RCRA Delisting Technical Support Document 

Chapter 2: Estimation of Chemical Releases and Media Concentrations Revised October 2008 

Chapter 2
 

Estimation of Chemical Releases and Media
 

Concentrations 


What’s covered in Chapter 2: 

� Source Release Systems 

� Releases to Groundwater and Estimation of Groundwater Waste Concentrations 

� Releases to Surface Pathways and Estimation of Waste Concentrations in Receiving Media 

This chapter describes the methodology and assumptions used in the Delisting Risk Assessment Software 

(DRAS) program to compute releases of waste constituents from petitioned wastes and estimate waste 

constituent concentrations in media at the POEs. Source release systems are discussed in Section 2.1, 

releases to groundwater are discussed in Section 2.2, and releases to surface pathways are discussed in 

Section 2.3. 

2.1 SOURCE RELEASE SYSTEMS 

The risk-based delisting process, as incorporated in DRAS, involves performing a risk assessment for 

petitioned wastes that are disposed of in two waste management units of concern to the U.S. EPA 

Delisting Program: surface impoundments and landfills. The process determines whether a waste that is 

petitioned for an exclusion (delisting) is not characteristically toxic and is thus exempt from Subtitle C 

disposal requirements, assuming that the petitioned waste meets all other criteria for delisting (see U.S. 

EPA 1996e for other delisting criteria). Once delisted, the petitioned waste may be disposed of in any 

municipal or industrial nonhazardous waste Subtitle D disposal unit. In preparing DRAS, U.S. EPA 

Region 6 focused its delisting petition evaluations on liquid and solid wastes disposed of in surface 

impoundments and landfills, respectively. Wastes disposed of in other waste management units will be 

evaluated on a case-by-case basis. 

In DRAS, U.S. EPA Region 6 assumed that petitioned liquid wastes will be disposed of in surface 

impoundments and that petitioned solid wastes will be disposed of in landfills. Whether a waste is liquid 

or solid is determined using methods specified in U.S. EPA’s “Test Methods for Evaluating Solid Waste” 

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 (SW-846) (U.S. EPA 1997g). U.S. EPA’s SW-846 also is available on the Internet at 

“http://www.epa.gov/epaoswer/hazwaste/test/main.htm”. The assumptions used to quantify releases of 

chemicals from liquid-phase wastes in surface impoundments and solid-phase wastes in landfills are 

described in Sections 2.1.1 and 2.1.2, respectively. 

2.1.1 Liquid-Phase Waste (Surface Impoundment) 

The method used to compute releases from liquid-phase wastes assumes that liquid industrial wastes are 

disposed of in an unlined surface impoundment with a sludge or sediment layer at the base of the 

impoundment. The determination of whether a waste is a liquid waste is made using U.S. EPA-approved 

Test Method 9095, referred to as the Paint Filter Test (U.S. EPA 1997g). The four parameters used to 

characterize a surface impoundment are (1) the area of the impoundment, (2) the ponding depth of the 

liquid in the impoundment, (3) the thickness of a relatively low-permeability sludge or sediment layer at 

the base of the surface impoundment, and (4) the hydraulic conductivity of this sludge or sediment layer. 

Additional information regarding characterization and modeling of liquid wastes disposed of in surface 

impoundments is provided in the “EPACMTP Technical Background Document” and the “EPACMTP 

Parameters/ Data Background Document” (U.S. EPA 2003a and 2003b). 

2.1.2 Solid-Phase Waste (Landfill) 

The method used to compute releases from solid-phase wastes assumes that solid wastes are disposed of 

in a Subtitle D landfill and are covered with a 2-foot-thick native soil layer. It is assumed that the Subtitle 

D landfill is unlined or that any liner at the base of the landfill will eventually completely fail. The two 

parameters used to characterize landfills are (1) area and (2) depth (the thickness of the waste layer). Data 

to characterize landfills were obtained from a nationwide survey of industrial Subtitle D landfills (Westat 

1987 and U.S. EPA 1997h). Parameters and assumptions used to estimate infiltration of leachate from a 

landfill are provided in the “EPACMTP Technical Background Document” and the “EPACMTP 

Parameters/ Data Background Document” (U.S. EPA 2003a and 2003b). 

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2.2	 RELEASES TO GROUNDWATER (EPACMTP) 

This section describes the method used to compute the release and transport of chemicals from a waste 

management unit to the subsurface and their subsequent transport through the unsaturated and saturated 

zones to a theoretical downgradient receptor well. 

2.2.1	 Methodology for Estimation of Waste Constituent Concentration in Groundwater 

This section summarizes the method used to calculate groundwater exposure concentrations resulting 

from release of waste constituents into the subsurface from two waste management units: surface 

impoundments and landfills. The exposure concentration is evaluated at a hypothetical groundwaterdrinking water well located a specific distance from the downgradient edge of the waste management 

unit. This well is referred to hereafter as the receptor well, and the exposure concentration measured at 

that well is referred to as the groundwater receptor well concentration (Cgw). The groundwater fate and 

transport model used was the EPACMTP (U.S. EPA 2003a). Receptor well concentrations for both 

carcinogens and noncarcinogens, and for both degraders and nondegraders were determined using the 

finite source option within the EPACMTP. 

2.2.2	 Overview of EPACMTP 

The EPACMTP is a fate and transport model that simulates one-dimensional, vertically downward flow 

and transport of contaminants in the unsaturated zone beneath a waste disposal unit as well as twodimensional or three-dimensional groundwater flow and contaminant transport in the underlying saturated 

zone. The model accounts for the following processes affecting contaminant fate and transport: 

advection, hydrodynamic dispersion, linear or nonlinear equilibrium sorption, chained first-order decay 

reactions, and dilution from recharge in the saturated zone. The EPACMTP incorporates a Monte Carlo 

module that allows assessment of the uncertainty associated with receptor well concentrations that result 

from both uncertainty and variability in the model input parameter values. 

The EPACMTP consists of four major components: 

•	 A module that performs one-dimensional analytical and numerical solutions for water 

flow and contaminant transport in the unsaturated zone beneath a waste management unit 

•	 A numerical module for steady-state groundwater flow subject to recharge from the 

unsaturated zone 

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•	 A module of analytical and numerical solutions for contaminant transport in the saturated 

zone 

•	 A Monte Carlo module for assessing the effect of the uncertainty resulting from 

uncertainty and variability in model parameter values on predicted receptor well 

concentrations 

The subsurface as modeled with the EPACMTP consists of an unsaturated (vadose) zone beneath a waste 

unit and an underlying water table aquifer. Contaminants move vertically downward through the 

unsaturated zone to the water table. The EPACMTP allows simulation of flow and transport in the 

unsaturated zone and in the saturated zone, either separately or combined. However, for the purposes of 

the delisting analysis, both the vadose zone and saturated zone were modeled. 

The EPACMTP is capable of simulating the fate and transport of dissolved contaminants from a point of 

release at the base of a waste disposal unit, through the unsaturated zone and underlying groundwater, to a 

receptor well at an arbitrary downstream location in the aquifer. The model accounts for the following 

mechanisms affecting contaminant migration: transport by advection and dispersion, retardation resulting 

from reversible linear or nonlinear equilibrium adsorption onto the soil and aquifer solid phase, and 

biochemical degradation processes. The latter may involve chain decay reactions if the contaminant or 

contaminants of concern form toxic daughter products that are of concern as well. As is true of any 

model, the EPACMTP is based on a number of simplifying assumptions that make the model easier to use 

and that ensure its computational efficiency. The major simplifying assumptions used in the EPACMTP 

are summarized below. 

1.	 Soil and Aquifer Medium Properties. It is assumed that the soil and aquifer are uniform, 

porous media and that flow and transport are described by Darcy’s law and the advectiondispersion equation, respectively. The EPACMTP does not account for the presence of 

preferential pathways such as fractures and macropores. Although the aquifer properties are 

assumed to be uniform, the model does allow for anisotropy in hydraulic conductivity. 

2.	 Flow in the Unsaturated Zone. Flow in the unsaturated zone is assumed to be steady-state, onedimensional, vertical flow from beneath the source toward the water table. The lower boundary 

of the unsaturated zone is assumed to be the water table. The flow in the unsaturated zone is 

assumed to be predominantly gravity-driven, and, therefore, the vertical flow component accounts 

for most of the fluid flux between the source and the water table. The flow rate is assumed to be 

determined by the long-term average infiltration rate through the waste management unit. In 

surface impoundments, this infiltration rate is assumed to be determined by the average depth of 

ponding. 

3.	 Flow in the Saturated Zone. The saturated zone module of the EPACMTP is designed to 

simulate flow in an unconfined aquifer with constant saturated thickness. The model assumes 

regional flow in a horizontal direction with vertical disturbance resulting from recharge and 

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infiltration from the overlying unsaturated zone and waste disposal facility, respectively. The 

lower boundary of the aquifer is assumed to be impermeable. Flow in the saturated zone is 

assumed to be steady-state. The EPACMTP accounts for different recharge rates beneath and 

outside the source area. Groundwater mounding beneath the source is represented in the flow 

system by increased head values at the top of the aquifer. This approach is reasonable as long as 

the height of the mound is small relative to the thickness of the saturated zone. 

4.	 Transport in the Unsaturated Zone. Contaminant transport in the unsaturated zone is assumed 

to occur by advection and dispersion. The unsaturated zone is assumed to be initially 

contaminant-free, and contaminants are assumed to migrate vertically downward from the 

disposal facility. The EPACMTP can simulate both steady-state and transient transport in the 

unsaturated zone with single-species or multiple-species chain decay reactions and with linear or 

nonlinear sorption. 

5.	 Transport in the Saturated Zone. Contaminant transport in the saturated zone is assumed to be 

a result of advection and dispersion. The aquifer is assumed to be initially contaminant-free, and 

contaminants are assumed to enter the aquifer only from the unsaturated zone immediately 

beneath the waste disposal facility, which is modeled as a rectangular, horizontal plane source. 

The EPACMTP can simulate both steady-state and transient three-dimensional transport in the 

aquifer. For steady-state transport, the contaminant mass flux entering at the water table must be 

constant with time; for the transient case, the flux at the water table may be constant or may vary 

as a function of time. The EPACMTP can simulate the transport of a single species or multiple 

species, chain decay reactions, and linear sorption. 

6.	 Contaminant Phases. The EPACMTP assumes that the dissolved phase is the only mobile 

phase and disregards interphase mass transfer processes other than adsorption onto the solid 

phase. The model does not account for volatilization in the unsaturated zone; this is a 

conservative approach for volatile chemicals. The model also does not account for the presence 

of a nonaqueous-phase liquid (such as oil) or for transport in the gas phase. When a mobile oil 

phase is present, significant contaminant migration may occur within it, and the EPACMTP may 

underestimate the movement of hydrophobic chemicals. 

