact of stack emissions from mass-
burn incinerators. The population of incinerators are too numerous to execute
October 1986 2-15 DRAFT—DO NOT QUOTE OR CITE
-------
TABLE 2-3. POLLUTANTS QUANTIFIED IN MUNICIPAL COMBUSTION EMISSIONS
THAT ARE POTENTIAL CANDIDATE POLLUTANTS FOR RISK ANALYSIS
Inorganic Compounds
Organic Compounds Metals Acid Gases
*Chlorobenzenes *Cadmium Hydrogen Chloride
Phenol *Beryllium Sulfur Dioxide
*Benzo(a)pyrene (PAH) Lead Hydrogen Fluoride
Naphthalene *Chromium
*Perchloroethylene Mercury
*Chlorodi benzodi oxi ns *Arseni c
Chlorodibenzofurans Selenium
*Benzene Nickel
*Formaldehyde Copper
*Carbon tetrachloride
*Chloroform
*Polychlorinated biphenyls
Chlorophenols
"'Estimates of carcinogenic potency are available.
TABLE 2-4. SUMMARY MATRIX OF EMISSIONS DATA GATHERED FOR MSW INCINERATORS0
Pollutant
Uncontrolled PM
Controlled PM
Metals
Arsenic
Beryllium
Cadmium
Chromium
Mercury
Nickel
Lead
Acid Gases
HC1
HF
Organic Acids
POM
Dioxins and Furans3
Number of Facilities Tested
Massburn
15
36
3
1
4
4
1
4
4
16
6
2
6
RDF
4
4
1
1
2
1
1
2
4
Modular Other
5 4
2 1
1
1 1
1
lb
Total
28
43
3
2
5
5
4
5
5
20
6
3
11
Includes U.S. and Canadian facilities only.
Coal -fired boiler firing 15 percent RDF/85 percent coal.
^
EPA is planning to stack test additional MWCs to determine more current
information regarding incinerator emissions.
October 1986 2-16 DRAFT—DO NOT QUOTE OR CITE
-------
the methodology on a pi ant-by-plant or technology-by-technology basis. EPA
does not intend to ignore the other technologies, however the state of develop-
ment of the various models to the methodology requires an extensive resource
commitment. Therefore, EPA has selected massburn incinerators for evaluation
of the methodology and to serve as an example of its application. Massburn
incinerators have more frequently been stack tested for emissions than
other technologies (refer to Table 2-4). In addition, the predominant technol-
ogy both in terms of existing and planned facilities are or will be massburn
MWCs. Therefore, discussion of potential stack emissions of pollutants will
focus specifically on the massburn technology, although EPA has gathered
emissions data on all MWC technologies (U.S. EPA, 1985a).
There are many factors that may influence the emissions of particulate
matter, metals, organic compounds and acid gases. Particulate matter (PM) is a
consequence of incomplete refuse combustion and the entrainment of
noncombustibles in the combustion gas stream. Particulate matter may exist in
a solid state or as an aerosol, and may contain adsorbed and absorbed heavy
metals and polycyclic organic compounds. Inorganic and organometallic sub-
stances in the refuse contribute to PM formation. The fuel molecules them-
selves contribute to PM formation via pyrolytic reactions, and inorganic
compounds may act as nucleation sites to induce PM formation. The size and
emission rate of uncontrolled PM is thought to depend on furnace residence
time, temperature, oxidation-reduction conditions and trace chemistries of the
PM and fuel. Increased residence times and temperature decreases particulate
size. When oxidation conditions predominate over reduction, particle size
decreases (U.S. EPA, 1985a).
Refuse may contain sources of trace metal air emissions from municipal
incineration. Chromium, lead, zinc, and selenium are used as surface coatings,
galvanizing compounds, and solders. Plastic objects composed of polyvinyl
chloride (PVC) may contain cadmium heat stabilizing compounds. Cadmium is also
found in inks, paints, and batteries. Lead can be found in certain types of
inks and paints, and batteries can be a source of nickel and mercury. High
temperature combustion tends to volatilize trace metals from the raw refuse.
In general trace organic emissions are a consequence of incomplete combus-
tion, and low combustion temperatures. Inefficient operation of the furnace
and the chlorine content of the fuel may contribute to the formation of chlori-
nated organics. The EPA has not parametrically tested a MWC to observe the
October 1986 2-17 DRAFT-DO NOT QUOTE OR CITE
-------
potential formation of products of incomplete combustion (PICs), however, the
EPA intends to do so in the winter months of 1987. Gross observations can be
made concerning combustion and the generation of PICs from the testing and
evaluation of hazardous waste incinerators (U.S. EPA, 1984a). Evaluation of 8
incinerators failed to define parametric relationships between residence time,
temperature, heat input, oxygen concentration and the distruction rated effi-
ciency of hazardous organic compounds. Carbon monoxide (CO) and total hydro-
carbons (THC) were monitored continuously during emissions testing to evaluate
as potential surrogates for organic compound emissions. The analysis indicated
that CO and THC may provide indication of changes in incinerator performance
and gross malfunctions in the combustion process, but were not good predictors
of PIC emissions. Generally the incinerators performed at a combustion effi-
ciency greater than 99 percent. Data from the tests suggested that complex
side reactions including recombination of molecular fragments may form actual
incomplete combustion products as by-products of the combustion of any organic
wastes, e.g., benzene, toluene, chloroform, tetrachlorethylene, and PAHs.
Polychlorodibenzo-p-dioxins (PCDDs) and polychlorodibenzofurans (PCDFs)
are a related group of tricyclic aromatic hydrocarbons that have been measured
in the stack exhaust gas of every MWC tested that exclusively incinerates
municipal solid waste. The presence of PCODs and PCDFs in flue gases of MWCs
can potentially be explained by several mechanisms: 1) PCDDs and PCDFs are
trace impurities in the fuel or feedstock used to sustain combustion, and are
not sufficiently thermally destroyed during combustion; 2) PCDDs and PCDFs are
produced during the combustion of chlorinated precursors, e.g., PCBs, chloro-
phenols, chlorobenzenes, phenoxyacetic herbicides, and fire retardants; 3)
PCDDs and PCDFs are formed as a consequence of a complex array of pyrolytic
processes involving chlorinated aliphatic and nonchlorinated aromatic compounds
that are chemically unrelated, e.g., PVC, DDT, polystyrene, cellulose, and
lignin, in conjunction with hydrogen chloride gas or free chlorine. The EPA
statistical analysis (using Spearman rank-order techniques) of MWC emissions
indicate that, as a general trend, emissions of PCDDs/PCDFs appear to be
inversely related to furnace temperature, that low combustion efficiency may
promote formation, and that fly ash emissions of PCDDs and PCDFs are not
related to furnace conditions alone, but may be dependent on a number of
complex variables related to function of the incineration process and design
(U.S. EPA, 1986c) including post-combustion formation phenomena outside the
firebox region.
October 1986 2-18 DRAFT—DO NOT QUOTE OR CITE
-------
The compounds found in MSW are complex and variable, and thus mapping any
mechanisms for formation of specific toxic organic compounds is an improbable
task. It is difficult to predict how constituents present in the fuel will
break apart, reform new compounds, dissociate into free radicals, or recom-
bine under the influence of oxygen, turbulent mixing and temperature. There-
fore, the aforementioned formation pathways are meant to be general in nature.
The EPA has compiled data on organic and inorganic emissions from massburn
incinerators to be used in the risk assessment. Table 2-5 summarizes the
limited organic emissions data found in published reports of massburn MWCs
operating in the United States. Emission values for Chicago NW incinerator and
the Peekskill, N.Y. incinerator were derived as average values from one test
report (U.S EPA, 1983a; NYDEC, 1986). The emission values for the Hampton, VA,
incinerator are mean values from three separate test reports and a total of
eleven independent analyses (Haile, 1984; U.S. EPA, 1983b; Scott Environmental
Services, 1985). The published data are limited. However, EPA has tested an
additional massburn MWC, but has only just published the results (U.S. EPA,
1986c). In order to extend the number of compounds to be included in the
assessment of potential risk to the population exposed to MWC emissions, recent-
ly published data of organic emissions will be utilized (U.S. EPA, 1986b).
Appendix A summaries these data.
Acid gases and sulfur dioxide are emitted from the combustion of municipal
solid waste. The chlorine content of MSW contributes to the emission of
hydrogen chloride gas (HC1). Conversion efficiencies of the fuel chlorine to
the emission of HC1 have been determined to be about 60-65 percent (U.S. EPA,
1985a). Chlorine in MSW is present in all combustible components of the
refuse, and ranges from about 0.5 percent mass to 0.9 percent mass-(Domalkski
et al., 1986). The National Bureau of Standards has determined that the paper
fraction of MSW contributes about one-quarter to one-half of the chlorine, and
the plastic fraction contributes about one-quarter to one-half of the chlorine
(Domalkski et al., 1986). The major amounts of the chlorine in MSW, therefore,
are contained in the paper and plastic fractions. Sulfur dioxide is the
predominant form of-sulfur emitted from MWCs, and is related to the sulfur
bound to the municipal solid waste. Mass balance analyses have shown that the
percentage of fuel sulfur converted to S0£ ranges from 14 to 90 percent (CARB,
1984), and is dependent on combustion conditions and refuse composition.
Compositional analyses have shown the refuse content of sulfur to range from
October 1986 2-19 DRAFT—DO NOT QUOTE OR CITE
-------
TABLE 2-5. EMISSIONS DATA FOR MASSBURN MWC FACILITIES
Chicaqo
Organic species
2,3,7,8-tetra-CDD
2,3,7,8-tetra-CDF
Total tetra-CDD
Total tetra-CDF
Total penta-CDD
Total penta-CDF
Total hexa-CDD
Total hexa-CDF
Total hepta-CDD
Total hepta-CDF
Total octa-CDD
Total octa-CDF
Total tetra thru octa CDD
Total tetra thru octa CDF
Total PCB
Formal dehyde
B(a)P
Total chlorinated benzene
^ Total chlorinated phenol
ro
0 Process Data
pg/sec
0.0101
0.155
2.21
—
0.403
1.53
0.187
0.184
0.0624
0.0148
—
—
1.04
—
—
43.6
88.2
pg/Mg
2.07
31.6
453
—
—
82.4
313
38.3
37.6
12.8
3.03
—
—
212
—
— —
8,920
18,000
Hampton
Mg/sec
0.142
1.19
2.78
12.6
6.43
19.4
6.44
8.45
6.00
6.17
1.56
0.433
23.2
47.1
5.17
—
71.4
291
1,050
pg/Mg
117
1,130
2,330
10,300
5,380
15,600
5,490
7,170
5,320
5,190
1,330
367
19,600
38,700
3,980
—
54,500
117,000
905,000
Peekskill
jjg/sec
0.0092
0.0705
0.0926
0.97
0.0923
0.572
0.126
0.590
0.181
0.343
0.291
0.0126
7.61
2.50
— —
19,800
"
——
~"
pg/Mg
1.17
8.95
11.8
124
11.7
72.6
16.0
74.9
23.0
43.6
36.9
1.60
966
317
_ "•
2,510,000
•" •
Stack flow rate
Waste feed rate
C02 concentration
Design feed rate
1,480 dNmVmin
17,600 kg/h
8.97%
400 tons/day
393 dNmVmin
4,380 kg/h
8.73%
125 tons/day
28,350 kg/h
9.61%
750 tons/day
CDD = chlorinated dibenzo-p-dioxins.
CDF =xchlorinated dibenzofurans.
PCB = polychlorinated biphenyls.
B(a)P = benzo[a]pyrene.
-------
0.2 percent to 0.37 percent on a dry weight basis (CARB, 1984). Mechanisms
involved in the release of fluorine and subsequent conversion to hydrogen
fluoride at MWCs are similar to HC1 formation, except conversion rates are not
known (CARB, 1984).
Levels of uncontrolled HC1 emissions at massburn MWCs have been measured
in a range of 110 to 605 ppm (v) (U.S. EPA, 1985a). The concentration of HF in
uncontrolled massburn emissions have been measured in a range of about 4.0 to 4
ppm (v) (U.S. EPA, 1985a; U.S. EPA, 1986b). Uncontrolled sulfur dioxide emis-
sions in massburn incinerators ranges from 0.72 to 159 ppm (v) (CARB, 1984).
Inorganic pollutants measured in MWCs are shown in Table 2-6. The speci-
fic tests of individual MWCs generating these data are reviewed in a separate
EPA report (U.S. EPA, 1985a). The mean values of metal emissions are a mathe-
matical average of reviewed MWC tests reported in the literature. In EPA's
risk assessment the mean values of all the tested massburn energy recovery
facilities in this inventory were used in assessing the potential risk from the
future population of MWCs. Because the Hampton incinerator has been tested
more frequently for emissions than any other MWC operating in the U.S., the EPA
has selected this particular facility to represent the potential environmental
impact of emissions from existing incinerators. The site is adequate in terms
of application of the multiple pollutant human risk and ecological effects
analysis since a regional drinking water reservoir, sensitive tidal basins,
sensitive agriculture, and residential areas are located nearby. Table 2-7
displays the emission rate of inorganic pollutants of the Hampton, VA, MWC from
a recently published report (U.S. EPA, 1986b). This information is used in the
dispersion modeling of the Hampton MWC to characterize the potential risk from
existing MWCs. These data will be used in exposure and risk assessment of the
existing population of MWC. The next section describes the modeling procedures
to determine human exposure to MWC emissions to be evaluated in the risk
assessment.
October 1986 2-21 DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 2-6. UNCONTROLLED METALS EMISSION FACTORS FOR MWCs
Weight Emission factor, mg metal/Mg feed
fracton, ug/g Mean
Metal
Arsenic
Beryllium
Cadmium
Chromi urn
Lead
Mercury
Nickel
Mean
140
3.9
1,100
330
18,000
640 .
890
Mass burn
2,500
70
20,000
5,900
320,000
12,000
16,000
Modular
110
3.0
850
250
14,000
490
690
RDI-
5,600
160
44,000
13,000
720,000
2,500
36,000
Notes:
mg: milligram (10~3 grams).
Mg: megagram (metric ton) of MSW incinerated.
ug: microgram (10 6 grams).
g: gram.
TABLE 2-7. CONTROLLED INORGANIC EMISSIONS FROM THE HAMPTON MWC
(average of 3 tests)
Arsenic
Beryllium
Chromi urn
Lead
Cadmium
Nickel
Mercury
ug/dscm
235.00
0.02
287.00
9613.00
503.00
227.00
2400.00
grams/sec
1.40 x 10"3
1.20 x 10"7
1.70 x 10 3
5.80 x 10"2
3.03 x 10"3
1.37 x 10"3
1.50 x 10~2
Notes:
|jg/dscm = micrograms of pollutant per day standard cubic meter of combustion
gas.
grams/sec = grams of pollutant emitted per second of plant operation.
Source: U.S. Environmental Protection Agency, 1986b.
October 1986 2-22 DRAFT—DO NOT QUOTE OR CITE
-------
3. EXPOSURE MODELING OF MUNICIPAL WASTE COMBUSTOR EMISSIONS
Typically a municipal waste combustor (MWC) is designed for a 30-year
working life. Most of these systems will, for the most part, be continuously
operated, except during periods of routine maintenance. Continuous operations
are preferred in order to maximize the production of steam or hot water neces-
sary to meet contract requirements in the sale of energy to the user. The
population surrounding the facility will be exposed to organic and inorganic
pollutants emitted from the stack during combustion. The degree of exposure
will be variable, and will be dependent on conditions of plant operations,
design of the incinerator, the character of the feed to the incinerator, the
technical specifications of the pollution control equipment, local meteorology,
local terrain, and the population density and distribution.
Pollutants emitted from MWCs into the atmosphere may be distributed across
environmental media (the atmosphere, soil and water) as a result of complex
mechanisms, most of which are just beginning to be understood. Some pollutants
are emitted in combustion gas adsorbed onto the surface of particulate matter,
while other pollutants remain in a gaseous or vaporous state. Some pollutants
may photodegrade, undergo complex chemical reactions, or combine with other
pollutants when transported through the atmosphere. Particulate-bound pollu-
tants may ultimately settle on the earth's surface by the forces of gravity.
Periods of precipitation may increase the rate of surface deposition by washout
of adsorbed and gaseous pollutants while passing through the emission plume.
Once deposited on the surface the pollutants may again be physically, chemical-
ly, or photolytically transformed, may be persistent, or may be transported by
the action of wind and precipitation to other environmental compartments.
Deposited materials may even be resuspended into the atmosphere, and once again
become air pollutants. The net effect is that human exposure to incinerator
emissions results not only from direct inhalation of ambient air concentrations
of the pollutants, but also indirectly from skin contact of the pollutants, and
ingestion of contaminated soil particles, water and food. Detailed
October 1986 3-1 DRAFT—DO NOT QUOTE OR CITE
-------
experimental evaluation of the environmental fate and environmental transport
of MWC emissions have not been done under actual conditions (Yoram, 1986).
Therefore, mathematical models of the fate and transport of pollutants entrained
in the stack exhaust gas are currently the most feasible alternative to the
assessment of human exposure to MWC emissions. These models can also be used
to estimate bioaccumulation in the natural ecosystem, potential accumulation of
pollutants adverse to the promotion of animal and plant life, and accumulation
of pollutants into the human food chain. The models specifically used in this
analysis of MWC emissions are: the Industrial Source Complex Short-Term Air
Dispersion Model; the Human Exposure Model; the Terrestrial Food Chain Model,
the Surface Runoff Model, and the Groundwater Contaminant Model.
These models vary somewhat in their approach to estimating potential
human exposure and risk. The preferred approach is to determine levels of
risk in the entire exposed population and numbers of individuals at each risk
level, or the aggregate risk. Another approach is to more narrowly focus on
describing the exposure and risk of only those individuals with the highest
exposure potential, or the most-exposed individuals (MEIs).
When assessing effects from inhalation of incinerator'emissions, it is
feasible to estimate the aggregate risk, since the ISC-ST (see Section 3.1)
estimates the pattern of ground-level emission concentrations within a radius
of 50 km from the facility, and the HEM (see Section 3.2) contains data on
human population distributions for the entire U.S. It is reasonable to assume
'that individuals residing at specific locations within this radius are exposed
at the concentrations predicted by the ISC-ST for those coordinates. The risk
to the MEI, or an individual residing in the area of highest predicted average
annual ground-level concentration, is also estimated.
For most other exposure pathways, however, it is not currently feasible
to estimate the aggregate risk, and therefore exposure and risk are quantified
for the MEI only. Human exposure to deposited contaminants via the terrestrial
food chain varies widely according to not only the pattern of deposition, but
also patterns of land use and individual living habits such as personal food
production. Therefore the TFC model (see Section 3.3) examines only a hypothet-
ical farm family living within the facility vicinity and producing a substan-
tial proportion of their own vegetables or meat and dairy products for consump-
tion. No attempt is made, however, to estimate the numbers of these individuals,
or to estimate the numbers' of people who could purchase smaller quantities of
October 1986 3-2 DRAFT—DO NOT QUOTE OR CITE
-------
these foods or who maintain only small gardens, as these practices would be
highly variable within the exposed population.
The Surface Runoff Model (see Section 3.4) presented in this document
estimates in stream contaminant concentrations resulting from runoff from a
watershed of a given size and average deposition rate. If the location of
drinking water intakes and number served by each is known,, it is possible
using this model to estimate the concentration and number exposed for each
drinking water source. Recreational fish consumption would be more variable
and therefore more difficult to quantify. In this document, however, it is
assumed for purposes of example that runoff from a watershed with an area of 1
km , having the MWC facility at its center, enters a stream of moderate size
o
(1 m /sec). The MEI is an individual receiving drinking water and fish solely
from this stream. No attempt is made to estimate aggregate risk.
Subsurface patterns of water movement are very complex, and thus it would
be nearly impossible to estimate groundwater contaminant concentrations
throughout a large area (radius = 50 km) over which deposition rates vary. It
would therefore be difficult to estimate exposures based on the numbers and
locations of all drinking water wells within this area. Therefore it is not
considered feasible to determine aggregate risk using the Groundwater
Contamination Model (see Section 3.5). In this document an example is shown
which assumes that an area with a radius of 200 m around the MWC facility
represents the recharge area for an aquifer feeding a small well. An
individual using this well as a drinking water source represents the MEI.
Dermal exposure to deposited particulate can also be expected to vary
greatly according to individual habits. It may be possible to assume that
some average amount of contact with soil can be applied to the whole
population in the area of deposition, and therefore dermal exposure patterns
for emitted contaminants can be estimated as is done for inhalation exposure.
However, it probably is not worthwhile to do so because other uncertainties in
the Dermal Exposure Model are so great by comparison (see Section 3.6).
Dermal exposure and risk are therefore estimated only for an MEI residing in
the area of maximum'-deposition.
The EPA is estimating the potential for adverse health effects in the
population exposed directly and indirectly to pollutant emissions from MWCs.
Given the complexities of predicting the environmental fate and transport of
specific chemicals emitted, as well as predicting multiple routes of human
October 1986 3-3 DRAFT-DO NOT QUOTE OR CITE
-------
exposure to specific chemicals, it is not currently feasible nor practical to
apply the models to every existing MWC, or planned MWC. Therefore, EPA has
decided to simplify the modeling process by using a model plant to characterize
the potential adverse impacts of emissions from technologies typical of MWCs
currently being planned or constructed, and the Hampton MWC to represent the
potential adverse impacts of air pollutant emissions from existing MWC technol-
ogy. A dichotomy is made between existing and planned technology with the
assumption that MWC technology currently marketed represents a distinct improve-
ment in design, operations and pollution control when compared to facilities
built a decade ago. In actual practice it is not only the design that is
essential to good incinerator performance, but operator training and equipment
maintenance are also important. The model plant configuration will closely
mimic actually planned massburn facilities in terms of stack height stack (and
diameters) building area, and pollutant emissions, but will reflect the upper-
end ,of capacities currently planned. The Hampton facility was planned in
the mid-1970's, and put into operation in September 1980. EPA has accumulated
several test reports of emissions of organic and inorganic pollutants at
Hampton, more so than any other specific incinerator site. Table 3-1 repre-
sents the descriptive parameters used in air dispersion modeling of the model
plant, and Table 3-2 are the modeling parameters of the existing MWC at
Hampton, Virginia.
"- The model plant is a massburn waterwall energy recovery MWC incinerating
2727 metric tons per day of raw, unprocessed municipal solid waste. The
configuration of the plant represents the upper end of the capacity range of
massburn MWCs EPA predicts will be the dominant technology through the year
2000. The EPA selected the stack height, diameter, exit velocity of the gases
and temperature of the exhaust gas based on actual measurements at facilities
of this size. The model plant will be located, for modeling purposes, in a wet
and humid climate in a geographical area where the MWC industry is expected to
grow rapidly over the next 15 years. The climate conditions were selected to
maximize the potential of-surface deposition of the emitted pollutants with the
physical action of washout during periods of precipitation. Another criteria
of selecting an appropriate site for the model plant was soil characteristics
that would enhance the movement of pollutants from the surface into underground
sources of drinking water by percolation through ground layers. On this basis
Western, Florida was selected as an appropriate site for the location of the
October 1986 3-4 DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 3-1. MODELING PARAMETERS FOR THE MODEL PLANT
Massburn Municipal Waste Combustor (heat recovery)
Capacity: 3000 TPD (2727 Mg/d)
Location: Western, Florida
Latitude: 27° 57'
Longitude: 82° 27'
Stack Height: 46 meters (m)
Stack Diameter: 3.1 m
Stack Exit Velocity: 11.3 m/s
Stack Temperature:
(a) 470° K with ESP
(b) 443° K with Fabric Filters (FF)
Building Configuration:
H: 35 m
W: 76 m
L: 42 m
TABLE 3-2. MODELING PARAMETERS FOR HAMPTON, VA
Massburn Municipal Waste Combustor (heat recovery)
Capacity: 120 TPD (109 Mg/d)
Latitude: 37° 6' 02"
Longitude: 76° 23' 28"
Stack Height: 27.44 m (2 stacks)
Stack Diameter: 1.22 m (2 stacks)
Stack Temp: 543° k with ESP
443° k with Fabric Filters
Stack Exit Velocity: 12 m/s
Building Configuration:
H: 27.28 m
W: 51.88 m
L: 59.5 m
hypothetical facility. The site has in excess of 48 inches of rainfall per
year, has a humid climate, and the soil characteristic is classified as sandy.
Additional data on soil characteristics, local meteorology, and population
distribution around the model plant will be available through the computer
files incorporate in the various environmental exposure models.
The existing population of massburn MWCs is being represented for the
purposes of the analysis by the Hampton facility. The Hampton facility has
been stack tested for pollutant emissions more frequently than other MWCs
operating in the United States, and therefore offers a comprehensive data base
of potential emissions. The Hampton facility consists of two identical
waterwall furnaces each fired at a charging rate of approximately 90,000
October 1986 3-5 DRAFT-DO NOT QUOTE OR CITE
-------
kilograms per day. Operational since September 1980, the facility achieves
approximately 85 percent volume reduction of the incoming refuse, thermally
converts the BTU value of the refuse at an efficiency of 65 percent, and
produces about 15,000 Kg steam per hour per unit. Last year the Hampton
incinerator supplied enough steam to the NASA-Langley Research Center to
provide about 82 percent of the heating and cooling needs.
The incinerator is charged with raw (unprocessed) refuse from an enclosed
storage pit by overhead crane. The refuse burns without the use of auxiliary
fuel as it travels down a series of three inclined reciprocating grates. The
residence time of the solids in the furnace is about 45 to 60 minutes. During
stable conditions the furnace temperature is approximately 800° C. The
combustion gas is passed through an electrostatic precipitator installed for
each unit before discharge into the atmosphere by two 27 meter smokestacks.
Measurements of particulate emissions have shown that each unit is capable of
emissions within the current requirements for MWCs of 180 mg/dscm corrected to
12 percent C0«.
3.1 THE INDUSTRIAL SOURCE COMPLEX MODEL
It has been customary for EPA only to make estimates of the ambient air
concentration attributable to stack emissions of airborne pollutants from a
point source. Recent analysis of principal sources of lead in soil of a
community in Montana disclosed the fact that the accumulation of lead was a
direct result of wet and dry deposition of lead adsorbed onto particulates and
emitted from the stack of a zinc smelter (MEA, Inc., 1982). Analysis of the
sources and losses of polychlorinated biphenyls (PCBs) in the sediments of
lakes remote from urbanization indicated that roughly 50 percent of the total
PCB input to the lake was due to atmospheric transport and deposition
(Swackhamer, 1986). Czuczwa and Hites (1984) reported on the analysis of
sediment core samples in the Great Lakes for the presence of polychlorinated
dioxins and dibenzofurans, and concluded that the atmospheric transport and
deposition of emissions from stationary combustion sources most likely was the
main source of PCDDs and PCDFs in the sediments. These studies provide
evidence that calculating only ambient air concentrations of specific pollut-
ants in the vicinity of a combustion point source may significantly under-
predict spatial and temporal exposures to pollutant emissions by not accounting
October 1986 3-6 DRAFT—DO NOT QUOTE OR CITE
-------
for the potential surface accumulation of pollutants throughout the operating
life of the facility. The EPA assessments of potential risk from air emissions
have primarily been concerned with carcinogenic and noncarcinogenic health
risks resulting from direct inhalation of ambient air concentrations of
pollutants. In order to extend the risk evaluation to a consideration of other
routes of population exposure to environmental pollutants, EPA will predict the
rate of deposition, overtime, of pollutants believed to be adsorbed onto
particulate matter in the smokestack exhaust gas, and will attempt to calculate
the spatial and temporal accumulation of these pollutants on the soil, and in
surface waters. These estimates of deposition will be factored into the
Surface Runoff Model, the Terrestrial Food Chain Model, and the Groundwater
Contaminant Model, to predict the migration of pollutants from the surface into
water bodies, the bioaccumulation of the pollutants into biota and up the
trophic system including the human food web, and indirect exposure to humans of
the deposited pollutants.
The Industrial Source Complex Model (ISC) will be used to predict the
dispersion of smokestack emissions from the model plant and Hampton facility
through the atmosphere as well as predict both wet and dry deposition of
pollutants onto the surface. The ISC model has been previously reviewed by the
Science Advisory Board in a separate document on the development of a risk
assessment methodology for sewage sludge incinerators (U.S. EPA, 1986d).
Therefore, this report will summarize the function of the ISC model in calcula-
ting ground level concentrations of emitted pollutants.
The Industrial Source Complex (ISC) Dispersion Model combines and enhances
various dispersion model algorithms into a set of two computer programs that
can be used to assess the air quality impact of emissions from the wide variety
of sources associated with an industrial source complex (U.S. EPA, 1986e). For
plumes comprised of particulates with appreciable gravitational settling
velocities, the ISC Model currently accounts for the effects on ambient partic-
ulate concentrations of gravitational settling and dry deposition. The ISC
short-term model (ISCST), an extended version of the Single Source (CRSTER)
Model (U.S. EPA, 1977) will be used in this analysis, and is designed to
calculate concentration or deposition values for time periods of 1, 2, 3, 4, 6,
8, 12 and 24 hours. If used with a year of sequential hourly meteorological
data, ISCST can also calculate annual concentration or deposition values. An
alternative model, the ISC long-term model (ISCLT), is a sector-averaged model
October 1986 3-7 DRAFT—DO NOT QUOTE OR CITE
-------
that extends and combines basic features of the Air Quality Display Model
(AQDM) and the Climatological Dispersion Model (COM). The long-term model uses
statistical wind summaries to calculate seasonal (quarterly) and/or annual
ground-level concentration or deposition values. Both ISCST and ISCLT use
either a polar or a Cartesian receptor grid.
The ISC Model programs accept the following source types: stack, area and
volume. The volume source option is also used to simulate line sources. The
steady-state Gaussian plume equation for a continuous source is used to calcu-
late ground- level concentrations for stack and volume sources. The area source
equation in the ISC Model programs is based on the equation for a continuous
and finite crosswind line source. The generalized Briggs (1971 and 1975)
plume- rise equations, including the momentum terms, are used to calculate plume
rise as a function of downwind distance. Procedures suggested by Huber and
Snyder (1976) and Huber (1977) are used to evaluate the effects of the aerody-
namic wakes and eddies formed by buildings and other structures on plume
dispersion. A wind-profile exponent law is used to adjust the observed mean
wind speed from the measurement height to the emission height for the plume
rise and concentration calculations. Procedures utilized by the Single Source
(CRSTER) Model are used to account for variations in terrain height over the
receptor grid. The Pasquill-Gifford curves (Turner, 1970) are used to calcu-
late lateral and vertical plume spread. The ISC Model has rural and urban
options.
For purposes of exposure analysis from MWC emissions the ISC Short-Term
(ISCST) model program will be utilized. The program makes mathematical calcu-
lations of dispersion and dry deposition and produces a printout of these
values. A more detailed description of ISCST appears in the user's guide
(U.S. EPA, 1986e). Because the current ISCST model has no provision for
calculating wet deposition of the emissions, EPA has had to develop a program
to estimate the effect of precipitation events on the rate of surface deposi-
tion. In general, wet deposition was modeled by incorporating Equation 3-1
into the ISC-ST.
A(n,j) OnKQ f / \1 r , / \*-j
WDep(n) = f - iexp - A(n,j) M— exp - ± PU Equation (3-1)
V2n o u(h) L \u(h)/J L \ V J
October 1986 3-8 DRAFT—DO NOT QUOTE OR CITE
-------
where:
f - fraction of time precipitation occurs (fraction of hour)
A(n,j) = the fraction of material removed per unit time in the nth
particle size category and jth precipitation intensity
category;
-------
TABLE 3-3. SUMMARY OF SCAVENGING COEFFICIENTS
EXPRESSED PER SECOND OF TIME, USED IN COMPUTING WET SURFACE DEPOSITION
Precipitation
Intensity
Heavy
Moderate
Light
<2
1.46 E-3
5.60 E-3
2.20 E-4
Particle Size
>2 and <10
4.64 E-3
8.93 E-4
1.80 E-4
Category (microns)
>10
9.69 E-3
9.69 E-3
9.69 E-3
Light = less than 0.10 inches per hour.
Moderate = 0.11 to 0.30 inches per hour.
Heavy = equal to or greater than 0.31 inches per hour.
has used the same washout coefficients for both rainfall and snowfall events.
Other frozen forms of precipitation, e.g., snow pellets, ice pellets, and hail,
present a relatively smooth surface to the aerosol particles and are not likely
to be effective scavengers of the pollutants. Therefore, periods and occur-
rences of these types of frozen precipitation are modeled in the dry deposition
mode.
The principal assumptions made in computing wet deposition are: (1) the
intensity of precipitation is constant over the entire path between the source
and receptor; (2) the precipitation originates at a level above the top of the
emission plume so that the hydrometeors pass vertically through the entire
plume, (3) the time duration of the precipitation over the entire path between
the source and receptor point is such that exactly f, a fraction between zero
and one, of the hourly emission Q in Equation (3-1) is subject to a constant
intensity for the entire travel time required to traverse the distance between
the source and receptor, and the remaining fraction (1-f) is subject only to
dry deposition processes.
Certain inferences must be made concerning particle size, particle distri-
bution, available surface area for adsorption, gravitational settling veloci-
ties and the potential for surface reflection of particles of a specific mass
before the ISC-ST program can predict deposition of a chemical. Unfortunately,
research has not provided adequate data on particle size f rationalization of
particulate matter entrained in MWC emissions. A series of stack tests con-
ducted by EPA at a refuse-to-energy MWC in Braintree, Massachusetts will
provide a conservative assumption regarding the distribution of particulate
October 1986 3-10 DRAFT—DO NOT, QUOTE OR CITE
-------
matter by particle size (U.S. EPA, 1980a). The MWC was tested at relatively
high participate emissions (6.7 kg/hour) in exceedance of EPA requirements, and
the participate matter control efficiency of the electrostatic precipitator was
calculated to be only 74 percent. Total particulate concentrations averaged
215 mg/dscm at 12% COp. Selection of the particle size distribution may or may
not represent the distribution of emissions from equipment marketed today, such
as multiple-field electrostatic precipitators, or from-fabric filters, but EPA
has made this selection on the basis of providing a reasonably conservative,
tending toward the worst-case, estimate of the deposition of particles in MWC
emissions. The performance of the pollution control equipment is dependent not
only on design, but on the degree of preventative and required maintenance and
cleaning and overhaul that is done to the apparatus throughout the working life
of the MWC. EPA has observed that emissions of particulate matter at energy
recovery MWCs equipped with ESPs have, on occasion, exceeded current Federal
standards, although most incinerators are able to comply (U.S. EPA, 1986b). It
is hoped that selection of a conservative distribution of particulate emissions
may account for potential errosion of pollution control efficiency of the
equipment over the facility life.
