EPA-453/R-94-080A
AIR EMISSIONS MODELS
FOR WASTE AND WASTEWATER
U.S. EPA Contract No. 68D10118
November 1994
prepared for the
U.S. Environmental Protection Agency
Office of Air Quality Planning and Standards
Research Triangle Park
North Carolina 27711
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TABLE OF CONTENTS
Chapter Page
1.0 INTRODUCTION 1-1
1.1 BACKGROUND 1-1
1.2 SCOPE 1-2
1.3 REPORT ORGANIZATION 1-3
2.0 DESCRIPTION OF PATHWAYS 2-1
2.1 GENERAL 2-1
2.2 VOLATILIZATION 2-2
2.3 ADSORPTION 2-5
2.4 MIGRATION 2-6
2.5 RUNOFF 2-7
2.6 BIOLOGICAL DECOMPOSITION 2-8
2.7 PHOTOCHEMICAL DECOMPOSITION 2-10
2.8 HYDROLYSIS 2-11
2.9 OXIDATION/REDUCTION 2-12
2.10 HYDROXYL RADICAL REACTIONS 2-13
2.11 REFERENCES 2-14
3.0 IMPORTANCE OF PATHWAYS 3-1
3.1 INTRODUCTION 3-1
3.2 THEORETICAL BASIS 3-1
3.2.1 Surface Impoundments 3-3
3.2.2 Aerated and Nonaerated Wastewater
Treatment 3-3
3.2.3 Land Treatment 3-3
3.2.4 Landfills 3-4
3.3 EMISSION MODELS 3-4
3.4 REFERENCES 3-7
4.0 COLLECTION SYSTEM AIR EMISSIONS 4-1
4.1 INTRODUCTION 4-1
4.2 COLLECTION SYSTEM EMISSION FACTORS 4-3
4.2.1 The use of emission factors 4-3
i
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4.2.2 Collection system units 4-5
4.2.3 A listing of emission factors 4-7
4.3 AIR EMISSION MODELS 4-12
4.3.1 Case Al. Air flow induced by waste flow
into drain 4-12
4.3.2 Case A2. Air flow in drain due to wind
pressure 4-14
4.3.3 Case A3. Airflow induced by density
differences 4-20
4.3.4 Case Bl. Manhole venting due to density
effects 4-24
4.3.5 Case B2. Manhole venting due to wind . . . 4-27
4.3.6 Case B3. Manhole venting due to wastewater
flow 4-30
4.3.7 Case Cl. Conduit air flow due to wind . . 4-34
4.3.8 Case C2. Conduit air flow induced by water
flow 4-38
4.3.9 Case C3. Conduit air flow due to
density differences 4-42
4.3.10 Case Dl. Stack vent air flow . . . . ,
4.3.11 Case D2. Stack vent working loss . . ,
4.3.12 Case D3. Trench volatilization loss . ,
4.3.13 Case El. J trap sealed with wastewater
with no wastewater flow ,
4.3.14 Case E2. J trap water sealed with
wastewater flow ,
4-45
4-49
4-51
4-54
4-57
4.3.15 Case E3. Lift station with periodic pumping of
wastewater 4-62
4.3.16 Case E4. Open surfaces in sumps 4-65
4.3.17 Case Fl. Primary clarifier weir 4-69
4.3.18 Case F2. Secondary clarifier weir .... 4-73
4.3.19 General weir model 4-73
4.4 REFERENCES 4-79
11
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5.0 SURFACE IMPOUNDMENTS AND OPEN TANKS 5-1
5.1 NARRATIVE DESCRIPTION OF EMISSIONS AND
MODEL UNITS 5-1
5.2 QUIESCENT SURFACES WITH FLOW 5-4
5.2.1 Emission Model Equations 5-4
5.2.2 Model Plant Parameters for Quiescent
Impoundments 5-12
5.2.3 Example Calculation for Storage
Impoundments 5-13
5.3 BI ODE GRADATION 5-17
5.3.1 Description of Biological
Active Systems 5-17
5.3.2 Rate of Biodegradation 5-23
5.3.3 Example Calculation for
Quiescent Impoundments 5-33
5.4 MECHANICALLY AERATED IMPOUNDMENTS AND
ACTIVATED SLUDGE UNITS 5-36
5.4.1 Emission Model Equations 5-36
5.4.2 Model Plant Parameters for Mechanically
Aerated Impoundments 5-39
5.4.3 Example Calculation for Mechanically
Aerated Treatment Impoundments 5-41
5.4.4 Example Calculation for Activated
Sludge Unit 5-48
5.5 DISPOSAL IMPOUNDMENTS WITH QUIESCENT SURFACES . . 5-48
5.5.1 Emission Model Equations 5-48
5.5.2 Model Plant Parameters for Disposal
Impoundments 5-52
5.5.3 Example Calculations for Disposal
Impoundments 5-52
5.6 DIFFUSED AIR SYSTEMS 5-58
5.6.1 Emission Model Equations 5-58
5.6.2 Model Unit Parameters for Activated
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Sludge Unit with Diffused Air 5-60
5.6.3 Example Calculation for Diffused
Air Activated Sludge Unit 5-61
5.7 OIL FILM SURFACES 5-62
5.8 DISCUSSION OF ASSUMPTIONS AND SENSITIVITY
ANALYSIS 5-64
5.8.1 Removal Mechanisms 5-64
5.8.2 Major Assumptions 5-67
5.8.3 Sensitivity Analysis 5-68
5.9 REFERENCES 5-74
6.0 WASTEWATER TREATMENT MODELS (WATERS) 6-1
6.1 UNITS FOR MODELING EMISSIONS OF VOLATILE COMPOUNDS . 6-1
6.1.1 Conventional Activated Sludge System .... 6-3
6.1.2 Sludge Handling 6-4
6.1.3 Conventional Activated Sludge
(Mechanical Aeration) 6-4
6.1.4 Conventional Activated Sludge
(Diffused Air: Coarse and Fine Bubble) 6-5
6.1.5 Aerated Lagoons (Mechanical Air) 6-6
6.1.6 Spray Evaporation Ponds 6-6
6.1.7 Dissolved Air Flotation (DAF) 6-7
6.1.8 Neutralization (Equalization)Process .... 6-7
6.1.9 Miscelaneous Physical-Chemical
Treatment Systems 6-7
6.2 AIR EMISSIONS OF VOLATILE COMPOUNDS FROM TRICKLING
FILTERS 6-8
6.3 AIR EMISSIONS OF VOLATILE COMPOUNDS FROM
COOLING TOWERS 6-11
6.3.1 Cooling Tower Default Parameters 6-14
6.3.2 Performance Data of Cooling Towers .... 6-14
6.3.3 Air Emission Modeling for Cooling Tower . . 6-14
6.3.4 Material Balance with Recycle 6-17
6.4 ESTIMATION OF AIR EMISSIONS FROM
iv
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API SEPARATOR UNITS 6-18
6.4.1 API Separator Model Elements 6-18
6.4.1.1 Region 1 flow distribution .... 6-19
6.4.1.2 Region 2 oil film separation . . . 6-21
6.4.1.3 Region 3 weir overflow 6-22
6.4.2 Example Calculation 6-25
6.4.2.1 Properties of benzene and unit
specifications 6-25
6.4.2.2 Region 1 calculations 6-25
6.4.2.3 Region 2 calculations 6-30
6.4.2.4 Region 3 Weir calculations .... 6-33
6.5 MODEL FOR PRETREATMENT UNITS 6-36
6.5.1 Pretreatment Equations 6-36
6.5.2 Pretreatment Examples 6-38
6.6 REFERENCES 6-43
7.0 LAND TREATMENT 7-1
7.1 NARRATIVE DESCRIPTION OF LAND TREATMENT
AND AIR EMISSIONS 7-1
7.2 LAND TREATMENT 7-4
7.2.1 Land Treatment Emission Model
Descriptions 7-4
7.2.1.1 Analytical Correlations 7-4
7.2.1.2 Biodegradation 7-8
7.2.1.3 Estimation of Equilibrium
Coefficient, Keq 7-9
7.2.1.4 Estimation of Effective
Diffusivity 7-11
7.2.1.5 Waste Partitioning 7-12
7.2.1.6 Emissions at Short Times 7-13
7.2.1.7 Estimating the Fraction
Emitted at Short Times 7-18
7.2.1.8 Estimating the Fraction
Emitted for Longer Times 7-22
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7.2.1.9 Tilling 7-24
7.2.1.10 Model Selection 7-28
7.2.2 Waste Application Model 7-30
7.2.3 Oil Film Model 7-30
7.2.4 Model Inputs 7-33
7.2.5 Estimation of Total Organic Compounds
Emissions 7-40
7.2.6 Example Calculations 7-41
7.2.6.1 Emissions from Land
Treatment Soil 7-41
7.2.6.2 Emissions from Waste Application .... 7-44
7.2.6.3 Emissions from an Oil Layer
on Soil Prior to Tilling 7-47
7.2.6.4 Examples of the Use of the Land Treatment
Model for Specific Cases 7-48
7.2.7 Assumptions and Sensitivity Analyses . . . 7-54
7.3 REFERENCES 7-56
0 LANDFILLS AND WASTEPILES 8-1
8.1 INTRODUCTION 8-1
8.2 CLOSED LANDFILLS 8-2
8.2.1 Emission Model Equations 8-2
8.2.2 Model Plant Parameters for Closed
Landfills 8-15
8.2.3 Example Calculation for Closed
Landfills 8-17
8.3 FIXATION PITS 8-23
8.3.1 Emission Model Equations 8-23
8.3.2 Model Plant Parameters for Fixation
Pits 8-30
8.3.3 Example Calculation for Fixation Pits . . . 8-31
8.4 OPEN LANDFILLS AND WASTEPILES 8-35
8.4.1 Emission Model Equations 8-35
8.4.2 Model Plant Parameters for Open
vi
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Landfills and Wastepiles 8-39
8.4.2.1 Parameters for Open Landfills 8-39
8.4.2.2 Parameters for Wastepiles 8-40
8.4.3 Example Calculation for Open
Landfill 8-42
8.5 REFERENCES 8-48
9.0 TRANSFER, STORAGE, AND HANDLING OPERATIONS 9-1
9.1 NARRATIVE DESCRIPTION OF MODEL PLANTS AND
EMISSIONS 9-1
9.2 CONTAINER LOADING 9-1
9.2.1 Emission Model for Container Loading .... 9-1
9.2.2 Model Parameters 9-2
9.2.3 Sample Calculation for Tank Loading .... 9-4
9.3 CONTAINER STORAGE 9-6
9.3.1 Emission Model for 55-Gal Drums,
Tank Trucks, and Railroad Tank Cars 9-6
9.3.2 Model Parameters for Drum Storage 9-7
9.3.3 Sample Calculations for Drum Storage .... 9-7
9.3.4 Emission Model for Open Dumpsters 9-7
9.3.5 Model Parameters for Open Dumpster
Storage 9-11
9.3.6 Sample Calculation for Open Dumpster
Storage 9-12
9.4 CONTAINER CLEANING 9-14
9.4.1 Emission Model for Container Cleaning . . 9-14
9.4.2 Model Parameters 9-15
9.4.3 Sample Calculation for Tank Truck
Cleaning 9-16
9.5 STATIONARY TANK LOADING 9-16
9.5.1 Emission Model for Stationary Tank
Model 9-16
9.5.2 Model Parameters 9-17
9.5.3 Sample Calculation for Tank Loading
vii
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Emission Model 9-18
9.6 STATIONARY TANK STORAGE 9-20
9.6.1 Model Description 9-20
9.6.2 Model Parameters 9-21
9.6.3 Sample Calculation for Tank Storage
Emission Model 9-21
9.7 SPILLS 9-23
9.7.1 Model Description 9-23
9.7.2 Model Parameters 9-24
9.7.3 Sample Calculation for Drum Storage
Model 9-25
9.8 FUGITIVE EMISSIONS 9-25
9.8.1 Emission Model for Fugitives 9-25
9.8.2 Model Parameters 9-25
9.8.3 Sample Calculation for Fugitive
Emission Model 9-26
9.9 VACUUM TRUCK LOADING 9-27
9.9.1 Emission Model for Vacuum Truck
Loading 9-27
9.9.2 Model Parameters 9-27
9.9.3 Sample Calculation 9-28
9.10 REFERENCES 9-29
10.0 COMPARISON OF MODEL RESULTS WITH FIELD TEST DATA . . . 10-1
10.1 INTRODUCTION 10-1
10.2 SURFACE IMPOUNDMENTS AND OPEN TANKS 10-1
10.2.1 Summary 10-1
10.2.2 Details of Comparisons 10-2
10.2.3 Recommendations for Additional Data . . . 10-20
10.3 LAND TREATMENT 10-23
10.3.1 Midwest Refinery—1985 (Case 1) 10-37
10.3.2 West Coast Refinery (Case 2) 10-37
10.3.3 Commercial Waste Disposal Test(Case 3) . . 10-39
10.3.4 Midwest Refinery—1979 (Case 4) 10-39
viii
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10.4 LANDFILLS AND WASTEPILES 10-39
10.5 TRANSFER, STORAGE, AND HANDLING OPERATIONS . . . 10-45
10.5.1 Models Documented in AP-42 10-45
10.5.2 Fugitive Emissions 10-45
10.5.3 Spillage 10-46
10.5.4 Open Dumpster Storage Emissions 10-46
10.6 REFERENCES 10-50
11.0 TECHNICAL SUPPORT FOR THE IDENTIFICATION OF
COLLECTION SYSTEMS AS MAJOR EMISSION SOURCES 11-1
11.1 INTRODUCTION 11-1
11.2 SUMMARY OF REFERENCES FOR AIR EMISSIONS FROM
COLLECTION SYSTEMS 11-2
11.3 ORGANIC COMPOUNDS WILL VOLATILIZE IN THE HEADSPACE
OF THE COLLECTION SYSTEM 11-17
11.3.1 Importance of Gas Concentrations 11-17
11.3.2 Industry Comments 11-17
11.3.3 Fingas, et al 11-18
11.3.4 Shell Petrochemical Facility 11-18
11.3.5 Ph.D. Dissertation of R. L. Corel 11-20
11.4 UNCONTROLLED WASTEWATER COLLECTION SYSTEMS CAN HAVE
SIGNIFICANT DISCHARGES OF HEADSPACE 11-23
11.4.1 Shell Facility 11-23
11.4.2 Velocity Screening at Pulp Mills 11-24
11.4.3 Rohm and Haas Chemical Plant
Collection System 11-25
11.4.5 Velocity Screening at Industrial Plants . 11-25
11.5 THE FRACTION OF ORGANIC COMPOUNDS LOST IN UNCONTROLLED
COLLECTION SYSTEMS 11-25
11.5.1 Shell Facility 11-23
11.5.2 Rohm and Haas Tracer Investigation .... 11-32
11.5.3 Amoco Refinery Material Balance 11-37
11.5.4 Amoco Refinery Vent Measurements 11-38
11.5.5 Tracer Testing at Coastal Eagle Point
ix
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Refinery 11-39
11.5.6 Ph. D. Dissertation of R. L. Corel .... 11-43
11.6 REFERENCES 11-44
APPENDIX A A GUIDE THROUGH THE LITERATURE A-l
APPENDIX B COMPREHENSIVE SOURCE LIST B-l
APPENDIX C PROPERTIES OF COMPOUNDS OF INTEREST C-l
x
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1.0 INTRODUCTION
1.1 BACKGROUND
This report was prepared for the U.S. Environmental
Protection Agency's (EPA's) Office of Air Quality Planning and
Standards (OAQPS) as part of the effort to develop air emission
models for hazardous waste treatment, storage, and disposal
facilities (TSDF). Basic to this effort is the determination of
the means by which organic compounds escape to the environment
from waste and wastewater.
Organic compounds in surface impoundments, land treatment
facilities, landfills, wastepiles, or wastewater collection and
treatment systems can depart through a variety of pathways,
including volatilization, biological decomposition, adsorption,
photochemical reaction, and hydrolysis. To allow reasonable
estimates of organic compounds disappearance, one must know which
pathways predominate for a given chemical, type of waste site,
and set of meteorological conditions.
Analytical models have been developed to estimate emissions
of organic compounds via various pathways from wastewater and
waste management units. Some of these models have been assembled
into a spreadsheet called CHEMDAT8 for use on an IBM PC, or
compatible, microcomputer. A user's guide for CHEMDAT8 is
included as a separate manual. Area emission sources for which
models are included on the diskette are as follows:
1-1
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• Nonaerated impoundments, which include quiescent
surface impoundments and open top WWT tanks;
• Aerated impoundments, which include aerated surface
impoundments and aerated WWT tanks;
• Disposal impoundments, which include nonaerated
disposal impoundments;
• Land treatment; and
• Landfills.
These models can be used to estimate the magnitude of site
emissions for regulatory purposes. Sample calculations using
each model are also included in this report.
A computer program "WATERS" is available for estimating the
fate of organic compounds in various wastewater treatment units,
including collection systems (Chapter 4), aerated basins (Chapter
5), and other units (Chapter 6). WATERS is written to run under
Microsoft's disk operating system DOS without the need to
purchase other programs (Windows or spreadsheets). WATERS
contains useful features such as the ability to link treatment
units to form a treatment system, the ability for recycle among
units, and the ability to generate and save site-specific
compound properties.
The terms "volatile" and "semivolatile" are used to describe
the tendency of an organic waste component to partition into the
headspace of the waste container. Waste constituents similar to
benzene and methylene chloride have relatively high vapor
pressures (>10 mm Hg) and relatively high Henry's law constants
(>10 mole fraction vapor/mole fraction liquid) and are considered
volatile. Other waste constituents similar to phenol do not have
high vapor pressures or Henry's law constants, but are considered
semivolatile because some part of the semivolatiles can be lost
to the atmosphere during waste handling and treating operations.
1-2
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1.2 SCOPE
This report briefly describes the chemical and physical
pathways for organic compounds and discusses their importance for
different types of sites and sets of conditions. Models
developed for estimating the relative magnitude of environmental
release in the presence of competing pathways are presented, and
physical
characteristics of the parameters that serve as inputs to the
models are identified.
The models provide an estimate of the relative magnitude of
organic compounds pathways on a compound-specific basis. Models
for aerated and nonaerated impoundments, lagoons, landfills,
wastepiles, and land treatment facilities have been installed in
an integrated spreadsheet program, CHEMDAT8, which allows a user
to calculate the partitioning of organic compounds among various
pathways depending on the particular parameters of the facility
of interest. The program is structured to allow new data (e.g.,
compounds and model facility parameters) to be added (see
CHEMDAT8 user's guide). The results of the calculated
partitioning may be used to identify those characteristics that
are important in determining relative organic compounds loss
rates.
Source variability will significantly influence the relative
importance of the pathways. For highly variable sources, it may
be possible to exclude insignificantly small pathways from
consideration. The relative magnitude of these pathways can
then be compared by applying the methodology to a model facility
to determine relative differences among various compounds.
1.3 REPORT ORGANIZATION
Chapter 2 describes each of the potential pathway mechanisms
that determine the fate of various chemical species. Chapter 3
discusses the importance of the pathways for surface impoundments
1-3
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and aerated and non-aerated WWT facilities, land treatment sites,
and landfills/wastepiles.
Chapter 4 presents air emission models that are applicable
to collection systems. A number of different collection system
elements are presented and a discussion of the use of the models
is provided.
Chapter 5 presents air emission models that are applicable
to conventional wastewater treatment units. A discussion of the
estimates of the effects of biological reactions on the air
emissions and water quality is presented. Recommendations for the
use of models are also presented.
Chapter 6 presents air emission models for trickling
filters, cooling towers, and API separators. In addition, this
section provides recommendations for the use of the air emission
models for a variety of waste management situations.
Chapters 7 and 8 describe the emission models applicable to
landtreatment and landfill sites. Models for estimating
emissions from transfer, storage, and handling operations are
described in Chapter 9. Chapter 10 compares treatment emission
model predictions with the field data that are available. Chapter
11 compares collection system model predictions with the field
data that are available from collection systems.
This report compares relative rates of organic compounds
destruction and volatilization to determine the most significant
pathways. The rate of organic compounds volatilization
destruction for any one pathway is calculated so that it can be
expressed as a fraction of the loss/destruction from all
pathways.
APPENDIX A presents an overview of the literature and
APPENDIX B presents comprehensive source list that includes
pertinent literature in addition to that cited in the sections
and Appendixes of this report.
1-4
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Properties of compounds of interest to TSDF pathways and
emission estimation are presented in APPENDIX C. A subset of
these compounds is a part of CHEMDAT8. The user's guide,
available separately, describes the procedures that are used in
estimating emissions using CHEMDAT8 or WATERS and other
procedures presented in the body of the report. The user's guide
also contains instructions for modifying CHEMDAT8 to include
additional compounds using the CHEM8 compound characteristics
presented in APPENDIX C.
1-5
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2.0 DESCRIPTION OF PATHWAYS
2.1 GENERAL
A pathway is considered here to be any process that removes
volatile organics from a site. The removal may be physical (as
in volatilization of a solvent from a surface impoundment) or
chemical (as in oxidation of an alcohol in a wastewater treatment
plant).
Pathways may be considered as rate processes, with rate
often strongly dependent on concentration of the disappearing
species and temperature of the system. Rates vary in order from
zero to mixed, with first order predominating at low
concentrations, that is:
= kc (2-1)
where,
c = concentration of disappearing substance, g/L;
t = time, s; and
kv = volatilization constant, s'1.
Half-life, the time required for one-half of the substance
to disappear, is a useful concept. It provides an easily
visualized measure of the time required for disappearance. For a
first-order rate process:
2-1
-------
tix, = (In2) k
_L / z
= 0.693 k
where, t1/2 = half-life, in seconds.
The half-life of a second-order equation is as follows:
t
-i
C
-i
(2-3;
where,
ks = second-order volatilization constant, L/(g»s); and
C0 = initial concentration, g/L.
Note that first-order half-lives are independent of initial
concentration while second order half-lives are not.
Much of the following material is taken from ICF.1 The
pathways described are physical (volatilization, adsorption,
migration, and runoff) and chemical (biological decomposition,
photochemical decomposition, hydrolysis, oxidation/reduction, and
hydroxyl radical reaction).
2.2 VOLATILIZATION
Volatilization occurs when molecules of a dissolved
substance escape to an adjacent gas phase. The driving force for
this process in nonturbulent liquids is molecular diffusion.
Equation (2-1) shows the rate of volatilization of an organic
chemical from water. For this case, the rate constant can be
estimated:2
k -
V
1
L
1
V
DiC]
Di°J
RT
+
(106) Hkw
.
-1
;2-4'
where,
L = mixing depth of water, cm;
k0! = mass transfer coefficient of oxygen in water, cm/s;
-------
D! = diffusion coefficient of the chemical (c) or oxygen (o)
in water, cm2/s;
m = liquid turbulence exponent, 0.5 to 1, dimensionless,
from TABLE 2-1;
R = ideal gas constant, atm cm3/ (mol»K) ;
T = temperature, K;
H = Henry's law constant, atm m3/mol;
kwg = mass transfer coefficient for water vapor in air,cm/s;
Dg = diffusion coefficient of the chemical (c) or water (w)
in air, cm2/s; and
n = gas turbulence exponent, 0.5 to 1.0, dimensionless,
from Table 2-1.
Equation (2-4) requires values of diffusion coefficients and
Henry's law constants. If tabulated values are not available,
the following estimations can be used. For the diffusion of a
chemical in air:3
D = 0.0067T1'5 (0.034+ M"1) °'5 M~°-17 [ (M/2.5d) °'33+1.81T2 (2-5)
where,
T = temperature, degrees Kelvin;
M = molecular weight of chemical, g/g mol;
d = density of liquid chemical, g/cm3.
For diffusion coefficients in water:
D1 = 1.518 (10~4) Vcm°-6 (2-6)
where Vcm = molar volume of chemical, cm3/g mol.
This equation assumes the system temperature to be 300 deg.
K. For other temperatures, a more rigorous form of the equation
should be used, as in Perry.4 Molar volume is estimated as the
ratio of molecular weight to liquid density at room temperature.
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TABLE 2-1. VALUES OF CONSTANTS FOR USE IN EQUATION 2-45
Value
Constant
Rivers
Lakes
L (cm)
k°1 (cm°s~
m
T (K)
RT
K«g
n
200
0.0022
0.7
293
2.40 x ID'2
0.58
0.7
200
0.0005
1.0
293
2.40 x 10'
0.58
0.7
2-4
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If ideal gases and solutions are assumed, Henry's law
constant can be estimated from:
H = P/(14.7s) (2-7)
where,
P = pure component vapor pressure, psia; and
s = solubility of chemical in water, g mol/m3.
Values for other terms in Equation 2-4 have been tabulated
by ICF and are given in TABLE 2-1.
In general, equations are available to estimate
volatilization from wastewater treatment systems and surface
impoundments.6'7 In the case of land treatment and landfills, the
models for volatilization are much less well developed and the
supporting data are more limited than those of the aqueous
systems. The rate of volatilization at a soil-air interface is a
function of the concentration and properties of the escaping
chemical, soil properties (moisture, temperature, clay, and
organic content), and properties of the air at soil level
(temperature, relative humidity, and wind speed).8
2.3 ADSORPTION
Adsorption takes place when molecules of a dissolved
chemical (in a liquid-solid system) become physically attached to
elements of the solid phase. Chemical bonding may also occur
(chemisorption). An example of adsorption is molecules of
solvent being sorbed by particles of silt in a surface
impoundment. If the adsorptive capacity of the solid material is
reached, no further net sorption will occur. With reductions in
concentration in the bulk liquid of the chemical being sorbed
(adsorbate), desorption may take place. The amount of material
adsorbed depends on (1) the concentration of adsorbate, (2) the
amount of solid phase (adsorbent), and (3) the temperature. For
systems with constant adsorbent properties, primarily surface
area per unit mass, the amount of material adsorbed at a
2-5
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particular concentration and temperature is proportional to the
mass of adsorbent. For example, the Freundlich adsorption
isotherm equation allows prediction of amount adsorbed as
follows:
- = KfC1/n (2-8)
m r
where,
x = mass of chemical adsorbed, g;
m = mass of adsorbent, g;
Kf = Freundlich adsorption coefficient, (g sorbate/
g sorbent)/(g sorbate/g solution);
C = concentration of chemical in solution at equilibrium,
g sorbate/g solution; and
n = empirical constant, ranging from 0.7 to 1.1, typically
1.0 for soils, dimensionless.
A Langmuir adsorption isotherm can be derived from a kinetic
rate theory describing the adsorption and desorption rates. The
rate of adsorption is proportional to the rate of collisions
between adsorbate molecules and free adsorbent surface. The rate
decreases with lowering adsorbate concentration and with
decreasing surface sites available for adsorbing molecules. The
following rate equation applies:
Rate of adsorption = k]_C (1-f) (2-9)
where,
kx = rate constant for adsorption, g/s;
f = fraction of adsorption sites occupied, dimensionless;
and
C = concentration of chemical in solution at equilibrium,
g sorbate/g solution.
For desorption:
Rate of desorption = k f (2-10)
2-6
-------
where,
f = fraction of adsorption sites occupied; and
k2 = rate constant for desorption, g/s.
At equilibrium the two rates are equal, and
kc
,
k2
(2-1D
Adsorption rates are usually rapid compared to the other
processes discussed here. However, mass transfer limitations may
reduce effective rates, especially for poorly mixed systems.
Lack of sorbent and its saturation may also reduce the
effectiveness of adsorption.
For estimating adsorption partitioning, a linear
relationship is assumed (n = 1 or kxC « k2) . The equilibrium
relationship for biomass is estimated from an equation of Matter-
Muller,9 based on the logarithm of the octanol-water partition
coefficient, LOW. For land treatment and land-fills, the only
partitioning of importance to fate predictions is gas-liquid
partitioning.