7.	 Adsorption Reactions. The EPACMTP computes chemical reactions involving adsorption. The 

EPACMTP assumes that sorption of organic compounds in the subsurface can be represented by 

linear adsorption isotherms in both the unsaturated and saturated zones. It is assumed that 

adsorption of contaminants onto the soil or aquifer solid phase occurs instantaneously and is 

entirely reversible. The effect of geochemical interactions is especially important in fate and 

transport analyses of metals. For simulation of metals, the EPACMTP uses one of two types of 

adsorption isotherms: 1) pH-dependent, empirically-derived isotherms (U.S. EPA 1990g); or 2) 

sorption isotherms generated by MINTEQA2 (Allison and others 1991). MINTEQA2 generates 

concentration-dependent effective partition coefficients for various combinations of geochemical 

conditions. This procedure is described in the background document for modeling of metal 

transport (U.S. EPA 1991d, 1996d, and 1999). 

The EPACMTP also accounts for chemical and biological transformation processes. All transformation 

reactions are represented by first-order decay processes. An overall decay rate is specified for the model; 

therefore, the model cannot explicitly consider the separate effects of multiple degradation processes such 

as oxidation, hydrolysis, and biodegradation. The user must determine the overall, effective decay rate 

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when multiple decay processes are to be represented. To maximize its flexibility, the EPACMTP has the 

capability of determining the overall decay rate from chemical-specific hydrolysis rate constants using 

soil and aquifer temperature and pH values (U.S. EPA, 1993e and 1996c). The EPACMTP assumes that 

reaction stoichiometry is prescribed for scenarios involving chain decay reactions. The speciation factors 

are specified as constants by the user (see the “EPACMTP Technical Background Document” and the 

“EPACMTP Parameters/ Data Background Document” (U.S. EPA 2003a and 2003b)). In reality, these 

coefficients may change as functions of aquifer conditions (for example, temperature and pH), 

concentration levels of other chemical components, or both. 

2.2.2.1 Contaminant Release and Transport Scenario 

Two source release scenarios are considered in the EPACMTP: continuous (infinite) and finite-source. 

Only the finite-source scenario is considered for delisting. For finite-source scenarios, the release of 

contaminants occurs over a finite period of time, after which the leachate concentration becomes zero 

(that is, all the contaminants in the waste disposed of in the waste management unit have leached out). 

Each type of waste management scenario is described by a relatively small number of parameters. The 

differences between waste management units are represented by different values or frequency 

distributions of the source-specific parameters. Source-specific stochastic parameters used by the 

EPACMTP for landfills include the capacity and dimensions of the waste management unit, infiltration 

and recharge rates, pulse duration, the fraction of hazardous waste in the waste management unit, and the 

density of the waste. The source-specific stochastic parameters used for surface impoundments include 

the area, the ponding depth (such as the depth of liquid in the impoundment), and the thickness and 

hydraulic conductivity of the sludge or sediment layer at the bottom of the impoundment. Data on the 

areas, volumes, and locations of landfills were obtained from the 1987 U.S. EPA survey of industrial 

Subtitle D waste facilities in the United States (Westat 1987 and U.S. EPA 1997h). Relevant information 

for surface impoundments was obtained from the 2001 EPA’s industrial Surface Impoundments Study 

(US EPA 2001). Derivation of the parameters for each type of waste management unit is described in the 

“EPACMTP Technical Background Document” and the “EPACMTP Parameters/ Data Background 

Document” (U.S. EPA 2003a and 2003b). 

For finite-source scenarios, simulations are performed for transient conditions, and the source is assumed 

to be a pulse of finite duration. In the case of landfills, the pulse duration is based on the initial mass of 

contaminant in the landfill, infiltration rate, landfill dimensions, and leachate concentration (U.S. EPA 

1996c). For surface impoundments, the duration of the leaching period is determined by the waste 

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management unit’s lifetime (the default value is 50 years for surface impoundments). For a finite-source 

scenario, the model can calculate either the peak receptor well concentration for noncarcinogens or the 

highest average concentration over a specified period for carcinogens. The finite-source methodology in 

the EPACMTP is discussed in detail in the finite-source background document (U.S. EPA 1996c). 

2.2.2.2 EPACMTP Modeling Assumptions and Input Parameters 

Specific EPACMTP modeling assumptions (in addition to the simplifying assumptions discussed in 

Section 2.2.1) are summarized in Table 2-1. This table also provides information on important input 

parameters as well as on their data sources. Overall, EPACMTP input parameters can be organized in the 

following four groups: 

C Source-specific parameters 

C Chemical-specific parameters 

C Unsaturated zone-specific parameters 

C Saturated zone-specific parameters 

For delisting, the EPACMTP is run in probabilistic (Monte Carlo) mode, and many of the source-, 

chemical-, unsaturated zone-, and saturated-zone specific parameters are represented by probability 

distributions reflecting variations on a national or a regional level. Specific capabilities and requirements 

associated with running the EPACMTP in the Monte Carlo mode are presented in Chapter 5 of the 

“EPACMTP Technical Background Document” (U.S. EPA 2003a). The Monte Carlo analysis enables 

parametric uncertainty and/or variability to be quantitatively accounted for. The flow and transport 

modules of EPACMTP are linked to a Monte Carlo driver that permits a probabilistic evaluation of 

variability and/or uncertainty in model input parameters, as described by specified (joint) probability 

distributions. The resulting sequence of receptor well concentrations (one for each iteration) are sorted 

and ranked from highest to lowest in order to obtain a probabilistic distribution of receptor well 

concentrations. The different groups of input parameters are summarized in Table 2-1 below. 

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TABLE 2-1 

 EPACMTP MODELING ASSUMPTIONS AND INPUT PARAMETERS 

Modeling Assumptions 

Modeling Element Description or Value 

Management Scenario Landfill 

Surface impoundment 

Modeling Scenario Finite-source Monte Carlo; 

LF: depleting source for organics, 

constant concentration pulse source for metals 

SI: constant concentration pulse for all constituents 

Exposure Evaluation Downgradient groundwater receptor well; maximum well concentration; 10,000-year 

modeling period 

Regulatory Protection Level 90 percent 

Source-Specific Parameters 

Parameter Description or Value 

Waste Unit Depth 

Waste Unit Area 

Waste Unit Volume 

Infiltration Rate 

Landfill 

Surface Impoundment 

Derived (for landfills), and site-based randomly selected from the EPA industrial SI data base 

for (surface impoundments) 

Site-based, randomly selected from OPPI database (for landfills), and from the EPA industrial 

SI data base (for surface impoundments) 

User-specified 

Site-based, derived from water balance using HELP model 

Site-based, derived from impoundment depth using Darcy’s law 

Leaching Duration 

Landfill 

Surface Impoundment 

Derived, continues until all constituents have leached out 

50 years (operational life of unit) 

Chemical-Specific Parameters 

Parameter Description and Source 

Decay Rate 

Organic Constituents 

Metals 

Hydrolysis rate constants compiled by U.S. EPA ORD (U.S. EPA 1993e) 

No decay 

Sorption 

Organic Constituents 

Metals 

Koc constants compiled by U.S. EPA ORD (U.S. EPA 1993e) 

MINTEQA2 sorption isotherm coefficients (K d) for Ba, Cd, Cr (III), Hg, Ni, Pb, Ag, Zn, Cu, 

V, Be, Mo, As(III), Cr(VI), Se(VI), Tl, Sb(V), Mn, F, As(V), and Se(IV) (U.S. EPA 2003b); 

d empirical K distributions for Fe and Sn 

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Unsaturated Zone-Specific Parameters 

Parameter Description and Source 

Depth to Groundwater Site-based, from API and USGS hydrogeologic database (Newell and others 1989 and 1990) 

Soil and Unsaturated Zone U.S. EPA ORD data, based on national distribution of three soil types (sandy loam, silt loam, 

Properties: silty clay loam) 

Bulk Density Carsel and others 1992 

Saturated Conductivity Carsel and Parrish 1992 

Moisture Retention Carsel and Parrish 1992 

Parameters (alpha and 

beta) 

Residual Water Content Carsel and Parrish 1992 

Saturated Water Content Carsel and Parrish 1992 

Percent Organic Matter Carsel and others 1992 

Dispersivity Gelhar’s analysis of EPRI field data (US EPA 2003b and EPRI 1985) 

Saturated Zone-Specific Parameters 

Parameter Description and Source 

Recharge Rate Site-based, derived from regional precipitation and evaporation data and soil type 

Aquifer Thickness Site-based, from API and USGS hydrogeologic database (Newell and others 1989 and 1990) 

Hydraulic Conductivity Site-based, from API and USGS hydrogeologic database (Newell and others 1989 and 1990) 

Hydraulic Gradient Site-based, from API and USGS hydrogeologic database (Newell and others 1989 and 1990) 

Porosity Effective porosity derived from national distribution of aquifer particle diameter 

Bulk Density Derived from porosity 

Dispersivity Derived from distance to receptor well (Gelhar and others 1992 and U.S. EPA 2003b) 

Groundwater Temperature Site-based, from USGS regional temperature map (Todd 1980 and US EPA 2003b) 

Fraction Organic Carbon National distribution, from U.S. EPA STORET database 

pH National distribution, from U.S. EPA STORET database 

Receptor Well Parameters 

Well Element Description and Source 

Radial Distance from Waste Nationwide distribution, from U.S. EPA OSW database (US EPA 1993) 

Management Unit 

Angle Off-Center Uniform within ± 90E from plume center line (no restriction to always lie within plume) 

Depth of Intake Point Uniform throughout saturated thickness of aquifer 

Notes: 

API = American Petroleum Institute 

HELP = Hydrologic Evaluation of Landfill Performance 

ORD = U.S. EPA Office of Research and Development 

STORET = Database Utility for STORage and RETrieval of Chemical, Physical, and Biological Data for 

Water Quality 

USGS = U.S. Geological Survey 

Source-Specific Parameters. Source-specific parameters give information about the waste management 

unit. The most sensitive parameters in this group include the waste management unit’s area and depth 

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and the infiltration rate (U.S. EPA 1996b and 1996c). It is important to note the difference in leaching 

duration for each waste management unit type (see Table 2-1). 

Chemical-Specific Parameters. Chemical-specific parameters describe the degradation (decay), 

adsorptive, and diffusive characteristics of each of the chemical species being simulated. Thus, the most 

important variables accounting for chemical characterization are the hydrolysis rate (ë) (for organic 

constituents), the normalized distribution coefficient for organic carbon (Koc) (for sorption estimation for 

organics), and the adsorption isotherm coefficient (Kd) (for linear and nonlinear sorption estimation for 

metals). Koc is also referred to as the organic carbon partition coefficient, and Kd is also referred to as the 

effective partition coefficient. For details on metals isotherms from MINTEQA2, see Appendix A-6, 

Application of CMTP and MINTEQA2 for determining DAFs. 