Table 3-4 displays the average particle-size distribution measured by EPA
at the Braintree MWC (U.S. EPA, 1980s). Recently tests have been conducted at
the outlet of fabric filters operating on a massburn heat recovery MWC in
Germany (Hahn, 1986). Table 3-5 summarizes the distribution of particulate
matter at the Wurzburg, West Germany massburn waste-to-energy MWC. In compari-
son to the Braintree MWC data, the emission of particulate matter outlet of the
fabric filters averaged 9.0 mg per dry standard cubic meter (at 12 percent
C02) over 3 test runs. Approximately one-half of the particulate emissions at
Wurzburg were less than 0.5 microns in diameter, although over one-third the
emissions were of a diameter greater than 12 microns.
For purposes of estimating deposition, estimations of the relationship
between the pollutant concentration and particle size were assumed. The
inorganic pollutants have been evaluated for their affinity for adsorption onto
surfaces of various particle diameters (U.S. EPA, 1985a). These observations
have been made from emissions of particulate matter from a limited number of
MWCs. In general the maximum mass of elemental emissions will be found associ-
ated with fine particles, e.g., two microns or less aerodynamic diameter.
Table 3-6 is a summary of the particle phase distribution of trace elements
adsorbed on particulate matter.
October 1986 3-11 DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 3-4. PARTICLE-SIZE DISTRIBUTION DETERMINED
IN PARTICIPATE HATTER EMISSIONS AT THE BRAINTREE MWC
Geometric Mean T *,
Particle Diameter Total Particle
(Microns) Mass
XL5.0
Notes:
1.
2.
3.
12.5
8.1
5.5
3.6
2.0
1.1
0.7
<0.7
Data taken from U.S. EPA, 1980a.
Data is an average of 6 stack test runs.
The facility is equipped with an electrostatic precipitator
12.8
10.5
10.4
7.3
10.3
10.5
8.2
7.6
22.4
rated at
an average particulate removal efficiency of 74% at time of test.
4. The facility recovers steam from the combustion of about 120 metric
tons of MSW per day.
TABLE 3-5. TYPICAL PARTICLE SIZE DISTRIBUTION DETERMINED
IN PARTICULATE EMISSIONS AT THE WURZBURG MASSBURN MWC
Particle Diameter Percent of
(Microns) Particle Mass
>12.0
7.5 - 12
5.1 - 7.5
3.5 - 5.1
2.3 - 3.5
1.1 - 2.3
0.7 - 1.1
37.8
0.0
1.0
1.5
3.6
2.0
0.5
0.47 - 0.7 l.o
<0.47 52.6
Notes:
1. Data taken from Hahn, 1986, and Hagenmaier, 1986 (unpublished).
2. Data represents one test day, but is typical of performance.
3. The facility is equipped with a dry venturi; scrubber coupled with
fabric filters for air pollution control.
4. The facility has a capacity of 600 metric tons MSW per day.
October 1986 3-12 DRAFT—DO NOT QUOTE OR CITE
-------
TABLE 3-6. RATIO OF METAL EMISSIONS AS A FUNCTION
OF PARTICLE DIAMETER (>2u/<2u)
Pollutant
Arsenic
Beryllium
Cadmium
Chromium
Lead
Nickel
Ratio
25/75
34/66
34/66
41/59
12/88
41/59
Source: U.S. Environmental Protection Agency (1985a).
The data on the distribution of organic compounds adsorbed onto variously
sized particulates is lacking. The deficiency of data on the relationships
between particle size and organic compound concentration may be overcome by
assuming that compounds will be distributed in proportion to the percent of
total particle surface area available for adsorption for each particle size
category (assuming particles are perfect spheres). The mass emission rate of
the organic pollutant may be distributed by particle size by computing propor-
tion and available surface area for a given particle size (Hart and Associates,
1985), and holding particle density constant. The particle weight is propor-
tional to volume if density is constant. Therefore, the ratio of the surface
area to volume is proportional to the ratio of surface area to weight for a
particle with a given radius (Hart and Associates, 1985). Multiplying this
proportion times the weight fraction of particles of a specific diameter
(microns) yields a number that estimates the amount of surface area available
for chemical adsorption. Example calculations are given in Appendix B.
The particle size and surface area distribution will be kept constant in
the analysis of inorganic and organic compound emissions resulting from the
application of two distinct sets of air pollution control systems: electro-
static precipitators (ESPs) and the combination dry venturi scrubber and fabric
filters. Comparisons between the particle size distributions depicted in
Tables 3-3 and 3-4 does show distinct differences between emissions resulting
from ESP control or fabric filter control. However, it was felt that not
enough particle mass fractionalization data exists to generalize a distribution
pattern for each control devise based on these limited measurements. There-
fore, EPA has assumed, for purposes of estimating deposition of specific
pollutants, a unitized particle size distribution of both pollution control
systems.
October 1986 3-13 DRAFT—DO NOT QUOTE OR CITE
-------
Three particle size categories were selected: greater than 10 microns,
2 to 10 microns, and less than 2 microns. The fraction of total surface area
available for chemical adsorption is calculated to be 0.03, 0.095, and 0.875,
respectively. The specific emission rate for each po.llutant considered in the
analysis is multiplied by the fraction of available surface area to estimate
the pollutant emission rate corresponding to the particle size distribution.
For example, the emission of formaldehyde may be estimated to be 80 milligrams
per second of plant operation. Then the formaldehyde emission rate in consid-
eration of available surface area would be: (0.030 x 80 mg/s) = 2.4 mg/s for
particles greater than 10 microns; (0.095 x 80 mg/s) = 7.6 mg/s for particles
sizes ranging from 2 to 10 microns; (0.875 x 80 mg/s) = 70 mg/s for particles
less than or equal to 2 microns in diameter. The organic pollutants emissions
will be determined on this basis for purposes of wet and dry surface deposition
modeling.
Table 3-6 summarizes the typical data on particle size distribution of
heavy metal emissions from MWCs. Two particle size categories are specified in
the data: less than 2 microns diameter, and greater than 2 microns diameter
(U.S. EPA, 1985a). In order to calculate emission rates of inorganic compounds
ratios of mass of the compound per particle size category were assumed. For
example Table 3-6 indicates that approximately 25 percent of arsenic has been
found adsorbed to particles greater than 2 microns diameter, and thus the ratio
is 25 percent/75 percent for >2/<2 microns. If arsenic were emitted at a rate
of 2.3 x 10 grams per second, then the emission rate per particle size
cutoff, taking into consideration available surface area for adsorption, would
be 1.4 x 10" g/s for particles greater than 10 microns; 4.4 x 10"4 g/s for
particles 2 to 10 microns, and 1.7 x 10 g/s for particles less than 2
microns.
If the assumptions regarding particle size distribution, and fraction of
total surface area for adsorption are held constant between particulate
emissions resulting from electrostatic precipitators or fabric filters, how
will EPA account for differences in emissions of organic and inorganic pollu-
tants with the application of different exhaust gas cleaning systems? Differ-
ences in emissions will be accounted for with differences in rates of emissions
of specific pollutants rather than differences in particle distribution. For
example electrostatic precipators have an ability to reduce inorganic compounds
by 97 percent, except for mercury which is only controlled about 30 percent.
October 1986 3-14 DRAFT—DO NOT QUOTE OR CITE
-------
By comparison, the combination dry scrubber and fabric fitters can control
inorganic compound emissions by 99 percent, except for mercury which can be
controlled by about 50 percent (Klicius et al., 1986).
Uncontrolled emissions of organic compounds could not be found in the
literature. The EPA will assume that organic emissions can be controlled or
reduced only by about 20 percent inlet to outlet of ESPs, and dry scrubbers
coupled with 'fabric filters will be capable of at least 95 percent control of
organic compound emissions. Recent analysis on the physical distribution of
PCDD and PCDF emissions from MWCs suggests that without appreciable cooling of
the combustion gas, about 75 percent of the congeners will be either in a
gaseous state or associated with aerosol particles less than 0.1 microns
aerodynamic diameter (Ministry of the Environment, Ontario, Canada, 1985;
Rappe and Ballschmidter, 1986; Ballschmidter, et al. 1984). The EPA is
currently testing across pollution control devices to gather better data on ESP
and dry scrubber - fabric filter control efficiencies for PCDDs/PCDFs on
municipal incinerators. When this data is available, EPA will be able to
refine estimates of control efficiency. Environment Canada has reported the
efficiencies of a pilot scale dry scrubber - baghouse system operating on a 227
metric ton massburn MWC (Klicius et al., 1986). At temperatures varying from
110°C to 200°C the overall removal efficiencies for PCDDs and PCDFs were
greater than 99 percent. Other organics such as chlorobenzenes, polychlorinated
biphenyls and chlorophenols were removed to a lesser extent, but generally
better than 95 percent at operating temperatures between 125°C and 140°C. The
exception were polycylic aromatic hydrocarbons which were removed at an effi-
ciency of above 84 percent at 140°C, and 98 percent at 200°C. Inherent in EPA's
current assumption of poor removal of orgnaic compounds with ESPs, and with no
prior gas cooling, is the assumption that at temperatures above 250°C charac-
teristic of combustion gas temperatures passing through ESPs, the organic
compounds will predominate in the gaseous state, and therefore will not be
efficiently collected by the system.
3.2 THE HUMAN EXPOSURE MODEL (HEM)
The Industrial Source Complex Model (ISC-ST) will be used to predict the
ambient air concentrations of the chemicals to be evaluated. The output will be
a concentration array for 160 receptors along each of 16 wind directions),
October 1986 3-15 DRAFT-DO NOT QUOTE OR CITE
-------
specified in concentric radial distances from each facility of 0.2, 0.5, 1, 2,
5, 10, 20, 30, 40 and 50 kilometers computed every 22.5° on a radius-polar grid
pattern. This output will be a suitable format for utilization with the Human
Exposure Model (HEM). HEM will be used to estimate the carcinogenic risk to
the population exposed by inhalation to the predicted ambient air concentra-
tions of the specified pollutants. The HEM also is capable of air dispersion
modeling and is often used in nationwide analyses of source categories.
Reference is made to the User's Manual for the HEM for a detailed description
of dispersion modeling capabilities (U.S. EPA, 1986f).
Exposure is the product of the population and the concentration to which
the population is exposed. To form this product both the concentration and the
population must be known at the same location or point. The HEM uses the
latitude and longitude of the facility to determine the population of the study
area. The permanent data base is comprised of the 1980 Census Data Base broken
down by block group/enumeration district (BG/ED). The population data base
contains the population centroid coordinates (latitude and longitude) and the
1980 population of each BG/ED in the country (about 300,000 centroids in 50
states plus the District of Columbia). A population centroid is the
population-weighted geographical center of a BG/ED for known geodetic coordi-
nates. For BG/ED centroids located between 0.2 Km and 3.5 Km from the facili-
ty, populations are apportioned among neighboring polar grid points. A polar
grid point is one of the 160 receptors at which concentrations are estimated by
the dispersion model. There are 64 (4 x 16) polar grid points within this
range. Both concentration and population counts are thus available for each
polar grid point. Log-log linear interpolation is used to estimate the concen-
tration of each ED/BG population centroid located between 3.5 Km and 50 Km from
the source. Concentration estimates for 96 (6 x 16) grid points (receptors at
5, 10, 20, 30, 40, and 50 Km from the source along each of the 16 wind
directions) resulting from dispersion modeling are used as reference points in
the interpolation. The User's Guide gives examples of these calculations
(U.S. EPA, 1986f, pp 2-12 to 2-19).
The HEM employs a number of simplifying assumptions when computing expo-
sures to a pollutant, including:
October 1986 3-16 DRAFT—DO NOT QUOTE OR CITE
-------
1. It is assumed that most exposure occurs at population weighted
centers (centroids) of block group and enumeration districts (BG/ED) because
the locations of actual residences are not contained in available databases.
The model relies on information provided in a database developed by the U.S.
Census Bureau.
2. It is assumed that people reside at these centroids for their entire
lifetimes (assumed to be 70 years for calculating cancer risk).
3. It is assumed that indoor concentrations are the same as outdoor
concentrations.
4. It is assumed that plants emit pollutants at the same emission rate
for 70 years. Long-term emission rates are not known.
5. It is assumed that the only source of exposure is the ambient air and
resuspension of pollutants via dust is not considered.
6. It is assumed that there is no population migration or growth.
7. The model does not provide for descriminating exposure situations
that may differ with age, sex, health status, or other situations. Susceptible
population subgroups are not considered.
The exposure estimates are combined with measures of carcinogenic potency
to estimate the probability of cancer by direct inhalation. The estimation of
carcinogenic potency is expressed as the "unit risk." The unit risk estimate
for an air pollutant is defined as the lifetime cancer risk occurring in a
population in which all individuals are exposed continuously from birth
o
throughout their lifetimes to a concentration of 1 ug/m of the agent in the
air they breathe. The measures of carcinogenic potency (unit risk estimates)
were derived by EPA's Carcinogen Assessment Group, and may be reviewed in a
Health Assessment Document prepared for each carcinogen. Generally, Health
Assessment Documents have been thoroughly reviewed by the Science Advisory
Board before final publication. The derivation of unit risk estimates depends
on a quantitative evaluation of adverse health outcomes statistically tied to
exposure to a chemical as observed in epidemiological, clinical, toxicological,
and environmental research. Evidence for carcinogenic action is based on the
observations of statistically significant tumor responses in specific organs or
tissues.
October 1986 3-17 DRAFT-DO NOT QUOTE OR CITE
-------
Evidence of possible carcinogenic!ty to humans is evaluated according to
specific criteria set forth in EPA's guidelines for Carcinogenic Risk Assess-
ment (U.S. EPA, 1986g; Federal Register, 1986). Since risks at low exposure
levels cannot be measured directly either by animal experiments or by epidemio-
logical studies, mathematical models must be used to extrapolate from high dose
to low doses characteristic of environmental exposures. Table 3-7 displays
representative unit cancer risks for inhalation exposure to specific chemicals.
There are differences in the amount, type and quality of the data used to
derive the unit risk estimates. Such differences may affect the confidence
that can be assumed in the quantitative estimate of carcinogenic potency and
subsequent quantification of cancer risks. The uncertainties associated with
cancer risk assessment vary with the chemical. Generally, human data pro-
vide the best evidence that a chemical is a carcinogen. In the absence of
human data, animal to man extrapolation must be performed. The unit risk
estimate based on animal bioassays is considered an upper bound estimate of
the excess cancer risk over a lifetime in populations exposed to the probable
carcinogen. The concept of equivalent doses for humans compared to animals,
for example, has little experimental verification regarding carcinogenic
response and this is another area of uncertainty. In addition, human popula-
tions are more variable than laboratory animals with respect to genetic consti-
tution, diet, living environment and activity patterns. The overall
uncertainties associated with unit risk estimates have not yet been statisti-
cally quantified. At best the linear extrapolation model used to derive unit
risk estimates provides an approximate but plausible estimate of the upper
limit of risk: It is not likely that the true risk would be much more than the
estimated risk, but the true risk could very well be considerably lower. The
quantitative aspects of risk assessments should not be construed as an accurate
representation of the true cancer risk, but are best utilized in the regulatory-
decision process, such as setting regulatory priorities based on a relative
indication of the potential magnitude of population risk from chemical expo-
sure.
By combining the estimates of public exposure with the unit risk estimate,
two types of quantitative estimates are produced. The first, called maximum
lifetime risk, relates the risk to the individual or individuals estimated to
live in the area of highest concentration as estimated by the dispersion model.
These are the "most-exposed individuals" (MEIs). The second type of risk
October 1986 3-18 DRAFT—DO NOT QUOTE OR CITE
-------
estimate, called aggregate risk, is a summation of all the risks to people
living within 50 kilometers of the facility. The aggregate risk is expressed
as incidences of cancer among all of the exposed population after 70 years of
exposure; for statistical convenience, it is often divided by 70 and expressed
as cancer incidences per year.
Two chemicals have been selected to provide example calculations of risk:
benzo(a)pyrene (B(a)P) and cadmium (Cd). Table 3-7 shows the unit risk esti-
mates for B(a)P and cadmium to be 1.7 x 10~3 (pg/m ) and 1.8 x 10* (ug/m ) ,
respectively. The ISC-ST dispersion model predicted the maximum annual average
ground level concentration within 50 kilometers of the model plant massburn
incinerator located in Western Florida. The maximum concentration of B(a)P was
«A ^J
determined to be 1.62 x 10 ug/m if only electrostatic precipitators are used
on the facility capable of only 20 percent control efficiency of B(a)P. If an
individual or individuals are exposed continuously for 70 years to the maximum
average annual ground-level concentration of B(a)P, then he or she would have a
maximum individual lifetime risk of about 3 chances in 10 million of contracting
cancer (3 x 10"7). The aggregate risk to the total population living within 50
kilometers is estimated by the HEM to be 0.0004 cancer incidences per year
(0.03 cancer cases in 70 years of B(a)P exposure), or roughly one cancer in the
exposed population of 1,650,000 people per 2500 years of operation of the
facility. Likewise the emission of cadmium from the stack of the model plant
' -3
results in a predicted maximum ground-level concentration of 1.81 x 10
ug/m3, assuming ESPs are used to control inorganic emissions by 95 percent.
The maximum individual lifetime risk from continuous (70 year) inhalation
exposure to the predicted concentration is estimated to be about 3 chances in 1
million (3.0 x 10~6) of contracting cancer. The aggregate risk to the popula-
tion of 1.65 million living within a radius of 50 km of the facility is
estimated by HEM to be 0.004 (0.28 cancers in 70 years), or about 1 cancer
every 250 years of facility operation associated with the stack emission of
cadmium.
If more advanced controls, e.g., dry scrubber coupled with fabric filters,
are used on the model plant instead of ESPs, then the risk calculations are
changed, reflecting 95 percent B(a)P control efficiency, and 99 percent cadmium
control efficiency. The ISCST air dispersion model predicts a maximum annual
average B(a)P concentration of 1.15 x 10" ug/m . The estimated maximum
individual lifetime risk of cancer now becomes a probability of about 2 chances
October 1986 3-19 DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 3-7. UNIT CANCER RISK ESTIMATES FOR INHALATION EXPOSURE
TO SPECIFIC CHEMICALS (RISK PER ug POLLUTANT/m3 OF AIR)
Arsenic 4.2 x 10l3
Benzene 8.3 x 10.J
Benzo(a)pyrene : 1.7 x 10_3
Beryllium 2.4 x 10.3
Cadmium 1.8 x 10_3
Carbon tetrachloride 1.5 x 10.5
Hexachlorobenzene 4.8 x 10_4
Chromium (VI) 1.2 x 10.2
*2378-TCDD 3.3 x 10.s (pg/m3) *
*HexaCDD 1.3 x 10 6 (pg/m3) *
Nickel 3.0 x 10l4
Formaldehyde 1.8 x 10 4
PCB 1.2 x 10"3
*The carcinogenic potency of TCDD and HexaCDD is such that the unit risk
estimates are based on inhalation exposure to 1 picogram (10-12g) per cubic
meter of air.
~8
in 100 million (2 x 10 ) from continuous exposure to B(a)P emissions. The
aggregate population risk from the stack emission of B(a)P is further reduced
from the ESP control scenario to an estimate of 0.00003 cancer incidences per
year (0.02 cancers in 70 years), or roughly 1 cancer in the exposed population
of 1.65 million every 33,000 years of facility operation. Additional emission
control of cadmium at the model plant would result in a predicted maximum
annual ground-level concentration of 6.6 x 10 ug Cd/m^ of air. This corre-
sponds to an estimated maximum individual lifetime risk of continuous inhala-
tion exposure of cadmium for 70 years to be about 1 chance in 1 million (1 x
10 ) of contracting cancer. The aggregate risk from population exposure (1.65
million people) residing within 50 km of the model plant is estimated by the
Human Exposure Model (HEM) to be 0.002 cancer cases per year (0.14 cancers in
70 years) directly attributable to emissipns of cadmium.
Table 3-8 illustrates the B(a)P output concentrations predicted by the ISC
air dispersion model when the model plant is controlled with ESPs. Table 3-9
illustrates the ground-level annual average concentrations of B(a)P when dry
scrubbers coupled with fabric filters are used at the facility. Table 3-10
depicts the distribution of cancer risk from B(a)P emissions resulting from ESP
control, and Table 3-11 illustrates the distribution of risk within 50 km of
the facility with dry-scrubber-fabric filter control. These tables serve to
illustrate the typical outputs of the ISC and HEM models.
October 1986 3-20
DRAFT-DO NOT QUOTE OR CITE
-------
fl • AMBIENT AIR CONCENTRATIONS OF
B(a)P AS PREDICTED BY THE ISC
* !*?!*._!!' MODEL (WITH ESP CONTROL)
m —
"iri 1 T 4 -i C r
,?o« ."lUfl I.AOA TTWira y.'WTin Tir.'AOir ?B.HATI tBTBmiflA.imr 11170 mi •
2.S477.0A9 l.'fcUA-BAS—TT
i.ai»-Afl«s ?;
r;inviiiifo»—<>;iui
f% Wli I.^niiT-OO* ".Pl'Jj-oo'i 5.795 i-no^ S.l«?5-<)h* T.
Tl USHflUAAB—T7
~wn—r.'«nyo-Boa—*i.'»7Ji-on5—a.s&vs-ffB*—»r
fK~ B~. 7rta»-ooS ^.S^bv-oaS j.vi^»-nn^ J^
TWf—t;q3ifc-ons—a.'usfc-Bns—3.'yqon-nnn—3.ni?3-oos
fc.nr*-qos
Tlf
IF
LUUIIUN
-g7 mut si»—P" n. L«I.—rrn
9t DEC. *»' 0" *. LQMC. t «2.
3IIUMHS
-------
, TABLE
AMBIENT AIR CONCENTRATIONS OF
B(a)P AS PREDICTED BY THE ISC
•T— — MODEL (WITH DRY SCRUBBER -
NC^..N>tTOH8 FABRIC FILTERS)
TTTU h 1 * T 4 N t »
t.OllOZ.nrtfl3. tut A
?i».ni»ii 4fl.0—007I .**! 11-007
*fii6.*7H«-i»064.*n9JI-iMi6 J.nA4|.Aiifc—77
Nw 5.5IT6-006 j;«4l-ii«6 5.'6^JI-od6—2.37«fl.fto6—l.iol^.oilfc—l.ftgAB.flflT—-fl06 £.*74i-oo6—?.'6« 5?-o«6—a;si5J-floA—2.Ho^-Aftl.—t.^Sfll-ngS—S.hlS*.A.|7—i.'23Hl-nB7
8t S.92ST
15f—5>1Bfc-o<»6—S.4«4*-4o6
SUUflPE LOCAriON
*7 HtC, 57' o" N. L*T. I ?i
6? PER, 27' 0" .«. LOKH. C H?,
-------
TABLE 3-10. HEM OUTPUT AS TO THE DISTRIBUTION OF RISK BY POPULATION
RESULTING FROM B(a)P EMISSIONS AT THE MODEL PLANT (WITH ESP CONTROL)
Estimated Risk Level
3.0 x 10"7
2.6 x 10~7
1.3 x 10"7
5.0 x 10"8
2.6 x 10"8
1.3 x 10" 8
5.0 x 10"9
2.2 x 10"9
Hypothetical Population
<1
<1
37
52,100
172,000
897,000
1,470,000
1,650,000
Note: The calculations of risk are meant to serve as an example output
to the Human Exposure Model, and do not apply to an existing MWC.
TABLE 3-11. HEM OUTPUT AS TO THE DISTRIBUTION OF RISK BY
POPULATION RESULTING FROM B(a)P EMISSIONS AT THE MODEL PLANT
(WITH DRY SCRUBBER - FABRIC FILTER CONTROL)
Estimated Risk Level
2.0 x 10"8
1.3 x 10"8
5.0 x 10"9
2.6 x 10"9
1.3 x 10"9
5.0 x 10"10
1.4 x 10"10
Hypothetical Population
<1
37
13,000
95,400
351,000
1,280,000
1,650,000
Note: The calculations of risk are meant to serve as an example output
to the Human Exposure Model, and do not apply to an existing MWC. '
There are uncertainties in the risk assessment calculations that must be
highlighted, but cannot be easily quantified. The basic assumptions implicit
in the methodology for inhalation cancer risk assessment are 1) that all
exposures occur at people's residences; 2) that people stay at the same loca-
tion continuously for 70 years; 3) that the ambient air concentrations and
emissions from MWCs which cause these concentrations persist for 70 years;
and 4) that the concentrations are the same inside and outside the residence.
In addition there is no numerical recognition of the risks to more susceptible
subgroups within the exposed population. Population growth or migration to or
from the modeling area is also not considered in the analysis. In the absence
of pollutant specific information on the treatment of concomitant exposure to
October 1986 3-23 DRAFT—DO NOT QUOTE OR CITE
-------
mixtures of carcinogens, EPA will assume that the risks associated with.differ-
ent pollutants from MWC emissions are additive. This conforms to EPA's recent
guidelines on cancer risk assessment (51 FR 33992, September 24, 1986).
Simultaneous exposure to several chemical carcinogens is a frequent occurrence
in the environment. The EPA is committed to toxicologic research on the health
risks posed to exposure of complex mixtures. The ability to predict how
specific mixtures of toxicants may interact ultimately must be based on an
understanding of the biological mechanisms involved in such interactions.
An example of the dilemma EPA often faces in predicting human risk to
complex mixtures is the chemical class polycyclic aromatic hydrocarbons (PAH).
This class of compounds are formed uniquely during the incomplete combustion or
pyrolysis of organic materials containing carbon and hydrogen. Potentially
several hundred distinct compounds may be formed, but only a few have been
speciated in MWC emissions (U.S. EPA, 1986b). These emissions are summarized
in Appendix A. Of all the PAH compounds investigated for carcinogenic proper-
ties, benzo(a)pyrene remains the most intensely studied compound (U.S., EPA,
1984b), and even it's metabolic pathway for carcinogenesis has been described.
The EPA has reviewed these data and has derived an inhalation unit risk esti-
mate for B(a)P that is appropriate to predict human cancer risk. The Inter-
national Agency for Research on Cancer (IARC) has provided additional scienti-
fic evaluation on the carcinogenicity of other PAH compounds that may be found
associated with benzo(a)pyrene when emitted from combustion sources (IARC,
1983). The IARC has carefully reviewed the scientific data, and has determined
the following specific PAH compounds to be probably carcinogenic to humans:
benz(a) anthracene, benzo(b)fluoranthene, benzo(j) fluoranthene, benzo(k)
fluoranthene, benzo(a)pyrene, dibenz(a,h)acridine, dibenz(a,j)acridine,
dibenz(a,h)anthracene, 7H-dibenzo(c,g)carbazole, dibenzo(a,e)pyrene, dibenzo
(a,h)pyrene, dibenzo(a,i)pyrene, dibenzo(a,l)pyrene, and indeno(l,2,3-cd)pyrene.
Additional PAH compounds have been found to be genotoxic, but have not been
adequately tested for carcinogenicity (IARC, 1983).
Typically B(a)P may be less than 10 percent of total PAH emissions from
MWCs (U.S. EPA, 1986b). If other PAH compounds that are also carcinogenic
have been identified in the emissions, then estimating the carcinogenic risk
from exposure only to B(a)P may potentially underestimate the risk posed by the
mixture of PAH compounds. The dilemma is presented by the fact that EPA has
only derived a carcinogenic potency estimate for B(a)P, and not the other PAH
October 1986 3-24 DRAFT—DO NOT QUOTE OR CITE
-------
compounds I ARC has determined to also be probable human carcinogens. As a
class of compounds, polycyclic aromatic hydrocarbons are a significant constit-
uent of total organic compound emissions from municipal waste combustors
(MWCs).
Given these set of circumstances, EPA is faced with certain choices with
respect to cancer risk assessment of the PAH mixture characteristic of combus-
tion emissions. EPA may:
1. Restrict PAH risk assessment only to the emission of benzo(a)pyrene,
the compound for which EPA has derived a cancer unit risk estimate.
2. Consider the PAH compounds identified by IARC as probably carcinogenic
to humans as potentially equal to the carcinogenic potency of B(a)P.
3. Consider total PAH compounds as potentially equal to the carcinogenic
potency of B(a)P.
Currently EPA favors scenario 2, that is, enfold the observed carcinoge-
nicity of other PAH compounds into the quantitative cancer potency
estimate for B(a)P. This scheme would only be applied to PAH mixtures as an
interim procedure until EPA could quantify the potency of the other carcino-
genic compounds or directly submit the mixture to an appropriate bioassay. The
EPA was faced with a similar dilemma in regarding the potential risk associated
with population exposure to a mix of structurally related polychlorinated .
dibenzo-p-dioxins and dibenzofurans. The EPA had derived cancer potency
estimates only for 2,3,7,8-tetrachlorodibenzo-p-dioxin and an isomer mixture of
hexachlorinated dioxins (U.S. EPA, 1985b), and yet EPA devised an interim
procedure for estimating risks associated with the mixture of PCDD and PCDF
compounds, potentially 210 distinct compounds, that relates the potency of the
mixture to the unit cancer risk estimate of 2,3,7,8-TCDD (Bellin and Barnes,
1986). The risk assessment methodology referred to as the 2378-TCDD Toxic
Equivalency Factor Method, was reviewed by the Science Advisory Board September
8, 1986.
The carcinogenic risk assessment evaluates the consequence of 70 year
exposure of the pollutant discharged in the stack emission from the MWC.
Potential exposure of the pollutant from other sources either natural or
man-made are not assessed by the HEM. Therefore the computed risk estimate is
considered attributable only to the emissions from the incinerator.
October 1986 3-25 DRAFT-DO NOT QUOTE OR CITE
-------
An estimation will be made of possible noncarcinogenic adverse health
effects resulting from inhalation of the maximum annual ground level (MAAGL)
concentration predicted by the ISC-ST model. A determination of potential
noncarcinogenic risks will be made regarding the mos,t exposed individual(s)
living at or near the ambient MAAGL concentration, and using a risk reference
dose as an appropriate indicator of risk. The EPA has developed risk
reference doses for threshold-acting toxicants which are more thoroughly
discussed in Section 4 of this report. Human exposure is characterized as a
daily intake of the pollutant by inhalation of the predicted MAAGL of the
pollutant. The appropriate toxicologic index, or risk reference dose, will be
compared with the estimated daily intake dose that has been adjusted to
account for intake of the pollutant from sources other than MWC emissions.
Other sources of exposure to the pollutant by inhalation may include a
monitored ambient air concentration at or near the MAAGL, or pollutant
emission from other stationary combustion sources located within the same
modeling radius of 50 km. Thus the background intake of the pollutant is
factored into the characterization of noncarcinogenic risk from inhalation
exposure in much the same manner as calculated in Equation 4-1 of Section 4 in
order to define the increment that could result from emission from the MWC
without exceeding the threshold. If the risk reference dose were exceeded,
then this would signal a potential for an adverse noncarcinogenic health
effect occurring in the population exposed to the stack emission of the
pollutant from the MWC.
3.3 TERRESTRIAL FOOD CHAIN MODEL
Contaminants associated with emissions from MWC are subject to deposition
on surfaces downwind from the MWC. The fallout may be deposited on soil and/or
vegetation. The Terrestrial Food Chain Model (TFC) refers to both the human
food chain and the ecological food chain. The pathways in the model assess
exposure to humans, animals, soil biota and vegetation. Terrestrial trophic
relationships are numerous and complex. These pathways have been selected for
assessing human exposure to deposited contaminants from emissions of MWC and
will be discussed in this section. Humans in the vicinity of the MWC have the
potential to ingest contaminated soil directly or consume vegetation and animal
tissues containing the contaminants. The ecological pathways involve plants,
soil biota and their consumers and are discussed in Section 5.
October 1986 3-26 DRAFT—DO NOT QUOTE OR CITE
-------
3.3.1 General Considerations
3.3.1.1 Most-Exposed Individuals (MEIs). Humans or other organisms may be
exposed to the soil-deposited contaminants of MWC emissions by several path-
ways. For each exposure pathway, it is important to identify the most-exposed
individual, or MEI. Occupational exposures are not considered. It is assumed
that workers involved in the operation of MWCs can be required to use special
measures or equipment to minimize their exposure to possibly hazardous materials.
This methodology is geared toward protection of the general public and the
environment. While many individuals of the general public may be exposed to a
varying degree, the MEI is that individual who would be expected to experience
the greatest risk and, therefore, require the greatest protection. The MEI is
a hypothetical individual, but care should be taken that the definition is
realistic. The definition of the MEI will vary with each pathway.
3.3.1.1.1 Crops for human consumption. This pathway (deposition-soil-plant-
human toxicity) is important wherever crops for human consumption are grown in
a vicinity where emissions from a MWC may be deposited. Uptake of the deposited
contaminants is assumed to occur through the plant roots. Direct adherence of
deposited contaminants or soil to crop surfaces is not considered here.
(Direct ingestion of deposited contaminants is discussed in Section 3.3.1.1.2.)
The MEI is defined as an individual residing in a region within 50 km of a MWC
in the area of maximal deposition of emissions. The individual who grows a
large proportion of his or her own food would result in highest risk since much
of the diet would be potentially affected.
3.3.1.1.2 Soil ingestion by children. Human adults may ingest some soil, but
the amounts consumed by young (i.e., preschool, 1-6 years of age) children are
much greater. This is especially true in children exhibiting the behavior
known as "pica," the ingestion of nonfood items; these children constitute the
MEI for this pathway (deposition-human toxicity). Preschool children with pica
for soil are assumed to be exposed in residential areas within 50 km of the MWC
in the areas of maximal deposition of emissions. The exposure is likely to
occur in gardens, lawns, landscaped areas, parks and recreational areas.
3.3.1.1.3 Herbivorous animals for human consumption. Two separate pathways
are considered whereby animal products may become contaminated: 1) deposition-
soil -pi ant-animal -human toxicity and 2) deposition-soil-animal (direct inges-
tion)-human toxicity. By the first pathway, row crops (i.e., grains) or other
forage crops (i.e., grasses) are grown on soils contaminated by MWC emissions
October 1986 3-27 DRAFT-DO NOT QUOTE OR CITE
-------
and take up contaminants through the roots. The crops are then harvested for
animal consumption. By the second pathway, the deposited contaminants adhere
to crop surfaces or remain in the thatch layer on the soil surface. The crop
is then harvested or grazed, resulting in ingestion of deposited particles or
pollutants. In addition to domestic grazers, wild herbivores, such as deer,
may forage grains or grasses within the range of emissions fallout and be taken
by hunters. The MEI for this pathway is the human consumer of these animal
products.
3.3.1.2 Soil Deposition Rate of Contaminants. The cumulative soil deposition
rate of contaminants (in kg/ha) is determined from the total (dry plus wet)
deposition rate of the pollutant [g/m2 (year)]"1 over the total period of
deposition from the MWC by the following equation:
CD = AD x T x 10 , Equation (3-2)
where:
CD = cumulative soil deposition of pollutant (kg/ha)_
AD = annual deposition rate of pollutant [g/m2 (year *)]
T = total period of deposition (years)
10 = conversion factor [m2 x kg/(ha x g)]
For this methodology, the lifetime of the MWC (or total period of
deposition) is considered to be >30 years; however, since the site is already
dedicated for MWC, it may be assumed that the MWC will be replaced. Therefore,
the lifetime of the MWC could be as great as 100 years.