2.4 MIGRATION
Migration occurs when chemicals applied to soils are
transported through the soils to groundwater. Leaching and
percolation are the mechanisms that physically remove chemical
molecules from a point of deposit and carry them toward a water
table. Capillary flow is a resisting mechanism that moves the
molecules upward through the soil. The leachability of a
chemical is a function of soil texture and cation exchange
capacity, amount of soil organic content, amount and intensity of
rainfall, and mechanical placement and adsorptive properties of
the chemical.10
2 . 5 RUNOFF
Chemicals at or near the soil may be washed away by rain.
The rate depends on soil and chemical characteristics and on
2-7
-------
rainfall rates and frequency. Clark, Viessman, and Hammer11
state that runoff in any drainage area is a function of climate
and the physical characteristics of the area. Significant
factors include precipitation type; rainfall intensity, duration,
and distribution; storm direction; antecedent precipitation;
initial soil moisture conditions; soil type; evaporation;
transpiration; and, for a given drainage area, its size, shape,
slope, elevation, directional orientation, and land use
characteristics. If rainfall is heavy shortly after application
of a chemical, runoff and erosion can physically remove it. The
chemical may be dissolved in runoff water, carried along by it,
or adsorbed on eroding soil particles that move with runoff. For
pesticide applications, about 3 to 10 percent of the applied
material appears in runoff water. Below a certain intensity,
rainfall will promote leaching of nonadsorbed chemical into the
ground rather than result in runoff.
2.6 BIOLOGICAL DECOMPOSITION
Biological decomposition takes place when microbes break
down organic compounds for metabolic processes. The rate of
decomposition depends on the structure of the compound and on the
needs of the microbes. If the compound is present in excess, the
rate of population increase is as follows:
dx/dt = Rx (2-12)
where,
t = time, sec;
x = concentration of biomass, g/L; and
R = specific growth rate coefficient, s'1.
-------
If the compound is present in limited amount, the rate becomes a
hyperbolic saturation function of the compound (substrate)
concentration:12
x =RmaxS/(Ks+S) (2-13)
where,
Rmax = maximum specific growth rate coefficient (where
substrate is in excess) s'1;
S = concentration of substrate, g/L; and
Ks = substrate concentration at which the rate of substrate
utilization is one-half of the maximum rate, g/L.
Because the microbial population increases at the expense of the
compound, the growth rate is proportional to the compound's rate
of disappearance. The rate process may be of zero, first, or
mixed order depending on concentration of the substrate. In the
presence of multiple substrates, kinetics become complex.
For the case of S much greater than Ks, the equation
approaches zero order, and Equation (2-13) becomes:
(2-14)
For cases where S is much less than Ks, the equation
approaches first order:
(2-15;
with Rmax/Ks being the first-order rate constant.
For intermediate values of S, the equation is mixed order,
with the order dependent on values of the constants Rmax and Ks.
2-9
-------
2.7 PHOTOCHEMICAL DECOMPOSITION
Photochemical decomposition may occur in two ways. A
chemical may absorb light and react (direct photolysis), or the
chemical may react because of light absorption by surrounding
elements (indirect photolysis).
For direct photolysis, the rate of reaction of a dilute
solution of chemical in pure water is as follows:
Kp = b • • •. I. [C] (2-16)
where,
kp = rate of direct photolysis, g/(L s);
b = unit conversion constant, 3.8 x 10~21 g mol cm4/
(L photon);
• = reaction quantum yield, dimensionless;
•. = light absorption coefficient at wavelength
interval •, L/(g mol°cm) ;
I. = light flux at wavelength interval •,
photons/ (cm3°s) ; and
C = concentration of the chemical in water g/L.
Lyman13 refers to Zeep and Cline;14 Zepp;15 and Mabey, Mill,
and Hendry16 for details of rate calculations in aquatic systems.
In these systems, the rate constant Kp varies with the
distribution of sunlight and its intensity. Time of day, season,
cloud cover, and latitude all affect kp so that a reference
condition must be stated; e.g., a light flux of photons per
second corresponding to a cloudless yearly average at a latitude
of 40°N.
Reactions may be photocatalyzed. For example, a Ti02
catalyst can be photoexcited by light at wavelengths less than
360 mm. Ollis17 examined the degradation of halogenated
2-10
-------
hydrocarbons with this catalyst and found a rate equation of the
form:
1
(2-17)
dC/dt k k K, C
P P b
where,
kp = photolysis rate constant, g chemical/ (L°s°g
catalyst) ; and
Kb = apparent binding constant of a reaction
intermediate adsorbed on the illuminated catalyst
surface, L/g chemical.
For 11 halocarbons, values of k ranged from 5.8 x 10~8 to 2.3 x
10~6 g/LoS°g of catalyst, with most about 2.8 x 10~7 to 1.7 x 10~6.
A twelfth halocarbon had a k value of 2.3 x 10~4. Values of kb
for the 12 compounds ranged from 2 to 20 L/g.
2.8 HYDROLYSIS
Hydrolysis occurs when a chemical reacts with water. For
organic compounds, the reaction usually replaces a functional
group (X) with a hydroxyl : 18
RX +H20 = ROH +HX . (2-18)
Reaction rate constants may be pH-dependent ; for a specific
pH:
~] (2-19)
where,
kH = first-order hydrolysis rate constant, s'1;
ka = second-order rate constant for acid-promoted
hydrolysis, L/ (g mol°s);
[H+] = hydrogen ion concentration, g mol/L;
kn = first-order rate constant for pH-independent
neutral hydrolysis, s'1;
2-11
-------
kb = second-order rate constant for base-promoted
hydrolysis, L/(g mol°s); and
[OH~] = hydroxyl ion concentration, g mol/L.
Equation (2-19) can be transformed to:
kw= [H+] [OH-] (2-20)
where kw = ionization constant for water ~ 10~14 g mo!2/L2.
The rate constant kH depends on system pH and on the relative
values of ka, kb, and kn.
kH=ka[H+] +kn+kbkw/[H+] . (2-21)
2.9 OXIDATION/REDUCTION
Organic compounds in aquatic systems may be oxidized by
oxygen (particularly as singlet oxygen, 102) or other oxidants
such as hydroxyl radicals (OH) and peroxy radicals (R02) . The OH
radicals tend to be very reactive, but present only in low
concentrations. The R02 radicals are less reactive than the OH
radicals, but are present in greater concentrations. Singlet
oxygen is highly reactive, but also selective. It has an affinity
for electron-rich structures such as dienes and substituted
olef ins .
The oxidation rate can be calculated as:19
T = C **>2 [*°2] + kso [1Q2] + kx [*] (2-22)
where,
k[R02] = rate constant for peroxy radicals, L/ (g mol-s);
[R02] = concentration of peroxy radicals, g mol/L;
kso = rate constant for singlet oxygen, L/ (g mol-s);
[102] = concentration of singlet oxygen, g mol/L;
kx = rate constant for "other" oxidants, L/ (g mol-s); and
2-12
-------
[X] = concentrations of "other" oxidants, g mol/L.
In anaerobic environments, reduction reactions may take
place. Organochlorines are particularly affected. The reduction
rate can be calculated as:20
~ = C£i k. [R.] (2-23)
where,
k± = rate constant for reductant i, L/g mol.s; and
[RJ = concentration of reductant i, g mol/L.
2.10 HYDROXYL RADICAL REACTIONS
Hydroxyl radical reactions may occur through addition of a
hydroxyl radical, abstraction of a hydrogen atom, or both. In the
addition, reaction molecules with high electron density portions
attract electrophilic hydroxyl radicals. Hydrogen abstraction
takes place when a carbon-hydrogen bond in an organic molecule is
easily broken; it is controlled by electronic configuration and
number of hydrogen reactions in the molecule. The rate constant
for the reaction is often in the range of 6 to 60 x 108
L/(g mol-s).
A hydroxyl radical reaction rate can be calculated as:21
1— = C kOH [OH' ] (2-24)
where
kOH = rate constant for hydrogen abstraction or hydroxyl
addition, L/(g mol-s).
2-13
-------
2.11 REFERENCES
1. ICE, Inc. The RCRA Risk-Cost Analysis Model Phase III
Report, Appendix E. Chemical and Physical Processes
Affecting Decay Rates of Chemicals in Aquatic Environments.
Draft. Economic Analysis Branch, U.S. Environmental
Protection Agency Office of Solid Waste. Washington, DC.
1984.
2. Reference 1, p. E-18, Equation (14).
3. Spivey, J. J., C. C. Allen, D. A. Green; J. P. Wood, and R.
L. Stallings. Preliminary Assessment of Hazardous Waste
Pretreatment as an Air Pollution Control Technique. Research
Triangle Institute. Research Triangle Park, NC. EPA Contract
No. 68-03-3149, Task 12-5. 1984.
4. Perry, R. H., and C. H. Chilton. Chemical Engineers'
Handbook, Fifth Edition. New York, McGraw-Hill. 1973.
5. Reference 1, p. E-18 - E-19.
6. Allen, C. C., D. A. Green, and J. B. White. Preliminary
Assessment of Aerated Waste Treatment Systems at
TSDFs--Phase I. Draft. Research Triangle Institute. Research
Triangle Park, NC. EPA Contract No. 68- 03-3149, Task
54-01F. 1985.
7. Farino, W., P. Spawn, M. Jasinski, and B. Murphy. Evaluation
and Selection of Models for Estimating Air Emissions from
Hazardous Waste Treatment, Storage, and Disposal Facilities.
GCA/Technology. EPA 450/3-84-020. 1984.
8. Hornick, S. B. In: Land Treatment of Hazardous Waste, Parr,
J. F. (ed). Noyes Data Company. Park Ridge, NJ. 1982.
9. Matter-Muller, C., W. Gujer, W. Geiger, and W. Stumm. The
Prog. Wat. Tech. (Toronto) 12:299-314. lAWPR/Pergamon Press,
Ltd., Great Britain. 1980.
10. Reference 8.
11. Clark, J. W., W. Viessman, Jr., and M. J. Hammer. Water
Supply and Pollution Control. Scranton, PA, International
Textbook Company. 1971.
12. Reference 1, p. E-16, Equation (11).
2-14
-------
13. Lyman, W. J., et al. Research and Development of Methods for
Estimating Physicochemical Properties of Organic Compounds
of Environmental Concern. Phase II, Part I. NTIS AD 11875A.
1981.
14. Zepp, R. G., and D. M. Cline. Rate of Direct Photolysis in
Aquatic Environment. Environ. Sci. Technol. 11(4):359-366.
1977.
15. Zepp, R. G. Quantum Yields for Reaction of Pollutants in
Dilute Aqueous Solution. Environ. Sci. Technol.
12(3):327-329. 1979.
16. Mabey, W. R., T. Mill, and D. G. Hendry. Photolysis in
Water. In: Laboratory Protocols for Evaluating the Fate of
Organic Chemicals in Air and Water. Draft. U.S.
Environmental Protection Agency. EPA Contract 68-03-2227.
1980.
17. Ollis, D. F. Contaminant Degradation in Water. ES&T.
19_(6) :480-484. 1985.
18. Reference 13.
19. Reference 1, p. E-12, Equation (2).
20. Reference 1, p. E-12, Equation (3).
21. Reference 1.
2-15
-------
3.0 IMPORTANCE OF PATHWAYS
3.1 INTRODUCTION
The importance of the nine pathways described in Chapter 2.0
for surface impoundments, open tanks, land treatment facilities,
landfills, and wastepiles is described in this section. The
discussion centers on the pathways used in the emission models
described in subsequent sections. The pathways described in
Chapter 2.0 are repeated below for convenience:
• Volatilization
• Adsorption
• Migration
• Runoff
• Biological decomposition
• Photochemical decomposition
• Hydrolysis
• Oxidation/reduction
• Hydroxyl radical reaction.
Section 3.2 presents the relative importance of these
pathways based on the theoretical discussions appearing in
Chapter 2.0, the data appearing in the literature, and
engineering judgment. Section 3.3 summarizes in tabular form the
results of the emission model analyses in Chapters 5.0 through
6.0 and the pathways forming the basis for the emission models.
3.2 THEORETICAL BASIS
3-1
-------
The relative importance of the nine pathways for TSDF is
discussed in the following text and summarized in Table 3-1.
TABLE 3-1. PATHWAYS FOR HAZARDOUS WASTE AREA EMISSION SOURCES5
Wastewater
treatment plants
Surface Non- Land
Pathway impoundments Aerated aerated treatment Landfill
Volatilization
Bi ode gradation
Photodecomposition
Hydrolysis
Oxidation/
reduction
Adsorption
Hydroxyl
radical
reaction
Migration13
Runoff b
I
I
S
S
N
N
N
N
N
I
I
N
S
N
S
N
N
N
I
I
N
S
N
S
N
N
N
I
I
N
N
N
N
N
N
N
I
S
N
N
N
N
N
N
N
I = Important.
S = Secondary.
N = Negligible or not applicable.
a Individual chemicals in a given site type may have dominant
pathways different from the ones shown here.
b Water migration and runoff are considered to have negligible
effects on ground and surface water in a properly sited,
operated, and maintained RCRA permitted hazardous waste
treatment, storage, and disposal facility.
3-2
-------
These data were used as the basis for the emission models
contained in CHEMDAT8.
Results of exercising these models to identify pathways of
importance are discussed in Chapters 4.0 through 10.0 and are
summarized in Section 3.3. A short discussion of the theoretical
basis for pathways selection follows.
3.2.1 Surface Impoundments
Data reported by ICF show predominant removal mechanisms and
halflives for 71 chemicals. Table 3-2 lists the mechanisms and
statistics for six surface water pathways. Average half-lives
range from about 1/2 to 8 days, with predominant mechanisms being
volatilization and biodegradation. The rate of photo-
decomposition depends on the depth of the surface impoundment.
The rate is negligibly low for depths as great as 3 meters and is
indicated in Table 3-1 as S for a secondary effect.
3.2.2 Aerated and Nonaerated Wastewater Treatment
As in the case of the surface impoundments, volatilization
and biodegradation are potentially significant mechanisms. The
relative rates of these mechanisms depend on the particular
component and treatment system. Photodecomposition is not
expected to be a significant pathway due to the opacity of the
system, the depth of the liquid, and the residence time of the
processes. Adsorption is not expected to be significant except
for large loadings of suspended solids and oils in the
wastewater. The concentrations for many organic compounds are
expected to be roughly the same in the biomass as in the aqueous
phase.
3.2.3 Land Treatment
Based on available emission data and literature sources,
volatilization and biodegradation are expected to be important in
land treatment.2'6 For highly volatile constituents,
volatilization is expected to be the predominant pathway; for low
volatile constituents, biodegradation is expected to be the
3-3
-------
predominant pathway. Adsorption of organic compounds onto
organic carbon in the soil also occurs at land treatment sites.
However, calculations of land treatment air emissions both with
and without consideration of adsorption show a difference of only
10 percent. Therefore, adsorption is not considered a major
pathway for organics removal.
The method of waste application and incorporation into the
soil influence the importance of photochemical reactions in the
degradation of organic wastes in land treatment facilities.7
Photodecomposition can occur in land treatment between
application and tilling (usually 24 hours), although exposure to
sunlight is limited to daylight hours. While exact rates of
photodegradation are not known, they are expected to be low. The
oil in which the hazardous materials are suspended is semiopaque
to sunlight, which would tend to keep photodecomposition low.
After tilling, photodegradation is nonexistent because sunlight
does not penetrate the soil surface.8 Consequently,
photodecomposition is not expected to be significant.
3.2.4 Landfills
Volatilization is expected to be a primary organic compounds
pathway for landfills. Biodegradation is expected to be
negligible for hazardous waste landfills. The toxic properties
of the water are expected to inhibit biological processes and
therefore biodegradation.9
Rates of diffusion in the gas phase may be important.
Components can diffuse through unsaturated soils (air pockets
present). Control of liquid infiltration into the landfill is
expected to keep migration into the soil at a negligible level.
3.3 EMISSION MODELS
3-4
-------
Based on the exercise of CHEMDAT8 in predicting and
comparing pathways for TSDF processes, the pathways shown in
Table 3-3 are used as the basis of the models. Insignificant
emissions or inadequate data upon which to develop the model
relationships are the principal reasons for limiting the models
to the pathways shown in Table 3-3.
It should be noted that CHEMDAT8 includes provisions to
activate the unused pathways should further investigations and
field tests indicate the desirability of incorporating additional
pathways in the emission models.
TABLE 3-2. STATISTICS FOR SURFACE WATER PATHWAYS
Pathway
Vola- Bio-
tiliza- degrada-
tion tion
Range of 0.9-15 0.04-96
half-lives,
days
Average 2.24 8.05
half-life
Standard 2.85 19.4
deviation
Number of 38 26
chemicals
Photo-
decompo-
sition3
0.04-900
76.3
1.37
259.0
1.82
12
Oxida-
tion/
Hydro- reduc-
lysis tion Adsorption
0.0003-35 0.1-5 0.04-1.5
5.39 2.05 0.55
10.8 2.40 0.83
11 43
^Statistics are given for chemicals with and without an outlier.
3-5
-------
TABLE 3-3. PATHWAYS FOR TSDF SITES
Type of facility
Pathways included in model
Quiescent storage and treatment impoundments Volatilization
Mechanically aerated impoundments
Quiescent disposal impoundments
Land treatment facilities
Closed landfills
Active landfills
Wastepiles
Volatilization
Biodegradation
Volatilization
Volatilization
Biodegradation
Volatilization
(diffusion
through cap)
Barometric
pumping
Volatilization
(diffusion
through waste)
Volatilization
3-6
-------
3.4 REFERENCES
1. ICE, Inc. The RCRA Risk-Cost Assessment Model Phase III
Report, APPENDIX E. Chemical and Physical Processes
Affecting Accurate Rates of Chemicals in Aquatic
Environments. Draft. Economic Analysis Branch, U.S.
Environmental Protection Agency Office of Solid Waste.
Washington, DC. 1984.
2. American Petroleum Institute. Land Treatment--Safe and
Efficient Disposal of Petroleum Waste. Undated.
3. Bossert, li, et al, Fate of Hydrocarbons During Oily Sludge
Disposal in Soil. Applied and Environmental Microbiology.
47_(4) :163-161 . 1984.
4. Pelter, P. Determination of Biological Degradability of
Organic Substances. Water Research. 10:231-235. 1976.
5. DuPont, R. Ryon, and J. A. Reinemon (Utah Water Research
Laboratory). Evaluation of Volatilization of Hazardous
Constituents at Hazardous Waste Land Treatment Sites.
Prepared for U.S. Environmental Protection Agency. Ada, OK.
August 1986. 157 p.
6. Eklund, B. M., T. P. Nelson, and R. G. Wetherold (Radian
Corporation). Field Assessment of Air Emissions and Their
Control at a Refinery Land Treatment Facility. Prepared for
U.S. Environmental Protection Agency. Cincinnati, OH. DCN
86-222-078-15-07. September 12, 1986. 330 p.
7. Kaufman, D. D. Fate of Toxic Organic Compounds in
Land-Applied Wastes. In: Land Treatment of Hazardous Wastes,
Parr, J. F., et al. (eds). Park Ridge, NJ, Noyes Data
Corporation. 1983. p. 77-151.
8. Reference 7.
9. Shen, T. T. Estimation of Hazardous Air Emissions from
Disposal Sites. Pollution Engineering, pp. 31-34. August
1981.
3-7
-------
4.0 COLLECTION SYSTEM AIR EMISSIONS
4.1 INTRODUCTION
This chapter presents the methods used to estimate air
emissions from wastewater collection systems. Air emission
factors are developed that can be used to predict the release of
volatiles to the atmosphere from liquid wastes discharged in the
waste collection system. As a waste stream containing a volatile
waste constituent is discharged into a collection system, the
volatile constituent can be emitted into the atmosphere through
the mechanism of mass transfer to the air flowing through the
collection system. Air can enter and leave a collection system
by openings in drains, open channels, channels with grates,
openings in manhole covers, junction boxes, sumps, and other
openings. Estimation of the flow of air in a collection system
unit (drain, manhole) relative to the flow of wastewater flowing
under the collection system unit permits an estimation of the
fraction of the volatile constituent lost to the atmosphere as it
passes under the unit.
The assumptions that were made to characterize chemical
collection conduit designs include the following:
• The design depth in the drain channel is assumed to be
half full.
• The flow in the channel for estimating fractional
emissions is assumed to be 80 percent of design depth.
(Lower depths result in higher emissions.)
• The air exiting the system is assumed to be at
equilibrium with the volatiles in the channels.
4-1
-------
• A typical wind velocity is assumed to be 1.6 m/s
(3.5 MPH).
The assumption of equilibrium in wastewater collection units
is considered an appropriate approximation for national emission
estimates. For certain site specific emission estimates mass
transfer may be a more suitable method. Additional information
concerning the unit to be modeled will be needed if the mass
transfer approach is taken.
The emission factors for the collection units are sensitive
to the magnitude of the flow rates in the channels. The loss of
volatiles in the channels could be less than the equilibrium
amount if the rate of mass transfer from the bulk of the
wastewater to the air was slow enough. This mass transfer rate
is expected to be sensitive to the depth in the channel, with
equilibrium not achieved for high flows of air across deep
channels. For the case of channel depths at a fraction of the
design depths and relatively low air rates (manhole covers and
enclosed collection systems), the assumption of equilibrium is
expected to be appropriate.
The assumption of equilibrium in wastewater collection units
is expected to be more accurate for systems with restricted
headspace ventilation. Mass transfer is not expected to be the
rate controlling mechanism in the situation of restricted
ventilation, and the assumption of equilibrium will limit air
emissions to the equilibrium value.
Since the air emission factors are sensitive to
environmental factors such as temperature, humidity, and wind
pressure, the measured air emissions from wastewater collection
systems are expected to be variable. Monte Carlo methods are
used to simulate the effect of variable environmental factors on
a waste collection system. Because of the degree of difficulty
in performing the Monte Carlo calculations with site specific
calculations, a short-cut technique using unit emission factors
4-2
-------
is presented here. The fraction lost from waste passing under
specific units is estimated for common unit types and waste
constituents of varying volatility.
For the model systems a temperature difference of 5 degrees
Celsius was chosen as a temperature difference between the
ambient air and the collection system temperature. This
temperature difference was used to estimate gas flows due to air
density differences. The actual temperature differences would be
site specific.
4.2 COLLECTION SYSTEM EMISSION FACTORS
4.2.1 The Use of Emission Factors
The emission factors developed in this document are
expressed in terms of the fraction of material in the collection
conduit main emitted per unit. The collection conduit is the
subsurface pipe or covered trench that the wastewater flows in by
gravity from unit to unit on the path of the waste to the
wastewater collection system. When the path of the waste placed
in the collection system is specified, the amount of material
remaining in the original waste stream is recalculated each time
the waste flows under a unit with a potential emission source
(drain connection, manhole, lift station, sump, etc.):
Emissions from unit = amount present x unit emission factor
New amount present = amount present - emissions from unit.
Table 4-1 illustrates how the toluene emissions from a waste
discharge into a collection system can be estimated. The waste
flows into an open trench drain. Forty feet downstream,
additional waste flows into the trench for an additional 20 ft.
The flow in the trench discharges into a drain. The subsurface
channel in the collection conduit has an additional drain
connection and a manhole before discharge into a covered sump
with a vent.
This application of the unit emission factors to a
wastewater collection system for toluene wastes indicates that a
4-3
-------
substantial fraction of the original toluene in the waste can be
lost due to mass transfer to the air that flows in the collection
TABLE 4.1 EXAMPLE OF THE USE OF EMISSION FACTORS
Unit
Emission
factor
Amount
present,
(g)
Emissions,
(g)
Open trench drain (40 ft)
Open trench drain (20 ft)
Drain
Drain connection
Manhole at junction
Covered sump with vent
Overall collection units
0
0
0
0
0
0
0
.045
.022
.08
.08
.0083
.11
.30
100
95.
93.
85.
79.
78.
70
5
4
9
2
5
4.
2.
7.
6.
0.
8.
30
5
1
5
7
66
6
system. Another way of interpreting these data is that for every
70 g of toluene that enter the wastewater treatment plant, 30 g
are emitted in the collection system before the waste reaches the
wastewater treatment plant (43 percent).
These emission factors for wastewater collection systems
are not expected to be applicable for all systems. They are for
a wastewater collection system designed to aerate the wastewater,
either for safety, for corrosion reduction, or for odor control.
There are a number of equipment changes that can reduce the air
emissions to levels much lower than can the system presented
here. Emissions can be reduced by using covers for sumps,
manhole covers with fewer and smaller openings, seals on drain
openings, or solid metal covers for trenches; by purging the
system with excess water; and by other methods. Increasing the
external wind speed will increase emissions from the collection
4-4
-------
systems, according to the models. The collection system air
emissions can be increased by any discharge of steam into the
collection conduit, the presence of open sumps or open junctions,
and the presence of a complex collection systems with many units
(potential emission sources) before discharge.
Computational techniques that have been used to improve the
accuracy of the estimates of emission factors include considering
mass transfer at the liquid-gas interface and using Monte Carlo
simulation of collection system characteristics.
4.2.2 Collection System Units
Ten cases for induced airflow in collection conduits are
illustrated with cases A1-D1. Cases Al, A2, and A3 illustrate
potential airflows from process drains. Cases Bl, B2, and B3
illustrate air emissions from manholes. Cases Cl, C2, and C3
illustrate airflow out of collection conduit lines. Case Dl
represents emissions from a covered sump with an open vent, and
Case D2 illustrates airflow out of drain grates. The following
brief explanations describe some of the assumptions used in
estimating the induced flow of air for each of these units:
Case Al estimates airflow into a drain annulus induced
by water flow. The air drawn in will escape somewhere
and be in equilibrium with the water at that point.
Case A2 estimates airflow into a collection conduit
through a drain annulus. No water is flowing into the
drain. The air comes to equilibrium with the water
flowing in the collection conduit and escapes at some
point upstream or downstream of the drain.
Case A3 estimates airflow from saturated air rising
from a drain annulus due to a density difference
between the air in the collection conduit and the
ambient air. No water is flowing through the drain.
The air is drawn in at a point upstream or downstream
of the drain and reaches thermal and chemical equi-
librium with the wastewater flowing in the collection
conduit by the time it reaches the drain.
4-5
-------
• Case Bl estimates airflow from manhole cover vents
caused by a density difference between air in the
collection conduit and the ambient air. The air
flowing out of the vents is in thermal and chemical
equilibrium with the water flowing in the collection
conduit at that point.
• Case B2 estimates airflow through manhole cover vents
induced by wind blowing in the upstream end of a
collection conduit that is blocked off after the
manhole. The air is in equilibrium with the water in
the manhole.
• Case B3 estimates the airflow from manhole cover vents
induced by wind blowing in one end of a collection
conduit and flowing past the manhole to some point
downwind. The air is in equilibrium with the water in
the collection conduit at the manhole. No drains or
vents are in the line between the upwind collection
conduit end and the manhole.
• Case Cl estimates the airflow induced by wind blowing
in one end of a collection conduit and out another.
The air is in equilibrium with water at the downwind
end of the collection conduit.
• Case C2 estimates the airflow into the collection
conduit from a junction box induced by water flow
through the junction box. This air escapes somewhere
(e.g., the next junction box downstream) in equilibrium
with the water flowing through at that point.
• Case C3 estimates airflow from the discharge end of a
partially filled collection conduit resulting from
density differences between the ambient air and the
warm humid air in equilibrium with the wastewater.