Unsaturated Zone-Specific Parameters. The unsaturated zone is the first subsurface level below the 

waste management unit. It is unsaturated with groundwater, and, therefore, it does not have a water table. 

The available soil moisture in this zone enables contaminants to move in the subsurface. The simulated 

process of fate and transport in this zone is based on the following assumptions: 

•	 The flow of the fluid phase is isothermal and is governed by Darcy’s law. 

•	 The flow is one-dimensional, vertically downward, and steady-state. 

•	 The fluid is slightly compressible and homogeneous. 

•	 The dynamics of a second phase (such as a vapor phase or nonaqueous liquid) can be 

disregarded. 

The solutions for transient and steady-state transport in the unsaturated zone are based on the following 

assumptions: 

•	 The leachate concentration entering the soil is either constant (with a finite or infinite 

duration) or decreasing with time following a first-order decay process. 

•	 Sorption of contaminants onto the soil solid phase is described by a linear or nonlinear 

(Freundlich) equilibrium isotherm. 

The required input values for this parameter group include depth to groundwater and soil hydraulic 

parameters and other soil properties, including fraction organic carbon (foc) and bulk density. 

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Saturated Zone-Specific Parameters. The saturated zone is located below the unsaturated zone and 

beneath the water table. Groundwater flow in the saturated zone is simulated using a steady-state solution 

for predicting hydraulic head and Darcy velocities in a constant-thickness groundwater system subject to 

infiltration and recharge along the top of the aquifer and a regional (ambient) gradient defined by 

upstream and downstream head boundary conditions. 

Simplifying assumptions used to simulate contaminant transport in the saturated zone are as follows: 

•	 The flow field is steady-state. 

•	 The aquifer is homogeneous and initially contaminant-free. 

•	 Adsorption onto the solid phase is described by a linear or nonlinear equilibrium 

isotherm. 

•	 Chemical and/or biochemical degradation of the contaminant can be described as a firstorder process. 

•	 For a multicomponent decay chain, the number of component species (parent and 

daughters) does not exceed seven. 

•	 The mass flux of contaminants through the source is either constant or controlled by firstorder decay until all mass has been released from the source. 

•	 The chemical concentration in groundwater is dilute, and the chemical is present in the 

aqueous and aquifer solid phases only. 

The required input parameters for this group include recharge rate, saturated thickness, hydraulic 

conductivity, hydraulic gradient, and groundwater temperature and pH. 

The receptor well location is also a saturated zone-specific parameter. As the location at which potential 

exposure to groundwater is measured, it can be anywhere downgradient of the waste management unit, 

within the areal and/or vertical extent of the contaminant plume, and/or along the contaminant plume 

center line. The receptor well location is determined using the following parameters: 

•	 Radial distance from the waste management unit (X-dimension) 

•	 Angle off of the plume centerline (Y-dimension) 

•	 Depth of the intake point (Z-dimension; defined as well depth within the saturated zone) 

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DAF = dilution attenuation factor (unitless) calculated 

CL = leachate concentration (that is, TCLP) waste- and chemical-specific 

(milligram per liter [mg/L]) 

 

Default 

C gw = waste constituent concentration in groundwater computed with 

 

 

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The EPACMTP fate and transport model was used to determine the degree of dilution and attenuation that 

a chemical will undergo as it leaches from a waste management unit and is transported in the subsurface, 

into the saturated zone, and to a theoretical downgradient receptor well. The decrease in chemical 

concentration during the chemical’s transport from the leachate to the receptor well is the DAF. The 

EPACMTP used Monte Carlo probability analyses to compute a DAF for each chemical in the DRAS 

chemical library. The EPACMTP determines the groundwater concentration at the theoretical drinking 

water well that is the 90th percentile of all predicted concentrations (C gw) resulting from the Monte Carlo 

analyses (that is, 90 percent of the resulting receptor well concentrations are less than this concentration). 

The DAF is then calculated by dividing the initial (C ) waste concentration by the predicted groundwater 

concentration at the 90th percentile as shown in Equation 2-1. 

L

(2-1) 

where: 

(mg/L) EPACMTP 

If the maximum allowable concentration of a chemical at the groundwater receptor well (Cgw) is assumed 

to be an HBN, Equation 2-1 can be rearranged, and the HBN (or a maximum contaminant level [MCL], if 

available) can be substituted for the receptor well concentration (Cgw) to back-calculate a maximum 

allowable leachate concentration (CLMAX ) for the ingestion exposure pathway. Equation 2-2 reflects this 

approach. 

(2-2) 

where: 

Default 

CLMAX = maximum allowable leachate concentration calculated 

(TCLP concentration) (mg/L) 

HBN = health-based number (or MCL) (mg/L) chemical-specific 

DAF = dilution attenuation factor (unitless) computed with EPACMTP 

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For further details about using the CMTP in conjunction with MINTEQA2 to derive DAFs, see Appendix 

A-6, Application of CMTP and MINTEQA2 for determining DAFs. 

2.2.2.3 Derivation Waste Volume-Specific DAFs 

Application of the EPACMTP to the U.S. EPA Delisting Program allows evaluation of specific waste 

volumes. To generate waste volume-specific DAFs, a number of EPACMTP runs were conducted, each 

run with a user-specified waste volume. For each Monte-Carlo realization, the user-specified waste 

volume was divided by a waste management unit area randomly drawn from a site in a regional site 

database, while retaining the site’s original area/depth ratio. The regional site database contains waste 

management unit depth and area data obtained from a national survey of Subtitle D industrial landfill 

facilities performed by the U.S. EPA Office of Policy Planning and Implementation (OPPI) (hereinafter 

referred to as the OPPI survey data) (Westat 1987; U.S. EPA 1988b), and from the 2001 EPA’s Industrial 

Surface Impoundment (SI) Study (US EPA 2001). The OPPI and SI survey data were compiled along 

with hydrogeologic data in the regional site database. The data for each waste management scenario are 

presented in “EPACMTP Parameters/ Data Technical Background Document” (U.S. EPA 2003b). For a 

complete description of the regional site-based modeling approach, refer to the “EPACMTP Technical 

Background Document” (U.S. EPA 2003a). 

Rather than compute DAFs for each of 326 chemicals for a range of waste volumes for landfills and 

surface impoundments, a method was developed to scale the DAFs based on a specific waste volume to 

the DAFs computed using the OPPI and SI data bases (U.S. EPA 1988b, 2001). The EPACMTP fate and 

transport model was used to compute waste volume-specific DAF scaling factors for the landfill and 

surface impoundment waste management scenarios (U.S. EPA 2003a, b). The DAF scaling factors and 

the regression equations developed for the landfill and surface impoundment waste management scenarios 

are presented in the following sections. 

2.2.2.3.1 Landfills 

DAF scaling factors were developed for lifetime landfill waste volumes ranging from 10,000 to 1,000,000 

3 cubic yards (yd ), based on the range of waste volumes encountered in the OPPI survey (U.S. EPA

1988b). The landfill DAF scaling factors are plotted in Figure 2-1. Figure 2-1 indicates that the DAF 

3 scaling factor is approximately 1.0 for landfill lifetime waste volumes greater than 154,000 yd . U.S. 

EPA did not consider DAF scaling factors less than 1.0. Therefore, for lifetime landfill waste volumes 

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3 greater than 154,000 yd , the waste volume-specific DAF is equal to the DAF computed based on the

OPPI data base. 

Equation 2-3 can be used to determine the DAF scaling factor (DAF ) as a function of lifetime landfill sf

waste volume for landfilled wastes. 

(2-3) 

where: 

Default 

DAFsf = DAF scaling factor (unitless) computed 3 V = landfill lifetime volume of waste (yd ) delisting petition-specific 

The correlation coefficient of this regression equation is 0.99, indicating that changes in the DAF scaling 

factor are explained by changes in the waste volume and that the DAF scaling factor can be predicted 

with confidence as a function of waste volume. 

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2.2.2.3.2 Surface Impoundments 

DAF scaling factors were developed for lifetime surface impoundment waste volumes ranging from 2,000 

3 to 1,000,000 yd based on the OPPI survey data (U.S. EPA 1988b). The surface impoundment DAF 

scaling factors are plotted in Figure 2-2. For surface impoundment lifetime waste volumes greater than 

3 about 40,000 yd , the DAF scaling factor is equal to 1.0.

Equation 2-4 can be used to determine the DAFsf as a function of waste volume for surface 

impoundments. 

(2-4) 

where: 

Default 
 

DAFsf = DAF scaling factor (unitless) computed
 3 V = impoundment lifetime volume of waste (yd ) delisting petition-specific 

The correlation coefficient of this regression equation is 0.99, indicating a good fit of this line to the data 

points and that the DAF scaling factor for surface impoundments can be predicted as a function of waste 

volume with a high level of confidence. 

2.2.3 Calculation of Groundwater Waste Constituent Concentration at the POE 

As discussed above, the EPACMTP model estimates fate and transport of chemicals in groundwater. 

Within this medium, three human health exposure pathways are evaluated by the U.S. EPA Delisting 

Program: direct ingestion, dermal absorption, and shower inhalation. This section describes how the 

waste constituent concentration at the groundwater POE is calculated. 

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Figure 2-1. DAF Scaling Factors for Landfills 

2-16
 
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Figure 2-2. DAF Scaling Factors for Surface Impoundments 

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(2-5) 

To adjust for the petitioned waste volume, the EPACMTP-modeled DAF is multiplied by the DAFsf 

determined as described above. The product is defined as the waste volume-adjusted DAF (DAFva), 

which is specific for each waste constituent. Equation 2-5 is used for this calculation.

 where: 

Default 

DAF va = waste volume-adjusted DAF (unitless) calculated 

DAFsf = DAF scaling factor (unitless) Equation 2-3 (landfills) or 2-4 

(surface impoundments) 

DAF = dilution attenuation factor (unitless) Computed with 

EPACMTP 

Using Equation 2-6, the TCLP leachate concentration of a petitioned waste constituent is divided by the 

waste volume-adjusted DAF to obtain the predicted groundwater receptor well concentration of the 

constituent. 

(2-6) 

where: 

Default 

C gw = waste constituent concentration in groundwater calculated 

(mg/L) 

TCLP = TCLP concentration of waste constituent (mg/L) waste constituent-specific 

DAF va = waste volume-adjusted DAF (unitless) Equation 2-5 

The waste constituent concentration in groundwater calculated using Equation 2-6 is used to determine 

exposure point concentrations for the three groundwater exposure pathways identified above: direct 

ingestion, dermal absorption, and shower inhalation. The exposures, in turn, are used to compute the risk 

and hazard associated with exposure to the waste constituent via each of these three pathways (see 

Chapter 4). It should be noted here that the EPACMTP DAFs may be very high (that is, greater than 

1,000,000) for some waste constituents. 