3.3.1.3 Soil Incorporation- of Deposited Contaminants. Following deposition,
contaminants from emissions of MWC may be incorporated into the upper layer of
soil where crops or other vegetation are grown. If incorporation is accom-
plished by disking or plowing of the upper layer of soil, it is assumed that
the deposited pollutants would be mixed into the soil to a depth of *20 cm (8
inches).
If the contaminants persist indefinitely in the upper soil layer, as is
the case for some inorganics, the following relationship exists between
contaminant deposition rates and concentration increment in soil:
LC = (CD x 10) x (B x D)"1 Equation (3-3)
October 1986 3-28 DRAFT-DO NOT QUOTE OR CITE
-------
where:
LC = maximal soil concentration increment of pollutant (pg/g DW)
CD = cumulative soil deposition of pollutant (kg/ha)
10 = conversion factor [(m2 x ha x ug)/(cm2 x m2 x kg)]
D = depth of soil layer (cm)
B = bulk density of soil (g/cm2)
From Equation (3-3), a soil concentration of 1 ug/g DW corresponds to a
deposition rate of 2.7 kg/ha. LC represents the concentration increment, not
the total concentration, because it it does not take into account background
concentrations of the contaminant that may already be present, whether natural
or from other pollution sources.
Where soil incorporation does not occur, particulate is assumed to be
retained in a shallower, uppermost soil layer. While the actual depth of this
uppermost layer retaining the unincorporated contaminant is unknown, a value of
1 cm will be assumed.
3.3.1.4 Contaminant Loss from Soils. Contaminants may be lost from soils as a
result of numerous processes, including leaching, volatilization and chemical
and biological degradation. These processes may occur simultaneously or at
different rates.
Organic contaminants and some inorganic contaminants may be subject to
some or all of these loss processes; thus, it may be extremely difficult to
model overall rate of loss. A simple means to estimate loss is based on
empirical data from soil systems where soil concentrations have been followed
over time. These data may be used to estimate a first-order loss rate constant
for the pollutant. The use of such a rate constant is recognized to be an
over-simplification since the processes involved are complex and not
necessarily or only first-order. Where no basis for an estimate is available,
no loss should be assumed. The maximal soil concentration of chemicals subject
to loss (for all k >0) may be calculated as a function of the annual deposition
rate constant as shown in the equation below:
LCT = AD x (l-e"kT) x 102 x (B x D x k)"1 Equation (3-4)
October 1986 3-29 DRAFT—DO NOT QUOTE OR CITE
-------
where:
LCT = maximal soil concentration of pollutant after time,
T (ug/g DW)
AD = annual deposition of contaminant [g/(m2 x year)]
MS = 2.7 x 103 Mg/ha = assumed mass of soil in upper 20 cm
k = loss rate constant (years) 1
T = total period of deposition (years)
102 = conversion factor [(m2 x mg)/(cm2 x g)]
B = bulk of density of soil g/cm2
D = depth of soil layer (cm)
This formula is derived from an environmental application of toxicokinetic
principles (O1Flaherty, 1981).
3.3.1.5 Contaminant Uptake Relationships in Plants
3.3.1.5.1 Plant uptake of inorganics. Uptake rates of inorganic chemicals,
especially uptake of metals by plants, have been recently reviewed (CAST, 1980;
Ryan et a!., 1982; Logan and Chaney, 1983). Ryan et al. (1982) used linear
regression (of plant tissue Cd concentration against applied Cd) to derive
uptake response slopes; i.e., the increase in tissue concentration for various
crops. [The term "uptake response" is used to denote an increase in tissue
concentration in response to exposure to a chemical; i.e., the difference in
pre- and post-exposure concentrations.] These authors stated that although the
uptake slope could be altered by several variables, it remained essentially
linear.
More recent work, as reviewed by Page et al. (1986), has shown that plant
response to metals from sludge-amended soil is curvilinear, approaching a
plateau concentration in tissue as sludge application rate increases; however,
metal-adsorptive materials present in the sludge matrix are thought to be
responsible for this effect. Since no such uptake limiting effect has been
demonstrated for deposited metals, uptake response slopes will be assumed to be
linear for this methodology. The assumption that response slopes are linear
means that dietary intake of a contaminant increases continually with
contaminant application (or deposition) to soil (if some or all of the diet
originates from these soils). Ryan et al. (1982) further assumed that response
was related to cumulative Cd application. Using this approach, a limit can be
derived for the cumulative application of a metal based on its dietary
threshold level in humans.
October 1986 3-30 DRAFT—DO NOT QUOTE OR CITE
-------
Linear response slopes can be calculated from any data set where tissue
analyses and cumulative metal deposition or application rates have been
recorded, assuming that the metal is not significantly lost over time. Plant
tissue contaminant concentration (in ug/g DW) is regressed against cumulative
contaminant application or deposition (in kg/ha) for the various treatment
levels, including the control, to calculate the uptake response slope. If
contaminant concentrations in soil, LC, were measured rather than deposition or
application rates, the cumulative deposition, CD, may be calculated based on
Equation (3-3). For inorganics that are lost over time, however, tissue
concentration should be regressed against LC.
Wherever possible, uptake response data derived from areas of emission
fallout will be used. For contaminants lacking such data, linear uptake slopes
derived from other types of chemical application (such as sludge or pesticide
additions), will be assumed to apply to deposited contaminants as well.
If the contaminant is phytotoxic, a maximum tissue concentration for a
given crop will be determined based on available phytotoxicity data, and
assumed as an upper limit to uptake. Phytotoxicity of metals may be altered by
soil pH. Phytotoxicity data chosen should be appropriate for the soil pH of
the fallout region of the MWC if possible. Maximum concentrations are those
associated with severe yield reduction (> 75%) or death of the plant, which
would preclude pollutant passage up the foodchain.
3.3.1.5.2 Plant uptake of orgam'cs. Linear uptake response is also assumed
for organic chemicals and is calculated as described for inorganics, but with
some important differences. Because organic compounds and some inorganics tend
to degrade in soil, plant tissue concentration is usually expressed as a
function of a measured soil concentration, rather than application or
deposition rate, for chemicals subject to loss. Therefore, the soil
concentration (in ug/g DW) is regressed against tissue concentration to
determine the uptake slope.
In addition, because soil concentration rather than deposition rate is
used, and because most of the compounds of concern are xenobiotics, tissue
concentration can be assumed to be zero when soil concentration is zero.
Therefore, the slope reduces to a bioconcentration factor that can be derived
from a single data pair.
3.3.1.6 Contaminant Uptake by Animal Tissues. Linear response slopes are
derived for uptake of inorganics or orgam'cs by animal tissues consumed by
October 1986 3-31 DRAFT-DO NOT QUOTE OR CITE
-------
humans. Tissue concentration is regressed against concentration in feed.
Tissue concentrations in the literature may be expressed in dry or wet weight,
but dry weight is preferred. For uniformity in applying this methodology, all
slopes should be derived based on dry weight (moisture free, but including fat)
concentrations in tissue and feed. Conversion from wet to dry weight for
various tissues should be made according to percent moisture values given in
USDA (1975) or another authoratative source.
. For lipophilic organics, tissue concentration is often expressed on a fat
basis (ug/g fat). If so, the uptake slope should also be expressed on a fat
basis rather than converted to a dry weight basis. Also, the slope for
organics may be the same as a bioconcentration factor derived for a single data
point (i.e., animal tissue concentration and feed concentration), as described
previously for plant uptake of organics (see Section 3.3.1.5.2.).
3.3.1.7 Human Diet. Humans may be exposed to crops or animal products that
have taken up the pollutants through the soil or diet, respectively. In order
to quantify potential dietary exposures, it is necessary to estimate the
amounts of various types of foods in the human diet. The most up-to-date and
detailed source of information regarding food consumption habits of the United
States population is the FDA Revised Total Diet Study food list (Pennington,
1983). While this food list provides a very detailed picture of the United
States diet, it cannot be used in its published form for risk assessments of
the present type. Many of the food items listed are complex prepared foods
(such as soup or pizza), rather than the raw commodities (such as crops or
meats) for which contaminant uptake data are available. Therefore, the
reanalysis of the Pennington (1983) diet that was used in U.S. EPA (1986h)
will also be used in this methodology. Each item in the Pennington diet
(including the infant/junior foods) was broken down into its component parts,
based on information available in FDA (1981) and USDA (1975). The percentages
of dry matter and fat for each component were also specified. These components
were then aggregated into the specific commodity groups required for this
methodology. A summary of consumption for each category by each age/sex group
is presented in Table 3-12. This analysis should be considered preliminary as
it has not been reviewed by the FDA.
3.3.2 Deposition-Soil-PIant-Human Toxicity Exposure Pathway
3.3.2.1 Assumptions. In addition to the assumptions listed in Table 3-13,
assumptions on the percent of diet affected by deposition of contaminants from
October 1986 3-32 DRAFT—DO NOT QUOTE OR CITE
-------
CO
CO
CO
inui_u o J.I..
BASED ON A
REANALYSIS OF THE FDA
REVISED TOTAL DIET FOOD LIST*
Consumption bv Aae-Sex Group (q dry weight/day)
Food Group
subgroup
Grains and cereals
wheat
corn
rice
oats
other grain
Potatoes
Leafy vegetables
Legume vegetables
Root vegetables
Garden fruits
Peanuts
Mushrooms
Vegetable oil
Meats
beef
beef fat
beef liver
beef liver fat
lamb
lamb fat
pork
pork fat
poultry
poultry fat
fish (including fat)
Dairy
dairy fat
Eggs
Other
6-11 Months
42.969
6.239
3.003
6.362
0.006
8.391
0.838
2.475
0.893
0.900
0.243
0.000
30.780
3.006
11.885
0.077
0.1012
0.0570
0.127
1.412
4.869
2.253
5.925
0.339
41.021
39. 196
3.395
130.992
2 Years
61.820
16.938
4.586
3.758
0.015
13.701
0.482
4.683
0.736
2.001
1.661
0.001
27.119
8.475
8.974
0.109
0.146
0.0314
0.070
4.994
8.264
4.515
1.412
1.200
31.854
16.252
6.788
253.974
14-16
Females
81. 980
20.148
5.137
1.779
0.018
21.505
1.141
6.459
1.422
3.766
1.374
0.002
41.938
15.322
14. 783
0.127
0.168
0.0253
0.056
7.892
10.556
6.997
1.602
2.540
32.088
18. 354
5.998
407.624
Years
Males
124.475
26.358
6.638
4.788
4.197
31.854
1.306
10.651
2.278
4.448
3.162
0.002
63.747
24.640
23.702
0.200
0.265
0.0180
0.040
10.942
15.698
8.258
1.809
2.676
49.858
29.677
8.103
527.718
25-30
Females
69.534
14.591
4.776
1.337
9.007
17.481
2.176
7.987
1.491
4.446
1.311
0.003
44.530
15.819
15.167
0.565
0.748
0. 1358
0.302
7.771
10.590
6.607
1.454
3.501
21.788
14. 545
6.286
379.042
Years
Males
102.207
22.439
6.783
2.044
77.912
28.289
2.164
11.891
2.126
5.943
2.840
0.004
75.746
25.050
27.952
0.423
0.5596
0.1064
0.237
13.846
20.046
9.681
2.061
4.670
31.772
21.991
9.197
482.937
60-65
Females
65.114
15.502
4.024
2.044
2.419
15.914
2.783
8.454
1.529
4.797
1.115
0.002
37.029
12.995
12.175
0.443
0.5854
0.0864
0.192
8.036
10. 554
6.736
1.464
3.629
19.108
12.145
7.000
198. 307
years
Males
89.525
20.347
4.396
2.293
25.644
23.837
2.526
11.098
1.866
5.442
2.137
0.002
55.621
20. 558
18.619
0.662
0.8762
0.0873
0.194
12.566
16.582
8.169
1.824
4.108
25.501
16.226
10.468
237.888
*Source: Pennington, 1983; Food and Drug Administration (1981).
-------
TABLE 3-13. ASSUMPTIONS FOR TERRESTRIAL FOOD CHAIN
Functional Area
Assumptions
Ramifications/Limitations
Soil incorporation of
contaminants
If soil incorporation is assumed,
incorporation depth is 20 cm, and the
upper 20 cm soil layer has a dry mass
of 2.7 x 103 Hg/ha.
Contaminant loss from soils
CO
i
CO
Soil background concentration is not
considered.
Trace metal contaminants are assumed to
be conserved indefinitely in the upper
layer unless loss constants are avail-
able.
Degradation of organic contaminants is
first-order.
Plant uptake of inorganics Uptake response slope may vary with pH.
Plant uptake is treated as linear with
application rate until highly toxic
concentrations in tissue are reached.
Plant uptake is treated as linear with
soil concentration until highly toxic
concentrations in tissue are reached.
Animal uptake of contami-
nants
Human diet
Animal tissue concentration is treated as
a linear function of feed concentration.
Pure chemicals added to diet can be used
to determine uptake response slope.
The FDA Revised Total Diet Study food
list (Pennington, 19B3) is representative
of the United States diet. The age/sex
group with the highest consumption of a
given crop group (typically the 25- to
30-year-old male) is the MEI for that item.
By this assumption, as soil concentration of 1 ug/g corresponds
to a pollutant application of 2.7 kg/ha. If actual depth (and
mass) is less, impacts on soil biota in the incorporation layer
could be underpredicted (and vice versa). Ramifications for
effects on plants is less clear, since less of the root zone is
contaminated as incorporation depth decreases.
Effects could be underpredicted if background concentration is
not considered.
Although most heavy metals are tightly bound to soil, measure-
ments tend to show that this assumption often overpredicts
concentrations and therefore, probably overpredicts certain
hazards.
Could over- or underpredict degradation rate, which is complex
and not necessarily first-order.
Uptake response slope determined for one region may over- or
underestimate uptake response slope for a region with different
soil pH.
Hay overpredict tissue response to contaminant application if
relationship is truly curvilinear.
May overpredict tissue response to contaminant application if
relationship is truly curvilinear.
Could under- or overpredict tissue concentration outside the
observed range of response.
Availability of chemicals in soil may differ; especially, nay be
lower, so that uptake is bverpredicted.
While very complete and detailed, the Pennington diet provides no
information on variability within age/sex groups.
-------
MWC emissions are made for this pathway. These assumptions and their potential
limitations are summarized in Table 3-14.
This exposure pathway deals with crops for human consumption. It will be
assumed that home gardeners produce and consume leafy, legume and root
vegetables, potatoes and garden fruits but not grains and cereals, peanuts or
mushrooms. The USDA (1966) survey of United States food consumption in
1965-1966 includes data on the percentages of foods consumed that were
homegrown for urban, rural nonfarm and rural farm households. The highest
percentages of homegrown foods were for rural farm households, which
constituted *6% of all United States households. The rural farm dweller in the
area of maximal deposition within 50 km of a MWC will be taken .as the MEI in
this pathway. All of the homegrown food by the MEI could come from soil
contaminated with deposited emissions from the MWC. The average percent of
annual consumption that is homegrown for various foods from rural farm
households, is shown in Table 3-15.
3.3.2.2 Calculation Method. Uptake response slopes, in units of ug/g DW
(kg/ha) for most inorganics or ug/g DW (ug/g) for organics or chemicals
subject to loss, are used to determine dietary response to the deposited
contaminant. The main disadvantage of using slopes is that the slopes for each
crop used will likely originate from different experimental conditions. In
order to examine total dietary response to a change in conditions (such as soil
pH), all of the response slopes may need to be changed. Lack of adequate
uptake data for a chemical would preclude this calculation.
Step A. Sort Available Uptake Response Data for All Food Crops
For chemicals showing increased uptake at lower pH (i.e., many metals),
the available response data for the crop should be grouped according to whether
soil pH was <6.0 or >6.0. Studies with pH >6.0 should be used only if natural
soils in the vicinity of the MWC have a neutral or alkaline pH.
Step B. Determine Uptake Response Slopes for Each Food Group
Response slopes with units of ug/g DW (kg/ha)" or ug/g DW (ug/g)~ , as
appropriate, are determined for each crop food group shown in Table 3-12. This
slope may be determined as a weighted mean of all the available response slopes
where weighting is according to the dry weight consumption of each crop. If
October 1986 3-35 DRAFT—DO NOT QUOTE OR CITE
-------
TABLE 3-14. ASSUMPTIONS FOR DEPOSITION-SOIL-PLANT-HUMAN
TOXICITY EXPOSURE PATHWAY
Functional Area
Assumptions
Ramifications/
Limitations
Fraction of diet
affected by deposition
of MWC emissions
All of an individual's
homegrown food could
come from soil with
deposited emissions.
The percentage of
homegrown food in
the diet of the MEI
can be estimated from
USDA (1966) survey
data on rural farm
households, which
constituted 6% of all
households.
May overpredict
exposure.
More recent information
if available, might show
significant changes in
both demographics and
gardening habits of
these households.
TABLE 3-15. AVERAGE PERCENT OF ANNUAL CONSUMPTION THAT IS HOMEGROWN FOR
VARIOUS FOODS, RURAL FARM HOUSEHOLDS3
Food Group
Percent Homegrown
Milk, cream, cheese
Fats, oil
Flour, cereal
Meat
Poultry, fish
Eggs
Sugar, sweets
Potatoes, sweet potatoes
Vegetables (fresh, canned, frozen)
Fruit (fresh, canned, frozen)
Juice (vegetable, fruit)
Dried vegetables, fruits
,2
,3
39.9
15.2
1.6
44.
34.
47.9
9.0
44.8
59.6
28.6
11.0
16.7
Calculated from data presented in U.S. Department of Agriculture (1966).
the data do not permit determination of a weighted mean, an unweighted mean may
be taken or, to be conservative, the highest single value may be chosen to
represent that food group.
If the data indicate that a food group consists of some crops that are
relatively high accumulators and some that are relatively low, it may be
worthwhile to subdivide the food group on this basis. If no response slope can
October 1986
3-36
DRAFT—DO NOT QUOTE OR CITE
-------
be determined for a particular crop food group, the highest value for any of
the other groups should be assigned to that group.
Step C. Determine Human Daily Intake
The increase in DI (in ug/day) of an inorganic contaminant due to
deposition of MWC emissions by this pathway is calculated from the response
data and deposition rate (see Section 3.3.1.2) by the following equation:
n
DI = CD x Z (UC. x FC. x DC.) Equation (3-5)
1=1 i i i
where:
DI = increment (above background) of daily intake of pollutants
(ug/day)
CD = cumulative soil deposition of pollutant (kg/ha)
UC.. = uptake response slope for ith food group [ug/g DW (kg/ha) *]
FC. = fraction of ith food group (crop) assumed to originate from
contaminated soil (unitless)
DC. = daily dietary consumption of ith food group (g DW/day)
The derivation of DI by the deposition-soil-plant-human toxicity pathway
for organic contaminants or chemicals subject to loss is largely the same as
the procedure for inorganics based on the linear response model. One differ-
ence is that uptake data are not segregated on the basis of soil pH since no
reason for doing this has been demonstrated. In addition, the procedure for
calculating response slopes for organics differs somewhat from that for
inorganics, in that tissue concentration is treated as a linear function of
soil concentration rather than application or deposition rate, as explained in
Section 3.3.1.5.2. Thus, the maximal soil concentration, LC (from Equation
3-4) of the contaminant is used in the above equation rather than CD for the
determination of the DI for organics or chemicals subject to loss. The LC
should be compared with the phytotoxicity threshold as described in Section
3.3.1.5.1.
Step D. Compare the DI with the RIA
The DI represents the increase (above background) in daily intake (in
ug/day) of contaminant resulting from MWC emissions. To assess whether the
contaminant poses a risk to human health, the DI is compared to an allowable
October 1986 3-37 DRAFT— DO NOT QUOTE OR CITE
-------
intake level based on health effects, referred to as the adjusted reference
intake, or RIA (also in ug/day). Assessment of risk to humans by indirect
exposure to these pollutants and the derivation of RIA is discussed in
Section 4.
3.3.2.3 Input Parameter Requirements. The following individual parameters are
required to calculate the human daily intake for the deposition-soil-plant-
human toxicity pathway.
3.3.2.3.1 Fraction of food group assumed to originate from soil contaminated
by MWC emissions (FC). All of the homegrown food is assumed to originate from
contaminated soil, but homegrown food comprises <60% of the diet of the rural
farm dweller (see Table 3-15). Therefore, values of FC (after rounding) are
0.60 for all vegetables (except dried legumes), 0.45 for potatoes and 0.17 for
dried legumes, based on the average percent of annual consumption of homegrown
foods. Some food groups (such as mushrooms, peanuts and grains and cereals)
are assumed to be unaffected and the FC is set at zero.
3.3.2.3.2 Uptake response slope (UC). Uptake response data are required for
as many crops as possible in the food groups for which FC ^ 0. If UC for the
contaminants varies with soil pH, slopes appropriate to local soil pH should be
chosen as discussed in Step A.
3.3.2.3.3 Daily dietary consumption of food group (DC). Values for DC (in g
DW/day) are needed for each food group for which FC ^ 0. The values chosen
should be appropriate for the MEI. Consumption data presented in Pennington
(1983) and reanalyzed in Table 3-12 are mean values for each of eight age/sex
group. One could define the MEI in terms of the age/sex group having the
highest consumption for all of the food groups combined or for each individual
food group. Alternatively, one could estimate a 95th-percentile consumption
level, based on variability of 1-day consumption; however, this procedure risks
overestimation of long-term consumption.
3.3.2.4 Example Calculations. In this section, examples will be provided for
this pathway using the metal cadmium, and the organic compound benzo(a)pyrene
[B(a)P]. Examples will be calculated for the model MWC facility assumed to be
located in western F.I or i da.
One of the most important steps in these calculations is the selection of
values for each of the parameters involved. Value selection must be based on
careful literature searches and evaluations of available data. For many parts
of this methodology, value selection will be a very data-intensive exercise.
October 1986 3-38 DRAFT—DO NOT QUOTE OR CITE
-------
Since the calculations presented in this section are only for the purpose of
examples, the values employed will not be based on literature searches, but
will rely on a few sources of readily available Information. Therefore, the
results do not constitute actual recommendations for risk assessment.
3.3.2.4.1 Cadmium
Steps A and B. Sort Available Uptake Response Data for All Food Crops and
Determine Uptake Response Slopes for Each Food Group
Studies are sorted according to soil pH. Natural soils in the vicinity of
the southeastern United States would tend to have pH >6.0. The data of Dowdy
and Larson (1975) (pH >6.0) will be used for this example in lieu of all of the
available studies. Response data on carrot and radish are used to derive a
weighted mean response slope for the "root vegetable" food group. Consumption
data are those for the 25- to 30-year-old male (U.S. EPA, 1986h).
Crop type carrot radish
Linear UC [ug/g (kg/ha)"1)] 0.20 0.056
Dry weight consumption (g DW/day) 0.55 0.021
Example Calculation of Weighted Mean
UC «. (0.20 x 0.55) + (0.056 x 0.021)
root veg. = 0.55 + 0.021
= 0.19 ug Cd/g [kg Cd/ha]"1
Step C. Determine Human Daily Intake
CD for cadmium for 30 and 100 years of total deposition is calculated from
Equation (3-2) using an AD = 1.088 x 10"2 g/m2-year. The DI is determined
using Equation (3-5). UC values for each food group have been derived from
the data of Dowdy and Lawson (1975) as illustrated above for root vegetables.
Values of DC are vegetables obtained from Table 3-12. DC values chosen are
those for the age/sex group with highest consumption of that food group.
Values of FC are those for the home garden scenario from Table 3-15.
October 1986 3-39 DRAFT-DO NOT QUOTE OR CITE
-------
Food Group UC DC _£C_ UC x DC x FC
Potatoes 0.038 31.85 0.45 0.545
Leafy vegetables 0.605 2.78 0.60 1.009
Legume vegetables, 0.0053 3.38 0.60 0.011
Legume vegetables, 0.0053 8.51 0.17 0.008
dried
Root vegetables 0.19 2.28 0.60 0.260
Garden fruits 0.073 5.94 0.60 0-260
Total 2.093
For T = 30 years, DI = CD (kg/ha) x 2.093 [ug x ha (kg x day)"1]
= 3.26 (kg/ha) x 2.093 [ug x ha (kg x day) *]
= 6.83 ug/day
For T = 100 years, DI = 10.88 kg/ha x 2.093 [ug x ha (kg x day)"1]
= 22.77 ug/day
This represents the increase in daily human intake due to cadmium from MWC
emissions with total depositions of 30 or 100 years. The DIs are compared with
the RIA for cadmium for adults (7.8 ug/day) and for children (2.4 ug/day), as
determined in Section 4.1.2.
3.3.2.4.2 Benzo(a)Pyrene
Steps A and B. Sort Available Uptake Response Data for AH Food Crops and
Determine Uptake Response Slopes for Each Food Group
Studies for organic compounds are not sorted according to soil pH. Using
the data in Connor (1984) and the consumption data for the 25- to 30-year-old
male (U.S. EPA, 1986h), the weighted mean response slope for the root
vegetables are calculated previously, in the example for cadmium. Leafy
vegetable UC were taken from Connor (1984). Data for other food groups were
not available and were assigned the highest available value (i.e., root
vegetables UC). Calculated values are shown in Step C below.
Step C. Determine Human Daily Intake
The DI is determined using Equation (3-5) with LC in place of CD. The
DC values are obtained from Table 3-12. DC values chosen are those for the
age/sex group with highest consumption of that food group. Values of FC are
those for the home garden scenario from Table 3-15.
October 1986 3-40 DRAFT—DO NOT QUOTE OR CITE
-------
Food Group
Potatoes
Leafy vegetables
Legume vegetables,
nondried
Legume vegetables,
dried
Root vegetables
Garden fruits
UC
1.74
0.42
1.74
1.74
1.74
1.74
Total
DC FC UC x DC x FC
31.85 0.45 24.939
2.78 0.60 0.701
3.38 0.60 3.529
8.51 0.17 2.517
2.28 0.60 . 2.380
5.94 0.60 6.201
40.267
The maximum soil concentration of B(a)P is determined from the annual
deposition using Equation (3-4). For B(a)P, the loss rate constant, k,
estimated from data for biotic loss only is 0.16 (year)" (Bossert and Bartha,
1986}, and may therefore underestimate total soiUloss. The AD is 5.66 x 10
g/(nr x year). The bulk density, B, of 1.5 g/cnr (NRC, 1984) for sandy clay
loam is used. For 30 years,
LC = 5.66 x 10"4 x [i-e'(0-16)(30)] x 102 x (1.50 x 20 x 0.16)"1
= 1.18 x 10"2 ug/g
For 100 years,
LC = 1.18 x 10"2 ug/g
The LC is compared with the phytotoxicity threshold, however, but data are
not available for benzo(a)pyrene. Therefore, for Equation 3-5 (as modified
for organics or chemicals subject to loss):
For 30 and 100 years,
DI = 1.18 x 10"2 ug/g x 40.267 (g/day) = 4.75 x 10"1 ug/day
This represents the increase in daily human intake due to B(a)P from MWC
emissions. The DI is compared with the RIA for B(a)P for adults (6.09 x 10"
ug/day) and for children (8.7 x 10"4 ug/day), as determined in Section 4.1.4.
October 1986
3-41
DRAFT-DO NOT QUOTE OR CITE
-------
3.3.3 Deposition-Human Toxicity ("Pica") Exposure Pathway
3.3.3.1 Assumptions. In addition to many of the assumptions listed in Table
3-13, some additional assumptions are made for this pathway, relating to the
degree of contaminated soil ingestion that could occur and the method of
assessing potential effects. These assumptions and their potential
ramifications are summarized in Table 3-16 and are further discussed in the
following sections.
3.3.3.2 Calculation Method. A human daily intake, 01 (in ug/day DW), is
determined as follows:
DI = LCT x Is x EDA Equation (3-6)
where:
9
EDA = exposure duration adjustment (unit!ess)
I = soil ingestion rate (g DW/day)
LCT = maximal soil concentration increment of pollutant after time,
1 T (ug/g)
Deposited contaminant is assumed to be within the uppermost 1 cm of soil
and ingested soil is assumed to originate from the same 1 cm layer. Soil
concentration, LC, is therefore calculated according to Equation (3-3) for most
inorganic chemicals or Equation (3-4) for chemicals that are subject to loss
processes using 1.50 g/cm as the bulk density and a depth, D, of 1 cm. The DI
determined by this pathway is then compared with the RIA for the contaminants.
3.3.3.3 Input Parameter Requirements
3.3.3.3.1 Soil ingestion rate (I ). Soil ingestion has been recognized as an
important source of exposure to several pollutants. For adults, a value of
0.02 g/day has been used to estimate ingestion (U.S. EPA, 1984c). Children may
ingest soil by either inadvertent hand-to-mouth transfer or by intentional
direct eating. Pica is the term for frequent, intentional eating of non-food
objects. Lepow et al. (1975) estimated that children frequently mouthing their
hands may inadvertently ingest >100 mg of soil/day. Children who eat soil
directly may ingest as much as 5 g/day (U.S. EPA, 1984c), thus establishing a
plausible typical-to-worse range of 0.1-5 g/day.
Studies aimed at more accurately determining the range of ingestion rates
have yielded some data, but are as yet inconclusive. Binder et al. (1985)
October 1986 3-42 DRAFT—DO NOT QUOTE OR CITE
-------
TABLE 3-16. ASSUMPTIONS FOR DEPOSITION-HUMAN TOXICITY ("PICA") EXPOSURE PATHWAYS
Functional Area
Assumptions
Rami fi cati ons/Limi tati ons
Exposure assessment
Effects assessment
Deposited emissions are not necessarily
soil incorporated and may be concentrated
in the uppermost soil layer. Deposited
contaminant is assumed to be distributed
within the uppermost 1 cm of soil, and
ingested soil is assumed to originate
from the same 1 cm layer.
Deposition-contaminated soil may be
ingested by children at the rates
observed in studies of pica for soil.
It is assumed that pica may occur from
1-6 years of age. Cancer potency is
adjusted to reflect a 5-year rather
than 70-year exposure.
Overestimates exposure in situations
where soil incorporation occurs to any
depth >1 cm. If incorporation is to a
lesser depth, exposure is under-
estimated. For example, if exposure
is to fallout dust directly, ingested
soil could approach 100% deposited
particulate.
Overestimates exposure to the extent
the pica child frequents some areas
where deposition has not occurred.
If the child is more susceptible to
chemical carcinogenesis than the
adult, a 5/70 adjustment could result
in underestimation of hazard.
CO
co
-------
conducted a pilot study to establish methods for determining soil ingestion
rates in children living near a lead smelter. Based on these preliminary data
and the value used in U.S. EPA (1986h), a value of 0.5 g/day is suggested as a
reasonably protective value for I . This value represents an estimate of the
95th percentile of soil ingestion in this study population.
3.3.3.3.2 Adjusted reference intake (RIA). Values for adjusted reference
intake (RIA, in ug/day) are derived, based on health effects data, as described
in Chapter 4. Some provisions may apply since this exposure occurs only in
children (1-6 years of age).
3.3.3.3.2.1 Human body weight (BW). A body weight of 10 kg, the
approximate mean body weight at 12 months of age, should be used as specified
in U.S. EPA (1986h).
3.3.3.3.2.2 Total background intake rate of pollutant (TBI). The value
of TBI should be based on a body weight of 10 kg. FDA data for infants/
toddlers may be available; otherwise a factor of 3 is suggested for a downward
adjustment (U.S. EPA, 1986h).
3.3.3.3.3 Exposure duration adjustment (EDA). An adjustment to the DI may be
required, based on the brief duration («5 years) of this exposure. Values for
health risk assessment are usually calculated to be representative of a
lifetime exposure. Derivation of these values are discussed in detail in
Chpater 4. An EDA value is suggested for use with carcinogenic chemicals. The
value is derived on the basis of exposure duration divided by assumed lifetime
(le/L), or 5 years/70 years = 0.07. This adjustment should be carefully
evaluated on a case-by-case basis to ensure that an exposure of DI/EDA will not
lead to toxic effects other than carcinogenesis. For non-carcinogenic
chemicals, the EDA value is equal to 1.
3.3.3.4 Example Calculations
3.3.3.4.1 Cadmium. The DI for cadmium is determined using Equation (3-6),
with I = 0.5 g/day. The LC for cadmium calculated according to Equation
(3-3) using a soil depth of 1 cm is 21.76 and 72.53 ug/g for 30 and 100
years, respectively. The EDA is equal to 1 since ingested cadmium is not
considered carcinogenic.
For 30 years, DI = 21.76 ug/g x 0.5 g/day x 1
= 10.88 US/day
October 1986 3-44 DRAFT—DO NOT QUOTE OR CITE
-------
For 100 years, DI = 72.53 |jg/g x 0.5 g/day x 1
= 36.27 pg/day
The DI is compared with the RIA for cadmium for children, 2.4 ug/day.
3.3.3.4.2 Benzo(a)Pyrene. The organic compound B(a)P is a carcinogen. For
this example, Ig is 0.5 g/day, EDA is 0.07 and LC is 2.36 x 10"1 pg/9 for both
30 and 100 years of total deposition. Therefore, the DI is:
DI = 2.36 x 10"l ug/g x 0.5 g/day x 0.07
= 8.26 x 10"3 ug/day
This value is compared with the RIA for B(a)P determined for children
(BW = 10 kg), 8.696 x 10~4 ug/day (see Section 4.14).
3.3.4 Exposure Pathways for Herbivorous Animals for Human Consumption
3.3.4.1 Assumptions. Animal forage may be contaminated by uptake through the
plant roots of deposited pollutants (deposition-soil-plant-animal-human
toxicity) or by adherence to plant surfaces or roots [deposition-soil-animal
(direct ingestion)-human toxicity] of deposited particulate or contaminated
soil. In both pathways, humans are exposed by consuming animal tissue, which
has taken up the contaminants. In the second pathway, soil incorporation is
not assumed and direct ingestion of the contaminant by farm animals may occur.
Direct ingestion is also possible by animals such as deer that will be taken by
hunters. The amount of game consumed by hunters, however, is assumed not to
exceed the consumption of home-produced meat by farm dwellers. Therefore, the
farm dweller is taken as the MEI for both pathways. Protection of the MEI is
assumed to be protective of hunters as well. Additional assumptions are listed
in Table 3-17.