• Case Dl estimates the airflow induced through a stack
on an enclosed sump. Air is in equilibrium with the
wastewater and is drawn into the system at some point
upstream or downstream of the sump.
• Case D2 estimates airflow from an open trench based
upon mass transfer in the rapid flowing water.
Enviromega of Burlington, Ontario, Canada recently measured
air emissions from a laboratory simulation of industrial
wastewater collection system elements1. The new measurement data
4-6
-------
has been used to test and refine existing analytical procedures
used to estimate emissions from air and water sources. Models
for four separate cases are developed on the basis of data
collected by Enviromega. The models are as follows:
• Case El estimates air emissions from wastewater in a J
trap without wastewater flow into the trap,
• Case E2 estimates air emissions from the wastewater
flowing into a J trap,
• Case E3 estimates air emissions from a lift station
where the pump is periodically lifting wastewater from
a splash filled covered sump, and
• Case E4 estimates air emissions from wastewater flow
into open sumps or junction boxes.
4.2.3 A Listing of Emission Factors
Air emissions factors are presented for induced airflow in
collection conduit systems accepting hazardous aqueous waste. The
major sources of induced airflow into and out of a collection
conduit system are process drains, manholes, and junction boxes.
Tables 4-2 through 4-8 describe the estimated fraction of the
organic emitted from the three units of the collection conduit.
The emission factors are listed for five different organic
compounds that differ in volatility: 1,3-butadiene, toluene,
naphthalene, 1-butanol, and phenol.
The airflow induced by the wind is sensitive to the geometry
of the source, the direction of flow of the wind, and the
velocity of the wind. Because of the large numbers of
significant factors that could conceivably influence the rate of
emissions due to wind, the emission estimates are presented as a
range, with zero as the lower bound of the range and a
combination of values from the three cases as the upper range.
The upper range is not the greatest possible value of the
estimated emissions, because higher collection system
temperatures or higher wind speed could increase the emission
4-7
-------
rate. The choice of a specific value to be used for estimating
emission factors from induced airflow in the collection conduit
component is also presented in Tables 4-2 to 4-8. In some cases,
the effects of the various mechanisms for airflow can be
additive, but in some cases the effects would tend to cancel each
other.
The summary of the result of the model weir calculations are
presented in Table 4-8.
TABLE 4-2. AIR EMISSION ESTIMATES FOR DILUTE AQUEOUS
1,3-BUTADIENE SOLUTIONS FLOWING THROUGH WASTE COLLECTION
SYSTEM NETWORKS3
(FRACTION EMITTED)
CASE
Case 1
Case 2
Case 3
Typical
value
DRAINS
(A)
0.63
0.73
0.54
0.63
MANHOLES
(B)
0.087
0.21
0.147
0.15
COLLECTION
CONDUITS
(C)
0.95
0.79
0.56
0.77
a Case Al is Unit A with Case 1 conditions. For a
discussion of units and cases, see the discussion on
pages 4-6 and 4-7.
4-f
-------
TABLE 4-3. AIR EMISSION ESTIMATES FOR DILUTE AQUEOUS
TOLUENE SOLUTIONS FLOWING THROUGH WASTE COLLECTION SYSTEM
NETWORKS3 (FRACTION EMITTED)
CASE
Case 1
Case 2
Case 3
Typical
value
DRAINS
(A)
0.073
0.113
0.053
0.08
MANHOLES
(B)
0.0045
0.0123
0.008
0.0083
COLLECTION
CONDUITS
(C)
0.48
0.148
0.057
0.23
a Case Al is Unit A with Case 1 conditions. For a
discussion of units and cases, see the discussion on
pages 4-6 and 4-7.
TABLE 4-4. AIR EMISSION ESTIMATES FOR DILUTE AQUEOUS
NAPHTHALENE SOLUTIONS FLOWING THROUGH WASTE
COLLECTION SYSTEM NETWORKS3
(FRACTION EMITTED)
CASE
Case 1
Case 2
Case 3
Typical
value
DRAINS
(A)
0.014
0.022
0.0098
0.015
MANHOLES
(B)
0.0008
0.0022
0.0014
0.0015
COLLECTION
CONDUITS
(C)
0.14
0.03
0.02
0.06
a Case Al is Unit A with Case 1 conditions. For a
discussion of units and cases, see the discussion on
pages 4-6 and 4-7.
4-9
-------
TABLE 4-5. AIR EMISSION ESTIMATES FOR
DILUTE AQUEOUS 1-BUTANOL SOLUTIONS
FLOWING THROUGH WASTE COLLECTION SYSTEM NETWORKS"
(FRACTION EMITTED)
CASE
Case 1
Case 2
Case 3
Typical
value
DRAINS
(A)
0.0001
0.00017
0.00007
0.00012
MANHOLES
(B)
0.000006
0.000017
0.000011
0.00001
COLLECTION
CONDUITS
(C)
0.00123
0.00023
0.00008
0.0005
a Case Al is Unit A with Case 1 conditions. For a
discussion of units and cases, see the discussion on
pages 4-6 and 4-7.
TABLE 4-6. AIR EMISSION ESTIMATES FOR
DILUTE AQUEOUS PHENOL SOLUTIONS
FLOWING THROUGH WASTE COLLECTION SYSTEM NETWORKS5
(FRACTION EMITTED)
CASE
Case 1
Case 2
Case 3
Typical
value
DRAINS
(A)
0.0000053
0.0000086
0.0000038
0.000006
MANHOLES
(B)
3 ID'7
8.5 ID'7
5.5 ID'7
6 ID'7
COLLECTION
CONDUITS
(C)
0.000063
0.000012
0.0000041
0.000026
a Case Al is Unit A with Case 1 conditions. For a
discussion of units and cases, see the discussion on
pages 4-6 and 4-7.
4-10
-------
TABLE 4-7. AIR EMISSION ESTIMATES FOR
WASTEWATER IN AN OPEN TRENCH SECTION
FLOWING THROUGH WASTE COLLECTION SYSTEM NETWORKS"
(FRACTION EMITTED)
COMPOUND
1, 3-Butadiene
Toluene
Naphthalene
Butanol
Phenol
PARTITION
COEFFICIENT
(Y/X)
7900
371
65.6
0.494
0.0252
FRACTION
EMITTED TO
AIR
0.059
0.045
0.025
0.0004
0.0002
a These compounds represents different compound types,
according to the value of the partition coefficient or
Henry's law constant.
TABLE 4-8. FRACTION OF A VOLATILE COMPONENT
EMITTED FROM A MODEL WEIR.
Component
1, 4 Butadiene
Toluene
Naphthalene
1-Butanol
Phenol
Fraction Emitted
0.35
0.20
0.056
0.00062
0.000033
4-11
-------
4.3 AIR EMISSION MODELS
4.3.1 Case Al Air Flow Induced by Waste Flow into Drain
Case Al considers airflow into a drain induced by wastewater
discharged to the collection conduit through a pipe inserted in
the drain. The air is assumed to be drawn into the annulus with
a velocity equal to that of the flowing water at the air/water
interface. The velocity of the induced air is assumed to
decrease to zero at the wall of the drain. The assumed air
velocity profile has not been experimentally confirmed. The air
drawn into the drain is assumed to escape at some other point in
the system after coming to equilibrium with the wastewater. In
relatively tight systems or systems with long runs of collection
conduit between openings, the resistance to airflow will inhibit
this mechanism of air induction. An illustration of this case is
presented in Figure 4-1.
The calculation requires the following inputs: flow rate of
wastewater, ratio of wastewater pipe area to drain pipe area,
partition coefficient applicable to the pollutant of interest at
the wastewater temperature, concentration of wastewater stream,
and temperature of the ambient air. The molar air density is
calculated at the ambient temperature based on the ideal gas law
assuming an ambient pressure of one atmosphere. The influent
flow rate of organics is calculated from the mass flow rate of
wastewater and the mass fraction of organics in the wastewater.
The influent air linear flow rate is calculated as one-fourth the
linear wastewater flow rate based on the assumed airflow profile.
This is converted to a molar airflow rate by multiplying by the
area ratio (drain pipe area to wastewater pipe area) and the
molar density of air.
The fraction emitted is calculated by multiplying the
dimensionless partition coefficient by the ratio of molar flows
of air to the total molar flow of air and water.
4-12
-------
,Air
Waste
wasleTlow"
Parameter
Ratio of the area of
waste to area of air
flow in drain
Fraction of entering
organic lost to
atmosphere
Partition coefficient
Temperature
Temperature
Symbol
Arr
F
K
Ta
T(C)
Units
dimensionless
mol fraction gas
per mol fraction
liquid
degrees K
degrees C
Value
4
0.21
371
298
25
Figure 4-1. Case Al. Air flow induced by waste flow into drain,
4-13
-------
ArrO.25 0'0121 K
F= ^ (4-1)
Arr 0.25 0-0121 K + 0.0555
Ta
The above symbols are defined in Figure 4.1. Note that,
within the limits of the assumption, a smaller wastewater pipe
flowing at an equivalent volumetric flow rate will induce a
greater airflow (and cause greater emissions) due to its higher
linear velocity.
Note also that slightly greater emissions will occur on
cooler days because more moles of denser ambient air will be
drawn in (it is assumed that this air will come to thermal
equilibrium with the wastewater before it escapes from the
system).
The calculation results are presented in Table 4-9.
4.3.2 Case A2 Air Flow in Drain Due to Wind Pressure
Case A2 considers airflow into a drain and through the
collection conduit. No water is flowing down the drain. The
pressure creating the airflow is due to changes in wind velocity.
The air pressure is estimated from the maximum pressure obtained
from wind flowing at 160 cm/s (3.5 mph) with the pitot tube
pointed into the wind. The drain would not normally be oriented
into the wind, but wind flow patterns and pressures are expected
to be influenced by the location of the drain relative to wind,
buildings, sumps, etc. An illustration of this case is presented
in Figure 4-2 .
The air flowing into the drain is assumed to escape at some
other point in the system after coming to equilibrium with the
wastewater. The frictional drag on the drain and in the headspace
of the collection conduit will determine the flow of air in
4-14
-------
response to the pressure exerted by the wind. The general
assumptions about case A2 are presented in Table 4-10.
TABLE 4-9. RESULTS OF CALCULATIONS FOR AIR EMISSIONS
FROM A DRAIN.
arr K Ta T (C) F
4
4
4
4
13.7
13.7
13.7
13.7
371
371
0.5
0.5
371
371
0.5
0.5
298
273
298
273
298
273
298
273
25
0
25
0
25
0
25
0
0
0
0
0
0
0
0
0
.21
.23
.00037
.0040
.48
.50
.00125
.00137
4-15
-------
Air
Air
wasTeflow~
Parameter
Length of collection conduit
Length of drain
Underflow rate
Diameter of drain
Radius of underflow conduit
Depth of liquid in underflow
Wind velocity
Relative humidity
Collection system temperature
Units
m
m
rnVs
m
m
m
m/s
percent
deg. C
Value
12.2
0.61
0.042
0.203
0.3048
0.244
1.56
50
25
Figure 4-2. Case A2. Air flow induced by wind
4-16
-------
TABLE 4-10. GENERAL ASSUMPTIONS AND CALCULATIONS FOR CASE A2
Air temperature
Relative humidity
Collection conduit temperature
Friction factor for air
Wind velocity
Radius of collection conduit
Depth of liquid in collection conduit
Headspace hydraulic radius
Flow of water in collection conduit
Headspace area in collection conduit
Density of air at 25 °C
K partition coefficient (Y/X)
Weight fraction organics in water
Flow of organics in collection conduit water
25 °C
50 percent
30 °C
0.006
156 cm/s (3.5 MPH)
30.48 cm (12 in)
24.4 cm (9.6 in)
10.9 cm
42,000 cmVs
1,830 cm2
0.0012 g/cm3
371
0.0005
21.1 g/s
Molar density of air in collection
conduit
0.00004 mol/cm3.
4-17
-------
The maximum pressure exerted by the wind is calculated
based on a solution of the Bernoulli equation:
• P = •2 (4-?1
2 g (q Z)
-1 c
where,
•P = calculated pressure, g force/cm2;
• = wind velocity, 156 cm/s (3.5 mph);
density of air at 25 °C, 0.0012 g/cm3; and
gc = 980.665 g-cm/gF-s2.
tp _1562 0.0012 _ Q Q15 g force
2 980.665 ' cm2
This value of the maximum pressure is equated to the energy
of the air velocity in the collection conduit and the frictional
losses in the collection conduit:
- - ( 1 + Ke + Arr- + + Kl ) — (4-3)
D D2 2 gc
where,
A? = pressure, 0.015 g force/cm2;
• = density of air, 0.0012 g/cm3;
Ke = diameter change coefficient, 0.31;
F = friction factor of air, 0.006;
L = length of collection conduit, (1,220 cm for
example case A2); and
D = equivalent diameter of the headspace in the
collection conduit, 40.4 cm (four times the
hydraulic radius).
4-18
-------
Arr = area ratio of collection conduit segment,
0.219
L2 = length of drain, (61 cm x 2 drains = 122 cm
for Case A2
D2 = diameter of drain, 20.3 cm
Kl = loss coefficient, 5
gc = 980.665 g-cm/gF-s2
Solving for • , the velocity of air in the drain is 62 cm/s
(122 ft/min). The sectional area of the drain is 324 cm2
permitting a calculated airflow of 20,000 cm3/s.
M = • Q = 0.81
where,
M = molar flow rate of air, mol/s;
• = density of air, 4 10~5 mol/cm3; and
Q = volumetric flow rate, 2 104 cm3/s.
The molar flow rate of the air is then calculated as 0.81 mol/s.
The flow rate of organics in the air at equilibrium with
the initial concentration of organics in the water is as follows
0 = M K C Mw (4-4;
2.7 = ( .81) (371) (0.0005) (18)
where,
0 = molar flow rate of organic in air, mol/s;
M = molar flow rate of air, 0.81 mol/s;
K = organic partition coefficient 371 mol/mol;
C = concentration of organic, 0.0005 g/g water;
and
Mw = molecular weight of water, 18 g/mol.
The flow of organic in the exit air is then 2.7 g/s.
The fraction of organics present in the air at equilibrium,
f, is independent of concentration (as long as K is a constant).
4-19
-------
The fraction of organics in the air is the ratio of the mass flow
in the air divided by the sum of the mass flow in the air and
water:
f = 0 / ( 0 + M ) (4-5)
f = 2.7/ (21.1 + 2.7)
f = 0.11.
The fraction of organics present in the air is also the fraction
lost as air emissions.
4.3.3 Case A3 Airflow Induced by Density Differences
Case A3 considers airflow up from a drain induced by
density differences between the ambient air outside the manhole
and the warm humid air in the collection conduit. No water is
flowing in the drain. The wastewater in the collection conduit
is assumed to be flowing in a direction perpendicular to the
airflow through the vents; the air is assumed to be saturated
with water and at chemical and thermal equilibrium with the
wastewater. In the case considered, the drain is assumed to be
10 cm (4 in.) in diameter and 61 cm (2 ft) long. Frictional
losses through both the drain and the collection conduit are
considered, based on a friction factor of 0.06. The height of
the "stack" is assumed to be 61 cm (2 ft). This is the vertical
distance between the level of the water in the collection conduit
and the drain. Ambient conditions are assumed to be 25 °C and 50
percent relative humidity. The wastewater temperature is assumed
to be 30 °C, and a greater difference in the ambient temperature
and the sewer temperature would tend to increase the effect of
density differences. An illustration of this case is presented
in Figure 4-3. The assumptions about Case A3 are presented in
Table 4-11. These assumptions are used for the calculations in
this section.
4-20
-------
Air
wasTeflow~
Parameter
Length of collection
conduit
Drain length
Underflow rate
Length of drain
Radius of underflow
conduit
Depth of liquid in
underflow
Ambient temperature
Relative humidity
Collection system
temperature
Units
m
m
rnVs
m
m
m
deg. C
percent
deg. C
Value
12.2
0.61
0.042
0.203
0.3048
0.244
25
50
30
Figure 4-3. Case A3. Air flow induced by density differences
4-21
-------
TABLE 4-11. GENERAL ASSUMPTIONS AND CALCULATIONS FOR CASE A3
Air temperature
Relative humidity
Collection conduit temperature
Friction factor for air
Radius of collection conduit
Depth of liquid in collection conduit
Headspace hydraulic radius
Flow of water in collection conduit
Headspace area in collection conduit
Density of saturated air at 40 °C
Density of air at 25 °C
K partition coefficient (Y/X)
Weight fraction organics in water
Flow of organics in
collection conduit water
Molar density of air in collection conduit
25 °C
50 percent
30 °C
0.006
30.48 cm
24.4 cm
10.9 cm
42,000 is
1,828 cm2
0.00117 g/cc
0.0012 g/cc
371
0.0005
21.3 g/s
0.00004
mo1/cm3.
4-22
-------
The densities of ambient air and warm humid collection
conduit air are calculated, and the density difference across the
drain system is calculated as 0.0000839 g/cm3. The maximum
pressure from density differences is the product of the density
difference and height. This value of the maximum pressure from
density differences is equated to the energy of the air velocity
in the collection conduit and the frictional losses in the
collection conduit:
:4-e;
where,
( 1 + Ke +
2 g • P
-1 c
•
4 F L * 2
ATT +
D
4 F L2
D2
+ Kl )
AP
•
Ke
F
L
D
Arr
L2
D2
Kl
the velocity of air exiting the drain hub
(cm/s);
pressure, 0.0019 g force/cm2;
density of air, 0.0012 g/cm3;
diameter change coefficient, 0.31;
friction factor of air, 0.006;
length of collection conduit, 610 cm;
equivalent diameter of the headspace in the
collection conduit, 43.6 cm (four times the
hydraulic radius);
area ratio of collection conduit segment,
0.219;
length of drain, 61 cm;
diameter of drain, 20.3 cm;
loss coefficient, 3; and
980.665 g-cm/gF-s2.
4-23
-------
Solving for •, the velocity of air in the drain hub is 26.9
cm/s (53 ft/min). The sectional area of the drain hub is 324 cm2
permitting a calculated airflow of 8,716 cm3/s. The molar density
of the air is 4.0 10~5 mol/cm3. a molar airflow rate is then
calculated as (8,716 cm3/s) (4.0 10~5 mol/cm3), or 0.35 mol/s. The
flow rate of organics in the air at equilibrium with the initial
concentration of organics in the water is calculated using
Equation 4-4:
0 = (0.35 mol/s) (371) (0.0005 g/g) (18 g/mol) .
0 = 1.17 g/s.
The fraction of organics present in the air at equilibrium
is independent of concentration (as long as K is a constant). The
fraction of organics in the air is the ratio of the mass flow in
the air divided by the sum of the mass flow in the air and water:
f = 0 / ( 0 + M ), where the variables are as
previously defined in case A2,
f = 1.17 / (1.17 + 21.3)
f = 0.053.
4.3.4 Case Bl Manhole venting due to density effects
Case Bl considers airflow from the vents in a manhole cover
induced by density differences between the ambient air outside
the manhole and the warm humid air in the collection conduit. An
illustration of this case is presented in Figure 4-4. The
wastewater in the collection conduit is assumed to be flowing in
a direction perpendicular to the airflow through the vents; the
air is assumed to be saturated with water and at chemical and
thermal equilibrium with the wastewater. In the case considered,
the manhole cover is assumed to have four vent holes of 2.5 cm (1
in.) diameter. Frictional losses through the manhole are assumed
negligible relative to losses through the manhole cover vents.
The height of the "stack" is assumed to be 67 cm (2 ft). This is
4-24
-------
the vertical distance between the level of the water in the
collection conduit and the manhole cover.
wasTe"fiow~
Parameter
Length of collection
conduit
cross-sectional area of
vent holes
Underflow rate
Height of manhole cover
above surface
Radius of underflow conduit
Depth of liquid in
underflow
Ambient temperature
Relative humidity
Collection system
temperature
Units
m
cm2
rnVs
m
m
m
deg.C
percent
deg.C
Value
12.2
20
0.0425
0.61
0.3048
0.244
25
50
30
Figure 4-4. Case Bl. Manhole venting induced by density
differences
4-25
-------
Ambient conditions are assumed to be 25 °C and 50 percent
relative humidity. The wastewater temperature is assumed to be
30 °C.
The densities of ambient air and warm humid collection
conduit air are calculated, and the density difference across the
manhole cover is determined. The gas velocity through the manhole
cover vents was then calculated from the density difference using
the equation for a sharp edged orifice:
• = 0.61 ( 2 gc h 11 )°'5 (4-7)
where,
• = linear velocity through the vent hole, cm/s;
gc = Gravitational constant, 981 cm/s2;
h = height of manhole above conduit, 61 cm (2 ft);
A» = density difference of air above and
below manhole, 3.2 10~5 g/cm3; and
• = density of warm humid air, 0.00117 g/cm3.
(Frictional losses through the thickness of the cover are
negligible.)
( 3 2 10~5 V'5
• = 0.61 2 ( 981 ) 61
^ 0.00117 j
The air velocity is converted to a volumetric flow rate by
multiplying by the cross-sectional area of the vent holes, 20 cm2
. = 34.9 ££!
s
(0.022 ft2) by the vent velocity 34.9 cm/s. Based on this
airflow, 710 cm3/s, the wastewater flow in the collection
conduit, and a partition coefficient appropriate for the compound
4-26
-------
of interest at the wastewater temperature, the fractional
emission is calculated. (The wastewater flow is 2,360 mol/s and
was calculated from an assumed collection conduit size, slope,
roughness, and an assumed wastewater depth in the collection
conduit.) The fraction emitted is calculated as
F = G K/(G K + L) (4-8) .
F =
0.0285 x 371
where,
F =
G =
K =
0.0285 x 371 + 2360
F = 0.0045
fraction emitted through cover vents,
dimensionless;
airflow rate from the cover vents, 0.0285 mol/s;
371, air/water partition coefficient for compound
of interest at wastewater temperature,
dimensionless; and
L = wastewater flow rate through collection conduit,
2,360 mol/s.
4.3.5 Case B2 Manhole Venting Due to Wind
Case B2 estimates airflow through manhole cover vents resulting
from wind blowing into the upstream end of a collection conduit.
An illustration of this case is presented in Figure 4-5. The air
flows down the collection conduit to the manhole where further
airflow is obstructed. This might occur where a collection
conduit ends at a pump sump or where a change in pipe size or
slope results in a completely filled pipe with no air space. The
airflow rate is estimated by calculating the air velocity through
the manhole cover vents that would result in a frictional head
loss equal to that available from the wind blowing into the
4-27
-------
Air
wasTeflow~
Parameter
Length of collection conduit
cross-sectional area of vent holes
Underflow rate
Height of manhole cover above surface
Radius of underflow conduit
Depth of liquid in underflow
Wind velocity
Collection system temperature
Units
m
cm2
m3/s
m
m
m
m/s
deg.C
Value
12.2
20
0.042
0.61
0.3048
0.244
1.56
30
Figure 4-5. Case B2. Manhole venting due to wind
4-28
-------
upstream end of the collection conduit. Frictional losses
through the collection conduit, the manhole, and the cover
thickness are assumed to be negligible in comparison to losses
through the cover vents.
Frictional losses through the cover vents are calculated
using an equation for flow through a sharp-edged orifice:
f ? • V-5
. = 0.61 • 2 — (4-9)
( " 's )
where,
• = linear velocity through vent cover, cm/s;
•w = wind velocity, 156 cm/s (3.5 mph);
•a = ambient air density, 0.0012 g/cm3; and
• s = density of warm humid air in collection
conduit, 0.00117 g/cm3.
Substituting into Equation 4-9 to estimate the velocity out
of the manhole vent cover openings,
• =0.61 1562 -
( .00117 )
=96.3 cm/s (2.1 mph)
The manhole cover is assumed to have four vents of 2.5 cm
(1 in.) diameter. The wind velocity in the direction of the
collection conduit is assumed to be 156 cm/s (3.5 mph). The
factor of 0.61 is an orifice coefficient that will be
approximately constant for the range of flows considered.
The molar airflow rate can be calculated from the linear
velocity through the cover vents (96.3 cm/s) by multiplying by
the total area of the four vents, 20 cm2 (0.022 ft2), and by the
molar density at the warm humid collection conduit conditions,
0.00004 mol/cm3.
0.078 mol/s = (96.3 cm/s) (20cm2) (0.00004)
4-29
-------
The wastewater flow rate in the collection conduit is implicitly
specified on the basis of assumed collection conduit depth,
diameter, slope, and roughness (2,360 mol/s).
The fractional emission of organics is calculated from
Equation 4-8, using the molar flow rates of air and water, and a
dimensionless partition coefficient appropriate for the compounds
of interest at the wastewater temperature:
GK + L
where,
G = airflow rate, 0.078 mol/s;
K = 371, dimensionless partition
coefficient; and
L = water flow rate, 2,360 mol/s.
F = (0.078) (371) / [ (0.078) (371)+ 2360]
F = 0.121
4.3.6 Case B3 Manhole venting due to wind underflow
Case B3 considers emissions from manhole cover vents over a
flowing, partially filled collection conduit. Air resulting from
wind blowing in one end of the collection conduit is flowing in
the upper portion of the collection conduit. The direction of the
airflow relative to the water flow is not considered; it is
assumed that the air in the collection conduit is at thermal and
chemical equilibrium with the wastewater at the location of the
manhole. An illustration of this case is presented in Figure 4-6.
The air velocity resulting from the wind pressure is
calculated from a Bernoulli equation based on frictional losses
through the unfilled section of the pipe;
D
4-30
-------
where,
a
s
L
D
linear velocity of air through unfilled
section of collection conduit, 80 cm/s
(1.79 mph);
velocity of wind, 156 cm/s (3.5 mph);
density of ambient air, 0.0012 g/cm3;
density of humid air in collection conduit,
0.00117 g/cm3;
friction factor for air, assumed constant at
0.006, dimensionless;
length of collection conduit, 4,570 cm; and
equivalent diameter (four times the
hydraulic radius) of unfilled section of
collection conduit, 40.4 cm.
156Z
.0012
.00117
+ 4 (.006) (4570)
40.4
v /
• = 80 cm/s (1.79 mph).
The velocity is then used to calculate the pressure drop
through the shorter length of collection conduit between the
manhole and the discharge end of the collection conduit:
4 f L
P =
where:
. P =
Ls =
pressure drop through collection conduit between
manhole and discharge end, g force/cm2
length of collection conduit between manhole and
gas exit, 3,050 cm
gravitational constant, 981 g cm/g force-52
4-31
-------
Air
wasTeflow~
Parameter
Length of collection conduit
Length after manhole
cross-sectional area of vent holes
Underflow rate
Height of manhole cover above surface
Radius of underflow conduit
Depth of liquid in underflow
Wind velocity
Collection system temperature
Units
m
m
cm2
m3/s
m
m
m
m/s
deg.C
Value
45.7
30.5
20
0.042
0.61
0.3048
0.244
1.56
30
Figure 4-6. CaseB3. Manhole venting due to wind underflow
4-32
-------
t = 4 (.006) (3050) (80)2 (.00117)
2 (981)(40.4)
•P = 0.0069
This pressure (0.0064 g force/cm2) is then used as the driving
force in the equation for flow through a square-edged orifice to
calculate the linear velocity of air emitted from the manhole
cover vents:
• =0.61 (2 g —) °'5
c c .
B
where:
•c = linear velocity through the cover vents, cm/s
0.61 = orifice coefficient (dimensionless) appropriate
for the velocity range expected.