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2.3 RELEASES TO SURFACE PATHWAYS 

This section describes the equations used to predict releases of chemicals from landfill and surface 

impoundment waste management units via the air and surface water pathways. The equations predict 

medium-specific (soil and air) concentrations of the waste constituent or constituents at the POE. The 

release of chemicals to air via particulate matter from landfills is discussed in Section 2.3.1, the release of 

volatile chemicals to air from landfills and surface impoundments is discussed in Section 2.3.2, and the 

methodology for calculating surface water concentrations for waste constituents eroded from landfills is 

presented in Section 2.3.3. 

2.3.1 Calculation of Waste Constituent Concentration in Air — Particles 

U.S. EPA Region 6 considers exposure to airborne particulate hazardous constituents released from 

wastes disposed of in landfills to be a function of (1) inhalation of particles and their absorption into the 

lungs at the POE and (2) air deposition of particles and subsequent ingestion of the soil-waste mixture at 

the POE. To address inhalation and ingestion of particles, U.S. EPA Region 6 calculates particulate 

emissions resulting from wind erosion of soil-waste surfaces and from vehicular traffic over the waste and 

waste loading and unloading. To estimate the respirable particulate emissions resulting from wind 

erosion of surfaces with an infinite source of erodible particles, U.S. EPA Region 6 used the methodology 

documented in “Rapid Assessment of Exposure to Particulate Emissions from Surface Contamination 

Sites (RAEPE)” (U.S. EPA 1985a). To calculate the dust and particulate emissions resulting both from 

vehicular traffic and from waste loading and unloading operations at a facility, U.S. EPA Region 6 used 

the methodologies documented in “Compilation of Air Pollutant Emission Factors, Volume 1: Stationary 

Point and Area Sources” (AP-42) (U.S. EPA 1985b). 

Particulate emission rates computed using these methodologies were summed and entered in the Ambient 

Air Dispersion Model (AADM), a steady-state, Gaussian plume dispersion model developed by U.S. EPA 

to predict the concentrations of constituents 1,000 feet downwind of a hypothetical land disposal facility 

(U.S. EPA 1985c). When evaluating delisting petitions, the DRAS assumes conservative values for all 

variables that are likely to influence the potential for soil erosion, including wind velocity and vegetative 

cover. U.S. EPA Region 6, however, modified the AADM assumptions regarding unit dimensions to 

more closely resemble a landfill’s dimensions. 

The method used to calculate total respirable particulate emissions from a landfill is described in 

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Section 2.3.1.1. The methods used to calculate the average downwind concentrations of particles and the 

downwind concentrations of respirable particulate emissions are described in Section 2.3.1.2. Calculation 

of air deposition rates and resulting soil concentrations is described in Section 2.3.1.3. 

2.3.1.1 Estimation of Particulate Emissions 

U.S. EPA Region 6 calculates the mass flux (the source term or amount of waste material that becomes 

airborne) associated with particulate emissions using the methodology presented in RAEPE (U.S. EPA 

1985a). The mass flux can be calculated using Equations 2-7 and 2-8. 

(2-7) 

where: 

Default 

Q p = emission rate of waste constituent particles calculated 

(milligrams per second [mg/s]) 

á = mass fraction of constituent in waste (unitless) waste-specific (equal to total 

concentration in waste; mg/mg) 

ET = total emission rate of particles that may be Equation 2-8

 inhaled (gram/hour [g/hr]) 

and: 

(2-8) 

where: 

Default 

ET = total emission rate of particles that may be calculated 

inhaled (g/hr) 

E w = particulate emission rate from wind erosion (g/hr) Equation 2-9 

E v = particulate emission rate from vehicular traffic (g/hr) Equation 2-17 

El = particulate emission rate from waste loading and Equation 2-21 

unloading operations (g/hr) 

The equations used to compute emissions resulting from wind erosion (Ew), vehicular traffic (Ev), and 

waste loading and unloading (El

) are presented in the following sections. 

Wind Erosion Emissions 

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Wind erosion emissions of respirable particles (those smaller than 10 microns [ìm] in diameter)a

 can be 

calculated using Equation 2-9. 

(2-9) 

where: 

Default 

Ew10 = wind erosion emission rate for particles up to 10 ìm calculated 

(g/hr) 

Vf = fraction of disposal site covered with vegetation (unitless) 0 (U.S. EPA 1994a) 

U = mean annual wind speed (meters/second [m/s]) 4 (discussed below) 

Ut = threshold value of wind speed at 7 m (m/s) 5.44 (Equation 2-10) 

F(X) = dimensionless function obtained from a plot in RAEPE 1.33 (Appendix B, (U.S. 

EPA 1985a)	 Figure B-6) 2 Aexposed = area of the waste management unit exposed (m ) Equation 2-15 

For the landfill waste disposal scenario, U.S. EPA assumes that no vegetative cover is present, thereby 

assuming enhanced erodibility of soil or waste. Therefore, the fraction of the disposal site covered with 

vegetation (Vf) is equal to 0. The mean annual wind speed is assumed to be 4 m/s. This value represents 

the average of the wind speeds registered at U.S. climatological stations as documented in Table 4-1 of 

RAEPE (U.S. EPA 1985a). This value is also assumed to be associated with climate stability class D.b 

The threshold value of wind speed, Ut , can be derived in the following manner as described in RAEPE 

(U.S. EPA 1985a) using Equation 2-10: 

•	 The waste is assumed to exhibit a particle size of 0.2 millimeter (mm) (a typical size for 

fine sand). The wind erosion threshold friction velocity (U*t) used in Equation 2-10 is 

derived from RAEPE (U.S. EPA 1985a) plots (see Appendix B, Figure B-6 of this 

DTSD) and is equal to 33 centimeter per second (cm/s). 

•	 A roughness height (Zo) of 1.0 cm (for a plowed field) is obtained from Appendix B. 

•	 Based on Z , a ratio of wind speed at 7 m (U ) to friction velocity (U ) can be obtained o	 t *t 

from Appendix B. In this case, U /U = 16.5. t *t 

Therefore: 

a Particles less than or equal to 10 :m in aerodynamic diameter are defined as the respirable fraction. 

Refer to discussions in RAEPE (U.S. EPA 1985a), AP-42 (U.S. EPA 1985b), and the “Draft Superfund Exposure 

Assessment Manual” (U.S. EPA 1986b).

b Stability class is a meteorological classification of atmospheric properties as those properties relate to 

dispersion of airborne materials. The classes range from A, extremely unstable, to F, moderately stable. The 

coefficients defined by the stability class are used in Equation 2-9 to calculate the downwind dispersion. 

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(2-10) 

Using the values identified above, X is calculated as shown in Equation 2-11. 

(2-11) 

Given the value of X computed using Equation 2-11, an F(X) value of 1.33 is obtained from the graph 

provided in RAEPE (U.S. EPA 1985a) (see Appendix B, Figure B-6 of this DTSD). 

The landfill area is determined using an equation developed for the U.S. EPA OSW Hazardous Waste 

Delisting Program (U.S. EPA 1991b). Waste volume is correlated with disposal unit area by means of a 

regression equation developed from national data on disposal unit dimensions obtained in the OPPI 

survey. A method to convert waste volume to disposal unit area was developed for both landfills and 

surface impoundments (U.S. EPA 1991b). Equation 2-12 below presents the regression equation to 

determine the disposal unit area for a one-time delisting of landfill waste or surface impoundment waste 

for a RCRA delisting. 

(2-12) 

where: 

Default 
 

A = area of the waste management unit (acres) calculated 
 3 V = total volume of waste (yd ) delisting petition-specific 

The area of the waste management unit is based on the volume of waste reported in the petition. 

U.S. EPA delisting program assumes that a petitioned waste is disposed of in one Subtitle D facility over 

a 20-year period (U.S. EPA 1988b). Based on maximum annual waste volumes and assuming waste 

disposal into the same landfill or surface impoundment over a 20 year period, Equation 2-13 below is 

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Chapter 2: Estimation of Chemical Releases and Media Concentrations Revised October 2008 

used to determine the disposal unit area for multi-year RCRA delistings to a landfill or surface 

impoundment. 

where: 

Default 
 

A = area of the waste management unit (acres) calculated 
 3 V = annual volume of waste (yd /year) delisting petition-specific 

U.S. EPA delisting program believes that, at most, a month’s (30 days’) worth of waste would be 

uncovered at any one time. Therefore, the fraction of the area in the landfill that would be exposed and 

available as a source for particulate emissions (Fexposed) is calculated as shown in Equation 2-14. 

(2-13) 

(2-14) 

This fraction exposed is used to compute the exposed area (Aexposed) and the site width as shown in 

Equation 2-15. 

(2-15) 

where: 

Aexposed 

Fexposed 

A 

= 

= 

= 

area of the waste management unit exposed (m ) 2 

fraction of the area exposed (unitless) 

area of the waste management unit (acres) 

Default 

calculated 

Equation 2-14 

Equation 2-12 (one-time) or 2­

13 (multi-year) 

Although particles greater than 10 ìm in size generally are not considered respirable, U.S. EPA Region 6 

also calculated the emission rate for particle sizes up to 30 ìm in order to assess the potential impact of 

deposition and ingestion of such particles (see Section 2.3.1.3). U.S. EPA Region 6 used the distributions 

of wind-eroded particles presented in AP-42 (U.S. EPA 1985b) to estimate an 

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RCRA Delisting Technical Support Document 

Chapter 2: Estimation of Chemical Releases and Media Concentrations Revised October 2008 

emission rate for particles up to 30 ìm in size .c Specifically, these distributions indicate that the release 

rate for particles up to 30 ìm in size should be approximately twice the release rate calculated for 

particles 10 ìm in size. Equation 2-16 shows this relationship. 

(2-16) 

where: 

Ew30 

Ew10 

= 

= 

wind erosion emission rate for particles up to 

30 ìm (g/hr) 

wind erosion emission rate for particles up to 

10 ìm (g/hr) 

Default 

calculated 

Equation 2-9 

Vehicle Emissions 

U.S. EPA Region 6 used Equation 2-17 as described in AP-42 (U.S. EPA 1985b) to calculate vehicle 

emissions: 

(2-17) 

where: 

Default 

Ev = vehicle emissions (g/hr) calculated 

k = constant; 0.36 for particles up to 10 ìm and (U.S. EPA 1986b) 

0.8 for particles up to 30 ìm 

s = percent silt content of waste (unitless) 8 (U.S. EPA 1986b) 

S = mean vehicle speed (kilometers per hours [km/hr]) 24 (U.S. EPA 1986b) 

W = mean vehicle weight (tons) 15 (U.S. EPA 1986b) 

w = mean number of wheels per vehicle 6 for W=15 tons 

(U.S. EPA 1986b) 

Np = number of days per year with at least 0.01 inch of 90 (U.S. EPA 1986b) 

precipitation (days/year) 

VKT	 = vehicle kilometers traveled = kilometers/trip x the Equation 2-18 

calculated number of round trips made per day 

(km-trips/day) 

Conservatively assuming that a minimum of 7.4 vehicles traverse the waste daily, based on the 95th 

percentile of vehicles per day reported in Table 3A-1 of the 1987 U.S. EPA survey of industrial Subtitle 

c Particles less than or equal to 30 ìm in diameter can be transported for considerable distances downwind; 

those larger than 30 ìm are likely to settle within a few hundred feet. See AP-42 (U.S. EPA. 1985b and the “Draft 

Superfund Exposure Assessment Manual” (U.S. EPA 1986b). 