3.3.4.2 Calculation Method
3.3.4.2.1 Uptake pathway: deposition-soil-plant-animal-human toxicity. To
determine the human daily intake of a contaminant by this pathway, an animal
feed concentration increment of the contaminant (AFC, in ug/g DW) first must
be determined. The equations differ for organics and inorganics since the
units of crop uptake differ. The AFC is calculated from the deposition rate
[for inorganics, Equation (3-2)] or the soil concentration [for organics or
October 1986 3-45 DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 3-17. ASSUMPTIONS FOR PATHWAYS DEALING WITH HERBIVOROUS ANIMALS
Functional Area
Assumptions
Ramifications/Limitations
Fraction of human diet
affected by MWC emissions
Ingestion by animals of
parti culate-contami nated
soil or forage crops
CO
-p.
CTl
More recent information,
if available, might show
significant changes in
both demographics and food
production habits of these
households.
Information to substan-
tiate this assumption is
not immediately available.
The MEI is assumed to be an individual raising
much of his own meats, poultry, eggs and dairy
products. The percentage of home-produced
foods in the diet of the MEI is estimated from
a USDA (1966) survey of rural farm households,
which constituted 6% of all households.
Individuals consuming wild game that forage in
emissions-contaminated areas are assumed to
have no greater exposure than the MEI identi-
fied above.
Contaminant uptake by crops may affect all
animal feeds, but only grazing animals are
affected by adherence. Adulteration by soil
of harvested crops such as grains fed to non-
grazing animals is assumed to be minimal.
Deposited contaminant is assumed to be distri-
buted within the upper most 1 cm of soil, and
ingested soil is assumed to originate from the
same 1 cm layer. Direct ingestion of soil may
occur, and only animals consuming pasture crops
are affected.
The linear response slope of the most responsive Will tend to overpredict
forage crop is used to represent all forage average crop response.
crops in the animal diet.
-------
chemicals that are subject to loss, Equation (3-3)] and the crop uptake
slope as follows:
(for most inorganics) AFC = CD x UC Equation (3-7)
(for chemicals subject to loss) AFC = LCT x UC Equation (3-8)
where:
CD = cumulative deposition of pollutant (kg/ha)
UC = linear response slope of forage crop [ug/g crop DW (kg/ha)'1]
' or [ug/g DW (ug/g) *]
= maximal soil concentration of pollutant, after time T (ug DW)
Following the calculation of the AFC, the human daily intake (DI ug/day)
is calculated as follows:
n
DI = AFC x Z (UA. x FA. x DA.) Equation (3-9)
i=l i i i
where:
AFC = animal feed concentration of pollutant (pg/g)
UA. = uptake response slope of pollutant in the ith animal tissue
[ug/g tissue DW (ug/g feed DW)"1]
DA. = daily dietary consumption of the ith animal tissue food group
1 (g DW/day)
FA. = fraction of ith food group assumed to be derived from animals
feeding on contaminated soil or feedstuffs
The DI determined above for this pathway is then compared with the RIA.
3.3.4.2.2 Adherence pathway: deposition-soil-animal (direct ingestion)-human
toxicity. The animal feed concentration (AFC, in ug/g DW) is calculated as
for the uptake pathway [see Equation (3-7)], but the values will differ
because fewer human food categories will be affected. Soil incorporation is
not assumed; therefore, the upper parts of pasture crops may be contaminated
and may be harvested as feed for cattle or sheep or grazed directly by these
animals. The deposited contaminant or particulate can also accumulate in the
thatch layer of pasture, and be directly consumed by grazing animals. Domestic
animals £such as pigs, poultry) that consume grains or other non-pasture crops
October 1986 3-47 DRAFT—DO NOT QUOTE OR CITE
-------
are assumed not to be affected. When the deposited contaminant is consumed
from plant or soil surfaces, contaminant intake by the grazing animal is
related to the fraction of the animal diet which that soil comprises.
Deposited contaminant is assumed to be within the uppermost 1 cm of soil and
ingested soil is assumed to originate from the same 1 cm soil layer. Soil
concentration (LC) is calculated according to Equation (3-3) for most
inorganic chemicals or Equation (3-4) for chemicals, which are subject to
loss processes, using 1.50 g/cm bulk density and a depth, D, of 1 cm. The
animal feed concentration is derived in terms of the soil concentration and
fraction of the animal diet that is adhering soil. AFC is derived as follows:
AFC = LCT x FS Equation (3-10)
where:
FS = fraction of animal diet that is adhering soil (unitless)
LCT = maximal soil concentration increment of pollutant after time,
T T (ug/g)
3.3.4.3 Input Parameter Requirements
3.3.4.3.1 Fraction of food group assumed to be derived from animals feeding on
soil or feedstuffs contaminated by MWC emissions (FA). As was the case with
FC in the crops-for-human-consumption pathway (see Section 3.3.2.3.1), this
parameter determines which food groups are included in the analysis.
For the first of these two pathways (deposition-soil-plant-animal-human
toxicity), all meat groups except fish will be assumed to be affected,
including beef, lamb, pork, poultry, dairy products and eggs (see Table 3-15).
The second pathway [deposition-soil-animal (direct ingestion)-human toxicity]
is assumed to affect only grazing animals (beef, lamb and dairy food groups are
included).
As for the pathways dealing with crops for human consumption, the MEI for
these two pathways is chosen as a farm family residing within 50 km of a MWC
(in the area of maximal contaminant deposition) raising a substantial
percentage of their own meat and other animal products. The choice of FA
values is based on the percentage of homegrown foods consumed by rural farm
households (see Table 3-15). As stated previously, it is assumed that these FA
values are sufficiently high that an individual consuming wild game from
contaminated areas will also be protected.
October 1986 3-48 DRAFT—DO NOT QUOTE OR CITE
-------
3.3.4.3.2 Dally dietary consumption of food group (DA). Values for daily
dietary consumption (DA, in g DW/day) are needed for each food group for which
FA ± 0. As was described for DC (see Section 3.3.2.3.3.), consumption data are
taken from Table 3-12. The food item "beef liver" includes various other organ
meats consumed by humans in smaller amounts, such as kidney, hearts, etc.
Individuals with a preference for those organs are expected to consume them at
the rates given for beef liver. It is also assumed that consumption of wild
game does not exceed the values of DA for other meats (beef, lamb) and will
also protect hunters.
3.3.4.3.3 Fraction of animal diet that is adhering soil (FS). Studies of
grazing animals indicate that soil ingestion ordinarily ranges from 1-10% of
dry weight of diet (but may range as high as 20%) for cattle and may be >30%
for sheep during winter months when forage is reduced (Thorton and Abrams,
1983). It will be assumed that 100% of deposited contaminant is retained in
the uppermost soil layer («1 cm depth) that animals may ingest. Since lamb
contributes relatively little to the United States diet, a value of FL = 10% or
0.10, based largely on cattle, will be used to represent a reasonably high
exposure situation.
3.3.4.3.4 Uptake slope of contaminant in animal tissue (UA). Uptake slopes
for affected tissues consumed by humans should be calculated as described in
Section 3.3.1.6. The form and availability of uptake response data will affect
the types of food groups included in the calculation of AFC [see Equations
(3-7) and (3-8)] and, therefore, the choice of values for FA and DA as well.
If values of UA can be calculated for several different types of meats, then
each of these meats can be separately included in Equation (3-9). If data
for only a single type of meat are available, it may be necessary to assume
that similar types of meat from different species have similar UA values.
These tissues would then be grouped in Equation (3-9).
This analysis also assumes that UA values for wild game species are no
greater than those for the domestic animals to which most available data apply.
If data contradicting this assumption are available, these values may be used
in Equation (3-9) to calculate DI.
3.3.4.3.5 Linear uptake response of forage crop (UC). The crop chosen for UC
should be the one showing the highest response slope appropriate for the
exposure scenario involved. Grains, legumes (such as soybeans), silage or
grasses could be affected in agricultural use.
October 1986 3-49 DRAFT-DO NOT QUOTE OR CITE
-------
3.3.4.4 Example Calculations
3.3.4.4.1 PI for uptake pathway
3.3.4.4.1.1 Cadmium. The animal feed concentration of cadmium is first
determined using Equation (3-7). The UC value of 0.14 ug Cd/g (kg Cd/ha)"
based on corn silage will be used (Telford et al., 1982). The CD for cadmium
is 3.26 and 10.88 kg/ha for 30 and 100 years, respectively.
Therefore:
For 30 years, AFC = 3.26 (kg/ha) x 0.14 ug Cd/g (kg Cd/ha)"1
AFC = 0.46 ug/g
For 100 years, AFC = 10.88 (kg/ha) x 0.14 ug/g (kg Cd/ha)"1
= 1.52 ug/g
• The DI for cadmium is calculated using Equation (3-9). Values of UA for
various animal tissues are taken from U.S. EPA (1985c) converted from wet to
dry weight basis. The value listed for "beef liver" is actually from sheep
kidney; it is the highest UA value for an organ meat, since the beef liver
consumption data are assumed here to represent any organ meat. No data were
available for pork, so an average of the values for beef and lamb was used.
Data were also unavailable for eggs and dairy products. In this example they
are assumed to be similar to poultry and beef muscle, respectively. The DA
values are derived from Table 3-12. They include fat since the UA values are on
a dry weight basis (including fat). The FA values are from Table 3-15; they
differ for the uptake and adherence pathways.
Animal Tissue Group
Beef
Beef liver
Lamb
Pork
Poultry
Dairy
Eggs
Total
UA
0.003
9.9
005
004
08
0.003
0.08
DA
FA
UA x DA x FA
(when FA 1 0)
53.0
1.54
0.44
33.9
11.7
79.5
0.44
0.44
0.44
0.44
0.34
0.40
0.070
6.71
0.001
,06
,32
.095
0.
0.
0.
8.1
0.48
0.31
7.566
Therefore, the DI for cadmium is:
For 30 years,. DI = 0.46 ug/g x 7.566 g/day = 3.48 ug/day
October 1986
3-50
DRAFT-DO NOT QUOTE OR CITE
-------
For 100 years, DI = 1.52 ug/g x 7.566 g/day = 11.50 ug/day
The DIs for this pathway are compared with the RIA for cadmium.
3.3.4.4.1.2 Benzo(a)Pyrene. The animal feed concentration of B(a)P is
determined using Equation (3-8). The UC value of 0.42 ug/g (ug/g) for
spinach leaves was chosen since it had the greatest UC value of available
foliage data (U.S. EPA, 1985d).
For both 30 and 100 years, AFC = LCj (ug/g) x 0.42 ug/g (ug/g)"1
= 4.96 x 10"3
-2
where LCj (1.18 x 10 ug/g) is calculated based on a soil depth of 20 cm.
B(a)P is not extensively bioaccumulated in animals (U.S. EPA, 1985d),
making UA approximately zero. Therefore, DI of B(a)P by the Uptake Pathway
from consumption of animal tissue is not increased due to MWC emissions.
3.3.4.4.2 DI for adherence pathway
3.3.4.4.2.1 Cadmium. The AFC for this pathway is derived using Equation
(3-10):
AFC = LCT x FS
where:
LCT = maximum soil concentration increment of pollutant after time
T T (ug/g)
FS = fraction of animal diet adhering soil (unitless)
The value of FS will be 0.10 and the soil concentration of Cd (in 1 cm of
uppermost soil layer) is 21.76 and 72.53 ug/g for 30 and 100 years, respectively.
For 30 years, AFC = 21.76 ug/g x 0.10 = 2.18 ug/g
For 100 years, AFC = 72.53 ug/g x 0.10 = 7.25 ug/g
Equation (3-9) 'and the UA and DA values from the uptake pathway are also
used. The FA values from Table 3-15 for meat and dairy (grazing animals only)
are used.
October 1986 3-51 DRAFT-DO NOT QUOTE OR CITE
-------
Animal Tissue Group
Beef
Beef liver
Lamb
Pork
Poultry
Dairy
Eggs
Total
UA
0.003
9.9
0.005
0.004
08
003
0.08
DA
FA
53.0
1.54
0.44
33.9
11.7
79.5
8.1
0.44
0.44
0.44
0
0
0.40
0
UA x DA x FA
0.070
6.71
0.001
0
0
0.
0
095
6.876
The DI for cadmium is 14.99 and 49.85 ug/day, for 30 and 100 years, respec-
tively. These are compared with the RIA for Cd.
3.3.4.4.2.2 Benzo(a)pyrene. The AFC for this pathway is also calculated
using Equation (3-10). The value of FS is 0.10 and the soil concentration of
B(a)P in the upper most 1 cm soil layer is 2.36 x 10 ug/g.
For both 30 and 100 years, AFC = 2.36 x 10"1 ug/g x 0.10 = 2.36 x 10"2 ug/g
Since UA for B(a)P is approximately zero, DI calculated according to
Equation (3-9) is zero, showing MWC emissions do not increase DI of B(a)P by
this pathway.
October 1986
3-52
DRAFT-DO NOT QUOTE OR CITE
-------
3.4 SURFACE RUNOFF
3.4.1 General Considerations. Contaminants associated with partlculates
emitted by municipal waste combustors are subject to deposition on surfaces
downwind from the MWC at rates determined by meteorology, terrain and
particle physics. This fallout 1s subsequently subject to dissolution
and/or suspension 1n runoff after precipitation events. Runoff moves over
the surface of the earth to a surface water body where 1t mixes with other
waters. As a consequence, humans utilizing water from the surface water
body or aquatic life living therein may be exposed to runoff transported
contaminants.
The methodology derived to calculate risks from the surface runoff
pathway was originally developed to evaluate Impacts from the application of
municipal wastewater sludge to land. A detailed discussion 1s available 1n
the document entitled "Development of Risk Assessment Methodology for Land
Application and Distribution and Marketing of Municipal Sludge," April, 1986
(U.S. EPA, 1986h). The methodology Is formulated 1n three successive tiers,
which begin with simple but very conservative estimates, and proceed to more
detailed analyses 1f the first tiers predict unacceptable risks. Both acute
events and chronic exposure are evaluated using standard approaches to cal-
culate runoff volume and associated erosion potential.
3.4.2 Assumptions. A number of assumptions were required to formulate the
risk-based methodology both with respect to runoff generation and subsequent
mixing In the receiving water. The key assumptions are provided In
Table 3-18, with a discussion of their Impact on the methodology. Because
the science Is not exact, the assumptions are mostly conservative (I.e.,
0492P 3-53 10/27/86
-------
TABLE 3-18
Surface Runoff Methodology Assumptions
Functional Area
Assumption
Ramifications
Long-Term Concentrations:
Tier 1
Tier 2/3
General
Source area
All contaminant deposited
on an annual basis 1s trans-
ported to the receiving
water In a dissolved form.
Loadings to the receiving
water can be described as
a function of solids
loading.
Facility operates over a
sufficient period for
surface soil levels to
reach equilibrium where
annual losses equal annual
Inputs.
No settling of particles
1n the deposition zone,
gross erosion reaches the
edge of field.
Event concentrations:
Tier 1
Tier 2/3
Source area
Stream
All contaminant emitted 1n
a year Is lost In a single
runoff event.
No klnetlcally limited
release of contaminant
from residual/soil mix-
ture; I.e., total con-
taminant concentration
Is fully equilibrated
Into adsorbed and dis-
solved phases.
Stream flow Is unchanged
by the storm unless arterial
velocity data are available
from the hydrograph.
Provides an extremely con-
servative estimate since
no losses are considered.
Mechanistically Inappro-
priate for contaminants
with low partition coef-
ficients.
Overpredlcts contaminant
loading by Ignoring loss
mechanisms other than
those used In formulation,
namely, runoff and Infil-
tration.
Maximizes contaminant loss,
Provides extremely conser-
vative estimate with no
provision for losses or In-
complete mobilization.
Maximizes contaminant con-
centration available for
runoff In dissolved phase,
thereby maximizing
loadings.
Overestimates stream
concentration, since the
storm will Increase stream
flow.
0492P
3-54
10/24/86
-------
overpredlct contaminant concentrations). In some Instances, however, the
nature of the effect of a given assumption may vary with site-specific
considerations.
3.4.3 Calculations. The methodology addresses both long- and short-term
exposures as Illustrated In Figure 3-1. In Tier 1. 1t Is assumed that all
contaminant emitted In a given year Is transported to the receiving water In
that year. The total mass flux of contaminant 1s distributed among the
watersheds In which the fallout Is deposited. The downstream boundary flow
Is then used to calculate the resulting concentration In the receiving water:
C1X • [(fa) (WAx)]/Vfx (Equation 3-11)
where: C1X = estimated receiving water concentration of contaminant for
watershed x (M/L»)
Fa = annual mass of contaminant 1n fallout/unit area (M/L2-T)
UAX = area of watershed receiving fallout (L2)
Vfx » total volume of flow of watershed outlet during the period of
observation; I.e., 1 year for chronic exposure (LVT).
For chronic exposure, total annual flow, Vf , 1s applied, while for acute
exposure the total flow over the duration of the storm event, VS , Is sub-
stituted (e.g., VSX » Vfx/365), 1f the 24-hour rainfall event Is evalu-
ated. (Ideally, V$x Is the actual hydrograph velocity for the event If It
Is known.) This very conservative calculation accounts for no losses during
transport and, therefore, overpredlcts contaminant levels. If these over-
predictions do not exceed health-based criteria (reference water concentra-
tions, RWC) or aquatic life-based criteria (ambient water quality criteria,
AWQC) the risks are deemed acceptable, and no further analysis Is required.
If the predictions exceed criteria, a more detailed Tier 2 or 3 analysis 1s
0492P 3-55 10/27/86
-------
Calculate Total Mitt «f
Contearfiiant UUted Annually
(La)
ftltrlottte He Aaena Tl»
Materthedi Khert Fall cut
Ifin Occur
Calculate Total Annual Vein* if flaw
At the Dotfnttreim lountftry of Each
Affected Viterthed
bit
Vts
1 .__ fcf lmfat ••** ta 1
Ulf>^ C*nd1tl«lll 1
i
Cttlwte Avtrage Sell L*«t1
For ConteBlnantc In Cach
IfitersHed fC.l
Jtevt*
Calculate Annual Sell ^articulate 1
Mt Due U Crotlon In Caen Hiterthed j
Calculate ContaKlnant Matt Input]
To Cach Mate.rtntd '
Calculate Total
Ho* At Dwnttrta*
Of Cacti Katertned
Mb •..•"IIK IT..._-|^MB—
Annual ffolna* Of
Calculate St»r« Cvent riowt
For Cacti Vattrthcd
CstlMte Cvtnt Sedivnt Lett
for Caeh Kitenhetf
Calculate Partition Ictwtcn
Paniculate and Plttelwd
Calculate Nisi of Paniculate and
•litolved Contartnant entering
Mater
Calculate to til font Ho* In
Hater (v$I)
Oerfvc Site Specific
Bate On Degradation
And fctmcff
lilt
FIGURE 3-1
Surface Runoff Pathway Methodology
0492P
3-56
10/24/86
-------
required to determine probable receiving water concentrations. The deriva-
tion of RWC 1s discussed In Section 4.3, and the AWQC 1s discussed In
Section 5.2.1.1.
Tier 2 and 3 calculations are Identical. They differ only 1n the
origin of Input values. Tier 3 Is based on site-specific data from empir-
ical observations for such parameters as degradation rate and partition
coefficient. For the long-term (chronic) exposure analysis, receiving water
contaminant Inputs are calculated from estimates of the soil contaminant
concentration and the bulk soil transport to the receiving water.
In order to estimate soil contaminant concentration, 1t Is assumed that
the MWCs are In operation long enough for a steady state concentration to be
reached. By definition, these conditions will prevail when soil levels are
high enough for the sum of zeroth and first-order losses to equal the annual
addition of contaminants 1n fallout. Maximum soil contaminant mass per area
during the period Is calculated as:
-»~ 1 ) (Equation 3-12)
Mm -
kl
where: Mm = maximum contaminant mass per area of soil (M/L2)
k? = annual fallout rate for contaminant less any zeroth order
losses (typically none) (M/L2-T)
ki m first-order loss rate Includes Infiltration losses, which are
controlled by partitioning, degradation and erosion losses
(T'»)
t = time span for analysis, typically the life of the combustor and
Its replacements.
The first-order loss mechanisms can be calculated as follows:
1) Infiltration (1f equilibrium between soil and water 1s assumed)
k * Rc/(B d Kd) (Equation 3-13)
0492P 3-57 10/24/86
-------
where: k-jj = first-order loss rate coefficient for Infiltration (T1)
Re = recharge (L/T)
B = the bulk density (M/L>)
d = depth of Incorporation (L)
Kd » the distribution coefficient (LVM)
Once again. If Infiltrate concentrations are thought to be solubility
limited, a zeroth order rate Is more appropriate.
2) Surface runoff
k1D « X /(B d) (Equation 3-14)
IK c
where: k = first order loss rate coefficient for surface runoff losses
Xe . sediment loss rate (M/L*-T)
B = bulk density (M/L»)
d a depth of Incorporation (L).
3) Degradation
k1Q * l"2/t1/2 (Equation 3-15)
where: k-jg = first-order loss rate for degradation (T'1)
t-|/2 = half-life due to degradation (T).
In general, volatilization would also represent a first-order loss
mechanism. For HWC participate fallout, however. It Is assumed that
volatile species will not be present 1n the solids settling from the atmos-
phere. Even If vapor pressures are measurable, the adsorption phenomena
will reduce their significance. Also, when field measurements are available
for deriving degradation rates, they may reflect all first-order losses
(k.j) and not Just degradation (LC^p). Therefore, care must be taken 1n
selecting Input values.
0492P 3-58 10/27/86
-------
If tilling 1s prevalent In the watershed, concentrations are reduced by
dividing by d (depth of tilling 1n cm) to reflect that runoff only affects
the top centimeter of soil. When no tilling Is practiced, d 1s set at
1.0 cm.
Sediment losses to the receiving water are computed using the Universal
Soil Loss Equation (USLE) (HVschmeler and Smith, 1978):
Xe = R K (LS) C P (Equation 3-16)
where: Xe = sediment loss rate (H/L*-T)
R « "eroslvlty" factor (I'*)
K » "erodabmty" factor (M/L*-T-un1t 'R')
LS » "topographic or slope length" factor (dlmenslonless)
C = "cover management" factor (dlmenslonless)
P - "supporting practice" factor (dlmenslonless).
Guidelines for selection of Input factor values for the USLE 1n each
watershed are provided 1n Ulschmeler and Smith (1978) as well as the
detailed runoff methodology discussion (U.S. EPA. 1986h).
The total annual flux of sediment to a receiving water 1s the product
of the unit area sediment loss. X , and the area of the watershed. HA .
The concentration of the contaminant In the receiving water (C1 ) Is
determined as the ratio of the total contaminant flux and the annual volume
of flow at the downstream watershed boundary. Vf , or:
C1x = Xe WAx Hm/(d Vfx B) (Equation 3-17)
where: Xe = sediment loss rate (N/L2-T)
UAX = area of watershed receiving fallout (L2)
Mm = maximum contaminant mass per area of soil (H/La)
0492P 3-59 10/27/86
-------
d = depth of tilling (L)
Vfx « total volume of flow at watershed outlet during the period of
observation (LVT)
B » bulk density of soil (M/L»)
The point for calculating the annual flow should be the furthest downstream
point where the contaminated watershed feeds the receiving water. For lakes
or estuaries, the outlet flow 1s applied as Vfx- Once again, these cal-
culations are made for each affected watershed or subwatershed unit.
For acute exposures, the methodology focuses on a specific storm
event. To estimate average contaminant mass per area of soil. It 1s once
again assumed that soil levels will Increase until balanced by losses. The
maximum level (Hm) at any time (t) after the combustion 1s Initiated can be
estimated as:
Mm « *20-e ) (Equation 3-18)
*1
where the terms are the same as defined In Equation 3-12. Once again, 1f
tilling Is practiced to a depth of d cm. the contaminant mass per unit
volume of soil (Mm1, In H/L") 1s:
Mm' , Mm/d . J^O-
d k!
Sediment loss to the receiving water Is calculated from the Modified
Universal Soil Loss Equation (MUSLE) (Williams, 1975; Halth, 1980). In this
approach. It Is necessary first to select a watershed retention factor (S)
from the Soil Conservation Service (SCS) runoff curve number (CN) according
to:
S - 2.54[(1000/CN)-10J (Equation 3-19)
where: S = watershed retention factor (cm).
0492P 3-60 10/24/86
-------
Mote that the units for S are centimeters and that S should be converted to
the proper length units for the remainder of the calculations. Using length
units of centimeters for DRt R . M and S 1s preferred. Next, a
runoff depth value Is estimated as:
D s (Rt » Ht - 0.2S)2 (Equation 3-20)
(Rt + Ht + 0.8S)
where: DR - depth of runoff In the watershed (L)
R{•« depth of total rainfall for the storm event (L)
HI s depth of snow melt during the storm event (L)
S = the watershed retention parameter (L)
The storm runoff volume 1s calculated from the runoff depth according to:
0 * (HAJDD (Equation 3-21)
X K
where: WAX = area of the watershed receiving fallout (L2)
Q « volume of runoff (L3)
OR * depth of runoff from the storm event (L)
A trapezoidal hydrograph Is assumed so that the peak runoff rate can be cal-
culated as:
qn = (HAX) PR **t (Equation 3-22)
Mp Tr(Rt - 0.2S)
where: qp = peak runoff rate (L*/T)
UAX = area of the watershed (L2)
Rt & depth of rainfall In the storm event (L)
DR s depth of runoff from the storm event (L)
Tr s duration of the storm event (T)
S = water retention parameter (L)
0492P 3-61 10/27/86
-------
Sediment losses from the storm event are estimated with the MUSLE according
to:
X$ » 11.8(0 qp)°*56K (LS) C P (Equation 3-23)
where: Xs « sediment loss from a single storm (metric tons)
0 - volume of runoff (ma) •
qp - peak runoff (mVsec)
K « "erodablllty" factor (tons/acre-year-unH "R")
LS « "topographic or slope/length" factor (dimenslonless)
C • "cover management" factor (dimenslonless)
P . "supporting practice" factor (dimenslonless)
Again, selection of K, LS, C and P 1s discussed elsewhere (H1schme1er and
Smith, 1978; U.S. EPA. 1986h). This equation 1s an empirical relationship
and the units must be consistent with those shown above.
If the risk evaluation Is to be based on total contaminant exposure,
receiving water concentrations are calculated as In Equation 3-17:
C1x « X$ WAx Hm'/VSx B (Equation 3-24)
using the maximum soil contaminant level Mm1 1n place of the average soil
contaminant level HC. The value for VS (used Instead of Vf ) would be
* X
the total volume of flow of the receiving water during the storm event
rather than a 1-year period. If criteria for risk differentiate between
dissolved and partlculate contaminant, then the total mass of contaminant
Mm' must be partitioned between the adsorbed portion, Aa, and the dissolved
portion. Da. These are derived as:
Aa . [l/(U(e/Kd B))]Mm' (Equation 3-25)
and
Da » n/(U(Kd B/e))]Hm' (Equation 3-26)
°492P 3-62
-------
where: Aa « adsorbed contaminant mass 1n top cm of soil (H/L2-L)
Da = dissolved contaminant mass In top cm of soil (H/L2-L)
e * available volumetric water capacity of the top cm of soil dif-
ference between wilting point and field capacity) (dlmen-
slonless)
Kd « distribution coefficient for contaminant In soil water system
(L*/M)
B * bulk density of soil (M/L>)
Mm1 s maximum level of contaminant In top centimeter of watershed
soil (H/LM.)
The contaminant losses by each route are defined as:
Pxt « [XS/(WAX B)]Aa (Equation 3-27)
and
Pqt « [DR/(Rt + Ht)]Da (Equation 3-28)
where: Pxt = loss of contaminant In adsorbed form (H/L2)
Pqt « loss of contaminant 1n dissolved form (N/L2)
Other terms are as defined In Equations 3-20, 3-24. 3-25 and 3-26. For
these cases, the receiving water concentrations would be derived from:
and
C1x « Pxt HAx/VSx (Equation 3-29)
C1x = P t HAx/VSx (Equation 3-30)
where: C1X - concentration 1n receiving water (H/L9)
pqt and pxt z 1oss rates from Equations 3-27 and 3-28 (M/L2)
WAX * area of watershed (L2)
VSX » total volumetric flow of receiving water during the
storm event (La)
0492P 3-63 10/27/86
-------
3.4.4 Required Inputs. For a Tier 1 analysis, the only Input parameters
required are the annual mass of contaminant fallout on each watershed or
subwatershed unit and the annual flow of the receiving water.
The equations for a Tier 2 and Tier 3 analysis are the same; therefore,
they have the same Input parameters. The difference Is that many of the
parameters used 1n a Tier 2 analysis would be obtained from the literature,
whereas Tier 3 requires all site-specific Input. All the Input parameters
required for the runoff pathway analysis are shown In Table 3-19. The only
Input required for the receiving water analysis 1s the stream flow or the
flow Into a lake or estuary.
3.4.5 Example. The methodology presented above can best be Illustrated
with example calculations. Two site-specific examples are provided below,
one for a long-term average case (chronic) and one for an event-loading
(acute) case. For both cases. Tier 1 and Tier 2/3 analyses are made for two
contaminants, benzo(a)pyrene and cadmium. Contaminant deposition patterns
were modeled for emissions representative of a MWC with an electrostatic
preclpltator.
The annual mass of contaminant fallout (Fa) was calculated by averaging
deposition estimates for a given area over a 5-year period. Deposition
rates were estimated at points on rays spaced every 22.5° at ranges varying
from 200-50,000 m from the MWC. Hence, the sample points form concentric
rings with the combustor at their center. The average annual mass of con-
taminant fallout was obtained by Integrating the measured data over the
watershed area (1 km2) being studied. The double Integral:
l/*R2JJf(r,e) r dr de,
where:
f (500.6) - -f (200.6) r t 5/3f(200.6)-2/3f (500.6)
0492P 3-64 10/27/86
-------
TABLE 3-19
Input Parameters for the Runoff Pathway Methodology
Symbol Function
foe So11 organic content (dlmenslonless)
MAX Area of land In a given watershed on which fallout will be
deposited (km2)
Fa Annual fallout rate/watershed unit (dry kg/ha-year)
t Time period over which facility and Us replacement will
operate (years)
Rc Recharge rate (m/year)
B Bulk density of the soil (g/cm3)
d Depth of soil Incorporation*
H| Total event snow melt (cm)
Rt Total storm rainfall depth (cm)
Tr Duration of storm event (hours)
Kd Distribution coefficient (cmVg)
Vfx Total volume streamflow for the receiving water (i/year)
VSX Event streamflow volume (t/event)
R, K, LS, Input variables for the Universal Soil Loss Equation (USLE)
C and P
6 Soil porosity (dlmenslonless)
t-j/2 Contaminant half-life (I/years)
*d Is used even If not tilled
0492P 3-65 10/27/86
-------
was evaluated for all data points on the 200 and 500 m rings (the area
within the 500 m ring 1s roughly the size of 1 km2 watershed). Integrat-
ing r's from 0-500 m and e's from 0-2*. the resulting average benzo(a)-
pyrene and cadmium fallout was 0.0020779 kg/ha-year and 0.039436 kg/ha-year,
respectively.
It 1s assumed that the partlculate emitted by the municipal waste
combustor 1s spread over various watersheds; however, for the purpose of
Illustration, the calculations are made for only a single watershed of an
area of roughly 1 km2 that contains the HWC facility at Us center.
All Input parameters for the example calculations are shown In
Table 3-20. Data values are for Illustrative purposes only, but are, for
the most part, representative of a real site.
3.4.5.1 Tier 1
3.4.5.1.1 Long-term average loading. For a Tier 1 analysis, calculate the
maximum possible receiving water concentration using Equation 3-11. For
benzo(a)pyrene:
C1 = [10"(Fa x HA )]/Vf
x l x x x
= 10«(0.0020779 kg/ha x 1.0 kma)/3.15 x 10" i/year
« 6.5965 x 10~« mg/l
For cadmium:
C1x = [10«{Fax x
10«(0. 039436 kg/ha x 1.0 km*}/3.15 x 1010 l/year
0.0001252 mg/l
3.4.5.1.2 Event loading. For event loading, VS 1s defined as the total
volume of flow during the selected storm event rather than the 1-year flow
0492P 3-66 10/27/86
-------
TABLE 3-20
Input Parameters for the Example Calculations
Location - western Florida
Soil Type - sandy clay loam
Organic matter content (foc) - 4%
Land use - agricultural, orchards
UA - watershed area. 1 km2 (100 ha)
Fa - annual mass fallout/unit area:
benzo(a)pyrene - 0.0020779 kg/ha-year
cadmium - 0.039436 kg/ha-year
t - elapsed time, 30 years
Re - recharge rate, 0.25 m/year
Vfx - annual streamflow volume, 3.15x1010 i/year
B - soil bulk density. 1.5 g/cm3
d - Incorporation depth, 1 cm
k0 - zeroth order loss rate, 0.0
k] - first-order loss rate
benzo(a)pyrene - degradation + Infiltration and surface runoff,
0.278 year"1
cadmium - Infiltration * surface runoff, 0.168 year'1
k2 - Fa-k0 = Fa
USLE Input Parameters:
R - 400 year"1
K - 0.21 tons/acre/year/unit R=0.21 tons/acre
LS - 0.179
C - 0.5
P - 1.0
CN - 78
Rt - total event storm rainfall, 5 cm
MI - total event snow melt, 0 cm
Tr - storm duration, 6 hours
VSX - event/storm streamflow volume, 8.63xl07 I/event
Kd - partition coefficient:
benzo(a)pyrene - 3000 cmVg
cadmium - 300 cmVg
6 - porosity, 0.2
0492P 3"67 10/27/86
-------
volume (Vfx) used for the long-term analysis. For this Illustration, a
24-hour event Is assumed. Therefore, the flow volume would be:
8.63xl07 l (3.15xl010 I/year * 365 day/year)
The concentration of benzo(a)pyrene 1n the receiving water would be:
C1x « [10«{Fax)(WAx)]/VSx
= 10"(0.0020779 kg/ha){1 km*}/8.63 x 10' i/event)
= 0.002408 mg/i
The concentration of cadmium 1n the receiving water would be:
C1x = [10-(Fax)(HAx)]/VSx
« 10«(0.0039436 kg/ha){l km*)/8.63 x 10' I/event)
« 0.04570 mg/l
3.4.5.2 Tier 2/3
3.4.5.2.1 Long-term loading. To estimate soil contaminant concentration 1t
1s first necessary to estimate rate constants for contaminant loss from
soils. For a degradable contaminant, k. Is the combined first-order loss
rate constant for degradation. Infiltration and surface runoff. For a non-
degradable contaminant, k.. 1s based on the loss due to Infiltration and
surface runoff only. For cadmium (nondegradable), k, was calculated as
follows:
1) Infiltration
k1} = Rc/(B d Kd)
= 0.25 m/year/(1.5 g/cm3)(l cm)(300 cm3/g)
= 0.0556/year
0492P 3-68 10/24/86
-------
2) Surface Runoff"
klR * VtB d)
• 10~« [1688 mt/km»-year/(1.5 g/cm3)(l cm)]
• 0.1125/year
kl = kll * klR
= 0.0556/year + 0.1125/year
= 0.168/year
For benzo(a)pyrene (degradable) a blodegradatlon rate In son of 0.16/year
was estimated from a recent study (Bossert and Bartha. 1986), and k] was
calculated as follows:
1} Infiltration
k-jj « Rc/(B d Kd)
« 0.25 m/year/(1.5 g/cma)d cm)(3000 cmVg)
= 0.0056/year
2) Surface Runoff
Same as 2) above where k1R » 0.1125/year
3) Degradation
k.. = 0.16/year
kl c kll * klR * klD
= 0.0056/year + 0.1125/year + 0.16/year
= 0.278/year
049 2 P
3-69 10/27/86
-------
Following estimation of degradation rates, the soil contaminant concen-
tration may be estimated.