.2
• =0.61 ( 4 f L —) °-5
C B J-)
rr cm
= 65.7
c s
Note that the above equations can be combined:
=0.6l2 (981)
( .00117)
The linear velocity can be converted to a molar flow rate by
multiplying by the cross-sectional area of the vents, 20 cm2
(four vents each 2.5 cm [1 in.] in diameter assumed in the
example), and the molar density of warm humid air at
4-33
-------
the wastewater temperature, 4*10 5 mol/cm3. The wastewater flow
rate, 2,360 mol/s, has been implicitly specified in the example
from the depth, diameter, slope, and roughness of the collection
conduit. The fraction of organics emitted is calculated from
Equation 4-8, using the molar flow rates and a dimensionless
partition coefficient appropriate for the compound of interest at
the wastewaters' temperature:
F=
GK + L
where,
F = fraction of organics emitted through manhole cover
vents;
G = airflow rate through manhole cover vents, 0.053
mo1/s;
K = partition coefficient, 371, dimensionless; and
L = wastewater flow rate, 2,360 mol/s.
(.053) (371)
F =
(.053) (371) + 2360
F = 0.008
4.3.7 Case Cl Conduit air flow due to wind
Case Cl considers air blowing directly into one end of a
collection conduit, reaching thermal and compositional
equilibrium within the collection conduit and exiting a junction
box. An illustration of this case is presented in Figure 4-7.
The general assumptions of Case Cl are presented in Table 4-12.
The maximum pressure exerted by the wind is calculated
based on a solution of the Bernoulli equation (see Equation 4-2):
4-34
-------
1/asEeflow"
Parameter
Length of collection
conduit
cross-sectional area of
vent holes
Underflow rate
Height of manhole cover
above surface
Radius of underflow conduit
Depth of liquid in
underflow
Wind velocity
Relative humidity
Collection system
temperature
Units
m
cm2
m3/s
m
m
m
m/s
percent
deg.C
Value
45.7
20
0.042
0.61
0.3048
0.244
1.56
50
30
Figure 4-7. Case Cl. Conduit air flow induced by wind
4-35
-------
TABLE 4-12. GENERAL ASSUMPTIONS AND CALCULATIONS FOR CASE Cl:
Air temperature 25 °C
Relative humidity 50 percent
Collection conduit temperature 30 °C
Friction factor for air 0.006
Wind velocity 156 cm/s (3.5 MPH)
Radius of collection conduit 30.48 cm (12 in.)
Depth of liquid in collection conduit 24.4 cm (9.6 in.)
Headspace hydraulic radius 10.9 cm
Flow of water in collection conduit 42,196 cm3/s
Headspace area in collection conduit 1,828 cm2
Density of air at 25 °C 0.0012 g/cm3
K partition coefficient (Y/X) 371
Weight fraction organics in water 0.0005
Flow of organics in collection conduit water 21.1 g/s
4-36
-------
Molar density of air in collection conduit 0.00004 mol/cm3.
2 g
— ' r
where,
•P = calculated pressure, g force/cm2;
• = wind velocity, 156 cm/s (3.5 mph);
density of air at 25 °C, 0.0012 g/cm3; and
gc = 980.665 g-cm/gF-s2.
,p _ 156^ 0.0012 _ Q Q15 g force
2 980.665 ' cm2
This value of the maximum pressure is equated to the energy
of the air velocity in the collection conduit and the frictional
losses in the collection conduit:
• P 4 F L • 2
-L = i + ±_i± — 4-i
D 2 gc
where,
A? = pressure, 0.015 g force/cm2;
• = density of air, 0.0012 g/cm3;
F = friction factor of air, 0.006;
L = length of collection conduit, 4570 cm;
D = equivalent diameter of the headspace in the
collection conduit, 40.4 cm (four times the
hydraulic radius), and
gc = 980.665 g-cm/gF-s2.
Solving for • , the velocity of air in the collection
conduit is 80 cm/s (1.8 mph) . The sectional area of the headspace
is 1,828 cm2 permitting a calculated airflow of 146,000 cm3/s.
The molar density of the air is 4 10~5 mol/cm3; a molar airflow
rate is then calculated as (1.46 105 cm3/s) (4 10~5 mol/cm3), or
5.8 mol/s. The concentration of organics in the air at equi-
4-37
-------
librium with the initial concentration of organics in the water
is as follows:
(5.8 mol/s) (371) (0.0005 g/g) (18 g/mol) or 19.4 g/s .
The fraction of organics present in the air at equilibrium
is independent of concentration (as long as K is a constant). The
fraction of organics in the air is estimated with Equation 4-8,
using the ratio of the mass flow in the air divided by the sum of
the mass flow in the air and water:
f = 0 / ( 0 + M ), where the variables are as
previously defined in case A2,
f = 19.4 / (21.1 +19.4)
f = 0.48.
4.3.8 Case C2 Conduit Air Flow Induced by Water Flow
Case C2 estimates airflow into a collection conduit from a
junction box, induced by water flow in the collection conduit.
This air reaches thermal and compositional equilibrium within the
collection conduit and is discharged from the collection conduit
at the next junction box. An illustration of this case is
presented in Figure 4-8. The assumptions are presented in Table
4-13.
The velocity profile for the surface of the water in the
collection conduit is assumed to be given by the following
empirical relationship:
Y > e ] (4-12'
{ e)
where:
•+ = velocity quotient, the ratio of the velocity to
the friction velocity,
e = surface roughness, cm
Y = distance from a point on the surface to the
nearest wall surface interface, cm.
4-38
-------
"wasfelovT
Parameter
Underflow rate
Radius of underflow conduit
Depth of liquid in
underflow
Units
rnVs
m
m
Value
0.042
0.3048
0.244
Figure 4-8. Case C2. Conduit air flow induced by water flow
4-39
-------
TABLE 4-13. GENERAL ASSUMPTIONS AND CALCULATIONS, CASE C2.
Air temperature 25 °C
Relative humidity 50 percent
Collection conduit temperature 30 °C
Friction factor for air 0.006
Radius of collection conduit 30.48 cm
Depth of liquid in collection conduit 24.4 cm
Headspace hydraulic radius 10.9 cm
Flow of water in collection conduit 42,196 cm3/s
Headspace area in collection conduit 1,828 cm2
Density of air at 25 °C 0.0012 g/cm3
g partition coefficient (Y/X) 371
Weight fraction organics in water 0.0005
Flow of organics in collection conduit water 21.1 g/s
Molar density of air in collection conduit 0.00004 mol/cm3
Reynolds number for airflow 2,100
Average velocity of water 39 cm/s
Surface velocity of water 42 cm/s
Roughness of collection conduit wall 0.21 cm
Slope of collection conduit 0.000431.
4-40
-------
The average velocity in the collection conduit is estimated
as 38.7 cm/s integrating the above equation for average flow. The
average surface velocity was 42.2 cm/s. The perimeter of the
surface was 60 cm, and the perimeter of the collection conduit
headspace was 108 cm. The average velocity of the airflow was
established as follows:
' Cm' = 42 2 C 60 cm |
s s \ 60 cm + 108 cm )
This average air velocity of 15.07 cm/s is 36 percent of
the water velocity in the collection conduit. This estimated
ratio of air velocity to water velocity compares favorably to a
reported percentage of 35 for laminar airflow due to liquid
drag2. The estimated Reynolds number for the above flow
conditions suggests that the flow of air may be in the
transitional zone. The assumption of laminar flow of air may have
overestimated the flow of air by 20 percent.
The estimated velocity of air in the collection conduit is
15 cm/s (0.33 MPH). The sectional area of the headspace is 1,828
cm2 permitting a calculated airflow of 27,000 cm3/s. The molar
density of the air is 4 10~5 mol/cm3; a molar airflow rate is then
calculated as (2.7 104 cm3/5) (4 10~5 mol/cm3), or 1.08 mol/s. The
flow rate of organics in the air at equilibrium with the initial
concentration of organics in the water is as follows:
(1.08 mol/s)(371)(0.0005 g/g)(18 g/mol) or 3.61 g/s .
The fraction of organics present in the air at equilibrium
is independent of concentration (as long as K is a constant). The
fraction of organics in the air is estimated with Equation 4-8,
using the ratio of the mass flow in the air divided by the sum of
the mass flow in the air and water:
4-41
-------
f = 0 / ( 0 + M ), where the variables are as
previously defined in case A2,
f = 3.61 / (21.1 + 3.61), and
f = 0.146.
4.3.9 Case C3 Conduit air flow due to density differences
This calculation considers airflow from the discharge end
of a partially filled collection conduit to the influent end of
the collection conduit resulting from a density difference
between the cooler ambient air and the warm humid air in equilib-
rium with the wastewaters. The air flowing from the collection
conduit is assumed to be in thermal and chemical equilibrium with
the wastewater. Air and water flow countercurrently. An
illustration of this case is presented in Figure 4-9.
The ambient temperature and relative humidity and the
wastewater temperature are used to calculate the density
difference; the slope and length of the collection conduit are
used to calculate the elevation difference producing the "stack
effect." Based on the length, diameter, and depth in the
collection conduit, the frictional resistance to airflow is
determined as a function of air velocity. The air velocity is
calculated from a balance of the "stack effect" and the
frictional losses using a form of the Bernoulli equation (see
Equation 4-6) :
2 gn •• h
' c
4FL
1 +
D
'4-13'
where,
4-42
-------
Air
wasteflow
Parameter
Length of collection
conduit
Underflow rate
Elevation of exit relative
to entrance
Radius of underflow conduit
Depth of liquid in
underflow
Ambient temperature
Relative humidity
Collection system
temperature
Units
m
rnVs
m
m
m
deg.C
percent
deg.C
Value
45.7
0.042
0.006
0.3048
0.244
25
50
30
Figure 4-9. Case C3. Conduit air flow induced by air density
differences
4-43
-------
• = velocity of air through the collection conduit
headspace, 5.3 cm/s;
gc = acceleration of gravity, 981 cm/s2
•• = density difference between ambient air and warm
humid air in collection conduit, 3.2 10~5 g
force/cm2;
h = elevation difference determined from
collection conduit length and slope,
2 cm;
• = density of warm humid air in collection conduit,
0.00117 g/cm3;
F = friction factor for airflow through
collection conduit, assumed constant
at 0.006, dimensionless;
L = collection conduit length, 4,570 cm; and
D = diameter of circle having equivalent
area to the cross section of the
collection conduit headspace, 40.4
cm.
This velocity is converted to a molar flow rate by
multiplying by the equation cross-sectional area of the headspace
in the collection conduit, 1,828 cm2 and the molar density of air
at the wastewater temperature, 4.0 10~5 mol/cm2.
M = • A V = 0.387
where,
M = molar flow rate of air, 0.382 mol/s;
density of air, 0.00004 mol/cm3;
A = cross-sectional area, 1828 cm2; and
V = velocity, 5.3 cm/s.
The water flow rate (2,360 mol/s) is specified implicitly
by the slope, diameter, depth, and roughness of the collection
conduit. The fraction of influent organics that is emitted is
calculated from the molar flow rates of water and air and the
dimensionless partition coefficient for the compound of interest
at the wastewater temperature. Air emitted from the collection
4-44
-------
conduit is assumed to be in equilibrium with influent wastewater.
The fractional emissions are calculated as:
F = G K / (G K+ L) = 0.057,
where,
G = airflow rate, 0.387 mol/s;
K = 371, dimensionless partition coefficient for
compound of interest; and
L = wastewater flow rate, 2,360 mol/s.
4.3.10 Case Dl Stack Vent Air Flow
Case Dl considers emissions from a stack on a sump. The
stack was designed to promote the discharge of fumes above
workers' heads so that their exposures to environmental releases
would be reduced. Case Dl uses a method identical to Case A3 to
estimate the airflow due to the stack effect. An illustration of
this case is presented in Figure 4-10. The assumptions are
presented in Table 4-14.
Case Dl considers airflow up from a sump; through a vent
induced by density differences between the ambient air outside
the sump and the warm humid air in the collection conduit. The
wastewater in the collection conduit is assumed to be flowing in
a direction perpendicular to the airflow through the vents; the
air is assumed to be saturated with water and at chemical and
thermal equilibrium with the wastewater. In the case considered,
the vent is assumed to be 10 cm (4 in.) in diameter and 366 cm
(12 ft) long. Frictional losses through both the drain and the
collection conduit are considered, based on a friction factor of
0.06. The height of the "stack" is assumed to be 366 cm (12 ft).
This is the vertical distance between the top of the sump and the
vent top. Ambient conditions are assumed to be 25 °C and 50
4-45
-------
Parameter
Length of collection
conduit
Underflow rate
Stack length
Stack diameter
Radius of underflow conduit
Depth of liquid in
underflow
Ambient temperature
Relative humidity
Collection system
temperature
Units
m
rnVs
m
cm
m
m
deg.C
percent
deg.C
Value
30.48
0.042
3.66
20
0.3048
0.244
25
50
30
Figure 4-10. Case Dl. Closed junction box vent induced by air
density differences from open exit conduit.
4-46
-------
TABLE 4-14. GENERAL ASSUMPTIONS AND CALCULATIONS, CASE D-l.
Air temperature
Relative humidity
Collection conduit temperature
Friction factor for air
Radius of collection conduit
Depth of liquid in collection conduit
Headspace hydraulic radius
Flow of water in collection conduit
Headspace area in collection conduit
Density of saturated air at 30 °C
Density of air at 25 °C
K partition coefficient (Y/X)
Weight fraction organics in water
Flow of organics in
collection conduit water
Molar density of air in collection
conduit
25 °C
50 percent
30 °C
0.006
30.48 cm
24.4 cm
10.9 cm
42,000 cmVs
1,830 cm2
0.00117 g/cm3
0.0012 g/cm3
371
0.0005
21.3 g/s
0.00004
mo1/cm3.
4-47
-------
percent relative humidity. The wastewater temperature is assumed
to be 30°C.
The maximum pressure from a warm humid collection conduit
air is calculated from the product of the density difference and
height. This value of the maximum pressure from density
differences is equated to the energy of the air velocity in the
collection conduit and the frictional losses in the collection
conduit (see Equation 4-3):
P , , . „ . 4 F L „ .2 . 4 F L2
„ 2
1 + Ke + Arr2 + + Kl
D D2 2 g
-1 c
where,
•P = pressure, 0.0117 g force/cm2;
• = density of air, 0.0012 g/cm3;
Ke = diameter change coefficient, 0.378;
F = friction factor of air, 0.006;
L = length of collection conduit, 3,048 cm;
D = equivalent diameter of the headspace in the
collection conduit, 43.6 cm (four times the
hydraulic radius);
Arr = area ratio of collection conduit segment, 0.055;
L2 = length of drain, 366 cm;
D2 = diameter of vent stack, 10 cm;
Kl = loss coefficient, 3; and
gc = 980.665 g-cm/gF-s2.
Solving for •, the velocity of air in the vent stack is 62
cm/s (118 ft/min). The sectional area of the vent is 324 cm2
permitting a calculated airflow of 20,000 cm3/s. The molar
density of the air is 4.0 10~5 mol/cm3. a molar airflow rate is
then calculated as (20,000 cm3/s) (4.0 10~5 mol/cm3), or 0.8 mol/s
The flow rate of organics in the air at equilibrium with the
initial concentration of organics in the water is as follows:
(0.8 mol/s)(371)(0.0005 g/g)(18 g/mol) or 2.7 g/s .
4-48
-------
The fraction of organics present in the air at equilibrium
is independent of concentration (as long as K is a constant). The
fraction of organics in the air is the ratio of the mass flow in
the air divided by the sum of the mass flow in the air and water:
f = 0 / ( 0 + M ), where the variables are as
previously defined in case A2,
f = 2.7 / (21.3 + 2.7)
f = 0.11.
4.3.11 Case D2 Stack Vent Working Loss
Case D2 considers emissions from a stack on a sump due to
working losses in the enclosed headspace of the sump. An
illustration of this case is presented in Figure 4-11. The stack
is designed to promote the discharge of working loss fumes above
workers' heads so that their exposures to environmental releases
would be reduced. Case D2 assumes that the working losses are
assumed to be saturated with water and at chemical and thermal
equilibrium with the wastewater. In the case considered, the
sump has a working loss due to flow variability in the system.
Q = 1.157 1(T5 V T
where,
Q = flow rate of air from vent, L/s;
V = volume of headspace in the sump, 10,000 L; and
T = turnovers of headspace, 3 per day.
Q=1.15710~5 (10,OOOL) (3 per day)
Q = 0.35 L/s (350 cm3/s).
4-49
-------
Parameter
Length of collection
conduit
Underflow rate
Stack length
Stack diameter
Radius of underflow conduit
Depth of liquid in
underflow
Turnovers
Volume of system
Collection system
temperature
Units
m
m3/s
m
cm
m
m
per day
m3
deg.C
Value
30.48
0.252
3.66
20
0.3048
0.244
3
10
30
Figure 4-11. Case D2. Closed junction box vent induced by
working losses from water seal on entrance and exit conduit.
4-50
-------
The molar density of the air is 4.0 10 5 mol/cm3; a molar airflow
rate is then calculated as (350 cm3/s) (4.0 10~5 mol/cm3) , or 0.014
mol/s. The flow rate of organics in the air at equilibrium with
the initial concentration of organics in the water is as follows:
(0.014 mol/s)(371)(0.0005 g/g)(18 g/mol) or 0.047 g/s .
The fraction of organics present in the air at equilibrium is
independent of concentration (as long as K is a constant). The
fraction of organics in the air is estimated from Equation 4-8,
using the ratio of the mass flow in the air divided by the sum of
the mass flow in the air and water:
f = 0 / ( 0 + M ) , where the variables are as
previously defined in case A2,
f = 0.047 / (21.3 + 0.047)
f = 0.0022.
4.3.12 Case D3 Trench volatilization loss
Case D3 considers emissions from open trenches around
process equipment. These trenches are used to collect process
wastes, tank cleaning wastes, unplanned leaks, and water. Air
blows across the top of a grate covering the open top channel of
the trench. An illustration of this case is presented in Figure
4-12.
The mass transfer coefficient of the gas phase is estimated
from a modified "j factor" equation. The average velocity is
required for that equation. The average wind speed is estimated
as one half the specified wind speed. The modified equation
converts a specified wind speed to an estimated average wind
speed.
4-51
-------
Parameter
Length of trench
Flow velocity
Depth of liquid in trench
Units
m
m/s
m
Value
12.2
0.4572
0.0762
Figure 4-12. CaseDS. Open trench.
The modified equation is as follows:
with,
kg =
V
1
Dv =
k = 2.3
9
mass transfer coefficient,
1.6 10~5 gmol/cm2 s (0.0038 m/s);
wind velocity, 447 m/s;
characteristic length, 1000 cm; and
diffusion coefficient, 0.088 cm2/s.
4-52
-------
The liquid mass transfer coefficient is calculated by Owens,
Edwards, and Gibbs3.
y0.67 R
Kal = 21.6 -—T (4-14)
•" o
where,
Kal = liquid mass transfer coefficient, 176 /day;
v = velocity, 1.5 ft/s;
H = depth, 0.25 ft; and
K/Ko = ratio of mass transfer coefficients of
toluene and air, 0.477.
Equation 4-14 was developed for highly volatile components and
therefore provides the liquid mass transfer coefficient. Because
the overall mass transfer coefficient is needed for more general
use, the overall mass transfer coefficient is obtained by
modifying Equation 4-14. The desired overall equation is
obtained by summing the resistance of the two regions of mass
transfer in series:
( 1 H V1
Ka = — + £? (4-15)
( Kal (C Kg K 0.000736) j
where:
Ka = the overall mass transfer coefficient, 150/day,
calculated from the above equation;
Kal = 176, calculated from Equation 4-14;
H = 0.25 ft;
C = a conversion factor, 24 x 3,600 / 12 / 2.54 x 100
(ft/day) (s/m);
Kg = 0.00325 m/s;
0.00736 = a conversion factor; and
K = 371 (Y/X) .
4-53
-------
The residence time in a 40-ft length of channel is 40 ft / 1.5
ft/s) or 27 s or 3.1 10~4 days. The fraction lost during flow
through the 40-ft channel is estimated with the following
equation:
f = 1- EXP (-Ka t)
f = 1- EXP {-(150/day) (0.00031 days)}
f = 0.045,
where f and Ka are defined above, and t is the time in days.
Therefore, 4.5 percent of toluene is estimated to be emitted over
the 40-ft section of channel.
4.3.13 Case El J Trap Sealed with Wastewater with No Wastewater
Flow
Case El applies to the case where a waste flows
periodically into a drain hub with a J trap. Waste is retained
in the J trap and emissions occur by diffusion and air flow out
of the drain hub. Air flow out of the hub is generated when
changes in the level of the liquid in the J trap displaces air in
the column above the waste. Fluctuations in the liquid level are
caused by pressure changes in the collection system induced by
either wind fluctuations or wastewater flow rate fluctuations.
The air displaced from over the surface in the J trap mixes with
the air in the drain hub in the proposed theoretical model. An
illustration of this case is presented in Figure 4-13.
Air emissions from a J trap during the time when no
wastewater flow is occurring into the hub can be described by the
following equations.
The equation for air loss is:
h
2
Deff = — (4-16;
4-54
-------
Parameter
Depth of liquid under top
of drain hub
Cross-section area of drain
Period of level fluctuation
distance of level
fluctuation
Units
m
cm2
minutes
cm
Value
.457
81
0.5
7.62
Figure 4-13. Case El. Open J drain trap with waste: no waste
flow.
4-55
-------
where,
Deff = the effective diffusivity, cm2/s;
h = the waste displacement, cm; and
t = the period of the displacement, s.
This equation can be used to estimate the emission rate with the
following equation:
D ff
e = ^££ A (4-17)
d
where,
Deff = the effective diffusivity, cm2/s;
d = the distance from the waste to the top of
the hub, cm;
A = the hub cross-section open area, cm2; and
e = the emission rate, cm3/s.
The emission rate in units more conventient for emissions
estimation is estimated from Equation 4-18.
1 8
E = K e (4-18)
24400 ^ ;
where,
K = the partition coefficient, y/x;
18 = the weight of a mol of water, g/mol;
24400 = the volume of a mol of gas, cm3/mol;
e = the emission rate, cm3/s; and
E = the emission rate, g/s per weight fraction
in the wastewater.
The following example calculation illustrates the use of
these equations to estimate emissions of toluene from a 4 inch
diameter hub.
4-56
-------
The following conditions are assumed for Case El.
h = 7.62 cm (3 in)
t = 30 s (0.5 min)
d = 45.7 cm (1.5 ft)
A = 81 cm2
K = 357 (y/x)
The effective diffusivity, Deff, is calculated as follows:
Dff- (7'62)2 =1.935-^
eff 30 s
The value of the emission rate, e, is estimated.
(1.935) (81) 43 cm3
6 45.7' ~s~
The emission rate E is calculated as follows.
E = (357) (18) (3'43) = 0.9034 ^- per weight tract, toluene water
24,400 s
The ratio of the emission rate of toluene to the underflow rate
is (0.9034 g/s)/(2833 g/s), or .00032.
4.3.14 Case E2 J Trap Water Sealed with Wastewater Flow
Case E2 applies to a case where a waste flows continuously
into a drain hub with a J trap. An illustration of this case is
presented in Figure 4-14. The flow of waste is exposed to the
wind for a short distance over the top of the drain hub. The
model approach is to identify correlations of the mass transfer
coefficients with the drain characteristics. These correlations
were developed empirically using Enviromega data1 rather than
4-57
-------
Parameter
Depth of liquid under top of drain hub
Cross-section area of drain hub opening
waste flow rate
diameter of waste pipe
drop distance
Units
m
cm2
cm2/s
cm
cm
Value
.457
81
250
5
3.81
Figure 4-14. Case E2. Open J drain trap with waste flow.
4-58
-------
from first principles. Extrapolation beyond the range of waste
flow and physical dimensions upon which the correlation is based
should be used with caution.
This case can be described by the following equations. The
equation for the gas phase mass transfer coefficient is composed
of an empirical constant with a correction for the gas diffusion
coefficient.
D
k = .178
9
where,
0.088
^ 0.088;
the gas phase mass transfer coefficient, m/s,
the reference gas diffusion coefficient,
cm2/s, and
Dg = the gas phase diffusion coefficient, cm2/s.
The equation for the liquid phase mass transfer coefficient
is composed of an empirical constant with a correction for the
liquid diffusion coefficient and a correction for the waste
velocity. The mass transfer coefficient is assumed to be
proportional to the velocity of waste entering the hub.
k1 = .0041 V
0.0000088
'4-20'
where,
V
0.0000088
the liquid phase mass transfer coefficient,
m/s;
the waste velocity, cm;
the reference liquid diffusion coefficient,
cm2/s; and
the liquid phase diffusion coefficient,
cm2 / s.
The overall mass transfer from a two-resistance model, K0 ,
is a combination of the gas and the liquid mass transfer
coefficients:
4-59
-------
/ 1 1 N -1
K = + (4-?1 '
K7 40.9 K K (4 Zl,
i g
where,
K0 = the overall mass transfer coefficient based
upon the liquid concentrations (m/s);
K! = the liquid phase mass transfer coefficient
(m/s);
Kg = the gas phase mass transfer coefficient
(m/s); and
K = the partition coefficient (atm-m3/mol) .
The air losses, falr, from the two-resistance model are as
follows:
K A
f . = I - EXP (- ——) (4-22)
air x _ '
where,
K0 = the overall mass transfer coefficient based
upon the liquid concentrations (cm/s);
q = the liquid flow rate (cm3/s);
A = the area of the exposed surface (cm2) ;
falr = the fraction of the component emitted to the
air.
The area of the mass transfer surface is assumed to equal
the product of the circumference of the inlet pipe and the depth
of fall of the wastewater stream before it enters the hub.
The following example calculation illustrates the use of
these equations to estimate air emissions of toluene.
The assumed conditions are:
4-60
-------
K = 357 (y/x)
q =250 cm3/s
pipe diameter = 2 in
pipe cross section area = 20.27 cm2
pipe circumference = 15.96 cm
distance from pipe exit to hub inlet = 3.81 cm(1.5 in)
The waste velocity can be calculated as follows:
3
250 ——
V= * = 12.33 ™ .
(20.27 cm2) s
The values of the gas and liquid diffusivities are almost
identical to the reference values in the correlations, and the
ratio can be assumed to equal unity. The liquid phase mass
transfer coefficient can be estimated as follows.
k, = (0.0041) (12.33) (I)'66 = 0.05 — .
1 s
This value of the liquid phase mass transfer coefficient is
approximately 6 times greater than an estimate for the turbulent
zone of an agitated aeration basin.
The gas phase mass transfer coefficient can be estimated:
k = (0.178)(1
g
-66
This value of the gas phase mass transfer coefficient is
approximately 25 times as much as the mass transfer surface of a
quiescent impoundment. The higher value for the gas phase mass
transfer coefficient may be related to the form induced eddies
around the stream of water and the much smaller eddy size near
the flowing water when compared to the quiescent impoundment.
The overall mass transfer coefficient can be calculated as
follows (based on liquid concentrations) .
4-61
-------
K = _ _
0 \ (0.178) (357/55555) (40.9) 0.05
The distance from the pipe exit to the entrance of the hub is
3.81 cm (1.5 in) . The circumference of the pipe opening is 15.96
cm (2 in dia.) . The value of the area of mass transfer is the
product of these two values, 61 cm2.
The fraction emitted is estimated as follows:
f =l
250
4.3.15 Case E3 Lift Station with Periodic Pumping of Wastewater
Case E3 applies to a case where a waste flows continuously
into an enclosed sump. The wastewater is lifted by pump from the
sump to a collection main at higher elevation. The pump rate is
substantially greater than the typical wastewater flow rate
(providing excess capacity) . The wastewater flows into the sump
by splashing at least part of the time. The gas flows out of the
headspace as the sump fills. The model approach is to assume a
fractional approach to equilibrium in the sump and in the exiting
gas. This fractional approach is assumed to be 50%, based in
part upon data provided by Enviromega1. An illustration of this
case is presented in Figure 4-15.