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D waste facilities in the United States (Westat 1987 and U.S. EPA 1997h), and adding the number of 

daily trips based on the petitioned waste volume, VKT can be calculated using Equation 2-18. 

(2-18) 

minimum round trips per day 

where: 

Default 
 

VKT = vehicle kilometers traveled (km-trips/day) calculated
 

Site Width = (m) Equation 2-15
 

TPDmin =
 7.4 (95th percentile from Subtitle 

D Survey) 

ñw = waste density (tons per cubic yard) 1.42 (U.S. EPA 1993a) 3 V = annual volume of waste (yd /year) delisting petition-specific 

Vehicle emissions, E and E , can then be calculated using Equation 2-17 as shown in Equations 2-19 v10 v30 

and 2-20. 

(2-19) 

and 

(2-20) 

U.S. EPA assumes a maximum total annual generation rate for the petitioned waste to estimate a daily 

disposal amount in tons per day. Therefore, assuming that 15-ton vehicles are used, the number of trips 

can be determined. U.S. EPA Region 6 realized that particulate emissions from vehicles are related to the 

size and weight of the vehicles used for waste transport as well as the number of vehicle trips made. The 

assumption of small vehicle use (with more trips) is conservative. 

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Chapter 2: Estimation of Chemical Releases and Media Concentrations Revised October 2008 

Waste Loading and Unloading Emissions 

U.S. EPA Region 6 used Equation 2-21, which is based on an AP-42 (U.S. EPA 1985b) methodology, to 

calculate emissions from waste loading and unloading operations, El

. In Equation 2-21, El is given in 

units of kilograms of waste released per metric ton of waste disposed of. 

(2-21) 

where: 

Default
 

El = emissions from waste loading and unloading calculated
 

operations (kg/metric ton)
 

Kp = batch drop particle size multiplier (dimensionless): U.S. EPA 1986b 

S 0.36 for particles up to 10 ìm 

S 0.73 for particles up to 30 ìm 

s = silt content of waste (percent) 8 (U.S. EPA 1986b) 

U = mean annual wind speed (m/s) 4 (U.S. EPA 1986b) 

DH = drop height of material from truck (m) 2 (U.S. EPA 1986b) 

M = moisture content of waste (percent) 1 (U.S. EPA 1986b) 3 Yd = dumping device capacity (m ) 10 (U.S. EPA 1986b) 

Based on the petitioned waste volume, emissions from waste loading and unloading operations are 

calculated for particles up to 10 and 30 ìm in size as shown in Equations 2-22 and 2-23, respectively. 

(2-22) 

where: 

Default 

= waste loading and unloading emission rate of calculated 

particles up to 10 ìm (kg/metric ton) 

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(2-23) 

where: 

Default 

= waste loading and unloading emission rate of calculated 

particles up to 30 ìm (kg/metric ton) 

To convert the emission rate from kilogram per metric ton to the required grams per hour, U.S. EPA 

Region 6 assumes that the waste has an average density of 1.686 g/cm3

 (U.S. EPA 1993a), which is equal 

to 1.42 tons per cubic yard (or 1.289 metric tons per cubic yard). Using the petitioned waste volume (V) 

in cubic yards per year and the waste density, the Agency determines the emission rate in grams per hour 

as follows: 

Total Respirable Particulate Emissions 

U.S. EPA Region 6 calculated the total annual average emissions of respirable particles (ET10) by 

summing E for wind erosion, E for vehicle travel, and E for waste loading and unloading w10 v10 l10 

operations. Equations 2-7 and 2-8 were then combined to produce Equation 2-24. 

(2-24) 

where: 

Default 

Qp10 = emission rate of waste constituent particles calculated 

up to 10 ìm (mg/s) 

á = mass fraction of waste constituent (unitless) waste-specific (equal to total 

concentration in waste in mg/mg) 

Ew10 = wind erosion emission rate of particles up Equation 2-9 

to 10 ìm (g/hr) 

Ev10 = vehicle travel emission rate of particles Equation 2-19 

up to 10 ìm (g/hr) 

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El10 = waste loading and unloading emission rate Equation 2-21 and 2-22 

of particles up to 10 ìm (g/hr) 

2.3.1.2 Calculation of Particulate Constituent Concentration in Air at the POE 

U.S. EPA Region 6 used the AADM, modified for a landfill source (U.S. EPA 1985c), to calculate the 

downwind concentrations of constituents released from a land disposal facility. The model assumes that 

(1) the emission rate is constant over time, (2) the emissions arise from an upwind virtual point source 

with emissions occurring at ground level, and (3) that no atmospheric destruction or decay occurs. 

U.S. EPA Region 6 calculated the downwind concentration of a constituent at the POE (Cavg) using 

Equation 2-25. 

(2-25) 

where: 

Default
 

Cavg = downwind concentration of waste constituent at POE calculated
 3 (mg/m )

Qp = emission rate of waste constituent particles (mg/s) Equation 2-24 

Óz = vertical dispersion coefficient (m) Equation 2-26 

U = mean annual wind speed (m/s) 4 (U.S. EPA 1986b) 

Lv = distance from the virtual point to the compliance point Equation 2-27 

located 1,000 feet (304.8 m) downwind (m)
 

F = frequency that wind blows from the sector of 0.15 


interest (unitless)
 

The Pasquill-Gifford vertical dispersion coefficient, Óz, can be calculated using Equation 2-26 as 

described in the “Industrial Source Complex (ISC) Dispersion Model User’s Guide” (U.S. EPA 1986a). 

(2-26) 

where: 

Default 

Ó = z vertical dispersion coefficient (m) calculated 

a = coefficient for stability class D 32.093 (from Appendix B, Table B-5) 

L = distance from center of the uncovered waste (site width)/2 + 0.3048 km 

area to compliance point 1,000 feet downwind (km) (U.S. EPA 1986b) 

b = coefficient for stability class D 0.81066 (from Appendix B, Table B-5) 

U.S. EPA Region 6 uses Equation 2-27 to calculate the distance from the virtual point to the compliance 

point 1,000 feet (304.8) downwind. 

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(2-27) 

where: 

Default 

L v = distance from the virtual point to the compliance calculated 

point located 1,000 feet downwind (km) 

L = distance from center of the uncovered waste area to (U.S. EPA 1986b) 

the (site width)/2 + 0.3048 km compliance point 

1,000 feet (304.8 m) downwind (km) 

L' = virtual distance (the distance necessary to convert Equation 2-28 

from an ideal point source to a volume source) (km) 

The virtual distance is calculated as the distance required for the transverse standard deviation of the 

Gaussian plume, óy, to grow to half the width of the site. This distance will be different for each stability 

class and is calculated for Pasquill Stability Category D using Equation 2-28. 

(2-28) 

where: 

Default 

L' = virtual distance (km) calculated 

)

0.5 x = 1⁄2 the width of the area exposed (m) 0.5*(Aexposed 

p = Pasquill Stability Category D coefficient (unitless) 68.26 (from Appendix B, 

Table B-6) 

q = Pasquill Stability Category D coefficient (unitless) 0.919 (from Appendix B, 

Table B-6) 

As stated previously, U.S. EPA Region 6 used a value of 4 m/s for wind speed (U). Based on information 

in the Climatic Atlas of the United States (Visher 1954), U.S. EPA Region 6 assumed that the average 

frequency (F) that the wind blows from various directions for many U.S. cities is 0.15. Using these 

values and the calculated values for Óz and Lv as inputs, the average downwind concentration of emissions 

can be calculated using Equation 2-25. 

The concentration of each hazardous constituent actually inhaled can be calculated using Equation 2-29. 

3 U.S. EPA Region 6 assumed that a moderately active person breathes 20 m of air per day based on

information in U.S. EPA’s “Exposure Factors Handbook” (U.S. EPA 1997b). 

(2-29) 

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where: 

Default 

Cinh = concentration of waste constituents inhaled (mg/day) calculated 3 Cavg = downwind conc. of waste constituent at POE (mg/m )	 Equation 2-25 3 20 = Adult inhalation rate (m /day)	 (US EPA, 1997b) 

2.3.1.3 Calculation of Air Deposition Rates and Resulting Soil Concentrations at the POE 

U.S. EPA Region 6 used the steps described below to investigate air deposition of the annual total 

emissions of particles less than or equal to 30 ìm in size (that is, total suspended particles [ET30]) to soil 

1,000 feet from the edge of a disposal unit. First, U.S. EPA Region 6 summed E w30 for wind erosion, Ev30 

for vehicle travel, and El30 for waste loading and unloading operations (calculated in Equations 2-16, 2­

20, and 2-23, respectively) to arrive at the total emission rate of particles up to 30 ìm in size (Qp30) as 

shown in Equation 2-30. 

(2-30) 

where: 

Default
 

Qp30 = emission rate of waste constituent particles calculated
 

up to 30 ìm (mg/s)
 

á = mass fraction of waste constituent (unitless)	 waste-specific (equal to total 

concentration in waste in 

mg/mg) 

Ew30 = wind erosion emission rate of particles up to Equation 2-16
 

30 ìm (g/hr)
 

Ev30 = vehicle travel emission rate of particles up to Equation 2-20
 

30 ìm (g/hr)
 

El30 = waste loading and unloading emission rate of Equation 2-23
 

particles up to 30 ìm (g/hr) 


U.S. EPA Region 6 calculated the average downwind particulate concentration in the air 1,000 feet from 

the disposal unit boundary using the AADM, as described previously (see Equation 2-25). U.S. EPA 

Region 6 then calculated the flux of particles hitting the ground at the downwind point using Equation 2­

31. 

(2-31) 

where: 

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Default 2 qd = rate of deposition (mg/m /s) calculated 

vd = deposition velocity (m/s) 0.03 (U.S. Department of 

Energy [U.S. DOE] 1984) 

Cavg = downwind concentration of waste constituent at POE Equation 2-25 3 (mg/m )

The effective deposition velocity is a function of friction velocity, surface roughness height, particle 

density, and particle size (U.S. DOE 1984). U.S. EPA Region 6 estimated a value of 0.03 m/s for the 

effective deposition velocity for all particles less than or equal to 30 ìm in size as described below. 