For a degradable contaminant, such as benzo(a)pyrene. the soil con-
taminant concentration can be estimated with Equation 3-12 as:
. 0.0020779 ka/ha-vear [(l-e-<0.278 year"1) (30 years)]
0.278 year"1
« 0.007473 kg/ha
For a nondegradable contaminant, cadmium, the maximum soil contaminant
concentration can be estimated with Equation 3-12 as:
= 0.039436 kg/ha-vear [(l-e-(0.168 year"1) (30 years)]
0.278 year"1
= 0.23322 kg/ha
Bulk soil losses to the receiving water are calculated using Equation
3-16 as follows:
X « 224.64(R)(K)(LS)(C)(P)
= (224.64)(400/year)(0.21 tons/acre/year)(0.179)(0.5)(1.0)
= 1688 mt/km2-year
0492P 3-70 10/27/86
-------
The total annual flux of sediment to a receiving water Is the product
of the unit area sediment loss (Xg) and the area of the watershed (1
km2). The contaminant concentration of benzo(a)pyrene In the receiving
water Is calculated with Equation 3-12 as:
no«)(Xe)(HAx)(Mm)
C1,
-------
If a soil Incorporation depth of 1 cm 1s assumed, then:
Mm' = Mm/(l cm)
Mm1 « 7.473 x 10~» kg/ha-cm
For the nondegradable contaminant cadmium:
. 0.039436 ko/ha-vear [(l-e-(°-™8 year"1) (30 years)]
0.168 year"1
. 0.2332 kg/ha
Mm1 = 0.2332 kg/ha-cm
Sediment loss to the receiving water 1s calculated using the HUSLE.
First, the watershed retention factor, S, Is calculated using Equation 3-19:
S = 2.54[(1000/CN)-10]
= 2.54[(1000/78)-10]
= 7.16 cm
where CN corresponds to hydrologlc soil Group B, moderately low runoff
potential, good hydrologlc conditions and straight row crops.
The runoff depth can be calculated using Equation 3-20 as:
Dp - («t * *t - 0-2S)2
(Rt * Ht * 0.8S)
[5 cm + 0 - (0.2)(7.16 cm)]2
5 cm + 0 + (0.8)(7.16 cm)
1.19 cm
0492P 3-72 10/27/86
-------
Calculate the storm runoff volume from Equation 3-21 as:
0 * 10«{HAX)DR
= (10«){1 km*)(1.19 cm)
* 1.19x10* m>
Calculate the peak runoff rate from Equation 3-22 as;
_ (2.78)(HAx)(Rt)(DR)
p rT {R - 0.2S)
(2.78)(1 km2)(5 cm)(1.19 cm)
(6 hours) [5 cm - 0.2 (7.16 cm)]
« 0.77 m'/sec
The sediment losses from the storm event are calculated from the MUSLE
Equation 3-23 as:
Xs » 11.8 (Q qp)°'56 (K) (LS) (C) (P)
» 11. 8[(1. 19x10* ma)(0.77 m»/sec)3°'56(0.21 tons/acre-year)(0.179)(0.5)(1.0)
= 36.7 mt
The adsorbed and dissolved fractions can be calculated as follows.
First, the total mass of contaminant, Hm', 1s divided Into the adsorbed and
dissolved portions using Equations 3-25 and 3-26.
Benzo(a)Pyrene
Aa * [l/(H(e/KdB))] Mm1
= [1/U(0.2/(3000 cmVg)(1.5 g/cm3))] 7.473 x 10"» kg/ha»cm
= 7.473 x 10"a kg/ha^cm
0492P 3.73 10/27/86
-------
.S g/cm3)/{0.2))] 7.473 x 1CT» kg/ha-cm
* 0.00 kg/ha*cm
Cadml urn
Aa = [1/(U(e/KdB))] Mm*
* [1/[U(0.2/(300 cmVg)(1.5 g/cm3))] 0.2332 kg/ha-cm
* 0.2331 kg/ha-cm
Da = [l/(U(KdB/e))] Mm1
= [1/(U((300 cmVg)(1.5 g/cm3)/(0.2))] 0.2332 kg/ha-cm
* 0.0001 kg/ha* cm
Based on these calculations, virtually all the contaminant loss for
both benzo(a)pyrene and cadmium 1s 1n the adsorbed form. The losses for
each route, adsorbed (P .) and dissolved (Pqt). are defined by Equations
3-27 and 3-28 as follows:
Benzo(a)pyrene
Pxl = [10-*Xe/(HAx)(B)] Aa
= [(10"«)(36.7 mt)/(l kma)(1.5 g/cm3)] 7.473 x!0"» kg/ha-cm
» 1.828 x 10'5 kg/ha
Pqt - lDR/(Rt
= [1.19 cm/ (5 cm + 0 cm)] 0.00 kg/ha*cm
=0.00 kg/ha
Cadmium:
Pxt = [10-*Xe/(WAx)(B)] Aa
= [{10"«)(36.7 mt)/0 km2)(1.5 g/cm3)] 0.2331 kg/ha-cm
« 5.7032 x 10~4 kg/ha
Pqt - [DR/dU * Ht)] Da
= [1.19 cm/ (5 cm * 0 cm)] 0.0001 kg/ha-cm
= 2.3800 x 10"5 kg/ha
0492P 3-74
10/27/86
-------
Once Pxl and Pqt are known, the receiving water concentrations due to
adsorbed and dissolved contaminant, respectively, can be calculated from
Equations 3-29 and 3-30 as follows:
Benzo(a)pvrene:
C1,
Cadmium:
(IO'HI.828 x 10"S kq/haHl km*)
(8.63 x 107 l)
2.1182 x 10'5 mg/l
VS
no'HO.OO kq/ha)(l km2)
(8.63 x 107 i)
0.00 mg/t
VS,
(1Q')(5.7032 x 10~4 kq/ha)(l km2)
(8.63 x 107 I)
6.6086 x 10~« mg/L
049 2 P
3-75
10/27/86
-------
vsx
nO*H2.380Q x IP"5 kq/haHl km*)
(8.63 x 107 I)
2.7578 x 1CT5 mg/l
The C1 for cadmium represents the maximum receiving water concentra-
tion to which humans would be exposed. The long-term C1X for cadmium
(8.3xlO~5 mg/i) is compared with the RUC values for cadmium as deter-
mined In Section 4.3.2. The proper value for comparison depends on whether
the receiving water body Is a source of drinking water (RWC ), fish or
other seafood (RWCf) or both (RWC ,). These RWC values for cadmium are
3.9xlO"» mg/l, 1.6xlO"» mg/l and 1.1 mg/l, respectively. The
long-term C1x for benzo(a)pyrene (2.7x10"* mg/l) 1s compared to RWC
values of 3.0xlO"« mg/l, 3.2xlO"« mg/l or 1.6x10"*, respectively.
Wherever values of C1X are determined to be below the appropriate RWC, It
may be necessary to determine site-specific Input data for Tier 3 cal-
culations.
3.5 GROUNDWATER INFILTRATION
3.5.1 General Considerations. Contaminants associated with partlculates
emitted from MWCs are subject to deposition on surfaces downwind from the
facility. This fallout 1s subsequently subject to dissolution In rain or
meltwater from precipitation events. The dissolved portion can follow one
of two pathways: either move over the surface as runoff to a surface water
body or Infiltrate Into the ground and recharge the groundwater. As a con-
0492P 3-76 10/27/86
-------
sequence, persons using the groundwater may be exposed to groundwater trans-
ported contaminants. Aquatic life Inhabiting surface water bodies fed by
the contaminated aquifer could be exposed as well.
The methodology derived to calculate risks from the groundwater pathway
was originally developed to evaluate Impacts from the landfllllng of muni-
cipal sludge and to evaluate the groundwater pathway associated with the
application of sludges to land. A detailed discussion 1s available In the
document entitled "Development of Risk Assessment Methodology for Municipal
Sludge Landfllllng,1 March 1986 (U.S. EPA. 19861). It Is formulated In
three successive tiers, which begin with simple but very conservative esti-
mates and proceed to more detailed analyses If the first tiers predict
unacceptable risks. Only chronic exposure Is evaluated using standard
approaches to calculate leachate generation and associated groundwater
transport In the unsaturated zone.
3.5.2 Assumptions. A number of assumptions were required to formulate the
risk-based methodology both with respect to leachate generation and subse-
quent transport 1n the unsaturated zone. The key assumptions along with a
discussion of their Impacts on the methodology are provided 1n Table 3-21.
Due to the Inexact nature of the science, any assumptions made are conserva-
tive (I.e., overpredlct contaminant concentrations). In some Instances,
however, the nature of the effect of a given assumption may vary with
site-specific considerations.
3.5.3 Calculations. The methodology for evaluating the groundwater pathway
Is Illustrated In Figure 3-2. Tier 1 Involves the comparison of projected
leachate concentrations with health-based criteria, as discussed In
Section 4.3. Leachate concentrations are predicted on the basis of annual
fallout and recharge rates. It 1s assumed that soil contaminant levels will
0492P 3-77 10/27/86
-------
TABLE 3-21
Assumptions for the Groundwater Pathway Hethodology _
Functional
Area
Assumption
Ramifications
Source term
Unsaturated
zone transport
Equilibrium will be reached
wherein annual Inputs from
fallout are lost In re-
charge as leachate
Leachate pattern 1s level
over the entire year.
One-dimensional flow In
the vertical direction.
Water flow Is steady-state.
Upper boundary has a con-
stant flux of recharge.
Soil characteristics are
constant with depth for
any layer.
Vertical hydraulic grad-
ient of unity (not assumed
In one of two alternative
approaches).
Overpredlcts to the extent
that some contaminants are
Irreversibly bound to
solids and neglects other
loss mechanisms such as
runoff.
Underpredlcts peak concen-
trations but provides
estimates of average
annual loading.
Overpredlcts concentration
since It Ignores horizontal
dispersion.
Overpredlcts concentration
by accelerating flow In a
compressed time period.
Overpredlcts or under-
predlcts, depending on
the soil type.
It 1s Impossible to deter-
mine the effect of this
assumption, since 1t will
vary from site to site.
Overpredlcts concentration
to the extent that grad-
ients may be <1.0 and,
therefore, time of travel
1s slower, allowing for
more degradation.
0492P
3-78
10/24/86
-------
TABLE 3-21 (cont.)
Functional
Area
Assumption
Ramifications
Unsaturated
zone transport
(cont.)
Saturated zone
(MINTEQ
Calculation)
Retardation of organic* 1s
related to soil organic
fraction only.
All adsorption Is revers-
ible.
First-order degradation
mechanism.
Groundwater conditions
dictate geochemistry.
The six groundwater pH-Eh
couplets modeled provide
an adequate set of alter-
natives.
Contaminant Interactions do
not affect geochemistry.
Overpredlcts contaminant
velocity for soils with
low organic content where
mineral Interaction may
predominate.
Overpredlcts concentration
arriving at aquifer.
Mlspredlcts degradation
where higher order rates
are functional: Overpre-
dlcts zero order mecha-
nisms.
Effects would be site-
specific depending on the
quantity and quality of
both the leachate and the
aquifer.
Some sites could have
extreme conditions beyond
those modeled. If pH
values are very low, the
model will underpredlct,
and 1f Eh conditions are
very low, the model will
overpredict contaminant
concentrations.
Effects would be source-
specific.
0492P
3-79
10/24/86
-------
for Each Contaminant 1
IIculm Conta*1nent Concentration
•I Fallout Fl u*/Hecharte_
Tier 1
Moalth
Criteria
It
> Health
CrlUMaV
011 trl but Ion
Coefficient
Lots Rate
Inorganic
Contaminant
tes
Deternlnc T1*c of Travel and
Losses 1n Unsiturated Zone
Considerations
IPtttmlne Concentration In the
'[Aquifer lased en HIH7EQ Cur»es
(•1)• • Concentration of Contaminant 1
-o»End
Tier 3
EioeHB»nta11y Oeterwlne Retardation
ane Oeoraditlon Values
Tier 2
Vcs
FIGURE 3-2
Logic Flow for Groundwater Pathway Evaluation
0492P
3-80
10/24/86
-------
Increase over time until leachate strength Is sufficiently high to deplete
the Input from fallout each year. Hence, leachate concentration Is defined
as:
X1 = Fa/R (Equation 3-31)
where: Xj « leachate contaminant concentration (M/L3)
Fa « annual fallout flux of contaminant/unit area (H/La-T)
Rc « annual depth of recharge (L/T).
Any contaminant found to have a leachate concentration below the relevant
health-based criteria can be eliminated from further consideration. Those
exceeding the criteria are carried forward to a Tier 2 or 3 analysis. Cur-
rently, the methodology can simulate geochemistry for seven metal contami-
nants (Table 3-22). For other Inorganics, It Is necessary to assume that no
precipitation reactions will occur.
The annual mass of contaminant fallout (Fa) should be calculated as the
average value that falls within a 200 m radius of the combust or. The 200 m
distance was chosen because 1t 1s the closest to the combustor that model
estimates can make accurately. In addition, the values at 200 m showed the
greatest deposition (most conservative). Also, a watershed of this size 1s
large enough to provide a potable water supply to several families (based on
a recharge rate of 0.25 m/year and the assumption that each person uses 100
gallons/day). Therefore, 1t 1s not unreasonable to assume that a single
well can be supplied by an area this size.
Presumably, the compliance point falls within the zone of deposition.
Hence, no saturated zone calculations are performed. The compliance point
Is assumed to be that point at the base of the unsaturated zone where leach-
ate enters the saturated zone. For metals, the contaminant concentration In
0492P 3_81 10/24/86
-------
TABLE 3-22
Metal Contaminants Simulated 1n the Geochemlcal
Portion of the Groundwater Pathway
1. Arsenic
2. Cadmium
3. Chromium
4. Copper
5. Lead
6. Mercury
7. Nickel
3-82 10/24/86
-------
the groundwater Is adjusted based on geochemlcal reactions. For organlcs,
the contaminant concentration Is adjusted due to degradation as 1t travels
through the unsaturated zone.
The Tier 2/3 analysis begins at the same point as Tier 1. calculation
of the source-term strength of the leachate from the fallout. Tier 2/3,
however, allows for site-specific Inputs to predict dispersion, degradation,
and retardation effects, which reduce resultant exposure levels. The dif-
ference between Tiers 2 and 3 lies In the number of Input parameters, which
are determined experimentally. In Tier 3, degradation rates and retardation
coefficients are measured directly.
The Initial step 1n Tier 2/3 1s to define the strength of contaminant
In recharge water. As with Tier 1, this 1s done by assuming that equi-
librium will be reached wherein the Inputs from fallout each year will be
transported away from the soil 1n leachate. Equation 3-31 1s used to cal-
culate the average leachate concentration over the life of the facility and
Us replacements.
For degradable contaminants. Equation 3-31 1s modified to:
X1 = Fa(l-e"kt)/t k RC (Equation 3-32)
where: Xj = average leachate contaminant concentration (M/L»)
Fa = fallout contaminant flux (H/L2-T)
k * first-order degradation rate for contaminant (T"1)
t = 1 year (T)
Rc = annual depth of recharge (L/T)
The next step Is to calculate the time required for leachate thus
formed to move downward through the unsaturated zone to the aquifer. This
Is accomplished by making a time of travel calculation with one of two
analytical approaches presented by the U.S. EPA (1985e). The selected
0492P 3-83 10/27/86
-------
approach assumes constant moisture content '1n the son with, steady-state
flow and a unit hydraulic gradient. It uses the soil moisture, pressure and
conductivity relationships described by Campbell (1974) to solve for the
moisture content of the soil:
f = fs(q/KSAT)1/(2b*3) (Equation 3-33)
where: f = field moisture content (LVL»)
q * moisture flux, recharge rate 1n this case (L/T)
KSAJ = saturated hydraulic conductivity of the soil 1n the
unsaturated zone (L/T)
fs = saturated moisture content of the soil (LVL8)
b = negative one times the slope of the log-log plot of metric
potential and saturated moisture content (dlmenslonless)
The velocity and travel times for flow In the unsaturated zone are related
according to:
V » q/f (Equation 3-34)
T = hy/V . (hy)(f)/q (Equation 3-35)
where: V « velocity of flow (L/T)
T = time of travel (T)
hy = depth of the unsaturated zone (L)
For multiple layer systems, a travel time Is calculated for each layer
and the total summed across the unsaturated zone. The total time Is then
divided Into the total depth to derive the equivalent velocity.
The velocity calculated from Equation 3-34 Is for water traveling
through the unsaturated zone. If the contaminant Interacts with the soil 1n
transit, H will travel at a retarded velocity, VB, defined as:
K
VR = V/[l * ((B/e)(Kd))] (Equation 3-36)
where: VR = velocity of the contaminant (L/T)
V = velocity of water calculated from Equation 3-34
B = average bulk density of soil 1n the unsaturated column (H/L9)
0492P 3'84 10/27/86
-------
e * average porosity of soil In the unsaturated column
(dlmenslonless)
Kd » soil-water partition coefficient for contaminant (LVH) ,
If the contaminant 1s degradable. Us concentration will chajige accord-
Ing to:
where: Xa
XV
k
hy
VR
e
Xa - X1 e[-k(hy)/VR] (Equation 3-37)
contaminant concentration upon entry In the aquifer (H/L")
concentration of leachate Initially from Equation 3-32
first-order degradation rate for contaminant (T~M
depth of unsaturated zone (L)
retarded velocity form Equation 3-36 (L/T)
base of natural logarithms, 2718 (unHless)
If the analyst wishes to account for dispersion as well as attenuation
and degradation 1n the unsaturated zone, the water velocity V from Equation
3-34 can be Input to the one-dimensional CHAIN code (Van Genuchton, 1985).
The CHAIN code 1s an analytical solution of the convectlve-dlsperslve trans-
port equation for a one-dimensional case that accounts for retardation and
degradation. The data Inputs to the model Include the average pore water
velocity, the dispersion coefficient, the water content, the pulse or
release time, the retardation factor, the decay rate, and several coeffi-
cients describing the source term. The output from the code Is the concen-
tration of the contaminant plume at the base of the unsaturated zone for a
period of time equal to several contaminant travel times through the
unsaturated zone.
For organic contaminants, the concentration calculated from Equation
3-37 Is the predicted concentration at the compliance point. For metals,
geochemistry can be considered using a series of Input-output curves gen-
erated by the MINTEQ code, which establishes the solubility limits for a
contaminant metal on the basis of groundwater (1on1c) composition, pH, and
Eh conditions. The MINTEQ geochemlcal code can be applied to generate pre-
0492P 3.85 10/27/86
-------
dieted contaminant concentrations under selected groundwater conditions.
M1NTEQ Is a hybrid code that combines an efficient mathematical structure
with a large, well documented thermodynamlc data base. Functionally, the
code models the mass distribution of a dissolved element between various
uncomplexed and complexed aqueous species; 1t also calculates the degree to
which the water Is saturated with respect to the solids 1n the thermodynamlc
data base. Adsorption, precipitation, and dissolution reactions can be
Included In calculations. Detailed documentation of the MINTEQ code and
data can be found In Felmy et al. (1983, 1984), Morrey (1985). and Deutsch
and Krupka (1985). Curves for selected pH-Eh couplets and an explanation of
how the relations were derived can be found In U.S. EPA (19861). An example
MINTEQ curve Is shown In Figure 3-3.
Output from the unsaturated zone (Equations 3-36 and 3-37, or the CHAIN
code and the MINTEQ Input-output curves) are the predicted contaminant con-
centration and timing at the compliance point. The pulse time Is assumed
continuous for the period of operation of the combustor. If the output from
the Tier 2 analysis exceeds health based criteria at the compliance point,
the analyst may choose to Initiate a Tier 3 evaluation of Inputs.
3.5.4 Required Inputs. For a Tier 1 analysis, the only data required are
the contaminant deposition rate, the recharge rate and the health-based
criteria.
Both Tier 2 and Tier 3 have the same Input parameter requirements. The
basic difference Is that Tier 3 uses more site-specific data than Tier 2.
In Tier 2, data on hydraulic conductivity, recharge, depth to groundwater
and soil type should be determined empirically from samples. Other soil-
related properties can be selected on the basis of soil type, while
0492P 3-86
10/27/86
-------
!.! -I
CD
CO
I
00
0.0
Unsatvrated tone Input Concentration (mg/1)
o
I\J
00
o»
FIGURE 3-3
Example HINTEQ Speclatlon Results for Entry of a Contaminant Into the Saturated Zone
for Conditions of pH . 7.0 and Eh = 1.50 mv
-------
degradation rate constants and partition coefficients can be taken from the
literature. For Tier 3, the latter properties should also be determined
empirically with samples from the site.
The Input parameters required for a Tier 2/3 analysis are shown 1n
Table 3-23.
3.5.5 Example. The methodology presented above can best be Illustrated
with example calculations. A site-specific example calculation Is provided
below for two contaminants, cadmium and benzo(a)pyrene. For both cases, a
Tier 1 and a Tier 2/3 analysis are Illustrated.
All Input parameters for the example calculation are shown 1n
Table 3-24. Data values are for Illustrative purposes only but are, for the
most part, representative of a real site. The annual mass of contaminant
fallout (Fa) was calculated by averaging electrostatic preclpltator measure-
ments over a 5-year period. Measurements were conducted along concentric
rings at ranges varying from 200-50,000 m from the combustor on rays spaced
every 22.5°. The average Fa at the 200 m radius was chosen for the reasons
discussed In Section 3.5.3.
3.5.5.1 Tier 1. For a Tier 1 analysis, the leachate contaminant concentra-
tion 1s calculated using Equation 3-31 as follows:
Benzo(a)pyrene
X, . Fa/Rc
= (0.33494 mg/m*.year)/(0.25 m/year)
= 1.340 mg/ma
Cadmium
Xj = Fa/Rc
• (6.42288 mg/m2^year)/(0.25 m/year)
= 25.692 rng/ni*
0492P 3-88
10/27/86
-------
TABLE 3-23
Input Parameters for Groundwater Pathway
PATHWAY DATA
Symbol Source Data
Fa flux rate for contaminant fallout (mg/m*-year)
Re net recharge (in/year)
Unsaturated Zone
hy depth to groundwater (m)
Xj leachate contaminant concentration, (mg/rn3)
Material Type
m Material layer thickness (m)
saturated hydraulic conductivity (m/year)
b slope of matrix potential and moisture content plot (dimension-
less)
fs saturated soil moisture content (mVm3)
B bulk density (kg/m3)
Saturated Zone
pH groundwater pH
Eh groundwater Eh (mvolts)
CHEMICAL-SPECIFIC DATA
Unsaturated Zone
Kd partition coefficient (I/kg)
k degradation rate constant (year"1)
0492P 3-89 10/27/86
-------
TABLE 3-24
Input Parameters for the Example Calculations
Rx Annual depth of recharge, 0.25 in/year
Fa Fallout contaminant flux:
benzo(a)pyrene - 0.56642 mg/m2-year
cadmium - 10.880992 mg/m9-year
fs Saturated moisture content of soil. 0.4 mVm3
q Moisture flux recharge rate. 0.25 m/year
KSAT Saturated soil hydraulic conductivity, 10* m/year
b Slope of matrix potential and moisture content plot, 4.0
hy Depth of unsaturated zone, 2 m
B Soil bulk density. 1.5 g/cm3
e Soil porosity, 0.2
Kd Soil-water partition coefficient:
benzo(a)pyrene - 3000 cmVg
cadmium - 300 cmVg
k First-order degradation rate:
benzo{a)pyrene - 0.16 year'1
cadmium - 0.00 year'1
pH Groundwater pH, 8.0
Eh Groundwater Eh. -200 mv
0492P 3-90 10/27/86
-------
The resulting leachate concentrations are then compared to the health-based
water concentration limits (RHC ) for benzo(a)pyrene (3.0xlO'» mg/l)
and cadmium (3.9xlO"» mg/i) as determined 1n Sections 4.3.2 and 4.3.4.
Because 1 wg/i = l mg/m3, 1t 1s evident that the benzo(a)pyrene con-
centration of 1.340 wg/i 1s greater than RWC limits; therefore. Tier 2/3
analysis would be necessary. Cadmium concentrations exceed the RWC limit
(25.692 vg/i as compared to 3.9 yg/l), requiring further studies to
be peformed.
3.5.5.2 Tier 2/3. A Tier 2/3 analysis begins at the same point as Tier 1,
calculation of the source-term strength of the leachate from the fallout.
Equation 3-31 Is used to calculate the average leachate concentration over
the life of the facility.
For cadmium, which Is nondegradable. Equation 3-31 1s used to calculate
the average leachate contaminant concentration just as 1n Tier 1 above. For
benzo(a)pyrene, which 1s degradable, Equation 3-32 1s used to calculate the
average leachate contaminant concentration as follows:
Fa(l-e-kt)
XI
-------
The next step Is to calculate the time required for leachate to move
downward through the unsaturated zone. First, the constant moisture content
Is calculated using Equation 3-33:
l/(2b+3)
fs
KSAT
= 0.4 mVm* 0.25 m/year
10* m/year
- 0.15
The water velocity travel time through the unsaturated zone can be cal-
culated as follows:
T - (hy)(f)/q
= (2 m) (0.15)7(0.25 m/year)
«1.2 year
Since both benzo(a)pyrene and cadmium travel with a retarded velocity
compared to the groundwater, their velocities through the unsaturated zone
can be calculated with Equation 3-36 as follows:
Benzo(a)pyrene
VR * V/[U((B/e)(Kd))]
= 1.67 m/year/[U((1.5 g/cmV0.2)(3000 cmVg))]
= 7.42 x 10~5 m/year
Cadmium
VR = V/[H((B/e)(Kd))]
= 1.67 m/year/[U((1.5 g/cm3/0.2)(3000 cmVg))]
= 7.42 x 10~4 m/year
Since benzo(a)pyrene 1s degradable. Us concentration will change as 1t
travels through the unsaturated zone. The benzo(a)pyrene concentration at
0492P 3.92 10/27/86
-------
the base of the unsaturated zone before 1t enters the aquifer can be cal
culated wUh Equation 3-37 as follows:
1.238 rag/in* e' «)/(7-42xlO-» in/year)]
= 0.00 ing/in3
Because the benzo(a)pyrene travel time through the unsaturated zone 1s so
long. It degrades before It reaches the aquifer. The cadmium will not
degrade, but travels much slower (1800 times slower) than water. Assuming
no dispersion, the cadmium concentration may still be reduced due to pre-
cipitation reactions In the saturated system. The MINTEQ results for cad-
mium for pH and Eh values close to the values specified In the example
problem are shown In Figure 3-4. An unsaturated zone concentration of 0.025
mg/l for cadmium (Y axis 1n Figure 3-4) corresponds to a concentration of
0.019 mg/l (19 vg/t) In the saturated zone after accounting for
*
speclatlon. The value of 19 pg/t at the compliance point 1s almost 5
times larger than the health based criteria (RWC ) for cadmium of 3.9
vg/l as discussed 1n Section 4.3.
0492P 3'93 10/27/86
-------
10
l\>
TJ
10
pH • 8.1, Eh « 206
0.01
0.125
0.250
0.499
Output Cadriim Concentration In the
Saturated Zone After Speclatlon (mg/1)
FIGURE 3-4
Groundwater Cadmium Speclatlon
-------
3.6 DERMAL EXPOSURE MODEL
The dermal exposure model refers to human skin contact with contaminants
from emissions of MWC deposited on the soil. The issue of dermal absorption
of deposited contaminants is very complex. There is a fundamental lack of
data for percutaneous absorption of chemicals in human skin from soil. Other
factors important for estimation of human exposure to contaminants by the
dermal route also have many uncertainties. The model described here is offered
as a possible approach for the estimation of human exposure and risk associated
with dermal exposure, but it is recognized that in most, if not all cases, the
available data will not provide a satisfactory basis for risk calculations.
3.6.1 General Considerations
3.6.1.1 Most-Exposed Individuals (MEIs). The MEI for this pathway of dermal
exposure is an individual residing within 50 km of a MWC (in the area of
maximal deposition of emissions) who spends a majority of his daily activity
outdoors. Preschool children (between 1-6 years of age) who play outdoors or
rural farmers would most likely be the MEIs, since these groups have the
greatest opportunity for dermal exposure to particulates deposited on the soil.
These children are likely to be exposed in residential areas (gardens, lawns,
parks, etc.). Farmers or individuals who garden would also have the potential
for substantial skin contact with soil. Occupational exposures of workers
involved in the operation of MWCs are not considered here since these workers
can be required to use special measures or equipment to minimize their exposure
to possibly hazardous materials.
3.6.2 Deposition-Human ("Dermal") Toxicity Exposure Pathway
3.6.2.1 Assumptions. For this methodology, it will be assumed that the daily
dermal intake of the contaminant increases continually with contaminant concen-
tration in soil particles contacting the skin. Using this approach, it would
be possible to derive a limit for the concentration of a contaminant in soil
based on a dermal threshold dose in humans. Systemic toxicity thresholds or
carcinogenic potencies of chemicals by a dermal route of exposure have not been
delineated by the U.S. EPA at the present time.
October 1986 3-95 DRAFT—DO NOT QUOTE OR CITE
-------
It will be assumed that 100% of the deposited contaminant participate is
retained in the uppermost 1 cm layer of soil. The maximum concentration of
deposited contaminant within 50 km of a MWC will be used. It is further
assumed that this soil layer contacts the exposed skin of individuals involved
in outdoor activities. Consideration of dermal exposure during outdoor
activity to contaminated particulate deposited on soil most likely accounts for
a more substantial exposure to the individual than contaminated particulate
deposited in indoor (residential) dust (Hawley, 1985). Indoor dust will
therefore be excluded from this model at the present time.
Dermal intake of contaminants will be assumed to be a function of the
fraction of the compound absorbed, contact time (or duration of daily expo-
sure), exposed skin surface area, contact amount (amount of soil accumulated on
skin) and soil contaminant concentration.
Calculation of the daily intake by dermal exposure is the same for
organics and inorganics, since there is immediate exposure potential and
therefore no soil incorporation. Background concentration of contaminants
will not be included because this methodology assesses only the risk associated
with the increase in exposure due to the contaminant from MWC emissions. The
assumptions and uncertainties relevant to this model and their ramifications/
limitations are shown in Table 3-25.
3.6.2.2 Calculation Method. A daily dermal intake (DDI, ug/day) is determined
as follows:
DDI = CT x SA x CA x AF x LC? x (10"3/24) x EDA Equation (3-38)
where:
DDI = human daily dermal intake (ug/day)
CT = contact time (hours/day),
SA = exposed skin surface area (cm2),
CA = contact amount (mg/cm2),
AF = absorption fraction (%/day),
LCj = maximal soil concentration at time, T (ug/g)
EDA = exposure duration adjustment (unitless),
(10 3/24) = conversion factor (10 6 g x 1 day x 10§ ug)/(ug x 24 hours
x rag).
The DDI represents the increase (above background) in daily human dermal
intake (ug/day) due to the contaminant from MWC emissions. To assess whether
the contaminant poses a risk to human health, the DDI is compared to the RIA
October 1986 3-96
DRAFT—DO NOT QUOTE OR CITE
-------
TABLE 3-25. ASSUMPTIONS AND UNCERTAINTIES FOR DERMAL EXPOSURE MODEL
Functional
Area
Assumption/
Uncertainty
Ramifications/
Limitations
Exposure
Assessment
Dermal intake
Fraction of
contami nant
absorbed
Contact time
Exposed skin
surface area
Deposited emissions are
not necessarily soil in-
corporated and may con-
centrate in the uppermost
soil layer. Deposited
contaminant is assumed to
be distributed within the
uppermost 1 cm of soil,
and the soil which con-
tacts the skin is assumed
to originate from the same
1 cm layer.
Linear with respect to
soil concentration.
Chemical and matrix
specific.
Data needed for
dermal absorption in
humans may not be
available.
Dermal absorption varies
with age.
May be concentration
dependent.
The MEIs are exposed to
contaminated particulate
12 hours/day during
outdoor activity.
Total intake is proportional
to exposed surface area;
adults = 2940 cm2:
children = 980 cm2.
Overestimates exposure in
situations where soil
incorporation occurs to any
depth >1 cm.
May overestimate exposure.
Uncertainty for many
contaminants and matrices.
Uncertainty in between
species extrapolation for
dermal exposure. Uncertainty
in route to route extrapola-
tion of absorption.
Absorption data for a
contaminant at one age
may over- or underestimate
absorption at another age.
Study data at one
concentration may over
or under estimate actual
absorption at another
concentration.
May underestimate exposure
by excluding indoor exposure
to dust.
May over-
intake.
or underestimate
October 1986
3-97
DRAFT—DO NOT QUOTE OR CITE
-------
TABLE 3-25. (continued)
Functional
Area
Assumption/
Uncertainty
Ramifications/
Limitations
Soil contact
amount
Effects
assessment
Children =1.5 mg/cm2;
Adults = 1.5-3.5 mg/cm2.
Substantial dermal exposure
may occur only from 1-6
years of age. Cancer
potency is adjusted to
reflect a 5-year rather
than 70-year exposure.
Lifetime exposure may
need to be adjusted
for days/year in contact
with contaminant during
70-year lifetime.
May over-
exposure.
or underestimate
Hazard could be underestimated
if child is more susceptible
to chemical carcinogenesis
than an adult and 5/70 is used
as an adjustment factor.
Uncertainty as to appropriate
adjustment.
(adjusted reference intake, ug/day). Assessment of risk to humans is
discussed in detail in Chapter 4.
3.6.2.3 Input Parameter Requirements
3.6.2.3.1 Absorption fraction (AF). Dermal absorption of a contaminant is
both chemical-specific and matrix-dependent (U.S. EPA, 1982; Hawley, 1985).
Physicochemical properties of the contaminant (e.g., lipid solubility) will
affect dermal absorption. Factors such as pH, molecular size, temperature and
humidity will also influence absorption. Absorption of a contaminant in a
matrix or vehicle is affected by the physicochemical properties of the matrix.
There is uncertainty as to how various soil or particulate types or matrices
would influence absorption of deposited contaminants. Poiger and Schlatter
(1980) demonstrated that a soil matrix reduced the absorption of TCDD when TCDD
was applied to the skin of rats in a soil-water paste.