This case can be described by the following equations. The
equation for the concentration in the gas phase is obtained from
the partition coefficient:
f H C1
C = - - (4-23)
9 0.0244
where,
Cg = the gas phase concentration of the constituent,
moles/m3 gas;
4-62
-------
G! = the liquid phase concentration of the constituent,
moles/m3 wastewater;
H = the partition coefficient, a tin m3 /mol; and
f = the fractional approach to equilibrium in the gas
phase.
The fraction emitted from lift stations is estimated from
the ratio of the concentration in the gas to the concentration in
the liquid, since it is assumed that the volumetric flow rate of
gas out of the sump equals the volumetric flow rate of the
wastewater into the sump.
The following example calculation illustrates the use of
these equations to estimate air emissions of toluene. The
assumed conditions are:
H = 0.00643 atm-mVmol
f = 0.5
The fraction of toluene that is emitted as air emissions is
estimated as follows:
f = (0.5) (.00643)
0.0244
4-63
-------
Parameter
Fraction equilibrium for headspace
Units
Value
0.5
Figure 4-15. CaseES. Lift station with periodic pumping.
4-64
-------
4.3.16 Case E4 Open Surfaces in Sumps
Case E4 applies to a case where a waste flows continuously
into a sump or junction box and the surface of the waste in the
sump or junction box is open to the atmosphere. Enviromega
collected data for two situations1, (1) flow of waste below the
surface and (2) the flow of waste into the surface from a
partially filled inlet wastewater conduit. The model approach is
to identify a method for applying the turbulent flow mass
transfer model (trench model) to flow in sumps and junction
boxes. The Enviromega data1 was used as the basis of the
development of this extension of the trench model. The equation
developed here may be applicable to other open surfaces in sumps
and junction boxes. An illustration of this case is presented in
Figure 4-16.
This case can be described by the following equations. The
equation for the liquid phase mass transfer coefficient is
composed of an empirical constant with a correction for the
liquid diffusion coefficient, the depth of flow, and the waste
velocity in the sump. The trench model for liquid phase mass
transfer is modified, equating the depth of flow to the depth of
entrance flow. The width of flow is assumed to equal the square
root of the surface area:
k, = 1.41 10~4 (4-25;
1 d-85 0.000021 j
where,
K! = the liquid phase mass transfer coefficient, m/s;
V = the waste velocity, cm/s, equals the flow rate
cm3/s divided by both the depth and the width;
d = the depth of liquid flow into the sump, cm;
0.000021 = the reference liquid diffusion coefficient, cm2/s; and
D! = the liquid phase diffusion coefficient, cm2/s.
4-65
-------
-A---K} --•
Parameter
Depth of inlet flow base
Flow rate of inlet waste
Area of surface
Depth of sump liquid
Units
m
rnVs
m2
m
Value
.19
0.00188
1.77
1.37
Figure 4-16. Case E4. Open junction box with submerged flow.
4-66
-------
The equation for the gas phase mass transfer coefficient is
composed of the empirical model developed by McKay4.
The overall mass transfer from a two-resistance model, K0 ,
is a combination of the gas and the liquid mass transfer
coefficients :
K = ( — + - - - ) -1
K, 40.9 K H
1 9
where,
K0 = the overall mass transfer coefficient based
upon the liquid concentrations (m/s);
K! = the liquid phase mass transfer coefficient
(m/s) ;
Kg = the gas phase mass transfer coefficient
(m/s) ; and
H = the partition coefficient (atm-m3/mol) .
The air losses, falr, from the two-resistance model are as
follows :
K A
(4-27
where,
K0 = the overall mass transfer coefficient based upon
the liquid concentrations (m/s);
q = the liquid flow rate (m3/s);
A = the area of the exposed surface (m2) ; and
falr = the fraction of the component emitted to the air.
The area of the mass transfer surface is the area of the
surface of the wastewater in the sump or junction box.
The following example calculation illustrates the use of these
equations to estimate air emissions of toluene.
4-67
-------
The assumed conditions are:
H = 0.00643 atm-m3/mol, 357 (y/x);
q = 0.001888 m3/s;
d =19 cm (15 cm inside pipe diameter + 4 cm
submerged) pipe fraction full on surface
inlet = 0.5;
h = 1.37 m; and
A = 1.77 m2.
In the case where the waste discharges below the surface of
the sump, the water is assumed to flow across the entire cross
section of the sump. As an approximation, the cross sectional
area of that flow is assumed equal to the product of the square
root of the sump area (the sump width) and the depth of
wastewater in the sump. The waste velocity in the subsurface
flow case can be calculated as follows:
V- —i_ = -001888 = 0.0074 ™ .
A172 d (1.77172) ( .19) s
where V = the waste velocity across the surface of the sump, m/s.
If the waste discharges from a pipe that is located at the
surface of the sump, the water will flow across the surface of
the sump. The depth of flow, d, is one half of the pipe
diameter, or 0.075 meters. As a rough approximation, the cross
sectional area of that flow is assumed equal to the product of
the width of the sump and the depth of flow in the entrance
conduit. The waste velocity in the surface flow case can be
estimated as follows:
V- -^— - -001888 = 0.01892 ™ .
A172 d (1.77)172 (0.075) s
The values of the gas and liquid diffusivities are almost
identical to the reference values in the correlations, and the
4-68
-------
ratio can be assumed to equal unity. The liquid phase mass
transfer coefficient can be estimated as follows using Equation
4-25.
kl - 1.41 10- ' - = 4.52 10-
1 ig-85 0.000021 s
The overall mass transfer from a two-resistance model, K0 ,
is obtained from a combination of the gas and the liquid mass
transfer coefficients:
k = 0.00521 —
9 a
The air losses, falr, from the two-resistance model are
estimated using the following equation with the previously
defined terms:
K = ( + r1 = 4.5 10
4.52 10"6 (40.9) (0.00521) (0.00643)
K A
f . = °-
g
f . - (4.510-6) i'77 = 0.00422
(4.5 10"b) 1.77 + 0.001888
4.3.17 Case Fl Primary Clarifier Weir
An illustration of this case is presented in Figure 4-17
Albert Pincinci5 (11/7/89) presented equations for the air
emissions from a primary clarifier weir:
4-69
-------
In (r) = 0.042
In (r) = 0.077
where,
72 0.509
.0.623 —0.66
1 y
primary
secondary
(4-28)
(4-29)
r = Ci/Co deficit ratio;
Ci = inlet concentration;
Co = outlet concentration;
Z = distance of fall (m) ; and
q = flow rate per length of weir (m3/h-m) .
The fraction lost to the air over the weir is calculated
from the ratio r:
f. =
air
;4-3o;
The equivalent liquid mass transfer coefficient is as follows:
Kl = (m/s) = falr q/(3600 Z) (4-31).
The overall mass transfer from a two-resistance model, K ,
is a combination of the gas and the liquid mass transfer
coefficients:
where,
K
K =
o
-1
40.9 K K
9
the overall mass transfer coefficient based
upon the liquid concentrations (m/s);
the liquid phase mass transfer coefficient
(m/s);
the gas phase mass transfer coefficient
(m/s); and
the partition coefficient (atm-m3/mol) .
4-70
-------
Parameter
Drop distance of waterfall
Flow rate of water over the weir
Length of weir
Tail water depth
Units
m
m3/s
m
m
Value
1.22
0.065
2
1
Figure 4-17. CaseFl. Flow over a weir.
4-71
-------
The air losses, falr, from the two-resistance model are as
follows:
K Z
o
f . = 1 - EXP (-
q
where,
K0 = the overall mass transfer coefficient based upon
the liquid concentrations (m/s);
q = the liquid flow rate per length of the weir
(mVh-m) ;
Z = the distance of fall (m) ; and
falr = the fraction of the component emitted to the r.
The input parameters are as follows:
Diameter of clarifier, d 19.5 m
Depth 2.4 m
Flow of liquid, q 0.07 m3/s
Height of waterfall, Z 0.3m
Clarifier weir/circumference 1
Partition coefficient (Y/X) 305
The circumference of the clarifier is • d = 61.30 m
Kl from secondary clarifier model 1.05 x 10-3 g-mol/cm2-s
The gas mass transfer coefficient 1.13 x 10-5 g-mol/cm2-s
The area of the waterfall (61 m) (0.3 m) 18.4 m2
K - _
1 1
0.00105 (1.13xlO~5) (305)
K0 = 8.0 x ID'4 g-mol/cm2-s (1.44 x lO'4 m/s
The fraction of VO lost to air is
falr = 1 - EXP [-Ko (area)/q]
= 1 - EXP (-1.44 x ID'4 x 18.4/0.07)
= 0.0371 .
4-72
-------
4.3.18 Case F2 Secondary Clarifier Weir
Albert Pincinci1 presented the following equation for the
air emissions from a secondary clarifier weir:
In (r) = 0.077 z°-623 g°-66 (secondary)
where,
r = Ci/Co deficit ratio;
Ci = inlet concentration;
Co = outlet concentration;
Z = distance of fall (m); and
q = flow rate per length of weir (m3/h-m).
4.3.18 Case F3 General Weir Model
The preceding two models were obtained by Princinci for
volatiles flowing over clarifier weirs. For the general weir
model, the model presented by Nakasone6 is adapted for use for
weirs. The model is modified to account for gas phase
resistance.
In (r) = 0.0785 Z1-31 g°-428 ^0.310 (4-33;
where,
r = Cs/(Cs-Co) deficit ratio, assumes that there is no
oxygen before the weir;
Cs = saturated oxygen concentration;
Co = outlet oxygen concentration;
Z = distance of fall (m), includes 1.5 times the
distance from the weir top to the critical depth
above the weir;
q = flow rate per length of weir (m3/h-m); and
h = the tailwater depth (m).
The constants in the above equation are a function of the
flow rate and the distance of fall. Table 4-15 presents
constants that can be used in the above equation.
4-73
-------
It is assumed that the rate limiting step for the diffusion
of oxygen is the mass transfer in the liquid phase (oxygen is
only slightly soluble in the water). From this equation, a value
of the liquid phase mass transfer coefficient can be estimated
for organics, after correcting for the relative diffusion
coefficient of the organic and oxygen in water.
,4-34;
where,
the mass transfer coefficient of the organic
in water (m/s) ;
ln(r) = the natural log of the deficit ratio for
oxygen in the water flowing over the weir;
Dlv = the diffusion coefficient of the organic in
water (cm2/s) ;
Z = distance of fall (m) , includes 1.5 times the
distance from the weir top to the critical
depth above the weir; and
q = flow rate per length of weir (m3/h-m) .
The gas phase mass transfer coefficient of the flow at the
weir is assumed to equal 0.05 for benzene. This is approximately
the magnitude for mechanically aerated systems. The assumption
of a gas phase mass transfer coefficient of this magnitude will
lower the estimate of the oxygen transfer from the correlation by
only a few percent. Significantly lower gas phase mass transfer
coefficients will lower the predicted oxygen transfer to an
extent not predicted by the correlation.
D
kg = 0.05 ( =21 )0'66 (4-35)
go
4-74
-------
where,
Kg
Dgv
Dgo
0.05 =
the mass transfer coefficient of the organic in
air (m/s);
the diffusion coefficient of the organic in air
(cm2/s);
the diffusion coefficient of the reference
material in air (0.088 cm2/s); and
the assumed mass transfer coefficient of a
turbulent surface.
TABLE 4-15 . PARAMETERS FOR USE IN THE EQUATION OF
H. NAKASONE.
applicable range
Z <= 1.2
m
Z > 1.2 m
Z <=1.2 m
Z >1.2 m
q<=235
q<=235
q >235
q >235
constant
0.0785
0.0861
5.39
5.92
Z exponent
1.31
0.816
1.31
0.816
q
exponent
0.428
0.428
-0.363
-0.363
The value of the overall mass transfer coefficient is
estimated by combining the liquid and gas mass transfer
coefficients.
1 1
where,
K0
K
K =
o
-i
40.9 K K
9
the overall mass transfer coefficient based upon
the liquid concentrations (m/s);
the liquid phase mass transfer coefficient (m/s);
the gas phase mass transfer coefficient (m/s); and
the partition coefficient (atm-m3/mol) .
4-75
-------
The fraction of the components that are emitted to the air
is estimated by the following relationship:
f . -1-EXP (- ^ 360°SeC) (4-36)
g hr
where,
K0 = the overall mass transfer coefficient based upon
the liquid concentrations (m/s);
q = the liquid flow rate per length of the weir
(mVh-m) ;
Z = the distance of fall (m); and
falr = the fraction of the component emitted to the air.
In the example case of toluene, the deficit ratio is
calculated using the following equations:
In (r) = 0.0861 z0'816 g°-428 ^o.aio
where the above constants are for the situation where Z > 1.2 m
and q < 235 (See TABLE 4-15) .
r = Cs/(Cs-Co) deficit ratio, assumes that there
is no oxygen before the weir;
Cs = saturated oxygen concentration;
Co = outlet oxygen concentration;
Z = distance of fall (1.219 m), includes 1.5
times the distance from the weir top to the
critical depth above the weir;
q = flow rate per length of weir (117 m3/h-m);
and
h = the tailwater depth (1m).
The natural log of the deficit ratio, ln(r), is calculated
as 0.7769. Next, the liquid phase mass transfer coefficient is
estimated.
K = ^ ( -JZ )°-66 ln( r ) ( ^£_) (4-37)
1 Z D1 3600 s
lo
4-76
-------
1.219
0.86
2.5
o.7769
3600 s
K1 = 0.0102 —
s
TABLE 4-16 . PLANT PARAMETERS FOR A MODEL WEIR.
Parameter
distance of fall
flow rate
length of weir
flow rate per length
tail water depth
Units
m
m3/s
m
m3/h-m
m
Symbol
Z
Q
q
h
Value
1.22
0.065
2
117
1
Using (0.86/2.5) as the ratio of the diffusion coefficients of
toluene and air, the estimated value of kl is 0.0102 m/s.
Next, the gas phase mass transfer coefficient of toluene is
estimated, based upon the reference mass transfer coefficient of
benzene.
D
k = 0.05
9
gv \ o. 6 6
D
go
where,
Kg =
Dgv =
Dgo =
0.05
the mass transfer coefficient of the organic
in air (m/s);
the diffusion coefficient of the organic in
air (0.087 cm2/s);
the diffusion coefficient of the reference
material in air (0.088 cm2/s); and
the assumed mass transfer coefficient of a
turbulent surface.
4-77
-------
k =0.05 (
9 0.088
k = 0.0496 —
9 a
The estimated gas phase mass transfer coefficient is 0.0496.
Next, the overall mass transfer coefficient is calculated:
K = ( — + - -
K7 40.9
9
where,
K0 = the overall mass transfer coefficient based
upon the liquid concentrations (m/s);
K1 = the liquid phase mass transfer coefficient
(0.0102 m/s) ;
Kg = the gas phase mass transfer coefficient
(0.49624 m/s) ; and
K = the partition coefficient (0.00668 atm-
mVmol) .
K - ( _ l— + _ I _ )-
0.01024 (40.9) (0.0496) (0.00668)
K = 0.00583 —
s
The overall mass transfer coefficient is 0.00583 m/s.
Next, the fraction of air emissions are estimated.
f . -1-EXP (- ^ 360°SeC)
hr
4-78
-------
where,
K0 = the overall mass transfer coefficient based upon
the liquid concentrations (0.00583 m/s);
q = the liquid flow rate per length of the weir(117
m3/h-m) ;
Z = the distance of fall (1.2192 m); and
falr = the fraction of the component emitted to the air.
= 1 - EXP (- °-00583 1-219 3600 sec
a-ir i i T far
The fraction of toluene that is emitted to the air is 0.20.
f . = 0.2
air
4.4 REFERENCES
1. "Measurement of Hazardous Air Pollutant Emissions from
Wastewater Collection System Components," Enviromega,
Burlington, Ontario, Canada, April 7, 1993.
2. Pescod and Price, Journal WPCF, vol. 54, no. 4, April 1982,
p. 393.
3. Owens, Edwards, and Gibbs, (Int. J. Air Wat. Poll., vol. 8
(1964) , pp. 469-486) .
4. McKay and Yeun. "Mass Transfer Coefficient Correlations for
Volatilization of Organic Solutes from Water, Environmental
Science and Technology, 1_7_: 211-217 . 1983.
5. Pincinci, A. "Calculating Emissions of Wolatile Organic
Compounds from Wastewater Treatment Plants", (4/28/88)
6. Nakasone, "Study of Aeration at Weirs and Cascades.",
Journal of Environmental Engineering, ASCE., 113, 64, 1987
4-79
-------
5.0 SURFACE IMPOUNDMENTS AND OPEN TANKS
This section discusses the approach used to estimate air
emissions from surface impoundments and open top tanks. The
emission models are described, model facilities are defined, and
example calculations are presented.
5.1 NARRATIVE DESCRIPTION OF EMISSIONS AND MODEL UNITS
Emissions from surface impoundments and open tanks originate
from the uncovered liquid surface that is exposed to the air. The
model used to estimate emissions from the liquid surface is based
on an overall mass transfer coefficient that incorporates two
resistances to mass transfer in series—the liquid-phase
resistance and the gas-phase resistance. Numerous correlations
are available to estimate the individual mass transfer coeffi-
cients (or resistances), and they depend upon the compound's
properties and the system's parameters. The recommended
correlations and their applicability are described in subsequent
sections. The emission estimating procedure also incorporates a
flow model that describes the method of operation. For
flowthrough systems, the impoundment's or tank's contents may be
completely mixed, plug flow, or somewhere in between with varying
degrees of backmixing or axial dispersion. Biologically active
impoundments and aeration tanks can be designed for either
completely mixed or plug flow, and both types of flow models are
discussed for these types of systems. For disposal impoundments,
the contents are assumed to be well mixed, and the bulk
concentration is expressed as a function of time. An expression
for biodegradation is incorporated for those units specifically
5-1
-------
designed for biodegradation, such as treatment impoundments or
wastewater treatment tanks. For these units, the relative rates
of air emissions and biodegradation are determined to assess the
predicted extent of each mechanism.
The general approach that is used to estimate emissions
compares the relative rates of air emissions, biodegradation, and
removal with the effluent. Several different types of model units
are presented and include mass transfer to the air from quiescent,
mechanically aerated, diffused-air, and oil-film liquid surfaces.
The other major difference among the types of model units is the
type of flow model that is used. For flow-through systems, the
degree of mixing can range from complete mixing to plug flow (no
mixing), and both cases are presented. For disposal units with no
flow out, emissions are a function of time, and average emissions
are estimated for some specified time since disposal. The major
difference in the emission equations is the liquid-phase
concentration that is used for the driving force for mass transfer
to the air. The simplest case is represented by well-mixed
systems in which the driving force is represented by CT, the
L
liquid-phase concentration in the bulk liquid, which is also equal
to the effluent concentration. Relative removal rates can be com-
pared for this well-mixed case from a simple material balance.
For plug flow, integration is required because the driving
force for mass transfer changes as the liquid flows through the
system. This concentration is a function of location or time
(which are equivalent in plug flow) and is expressed as C,
(denoting a dependence on time). The effluent from a plug flow
system is denoted as C . For disposal impoundments, the driving-
force concentration changes with time and is also denoted as C,;
however, there is no effluent from a disposal impoundment. The
integration required for plug flow is from t = 0, when the
material first enters the unit, to t = residence time, when the
material leaves the unit. For disposal units, the integration is
from t = 0, when the material is first placed in the unit, to
5-2
-------
t = time since disposal, which must be specified to estimate
average emissions. The integrated forms of these emission equa-
tions are very similar.
The well-mixed flow model is recommended and is the model
used in the computer program accompanying this report. This flow
model is more generally applicable than plug flow, the
calculations are more straightforward, and the two types give
similar results. The only exception is a flowthrough impoundment
with an oil film surface, which uses the plug flow model because
the oil film inhibits mixing. Both models yield an estimate of
air emissions, biodegradation, and the quantity leaving with the
effluent. It is important to recognize that the quantity leaving
with the effluent may also eventually contribute to air emissions,
especially for treatment units in series or for discharges to
streams or publicly owned treatment works.
Equations are presented for estimating the various removal
rates, and example calculations for different types of
impoundments are also provided. Example calculations are not
presented separately for open tanks because the procedure is
analogous to that used for impoundments. In general, open tanks
will have different input parameters that will account for
differences in emission rates compared to impoundments. For
example, the liquid surface area for open tanks will be less, and
the fetch-to-depth (F/D) ratio will be much lower for tanks. If
the open tank has a wind barrier to reduce the wind velocity, the
reduced wind velocity can be used in the mass transfer
correlations. In addition, the modeling approach accounts for the
shorter retention times in tanks (on the order of hours) compared
to impoundments (on the order of days). For open tanks, the mass
transfer correlation of Springer is recommended for windspeeds
less than 3.25 m/s, and the correlation of MacKay and Yeun is
recommended for windspeeds greater than 3.25 m/s. Both are
discussed in the following section.
5-3
-------
5.2 QUIESCENT SURFACES WITH FLOW
5.2.1 Emission Model Equations
The primary focus on emissions from impoundments and
wastewater treatment tanks is on aqueous solutions contaminated
with organics because aqueous waste is the most common waste type
handled in these facilities. For aqueous systems, the basic
relationship describing mass transfer of a volatile constituent
from the open liquid surface to the air is:
E = KACL (5-1)
where,
E = air emissions from the liquid surface, g/s;
K = overall mass transfer coefficient, m/s;
2
A = liquid surface area, m ; and
CT = concentration of constituent in the liquid phase,
/ 3
g/m .
The overall mass transfer coefficient (K) is estimated from a
two-phase resistance model that is based on the liquid-phase mass
transfer coefficient (k in m/s), the gas-phase mass transfer
coefficient (k in m/s), and Henry's law constant in the form of a
partition coefficient (Keq). The two resistances act in series
and the overall resistance is expressed as:
11 i
(5-2)
_ _
K kT k,, Keq
J_i (j
where,
K = overall mass transfer coefficient, m/s;
k = liquid-phase mass transfer coefficient, m/s;
j_i
k,, = gas-phase mass transfer coefficient, m/s; and
(j
Keq = equilibrium constant or partition coefficient,
concentration in gas phase/concentration in liquid
phase where both concentrations are in the same
units.
5-4
-------
Henry's law constant (H in atm»m /g mol) is estimated for the
constituents of interest by dividing the constituent's vapor
pressure (in atmospheres) by its solubility in water (in
g mol/m ). The equilibrium constant is estimated by:
Keq = H/RT (5-3)
where,
H = Henry's law constant, atm»m /g mol;
R = universal gas constant, 8.21 x 10 atm»m /g mol»K;
and
T = temperature, K.
For a standard temperature of 25 °C, the expression for Keq
reduces to:
Keq = 40.9 x H (5-4)
The units associated with Keq in Equation (5-4) are the ratio of
gas-phase to liquid-phase concentrations and require that both be
expressed in the same units of mass/volume.
Several mathematical models have been developed to estimate
the individual liquid- and gas-phase mass transfer coefficients.
The models are based on different systems, constituents, and
sometimes different theoretical considerations. Many of these
models yield similar results. The procedures used in this section
to estimate the individual mass transfer coefficients rely
primarily on existing mass transfer correlations that are believed
to be generally applicable.
The liquid-phase mass transfer coefficient (k ) has been
j_i
shown to be a function of the constituent's diffusivity in water,
1 2
windspeed, and liquid depth. ' Work performed at the University
of Arkansas by Springer et al. confirmed these effects and
resulted in the correlations given in Table 5-1. Springer used
simulation studies in a wind tunnel water tank of a constant fetch
(2.4 m) and variable depth (4.7 cm to 1.2 m). Fetch is defined as
the linear distance across the liquid surface in the direction of
the wind flow, and the F/D ratio is defined as the fetch divided
5-5
-------
by the depth of the impoundment. Ethyl ether was used as the
volatile component in the desorption experiments, in which the
wind velocity and F/D ratio were varied. Springer's results shown
in Table 5-1 yield three different correlations for k that depend
j_i
upon the combination of windspeed and F/D ratio of interest.
Springer's model implies that k is constant for windspeeds of 0
j_i
to 3.25 m/s. Although Springer examined only the mass transfer of
ethyl ether, his results are extrapolated to other compounds by
the ratio of the compound's and ether's diffusivities in water to
the 2/3 power. The windspeed in Springer's correlation is defined
as the windspeed 10 m above the liquid surface. For practical
application of his correlation, typically reported values of
windspeed are used. Springer's model does not include the case in
which the F/D ratio is less than 14 and the windspeed is greater
than 3.25 m/s. For this specific case, k was estimated from
78
MacKay and Yeun's correlation shown in Table 5-1. ' MacKay and
9
Yeun did not address the effect of depth; however, their
correlation is based on data from 11 organic compounds in a well-
mixed system, the compounds represent a broad range of Henry's law
constants, and their general correlation is applicable for the
case described above that is not covered by Springer's
correlation.
The gas-phase coefficient (]<:„) was estimated from the
10
correlation of MacKay and Matasugu as shown in Table 5-1. This
correlation was developed from experiments on the evaporation of
isopropyl benzene, gasoline, and water into air. These
researchers verified that previous work, which assumed that the
wind velocity profile follows a power law, could be used to
quantify the rate of evaporation from a smooth liquid surface.
The result was a correlation that expressed ]<:„ as a function of
(j
windspeed and the fetch or effective diameter of the liquid
surface.
The individual mass transfer coefficients estimated from the
correlations in Table 5-1 are used in Equation (5-2) to estimate
5-6
-------
TABLE 5-1. EQUATIONS FOR CALCULATING INDIVIDUAL MASS
TRANSFER COEFFICIENTS FOR VOLATILIZATION OF ORGANIC
SOLUTES FROM QUIESCENT SURFACE IMPOUNDMENTS
Liquid phase
4
et al. (for all
2/3
4
Springer et al. (for all cases except F/D3.25 m/s):
_ D ^
k = 2.78 x 10 *- —| (0 < U10 < 3.25
* ether* (All F/D ratios)
D 2/3
-Q - 7 9 * w *
kT = [2.605 x 10 (F/D) + 1.277 x 10 '] uf_ *- —*
L 10 *ether*
(U10 >3.25) (m/s) (14 < F/D <51.2)
_7 9 # D :,2/3 (Uin>3.25) (m/s)
kT = 2.611 x 10 U, *— *
10 *D ., * (F/D>51.2'
ether v
where,
U1n = windspeed at 10 m above the liquid surface, m/s;
2
D = diffusivity of constituent in water, cm /s;
2
D ,, = diffusivity of ether in water, cm /s; and
F/D = fetch-to-depth ratio (fetch is the linear distance
across the impoundment).
Gas phase
MacKay and Matasugu (in Hwang ):
i A QO in~3 TT°-78 c -0.67 , -0.11 , , .
k_= 4.82 x 10 U Sc« d (m/s)
(j (j e
where,
U = windspeed, m/s;
u G
Sc« = Schmidt number on gas side = —^ ;
G ' DG a
LU = viscosity of air, g/cm»s;
G
(continued)
5-7
-------
TABLE 5-1 (continued)
. 3
G = density of air, g/cm ;
2
D = diffusivity of constituent in air, cm /s;
a
0.5
d = effective diameter of impoundment = |-—J , m; and
2
A = area of impoundment, m .