The ranges of values for friction velocity and surface roughness height can be obtained from Figure B-4 

in Appendix B of this document. To be conservative, U.S. EPA Region 6 assumed that the terrain on 

which particles are deposited is a field of grass with blades up to about 5 cm high (a surface not 

susceptible to erosion but with a capacity for trapping particles) and assumed a wind speed of 4 m/s. 

These assumptions led to a roughness height of 2.0 cm and a friction velocity of 50 cm/s (see Figure B-3 

in Appendix B). A particle density of 2.61 g/cm3

 (an average particle density for clay particles) was 

assumed for the petitioned waste (U.S. EPA 1994a). Figure B-6 in Appendix B presents curves of 

deposition velocity as a function of particle size for several particle densities and roughness heights. U.S. 

EPA Region 6 used these curves to obtain deposition velocities for particle sizes of 1, 10, and 30 ìm 

(0.001, 0.018, and 0.07 m/s, respectively). 

To obtain an effective deposition velocity for particles less than or equal to 30 ìm in size, U.S. EPA 

Region 6 calculated the mean velocities for the two particle size ranges, 1 to 10 ìm (0.0095 m/s) and 10 

to 30 ìm (0.044 m/s), and then used the average of these two mean velocities to derive an effective 

deposition velocity of 0.03 m/s for use in Equation 2-31. 

Finally, U.S. EPA Region 6 calculated the resulting soil concentration (Csoil) after 1 year of accumulation. 

U.S. EPA Region 6 conservatively assumed no constituent removal for 1 yr (that is, no leaching, 

volatilization, soil erosion, or degradation). To calculate Csoil, U.S. EPA Region 6 used Equation 2-32. 

(2-32) 

where: 

Default 

Csoil = concentration of constituent in soil at calculated 

the POE (mg/kg) 

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2 qd = rate of deposition in mg/m /s Equation 2-31 3 ñb = soil bulk density (kg/m ) 1450 (Brady 1984) 

t = soil thickness from which particles can be ingested 0.01 (U.S. EPA 1994a) 

(m) 

U.S. EPA Region 6 selected a value of 1450 kg/m3

 for soil bulk density; this value is the midpoint of the 

3 range for soil bulk density (1200 to 1700 kg/m ) cited in The Nature and Properties of Soils (Brady 1984). 

U.S. EPA Region 6 also selected a value of 1 cm as the thickness of the soil surface that will be available 

for mixing and ingestion. U.S. EPA Region 6 believes that a value of 1 cm is reasonable, given the 

assumption that no constituent removal would occur for 1 yr (U.S. EPA 1994a). 

2.3.2 Calculation of Chemical Concentration in Air at the POE-Volatiles 

Petitioned wastes potentially contain volatile organic compounds (VOC). Therefore, U.S. EPA Region 6 

evaluates the potential threat to human health resulting from atmospheric transport and inhalation of 

volatile constituents from a petitioned waste. U.S. EPA Region 6 evaluates the potential influence of 

volatiles in petitioned waste on air quality for two source terms: landfills and surface impoundments. 

U.S. EPA Region 6 derives an annual waste generation rate and estimated emissions from landfills using 

Farmer’s equation (Farmer and others 1978). Estimates of emissions of VOCs from disposal of 

wastewaters in surface impoundments are computed with U.S. EPA’s Surface Impoundment Modeling 

System (U.S. EPA 1990a, 1990b). 

The emission rates derived for the two disposal scenarios are entered in U.S. EPA’s AADM, a steadystate, Gaussian plume dispersion model, to predict the concentrations of constituents 1,000 feet 

downwind of a hypothetical disposal facility. For a complete description and discussion of the AADM, 

refer to (U.S. EPA 1985c). 

2.3.2.1 Calculation of Volatile Emissions from a Landfill Using Modified Farmer's Equation 

Shen’s modification of Farmer’s equation, which was developed by U.S. EPA’s Office of Air Quality 

Planning and Standards (OAQPS), is used to estimate the rate of emission of volatiles from a covered 

landfill (U.S. EPA 1984). This equation provides the rate of volatile emission instead of the flux rate by 

multiplying by the landfill surface area. U.S. EPA Region 6 determined that Farmer’s equation would 

provide a reliable estimate of volatile emissions from a landfill. The rate of emission from the landfill is 

calculated using Equation 2-33. 

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(2-33) 

where: 

Default
 

Ei = emission rate of chemical I (g/sec) calculated
 2 Da = diffusion coefficient of constituent in air (cm /sec) chem-specific (Appendix A-1) 4 2 2 1/1x10 = conversion factor for diffusivity from cm to m 2 Aexposed = surface area (m ) Equation 2-15 3 Csi = saturation vapor concentration of I in landfill (g/m ) Equation 2-35


L = depth of soil cover (m) 0.1524 (U.S. EPA 1994a)
 

Pa = air-filled sand porosity (dimensionless) 0.40 (U.S. EPA 1994a)
 

PT = total sand porosity (dimensionless) 0.40 (U.S. EPA 1994a)
 

W /W = weight fraction of waste constituent I (mg/Kg) waste-specific
 i

1kg/1000g = weight fraction conversion factor for (g/g) to (mg/Kg)
 

1g/1000mg =weight fraction conversion factor for (g/g) to (mg/Kg)
 

It may be assumed that (1) the total porosity of dry sand is about 40 percent (PT = 0.40) (U.S. EPA 

1994a), (2) the percentage of air-filled pore space in the sand above the landfill is 40 percent (Pa = 0.40) 

(U.S. EPA 1994a), and the concentration of the constituent at the surface is negligible. Therefore, 

Equation 2-33 can be simplified to: 

(2-34) 

A conservative assumption is made that the waste constituent I is in a state of pure component at 

saturation vapor concentration. The saturation vapor concentration of a waste constituent (Csi) in the 

landfill can be calculated as shown in Equation 2-35. 

(2-35) 

where: 

Default 3 Csi = saturation vapor concentration of I in landfill (g/m ) calculated
 

pi = vapor pressure of waste constituent I (atm) chem-specific (Appendix A-1)
 

MWi = mole weight of waste constituent I (g/mole) chem-specific (Appendix A-1)
 3 -5 R = molar gas constant (atm-m /mole-EK) 8.21 x 10
 

T = standard temperature (EK) 298 K (U.S. EPA 1994a)
 

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2.3.2.2 Calculation of Downwind Volatile Waste Constituent Concentration in Air at the 

POE—Landfill 

A landfill system may include a cover above the waste to limit volatile emissions. Therefore, U.S. EPA 

Region 6 conservatively assumed that a minimum of a 6-inch daily cover was present. U.S. EPA Region 

6 used the area of the landfill to calculate atmospheric emissions and transport because the entire surface 

area will form a source of volatile waste constituents. 

U.S. EPA Region 6 used the AADM, modified to estimate emissions from a landfill (U.S. EPA 1994a), to 

calculate the downwind concentrations of volatile constituents released from the landfill. The model 

assumes that (1) the emission rate is constant over time, (2) the emissions arise from an upwind virtual 

point source with emissions occurring at ground level and (3) no atmospheric destruction or decay of the 

constituent occurs. U.S. EPA Region 6 calculated the average downwind concentration of a constituent 

as shown in Equation 2-25. The volatile emission rate, Q , is calculated using Equation 2-36. v 

(2-36) 

where: 

Default 

Q v = volatile emission rate (mg/sec) calculated 

Ei = chemical flux (g/sec) Equation 2-34 

The emission rate is then used to compute the downwind concentration (Cavg) by employing the AADM 

(Equation 2-25). To calculate the POE and the mass inhaled (Cinh ), U.S. EPA Region 6 assumes that a 

3 moderately active person breathes 20 m of air per day (U.S. EPA 1997b). The mass inhaled (Cinh ) is 

calculated as shown in Equation 2-37. 

(2-37) 

where: 

Default 

Cinh = mass of waste constituent inhaled (mg/day) calculated 3 Cavg = downwind conc. of waste constituent at POE (mg/m ) Equation 2-25 3 20 = Adult inhalation rate (m /day) (US EPA, 1997b) 

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Cinh is computed for each contaminant in the petitioned waste and is then compared to the inhalation 

levels of concern. 

2.3.2.3 Calculation of Volatile Emissions from a Hypothetical Surface Impoundment 

U.S. EPA Region 6 uses the Surface Impoundment Modeling System (SIMS) (U.S. EPA 1990a, 1990b) 

to calculate emissions from a hypothetical disposal impoundment. SIMS was developed by U.S. EPA’s 

OAQPS and is available for downloading along with its documentation from the OAQPS web site at 

“www.epa.gov.oar.oaqps”. The equations contained in SIMS to estimate emissions are presented in this 

section. 

Surface impoundments can be divided into two general categories: treatment and disposal 

impoundments. In treatment impoundments, wastewater containing particles, biochemical oxygen 

demand (BOD), or photodegradable constituents is introduced into a lagoon. The particles settle to the 

bottom of the unit, and a combination of biological, photochemical, and volatilization mechanisms causes 

destruction or removal of dissolved contaminants. An impoundment may be artificially aerated to speed 

up these processes. The partially treated and clarified wastewater is then drawn off for further treatment 

and is discharged via a permitted outfall (such as an [NPDES] outfall). The settled sludge may be 

dredged and landfilled separately. In disposal impoundments, the wastewater does not flow through the 

unit; instead, the impoundment is sized such that all the water evaporates or infiltrates to groundwater. 

Disposal impoundments are defined as units that receive wastewater for ultimate disposal rather than for 

storage or treatment. Generally, wastewater is not continuously fed to or discharged from these types of 

impoundments. Therefore, the assumption of an equilibrium bulk concentration, which is applicable for 

flow-through impoundments, is not applicable for disposal impoundments; the concentrations of VOCs in 

a disposal impoundment decrease with time. The emission estimating procedure accounts for the 

decreasing liquid-phase concentrations the driving force for air emissions. For a disposal impoundment 

that contains no biomass, the biomass concentration equals zero, and no biodegradation of pollutants 

occurs in the impoundment. 

Further information can be found in the “Background Document for the Surface Impoundment Modeling 

System (SIMS) Version 2.0" (U.S. EPA 1990b). U.S. EPA Region 6 runs the SIMS model for each 

chemical until the fraction emitted to air reaches 1.0 percent, indicating that a significant fraction of the 

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waste constituent has left the unit. For the lighter nonpolar compounds, this generally occurs after 

approximately 11 days. U.S. EPA Region 6 computes the average rate of emission from the 

impoundment by dividing the total mass of each contaminant by the number of days required for all of the 

contaminant to volatilize. The total rate of volatile emissions from the surface impoundment Qv is 

computed using Equation 2-38. 