Dermal absorption in this case refers to the fraction of the applied
dose of the compound absorbed by human skin within a day (i.e., 6%/24 hours).
Such data for a contaminant in a particulate or soil matrix should be used but
are rarely available. Absorption studies of the contaminant by the dermal
route of exposure may not be available, while exposure by other routes may be
present in the literature. Guidelines, however, for route-to-route extrapola-
tion of absorption of toxicants have not been clearly delineated and pose
October 1986
3-98
DRAFT—DO NOT QUOTE OR CITE
-------
uncertainties, especially where dermal absorption is concerned. Extrapolation
of percutaneous absorption data among animal species would introduce additional
uncertainty.
Absorption of contaminants may be altered if the skin is damaged
(diseased, lacerated or abraded) (U.S. EPA, 1982). Percutaneous absorption
may also vary with age. These issues pose uncertainties and need further
research.
The fraction of the contaminant absorbed may vary with concentration
(Feldman and Maibach, 1974). The assumption of complete absorption of a
contaminant irrespective of dose is the most conservative approach (AF = 1),
but may be unrealistic. For example, Kimbrough et al. (1984) suggested human
dermal absorption of TCOD from soil was *1£.
The estimation of the fraction absorbed of a contaminant is very complex
and dependent on many factors as discussed above. A paucity of data regarding
the dermal absorption of chemicals in humans, particularly in particulate or
soil matrices, will make estimation of the fraction of contaminant absorbed
from soil difficult for this methodology.
3.6.2.3.2 Contact time (CT). Contact time is defined as the amount of time/
day the MEIs would spend in association with contaminated soil. Children
playing outdoors and adult farmers would contact soil very frequently, and soil
contact would continue until the cleansing of exposed skin. The maximum
contact time for these individuals will be assumed to be 12 hours/day (Hawley,
1985). Contact time would probably be greater if indoor dust exposure was also
considered.
3.6.2.3.3 Surface area (SA). Total intake of a contaminant would be approxi-
mately proportional to the exposed surface area for absorption. The anatomical
region and surface area of skin that is directly exposed to the contaminant
will affect dermal intake of the contaminant since there is anatomical varia-
bility in percutaneous absorption (Maibach et al., 1971). Exposed surface area
of adults wearing short-sleeved, open-necked shirts, pants and shoes, with no
gloves or hat, is *2940 cm, whereas that of children wearing the same
clothing is «980 cm2 (U.S. EPA, 1984d).
3.6.2.3.4 Contact amount (CA). The contact amount or amount of soil accumula-
ting on skin is considered to have an upper limit of 1.5 mg/cm for children
(U.S. EPA, 1984d; Hawley, 1985) based on the reports of Lepow et al. (1975) and
Roels et al. (1980). The contact amount for children was assumed to also apply
October 1986 3-99 DRAFT—DO NOT QUOTE OR CITE
-------
to adults (as used in U.S. EPA, 1984d). Hawley (1985), however, suggested the
O
soil coating on adults could be as great as 3.5 mg/cm .
3.6.2.3.5 Adjusted reference intake (RIA). Values for adjusted reference
intake (RIA, in ug/day) are derived based on health effects data as detailed in
Chapter 4. Some special provisions may apply when dermal exposure occurs only
during early childhood (1-6 years).
3.6.2.3.5.1 Human body weight (BW). A body weight of 10 kg, the approxi-
mate mean child's body weight at 12 months of age should be used in accordance
with U.S. EPA (1986h).
3.6.2.3.5.2 Total background intake of pollutant (TBI). The total value
of TBI should be based on a child's body weight of 10 kg. Background intake of
pollutants from all routes of exposure should be included.
3.6.2.3.6 Exposure duration adjustment (EDA). An adjustment to the
DDI for dermal exposure to contaminants may be required based on the brief
or intermittent duration of this exposure (U.S. EPA, 1984d). Values of RIA
for threshold and non-threshold (carcinogens) toxicants are calculated to be
representative of lifetime exposure. For dermal exposure to a carcinogen
during childhood, for example, the DDI value should be adjusted on the basis
of exposure duration divided by assumed lifetime (le/L) or 5 years/70 years
= 0.07.
An exposure duration adjustment for lifetime dermal exposure to contami-
nated particulates from MWC should also consider the days/week the MEI would be
in contact with the soil. Seasonal and climatic conditions that might affect
annual exposure duration should also be taken into account for adjustment of
lifetime exposure. U.S. EPA (1984d) suggested 247-365 days/year as a range for
annual exposure duration while Kimbrough et al. (1984) assumed 6 months/year.
Hawley (1985) varied estimates for fraction of week (2/7 to 5/7 days) and
fraction of year (12 days to 6 months/year) for exposure to contaminants based
on age (children, adults) and outdoor or indoor exposure. Individuals such as
gardeners or farmers in southern climates in the United States would be
expected to have a longer annual duration of exposure than those in the
northern climates.
The appropriate factor for the exposure duration adjustment for lifetime
exposure of the MEI is dependent on many of the assumptions discussed above.
An assumption of 365 days/year is obviously the most conservative for esti-
mating lifetime exposure.
October 1986 3-100 DRAFT—DO NOT QUOTE OR CITE
-------
3.6.2.3.7 Relative effectiveness (RE). In the case of dermal exposure,
RE refers to the effectiveness of the absorbed dermal dose relative to an
unabsorbed ingested oral dose. This is because DDI is an estimate of absorbed
dose, whereas RIA is based on ingested dose. Very,little information is
available concerning the toxicological effectiveness of dermally-absorbed
contaminants. If it is assumed that an absorbed dose is equivalently effective
regardless of exposure route, then RE for an absorbed dermal dose relative to
an unabsorbed ingested dose is equivalent to the reciprocal of the gastro-
intestinal absorption fraction. It is recognized, however, that this assump-
tion does not hold in many cases, such as when effects occur at the portal of
entry or when removal, inactivation or activation of the compound before
reaching the target organ varies with exposure route.
3.6.2.4 Example Calculations
3.6.2.4.1 Cadmium. The DDI for cadmium for adults can be estimated using
Equation 3-38 (see Section 3.6.2.2). The values for SA, CA and CT are 2940
n n
cm, 1.5 mg/cm and 12 hours/day, respectively. For the purpose of this
example only, the AF will be arbitrarily set at 1%/day. The LC is 21.76 and
72.53 ug/g for 30 and 100 years, respectively.
For 30 years, DDI = 12 hours/dav x 2940 cm2 x 1.5 mg/cm2 x 0.01/day x
21.76 ug/9 x (10V24) x 1 = 0.48 ug/day.
For 100 years, DDI = 12 hours/day x 2940 cm2 x 1.5 mg/cm2 x 0.01/day x 72.53
ug/g x (10 V24) x 1 = 1.60 ug/day.
To compare the DDI for adults to the RIA, the DDI is converted to an
equivalent DI using the following equation:
DI = DDI x (RE)"1 Equation (3-39)
For 30 years, DI = 0.48 ug/day x (1/0.045)"1 = 0.02 ug/day
For 100 years, DI = 1.60 ug/day x (1/0.045)"1 = 0.07 ug/day
The value of RE is derived as the reciprocal of the absorption fraction for
ingested cadmium. An absorption fraction of 4.5% was assumed in the derivation
of the provisional RfD (Federal Register, 1985) (Section 4.1.2). The equiva-
lent DI is compared with RIA for cadmium for adults.
October 1986 3-101 DRAFT-DO NOT QUOTE OR CITE
-------
The DDI for children Is determined by using an SA of 980 cm2. EDA remains
set at 1 since the RIA is based on noncarcinogenic effects. The DDI for
children, 0.18 and 0.60 ug/day for 30 and 100 years, respectively, is converted
to an equivalent DI for children and compared to the RIA for cadmium for
children.
3.6.2.4.1 Benzo(a)Pyrene. Values of DDI for B(a)P for adults and
children are calculated in the same manner as above for cadmium; however,
the LC for B(a)P, 1.17xlO"2 ug/g for 30 and 100 years, is substituted. For
this example, AF will be arbitrarily set at 1% day. The RE is derived as the
reciprocal of the absorption fraction (50%) for ingested B(a)P (U.S. EPA,
1985c)., The EDA remains at 1 for lifetime exposure.
DDI = 12 hours/day x 2940 cm2 x 1.5 mg/cm2 x 0.01/day x
2.36 x 10"1 ug/g x (10~3/24) x 1 = 5.20 x 10"3 ug/day
To compare the DDI to the RIA, the DDI is converted to an equivalent DI as
follows:
DI = DDI x (RE)"1
DI = 5.20 x 10"3 ug/day x (1/0.5)"1 = 2.60 x 10"3 ug/day
For children, the DDI is calculated using a surface area of 980 cm2 and an
EDA of 0.07.
DDI = 12 hours/day x 980 cm2 x 1.5 mg/m2 x 0.01/day x
2.36 x 10"1 ug/g x (10"3/24) x 0.07 = 1.21 x 10"4 ug/day
The DI for children using the same RE (1/0.5) as for adults is:
DI = 1.2 x 10" 4 ug/day x (1/0.5)"1 = 6.07 x 10"5 ug/day
The DIs are compare'd with the RIA for benzo(a)pyrene.
October 1986 3-102
DRAFT-DO NOT QUOTE OR CITE
-------
4. ESTIMATING CARCINOGENIC AND NONCARCINOGENIC RISKS TO HUMANS
BY INDIRECT EXPOSURE
Risk assessments ordinarily proceed from source to receptor. The source
is first characterized and contaminant movement away from the source is then
modeled to estimate the degree of exposure to the receptor, or MEI. Health
effects are then predicted based on the estimated exposure.
Health effects amy vary according to exposure route. Humans may be ex-
posed directly to MWC emissions by inhalation, or may be indirectly exposed
to pollutants that are deposited and then subsequently enter the human food
chain or a water source or contact the skin. Health effects from direct inha-
lation are discussed in Section 3.2. This chapter addresses effects from in-
direct exposures.
In this methodology, indirect human exposure to a contaminant from the
emissions of a MWC is characterized by daily intake or water concentration of
the contaminant. Daily intake by ingestion or dermal exposure (DI, in ug/day)
will be compared to the adjusted reference intake (RIA, in ug/day) in order to
characterize whether a human health risk exists. The concentrations of pol-
lutants in surface water (Ci'x in mg/£) or groundwater (X.., in mg/£) are com-
pared with the reference water concentration (RWC, in mg/£).
The RIA will be defined as the increase in dietary intake of a contaminant
that is used to evaluate the potential for adverse effects on human health. To
exceed the RIA would be a basis for concern that adverse health effects may
occur in those individuals. The RIA is termed "adjusted" because it is a
health-based reference intake value that has been adjusted from a pre-weight
basis to a particular human body weight and also to account for contaminant
intake from other sources.
The RWC (in mg/£) is defined as a surface water or groundwater concentra-
tion of pollutant used to evaluate the potential for adverse effects on human
health. If a particular concentration in MWC emissions results in surface water
concentration of a pollutant that is due to runoff greater than the RWC, adverse
health effects may occur in a human population using the surface waters as a
October 1986 4-1 DRAFT—DO NOT QUOTE OR CITE
-------
source of drinking water or consuming fish from these waters. Similarly, the
groundwater concentration of a contaminant is compared with the RWC to ascertain
if any adverse health effects might occur as a result of MWC emissions leaching
into the aquifer. Since groundwaters often feed surface water bodies, in some
instances constituting the major source, human exposures could likewise result
from drinking these waters from a surface source or through ingestion of con-
taminated fish.
4.1 DETERMINATION OF THE ADJUSTED REFERENCE INTAKE (RIA)
The procedure for determining RIA varies according to whether the pollutant
acts by a threshold or nonthreshold mechanism of toxicity.
4.1.1 Threshold-Acting Toxicants
Threshold effects are those for which a safe (subthreshold) level of toxicant
exposure can be estimated. For these toxicants, RIA is derived as follows:
RIA
= RfD_xJW - TBII 103 Equation (4-1)
where:
RIA = adjusted reference intake (ug/day)
RfD = risk reference dose (mg/kg/day)
BW = human body weight (kg)
TBI = total background intake of pollutant (mg/day) from all other sources
of exposure
RE = relative effectiveness of exposure (unitless)
103 = conversion factor (ug/mg)
The definition and derivation of each of the parameters used to estimate RIA
for threshold-acting toxicants are further discussed in the following sections.
4.1.1.1 Risk Reference Dose (RfD). When toxicant exposure is by ingestion,
the threshold assumption has traditionally been used to establish an "accept-
able daily intake," or ADI. The Food and Agricultural Organization and the
World Health Organization have defined ADI as "the daily intake of a chemical
which, during an entire lifetime, appears to be without appreciable risk on
the basis of all the known facts at the time. It is expressed in milligrams
of the chemical per kilogram of body weight (mg/kg)" (Lu, 1983). Procedures
for estimating the ADI from various types of toxico.logical data were outlined
October 1986 4-2 DRAFT—DO NOT QUOTE OR CITE
-------
by the U.S. EPA In 1980 (Federal Register, 1980). More recently, the Agency
has preferred the use of a new term, the "risk reference dose," or RfD, to
avoid the connotation of acceptability, which is often controversial.
4.1.1.2 Human Body Weight (BW). The choice of body weight for use in risk
assessment depends on the definition of the individual at risk, and that in
turn depends on exposure and susceptibility to adverse effects. The RfD (or
ADI) was defined as the dose on a body-weight basis that could be safely toler-
ated over a lifetime. Food consumption on a body-weight basis is substantially
higher for infants and toddlers than for teenagers or adults. Certain beha-
viors, such as mouthing of dirty objects or direct ingestion of soil, which
could also contribute to exposure, are also much more prevalent in children than
adults. Therefore, infants and toddlers would be at greater risk of exceeding
an RfD when exposure is by food or soil ingestion. The effects, however, on
which the RfD is based may have a long latency period, in some instances ap-
proaching the human lifespan. In these cases it may be reasonable to base the
derivation of criteria upon adult values of BW (70 kg). In cases where effects
have a shorter latency (<10 years) and where children are known to be at special
risk, it may be more appropriate to use a value of 10 kg as the BW for toddlers
or infants.
4.1.1.3 Relative Effectiveness of Exposure (RE). RE is a urn*tless factor that
shows the relative toxicologies! effectiveness of an exposure by a given route
when compared with another route. The value of RE may reflect observed or esti-
mated differences in absorption rates between different exposure routes, that
are then assumed to translate into a difference in the toxicant's effectiveness.
In addition to route differences, RE can also reflect differences in the expo-
sure conditions, such as ingestion of food or water. Since most exposures in
this group of pathways occur by food consumption, the RE factors applied are
all with respect to ingestion in food.* Therefore, the value of RE in Equation
(4-1) gives the relative effectiveness of the exposure route and circumstances
on which the RfD was based when compared with food.
A RE factor should only be applied where well-documented and referenced
information is available on the contaminant's inhalation and oral pharmaco-
kinetics. When such information is not available, RE is equal to 1.
*The only exception is exposure from soil ingestion. In this case, RE values
should take into account the soil matrix if supporting data are available.
October 1986 4-3 DRAFT-DO NOT QUOTE OR CITE
-------
4.1.1.4 Total Background Intake of Pollutant (TBI). It is important to recog-
nize that sources of exposure other than the soil deposition of MWC emissions
may exist, and that the total exposure should be maintained below the RfD.
Other sources of exposure include background levels (whether natural or anthro-
pogenic) in drinking water, food or air. Other types of exposure that are due
to occupation or habits such as smoking might also be included, depending on
data availability and regulatory policy. These exposures are summed to esti- .
mate TBI.
Data for estimating background exposure usually are derived from analytical
surveys of surface, ground or tap water, from FDA market basket surveys, and
from air monitoring surveys. These surveys may report means, medians, percen-
tiles or ranges, as well as detection limits. Estimates of TBI may be based
on values representing central tendency or on upper-bound exposure situations,
depending on regulatory policy. Data chosen to estimate TBI should be consist-
ent with the value of BW. Where background data are reported in terms of a con-
centration in air or water, ingestion or inhalation rates applicable to adults
or children can be used to estimate the proper daily background intake value.
Where data are reported as total daily dietary intake for adults and similar
values for children are unavailable, conversion to an intake for children may
be required. Such a conversion could be estimated on the basis of relative
total food intake or relative total caloric intake between adults and children.
. As stated in the beginning of this subsection, the TBI is the summed esti-
mate of all possible background exposures, except exposures resulting from
a deposition of MWC emissions.
To determine the effective TBI, background intake values (BI) for each
exposure route must be divided by that route's particular relative effective-
ness (RE) factor. Thus, the TBI can be derived after all the background expo-
sures have been determined, using the following equation:
TRT (mn/c\*^ - Bl(food) + Bl(water) Bl(air) Bl(dermal) . BI(n)
mi tmg/aay; - RE(food) + RE(water) RE(air) RE(dermal) '*'' RfOO
Equation (4-2)
where:
TBI = total background intake rate of pollutant from all other sources of
exposure (mg/day)
October 1986 4-4 DRAFT—DO NOT QUOTE OR CITE
-------
BI - background Intake of pollutant from a given exposure route (indi-
cated by subscript) (rag/day)
RE = relative effectiveness, with respect to the pertinent route of
exposure (indicated by subscript) (unitless)
When TBI is subtracted from the weight-adjusted RfD, the remainder (after
adjusting for RE) defines the increment that can result from MWC emissions with-
out exceeding, the threshold. If upper-bound data (such as 95th percentiles)
were used to estimate TBI, then an increase in exposure corresponding to this
increment, if realized, would cause the RfD to be approached or exceeded in a
relatively small percentage (5%) of the exposed population. If central-tendency
data (the median) were used to estimate TBI, such an increase would cause the
RfD to be approached or exceeded in about half of the exposed population. If
TBI were set at zero for lack of exposure data, the allowed increase would re-
sult in an unknown degree of exceeding the RfD, depending on whether other
sources of exposure exist.
4.1.2 Calculation of RIA for Cadmium
The total background intake (TBI) of cadmium for adults is 27.2 ug/day
(Federal Register, 1985). The provisional RfD for cadmium adjusted for a BW of
70 kg and a RE of 1, is equal to 35 ug/day (Federal Register, 1985). The RIA,
calculated according to Equation (4-1), is 7.8 ug/day for adults.
The provisional RfD for Cadmium of 0.50 ug/kg/day was based on a human
adult oral exposure regime of 350 ug/day, which was estimated to represent a
threshold effect level. The RfD was derived using an uncertainty factor of 10
and dividing by a body weight of 70 kg. It is implied that children exposed to
the same food and water that resulted in this threshold level in adults would
not be adversely affected. Therefore, the exposure limit for children will be
derived based on relative cadmium exposure assuming the same food and water
sources.
Relative cadmium intake for toddlers and adults for the fiscal years 1975-
1977 are available from FDA (1980a,b). Mean daily ingestion was 10.9 ug (todd-
lers) and 34.6 ug (adults). The adult daily ingestion limit (350 ug/10 = 35 ug)
may be scaled to an intake for toddlers as follows:
35 x (10.9/34.6) ug/day = 11.0 ug/day
October 1986 4-5 DRAFT-DO NOT QUOTE OR CITE
-------
This procedure is considered preferable to scaling the value strictly on body
weight (i.e., 0.5 ug/kg/day x 10 kg = 5 ug/day) as would be done if the RfD
originated from animal data expressed on a body weight basis. The distinction
is important because the actual ingested dose is quite close to the RfD.
Although mean adult intake listed above was approximately equal to the RfD, a
more recent estimate of 27.2 ug/day, employed by the U.S. EPA Office of Drinking
Water (Federal Register, 1985), is somewhat lower. If this intake value is
similarly scaled for toddlers, the resulting average daily intake-is 8.6 ug/day.
RIA for the toddler may then be calculated as:
(11.0-8.6) ug/day = 2.4 ug/day
4.1.3 Non-Threshold Toxicants-Carcinogens
For carcinogenic chemicals, the Agency considers the risk of cancer to be
linearly related to dose (except at high dose levels) (Federal Register, 1986).
The threshold assumption, therefore, does not hold, as risk diminishes with
dose but does not become zero or background until dose becomes zero.
The decision whether to treat a chemical as a threshold- or nonthreshold-
acting (carcinogenic) agent depends on the weight of the evidence that it may
be carcinogenic to humans. Methods for classifying chemicals as to their weight
of evidence have been described by the U.S. EPA (Federal Register, 1986). .
To derive values of RIA, a decision must be made as to which classifica-
tions constitute sufficient evidence for basing a quantitative risk assessment
on a presumption of cardnogenicity. Chemicals in classifications A and B,
"human carcinogen" and "probable human carcinogen," respectively, have usually
been assessed as carcinogens, whereas those in classifications D and E, "not
classifiable as to human carcinogen!city because of inadequate human and animal
data" and "evidence of noncarcinogenicity for humans," respectively, have usual-
ly been assessed according to threshold effects. Chemicals classified as C,
"possible human carcinogen," have received varying treatment. For example,
lindane, classified by the Carcinogen Assessment Group (CAG) of the U.S. EPA as
"B2-C," or between the lower range of the B category and category C, has been
assessed using both the linear model for tumorigenic effects (U.S. EPA, 1980b)
and based on threshold effects (Federal Register, 1985). The use of the
weight-of-evidence classification, without noting the explanatory material for
October 1986 4-6 DRAFT—DO NOT QUOTE OR CITE
-------
a specific chemical, may lead to a flawed conclusion since some of the classi-
fications are exposure route dependent. Table 4-1 gives an illustration of
these EPA classifications based on the available weight of evidence.
TABLE 4-1. ILLUSTRATIVE CATEGORIZATION OF CARCINOGENIC EVIDENCE
BASED ON ANIMAL AND HUMAN DATA*
Human
Evidence
Sufficient
Limited
Inadequate
No data
No evidence
Animal Evidence
Sufficient
A
Bl
62
B2
B2
Limited
A
Bl
C
C
C
Inadequate
A
Bl
D
D
D
No Data
A
Bl
D
D
D
No Evidence
A
Bl
D
E
E
The above assignments are presented for illustrative purposes. There may
be nuances in the classification of both animal and human data indicating
that different categorizations than those given in the table should be
assigned. Furthermore, these assignments are tentative and may be modified
by ancillary evidence. In this regard all relevant information should be
evaluated to determine if the designation of the overall weight of evidence
needs to be modified. Relevant factors to be included along with the tumor
data from human and animal studies include structure-activity relationships,
short-term test findings, results of appropriate physiological, biochemical
and toxicological observations, and comparative metabolism and pharmaco-
kinetic studies. The nature of these findings may cause an adjustment of
the overall categorization of the weight of evidence.
If a pollutant is to be assessed according to nonthreshold carcinogenic
effects, the adjusted reference intake, RIA (in ug/day), is derived as follows:
RIA = %*|W - TBI x 103 Equation (4-3)
where:
RIA = adjusted reference intake (ug/day)
qf = human cancer potency [(mg/kg/day) *] _
RL = risk level (unitless) (e.g., 10 5, 10 6, etc.)
BW = human body weight (kg)
RE = relative effectiveness of exposure (unitless)
TBI = total background intake of pollutant (ing/day); from all
other sources of exposure
103 = conversion factor (ug/mg)
October 1986 4-7 DRAFT-DO NOT QUOTE OR CITE
-------
The RIA, in the case of carcinogens, is thought to be protective recognizing
that the estimate of carcinogenicity is an upper limit value. The definition
and derivation of each of the parameters used to estimate RIA for carcinogens
are further discussed in the following sections.
4.1.3.1 Human Cancer Potency (q*). For most carcinogenic chemicals, the line-
arized multistage model is recommended for estimating human cancer potency from
animal data (Federal Register, 1986). When epidemiological data are available,
potency is estimated based on the observed relative risk in exposed versus non-
exposed individuals, and on the magnitude of exposure. Guidelines for use of
these procedures have been presented in the Federal Register (1980, 1985) and
in each of a series of Health Assessment Documents prepared by the Office of
Health and Environmental Assessment, Office of Research and Development (such
as U.S. EPA, 1985b). The potency value normally used in risk assessments is
the upper-bound estimate of the slope of the dose-response curve in the low
dose range, and it is expressed in terms of risk-per-dose, where dose is in
units of mg/kg/day. Thus, q? has units of (mg/kg/day) .
4.1.3.2 Risk Level (RL). Since,by definition no "safe" level exists for expo-
sure to nonthreshold agents, values of RIA are calculated to reflect various
levels of cancer risk. If RL is set at zero, then RIA will be zero. If RL is
set at 10 , RIA will be the concentration that, for lifetime exposure, is cal-
culated to have an upper-bound cancer risk of one case in one million indivi-
duals exposed. This risk level refers to excess cancer risk; that is, over
and above the background cancer risk in unexposed individuals. By varying RL,
RIA may be calculated for any level of risk in the low-dose region; that is,
_2
RL <10 . Specification of a given risk level on which to base regulations is
a matter of policy- Therefore, it is common practice to derive criteria repre-
senting several levels of risk without specifying any level as "acceptable."
4.1.3.3 Human Body Weight (BW). Considerations for defining BW are similar to
those stated in Section 4.1.1.2. The CAG assumes a value of 70 kg to derive
unit risk estimates for air or water. As discussed previously, ingestion
exposures may be higher in children than in adults when expressed on a body
weight basis; however, if exposure is lifelong, values of BW are usually chosen
so as to be representative of adults.
4.1.3.4 Total Background Intake of Pollutant (TBI). As discussed in Section
4.1.1.4, it is important to recognize that sources of exposure other than the
deposited MWC emissions may exist.
October 1986 4-8 DRAFT—DO NOT QUOTE OR CITE
-------
4.1.3.5 Relative Effectiveness of Exposure (RE). In some cases, potency
estimates have been derived on the basis of a different type of exposure than
may occur from food chain contamination. In these cases, the use of RE for
carcinogens is similar to that described earlier for threshold- acting toxicants
(see Section 4.1.1.3). As stated in that section, an RE factor should only be
applied where we 11 -documented and referenced information is available on the
contaminant's pharmacokinetics. When such information is not available, RE is
equal to 1.
4.1.4 Calculation of RIA for Benzo(a)Pyrene
The human cancer potency (q) for benzo(a)pyrene (B(a)P) has been determined by
the U.S. EPA to be 11.5 (mg/kg/day)"1 (U.S. EPA, 1980c). RL, BW and RE are set
at 10 , 70 kg and 1, respectively, for this example. The TBI for adults is
estimated to be *0.88 ug/day (U.S. EPA, 1980c). The RIA is then calculated as
follows using Equation (4-3):
RIA = [ 11.5° (mgAgX) x 1 "0-88xl°3 mg/day]X 1Q3 ^
= (6. 09. 10" 6 mg/day - 0.88xlO"3 mg/day) x 103 ug/mg = -0.87 ug/day
The result obtained is negative because the estimated background intake exceeds
by a factor of >100 the lo"S incremental risk level based on the calculated
canceY potency. This indicates that the cancer risk due to existing levels of
™ A
B(a)P exposure may be as high as on the order of 10 . If so, then it is not
meaningful to calculate RIA for lower levels of RL unless one wishes to deter-
mine an intake based solely on MWC emissions and ignoring other sources. In
the latter case, TBI is set at zero:
RIA "[11.5 (mg/Rgyday? * 1 ' °\* «3 = 6'087><10 * **'**
The RIA for B(a)P for children is derived using a body weight of 10 kg.
The TBI is 0.29 ug/day (U.S. EPA, 1980c); however, the TBI is set at 0 for this
example. The RL, qj and RE are 10"6, 11.5 (mg/kg/day)"1 and 1, respectively.
Using Equation 4-3:
OTA - 10 6 x y> k9 - Ox 103 = 8.696xlO~4 ug/day
RIA 111.5 (mg/kg/day) x 1 J HU y
October 1986 4-9 DRAFT-DO NOT QUOTE OR CITE
-------
4.2 COMPARISON OF DAILY INTAKES (DI) AND DERMAL DAILY INTAKE (DDI) WITH
THE ADJUSTED REFERENCE INTAKE (RIA)
Human dally intakes for the various pathways of the Terrestrial Food Chain
Model (DI, in ug/day) estimate the increase in contaminant intake by ingestion
due to emissions from MWC. The DIs can be directly compared with the RIA (in
jjg/day), which represents the increase in dietary intake of a contaminant that
is used to evaluate the potential for adverse effects on human health. When
comparing the DIs for ingestion with the RIA, the RE value in Equation (4-1) is
equal to 1.
In order to determine the health risk associated with increased daily
dermal intake (DDI, in ug/day) to contaminants from MWC emissions, the DDI must
be compared to a toxic threshold or carcinogenic potency of contaminants by a
dermal route of exposure. Such information for dermal exposure of toxicants
has not been defined by the U.S. EPA at the present time. As shown in Section
3.7.2.4, the DDI can be transformed to an equivalent DI, by multiplying the DDI
by (RE) , or the ratio of the oral dose to 1
equivalent DI can then be compared to the RIA.
by (RE)"1, or the ratio of the oral dose to the dermally absorbed dose. The
4.3 DETERMINATION OF THE REFERENCE WATER CONCENTRATION (RWC)
The procedure for determining RWC varies according to whether the pollutant
acts by a threshold or non-threshold mechanism of toxicity.
4.3.1 Threshold-Acting Toxicants
Threshold effects are those for which a safe (subthreshold) level of toxicant
exposure can be estimated. For the groundwater and surface runoff pathways (if
the source of contaminant is only drinking water), RWC (in mg/£) is derived as
follows:
where:
RfD = risk reference dose (mg/kg/day)
BW = human body weight (kg)
I = total water ingestion rats (A/day)
TBY = total background intake of pollutant (mg/day) from all other
sources of exposure
October 1986 4-10 DRAFT—DO NOT QUOTE OR CITE
-------
RE = relative effectiveness of exposure with respect to drinking
water exposure of the exposure route indicated by subscript
(umtless)
If the only source of pollutant is fish living in,polluted surface waters,
the reference concentration in water is calculated according to the following
equation:
RWCf =!B!d_x_BW . TBIL (BCF x If) Equation (4-5)
where:
BCF = 'biconcentration factor in fish (A/kg)
If = human consumption of fish (kg/day)
If the source of pollutant is both drinking water and fish from polluted
surface water, the reference concentration is calculated according to Equation
(4-6):
RWCwf
jRfD_x__BW - TBI -r [Iw + (BCF x If)] Equation (4-6)
The definition and derivation of each of the parameters used to estimate vari-
ous. RWCs for threshold acting toxicants are further discussed below.
4.3.1.1 Risk Reference Dose (RfD). The RfD is defined in Section 4.1.1.1.
4.3.1.2 Human Body Weight (BW). The BW is defined in Section 4.1.1.2.
4.3.1.3 Water Ingestion Rate (!,,)• It varies widely among individuals accord-
ing to age and sex. Table 4-2 shows the variation of adult drinking water in-
take within and among several studies. Mean intakes in New Zealand, Great Bri-
tain, The Netherlands and Canada varied from 0.96-1.30 £/day, and 90th percen-
tiles varied from 1.64-1.90 A/day. The variation of mean drinking water intake
and body weight with age and sex for the United States population are illus-
trated in Table 4-2. The choice of values for use in risk assessment depends
on the definition of the individual at risk, which in turn depends on exposure
and susceptibility to adverse effects. The RfD (or ADI) was defined above as
the dose on a body weight basis that could be safely tolerated over a lifetime.
As shown in Table 4-2, water consumption on a body-weight basis is substantially
higher for infants and toddlers than for teenagers or adults. Therefore,
infants and toddlers would be at greater risk of exceeding an RfD when exposure
is by drinking water; however, the effects on which the RfD is based may have
October 1986 4-11 DRAFT—DO NOT QUOTE OR CITE
-------
TABLE 4-2. WATER INGESTION AND BODY WEIGHT BY AGE-SEX GROUP
IN THE UNITED STATES
Age-Sex Group
6-11 months
2 years
14-16 years, female
14-16 years, male
25-30 years, female
25-30 years, male
60-65 years, female
60-65 years, male
Mean Water
Ingest ion
(mA/day)
308
436
587
732
896
1050
1157
1232
Median
Body Weight
(kg)
8.8CC
13.5^
51. 3^
54. 2^
58-5d
67.6°
67. 6*
73. 9d
Water Ingest ion
per Body Weight
(m£/kg/day)D
35.1
32.2
11.4
13.5
15.3
15.5
17.1
16.7
aSource: Pennington, 1983. From the revised FDA Total Diet Study.
Includes categories 193, 195-197, 201-203.
The water ingestion per body weight ratios have been derived from the
referenced values for illustrative purposes only.
Source: Nelson et al., 1969. Calculated by averaging several age or sex
groups.
Source: Society of Actuaries, 1959. Average body weights for median
heights of 156 cm (5 feet 5 inches) and 173 cm (5 feet 8 inches) for
females and males, respectively.
a long latency periods, in some instances approaching the human lifespan. In
these cases, it may be reasonable to base the derivation of criteria upon adult
values of BW and Iw- In cases where effects have a shorter latency (i.e., <10
years) and where children are known to be at special risk, it may be more appro-
priate to use values for toddlers or infants.
The approach currently employed in the derivation of recommended maximum
contaminant levels (RMCLs) by the U.S. EPA Office of Drinking Water is to
assume an Iw of 2.0 £/day (Federal Register, 1985) and an Iw of 1.0 £/day for a
10 kg child.
4.3.1.4 Relative Effectiveness of Exposure (RE). The RE is defined in Section
4.1.1.3. The RE factors for the Surface Runoff and Groundwater Pathways are
applied with respect to drinking water exposure.
4.3.1.5 Total Background Intake Pollutant (TBI). The TBI is defined in Sec-
tion 4.1.1.4.
4.3.1.6 Biconcentration Factor (BCF). Bioconcentration is the ability of
living organisms to accumulate substances to higher than ambient level
concentrations. The degree to which a chemical accumulates in an aquatic
October 1986 4-12 DRAFT—DO NOT QUOTE OR CITE
-------
organism above ambient concentrations is indicated by BCF. Specifically, it is
defined as the quotient of the concentration of a substance in all or part of
an aquatic organism (mg/kg fresh weight) divided by the concentration in water
to which the organism has been exposed (mg/£). The BCF is usually determined
at equilibrium conditions, or for 28-day exposures, and is based upon the fresh
weight of the organism. The BCF therefore has units of mg/kg (mg/4)"1 or £/kg.
Biconcentration is distinguished from other terms commonly used to de-
scribe increases in the concentration of chemicals in an organism, such as bio-
magnification, bioaccumulation or ecological magnification, in that bioconcen-
tration considers only the uptake of a pollutant by an organism from the ambient
water. The other similar processes are associated with increases in the concen-
tration of chemicals resulting from consumption of contaminated food sources
as well as accumulation from water.
Although it has been documented in numerous studies that bioconcentration
may be the primary pathway for accumulation (Marcelle and Thome, 1984; Bahner
et al., 1977; Clayton et al., 1977), there is also evidence that biomagm'fica-
tion by aquatic food chains can be important under certain environmental cir-
cumstances (Lee et al., 1976).