Liquid phase
MacKay and Yeun6 (for F/D 3.25 m/s):
kL = 1.0 x 10~6 + 34.1 x 10~4 U* ScL~°'5 (U*>0.3) (m/s)
kL = 1.0 x 10~6 + 144 x 10~4 U*2'2 ScL~°'5 (U*< 0.3) (m/s)
where,
U* = friction velocity (m/s) = 0.01 UIQ (6.1 + 0.63 U^)0'5,
U,„ = windspeed at 10 m above the liquid surface, m/s;
ScT = Schmidt number on liquid side = —;=— ;
J_i -j- JJW
UT = viscosity of water, g/cm»s;
•T = density of water, g/cm ; and
2
D = diffusivity of constituent in water, cm /s.
w
-------
the overall mass transfer coefficient. The equilibrium constant
for a constituent dissolved in water at 25 °C is estimated from
Equation (5-4) . However, an estimate of the concentration in the
liquid phase (CT ) is needed in Equation (5-1) to estimate
j_i
emissions .
The concentration CT in Equation (5-1) is the driving force
j_i
for mass transfer. For an impoundment that is instantly filled
with waste, the driving force (CT ) is the initial concentration
j_i
in the waste. However, this concentration will decrease with
time as the constituent is lost to the air, which suggests that
emissions may also decrease with time (assuming constant K and
A) . For flowthrough systems, the concentration may be cyclical
if the loading of the process is cyclical. Continuous
flowthrough systems may attain some equilibrium concentration.
The flow model assumed for quiescent impoundments and tanks
with no biodegradation is that the contents of the system are
well mixed and that the bulk concentration (driving force) in the
system is equal to the effluent concentration, CT . A material
j_i
balance around this system yields:
QC. = KACT + QCT (5-5)
or
= QCi/(KA+Q) (5-6)
where,
Q = volumetric flow rate, m /s;
C. = initial concentration in the waste, g/m ;
„ = equilibrium or bulk concentration in the
L impoundment, g/m ;
K = overall mass transfer coefficient, m/s; and
2
A = liquid surface area, m .
The well-mixed assumption is made for the sake of simplicity
and assumes that bulk convection and wind-induced eddies combine
to mix the basin contents. Axial dispersion in the flow
5-9
-------
direction is also possible, and some systems may be designed
specifically for plug flow (e.g., some biological treatment
tanks). An assumption of plug flow instead of well-mixed flow
would yield slightly higher estimates of emissions; however, the
difference is small. Calculations presented by Thibodeaux for an
aerated basin that was well-mixed or had plug flow showed that
the plug-flow assumption yielded estimates that were higher by
11 percent for acetaldehyde, 5 percent for acetone, and 0 percent
for phenol.
The approach described to estimate emissions from quiescent
impoundments with no biodegradation includes the following steps:
1. Estimate the individual mass transfer coefficients from
Table 5-1.
2. Estimate the equilibrium constant from Equation (5-3).
3. Estimate the overall mass transfer coefficient from
Equation (5-2).
4. Estimate the liquid-phase concentration from Equation
(5-6) .
5. Estimate emissions from Equation (5-1) .
The major assumptions associated with this procedure are:
• The two-resistance model and the correlations for the
individual mass transfer coefficients are applicable to
the system of interest.
• The impoundment's contents are well mixed.
• There is no significant removal by biodegradation,
seepage, adsorption, or other forms of degradation.
• The waste material of interest is aqueous waste with no
separate organic phase.
• The estimate of Henry's law constant (equilibrium
partitioning between the vapor and liquid) is
reasonably accurate.
The recommended procedure for quiescent impoundments is to
assume that the liquid is well mixed. This assumption is used in
5-10
-------
the computer model accompanying this report and is illustrated in
the example calculations. However, impoundments and tanks with
quiescent surfaces can also be designed for plug flow with the
use of baffles or other design techniques to reduce the extent of
backmixing. In a plug-flow system, the rate of air emissions at
any point in the system changes as the material flows through the
system. There is no uniform liquid concentration within the
plug-flow unit as there was in the well-mixed system, and the
lowest concentration occurs in the effluent (i.e., there is no
backmixing of the effluent with the influent) . For plug flow,
the rate of disappearance of a compound by air emissions is given
by:
, V }
I - /
= K A Ct (5-7)
where,
t = time, s;
V = volume,
and with the other symbols as previously defined.
C, = concentration after the plug has traveled t seconds;
= time, s;
V = volume, m ;
Rearranging Equation (5-7) yields:
KA e.
Ct V
Integrating Equation (5-8) from C, = C. at t = 0 to C, = CT at
HI H J-i
t = V/Q (one residence time) gives:
- ^p (5-9)
where CL = effluent concentration, g/m ,and with the other
symbols as previously defined.
5-11
-------
The residence time, • in seconds, equals V/Q and V = AD
(area times depth); consequently, A/Q = */D. Substituting into
Equation (5-9) yields an equivalent expression:
The ratio CT/C. represents the fraction removed with the
j_i i
effluent; therefore, 1 - CT/C. represents the fraction that is
j_i i
emitted (f . ) from the plug-flow system:
air ^ ^ •*
Jair
c,
(5-11
;
The average emission rate is calculated from:
E=fairQC, (5-12)
where,
E = emissions, g/s;
f . = fraction emitted from Equation (5-11);
air M
Q = flow rate, m /s; and
3
C. = influent concentration, g/m .
5.2.2 Model Plant Parameters for Quiescent Impoundments
A model facility was developed for quiescent impoundments to
illustrate the emission estimating procedure. A 1981 survey
12
compiled by Westat showed that the median surface area for
2
storage impoundments was approximately 1,500 m and that the
median depth was 1.8 m. Detention times ranged from 1 to 550
days, with over half of the values at 46 days or less. For this
example, a detention time of 20 days was chosen. The area and
depth yield a total volume of 2,700 m , and the detention time of
3
20 days yields a flow rate of 1.6 L/s (0.0016 m /s).
Meteorological conditions are also needed as input
parameters for the emission models. For this emission estimate,
5-12
-------
a standard temperature of 25 °C and a windspeed of 4.47 m/s
(10 mi/h) were used. Benzene was chosen as an example
3
constituent at a concentration of 10 g/m (10 ppm) to estimate
emissions from the model facility. The properties of benzene
-3
that are used include Henry's law constant (5.5 x 10
3 2
atm»m /g mol), diffusivity in air (0.088 cm /s), and diffusivity
/~ O
in water (9.8 x 10 cm /s). Table 5-2 lists the input
parameters for the estimate of emissions given in Section 5.2.3.
5.2.3 Example Calculation for Storage Impoundments
This section presents a step-by-step example calculation for
emissions from storage impoundments. The equations described in
Section 5.2.1 are used with the model unit parameters given in
Section 5.2.2 to estimate emissions from an aqueous waste
containing 10 g/m of benzene.
a. Calculate liquid-phase mass transfer coefficient, k . Use
Springer's model (see Table 5-1) :
0.5 0.5
a -TQ 3 i c. n n
Effective diameter = *AJied * x 2 = *1'ouu* x 2 = 43.7 m
F/D = Effective diameter/depth = ^3'7 =24.3 .
1 . o
Windspeed = 4.47 m/s (UIQ > 3.25 m/s)
F/D = 24.3
* D * 0-67
- Q -79* w *
kT = [2.605 x 10 (F/D) + 1.277 x 10 '] U., n *^ —* m/s
L 10 *ether*
where,
U1n = windspeed = 4.47 m/s
-L U ,- f-*
D = 9.8 x 10 cm /s (benzene)
w _, „
D , , = 8.5 x 10 = cm /s (ether)
ether
F/D = 24.3.
5-13
-------
TABLE 5-2. INPUT PARAMETERS—STORAGE IMPOUNDMENT
Area
Depth
Volume
Retention time
Flow
Temperature
Windspeed
Constituent
Influent concentration
Henry's law constant
Diffusivity in air (benzene)
Diffusivity in water (benzene]
Diffusivity in water (ether)
Viscosity of air
Density of air
1,500 m2
1.8m
2,700 m3
20 days
0.00156 i
25 °C
4.47 m/s
Benzene in water
10 g/m3
-3 3
5.5 x 10 atm»m /g mol
0.088 cm2/s
9.8 x 10~6 cm2/s
8.5 x 10~6 cm2/s
1.81 x 10 g/cm»s
-3 3
1.2 x 10 g/cm
5-14
-------
Then,
_6* 0-67
-Q -7 9*Q8v~IO*
k = [2.605 x 10 (24.3) + 1.277 x 10 ](4.47) * -K*
*8.5 x 10 *
kL = [2.605 x 10~9 (24.3) + 1.277 x 10~7] (4.47)2 (1.1)
kT = 4.2 x 10~6 m/s.
J_i
b. Calculate gas-phase mass transfer coefficient, k . Use
MacKay and Matasugu (see Table 5-1):
i A QO i n~3 TT°-78 c -0.67 , -0.11 , , .
k_ = 4.82 x 10 U Sc~ d (m/s)
(j (j e
where,
U = windspeed, 4.47 m/s
^ _ Schmidt No. _ viscosity of gas
G for gas (gas density)(diffusivity of i in gas)
Gas = air
Viscosity (air) = 1.81 x 10 g/cm»s
Density (air) = 1.2 x 10 g/cm
2
Diffusivity (benzene in air) = 0.088 cm /s
-4
s = 1.81 x 10 g/cm»s = 1 71
G (1.2 x 10~3 g/cm3) (0.088 cm2/s)
Then,
d = effective diameter = 43.7 m
e
^ = (4.82 x 10 3) (4.47)0'78 (1.71) °'67 (43.7
-3
=7.1x10 m/s
c. Calculate overall mass transfer coefficient (K) from
Equation (5-2 ) :
1 1 1
K
ea
where
-3 3
„ _H _ 5.5x10 m «atm/mol _ „ 295
R^T 3
(8.21 x 10~5) ( a ) (298 K)
5-15
-------
Then,
^ = l + l_ = 2.39 x 105
K 4.2 x 10~6 (0.225)(7.1 x 10~3)
K = 4.2 x 10~6 m/s .
d. Estimate emissions for a well-mixed system:
QC. = KC A + QC (from material balance of Equation (5-5))
QC.
C =
o KA + Q
Retention time = 480 h
Volume = 2,700 m3
Where,
*9700m**1h* "3
Q = flow rate = r^O h * *3,600 s* = °-00156 m /s-'
C. =10 g/m3; and
K = 4.2 x 10~6 m/s.
c = (0.00156 m3/s) (10 g/m3) =1.98 g/m3
0 (4.2 x 10~6 m/s) (1,500 m2) + (0.00156 m3/s)
A = 1,500 m2
Air emissions = KC A (Equation 5-2)
= (4.2 x 10~6 m/s)(1.98 g/m3)(1,500 m2) = 0.012 g/s
=3.8 Mg/yr
Estimate emissions for a plug-flow system:
f . = 1 - exp (-K»/D) (Equation 5-11)
cL 1 IT
K = 4.2 x 10~6 m/s (Step c)
. = 480 h = 1.73 x 106 s
D = 1 . 8 m
f . = 1 - exp (-4.2 x 10~6 m/s»1.73 x 106s/1.8 m) = 0.98
cL 1 IT
5-16
-------
E = f . Q C (Equation 5-12)
3.1 IT O
f . =0.98;
air _
Q = 0.00156 m /s; and
C = 10 g/m ?
E = (0.98) (0.00156 m3/s) (10 g/m3)
E = 0.015 g/s = 0.47 Mg/yr.
5.3 BIODEGRADATION
This section identifies some of the major design features of
biological treatment processes, such as activated sludge units
and impoundments designed for biodegradation. Mathematical
models for biodegradation are also presented and incorporated
into predictive fate models.
5.3.1 Description of Biological Active Systems
The activated sludge process is an aerobic biological
treatment in which the pollutants are degraded by microorganisms
suspended uniformly in the reaction tank. Oxygen is introduced
by mechanical means, and the microorganisms are maintained by
recycling the activated sludge that is formed. In most units,
the sludge is removed by settling in a separate unit, a portion
of the sludge is recycled, and a small portion is wasted (removed
from the system) on a continuous basis. Oxidation or stabiliza-
tion impoundments and aerated impoundments are used to treat
entire plant wastes as well as to polish the effluent from other
treatment processes. Solids usually settle out in the
impoundment or are removed in a separate vessel. Generally, the
solids are not recycled; however, if the solids are returned, the
13
process is the same as a modified activated sludge process.
Typical design parameters for an activated sludge process
are given in Table 5-3. Two of the most commonly used parameters
are the food-to-microorganism (F/M) ratio and residence time.
The F/M ratio describes the organic loading on the biological
system and is calculated as the weight of BOD,- (biochemical
5-17
-------
TABLE 5-3. DESIGN PARAMETERS FOR ACTIVATED SLUDGE PROCESSES
14
Process
kg BOD/kg
F/M,d
Loading,
3
MLSS,
b
biomass»day kg BOD/m »day g/L
Retention
time, h
Conventional0 0.2-0.4
CSTRd 0.2-0.6
Contact 0.2-0.6
Stabilization
Extended aeration 0 . 05-0.15
09 systems 0.25-1.0
0.3-0.6 1.5-3.0 4-f
0.8-2.0 3.0-6.0 3-5
1.0-1.2 1.0-3.0e 0.5-le
4.0-10f 3-6f
0.1-0.4 3.0-6.0 18-36
1.6-3.3 6.0-8.0 1-3
f
F/M = Food to microorganism ratio.
'MLSS = Mixed liquor suspended solids.
Plug flow design.
CSTR = Continuous stirred-tank reactor
^
'Contact unit.
Solids stabilization unit.
5-1!
-------
oxygen demand from a 5-day test) that enters the system in a 24-
hour period divided by the total weight of biological solids in
the system. The biological solids may be roughly estimated from
the mixed liquor suspended solids (MLSS) if substantial
quantities of inorganics (such as silt) are not present. If
inorganic solids are present, the biological solids may be better
approximated by the mixed liquor volatile suspended solids
15
(MLVSS). For municipal wastewater systems, the volatile solids
comprise about 60 to 80 percent of the total suspended solids in
the sludge; consequently, in the absence of a direct measurement
of MLVSS, the biological solids in municipal wastewater can be
estimated as 60 to 80 percent of the total suspended solids.
Conventional plants, which use an activated sludge process that
has long and narrow basins designed to approach plug flow,
operate with an F/M ratio of 0.2 to 0.4, but values as low as
0.05 are not unusual. High F/M values indicate a high loading,
as from a sudden influx of organics or the loss of biological
solids, and will lead to a deterioration in effluent quality.
Aeration tanks are usually constructed of reinforced
concrete, are open to the atmosphere, and are usually rectangular
in shape. Treatment plants may consist of several tanks,
operated in series or parallel. Some of the largest treatment
plants may contain 30 to 40 tanks arranged in several groups or
, . . 18
batteries.
Typical parameters associated with biologically active
impoundments are given in Table 5-4. The loading parameter is
expressed in terms of area or volume, and typical retention times
in aerated impoundments range from 7 to 20 days. The level of
suspended solids in these impoundments is over an order of
magnitude less than the level in activated sludge processes.
Although the parameters in Table 5-4 are listed as "typical,"
large variations exist among real facilities, and at a single
5-19
-------
TABLE 5-4. IMPOUNDMENTS DESIGNED FOR BIODEGRADATION19
Type
Application
Typical daily
loading,
kg BOD5/m »day
Retention Typical Suspended
time, day depth,m solids, g/L
Facultative
Aerated
Aerobic
Anaerobic
Raw municipal wastewater
Effluent from primary
treatment, trickling
filters, aerated ponds,
or anaerobic ponds
Industrial wastes
Overloaded facultative
ponds
Situations where limited
land area is available
Generally used to treat
effluent from other
processes, produces
effluent low in soluble
BOD,- and high in algae
solids
Industrial wastes
0.0011 - 0.0034 25-180
0.008 - 0.32
0.16 - 0.80
7-20
0.021 - 0.043b 10-40
20-50
1.2-2.5 0.11-0.40
2-6
0.26-0.30
0.3-0.45 0.14-0.34
2.5-5
0.08-0.16
b
Based on a typical depth of 2 m.
Based on a typical depth of 0.4 m.
5-20
-------
facility the values may change with time. For example, a study
conducted over 12 months at an aerobic impoundment used to treat
municipal wastewater reported suspended solids levels of 0.02 to
21
0.1 g/L and volatile suspended solids of 0.01 to 0.06 g/L.
Another study of eight quiescent impoundments at four
different sites with confirmed biological activity estimated
active biomass concentrations from the rate of oxygen
consumption that ranged from 0.0014 to 0.22 g/L with an average
of 0.057 g/L.22
The biomass concentration is an important parameter in
estimating biodegradation rates. The best value to use for a
specific site is a direct measurement such as volatile
suspended solids for the system of interest. In the absence of
site-specific data, a number may be chosen from the ranges for
suspended solids given in Tables 5-3 and 5-4. Alternatively,
typical or default values for biomass concentration given in
Table 5-5 may be used.
TABLE 5-5. TYPICAL OR DEFAULT VALUES FOR BIOMASS CONCENTRATION3
Unit Biomass concentration, g/L
Quiescent impoundments 0.05
Aerated impoundments 0.25
Activated sludge units 4.0
These values are recommended for use in the emission equa-
tions when site-specific data are not available.
Based on the range (0.0014 to 0.22) and average (0.057)
from actual impoundments as discussed in the text.
From the data in Table 4-4 for aerated impoundments.
Assumes biomass is approximated by the suspended solids
level. Range is typically 0.05 to 0.30.
nyiidrange value from Table 4-3 for CSTR based on mixed
liquor suspended solids.
5-21
-------
The major mechanisms of organic removal in biologically
active systems include biodegradation, volatilization, removal
with the effluent, and removal by adsorption on the waste
sludge. A study by Petrasek et al. of purgeable volatile
organics in a pilot-scale wastewater treatment system showed
that less than 0.4 percent (generally less than 0.1 percent) of
23
the volatiles were found in the waste-activated sludge.
Bishop, in a study of municipal wastewater treatment, concluded
that only a modest amount of purgeable toxics were transferred
24 25
to the sludge. Hannah et al. found that the concentrations
of volatile organics in sludges from pilot-scale systems were
generally comparable to or less than the corresponding concen-
trations in the process effluent. This indicated that volatile
organics do not have a high affinity for wastewater solids and
do not concentrate in the sludges. Kincannon and Stover found
that 0 to 1 percent of three compounds (1,2-dichloroethane,
O £.
phenol, and 1,2-dichlorobenzene) was adsorbed on the sludge.
Melcer, in a review of biological removal studies, concluded
that polycyclic aromatic hydrocarbons, pyrene, anthracene,
fluoranthene, and chrysene were the most commonly occurring
27
priority pollutants found in sludges. These studies suggest
that the compounds most likely to be emitted to the air
(volatiles) do not concentrate on sludges; however, some of the
relatively nonvolatile organics may be adsorbed. Consequently,
the modeling approach presented in this section assumes that
the removal of volatile organics with the waste sludge is not
significant. The major removal mechanisms that are considered
include volatilization, biodegradation, and removal with the
effluent.
5-22
-------
5.3.2 Rate of Biodegradation
Numerous models have been proposed for the removal of
organic compounds by biodegradation and include design
equations for activated sludge systems and stabilization or
28 29
oxidation impoundments. ' There is general agreement in the
literature that, for high organic loadings relative to biomass,
the biodegradation rate is zero-order with respect to
concentration (i.e., the rate is independent of organic
concentration). For lower residual levels, the rate
becomes first order with respect to concentration or follows
Monod-type kinetics. ' ' The Monod-type biodegradation
rate equation can be written as follows:
-v- — \T IS V\ C* / { IS J-C* \ / R 1 ^ \
r_ — V rvm = Y D.U/lrv+U-r) ( O~ ±j)
TJ max -i o T
D l o i i
where,
r_ = biodegradation rate, g/s;
JD
V = volume, m ;
b. = biomass concentration, g/m ;
K = maximum rate constant, g/s-g biomass;
max ' ^ y_
CT = component concentration, g/m ; and
3
K = half saturation constant, g/m .
s
The Monod model was originally developed to describe
microbial growth rates for a single microbial population
based upon a single, rate-limiting substrate. A yield
coefficient was subsequently employed to determine the
utilization rate of that substrate. For convenience of use,
the biodegradation rate model given in Equation (5-13) has
been written directly for component disappearance in terms of
overall biomass concentration. It is assumed that Equation
(5-13) applies to each organic constituent in the waste
(although the rate constants will be different for each
5-23
-------
constituent), and that the biodegradation of any one
constituent is independent of the concentrations of other
constituents. Subsequent references to the Monod or the
Monod-type model in this report refer specifically to
Equation (5-13). The significant features of this model are:
1. At high concentrations (specifically, C » K ), C
dominates the denominator and can, therefore, be
eliminated from Equation (5-13) . The
biodegradation rate is then independent of (i.e.,
zero order with respect to) the component
concentration.
2. At low concentrations (C < K , and the
biodegradation rate becomes directly proportional
(i.e., first order with respect) to the component
concentration. The apparent first-order rate
constant is: K1 = K /K .
_L max s
Theoretical Monod curves for several different compounds
are presented in Figure 5-1 to illustrate these features.
A literature review was conducted to determine
appropriate rate constants for the Monod model. References
that served as primary sources of biodegradation rate data
included: Fitter, Kincannon et al., Petrasek et al.,
Q £.
and Hannah et al. Data obtained from each reference
included rate constants as reported, influent concentrations,
effluent concentrations, biomass concentration, retention
time (RT), and fraction of the amount of component removed by
biodegradation (F_). Using this information and field data
JD
collected during specially designed biodegradation rate
studies, Coburn et al. developed a base of component-specific
biological removal rates that contains nearly 500 entries and
37
removal data for 90 different organic constituents.
5-24
-------
Biorates of three compounds
I
o
2
8
5 -r
4
3
2
1
0
1
-2
-3
-2-10123
Natural log of concentration (in mg/L)
Phenol
Legend
Benzene
Chloroform
Figure 5-1. Theoretical relationship between concentration
and biodegradation rates normalized by the amount of biomass
as predicted using the Monod model for phenol, benzene, and
chloroform.
5-25
-------
Appendix C contains a listing of the Monod parameters
for 88 compounds. Some of these 88 compounds do not have a
listing for both Monod parameters. The values of the Monod
parameters presented in Appendix C were selected primarily
from the Coburn data base. The following paragraphs describe
how these biorates were derived.
For most compounds, there were inadequate biodegradation
rate data to determine the Monod rate constants using
traditional methods (e.g., Lineweaver-Burke plot). However,
when reported, values for K were generally between 1 and 10
o
mg/L for a variety of different compounds. Thus, the Monod
constant, K , was calculated from organic removal data when
max
high concentrations (CT > 10 mg/L) were employed by assuming
J_i
strict zero-order kinetics as follows:
K = F_.(C. - CT)/[(RT) b.] (5-14
max B i L i
where,
F = the fraction of component removal attributed
to biodegradation;
C = inlet concentration, g/m ;
CT = bulk liquid and effluent concentration, g/m ;
and
(RT) = residence time, s.
Note that, with zero-order kinetics, Equation (5-14) applies
to both continuous, well-mixed systems and to plug-flow and
batch systems.
5-26
-------
The half-saturation constant K was estimated (knowing
K ) from the apparent first-order rate constants when low
max
concentrations were present (specifically, K = K /K1).
S IftcLX _L
The equation used to calculate the apparent first-order rate,
KI, depends on the type of experimental system that was
employed. For continuous, well-mixed systems, KI was
calculated as follows:
K.. = FR(C. - CT)/[(RT)b.CT] . (5-15)
1 D 1 J_i 1 J_i
For batch systems and for continuous, plug-flow systems, the
equation used to calculate K., was:
K1 = F0 ln(C /CT ) / [ (RT)b. ] . (5-16)
J_ JD O J_i 1
Using this approach, rate constants for specific
compounds in the biodegradation rate data base were
determined. These rate constants are provided in Appendix C,
Table C-2. Upon evaluating the biodegradation rate data from
several different laboratory and field studies, it is
recognized that biodegradation rates can vary widely from
site to site. Therefore, the following priority schedule is
provided as guidance in determining the appropriate
biodegradation rate constants to be employed in the emission
models:
• Use site-specific biodegradation rate data in
experiments controlled for air emissions where
available.
• Use the rate constants suggested in Appendix C,
Table C-2, as available.
• Estimate the biodegradation rate constants using
the following methodology:
5-27
-------
Approximate Kmax from available data for Kmax for
compounds of similar structure and/or
functional groups; and
Approximate KI either by using the
correlation:
K., = 3.75 x 10~8 K °'38 (5-17)
1 ow ^ '
where
Kow = octanol-water partitioning coefficient,or by
using the default (average) value for KIr which is:
K! = 1 L/h/g (2.78 x 10~7 m3/s/g) , and then calculate
Ks as: Ks = K^/iq.
The correlation provided in Equation (5-17) was developed
based upon the assumption that biodegradation was primarily an
intracellular phenomenon. As such, the first-order
biodegradation rate can be limited either by the rate of the
internal reaction or by the rate of diffusion of the chemical
through the cell membrane and into the cell. If the internal
component concentrations are assumed to be proportional to the
concentration of components absorbed onto the cell membranes,
then, regardless of what limits the first-order biodegradation
rate, the limiting first-order biodegradation rate will be
directly proportional to the concentration of constituent
absorbed onto the membrane. Because the octanol-water parti-
tioning coefficient has been used to correlate the absorption
partitioning of organic chemicals onto biomass, ' it follows
that the octanol-water partitioning coefficient may also be used
to correlate the limiting first-order biorate constant since the
observed biorate is based on bulk liquid concentrations. To that
end, the limiting first-order rate constants for a variety of
compounds were plotted versus their corresponding octanol-water
partitioning coefficient. The results, presented in Figure 5.2,
indicate a fair correlation between the octanol-water
partitioning coefficients and the limiting first-order rate
5-28
-------
constants for most compounds. The primary discrepancies are for
ionizable or polar compounds.
The simple correlation with K should be used with caution.
ow
Figure 5-2 indicates a range of 25, with most of the data
scattered between a line five times the correlation and another
line one-fifth of the correlation. Some compounds may
biologically react slowly. For those compounds, the K
ow
correlation would significantly overpredict the biorate.
Activated sludge biorates are published in the literature
and can be a useful data source. Published biorates can be
useful if the biorate accounts for volatilization, if the waste
treatment system is the same as the system used for the published
biorates, and if the waste and operational parameters are similar
to the system as used for the published biorates. The biorate is
expected to be a strong function of several system variables.
The recommended priority schedule for the selection of biorates
reflects procedures that are based on an average biorate for many
different systems. It is possible that the literature biorate
may not accurately reflect the performance of specific systems,
and the error could possibly be greater than some of the simple
correlations presented in the priority schedule.
Assuming continuous, steady-state operation for a system
that is well-mixed, a mass balance on the system can be written
as follows:
QC. = QCT + VK b.CT/(K +CT ) + K ,, V CT (5-18)
i L max i L s L other L
where,
3
Q = flow rate, m /s;
Kother = sum of apparent first-order rate constants for
competing mechanisms, 1/s;
and the other symbols are as previously defined.
Note that Equation (5-18) was written in a general fashion
so that, if desired, the rate of removal via adsorption onto
5-29
-------
M
Figure 5.2. Correlation of Octanol-water Partition Coefficient
and the first order biorate constant.
5-30
-------
biomass solids can be included. For most volatile organics,
however, the adsorption pathway is negligible so that K ,, is
dominated by the volatilization rate. Consequently,
K , , = KA/V (5-19)
other v '
where,
K = overall mass transfer coefficient, m/s; and
A = area, m .