(2-38) 

where: 

Default 

Qv = total emission rate of volatiles (mg/sec) calculated 3 Co = initial surface impoundment concentration (gm/m ) waste-specific 

tf = time for constituent concentration to reach 1 Equation 2-39 

percent of Co (days) 

K = overall mass transfer coefficient (m/s) Equation 2-41 

Asi = area of the surface impoundment (acres) Equation 2-12 or 2-13 

tr =	 retention time for liquid in a surface impoundment Equation 2-40 

(days) 3 Vsi = volume of liquid in the surface impoundment (m ) delisting petition-specific 

4046.8 = conversion of acres to square meters 

86,400 = conversion of days to seconds 

1000 = conversion of grams to milligrams 

The time for the constituent concentration to reach 1 percent of Co is computed with Equation 2-39. 

(2-39 ) 

where: 

Default 

tf =	 time for constituent concentration to reach 1
 calculated 

 percent of Co (days)
 3 Vsi = volume of liquid in the surface impoundment (m ) waste-specific 

K = overall mass transfer coefficient (m/s) Equation 2-41 2 Asi = area of the surface impoundment (m )	 waste-specific 

The retention time for liquid in a surface impoundment (tr) is computed with Equation 2-40. 

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(2-40) 

where: 

Default 

rt = retention time for liquid in surface impoundment calculated 

(days) 

Vsi = volume of liquid in the surface impoundment (m ) 3 waste-specific (user-provided) 

Qf = flow rate of liquid into surface impoundment (m /yr) waste-specific (user-provided) 3 

The overall mass transfer coefficient is calculated using Equation 2-41. 

(2-41) 

where: 

Default 

K = overall mass transfer coefficient (m/s) calculated 

Kl = liquid side mass transfer coefficient (m/s) Equation 2-42 

Keq = equilibrium constant (unitless) Equation 2-43 

Kg = gas mass phase transfer coefficient (m/s) Equation 2-44 

Kl

 is calculated using Equation 2-42. 

(2-42) 

where: 

Default 

Kl = liquid side mass transfer coefficient (m/s) calculated 

U10 = wind speed at 10 m (m/s) 5.73 (U.S. EPA 1994a) 2 Dw = diffusion coefficient in water (cm /s) chemical-specific 2 -6 Dether = diffusion coefficient of ether (cm /s) 8.56 x 10 (U.S. EPA 1994a) 

Keg, also called the dimensionless Henry’s Law Constant (H’), is calculated using Equation 2-43. 

(2-43) 

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Chapter 2: Estimation of Chemical Releases and Media Concentrations Revised October 2008 

where: 

Default 

Keq = equilibrium constant (unitless) calculated 3 H = Henry’s Law constant (atm-m /g-mol) chemical-specific 3 -5 R = universal gas law constant (atm-m /g-mol K) 8.21 x 10
 

T = standard temperature (K) 298
 

K is calculated using Equation 2-44. g

(2-44) 

where: 

Default 

Kg = gas mass phase transfer coefficient (m/s) calculated 

U10 = wind speed at 10 m (m/s) 5.73 (U.S. EPA 1994a) 

de = effective diameter of surface impoundment (m) Equation 2-45 

Scg = Schmidt number on gas side (unitless) Equation 2-46 

The effective diameter of the surface impoundment is a function of the size of the surface impoundment, 

and therefore, is directly related to the to the volume of liquid waste petitioned for the delisting (Vsi). The 

effective diameter is calculated in Equation 2-45. 

(2-45) 

where: 

Default 

de = effective diameter of surface impoundment (m) calculated 

Asi = area of the surface impoundment (acres) Equation 2-12 or 2-13 

4,046.8 = conversion from acres to square meters 

Scg is calculated using Equation 2-46. 

(2-46) 

where: 

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Default 
 

Scg = Schmidt number on gas side (unitless) calculated
 

μa = viscosity of air (gm/cm-s) 0.000181 (U.S. EPA 1990b) 3 -3 ña = density of air (g/cm ) 1.2 x 10 (U.S. EPA 1990b) 2 Da = diffusivity of constituent in air (cm /s) chem-specific (Appendix A-1) 

Average emission rates (in gm/hr) are computed using this methodology and are then used as input for the 

atmospheric dispersion modeling analysis described below. 

2.3.2.4 Calculation of Downwind Waste Constituent Concentration in Air at the POE Surface 

Impoundment 

Calculation of the downwind waste constituent concentration at the POE depends on the assumed size of 

the disposal unit. The assumptions necessary to determine the size of the unit depend on whether the unit 

is considered to be a covered landfill or a surface impoundment. U.S. EPA Region 6 calculates 

downwind concentrations separately for each configuration of the waste management unit. U.S. EPA 

Region 6 uses Equation 2-25 in Section 2.3.1.2 to calculate downwind concentrations for a landfill. 

Section 2.3.2.4 presents U.S. EPA Region 6’s calculation of downwind waste constituent concentrations 

at the POE for a hypothetical surface impoundment. 

U.S. EPA Region 6 uses the AADM (U.S. EPA 1985c) to calculate the downwind concentrations of 

constituents released from a surface impoundment. The model assumes that (1) the emission rate is 

constant over time, (2) the emissions arise from an upwind virtual point source with emissions occurring 

at ground level, and (3) no atmospheric destruction or decay of waste constituents occurs. U.S. EPA 

Region 6 calculates the average downwind concentration of a constituent as shown in Equation 2-25. The 

downwind concentrations of emissions at the POE for a hypothetical surface impoundment (C avg) can

 be calculated as shown in Equation 2-47. 

(2-47) 

where: 

C avg

Q v

3z

= 

= 

= 

downwind conc. of waste constituent at POE (mg/m ) 3 

total emission rate of volatiles (mg/s) 

vertical dispersion coefficient (m) 

Default 

calculated 

Equation 2-38 

Equation 2-26 

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U = mean annual wind speed (m/s) 4 (U.S. EPA 1986b) 

Lv = distance from the virtual point to the compliance point Equation 2-27

 located 1,000 feet (304.8 m) downwind (m) 

F = frequency that wind blows from the sector of interest 0.15 (U.S. EPA 1986b) 

(unitless) 

The mass of a constituent inhaled (Cinh) downwind from a hypothetical surface impoundment is computed 

with the equation that is used to compute inhaled concentrations of volatiles emitted from landfills 

(Equation 2-37). 

2.3.3 Calculation of Waste Constituent Concentration in Surface Water 

Exposure through the surface water pathway results from erosion of hazardous materials from the surface 

of a solid waste landfill and transport of these constituents to nearby surface water bodies. U.S. EPA 

Region 6 uses the universal soil loss equation (USLE) (Wischmeier and Smith 1978) to compute longterm soil and waste erosion from a landfill in which delisted waste has been disposed. The USLE is used 

to calculate the amount of waste that will be eroded from the landfill. In addition, the size of the landfill 

is computed using the waste volume estimate provided by the petitioner and the preprocessing calculation 

for the EPACML (U.S. EPA 1991b). The volume of surface water into which runoff occurs is determined 

by estimating the expected size of the stream into which the soil is likely to erode (Keup 1985). The 

amount of soil delivered to surface water is calculated using a sediment delivery ratio (Mills and others 

1982). Finally, a portion of the solid phase that is transported to a surface water body is assumed to be 

dissolved in surface water column. The dissolved fraction of the waste constituent in the water column is 

determined by a partitioning equation (U.S. EPA 1998b). U.S. EPA Region 6 uses conservative values 

for all variables likely to influence the potential for soil erosion and subsequent discharge to surface 

water. By using conservative values, U.S. EPA Region 6 is providing reasonably conservative estimates 

of the concentrations of waste constituents in surface water. 

U.S. EPA Region 6 calculates erosion and discharge of contaminants to surface water using the USLE 

(Wischmeier and Smith 1978). The USLE is used to calculate the annual amount of soil and waste 

eroded, as shown in Equation 2-48. 

(2-48) 

where: 

Default 

Aeroded = soil and waste eroded (tons/acre/yr) calculated 

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RF = rainfall erosion factor (1/yr) 300 (Wischmeier and Smith 1978) 

Kef = soil erodibility (tons/acre) 0.3 (Wischmeier and Smith 1978) 

LS = slope length, or topographic factor petition-specific; see Appendix B, Table

 (dimensionless) B-2 (Wischmeier and Smith 1978) 

CM = cover & management factor (dimensionless) 1 (Wischmeier and Smith 1978) 

P = support practice factor (dimensionless) 1 (Wischmeier and Smith 1978) 

Rainfall erosion factor (RF) values range from 20 to 550 per year. A value of 300 was chosen as a 

conservative estimate (Wischmeier and Smith 1978). The distribution of rainfall erosion factor values for 

the United States is given in Figure B-1 Appendix B (Wischmeier and Smith 1978). Values greater than 

300 occur in only a small portion of the southeastern United States. By selecting a value of 300 for its 

analysis, U.S. EPA Region 6 ensures that a reasonable worst-case scenario is provided for most possible 

landfill locations within the United States. 

Soil erodibility (K ) factors range from 0.1 to 0.4 ton per acre (see Table B-1 in Appendix B). A value of ef

0.3 was selected for the analysis; it corresponds to the clay loam, clay, and silty clay loam soil types 

(Wischmeier and Smith 1978). U.S. EPA Region 6 believes that these soil types represent a reasonable 

worst-case approximation of the types of soil material present at a landfill. 

Topographic factor (LS) values range from 0.06 to 12.9 (see Table B-2 in Appendix B) and account for 

the influence of slope length and steepness on erosion potential (Wischmeier and Smith 1978). The value 

is a function of the slope and slope length (see the equation at the bottom of Table B-2 in Appendix B). 

This parameter is partly petition-specific in that the slope length is calculated from the area of the landfill. 

The slope, however, is not petition-specific and is conservatively assumed to be 5%. 

Cover and management factors (CM) range from 0.4 to 1.0 (see Table B-3 in Appendix B). A value of 

1.0 reflects dedicated disposal practices at a facility. 

Support practice factor (P) values range from 0.25 to 1.0 (see Table B-4 in Appendix B). This factor 

reflects the influence of conservation practices on erosion potential. If conservation practices are used 

(for example, contouring or terracing), the potential for erosion is lower. A support practice factor value 

of 1.0 means that no support practice is used. This value was therefore chosen as the most conservative 

value for erosion potential. 

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2.3.3.1 Computing the Amount of Soil Delivered to Surface Water 

U.S. EPA Region 6 computes the percentage of eroded material that is delivered to surface water (the 

sediment delivery ratio, Sd) based on the assumption that some eroded material will be redeposited 

between the landfill and the surface water body. U.S. EPA Region 6 assumes a distance (D) of 100 m to 

the nearest surface water bodyd

 and uses the sediment delivery ratio equation developed by Mills and 

others (1982) as shown in Equation 2-49. 