Bioconcentration factors are specific for the compound and the species
absorbing the compound. The compounds with the greatest tendency to bioaccum-
ulate are those that are lipophilic and resistant to biological degradation.
Initial diffusion into the organism occurs by rapid surface adsorption or parti-
tioning to the lipoprotein layer of cell membranes. Once in the bloodstream,
subsequent accumulation of the chemical into particular compartments of the
organism is dependent upon the metabolic capabilities of the organism and the
lipid content of the individual organs. With continuous exposure to a compound,
the condition is eventually reached when the rate of excretion is equal to the
rate of uptake.
Bioconcentration factors can be estimated through laboratory experiments,
field studies, correlations with physicochemical factors such as octanol/water
partition coefficients, and models based upon pollutant biokinetics coupled to
fish energetics. In the development of the ambient water quality criteria, the
U.S. EPA used mostly laboratory data in the calculation of BCFs. Field data
are often less reliable than laboratory data because it cannot usually be shown
that constituent concentrations in field situations have been held constant for
a long period of time or over the range of territory inhabited by the organism.
October 1986 4-13 DRAFT—DO NOT QUOTE OR CITE
-------
BCFs calculated from field data also may be greater than those calculated from
laboratory data, which is apparently due to ingestion of the compound through
prey, sediments and water, in addition to absorption from water.
Where laboratory and field data are not available, BCFs can be estimated
by several methods. Correlations between BCFs and octanol/water partition co-
efficients, water solubility and soil adsorption coefficients have been docu-
mented. Veith et al. (1979) developed the following equation using the corre-
lation between the BCF and the n-octanol/water partition coefficient (P) to
estimate BCFs to within 60% before laboratory testing:
Iog10 = 0.85 Iog10 P - 0.70 Equation (4-7)
The equation was developed using data from whole-body analyses of *7.6% lipids
(Federal Register, 1980). The U.S. EPA adopted the equation developed by Veith
et al. (1979) for use in determining BCFs for use in the exposure sections of
the health effects chapter of the AWQC documents in those cases where an appro-
priate BCF is not available. In a later study, Veith et al. (1980) used the
results of their own laboratory experiments and data from other laboratories
for a variety of fish species and 84 different organic chemicals to obtain the
following modification of their original equation:
Iog10 BCF = 0.76 Iog10 P - 0.23 Equation (4-8)
Equations similar to the ones developed by Veith et al. (1979, 1980) have been
developed for more specific chemical classes and particular aquatic species
(Veith et al., 1979; Neeley et al., 1974). Other investigators (Norstrom et
al., 1976) have developed more elaborate models using pollutant biokinetics and
fish energetics in addition to using octanol/water partition coefficients to
predict BCFs.
Since bioconcentration for lipophilic compounds depends on lipid content
of the fish, it is important to adjust measured or estimated BCF values for
these compounds to reflect the lipid content of seafood in the United States
diet. The U.S. EPA determined in 1980 that the average lipid content of fresh-
water and estuarine species, weighted by average daily consumption, was 3.0%
(Federal Register, 1980). Since fresh and estuarine waters would be those
impacted by runoff from areas of MWC deposition, a lipid content of 3% should
be assumed. The adjustment is made as follows:
October 1986 4-14 DRAFT—DO NOT QUOTE OR CITE
-------
LC.
BCFg = BCFu j-gS Equation (4-9)
e
where:
BCFa = adjusted BCF (£/kg)
BCFy = unadjusted BCF (£/kg)
LCd = lipid content of dietary seafood (kg/kg)
LCg = lipid content of experimental organism (kg/kg)
4.3.1.7 Fish Consumption Rate (If). Several recent publications have provided
estimates of average daily intake of fish. A USDA survey conducted 1977-1978
estimated mean intake to range from 9-r 14 g/day (including all types of fish
such as shellfish and canned fish) depending on geographic region, with the
Northeast showing the highest value (USDA, 1985). Another survey (USDA, 1984)
estimated the national average for fish consumption to be 12.9 pounds/year at
16 g/day. This latter document separates fish into categories, including fish,
shellfish, canned fish, etc. Daily intake of fresh or frozen fish was 6.46
g/day.
The most recent fish consumption document from the U.S. Department of
Commerce (1985) reports total per capita fish and shellfish consumption ranging
from 12.8 pounds/year in 1980 to 13.6 pounds/year in 1984, which is the highest
consumption on record (Table 4-3). The latter value is a daily intake of 16.9
g of fish (all kinds). These figures do not include any recreational catch,
which is estimated to be an additional 3-4 pounds/year or 3.7-5 g/day (U.S.
EPA, 1980d). If one assumes a value of 3.5 pounds/year (4.35 g/day) from recre-
ational fishing, the total average per capita intake of all types of seafood
is «21.25 g/day.
Runoff containing MWC emissions could affect freshwater and estuarine
species, but not the marine species that constitute the greater portion of
seafood in the U.S. diet. To estimate average daily consumption of the former,
the U.S. EPA examined data from a survey of fish consumption in 1973-1974 (as
reanalyzed in U.S. EPA, 1980d) and eliminated all species not taken from fresh
or estuarine waters (Stephan, 1980). Per capita consumption was reduced from
13.4 to 6.5 g/day, or by a factor of 2.1. Therefore, it seems reasonable to
assume that in most instances freshwater and estuarine species will constitute
»50% of total consumption or *10.6 g/day.
October 1986 4-15 DRAFT—DO NOT QUOTE OR CITE
-------
TABLE 4-3. UNITED STATES ANNUAL PER CAPITA CONSUMPTION OF COMMERCIAL
FISH AND SHELLFISH, 1960-1984*
Per Capita Consumption
Year
1960
1961
1962
1963
1964
1965
1966
1967
1968
1969
1970
1971
1972
1973
1974
1975
1976
1977d
1978d
1979d
Civilian Resident
Population
Million Persons
178.1
181.1
183.7
186.5
189.1
191.6
193.4
195.3
197.1
199.1
201.9
204.9
207.5
209.6
211.6
213.8
215.9
218.1
220.5
223.0
Fresh .
and Frozen
Pounds
5.7
5.9
5.8
5.8
5.9
6.0
6.1
5.8
6.2
6.6
6.9
6.7
7.1
7.4
6.9
7.5
8.2
7.7
8.1
7.8
Canned0
of Edible
4.0
4.3
4.3
4.4
4.1
4.3
4.3
4.3
4.3
4.2
4.5
4.3
4.9
5.0
4.7
4.3
4.2
4.6
5.0
4.8
Cured
Meat
0.6
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.4
0.4
0.5
0.5
0.4
0.5
0.4
0.5
0.4
0.3
0.4
Total
10. J
10.7
10.6
10.7
10.5
10.8
10.9
10.6
11.0
11.2
11.8
11.5
12.5
12.8
12.1
12.2
12.9
12.7
13.4
13.0
(continued on following page)
October 1986
4-16
DRAFT—DO NOT QUOTE OR CITE
-------
TABLE 4-3 (continued)
Per Capita Consumption
Year
1980d
1981d
1982d
1983d
1984d
Civilian Resident
Population
Million Persons
225.6
227.7
229.9
232.0
234
Fresh .
and Frozen
Pounds
8.0
7.8
7.7
8.0
8.3
Canned
of Edible
4.5
4.8
4.3
4.8
5.0
Cured
Meat
0.3
0.3
0.3
0.3
0.3
Total
12.8
12.9
12.3
13.1
13.6
aSource: U.S. Department of Commerce (1985).
Beginning in 1973, data include consumption of artificially cultivated
catfish.
C8ased on production reports, packer stocks and foreign trade statistics
for individual years.
Domestic landings data used in calculating these data are preliminary.
Note: The consumption figures refer only to consumption of fish and shell-
fish entering commercial channels, and they do not include data on
consumption of recreationally caught fish and shellfish, which since
1970 is estimated to be between 3 and 4 pounds (edible meat)/person
annually. The figures are calculated on the basis of raw edible meat
(e.g., .excluding bones, viscera, shells). The U.S. Department of
Agriculture (USDA) consumption figures for red meats and poultry are
based on the retail weight of the products, as purchased in retail
stores. The USDA estimates are the net edible weight to be *70-95% of
the retail weight, depending on the cut and type of meat. From 1970
through 1980, data were revised to reflect the results of the 1980
census.
There is a great disparity from the national average, depending on region,
age, race and religion. SRI reported fish intake by the Black and Jewish
populations to be double the average value (U.S. EPA, 1980d). The New England
and East South Central regions had the highest fish consumption regionally.
Consumption levels in the upper 95th percentile were typically 300-400% of the
national average. The highest value in the upper 95th percentile was for
Orientals at 67.3 g/day. This is 502% of the national average reported by SRI.
October 1986 4-17 DRAFT-DO NOT QUOTE OR CITE
-------
Applying the same percentage increase to the revised daily average consumption
of 10.6 g/day, the 95th percentile is estimated to be «53 g/day or *43 pounds/
year.
4.3.2 Calculation of RWC for Cadmium
The total background intake (TBI) of cadmium for adults is 27.2 pg/day
(Federal Register, 1985). An RfD for cadmium (adjusted for a BW of 70 kg and a
RE of 1) equal to 35 ug/day has been proposed (Federal Register, 1985) but has
not yet been officially adopted by the Agency. The Iw for an adult is 2.0
A/day (Federal Register, 1985). Using Equation (4-4), the RWCW for cadmium is
3.9 ug/£.
A value for BCF is needed to calculate RWCf or RWCwf. Biconcentration
data for cadmium are surveyed in the document, Ambient Water Quality Criteria
for Cadmium - 1984 (U.S. EPA, 1985f). The geometric mean of all values avail-
able for edible parts of bivalve molluscs is 659, whereas that for edible parts
of all other species consumed by humans is 13. If these values are weighted
according to relative consumption of bivalves (12.2%) and other organisms
(87.8%), a weighted average BCF for cadmium in the edible portion of all fresh-
water and estrarine aquatic organisms consumed in the United States is calcu-
lated to be 92. As stated earlier, daily consumption of freshwater and estua-
rine species at the 95th percentile is estimated to be 53 g/day (I. = 0.053
kg/day).
Using Equations 4-5 and 4-6, respectively, RWC- is calculated to be 1.6
ug/2 and RWCwf is 1.1 ug/2.
4.3.3 Non-Threshold Toxicants-Carcinogens
The Agency's classification of chemicals as carcinogens are discussed in Sec-
tion 4.1.3.
If a pollutant is to be assessed according to non-threshold, carcinogenic
effects, the reference concentration in surface water or groundwater used for
drinking, but not supplying fish for humans consumption (RWC in mg/2), is
calculated as follows:
RWCW = |nrll " TBI|* !w Equation (4-10)
October 1986 4-18 DRAFT—DO NOT QUOTE OR CITE
-------
where:
qj = human cancer potency [(mg/kg/day)"1]
RL = risk level (unitless); e.g., 10 5, 10~6, etc.
BW = human body weight (kg)
I = water ingestion rate (Jd/day)
Rt = relative effectiveness of exposure (unitless)
TBI = background intake of pollutant (mg/day); from all other sources
of exposure
If the only source of pollutant is fish from polluted surface waters, the
reference concentration (RWCf) is calculated according to the following:
RWC =
- TBIJT (BCF x If) Equation (4-11)
If the sources of pollutant are both drinking water from surface waters and
fish living in these surface waters, then the reference concentration (RWCw^)
can be calculated according to the following:
E-i
% * pc - TBlU [I + (BCF x If)] Equation (4-12)
-)l X KC I W I
where:
TBI = background intake of pollutant (mg/day) from all other sources
of exposure
RE = relative effectiveness of exposure (unitless)
BCF = biconcentration factor in fish (2/kg)
I = water ingestion rate (A/day)
Ij = human consumption of fish (kg/day)
q* = human cancer potency ([mg/kg/day] *)
Rt = risk level (unitless)
TBI, RE, BCF, I and If are defined as for threshold-acting toxicants (see
Sections 4.1.1.3, 4.1.1.4, 4.3.1.3, 4.3.1.6, 4.3.1.7) and q* and RL are
discussed in Sections 4.1.3.1 and 4.1.3.2.
4.3.4 Calculation of RWC for Benzo(a)Pyrene
The q* for orally ingested B(a)P has been determined to be 11.5 (mg/kg/
day)"1 (U.S. EPA, 1980c). RL, BW and RE are set at 10"6, 70 kg and 1 for this
example. The TBI for adults is estimated to be 0.88 ug/day (U.S. EPA, 1980c).
For this example, however, the TBI will be set at 0. The Iw is 2 £/day for
October 1986 4-19 DRAFT-DO NOT QUOTE OR CITE
-------
adults (Federal Register, 1985). Using equation 4-10, the RWCw for B(a)P is
S.Oxlo"6 mg/£.
In the Ambient Water Qualtiy Criteria Document for Polynuclear Aromatic
Hydrocarbons (U.S. EPA, 1980c), a BCF of 30 was determined as representative of
fish, based on very limited available data. A geometric mean BCF of 79 for
bivalves may be calculated from data in the same document. A weighted average
BCF of 36 is calculated following the procedure used previously for cadmium.
Using this value and an If of 0.053 kg/day in Equations 4-11 and 4-12, RWCf and
RWCwf are calculated to be 3.2 x 10"6 mg/2 and 1.6 x 10"6 mg/£, respectively.
4.4 COMPARISON OF CONTAMINANT CONCENTRATIONS (Ci and Xi) WITH REFERENCE
WATER CONCENTRATIONS (RWC)
Contaminant concentrations for the Surface Runoff and Groundwater Pathways
(mg/£) estimate the increase in contaminant water concentrations due to
emissions from MWC. The calculated concentration of contaminants in receiving
water (Ci) can be compared with the chronic RWC . The predicted leachate
w
concentration (Xi) for a contaminant entering an aquifer is also compared
directly to the RWC . The concentration of the contaminant in runoff water
after storm events may result in concentrations that exceed the long-term Ci
for brief periods. While this methodology is geared only toward examining
long-term effects, the event concentrations may be compared to existing
short-term assessment values such as 1-day Drinking Water Health Advisories
(see Federal Register, 1985, for example).
October 1986 4-20 DRAFT—DO NOT QUOTE OR CITE
-------
5. ECOLOGICAL EFFECTS FROM MWC EMISSIONS
5.1 TERRESTRIAL FOOD CHAIN MODEL
5.1.1 General Considerations
5.1.1.1 Most-Exposed Individuals (MEIs). For the ecological effects or
pathways of the terrestrial food chain model, the MEI is a plant or animal
rather than a human.
5.1.1.1.1 Toxicity to herbivorous animals. The pathways for exposure for
herbivorous animals are 1) deposition-soil-plant-animal toxicity, and 2)
deposition-soil-animal toxicity (direct ingestion). For these pathways, it
does not matter whether the animals are subsequently consumed by humans. The
end point of concern is toxicity to the animals themselves, which constitute
the MEI. It is assumed that wildlife may forage on lawns, gardens, agricul-
tural areas or forests within the region of maximal deposition of emissions of
the MWC.
5.1.1.1.2 Phytotoxicity. This pathway is described as deposition-soil-piant
toxicity. Toxic effects in plants are of concern since plants have an integral
role in the terrestrial food chain. The MEI, or vegetation type to be pro-
tected, will ordinarily be the most sensitive plant species for which data are
available.
5.1.1.1.3 Toxicity to soil biota or their predators. Two pathways are
considered here: 1) deposition-soil-soil biota toxicity, and 2) deposition-
soil-soil biota-predator toxicity. The term "soil biota" is intended to be
interpreted broadly. The first pathway examines effects on a broad range of
organisms including microorganisms, soil invertebrates such as earthworms, and
various anthropods living in or near the soil, as long as potential effects in
these organisms can .be related to soil concentrations. The second pathway
examines effects on predators of these organisms, especially small mammals and
birds. These predators could include insectivores as long as available data
permit the contaminant concentration in the prey to be related to contaminant
soil concentration.
October 1986 5-1 DRAFT—DO NOT QUOTE OR CITE
-------
5.1.1.2 Soil Deposition Rate of Contaminants. Cumulative soil deposition rate
of contaminants is determined in the same manner as described in Section
3.3.1.2.
5.1.1.3 Soil Incorporation of Deposited Contaminants. The considerations and
equations for soil incorporation of the deposited contaminants from MWCs are
the same for the ecological effects of the Terrestrial Food Chain Model as
those described in Section 3.3.1.3.
5.1.1.4 Contaminant Loss From Soils. The considerations are the same as those
enumerated in Section 3.3.1.4.
5.1.1.5 Contaminant Uptake Relationship. Uptake relationships for inorganic
and organic pollutants in both plants and animals are the same as in the ,
Terrestrial Food Chain Model.
5.1.1.6 Toxicity Thresholds for Nonhuman Organisms. When emissions from MWC
are deposited on land, organisms such as soil biota and their predators, plants
and grazing animals could be at risk as well as humans. Specific methods,
however, for selecting threshold levels for protecting these diverse groups
have not been articulated. It may be difficult to determine what studies are
most appropriate, what effects are of concern and how to select protective
values based on the available information.
The following general guidelines were suggested for determining toxicity
thresholds in plants, invertebrates and vertebrates (U.S. EPA, 1986h) and will
also be suggested here.
1. All inhibitory effects should be considered adverse unless evidence to the
contrary is available. These effects usually include reductions in
growth, fecundity, lifespan or performance, as well as symptomatic
manifestations of toxicity. For example, reductions in soil microbial
activity or diversity that can be attributed to a given chemical should be
considered adverse in lieu of information to the contrary. Where effects
cannot be attributed to one chemical, thresholds ordinarily cannot be
determined.
2. The geometric mean of exposure levels bracketing an adverse effect should
be used as the threshold. For example, if exposure levels are 1, IQ/and
100 and effects are significant at 100, a value of *30 ([10 x 100] ' )
should be used. Where effects occur at the lowest exposure level, and
other studies better defining the threshold value are unavailable, no
threshold can be determined. Where results were not statistically tested,
careful judgment should be used to determine the biological significance
of a given change.
3. The chemical form of a contaminant used in a study may not be equivalent
in bioavailability to the form present in the deposited emissions or
October 1986 5-2 DRAFT—DO NOT QUOTE OR CITE
-------
migrating from soil following deposition. Careful scientific judgment
should be used to determine toxicity thresholds of different chemical
forms.
4. Where studies with few species are available for a given chemical, the
results with the most sensitive species should be used to determine the
threshold. If many species have been studied, procedures for estimating
the fifth percentile of response may be used as described in the Aquatic
Life Guidelines (U.S. EPA, 1984e).
5. Where several tests have been conducted for a given species, results
appearing to be outliers should be disregarded.
5.1-2 Exposure Pathways for Toxicity to Herbivorous Animals
These two pathways, deposition-soil-plant-animal toxicity and deposition-
soil-animal toxicity (direct ingestion), are similar to the pathways for con-
taminant uptake by animal tissues consumed by humans, as discussed in Section
3.3.4; however, since toxicity to the animal itself is now the endpoint of con-
cern, the list of animals to be considered is broadened to include all herbi-
vores. Herbivorous rodents and birds should be considered, as well as large
herbivores and other domestic animals. The pathway for adherence of soils to
plant roots is limited to grazing animals that ingest significant amounts of
soil (cattle, sheep).
5.1.2.1 Assumptions. The assumptions pertaining to this pathway have been
stated in Tables 3-12 to 3-17.
5.1.2.2 Calculation Method. The first set of calculations for the
deposition-soil-plant-animal and deposition-soil-animal pathways are the same
as given in Section 3.3.4.2 for the determination of an animal feed
concentration (AFC, in ug/g DW) from the cumulative deposition (CD, for
inorganics) or soil concentration (LC, for organics and/or chemicals subject to
loss) and linear uptake response slope for crops (UC) [Equations (3-7, 3-8)].
The AFC is calculated for the deposition-soil-animal pathway from the soil
concentration (LC) and fraction of soil constituting the animal diet (FS) as
in Section 3.3.4.2.2. [Equation (3-10)].
The AFC is then compared to the reference feed concentration (RFC, in ug/g
DW) in a sensitive herbivore. Since AFC is actually the increase in feed
concentration resulting from MWC emissions, it is compared with the RFC, which
is defined as follows:
RFC = TA-BC Equation (5-1)
October 1986 5-3 DRAFT-DO NOT QUOTE OR CITE
-------
where:
TA = threshold feed concentration in a sensitive herbivore (ug/g DW)
BC = background concentration in feed crop (ug/g DW)
5.1.2.3 Input Parameter Requirements. The input parameters, LC and CD, were
defined in Section 3.3.1, UC was defined in Section 3.3.2.3, and FS was defined
in Section 3.3.4.3.3. The appropriate threshold feed concentration, TA, must
be identified. Procedures for selecting TA are discussed above in Section
5.1.6. An important source of relevant data for inorganic chemicals is found
in NRC (1980), "Mineral Tolerances of Domestic Animals." Values of TA are
needed for various types of herbivorous animals, including grazing animals,
many birds and many small mammals.
5.1.2.4 Example Calculations
5.1.2.4.1 Cadmium. The calculated AFC for cadmium for the uptake pathway,
as described in Section 3.3.4.4.1.1, is 0.46 and 1.52 ug/g for 30 and 100
years, respectively. The AFC for the adherence pathway for cadmium is 2.18 and
7o25 ug/g for 30 and 100 years, respectively, calculated as described in
Section 3.3.4.4.2.1.
For the uptake pathway (deposition-soil-plant-animal toxicity), a wider
variety of herbivores may be affected than for the adherence pathway
(deposition-soil-animal), which affects only grazing animals.
The calculated AFC is compared with the reference feed concentration,. RFC,
determined as the threshold concentration (TA) minus the background
concentration (BC). TA is calculated as a geometric mean of levels just
causing and not causing an adverse effect as described in Section 5.1.6. In a
48-week study with chickens, feed concentrations of 3 and 12 ug Cd/g (added as
CdSO.) showed no adverse effect and decreased eggshell thickness, respectively
1/2
(Leach et al., 1979). A geometric mean of (3 x 12) = 6 ug Cd/g is calculated,
and represents the TA for the uptake pathway. A BC of 0.1 ug Cd/g is typical
of corn grain. An RFC of 5.9 ug Cd/g is thus derived. The RFC is then
compared with the AFC for cadmium by this pathway.
For the adherence pathway, the available data for cadmium (U.S. EPA,
1985c) does not allow calculation of a TA.
5.1.2.4.2 Benzo(a)pyrene. The AFC for B(a)P for the uptake pathway is calcu-
lated as 4.96 x 10 ug/g for both 30 and 100 years, as described in Section
3.3.4.4.1.2. Carcinogenicity studies in mice indicate the TA for B(a)P to be
<40 ug/g (U.S. EPA, 1985d). The RFC for B(a)P is 40 ug/g (the TA) minus 0.005
.October 1986 5-4 DRAFT—DO NOT QUOTE OR CITE
-------
(the BC, using a plant/soil ratio of 0.05 and a background soil concentration
of 0.01 ug/g DW) (Connor, 1984; U.S. EPA, 1985d) = 39.995 ug/g.
The AFC for the adherence pathway for B(a)P calculated as described in
Section 3.3.4.4.2.2 is 2.36 x 10"2 ug/g for both 30 and 100 years. The paucity
of data for toxicity of B(a)P in grazing animals prohibits the calculation of a
TA for the adherence pathway.
5.1.3 Deposition-Soil-Plant Toxicity Exposure Pathway
5.1.3.1 Assumptions. The assumptions pertinent to this pathway have been
stated in Table 3-13.
5.1.3.2 Calculation Method. This pathway involves the direct comparison of
deposition rates or soil concentrations of contaminants deposited from MWC
emissions with a threshold phytotoxic application (deposition) rate or soil
concentration of the pollutant.
5.1.3.3 Input Parameter Requirements. The appropriate deposition rate or soil
concentration that corresponds to the threshold level for adverse effects in
plants must be selected. The threshold is generally defined in Section 5.1.6
as the geometric mean of the lowest exposure level causing, and the highest
level not causing, a significant adverse effect. The most sensitive species is
used for determination of the threshold, unless that species appears to be
unusally sensitive and is not found in the area of concern. Phytotoxicity of
metals may be altered by soil pH. Phytotoxicity data chosen should be
appropriate for the soil pH of the fallout region of the MWC, if possible.
5.1.3.4 Example Calculations. No calculations are required for this pathway,
except in some cases to convert CD to LC.
5.1.4 Exposure Pathways for Toxicity to Soil Biota and Their Predators
This section deals with two pathways: toxicity to soil biota
(deposition-soil-soil biota toxicity) and toxicity to predators of soil biota
(deposition-soil-soil biota-predator toxicity). As explained in Section
5.1.1.1.3, the term "soil biota" refers to a broad range of organisms including
microorganisms and various invertebrates living in or on the soil. Their
predators similarly include a variety of organisms. The availability of data
determines what species are considered.
5.1.4.1 Deposition-Soil-Soil Biota Toxicity Exposure Pathway. Procedures for
this pathway are identical to those described for phytotoxicity (Section 5.3),
October 1986 5-5 DRAFT-DO NOT QUOTE OR CITE
-------
except that the threshold levels pertain to effect thresholds in soil biota
rather than in plants. "Adverse" effects can be particularly difficult to
define where microorganisms are concerned. It is assumed that reductions in
soil microbial activity or diversity that can be attributed to the chemical
should be considered adverse in lieu of information to the contrary. The depo-
sition rate or soil concentration is then compared with the threshold concen-
tration in the soil biota.
5.1.4.1.1 Example calculations. This pathway does not require calculations,
except conversion of CD to LC.
5.1.4.2 Deposition-Soil-Soil Biota Predator Toxicity Exposure Pathway.
5.1.4.2.1 Assumptions. In addition to many of the assumptions listed in Table
5-1, some additional assumptions relevant to this pathway are stated in Table
5-2.
5.1.4.2.2 Calculation method. Calculations of criteria for this pathway may
take either of two forms, depending on the type of data available concerning
contaminant uptake by soil biota. If uptake response (increase in
concentration) in soil biota, UB, can be expressed in terms of a contaminant
deposition rate, the predator intake is calculated as follows:
RFC = CD x UB Equation (5-2)
where:
RFC = predator feed concentration (ug/g DW)
CD = cumulative soil deposition of pollutant (kg/ha)
UB = uptake response slope in soil biota (ug/g [kg/hg]"1)
If the chemical is not subject to degradation or loss in soil, PFC is a
cumulative concentration. If soil biota response is measured in terms of a
soil concentration, the following equation is used:
PFC = LCT x UB Equation (5-3)
where:
LCj = maximal soil concentration of pollutant at time, T (ug/g DW)
UB = uptake response slope in soil biota (ug/g [ug/g]"J).
LC represents cumulative soil concentration (above background) due to emis-
sions from MWC. The PFC is then compared with a reference feed concentration,
October 1986 5-6 DRAFT—DO NOT QUOTE OR CITE
-------
TABLE 5-1. ASSUMPTIONS FOR ECOLOGICAL EFFECTS FOR TERRESTRIAL FOOD CHAIN
Functional Area
Assumptions
Ranri fi cati ons/Lirai tati ons
Toxicity thresholds
for nonhuman organisms
All inhibitory effects should be
considered adverse.
The geometric mean of exposure
levels bracketing the appearance
of an adverse effect should be
used as the threshold.
The form of a contaminant used
in a study should not be con-
sidered equally bioavailable
and toxic as form in soil, unless
suitable data using soil are not
available.
Some "inhibitory" changes might not
significantly affect the individual
survival or population dynamics.
Conversely, some ecologically important
effects might not be observed in toxicity
tests.
The true threshold may lie at any point
between these two levels, and may be over-
or underpredicted by this method.
Availability of chemicals in soil or
deposited on soil may differ; particularly,
it may be lower, so that toxicity is over-
predicted.
01
-xl
-------
TABLE 5-2. ASSUMPTIONS 'FOR DEPOSITION-SOIL-SOIL BIOTA-PREDATOR TOXICITY EXPOSURE PATHWAY
Functional Area
Assumptions
Ramifications/Limitations
Contaminant uptake by
soil biota
Use of available data
to protect a variety
of species
Response in tissues of soil biota
can be represented by a linear
function.
It is assumed that use of the
highest available response slope
in soil biota and the lowest
available dietary threshold in
predators will result in protection
of untested species.
Probably oversimplifies a more
complex relationship.
This conservative assumption could
be overprotective of some species;
the extent to which it underprotects
others is unknown. A "match" of
data for a consumed organism and its
predator usually is not possible.
en
i
00
-------
RFC, for the predator (ug/g DW), determined as the threshold concentration
(TA) minus the background concentration (BB) in the soil biota.
5.1.4.2.3 Input parameter requirements.
5.1.4.2.3.1 Uptake response slope in soil biota (UB). UB may take any of
several forms, depending on the characteristics of the chemical and the data
available. UB is derived by linear regression of tissue concentration by
either contaminant deposition rate or soil concentration. Two or more points
are needed to derive the slope for inorganics, while for organic compounds a
single data pair can be used to derive a bioconcentration factor for plants and
animal tissues.
, The highest available uptake response slope will be used to estimate the
level of dietary contamination to which predators of soil biota would be
subject. This fact is important when evaluating data for earthworms. Some
studies distinguish between contaminant physiologically absorbed and that which
is due to gut contents or soil contamination of the sample. There is no need
to distinguish between absorbed and unabsorbed contaminants, since a predator
would ingest the gut contents as well as the rest of the organism. Whenever
possible, analyses used should be based on the whole organism.
5.1.4.2.3.2 Threshold feed concentration for the predator (TA). A wide
variety of organisms may prey on soil invertebrates, including birds and
insectivorous mammals. For a given chemical, information on subchronic or
chronic toxicity for oral administration may be available for only a few
species, and thus the value chosen for the threshold feed concentration may be
for a species not actually preying on soil biota. In general, the lowest
dietary adverse effect level and the no adverse effect level found for birds or
small mammals are used to determine the threshold. The threshold is calculated
as outlined in Section 5.1.5.
5.1.4.2.4 Example calculations.
5.1.4.2.4.1 Cadmiurn. The predator feed concentration (RFC) for cadmium
is calculated using Equation 5-2:
RFC = CD x UB
where:
CD = cumulative soil deposition of pollutant (kg/ha)
UB = uptake response slope in soil biota (ug/g[kg/ha] i)
October 1986 5-9 DRAFT-DO NOT QUOTE OR CITE
-------
CD is 3.26 and 10.88 kg/ha for 30 and 100 years, respectively, for
cadmium, and the UB is 13.7 ug Cd/g (kg/ha]"1 (U.S. EPA, 1985c). The RFC for
cadmium is:
For 30 years, RFC = 3.26 kg/ha x 13.7 ug/g (kg/ha)"1 = 44.66 ug/g
For 100 years, RFC = 10.88 kg/ha x 13.7 ug/g (kg/ha)"1 = 149.06 ug/g
The RFC is compared with the RFC for cadmium by this pathway. The RFC for
cadmium is 1.2 ug/g based on a TA of 6 ug/g (see Section 5.1.2.4.1) and a
background in soil biota of 4.8 ug/g (U.S. EPA, 1985c).
5.1.4.2.4.2 Benzo(a)pyrene. Lack of data availability "for UB of B(a)P
prohibits calculation of a RFC.
5.2. SURFACE RUNOFF AND GROUNDWATER MODELS
5.2.1 General Considerations
Toxic pollutants that end up in surface or groundwater can cause adverse
health or environmental effects in several ways:
1. Adverse effects on fish and other biota inhabiting streams, lakes and
estuaries. For example, if the concentration of a particular pollutant
exceeds a certain reference value, fish and other biota will either die or
experience some other adverse effects (reproductive effects, growth
retardation, etc.).
2. Adverse effects on wildlife consuming fish from polluted surface waters.
Some pollutants may accumulate in surface water biota (bioaccumulation).
Although these pollutants may not cause problems in fish and other water
biota, they may adversely affect animals consuming such polluted fish and
other biota.
5.2.1.1 Aquatic Life Protection. For protection of aquatic life from
long-term effects, the AWQC should be used (Federal Register, 1980). The AWQC
contains two concentrations, one that should not be exceeded at any time and
another that should not be exceeded, on an average, in a 24-hour period.
Criteria for acute exposures normally utilize tests of 96-hour duration or
less. For chemicals for which criteria are not available, the literature
should be evaluated to determine whether useful data have become available
since the AWQC was developed. The concentration increase of a pollutant in
surface water (Ci) or groundwater (X..) due to MWC should be added to the
October 1986 5-10 DRAFT—DO NOT QUOTE OR CITE
-------
existing background concentration and then compared with the AWQC in order to
evaluate whether risk to aquatic life exists.
5.2.1.2 Wildlife Protection. The comparison of Ci or X. (plus the background
concentration) with AWQC should also suffice to evaluate potential for effects
on wildlife, since the AWQC are designed to prevent chronic toxicity in wild-
life that consume aquatic organisms (U.S. EPA, 1984e); however, for many chemi-
cals, sufficient data on wildlife toxicity were not available for incorporation
in AWQC derivation. Therefore, toxicity thresholds applicable to wildlife spe-
cies that prey on aquatic organisms should be determined. Threshold feed con-
centrations (AFC in ug/g wet weight) in wildlife species
should be divided by the BCF (in £/kg) to determine a new estimate of AWQC. If
lower than the previous value, this number should be substituted for evaluation
of wildlife effects.
October 1986 5-11 DRAFT-DO NOT QUOTE OR CITE
-------
6. REFERENCES
Anonymous. (1985) Update: resource recovery activities report. Waste Age.
Banner, L. H.; Wilson, A. J.; Sheppard, J. M.; Parick, J. M.; Goodman, L R.;
Walsh, G. E. (1977) Kepone, bioconcentration, accumulation, loss and
transfer through estuarine food chains. Chesapeake Sci. 18: 297-308.
Ballschmidter, K. (1984) Distribution of polychlorodibenzodioxins and - furan
emissions between particulates, flue gas condensate and impinger adsorp-
tion in stack gas sampling. Presented at the international symposium on
chlorinated dioxins and related compounds; October; Ottawa, Canada.
Bellin, J.; Barnes, D. (1986) Procedures for estimating risks associated with
exposures to mixtures of chlorinated dibenzo-p-dioxins and dibenzofurans.
Prepared for Risk Assessment Forum, U.S. EPA, Washington, D.C.
Binder, S.; Sokal, D.; Maughan, D. (1985) Estimating the amount of soil ingest-
ed by young children through trace elements. Draft. Atlanta, GA: Centers
for Disease Control, Center for Environmental Health, Special Studies
Branch.
Bogert, L. J.; Biggs, G. M.; Calloway, D. H. (1973) Nutrition and physical
fitness, 9th ed. Philadelphia, PA: W. B. Saunders Co.; p. 578.