To determine the fraction of the organic compound emitted or
biodegraded using the Monod model, one first has to solve for the
effluent concentration. The effluent concentration can be
determined by rearranging Equation (5-18) as follows:
K'CT2 + [K K' + (V/Q)K b. - C. ] CT -KG. =0 (5-20)
L s max i i L s i
where
K' = (K , , ) (V/Q) + 1, dimensionless.
other
Equation (5-20) is easily solved using the quadratic formula
as follows:
CT = [-b + (b2 - 4ac)°-5]/2a (5-21)
j_i
where,
a = K' = (K ,, )(V/Q) + 1;
other
b = K K' + (V/Q)K b. - C.; and
s max i i
c = -K C..
s i
The plus sign is selected in Equation (5-21) to ensure
positive effluent concentrations. Note that, because all of the
rate constants and concentrations must have positive values, the
constant, c, must be negative so that the quadratic equation
always has real, positive roots.
5-31
-------
Once the effluent concentration is calculated, the fraction
of the component feed emitted to the air (f . ) is:
^ air
f . = K A CT/Q C. . (5-22)
air L i
Emissions (E, g/s) are calculated from:
E = f . QC. . (5-23)
air i
Similarly, the fraction of the component feed biodegraded (f, . )
is:
fbio = VKmaxbi CL/[(Ks + CL)QCi] ' (5~2^
If the biological system is operated with plug flow, the
treated wastewater does not mix with the influent. The
biodegradation rate and air emission rate change as the treatment
progresses toward completion. For plug flow, the rate of
disappearance of a compound by biodegradation and air emissions
is given by:
-d C, (V) VK b.C,
t _ max i t + ^ c (5-25)
dt (Ks + Ct)
where,
C, = concentration at time = t; and
t = time, s.
and with the other symbols as previously defined. Due to the
nonlinear nature of the biodegradation rate term, Equation (5-25)
cannot be directly integrated. Therefore, it is further assumed
that first-order kinetics dominates the system's biodegradation.
Equation (5-25) can then be rearranged as follows:
d C
= (-Kb. - KA/V) dt (5-26)
ct
5-32
-------
3
where K., = K /K , m /q biomass.
1 max s
Integrating Equation (5-26) from C , = C . at t = 0 to C, = CT
HI H J-i
(effluent concentration) at t = V/Q (one residence time) gives:
Ce/CQ = exp (-K1biV/Q - KA/Q) . (5-27
The ratio CT/C. represents the fraction leaving with the
j_i i
effluent; consequently, 1 -~CT/C. represents the sum of the
j_i i
fractions that are biodegraded and emitted to the air. The
fractions of component feed emitted to the air and biodegraded
are calculated from their relative rates:
f . = (1 - C /C.) (KA)/(KA + K b.V) (5-28
d -L -L J_i _l_ _L _l_
fbio = d - CL/Ci) (KlbjV)/(KA + K1biV) . (5-29)
The average emissions rate (E, g/s) is:
E = f . QC. . (5-30)
air ^ i ^ '
5.3.3 Example Calculation for Quiescent Impoundments
The application of the biodegradation model to quiescent
impoundments is presented in the form of an example calculation.
The calculation is based on the quiescent impoundment's operating
parameters from Table 5-2. For other types of impoundments, the
application of the biodegradation model is illustrated in
subsequent sections.
The waste stream for the example calculation is defined as
containing benzene at 10 ppm with a total organic content of
250 ppm (0.25 g/L) . The resultant organic loading on the
3
impoundment on a daily basis is 12.8 kg/1,000 m . The active
biomass is assumed to be 0.05 g/L from a reported range from
eight quiescent impoundments of 0.0014 to 0.22 g/L.
a. Calculate the effluent concentration of benzene for a
well-mixed system from Equation (5-21) :
CT = [-b + (b2 - 4ac) 0>5] /2a
5-33
-------
where
a = K' = (KA/V) (V/Q) + 1 = KA/Q + 1
b = K K' + (V/Q) K b. - C.
s max i i
c = -K C.
-6
K = 4.2 x 10 m/s (Section 5.2.3, Step c)
A = 1, 500 m2
Q = 0.00156 m3/s
K =19 mg/g/L = 5.28 x 10 g/g/s (from Appendix C,
ITlcLX
Table C-2)
3
K = 13.6 mg/L = 13.6 g/m (from Appendix C,
O
Table C-2) , ks = k^/^
b. = 0.05 g/L = 50 g/m3
V = 2,700 m3
C = 100 ppm = 100 g/m3
KA = (4.2 x 10~6 m/s) (1,500 m2 ) = 6.3 x 10~3 m3/s
a = K' = (6.3 x 10~3 m3/s) / (0.00156 m3/s) + 1 = 5.0
b = (13.6 g/m3) (5.0) + (2,700 m3/0. 00156 m3/s)
(5.28 x 10~6 g/g/s) (50 g/m3) - (100 g/m3)
b = 425 g/m3
c = -(13.6 g/m3) (100 g/m3) = -1,360 g2/m6
CL = {-[425 g/m3] + [(425 g/m3)2 - 4(5.0) (-1,360
g2/m6)]0'5}/ [2(5.0)]
CT = (-425 g/m3 + 455.9 g/m3)/10
3
CT = 3.09 g/m
b. Calculate the fraction emitted for a well-mixed system
from Equation (5-22) :
f . = KACT/(QC ;
air L o
where,
f . = (6.3 x 10 3 m3/s) (3.08 g/m3)/
cL 1 IT
[(0.00156 m/s) (100 g/m ) ]
f . = 0.124 .
air
c. Calculate benzene emissions for well-mixed system:
5-34
-------
E(g/s) = f Q
= (0.124) (0.00156 m/s) (100 g/m )
= 1.93 x 10~2 g/s = 0.61 Mg/yr.
d. For a plug-flow system, calculate fraction removed with
the effluent from Equation (5-27) :
CT/C. = exp (- K, b. V/Q - KA/Q)
j_i i _L i
where,
KI
C, Table C-l) ;
= 1.4 L/g-h = 3.89 x 10 m /g-s (from Appendix
b. = 0.05 g/L = 50 g/m3;
V = 2,700 m3 ;
Q = 0.00156 m3/s
C. =10 ppm =10 g/m3;
-L _ £•
K = 4.2 x 10 m/s; and
A = 1, 500 m2.
KlbiV = (3'89 x 10~7 m3/s/g biomass) (50 g/m3) (2,700 m3)
= 5.25 x 10~2 m3/s
KA = (4.2 x 10~6 m/s) (1,500 m2 ) = 6.3 x 10~3 m3/s
-2 3 -3 3
,„ _ ^ -5.25 x 10 m /s 6.3 x 10 m /s *
^L/^i ~ exP * _ _ if
1.56 x 10~3 m3/s 1.56 x 10~3 m3/s
CT/C. = exp (-37.7) = 0.00 .
j_i i
e. Calculate fraction emitted from Equation (5-28) :
f . = (1 - CT/C.) (KA) / (KA + K, b. V)
air L i 1 i
f . = (1 - 0) (6.3 x 10~3 m3/s) / (6.3 x 10~3 m3/s +
cL 1 IT O Q
5.25 x 10 m/s)
f . = 0.107 .
air
f. Calculate benzene emissions for plug flow:
E(g/s) = f Q G
= (0.107) (0.00156 m3/s) (10 g/m3)
= 1.67 x 10~3 g/s = 0.053 Mg/yr.
5-35
-------
5.4 MECHANICALLY AERATED IMPOUNDMENTS AND ACTIVATED SLUDGE
UNITS
Some impoundments and tanks are mechanically agitated to
improve mixing or to transfer air to the liquid (e.g., treatment
tanks designed for biodegradation). The agitation creates a
turbulent liquid surface that enhances mass transfer to the air.
A significant difference from the approach for quiescent surfaces
discussed in Section 5.2 is the appropriate correlations for the
individual mass transfer coefficients.
5.4.1 Emission Model Equations
The calculation of the overall mass transfer coefficient for
mechanically aerated systems considers that the liquid surface is
composed of two zones, quiescent and turbulent. The individual
mass transfer coefficients for the turbulent zone are based on
40 41
the correlations of Thibodeaux and Reinhardt. Thibodeaux's
model was developed from accepted interphase mass transfer
concepts, published rate coefficient correlations, and existing
operating data on 13 aerated basins at 11 pulp and paper mills.
The basins represented a wide range of design and operating
parameters, in spite of being from only one industry type. The
simulation employed 11 organic chemical species common to
industrial wastewater.
Reinhardt absorbed ammonia in aqueous sulfuric acid to
measure the gas-phase mass transfer coefficient associated with
flat-blade surface agitators in developing his correlation to
42
calculate the gas-phase mass transfer coefficient.
Table 5-6 summarizes the correlations developed by
Thibodeaux and Reinhardt. These correlations are used to
estimate the individual mass transfer coefficients for the
turbulent portion of the liquid surface. The individual
coefficients are then used in Equation (5-2) to calculate an
overall mass transfer coefficient for the turbulent zone. An
overall mass transfer coefficient for the quiescent zone is
5-36
-------
TABLE 5-6. EQUATIONS FOR CALCULATING INDIVIDUAL MASS TRANSFER
COEFFICIENTS FOR VOLATILIZATION OF ORGANIC SOLUTES FROM
TURBULENT SURFACE IMPOUNDMENTS
Liquid phase
Thibodeaux: '
k=[8.22 10"9J(POWR)(1.024)t"2° 0. 106 MWT /(Va •)] (D /D _ )
L t L v L w (j~, w
where,
k = mass transfer coefficient based on liquid, (m/s);
j_i
J = oxygen transfer rating of surface aerator, Ib
02/h»hp;
POWR = total power to aerators, hp;
T = water temperature, °C;
Ot = oxygen transfer correction factor;
MW = molecular weight of liquid;
j_i
V = volume affected by aeration, ft ;
a = surface-to-volume ratio of surface impoundment,
• = density of liquid, g/cm ;
L
2
D = diffusivity of constituent in water, cm /s; and
w
-5 2
D = diffusivity of oxygen in water = 2.4 x 10 , cm/s.
0 , w
Gas phase 45 46
Reinhardt: '
k,, = 1.35 x 10~7 R1'42 p°'4 Sc°'5 F~°'21 D MW /d (m/s)
G e G r a a
where, „
R = d w» /u = Reynold's number;
e a a
d = impeller diameter, cm;
w = rotational speed of impeller, rad/s;
(continued)
5-37
-------
TABLE 5-6 (continued)
3
• = density of air, g/cm ;
a
u = viscosity of air, g/cm»s; _.
a = 4.568 x 10 T(°C) + 1.7209 x 10 ;
* 5 3
p = P g / (• d w ) = power number;
_L C J-i
P = power to impeller, ft»lbf/s
=0.85 (POWR) (550 ft»lbf/s»hp)/number of aerators,
where 0.85 = efficiency of aerator motor;
2
g = gravitation constant, 32.17 Ib »ft/s /lbf;
• = density of liquid, Ib/ft ;
j_i
d* = impeller diameter, ft;
Sc« = Schmidt number on gas side = u /• D ;
(j a a a
2
F = d*w /g = Froude number ;
r yc
2
D = diffusivity of constituent in air, cm /s; and
a
MW = molecular weight of air.
a
5-31
-------
calculated as described in Section 5.2. The two overall
coefficients are combined to obtain a single coefficient for
the system based on the relative areas of the turbulent and
quiescent zones. For example, if 25 percent of the surface of
the impoundment is turbulent, the overall coefficient would be
the sum of 25 percent of the value for the turbulent area
coefficient plus 75 percent of the value for the quiescent
zone.
The model for mechanically aerated systems also incorporates
biodegradation as a competing mechanism. The extent of
biodegradation is difficult to predict in a generally applicable
form because it is very dependent upon the constituent of
interest, the waste matrix, the design and operation of the
biodegradation unit, and the concentrations and properties of the
microorganisms.
5.4.2 Model Plant Parameters for Mechanically Aerated
Impoundments
The dimensions of the treatment impoundment used as an
example to estimate emissions were derived from the Westat data as
described in Section 5.2.2 for storage impoundment. A median area
2
of 1,500 m and a depth of 1.8 m were chosen, which yields a total
3
volume of 2,700 m . The retention time in treatment impoundments
is expected to be less than the retention time in storage
impoundments. Two design manuals listed typical retention times
47
for aerated (biologically active) ponds as 7 to 20 days and 3 to
48
10 days. For the example case, a retention time of 10 days was
chosen from the design range of 3 to 20 days. The resulting flow
rate is 3.1 L/s (0.0031 m3/s).
The correlations of Thibodeaux and Reinhardt given in
Table 5-6 require values for the parameters that describe the
49
mechanical aeration system. Metcalf and Eddy, Inc., suggest a
3
range of 0.5 to 1.0 hp/1,000 ft for mixing in an impoundment.
However, more power may be needed to supply additional oxygen or
to mix certain treatment solutions. A review of trip reports
showed power usage as high as 3.5 hp/1,000 ft at a specific TSDF
5-39
-------
impoundment.50 For this analysis, a midrange value of
0.75 hp/1,000 ft from Metcalf and Eddy was used to generate an
estimate of 75 hp required for mixing in the model unit.
Data from Reference 51 indicated that five aerators with 15-
hp motors and 61-cm diameter propellers turning at 126 rad/s would
agitate a volume of 441 m (15,590 ft ). Assuming a uniform depth
in the impoundment of 1.8 m, the agitated surface area was
2
estimated as 245 m (441/1.8). The agitated surface is assumed to
be turbulent and comprises 16 percent (245/1,500) (100) of the
total area. The balance of the surface area of the impoundment
(84 percent) is assumed to be quiescent. As a comparison,
2
Thibodeaux reported a turbulent area of 5.22 m /hp and
2
investigated a range of 0.11 to 20.2 m /hp. The value of 5.22
2
m /hp and a total of 75 hp yields an estimated turbulent area of
2
392 m (26 percent), which is greater than the 16-percent
52
turbulent area calculated by the above procedure. (Very few
data are available on the distribution of turbulent areas for
aerated impoundments. The extent of turbulence depends in part on
the number, size, and placement of aerators. The example is based
on typical aerator requirements to mix the contents of the
impoundment.)
Typical values were chosen for the oxygen transfer rating of
the aerator and the oxygen transfer correction factor. A value of
3.0 Ib Cu/hp/h was chosen for oxygen transfer rating from a range
53
of 2.9 to 3.0. A value of 0.83 was used for the correction
54
factor from a typical range of 0.80 to 0.85. The transfer of
power to the impeller was assumed to be 85 percent efficient,
yielding an estimate of 64 hp for the impeller power.
The model for biodegradation requires the system's biomass
concentration as an input parameter. The concentration of biomass
in real systems can be highly variable depending upon the system's
design and method of operation. For this analysis, the specified
biomass is assumed to be actively degrading the constituent of
5-40
-------
3
interest. A value of 250 g/m (0.25 g/L) of biomass was chosen
from the values presented in Table 5-5.
The example constituent (benzene) and the meteorological
conditions chosen for the example calculation are the same as
those chosen for storage impoundments. Input parameters for
the mechanically aerated model unit are summarized in
Table 5-7.
5.4.3 Example Calculation for Mechanically Aerated Treatment
Impoundments
The example calculation for emissions from a mechanically
aerated impoundment includes an estimate of the overall mass
transfer coefficient for the turbulent zone. The overall mass
transfer coefficient for the quiescent zone for storage
impoundments is calculated as illustrated in Section 5.2.3 and
will not be repeated here. Biodegradation is included as a
competing removal mechanism.
a. Calculate turbulent liquid-phase mass transfer coefficient,
kT . Use Thibodeaux (Table 5-6) :
D
-9 T-20 6 * w *
kT (m/s) = [8.22 10 J(POWR) (1.024) 0.10 MWT / (Va • ) ] *=— ^ - *
L t L v L *DQ^w^
where,
J = Op transfer rating, use 3.0 Ib Cu/h»hp
POWR = 75 hp
T = water temperature = 25 °C
0, = Op transfer correction factor, use 0.83
MW = molecular wt of liquid (water) = 18 g/g mol
j_i
(Va ) = agitated area in ft2 = 240.0 m 2 *0.0929 m2*
= 2,583 ft2
3
• T = water density = 1 g/cm
j_i
D = 9.8 x 10~6 cm2/s
w
D_ = 2.4 x 10~5 cm2/s
Op, w
5-41
-------
TABLE 5-7. INPUT PARAMETERS — TREATMENT IMPOUNDMENTS
(MECHANICALLY AERATED)
2
Area: 1,500 m
Depth: 1.8m
Volume: 2,700 m
3
Retention time: 10 days
3
Flow: 0.0031 m /s
Turbulent area: 240 m
Quiescent area: 1,260 m
2
Number of impellers: 5
Total power: 75 hp
Power to impeller: 13 hp
Impeller speed: 126 rad/s
Impeller diameter: 61 cm
)
02 transfer: 3 Ib/h/hp
Op correction factor: 0.83
Temperature: 25 °C
Windspeed: 4.47 m/s
Viscosity of air: 1.8 x 10
-3
-4
g/cm»s
3
-
Density of air: 1.2 x 10 g/cm
-5 2
Diffusivity of Cu in water: 2.4 x 10 cm /s
3
Density of liquid: 1 g/cm
Molecular weight of liquid: 18 g/g»mol
Molecular weight of air: 29 g/g»mol
Constituent: benzene with other biodegradable organics in water
3
Concentration (benzene): 100 g/m (100 ppm)
3
Concentration (total organics): 250 g/m (250 ppm)
-3 3
Henry's law constant (benzene): 5.5 x 10 atm»m /g mol
(continued)
5-42
-------
TABLE 5-7 (continued)
2
Diffusivity in air (benzene): 0.088 cm /s
/~ O
Diffusivity in water (benzene): 9.8 x 10 cm /s
Maximum biorate (benzene and other organics):
19 mg/h/g of biomass = 5.28 x 10 g/g biomass»s
Limiting first-order biorate constant:
1.4 L/h/g = 3.89 x 10 m /s/g biomass
3
Biomass concentration: 0.3 g/L = 300 g/m
5-43
-------
6 6 °'5
k = (8.22 x 10~9) (3) (75) (1.024)5|(0'83) (1° ) (18)* *9'8 x 10_ *
(2,583) (1) - .2.4 x 10 -
=7.7x10 m/s
b. Calculate turbulent gas-phase mass transfer coefficient, k .
Use Reinhardt (see Table 5-6):
k,,(m/s) = 1.35 x 10~7 Re1'42 p°'4 Sc°'5 Fr~°'21 D MW /d
(j baa
where
_ _ , , . , d w a
Re = Reynold s number = -
d = impeller diameter = 61 cm
w = impeller speed = 126 rad/s
• = 1.2 x 10~3 g/cm3
cL
yi =1.81x10 g/cm»s
3.
Re = (6^ (126) (1'2 X 10"3) = 3.1 x 106
1.81 x 10
,
p = power number = - J-F — r
• d w
550 ft lbf
PT = 12.8 hp - T- - - = 7,040
I ^ s»hp '
g = 32.17 j;b'ft
s2 lbf
• = 62.37 lb/ft3
d* = impeller diameter in feet = 2.0
w = 126 rad/s
(7,040) (32.17) , r in-5
p = -—'- g- -^= 5.6 x 10
(62.37) (2^) (126J)
3cr = 1.71 (from Section 5.2.3, part b)
5-44
-------
* 2 2
„ _ , , dw (2) (12 6) „ „ -,^2
Fr = Froude number = = v 'v i n = 9-9 x 1°
gc 32.17
D = 0.088 cm2/s (benzene)
3.
MW = 29 g/g mol
3.
d = impeller diameter in cm = 61 cm
(1.35xl07)(3.1xl06)1>42(5.6xl05)°'4(1.7lP'5(9.9xl§ y°-21
(0.088) (29)/61
= 5.7 x 10~2 m/s .
c. Calculate overall mass transfer coefficient for turbulent
area, K:
L_ = -J- + I = I + I = 201
K kL Keq kQ ?>? x 1Q 3 (Q.225)(5.7 x 10 )
K = 0.0048 m/s .
d. Calculate overall mass transfer coefficient for combined
quiescent and turbulent areas, K:
From Section 5.2.3, K for quiescent area = 4.2 x 10 m/s;
_3
From Part C, K for turbulent area = 4.8 x 10 m/s;
2
Turbulent area = 240 m ; and
Quiescent area = 1,260 m .
„ , , . (4.2 x 10~6) (1,260) +(0.0048) (240) _ _ ._-4 ,
K ( Tf\ I Q I — —- - - - - —— / / V I 1 ID / Q
K (m/s) - (1,260 + 240) ~ /'/ X 1U m/S'
(weighted by area)
e. Calculate the effluent concentration for benzene for a well-
mixed system from Equation (5-21):
CT = [-b + (b2 - 4ac)0>5] /2a
5-45
-------
where
a = K' =
b
c =
K
A
Q
K
max
b.
V
C
o
K
s
K
s
K
s
KA
(KA/V) (V/Q) + 1 = KA/Q +
K K' + (V/Q) K b. - C
s max i o
-K C
so .
7.7 x 10 m/s
1,500 m2
0.0031 m3/s
5.28 x 10 g/s/g biomass
0.3 g/L = 300 g/m3
2,700 m3
100 ppm = 100 g/m3
K /K1
max 1 „
(5.28 x 10 g/s/g)/(3.89 x 10 m
13.6 g/m
(7.74 x 10~4 m/s) (1,500 m2 ) = 1.16
1
3/s/g)
m/s
a = K' =(1.16 m3/s)/(0.0031 m3/s) + 1 = 375
b = (13.6 g/m3)(373) + (2,700 m3/0.0031 m3/s)
(5.28 x 10~6 g/s/g) (300 g/m3) - (100 g/m3)
3
b = 6,352 g/m
c = -(13.6 g/m3)(100 g/m3) = -1,360 g2/m6
CL = {-[6,352 g/m3] + [(6,352 g/m3)2
- 4(373) (-136 g2/m6)]0'5}/ [2(373)]
CT = (-6,352 g/m3 + 6,509 g/m3)/746
3
CT = 0.021 g/m .
j_i
f. Calculate the fraction emitted for a well-mixed system from
Equation (5-22):
f . = KACT/(QC ;
air L o
where
f . = (1.15 m3/s)(0.21 g/m3)/[(0.0031 m3/s)(100 g/m3)]
cL 1 IT
f . = 0.78 .
air
g. Calculate benzene emissions for well-mixed system:
E(g/s) = f Q CQ
5-46
-------
= (0.79) (0.0031 m3/s) (100 g/m3)
= 0.24 g/s = 7.7 Mg/yr
h. For a plug-flow system, calculate the fraction removed with
the effluent from Equation (5-27) :
CT/C. = exp (-K, b. V/Q - KA/Q)
-73
KI = 3.89 x 10 m /s/g biomass
b. = 0.3 g/L = 300 g/m3
V = 2,700 m3
Q = 0.0031 m3/s
C. =10 ppm =10 g/m3
K = 1.0 x 10~3 m/s
A = 1,500 m2
KlbiV = (3'89 x 10~7 m3/s/g biomass) (300 g/m3) (2,700 m3)
= 0.315 m3/s
KA = (7.7 x 10~4 m/s) (1,500 m2 ) = 1.15 m3/s
C /C = exp
L i 0.0031 m/s 0.0031 m/s
Calculate fraction emitted from Equation (5-28) :
f . = (1 - CT/C.) (KA)/(KA + ICb.V)
d _L -L J_i _L _L _l_
f . = (1 - 0) (1.15 m3/s)/(1.15 m3/s + 0.315 m3/s)
3. 1 IT
f . = 0.78 .
air
Calculate benzene emissions for plug flow:
E(g/s) = f , Q C,
air i _ _
= (0.78) (0.0031 m/s) (100 g/m )
= 0.24 g/s = 7.7 Mg/yr
5-47
-------
5.4.4 Example Calculation for Activated Sludge Unit
As discussed in Section 5.2, an activated sludge unit usually
consists of a concrete tank that is aerated and contains a
relatively high concentration of active biomass. A model unit is
defined in this section for this process, and the results of
intermediate and final calculations are given. Detailed example
calculations are not presented because the approach is exactly the
same as that used for the mechanically aerated impoundment. The
only significant difference in the method of operation is the
recycle of solids back to the activated sludge unit, which results
in a higher biomass concentration. For this model unit, a biomass
concentration of 4 g/L (4,000 g/m ) was chosen from the range of
1.5 to 6 g/L in Table 5-3 and the recommended values in Table 5-5.
Other differences between the aerated impoundment and activated
sludge tank include, for the tank, a smaller surface area, a
shorter retention time, a greater turbulent area, and a smaller
F/D ratio. 70 percent of the unit surface is assumed to be
turbulent. The aerated surface area was estimated as described in
Section 5.4.2. An aerator with a 7.5-hp motor will agitate a
m,
2
volume of 56.9 m (2,010 ft ). For a uniform depth of 4 m, the
agitated volume yields an agitated surface area of 14.2 m
(56.9 m /4 m). The input parameters are defined for this model
unit in Table 5-8, and the results of the calculations are
presented in Table 5-9.
5.5 DISPOSAL IMPOUNDMENTS WITH QUIESCENT SURFACES
5.5.1 Emission Model Equations
A disposal impoundment is defined as a unit that receives a
waste for ultimate disposal rather than for storage or treatment.
This type of impoundment differs from the storage and treatment
impoundments in that there is no liquid flow out of the
impoundment (seepage into the ground is neglected). For this
case, the well-mixed system with a bulk concentration that is at
equilibrium (i.e., the bulk concentration does not change
5-4!
-------
TABLE 5-8. INPUT PARAMETERS—MECHANICALLY AERATED
ACTIVATED SLUDGE UNIT
Area: 27 m
Depth: 4m
Volume: 108 m
Retention time:
Flow: 0.0075 m^/s
Turbulent area:
Quiescent area:
3
4 h
2
19 m (70^
8.0m
Total power: 7.5 hp
Power to impeller: 6.4 hp
Impeller speed: 126 rad/s
Impeller diameter: 61 cm
02 transfer: 3 Ib/h/hp
Op correction factor: 0.83
Temperature: 25 °C
Windspeed: 4.47 m/s
-4
Viscosity of air: 1.8 x 10_^ g/cm»s
Viscosity of water: 9 x 10 g/cm»s
Density of air: 1.2 x 10 g/cm
Diffusivity of 09 in water' 2.4 x 10
Density of liquid: 1 g/cm
Molecular weight of liquid: 18 g/g»mol
Molecular weight of air: 29 g/g»mol
Constituent: benzene with other^biodegradable organics in water
Concentration (benzene): 10 g/m (10 ppm)
Concentration (total organics): 250 g/m (250 ppm)
-5
cm /s
5.5 x 10
-3
0.088 cm /Sf.
9.8 x 10
Henry's law constant (benzene
Diffusivity in air (benzene) :
Diffusivity in water (benzene
Maximum biorate (benzene and other organics
biomass
Limiting first-order biorate constant = 3.8
Biomass concentration: 4.0 g/L = 4,000 g/m
3
atm»m /g»mol
2
cm /s
: 5.2
x 10 g/s/g
x 10
-7 3
m /s/g biomass
5-49
-------
TABLE 5-9. INTERMEDIATE AND FINAL CALCULATION RESULTS
FOR ACTIVATED SLUDGE MODEL UNIT
Quiescent zone:
kL = 6.5 x 10~6 m/s
k = 8.9 x 10~3 m/s
K = 6.5 x 10~6 m/s
Turbulent zone:
kL = 9.7 x 10~2 m/s
k = 4.3 x 10~2 m/s
g _
K = 4.88 x 10 m/s
_3
Overall mass transfer coefficient = 3.4 x 10 m/s
For well-mixed system:
CT = 3.17
J_i
f . = 0.391
air
Emissions = 0.30 g/s = 9.3 Mg/yr
For plug-flow system:
f . = 0.391
air
Emissions = 0.30 g/s = 9.3 Mg/yr
5-50
-------
with time) is not applicable. The quantity of a constituent in a
disposal impoundment will decrease with time after the waste is
placed in the impoundment because of the loss of volatiles to the
air.