(2-49) 

where: Default 

Sd = sediment delivery ratio (unitless) calculated 

0.77 = constant (unitless) assumed (U.S. EPA 1994a) 

D = 100, distance to stream or river (m) assumed (U.S. EPA 1994a) 

U.S. EPA Region 6 multiplies the total annual mass of eroded material by the sediment delivery ratio to 

the determine the mass of soil and waste delivered to surface water (A )s . Using a sediment delivery ratio 

of 0.28, As can be calculated as shown in Equation 2-50. 

(2-50) 

where: 

Default 

As = soil and waste mass delivered to surface water (kg/acre/yr) calculated 

Aeroded = amount of soil and waste eroded (tons/acre/yr) Equation 2-48 

Sd = sediment delivery ratio (unitless) 0.28 (Equation 2-49) 

The total annual amount of soil and waste eroded from the landfill is then calculated using Equation 2-51. 

(2-51) 

d Refer to the draft “National Survey of Solid Waste (Municipal) Landfill Facilities” (U.S. EPA 

1988b). This report shows that 3.6 percent of the surveyed landfill facilities are located within 1 mile (1,609 m) 

of a river or stream and that the average distance from these facilities to the closest river or stream is 1,936 feet 

(586 m). Therefore, the assumption of D = 100 m is conservative. 

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where: 

Default 

Aw = rate of soil and waste erosion from landfill calculated

 (kg/acre/yr) 

As = soil and waste mass delivered to surface water Equation 2-50

 calculated (kg/acre/yr) 

Fexposed = fraction of area exposed to erosion (unitless) 0.0041 (U.S. EPA 1994a) 

The fraction of waste exposed to erosion, Fexposed, is based on the assumption that 1 month’s worth of 

waste is uncovered at any one time that is, 30 days/(365 days/yr x 20 yr) = 0.0041. U.S. EPA Region 6’s 

assumption that 1 month’s worth of waste would be left uncovered at any one time and thus would be 

readily transportable by surface water runoff is conservative. The minimum criteria for municipal solid 

waste landfills (MSWLF) set forth in 40 CFR Part 258 require that disposed waste be covered with 6 

inches of earthen material at the end of each operating day or at more frequent intervals (40 CFR 258.21). 

Although a facility might request a temporary waiver of this cover requirement because of extreme 

seasonal weather conditions, it is highly unlikely that 1 month’s worth of waste would be exposed all the 

time during the active life of the facility (that is, 20 years). 

2.3.3.2 Determining the Volume of Surface Water 

U.S. EPA Region 6 selected a representative volume or flux rate of surface water based on stream order, 

which is a system of taxonomy for streams and rivers. A stream that has no other streams flowing into it 

is referred to as a first-order stream. Where two first-order streams converge, a second-order stream is 

created. Where two second-order streams converge, a third-order stream is created. Stream order has 

proven to be a good predictor of flow parameters, including average length, drainage area, mean flow, 

width, depth, and velocity (Keup 1985). Data indicate that second-order streams have a flow rate of 

9 about 3.7 cubic feet per second (3.3 x 10 L/yr). The second-order stream was selected for analysis as the 

smallest stream capable of supporting recreational fishing. Fifth-order streams were also chosen for 

analysis as the smallest streams capable of serving as community water supplies. Fifth-order stream flow 

is on the order of 380 cubic feet per second (3.4 x 1011 L/yr) (U.S. EPA 1994a). 

2.3.3.3 Computing the Waste Constituent Concentration in Surface Water 

The waste constituent concentration in a surface water body near a land disposal facility is calculated 

using Equation 2-52. 

(2-52) 

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where: 

Default 

C sw = concentration of waste constituent in surface water calculated 

(mg/L) 

A = area of the waste management unit (acres) Equation 2-12 or 2-13 

A w = rate of waste erosion from landfill (kg/acre/yr) Equation 2-51 

Q stream = volume of stream (L/yr) (U.S. EPA 1994a) 

- for 5 order stream th 3.4 x 10 11 

- for 2 order stream nd 3.3 x 10 9 

Ctotal waste= waste constituent concentration in delisted waste (mg/kg) waste-specific 

U.S. EPA Region 6’s assumption that surface water runoff from a land disposal facility will be 

uncontrolled and will enter an adjacent surface water body also is conservative because 40 CFR 258.26 

requires adequate run-on and run-off controls at MSWLFs. In addition, 40 CFR 257.33 and 258.27 

prohibit a point source or nonpoint source discharge of pollutants that violates any requirements of the 

Clean Water Act, including NPDES requirements and any requirement of an approved area-wide or statewide water quality management plan. 

2.3.4. Calculation of Dissolved Phase Waste Concentration in Surface Water (Cdw) 

U.S. EPA OSW (U.S. EPA 1998b) recommends the use of Equation 2-53 to calculate the fraction of the 

surface water concentration (C ) of a waste constituent that is dissolved in the water column (C ). sw dw 

(2-53) 

w 

here: 

Default 

Cdw = Dissolved phase water concentration (mg/L) Calculated 

Csw = Waste concentration in water column (mg/L) Equation 2-52
 

Kdsw = Suspended sediments/surface water partition chem-specific (Appendix A-1)


 coefficient (L water/kg suspended sediment)
 

TSS = Total suspended solids concentration (mg/L) Equation 2-54
 

1 x 10-6 Units conversion factor (kg/mg)
 

The use of Equation 2-53 to calculate the concentration of the waste constituent dissolved in the water 

column is consistent with U.S. EPA (1994e) and U.S. EPA (1998b). The total suspended solids 

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concentration (TSS) is derived as a function of the soil and waste mass delivered to the surface water from 

the landfill and the background suspended solids concentration. The TSS is calculated as follows: 

(2-54) 

where: 

Default 

TSS = total suspended solids concentration (mg/L) calculated 

Aw = rate of soil and waste from the landfill (kg/acre/yr) Equation 2-51 nd 9 Q = volume of stream (L/yr) - for 2 order stream 3.3 x 10 (U.S. EPA 1994) stream 

A = area of the waste management unit (acres) Equation 2-12 or 2-13
 

SSb = background suspended solids concentration (mg/L) 10 ppm (U.S. EPA 1994a)
 

The total waste constituent concentration in the water column (Csw) in a second order stream is calculated 

by using the Equation 2-52. The surface water partition coefficient (Kdsw) is discussed below. 

2.3.4.1 Partitioning Coefficients for Suspended Sediment-Surface Water (Kdsw) 

Partition coefficients (Kd) describe the partitioning of a compound between sorbing material, such as soil, 

soil pore-water, surface water, suspended solids, and bed sediments. For organic compounds, Kd has 

been estimated to be a function of the organic-carbon partition coefficient and the fraction of organic 

carbon in the partitioning media. For metals, Kd is assumed to be independent of the organic carbon in 

the partitioning media and, therefore, partitioning is similar in all sorbing media. 

The soil-water partition coefficient (Kds) describes the partitioning of a compound between soil 

pore-water and soil particles, and strongly influences the release and movement of a compound into the 

subsurface soils and underlying aquifer. The suspended sediment-surface water partition coefficient 

(Kdsw) coefficient describes the partitioning of a compound between surface water and suspended solids 

or sediments. 

Organics For organics (including PCDDs and PCDFs), soil organic carbon is assumed to be the 

dominant sorbing component in soils and sediments. Therefore, Kd values were calculated using the 

following fraction organic carbon (fOC) correlation equations from Review Draft Addendum to the 

Methodology for Assessing Health Risks Associated with Indirect Exposure to Combustor Emissions 

(U.S. EPA 1993d): 

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(2-55) 

U.S. EPA (1993d), from literature searches, states that fOC could range as follows: 

C 0.05 to 0.1 in suspended sediments - for which a mid-range value of f = 0.075 oc,sw 

generally can be used. 

Consistent with the Region 6 Combustion Risk Assessment Protocol (U.S. EPA 1998b) guidance 

document, this DTSD uses mid-range foc values recommended by U.S. EPA (1993d). Kdsw values were 

calculated using Koc values recommended for each compound in this DTSD. 

The variables (Kdsw, TSS) in the Equation 2-53 are site-specific. Therefore, the use of the default values 

will introduce an under- or overestimation of Cdw. The degree of uncertainty associated with TSS is 

dependent on the suspended solids of the nearest second order stream and how this value compares to the 

default background suspended solids value (20 mg/L). Uncertainty associated with the variable Kdsw is 

associated with estimates of the fraction organic content (fOC). Because fOC values can vary widely for 

different locations in the same medium, using default fOC values may result in significant uncertainty in 

specific cases. 

2.3.4.2 Concentration of Waste Constituent in Fish (Cfish) 

The waste constituent concentration in fish is calculated using either a waste-specific bioconcentration 

factor (BCF) or a waste-specific bioaccumulation factor (BAF). For compounds with a log Kow less than 

4.0, BCFs are used. Compounds with a log Kow greater than 4.0 are assumed to have a high tendency to 

bioaccumulate, therefore, BAFs are used. Appendix A-1 provides a detailed discussion on the sources of 

the waste constituent-specific BCF and BAF values, and the methodology used to derive them. BCF and 

BAF values are generally based on dissolved water concentrations. Therefore, when BCF or BAF values 

are used, the waste constituent concentration in fish is calculated using dissolved water concentrations. 

The equations used to calculate fish concentrations are described in the subsequent subsections. 

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2.3.4.3 Calculation of Fish Concentration (Cfish ) from Bioconcentration Factors Using Dissolved 

Phase Water Concentration 

U.S. EPA OSW recommends the use of Equation 2-56 to calculate fish concentration from BCFs using 

dissolved phase water concentration. The use of this equation is further described in Appendix A-1. 

(2-56) 

where: 

Default 

Cfish = Concentration of waste constituent in fish calculated

 (mg/kg FW tissue) 

Cdw = Dissolved phase water concentration of waste Equation 2-53

 constituent (mg/L) 

BCF = fish Fish bioconcentration factor of waste constituent (L/kg) chem-specific (Appendix A-1) 

The dissolved phase water concentration (Cdw) is calculated by using Equation 2-53 above. 

Chemical-specific BCFfish values are presented in Appendix A-1. The use of Equation 2-56 to calculate 

fish concentration is consistent with U.S. EPA (1994e) and U.S. EPA (1998b) . 

2.3.4.4 Calculation of Fish Concentration (Cfish) from Bioaccumulation Factors Using Dissolved 

Phase Water Concentration 

U.S. EPA OSW recommends the use of Equation 2-57 to calculate fish concentration from BAFs using 

dissolved phase water concentration. 

(2-57) 

where: 

Default 

Cfish = Concentration of waste constituent in fish calculated

 (mg/kg FW tissue) 

Cdw = Dissolved phase water conc. of waste constituent (mg/L) Equation 2-53 

BAFfish = Fish bioconcentration factor of waste constituent (L/kg) chem-specific (Appendix A-1) 

The dissolved phase water concentration (Cdw) is calculated by using Equation 2-53. Chemical-specific 

bioaccumulation factor (BAF fish ) values are presented in Appendix A-1. 

2-47