Bossert, I. D.; Bartha, R. (1986) Structure-biodegradability relationships of
polycyclic aromatic hydrocarbons in soil. Bull. Environ. Contam. Toxicol.
37: 490-495.
Briggs, C. A. (1971) Some recent analyses of plume rise observations. In:
Proceedings of the second international clean air congress. New York, NY:
Academic Press.
Briggs, G. A. (1975) Plume rise predictions. In: Lectures on air pollution and
environmental impact analysis. Boston, Massachusetts: American Meteorolog-
ical Society.
California Air Resources Board (CARB). (1984) Air pollution control at resource
recovery facilities.
Campbell, G. S. (1974) A simple method for determining unsaturated conductivity
from moisture retention data. Soil Sci. 117: 311-314.
October 1986 6-1 DRAFT-DO NOT QUOTE OR CITE
-------
Clayton, J. R.; Pavlou, S. P.; Breitner, N. F. (1977) Polychlorlnated biphen-
yls in coastal marine zooplankton: bioaccumulation by equilibrium parti-
tioning. Environ. Sci. technol. 11: 676-682.
Connor, M. S. (1984) Monitoring sludge-amended agricultural soils. Biocycle 25:
47-51.
Council for Agricultural Science and Technology (CAST). (1980) Effects of
sewage sludge on the cadmium and zinc content of crops. Washington, DC:
U.S. Environmental Protection Agency; EPA report no. EPA-600/8-81-003.
Czuczwa, J. M.; Hites, R. (1984) Environmental fate of combustion - generated
polychlorinated dioxins and furans. Environ. Sci. Technol. 20: 444-450.
Deutsch, W. J.; Krupka, K. M. (1985) MINTEQ geochemical code: compilation of
thermodynamic database for the aqueous species, gases, and solids con-
taining chromium, mercury, selenium, and thallium. Draft. Battelle
Memorial Institute.
Domalski, E. S.; Ledford, A. E.; Bruce, S. S.; Churney, K. L. (1986) The
chlorine content of municipal solid waste from Baltimore County, Maryland
and Brooklyn, New York. In: Proceedings of 1986 national waste processing
conference; June; Denver, CO. The American Society of Mechanical Engineers;
pp. 435-448.
Donahue, R. L.; Miller, R. W.; Shickluna, I. C. (1983) Soils, 5th ed. Englewood
Cliffs, NJ: Prentice Hall, Inc.
Dowdy, R. H.; Larson, W. E. (1975) The availability of sludge-borne metals to
various vegetable crops. J. Environ. Qua!. 4: 278-282.
Federal Register. (1980) Water quality criteria documents. Availability. F.R.
(November) 45: 79318-79379.
Federal Register. (1985) National primary drinking water regulations; synthetic
organic chemicals, inorganic chemicals and microorganisms, proposed rule.
F. R. (November 13) 50: 46936-47022.
Federal Register. (1986) Guidelines for carcinogen risk assessment. F. R.
(September 24) 51: 33992-34003.
Feldman, R. J.; Maibach, H. I. (1974) Percutaneous penetration of some pesti-
cides and herbicides in man. Toxicol. Appl. Pharmacol. 28: 126-132.
Felmy, A. R.; Brown, S. M.; Onishi, Y.; Yabusaki, S. B.; Argo, R. S. (1983)
MEXAMS — the metals exposure analysis modeling system. Washington, DC:
U.S. Environmental Protection Agency.
Felmy, A .R.; Girvin, D. C.; Jenne, E. A. (1984) MINTEQ — a computer program
for calculating aqueous geochemical equilibria. Washington, DC: U.S.
Environmental Protection Agency; EPA report no. EPA-600/3-84-032.
Food and Drug Administration (FDA). (1980a) FY77 Diet studies - infants and
toddlers (7320.74). FDA Bureau of Foods. October 22.
October 1986 6-2 DRAFT—DO NOT QUOTE OR CITE
-------
Food and Drug Administration (FDA). (1980b) FY77 total diet studies - adult
I/3ZO.73). FDA Bureau of Foods. December 11.
Food and Drug Administration (FDA). (1981) Documentation of the revised total
diet study food lists and diets. NTIS PB 82 192154. Springfield, VA.
Franklin, W. E.; Franklin, M. A.; Hunt, R. G. (1982) Waste-paper -- the future
of a resource, 1980-2000. Prepared for the Solid Waste Council of the
paper industry. Prairie Village, Kansas.
Frounfelker, R. (1979) A technical environmental and economic evaluation of
small modular incinerator systems with heat recovery. Prepared by Systems
Technology Corp. for U.S. EPA, Cincinnati, Ohio. EPA contract no.
68-01-3889.
Graedel, T. E.; Franey, J. P. (1977) Field measurements of sub-micron aerosol
washout by rain. In: Proceedings of the symposium on precipitation scav-
enging; 1974. ERDA Symp. Ser. 41: 503-523.
Hagernmaier, H. (1986) Preliminary letter report: air emissions testing at the
Martin GMBH waste-to-energy facility in Wurzburg, West Germany.
Unpublished.
Hahn, J. L. (1986) Air emissions testing at the Wurzburg, West Germany
waste-to-energy facility. Presented at the annual meeting of the Air
Pollution Control Association; June.
Haile, C. L.; Blair, R.; Lucas, R.; Walker, T. (1984) Assessment of emissions
of specific compound from a refuse fired waste-to-energy system. Prepared
by Midwest Research Institute for the U.S. EPA; EPA report no.
EPA-560/5-84-002.
Haith, D. A. (1980) A mathematical model for estimating pesticide losses in
runoff. J. Environ. Quality 9: 428-433.
Hart, Fred C. and Associates. (1984) Assessment of potential public health
impacts associated with predicted emissions of polychlorinated dibenzo-
dioxins and polychlorinated dibenzofurans from the Brooklyn navy yard.
resource recovery facility. Prepared for the New York City Department of
Sanitation.
Hawley, J. K. (1985) Assessment of health risk from exposure to contaminated
soil. Risk Analysis 5: 289-302.
Huber, A. H.; Snyder, W. H. (1976) Building wake effects on short stack efflu-
ents. In: Preprint volume for the third symposium on atmospheric diffusion
and air quality. Boston, Massachusetts: American Meteorological Society.
Huber, A. H. (1977) Incorporating building/terrain wake effects on stack
effluents. In: Preprint volume for the joint conference on applications of
air pollution meteorology. Boston, Massachusetts: American Meteorological
Society.
October 1986 6-3 DRAFT-DO NOT QUOTE OR CITE
-------
International Agency for Research on Cancer (IARC). (1983) Polynuclear aromatic
compounds. Part 1: Chemical, environmental, and experimental data. Lyon,
France: International Agency for Research on Cancer.
Kimbrough, R. D.; Falk, H.; Stehr, P.; Fries, G. (1984) Health implications of
2,3,7,8-tetrachlorodibenzodioxin (TCDD) contamination of residential soil.
J. Toxicol. Environ. Health 14: 47-93.
Klicius, R.; Hay, 0. J.; Finkelstein, A. (1986) The national incineration
testing and evaluation program. An assessment of A) two-stage incinera-
tion, B) pilot scale emission control. Presented for use of EPA's Science
Advisory Board by Environment Canada, Ottawa, Ontario.
Leach, R. M., Jr.; Wang, K. W. L; Baker, D. E. (1979) Cadmium and the food
cham: The effect of dietary cadmium on tissue composition in chicks and
laying hens. J. Nutr. 109: 437. (Cited in U.S. EPA, 1985c).
Lee, R. F.; Ryan, C.; Neuhausan, M. L. (1976) Fate of petroleum hydrocarbons
taken up from food and water by the blue crab, Callinectes sapidus. Mar.
Biol. 37: 363-370.
Lepow, M. L; Gillette, M.; Markowitz, S.; Robino, R.; Kapish, J. (1975)
Investigations into sources of lead in the environment of urban children.
Environ. Res. 10: 415-426.
Logan, T. J.; Chaney, R. L. (1983) Utilization of municipal wastewater and
sludge on land -- metals. In: Proceedings of the workshop on utilization
of municipal wastewater and sludge on land. Riverside, CA: University of
California; p. 235-323.
Lu, F. C. (1983) Toxicological evaluations of carcinogens and noncarcinogens.
Pros and cons of different approaches. Reg. Toxicol. Pharmacol. 3:
121-132.
Maibach, H. I.; Feldmann, R. J.; Milby, T. H.; Serat, W. F. (1971) Regional
variation in percutaneous penetration in man. Arch. Environ. Health 23:
208-211.
Marcelle, C.; Thome, J. P. (1984) Relative importance of dietary and environ-
mental sources of lindane in fish. Bull. Environ. Contam. Toxicol. 33:
423-429.
MEA, Inc. (1982) East Helena source apportionment study. Particulate source
apportionment using the chemical mass balance receptor model. Volume 1,
Draft Report. Prepared for the Department of Health and Environmental
Sciences, State of Montana.
Ministry of the Environment, Ontario, Canada. (1985) Scientific criteria
document for standard development, no. 4-84. Polychlorinated dibenzo-p-
dioxins (PCDDs) and polychlorinated dibenzofurans (PCDFs).
Morrey, J. R. (1985) PRODEF: A code to facilitate the use of the geochemical
code MINTEQ. Draft. Battelle Memorial Institute.
October 1986 6-4 DRAFT—DO NOT QUOTE OR CITE
-------
National Research Council (NRC). (1980) Mineral tolerances of domestic animals.
Washington, DC: Subcommittee on mineral toxicity in animals.
National Research Council (NRC). (1984) Mitigative techniques and analysis of
generic site conditions for groundwater contamination associated with
severe accidents. Battelle Memorial Institute. P. 3.33.
Neeley, W. B.; Bronson, D. R.; Blau, G. E. (1974) Partition coefficient to
measure bioconcentration potential of organic chemicals in fish. Environ.
Sci. Technol. 8: 1113-1115.
Nelson, W. E.; et al., Eds. (1969) Textbook of pediatrics, 9th ed. Philadelphia,
PA: W. B. Saunders Co. (Cited in Bogert et al., 1973).
New York State Department of Environmental Conservation (NYDEC). (1986) Emis-
sion source test report ~ preliminary test on Westchester RESCO
(Peekskill, N.Y.).
Norstrom, R. J.; McKinnon, A. E.; DeFrietas, A. S. W. (1976) A bioenergetics-
based model for pollutant accumulation by fish. Simulation of PCB
and methylmercury residue levels in Ohawa River yellow perch (Perca
flavescens). J. Fish. Res. Board Can. 33: 248-267.
0'Flaherty, E. J. (1981) Toxicants and drugs. Kinetics and dynamics. New York,
NY: John Wiley and Sons.
Page, A. L.; Logan, T. J.; Summers, L. E. (1986) Report of the workshop on
effects of sewage sludge quality and soil properties on plant uptake of
sludge-applied trace contaminants. November 1985; Las Vegas, NV. Draft.
Pennington, J. A. T. (1983) Revision of the total diet study food list and
diets. J. Am. Diet. Assoc. 82: 166-173.
Poiger, H.; Schlatter, C. (1980) Influence of solvents and adsorbents on dermal
and intestinal adsorption of TCDD. Food Cosmet. Toxicol. 18: 477-481.
Radke, L. F.; Hobbs, P. V.; Eltgroth, M. W. (1980) Scavenging of aerosol
particles by precipitation. J. Appl. Meteorol. 19: 715-722.
Rappe, C.; Ballschmidter, K. (1986) The chemistry of dioxins. Working paper
prepared for the World Health Organization working group on risks to
health of dioxins from incineration of sewage sludge and municipal waste,
March; Naples, Italy.
Roels, H. A.; Buchet, J. P.; Lauwerys, R. R. (1980) Exposure to lead by the
oral and pulmonary routes of children living in the vicinity of a primary
lead smelter. Environ. Res. 22: 81-94.
Ryan, J. A.; Pahren, H. R.; Lucas, J. B. (1982) Controlling cadmium in the
human food chain. A review and rationale based on health effects. Environ.
Res. 28: 251-302.
October 1986 6-5 DRAFT-DO NOT QUOTE OR CITE
-------
Scott Environmental Services. (1985) Sampling and analysis of chlorinated
organic emissions from the Hampton waste-to-energy system. Prepared for
the Bionetics Corporation.
Society of Actuaries (1959) Build and blood pressure study. (Cited in Bogert
et al., 1973).
Stephan, C. E. (1980) [Memorandum to J. F. Stara] Duluth, MM: U.S. Environmen-
tal Protection Agency, Environmental Research Center; July 30.
Swackhamer, D. L. (1986) Estimation of the atmospheric and nonatmospheric
contributions and losses of polychlorinated biphenyls for Lake Michigan on
the basis of sediment records of remote lakes. Environ. Sci. Techno!: 20:
879-883.
Telford, J. N.; Thonney, M. L.; Hogue, D. E.; et al. (1982) lexicological
studies in growing sheep fed silage corn cultured on municipal sludge-
amended acid subsoil. J. Toxicol. Environ. Health 10: 73-85. (Cited in
U.S. EPA, 1985c).
Thorton, I.; Abrams, P. (1983) Soil ingestion: a major pathway of heavy metals
into livestock grazing contaminated land. Sci. Total Environ. 28: 287-294.
Turner, D. B. (1970) Workbook of atmospheric dispersion estimates. Cincinnati,
Ohio: U.S. Department of Health, Education and Welfare, National Air
Pollution Control Administration; PHS publication no. 999-AP-26.
U.S. Department of Agriculture (USDA). (1966) Household food consumption
Survey, 1965-1966. Report 12. Food consumption of households in the United
States, seasons and year 1965-1966. Washington, DC: U.S. Government
Printing Office.
U.S. Department of Agriculture (USDA). (1975) Composition of foods.
Agricultural Handbook No. 8.
U.S. Department of Agriculture (USDA). (1984) Food consumption, price, and
expenditure 1963-1983. National Economic Division. Economic Research
Service. November.
U.S. Department of Agriculture (USDA). (1985) Food and nutrient intakes: indi-
viduals in four regions, 1977-1978. Report 1-3. Nationwide Food Consump-
tion Survey, 1977-1978. July.
U.S. Department of Commerce (1985). Fisheries of the United States, April 1985.
Current fisheries statistics no. 8360. Washington, DC.
U.S. Environmental Protection Agency. (1977) User's manual for single source
(CRSTER) model. Research Triangle Park, NC. EPA report no.
EPA-450/2-77-013.
U.S. Environmental Protection Agency. (1980a) Environmental assessment of a
waste-to-energy process. Braintree municipal incinerator. Prepared by
Midwest Research Institute for Industrial Environmental Research Labora-
tory; EPA report no. EPA-600/7-80-149.
October 1986 . 6-6 DRAFT—DO NOT QUOTE OR CITE
-------
U.S. Environmental Protection Agency. (1980b) Ambient water quality criteria
document for hexachlorocyclohexane. Cincinnati, OH: Environmental Criteria
and Assessment Office; Cincinnati, OH. EPA report no. EPA-440/5-80-054.
U.S. Environmental Protection Agency. (1980c) Ambient water quality criteria
for polynuclear aromatic hydrocarbons. Cincinnati OH: Environmental
Criteria and Assessment Office; EPA report no. EPA-440/5-80-069.
U.S. Environmental Protection Agency. (1980d) Seafood consumption data
analysis: final report. Prepared by SRI International, Menlo Park, CA,
under contract no. 68-01-3887. U.S. EPA, Washington, DC.
U.S. Environmental Protection Agency. (1982) Dermatotoxicity. Washington, DC:
Office of Pesticides and Toxic Substances; EPA report no.
EPA-560/11-82-002.
U.S. Environmental Protection Agency- (1983a) Comprehensive assessment of the
specific compounds present in combustion processes. Volume I. Pilot study
of combustion emissions. Prepared by Midwest Research Institute for the
Office of Toxic Substances; EPA report no. EPA-560/5-85-004.
U.S. Environmental Protection Agency. (1983b) Evaluation of HC1 and chlorinated
organic compound emissions from refuse-fired waste-to-energy systems.
Prepared by Scott Environmental Services for Environmental Sciences
Research Laboratory Research Triangle Park, NC.
U.S. Environmental Protection Agency. (1984a) Performance evaluation of
full-scale hazardous waste incinerators. Volume II. Prepared by Midwest
Research Institute for Incineration Research Branch, Cincinnati, Ohio. EPA
contract no. 68-02-3177.
U.S. Environmental Protection Agency. (1984b) Health effects assessment for
polycyclic aromatic hydrocarbons (PAHs). Cincinnati, OH: Environmental
Criteria and Assessment Office; EPA report no. EPA-540/1-86-013.
U.S. Environmental Protection Agency. (1984c) Air quality criteria for lead.
Research Triangle Park, NC: Environmental Criteria and Assessment Office;
EPA report no. EPA-600/8-83-0288.
U.S. Environmental Protection Agency. (1984d) Risk analysis of TCDD
contaminated soil. Washington, DC: Office of Health and Environmental
Assessment; EPA report no. EPA-600/8-84-031.
U.S. Environmental Protection Agency. (1984e) Guidelines for deriving numerical
national water quality criteria for the protection of aquatic organisms
and their uses. Available from NTIS, Springfield, VA; PB85-121101.
U.S. Environmental Protection Agency. (1985a) Municipal waste combustion study
data gathering phase. Preliminary Draft. Prepared by Radian Corp. for
Office of Air Quality Planning and Standards; EPA contract no. 68-02-3818.
U.S. Environmental Protection Agency. (1985b) Health assessment document for
polychlorinated dibenzo-p-dioxins. Office of Health and Environmental
Assessment; EPA report no. EPA-600/8-84/014F.
October 1986 6-7 DRAFT-DO NOT QUOTE OR CITE
-------
U.S. Environmental Protection Agency. (1985c) Environmental profiles and hazard
indices for constituents of municipal sludge: cadmium. Washington, DC:
Office of Water Regulations and Standards.
U.S. Environmental Protection Agency. (1985d) Environmental profiles and hazard
indices for constituents of municipal sludge: benzo(a)pyrene. Washington,
DC: Office of Water Regulations and Standards.
U.S. Environmental Protection Agency. (1985e) Technical support for development
of guidance on hydrogeologic criterion for hazardous waste management
facility location. Cincinnati, OH: Hazardous Waste Environmental Research
Laboratory. Draft.
U.S. Environmental Protection Agency. (1985f) Ambient water quality criteria
for cadmium — 1984. EPA report no. EPA-440/5-84/032.
U.S. Environmental Protection Agency. (1986a) Characterization of the municipal
waste combustion industry. Final report. Prepared by Radian Corp. for the
Office of Air Quality Planning and Standards; EPA contract no. 68-02-3889.
U.S. Environmental Protection Agency. (1986b) Characterization of stack emis-
sions from municipal refuse-to-energy systems. Prepared by Battelle
Columbus Laboratories for Atmospheric Sciences Research Laboratory; EPA
report no. EPA-600/3-86/055.
U.S. Environmental Protection Agency. (1986c) Engineering analysis report --
national dioxin study TIER 4-combustion sources. Draft report. Prepared by
Radian Corp. for Office of Air Quality Planning and Standards; EPA report
no. EPA-450/4-84-014h.
U.S. Environmental Protection Agency. (1986d) Development of risk assessment
methodology for municipal sludge incineration. Prepared for Office of
Water Regulations and Standards by the Environmental Criteria and Assess-
ment Office - Cincinnati, Ohio; ECAO-CIN-486. Final report.
U.S. Environmental Protection Agency. (1986e) Industrial source complex (ISC)
dispersion model user's guide-second edition. Office of Air Quality
Planning and Standards; EPA report no. EPA-450/4-86-005a.
U.S. Environmental Protection Agency. (1986f) User's manual for the human
exposure model (HEM). Office of Air Quality Planning and Standards; EPA
report no. EPA-450/5-86-001.
U.S. Environmental Protection Agency. (1986g) Environmental news release,
August 25, 1986 announcing the publication of risk assessment guidelines
to evaluate the public health risk of environmental pollutants.
U.S. Environmental Protection Agency. (1986h) Development of risk assessment
methodology for land application and distribution and marketing of munici-
pal sludge. Prepared by the Environmental Criteria and Assessment Office,
Cincinnati, OH for the Office of Water Regulations and Standards, Washing-
ton, DC. Draft Final'.
October 1986 6-8 DRAFT—DO NOT QUOTE OR CITE
-------
U.S. Environmental Protection Agency. (1986i) Development of risk assessment
methodology for municipal sludge landfill ing. Prepared by the
Environmental Criteria and Assessment Office, Cincinnati, OH for the
Office of Water Regulations and Standards, Washington, DC. Draft Final.
Van Genuchton, M. (1985) Concentive-dispersive transport of solutes involved in
sequential first-order decay reactions. J. Computers Geosci. 11: 129-147.
Veith, G. D.; Foe, D. L.; Bergstedt, B. V. (1979) Measuring and estimating the
bioconcentration factor of chemicals in fish. J. Fish. Res. Board Can. 36:
1040-1048.
Veith, G. D.; Macek, K. J.; Petrocelli, S. R.; Carroll, J. (1980) An evaluation
of using partition coefficients and water solubility to estimate BCFs for
organic chemicals in fish. In: Eaton, J. G.; Parrish, P. R.; Hendricks, A.
C., eds. Aquatic toxicology, pp. 116-129. ASTM STP 707.
Williams, F. R. (1975) Sediment and yield prediction with universal equation
using runoff energy factor. In present and prospective technology for
predicting sediment yields and sources. USDA ARS-S-40.
Wischmeier, W. H.; Smith, D. D. (1978) Predicting rainfall erosion losses — a
guide to conservation planning. USDA Handbook No. 537.
Wolf, M. A.; Dana, M. T. (1969) Experimental studies on precipitation scaveng-
ing. Battelle-Northwest annual report. USAEC report BNWL - 1051 (Part 1),
18-25.
Yoram, C. (1986) Organic pollutant transport. Improved multimedia modeling
techniques are the key to predicting the environmental fate of organic
pollutants. Environ. Sci. Techno!. 20: 538.
October 1986 6-9 DRAFT-DO NOT QUOTE OR CITE
-------
APPENDIX A
Organic and inorganic emissions from a mass burning MWC recently reported
in: U.S. EPA, 1986b.
October 1986 A"1 DRAFT-DO NOT QUOTE OR CITE
-------
TABLE A-l. ELEMENTAL COMPOSITION OF PARTICULATE
STACK EMISSIONS FROM MASS BURNING UNIT3
Element(b)
H+
L1
Be
8
O
N+
F
Na*
Mg*
Al*
Si*
P*
S*
£1*
K*
•Ca*
Sc
T1*
V
Cr*
Mn*
Fe*
Co
N1*
Cu*
Zn*
Sa
Ge
As*
Se
8r*
Rb
SP
Y
Zr
Nb
MO
Tes^t
ng/g
4000
10
0.05
30
254000
<1000
80
46800
4900
23700
53100
4900
22600
136000
60000
36200
<3
9200
40
500
900
7900
20
400
780
44000
30
10
300
<30
1700
100
70
4
60
6
20
No. 1
tig/dscm
2100
5
0.03
16 •
135000
<530
40
24800
2600
12600
28100
2600
12000
72100
31800-
19200
<2
48800
20
270
480
4190
11
210
410
23300
16
5
160
<16
900
53
37
2
32
3
11
Test
.H9/9
3000
10
0.03
30
191000
<1000
200
40500
6200
27700
64200
6200
27200
110800
69600
50400
<3
9900
100
450
980
8400
40
400
770
26600
30
10
130
<30
1300
70
50
2
30
4
20
No. 2
ng/dson
1490
5
0.01
15
94700
<500
100
20100
3080
13700
31800
3080
13500
55000
34500
25000
<1
4910
50
220
490
4170
20
200
380
13200
15
5
64
<15
640
35
25
1
15
2
10
Test
ng/9
3000-
, 20
0.03
60
235000
<1000
80
46100
5600
26500
65600
5900
20300
125000
55200
42400
<3
8200
40
560
850
9000
40
400
950
30000
15
10
710
. <30
1500
100
50
2
20
2
15
No. 3
tig/dscm
2000
13
0.02
40
157000
<670
50
30900
3570
17700
43900
• 3950
13600
83700
36900
28400
<2
5490
27
370
570
6020'
27
270
640
20100
10
7
480
<20
1000
67
33
1
13
1
10
A-2
-------
TABLE A-l. (continued)
El em
Ru
Rh
Pd
Ag
Cd*
In
Sn*
Sb*
Te
I
Cs
Ba*
La
Ce
Pr
Nd
Sm
Eu
Sd
Tb
Oy
Ho
Er
Tm
Yb
Lu
Hf
Ta
U
Re
Os
Ir
Pt
Au
Tl
Pb*
81
Th
U
(a)
(b)
.nt(b) J6St
entvu/ i*g/g ;
O.6
<8
<2
150
970
40
5000
880
O.6
20
5
2900
4
10
2
20
<1
O.4
O.6
0.2
O.6
0.2
O.4
O.2
<1.5
O.2
<3
O.4
O.6
O.4
O.6
O.4
O.6
O.2
1
20700
150
O.4
0.2
Reported as tig
cubic meter of
No. 1
WJ/dson
0.3
<3
<1
80
510
21
2650
470
0.3
11
2
1540
3
5
1
11
O.5-
O.2
O.3
O.I
O.3
0.1
O.2
0.1
O.8
0.1
<2
O.2
0.3*
O.2
O.3
0.2
O.3
0.1
0.5
11000
80
0.2
O.I
of element/gram of
stack gas
Test
w/g
0.6
<6
<2
100
780
40
4300
740
•O.5
15
5
980
4
10
2
10
<1
O.4
O.6
0.2
O.6
O.2
O.4
0.2
<1.5
O.2
<3
O.4
O.6
O.4
O.6
O.4
O.6
0.2
3
14400
40
O.4
0.2
fly ash
No. 2
(ig/dson
0.3
<3
<1 '
50
390
20
2130
370
0.3
7
2
490
2
5
1
5
O.5
O.2
O.3
0.1
O.3
0.1
O.2
O.I
0.7
O.I
<1
O.2
0.3
O.2
0.3
O.2
O.3
0.1
<1
7140
20
0.2
0.1
and i*g of
Test
M/9
0.6
<6
<2
' 40
910
20
3600
990
O.6
6
2
1300
2
2
1
10
O.4
O.6
O.2
O.6
O.2
O.4
O.2
<1.5
0.2
<3
O.4
O.6
O.4
0.6
O.4
O.6
0.2
1
16000
60
O.4-
O.2
No. 3
jig/dson
O.4
<4
27
610
13
2410
660
O.4
4
1
870
1
1
0.5
7
O.7
O.3
O.4
O.I
O.4
0.1
O.3
o.r
O.I
<2
O.3
O.4
O.3
O.4
0.3
O.4
O.L
0.5
10700
40
O.3
O.I
element/dry, standard
+ Elements determined by combustion with Perkln Elmer 240 Analyzer
* Elements determined by XRF
AIT other elements determined by spark source mass spectrometry.
A-3
-------
TABLE A-2. PAH CONCENTRATIONS IN STACK EMISSIONS FROM MASS BURNING UNIT
Concentration In Stack Emissions. ng/dscm(*)
Test No,
Compound
Naphthalene
Methyl Naphthalenes
D (methyl naphthal enes
Acenaphthene
Acenaphthylene
Fluorene
Phenanthrene/Anthracene
Methyl Phenanthrene/Anthracene
Fluor anthene
Pyrene
Benzo(a)anthracene/chrysene
Benzo(a)pyrene
Benzo(e)pyrene
Perylene
Benzo(b or k)f luoranthene
Indenol(l,2.3.-cd)pyrene
Benzol g,h. I tperylene
Part.
9000
NO
NO
NO
NO
NO
12000
NO
NO
NO
NO
NO
NO
NO
NO
NO
NO
Gas
390000
NO
NO
NO
NO
NO
150000
NO
83000
90000
4500
6800
7700
1600
NO
NO
NO
1 Test No. 2 Test No
Total Part.
399000
NO
NO
NO
NO
NO
162000
NO
83000
90000
4500
6800
7700
1600
NO
NO
NO
Gas Total Part. Gas
- 5600000
160000
NO
NO - j -
520000
NO - -
- 2400000
NO - -
1500000 • -r
1600000
52000 - *
18000
45000
22000
NO -
NO - -
NO * -
. 3
Total
2100000
8900
NO
NO
1900000
NO
410000
NO
210000
200000
5900
2300
4700
2800
NO
NO
NO
(a) All data are corrected for blank levels.
NO - Not Detected; estimated minimum detectable concentration Is 10 ng/dscm.
-------
TABLE A-3. ALDEHYDE EMISSION DATA
ui
Test
Unit
RDF 1
2
MASS 1
2
2
2
3
MOO 1
2
Concentration In Stack
No. Formaldehyde Ace t aldehyde Propanal
230(180)
3280(2630)
Imp 1) 550
Imp 2) 32
Total) 532(466)
1310(1050)
27(22)
11(9)
1070(580)
460(250)
1010(550)
130
10
280(150)
12(7)
5(3)
340(140)
54(22)
110(46)
13
<5
~T3(5)
25(10)
Gas. ug/dscm (ppb)
Aero le In Pentanal Benz aldehyde
' 45(19
110(47
!
1230(530) <6(<2)
30 <5
5 <5
ISjlSj <5!<2i
70(30) <6(<2)
<3(<1) <2(<1) <6(<2)
<3(<1) <2(<1) <6(<2)
1200(270)
3470.
10
3480(790)
230(52)
460(100)
810(180)
-------
TABLE A-4. ORGANIC COMPOUNDS IDENTIFIED IN
MASS EMISSION SAMPLES3
•Tentative Estimated Concentration In
Compound Identification^) . Stack Emissions,
Styrene 510
Ethyl benzene 260
Benzaldehyde ' 770
Propynyl benzene 260
Napthalene 2560
Isoqu1nol1ne 130
Benzoqulnone 130
2-methyl naphthalene 380
i-methyl naphthalene 380
THchlorophenol 80
31pheny1 800
Acenaphthalene 1520
1,4-naphthalenedlone 130
Olbenzofuran 960
Fluorene 510
9H-fluoren-9-one 130
Anthracene/phenanthrene 1280
1-phenyl napthalene 180
2-phenyl naphthalene 130
8enzo(c)c1nno11ne 800
Fluoranthene 1280
Pyrene 960
Benzo(g,h,1)fluoranthene 130
Chrysene/3enzo(a)anthracene 60
Benzo(e)pyrene 40
Benzo(a)pyrene 40
(a) These data provide tentative identities and concentrations of
organic compounds which may be present In the stack emissions.
The identity of the compounds was not-been confirmed.
Concentration data may be accurate only to a factor of ± 5.
(b) From analysis of combined partlculate and XAO-2 sample extracts.
A-6
-------
TABLE A-5. VOLATILE HYDROCARBON EMISSION DATA-MASS BURNING UNIT
Compound
Sample No. 1 Sample
No. 2 Sample No. 3
*
Concentration 1n Samples, uq/m3
Chloroform
1,2 Olchloroethane
Tetrachl oroethyl ene
o-D1chlorobenzene
Olchlorobenzene
Hexachloroethane
1,3,5-Trlchlorobenzene
1,2,4-Trlchlorobenzene
Chlorobenzene
44
9,287 9,
1,120
54
34
30
6
38
47
163 248
572 . 19,000
861 856
41 20
49 319
27 45
6 5
41 31
58 156
Concentration in Samples, ppb C
Isobutane
n-pentane
Benzene
Toluene
Benzaldehyde
p-ethyl toluene
1,2, 4-Tr Imethy 1 benzene
Nap thai ene
m+p xylene
263
625 . 1,
11,735 16,
66
128
<50
70
279
<50
584 160
030 1,362
039 36,831
92 225
591 277
152 184
136 229
663 1,169
<50 304
Total volatile hydrocarbons 74,000
93,000
54,000
A-7
-------
APPENDIX B
CALCULATION OF PARTICLE SURFACE AREA DISTRIBUTION FOR DEPOSITION MODELING
The following represents an example of the calculation of available
surface area for adsorption of a pollutant for a given particle size. The
method of calculation follows the procedure described by Hart (1986) in a
report of the emission impacts of a proposed MWC in New York City. That report
has been peer reviewed by a panel of local, national, and international author-
ities in both science and engineering.
(a) Assume aerodynamic spherical particles
(b) Specific surface area of a spherical particle with radius, r:
S = 4 n r2
(c) Volume of spherical particle with radius, r:
V = 4/3 n r3
(d) Then the ratio of surface area to volume is:
S/v = 4 n r2/(4/3 0 r3) = 3/r
If particle density is held constant, then it can be assumed that particle
weight is proportional to particle volume. Hart (1986) further postulates that
the ratio of surface area to volume is proportional to the ratio of surface
area to weight for a particle with a given radius. Therefore, the ratio of
surface area to volume represents the potential relationship between the
surface area and the weight of the particle. Multiplying the ratio of the
surface area to volume calculation by the percent weight fraction of particles
emitted in a given size category (microns) should approximate the amount of
surface area available for adsorption in that particle size category. When
these calculations are summed for all particle size categories, total surface
area is assumed for total particle emissions. Dividing the surface area for
each particle category by the total available surface area for all particles
gives an estimation of the fraction of total area on any size particle. If the
October 1986 B-l DRAFT-DO NOT QUOTE OR CITE
-------
emission rate of a pollutant in grains per second is known, then the multiplica-
tion of the emission rate times the fraction of available surface area will
determine the emission rate of the pollutant per particle size.
Example
(1) Assume 15 micron size particles with a radius of 7.5 microns.
Then the ratio of surface area to volume is:
S/A = 3
775
S/A = 0.4
(2) Then multiplying the S/A ratio by the fraction of total weight
corresponding to 15 micron size particles equals the relative propor-
tion of total surface area for that given size.
Given: Fraction of weight of 15 micron particles = 12.8 percent.
Then, S/A ratio x 0.128 = proportion of total surface area.
0.4 x 0.128 = 0.512
(3) If the sum of the computed relative proportion of total surface area
for all particles sizes emitted is 3.4423, then the fraction of total
surface area comprised of 15 micron particles is:
0.512/3.4423 = 1.49%
(4) If the emission rate of a pollutant is 10 mg/second, then the emis-
sion rate for the pollutant adsorbed to 15 urn particles is:
10 mg/second times 0.0149 = 0.149 mg/sec.
The fraction of total surface area was computed in the same manner for all
particle diameters in Table 3-4 of the report. For convenience in deposition
modeling, three particle size categories were chosen from Table 3-4: greater
than 10 microns; 2 to 10 microns; less than 2 microns. The fraction of total
surface areas for these ranges were summed with each particle size category to
represent a single fraction of total surface area for the given particle size
category, e.g., 0.03 for > 10 microns; 0.095 for 2 to 10 microns; and 0.875 for
less than 2 microns.
October 1986 B-2 DRAFT—DO NOT QUOTE OR CITE
------- |