The calculation of the overall mass transfer coefficient is
the same as that presented for impoundments with quiescent
surfaces. If the disposal impoundment is aerated, K is calculated
as described for aerated impoundments in Section 5.4. The
emission estimating procedure differs in the calculation of the
liquid-phase concentration that is the driving force for mass
transfer to the air. For a disposal impoundment that is filled
with a batch of waste, the rate of disappearance of a compound by
biodegradation and air emissions is described by Equation (5-26) .
Integrating Equation (5-26) from C, = C. at t = 0 to C, =
C, at t = t gives:
C /C = exp (-K b t - KAt/V) . (5-31)
t i 1 i
For an impoundment with a uniform depth, V/A = D. Substituting
V/A = D into Equation (5-31) yields:
Ct/Ci = exp (-K1bi t - Kt/D) (5-32)
When Equation (5-32) is evaluated after some fixed time t, the
ratio C,/C. represents the fraction of the compound remaining in
the impoundment; consequently, 1 - C./C. represents the fraction
that has been removed by biodegradation and air emissions. The
fractions emitted to the air and biodegraded after some time (t)
are calculated from their relative rates:
f . = (1 - C./C.) (KA)/(KA + ICb.V) (5-33)
d -L -L L -L -L -L
fbio = d - Ct/Ci)(K1biV)/(KA + K1biV) (5-34)
The quantity emitted after some time (t) is given by:
Emitted quantity (g) = f . V C.. (5-35)
air i
The average emission rate over the period of time = t is:
5-51
-------
E (g/s) = f . V C./t . (5-36)
a -L -L _L
Alternatively, a simplifying assumption may be made that,
because the impoundment is designed for disposal, all
significantly volatile compounds are eventually emitted to the
air. Emissions under this assumption would simply be QC. where Q
equals the disposal rate in cubic meters/second. This assumption
is probably valid for volatile compounds; however, compounds that
are relatively nonvolatile may be removed slowly and the
assumption may result in an overestimate of emissions.
5.5.2 Model Plant Parameters for Disposal Impoundments
The Westat data summary for impoundments indicated that
disposal impoundments generally have higher surface areas and
shallower depths than storage and treatment impoundments. The
median surface area for disposal impoundments was approximately
2 2
9,000 m (compared to 1,500 m for storage impoundments), and the
median depth was approximately 1.8 m. The disposal impoundment is
assumed to be filled with waste every 6 months (two turnovers per
year) .
The meteorological conditions and type of waste (water
containing benzene and other organics for the example calculation
are the same as those used for quiescent and aerated impoundments
with biodegradation. The inputs for the example calculation of
emissions from disposal impoundments are summarized in Table 5-10.
5-52
-------
5.5.3 Example Calculations for Disposal Impoundments
Example calculations are presented below for the model unit
defined to represent disposal impoundments.
a. Calculate liquid-phase mass transfer coefficient, k . Use
Springer's model (see Table 5-1):
0.5 0.5
*ft-ra=i# *Q 000*
Effective diameter = |^^^| x 2 = * '. * x 2 = 107 m
F/D = Effective diameter/depth = -=—„• = 59.5
1 . o
Windspeed = 4.47 m/s (UIQ > 3.25 m/s)
* D * 0-67
kT = 2.611 x 10"7 tLn |^—-—| m/s
L 10 *Dether*
5-53
-------
TABLE 5-10. INPUT PARAMETERS—DISPOSAL IMPOUNDMENTS
Area: 9,000 m2
Depth: 1.8m
Volume: 16,200 m3
Turnovers per year: 2
Temperature: 25 °C
Windspeed: 4.47 m/s
/~ O
Diffusivity in water (ether): 8.5 x 10 cm /s
Viscosity of air: 1.81 x 10 g/cm»s
-3 3
Density of air: 1.2 x 10 g/cm
Constituent: benzene with other biodegradable organics in water
Concentration (benzene): 100 g/m (100 ppm)
3
Concentration (total organics): 250 g/m (250 ppm)
Henry's law constant (benzene): 5.5 x 10 atm»m /g mol
2
Diffusivity in air (benzene): 0.088 cm /s
f~ O
Diffusivity in water (benzene): 9.8 x 10 cm /s
/~ Q
Limiting first-order biorate constant: 3.89 x 10 m /s/g biomass
Biomass concentration: 0.05 g/L = 50 g/m
5-54
-------
TABLE 5-11. INPUT PARAMETERS—DIFFUSED AIR ACTIVATED SLUDGE UNIT
Area: 27 m
Depth: 4m
Volume: 108 m
Retention time:-, 4 h
Flow: 0.0075 m /s 2
Quiescent area: 8.0m _
Diffused air rate: 0.04 m /s
Temperature: 25 °C
Windspeed: 4.47 m/s
-4
Viscosity of air: 1.81 x_10 g/cm»s
Density of air: 1.2 x 10 g/cm _,- „
Diffusivity of Cu in water' 2.4 x 10 cm /s
Density of liquid: 1 g/cm
Molecular weight of liquid: 18 g/g»mol
Molecular weight of air: 29 g/g»mol
Constituent: benzene with other biodegradable organics in water
Concentration (benzene): 100 g/m (100 ppm)
Concentration (total organics): 250 g/m (250 ppm)
-3 3
Henry's law constant (benzene): 5.5 x 10 atm»m /g»mol
Diffusivity in air (benzene) : 0.088 cm /s,- 9
Diffusivity in water (benzene) : 9.8 x 10 cm /s _,.
Maximum biorate (benzene and other organics): 5.28 x 10 g/s/g
biomass _7 .-,
Limiting first-order biorate constant: 3.89 x 10 m /s/g biomass
Biomass concentration: 4.0 g/L = 4,000 g/m
5-55
-------
where,
U1 n = windspeed = 4.47 m/s;
-L U _ x~ Q
D = 9.8 x 10 cm /s (benzene) ; and
w _ , „
D , , = 8.5 x 10 = cm /s (ether) .
ether
Then,
•—} i~\ OQ "1 O tf^
k = 2.611 x 10 (4.47 *
* -6*
*8.5 x 10 *
k = 5.7 x 10 m/s
b. Calculate gas-phase mass transfer coefficient, k . Use
MacKay and Matasugu (see Table 5-1) : "
i A oo i n~3 TT 0.780 -0.67 , -0.11, , .
k_ = 4.82 x 10 U Sc« d (m/s)
b be
where,
U = windspeed = 4.47 m/s
^ _ Schmidt No. _ _ viscosity of gas _
G for gas (gas density) (diffusivity of i in gas]
Gas = air
-4
Viscosity (air) = 1.81 x 10 g/cm»s
Density (air) = 1.2 x 10 g/cm
2
Diffusivity (benzene in air) = 0.088 cm /s
_ _ 1.81 x 10 g/cm*s
Then,
G (1.2 x 10 3 g/cm3) (0.088 cm2/s)
d = effective diameter = 107 m
e
k,, = (4.82 x 10~3) (4.47)0'78 (1.71)~°'67 (107)"0'11
G
=6.5x10 m/s
c. Calculate overall mass transfer coefficient, K:
K k~ + Keq kr
5-56
-------
where,
-3 3
H 5.5 x 10 m *atm/mol
TD m ~
(8.21 x 10-5
Then
F = - - - rz + - - - TQ- = i-76 x lo5
5.7 x 10 (0.225) (6.5 x 10 )
K = 5.7 x 10~6 m/s
d. Calculate the fraction remaining from Equation (5-32) . The
impoundment is filled with waste initially, and 6 month later
it will be filled again. Calculate the fraction remaining
after the initial 6-month period:
Ct/Ci = exp (-K1bi t - Kt/D) ;
KI = 3.89 x 10 m /s/g biomass;
b. = 50 g/m3;
t = 6 mo = 1.58 x 107 s;
C. = 100 g/m3;
K = 5.7 x 10~6 m/s;
D = 1 . 8 m;
K1bit = (3.89 x 10~7 m3/s/g biomass) (50 g/m3) (1.58 x 107 s);
= 307;
Kt/D = (5.7 x 10~6 m/s) (1.58 x 107 s) / 1.8 m = 50.0; and
Ct/Ci = exp (~307 - 50) = 0 .
5-57
-------
f. Calculate the fraction emitted from Equation (5-33) :
f . = (1 - C./C. ) (KA) / (KA + ICb.V)
d -L -L L _L -L _L
ct/ci = °
KA = (5.7 x 10~6 m/s)(9,000 m2) = 0.051/m3/s
Since the concentration is high enough for zero-order
kinetics, K.,b.V is replaced with K
1 i max
f . = (1 - 0) (0.051 m3/s) / (0.051 m3/s + 0.315 m3/s)
3.1 IT
f . = 0.14 .
air
g. Calculate the average emission rate over the 6-mo period from
Equation (5-36):
E (g/s) = f . V C./t
^ air i
= (0.14) (16,200 m3) (100 g/m3)/1.58 x 107 s
= 1.4 x 10~2 g/s.
5.6 DIFFUSED AIR SYSTEMS
5.6.1 Emission Model Equations
Some impoundments and open tanks (e.g., activated sludge
units) are sparged with air to promote biodegradation or air
stripping. To estimate emissions from diffused air systems, the
model assumes that the air bubbling through the liquid phase
reaches equilibrium with the liquid-phase concentration of the
constituent. The emissions leaving with the diffused air are
estimated by:
E = QaKeqCL (5-37)
where,
E = emissions, g/s;
Q = air flow rate, m /s;
a
Keq = equilibrium constant; and
CT = concentration in the liquid phase, g/m .
5-51
-------
Emissions can also occur from wind blowing across the
surface. If the air sparging creates a very turbulent surface
similar to the surface of mechanically aerated systems, then the
emission rate should be based on values of K typical for
mechanically aerated systems. If the air sparging rate does not
result in a turbulent surface, then K can be estimated from the
correlations given for quiescent surfaces in Section 5.2.
The approach to estimate total emissions for flowthrough
tanks and impoundments sparged with diffused air is similar to
that described for quiescent and aerated systems. Because the
unit is sparged with air, the liquid phase is assumed to be well
mixed and the plug-flow model is not used. A material balance
around this well-mixed system is identical to Equation (5-18) in
Section 5.3.2, but now K ,, is:
' other
K , , = (KA + Q Keq)/V (5-38J
other a M
where all of the symbols have been previously defined. The
steady-state liquid phase concentration (CT) is then calculated
j_i
using Equation (5-21). Air emissions are estimated as the sum
from wind blowing across the surface and from the diffused air:
E = KACT + Q Keq CT . (5-39;
is:
The fraction of the component feed emitted to the air (f . )
air
f . = (KCTA + Q KeqCT)/QC. . (5-40]
a -L L j_i a j_i -L
For disposal impoundments with diffused air systems, the
steady-state assumptions of the flowthrough models do not apply.
Emissions are greatest when the waste is first placed in the
impoundment and gradually decrease with time. To incorporate the
contribution to mass transfer from diffused air, an equivalent
mass transfer coefficient is defined:
KD = KeqQa/A (5-41]
where
5-59
-------
K = equivalent mass transfer coefficient for diffused air,
m/s and all of the other symbols are as previously defined.
The mass transfer coefficient for wind blowing across the
surface (K) is calculated as described previously for flowthrough
systems. A combined overall mass transfer coefficient (K ) is
defined as:
KC = KD + K . (5-42)
The overall mass transfer coefficient (K ) is used in the
c
equations for disposal impoundments (Section 5.5.1) to estimate
the fraction emitted (Equation 5-33) and the average emission rate
(Equation 5-36). The combined overall mass transfer coefficient
defined above includes the mass transfer effects from both removal
mechanisms (wind and diffused air) .
5.6.2 Model Unit Parameters for Activated Sludge Unit with
Diffused Air
A model unit for the activated sludge process was defined in
Section 5.4.4 and Table 5-8. The same dimensions are used here to
define an activated sludge unit that uses diffused air instead of
mechanical aeration. The only additional parameter that must be
specified is the diffused air rate, which typically ranges from
0.3 to 0.5 m/s per 1,000 m of volume (20 to 30 ft /min per 1,000
355 3
ft of volume). For the model unit with a volume of 108 m , an
3
estimate of 0.04 m /s is recommended based on the mid-point of the
design range. The model unit input parameters are summarized in
Table 5-11.
5-60
-------
5.6.3 Example Calculation for Diffused Air Activated Sludge Unit
An example calculation is presented below for the model unit
defined in Table 5-11.
a. Calculate the liquid-phase, gas-phase, and overall mass
transfer coefficients. This procedure was illustrated
for quiescent surfaces and the results for this model
unit are given in Table 5-9:
k = 6.5 x 10 m/s;
-3
k =8.7x10 m/s; and
g _2
K = 3.42 x 10 m/s (weighted by area).
b. Calculate the equilibrium constant, Keq. The compound
is benzene in water, and Keq has been presented as 0.225
in the previous sample calculations (from Equation 5-5).
c. Calculate the equilibrium liquid concentration in the
unit (CT) from Equation (5-21):
j_i
Q = 0.0075 m3/s
C. = 100 g/m3
K = 3.42 x 10~3 m/s
A = 27 m2
Qa = 0.04 m3/s
Keq = 0.225
K = 5.28 x 10 g/s/g biomass
max _£ _7
K = K /K, = (5.28 x 10 g/s/g)/(3.89 x 10
s max 1 _ ^ ^
m /s/g)
= 13.6 g/m
b. = 4,000 g/m3
V = 108 m3
QC. =(0.0075 m3/s)(100 g/m3) = 0.75 g/s
KA =(3.42 x 10~3 m/s)(27 m2) = 9.23 x 10~2 m3/s
QaKeq =(0.04 m3/s) (0.225) = 9.0 x 10"3 m3/s
Kother =(KA + Q Keq)/V (from Equation 5-38)
= [(9.23xlO~2 m3/s)+(9.0xlO~3 m3/s)]/(108 m3)
= 9.39 x 10~4 1/s
V/Q =(108 m3)/(0.0075 m3/s) = 14,400 s
a = K' = (9.29 x 10~4 1/s) (14,400 s) + 1 = 14.5
5-61
-------
b = (13.6 g/m3) (14 . 5) + (14,400 s) (5.28 x 10 6 g/s/g)
(4,000 g/m3) - 100 g/m3
= 401 g/m3
c = -(13.6 g/m3)(100 g/m3) = -1,360 g2/m6
CT = {[-401 g/m3] + [(401 g/m3)2 - 4(14.5)
J_i
(-1,360 g2/m6) ]° ' 5 } / [2 (14.5) ]
= [(-401 g/m3) + (490 g/m3)]/29
3
= 3.06 g/m
d. Calculate air emissions from Equation (5-39).
E = (9.23 x 10~2 m3/s)(3.06 g/m3) + (9.0 x 10~3
m3/s)(3.06 g/m3)
= 0.31 g/s = 9.7 Mg/yr.
5.7 OIL FILM SURFACES
Some wastes discharged into impoundments may contain volatile
organics and oil. Many volatile organics will partition mostly
into the oil, so the oil phase can contain most of the volatiles.
The oil phase will rise to the surface of the impoundment where it
is exposed to the atmosphere.
Some impoundments may have a floating film of oil on the
surface. A rigorous approach to estimating emissions from this
type of source would consider three resistances acting in series:
• From the aqueous phase to the oil
• Through the oil
• From the oil to the air.
Such an approach would require estimates of these three
resistances and estimates of the equilibrium partitioning between
both the aqueous and oil phases and the oil and air phases.
Because these estimates are not generally available, a simplifying
5-62
-------
assumption is that the oil film is relatively thin, that the oil
originally contains the volatile constituents, and that mass
transfer is controlled by the gas-phase resistance. For this
case, Equation (5-2) reduces to:
K = k_ Keq (5-43)
(j
where ]<:„ is calculated from the correlation of MacKay and Matasugu
(j
(Table 5-1) and Keq is calculated from Raoult's law by:
*
Keq = P» MW ^7 (• MW P ) (5-44)
where
Keq = dimensionless equilibrium constant
*
P = vapor pressure of the volatile compound
of interest, atm
P = total pressure, 1 atm
° 3
• = density of air, g/cm
a 3
• = density of oil, g/cm
j_i
MW . , = molecular weight of oil, g/g mol
MW = molecular weight of air, 28.8 g/g mol.
a
The value of K calculated above is substituted into the equations
for flow-through systems to estimate emissions. For the well-
mixed flow models, C. and CT in Equations (5-1) and (5-6)
i J_i
represent the organic compound concentration in the oil phase
(entering and leaving the impoundment, respectively) , and the
flowrate Q is the volumetric flow rate of oil. Biodegradation is
neglected because the oil film inhibits the transfer of oxygen.
The procedure described above assumes that the oil layer in
the impoundment is well mixed. For example, changes in wind
direction in units with retention times on the order of days may
tend to move the oil layer in different directions and result in
mixing. However, some systems may be designed for or
characterized by plug flow. This flow model assumes that the oil
film moves across the impoundment's surface without backmixing.
For plug flow of the oil film in flowthrough impoundments and
tanks, the fraction of organic compound in the oil layer emitted
to the air is given by Equation (5-11), and air emissions are
5-63
-------
estimated from Equation (5-12). In these equations, CT is the
j_i
organic compound concentration in the oily effluent, C. is the
initial concentration in the oil layer entering the impoundment, •
is the residence time, D is the oil-film thickness, and Q is the
volumetric flowrate of oil.
For an oil film on a disposal impoundment, emissions are
calculated as described in Section 5.5. However, biodegradation
is neglected and Equation (5-32) reduces to:
CT/C. = exp (-Kt/D) (5-45)
j_i i
and the fraction emitted to the air is:
f . = 1 - exp (-Kt/D) (5-46)
air r \ ' i
where,
C, = concentration in the oil film at time = t;
CT = initial concentration in the oil film; and
j_i
D = oil-film thickness.
and with the other symbols as previously defined. The average
emission rate E, in units of g/s, over the period of time equal to
t is:
E = f V Ci/t (5-47
3
where V = volume of oil in the impoundment, m and with the other
symbols as previously defined. An example calculation of this
approach is given in Section 7.0 for applying an oil film to soil,
which is analogous to an oil film on a disposal impoundment
because there is no flow out in either case and emissions are a
function of the time since disposal.
5.8 DISCUSSION OF ASSUMPTIONS AND SENSITIVITY ANALYSIS
5.8.1 Removal Mechanisms
The organic constituents present in wastes that are treated,
stored, or disposed of in surface impoundments and open tanks may
5-64
-------
leave the unit by any of several mechanisms. Because of the large
open surface area and relatively high volatility of many organic
constituents, emissions to the air may be a primary removal
mechanism for certain constituents. Other constituents may be
destroyed in impoundments and tanks specifically designed for
biodegradation. Aeration is often used to supply oxygen to
biologically active systems. Unfortunately, aeration also greatly
enhances the mass transfer of organic constituents to the air.
Other removal mechanisms include adsorption on solids, seepage
through the ground, or degradation (e.g., by photolysis or
hydrolysis). For flowthrough systems, the organic constituents
may leave the unit with the effluent that will subsequently be
treated, stored, or disposed of.
Initial studies suggest that emission to air is a primary
removal mechanism, especially for volatile constituents.
Biodegradation in specific systems, particularly for
semivolatiles, may also be significant. For flowthrough systems,
the removal of semivolatiles with the effluent may also be a
primary removal mechanism. Other forms of degradation, adsorp-
tion, and seepage are neglected in this analysis for several
reasons. These mechanisms are not believed to be significant for
most systems and most constituents; however, they may be removal
routes in a specific system or for a specific constituent. For
example, an open tank may be designed specifically for liquid-
phase carbon adsorption. These mechanisms are also difficult to
model in a manner that is generally applicable considering the
relatively sparse data on such removal mechanisms, especially in
hazardous waste impoundments and tanks. Consequently, the
modeling effort focuses on mass transfer to the air and some
consideration of biodegradation.
Numerous studies have been conducted to assess mass transfer
to the air; these include theoretical assessments, correlations
based on laboratory and bench-scale measurements, and field
measurements at actual sources. Additional data on specific
5-65
-------
wastes have been collected in air- stripping studies as more air-
stripping columns have been used to remove organic constituents
from water. The result is that the state of knowledge of mass
transfer from the liquid to the gas phase (e.g., ambient air) is
probably advanced compared to the state of knowledge of other
removal mechanisms. The level of confidence in the air emission
models is probably highest for the volatile constituents because
of very high mass transfer rates. The level of confidence is
somewhat lower for the relatively nonvolatile constituents because
of potentially significant rates of removal by other mechanisms.
Much of the data on the performance of systems designed for
biodegradation are reported as total removal from measurement of
the influent and effluent concentrations. This total would
include removal to the air and biodegradation. Some studies have
been conducted in closed systems in which the biodegradation rate
may be measured directly (loss to the air is deliberately
prevented). These data are useful for comparing the relative
rates of removal by biodegradation among constituents and make
possible a ranking of these constituents with respect to
biodegradability. In addition, the estimated rate of
biodegradation may be compared to the estimated rate of air
emissions to assess the relative extent of each.
The biodegradation model has not been validated and is used
in this report as an approximate measure of the extent of
biodegradation. For any specific treatment system, measurements
of actual biodegradation rates should be used if available. Any
user of the biodegradation model should be aware that the
predicted rate is very sensitive to the choice of values for the
biorate, biomass concentration, and the concentration of organic
constituents in the waste. An environmentally conservative
approach with respect to air emissions would be to neglect
biodegradation (assume the rate is zero). This approach is
probably valid for volatile constituents in aerated systems;
however, the approach may tend to overestimate emissions of
5-66
-------
relatively nonvolatile constituents that are destroyed in treat-
ment systems specifically designed for biodegradation.
5.8.2 Major Assumptions
An inherent assumption in the emission estimating procedure
is that the mass transfer correlations chosen earlier are
generally applicable. A paper that compares several different
models concludes that, in most cases, many different models yield
56
comparable results for volatile constituents. The choice of
models may affect the estimated mass transfer coefficients for
semivolatiles more than those for volatiles. The calculations
indicate that emissions of volatiles are controlled by the liquid-
phase resistance. Consequently, the value for the overall mass
transfer coefficient (K) is primarily determined by the
correlation used for the liquid-phase mass transfer coefficient
(k ). For constituents with decreasing volatility, both the
j_i
liquid-phase and gas-phase resistance begin to contribute to the
overall resistance to mass transfer. For these constituents, the
choices of correlations for both ]<:„ and kT become important, and
(j LI
the choice of correlations may significantly affect the emission
estimates.
The flow model chosen for storage and treatment impoundments
assumes that the impoundment's contents are well mixed and that
the system is operated at steady-state conditions. The flow for
specific facilities may be better represented by plug flow or a
model that accounts for axial dispersion. The choice of flow
model does not make a significant difference in the estimated
emissions. However, if the loading of the impoundment is cyclical
or intermittent instead of continuous, the emissions from the
impoundment are likely to be cyclical or intermittent. Estimates
of short-term emission rates are very dependent upon the method of
operation of the system. For disposal impoundments, peak
emissions occur when the waste is first placed in the impoundment
and then decrease with time. The approach used in this report
5-67
-------
estimates the average emission rate over a given period of time
and does not provide an estimate of the initial peak emissions.
The calculation of Henry's law constant also contains
inherent assumptions. The approach is valid for dilute solutions
and has been applied successfully in the design of air-stripping
columns. However, specific mixtures may deviate from Henry's law
because of component interactions or because of concentrations
outside the range of applicability. Errors in applying Henry's
law are generally environmentally conservative; i.e., the actual
gas-phase concentration is not likely to be underestimated.
For concentrated mixtures of organics in a separate oil
layer, the use of Raoult's law is recommended. This approach is
valid for mixtures of constituents with similar properties,
especially when the concentration of the component of interest is
very high. A preferred approach would be to avoid the use of
Henry's law or Raoult's law and actually measure the equilibrium
partitioning between the liquid and gas phase of a waste.
However, very few data are available for equilibrium partitioning
that can be applied generally to hazardous waste mixtures.
5.8.3 Sensitivity Analysis
The emission correlations were evaluated for sensitivity to
57
each of the input parameters. In the analysis, each input
parameter was varied individually over the entire range of
reasonable values. The effect on emissions was noted, and the
most sensitive parameters were identified.
Detention time is an important parameter that affects
emissions from the impoundment. The emission estimates for
volatile constituents are sensitive to short detention times, and
the estimates for semivolatiles are sensitive to long detention
times. Essentially all of the volatile constituents are emitted
for longer detention times (several days), and very little of the
semivolatiles are emitted for short detention times (a few days) .
However, significant emissions of the semivolatiles may occur for
5-61
-------
long detention times in storage impoundments or in disposal
impoundments.
The value of Henry's law constant was not important for
volatile constituents. The correlations indicated that these
constituents are controlled by the liquid-phase resistance, which
is not affected by Henry's law constant. The value of Henry's law
constant has a direct effect on the emissions of semivolatiles
(such as phenol), and the greatest effect is on those relatively
nonvolatile compounds for which mass transfer is controlled by the
gas-phase resistance.
Windspeed has a direct effect on the emission estimates for
quiescent surfaces and has little effect on those from aerated
systems. The results showed that a standard windspeed of 5.5 m/s
was reasonable compared with the results for windspeed
distributions at actual sites.
Temperature did not affect the emission estimates for the
volatile constituents. However, temperature did affect the
emission estimates for nonvolatile constituents with mass transfer
controlled by the gas phase. The temperature dependence of
Henry's law constant accounts for this effect.
The diffusivity in air and water for a wide variety of
constituents spans a relatively narrow range of values. The
analysis showed that the emission estimates were not sensitive to
the choice of values for diffusivity.
For mechanically aerated systems, the choice of values for
impeller diameter, impeller speed, oxygen transfer rate, and
oxygen correction factor did not affect the emission estimates
significantly. The total horsepower and turbulent area had a
direct effect on emissions of semivolatiles (e.g., phenol).
However, there was no significant effect on emissions of volatile
constituents because the models predicted that they would be
stripped almost completely from the water over the full range of
aeration values.
5-69
-------
The biodegradation model was very sensitive to all parameters
investigated. The sensitive parameters include organic
concentration, biomass concentration, and biorate.
Two meteorological parameters required in the models are
temperature and windspeed. The emission estimates are based on a
standard temperature of 25 °C and a windspeed of 4.47 m/s
(10 mi/h). These standard values were evaluated by estimating
emissions for windspeed/temperature combinations at actual sites
based on their frequency of occurrence. Over a 1-yr period, the
results from site-specific data on windspeed and temperature were
not significantly different from the results using the standard
values. Consequently, the standard values were judged adequate to
estimate annual emissions. For short-term emissions, the actual
temperature and windspeed over the short-term interval should be
used to avoid underestimating emissions during high-windspeed
/high-temperature conditions.
A sensitivity analysis was performed for three impoundment
model units (storage, mechanically aerated, and disposal)
presented in the example calculations in this section. Three
compounds were chosen to represent relatively nonvolatile
compounds (p-cresol), moderately volatile compounds (acetone), and
relatively volatile compounds (benzene). Each of these compounds
can be biodegraded. The results are given in Tables 5-12, 5-13,
and 5-14. The key input parameters identified in the tables were
increased by 50 percent from the base case to determine the effect
on the percent of the compound in the waste that is emitted to the
air.
For each of the different types of impoundments, the
volatility appears to be important only for the low volatility
category. As discussed previously, the windspeed (air turbulence)
has a direct effect for each of the compounds in a storage
impoundment and does not affect the mechanically aerated unit's
results. The low volatility compounds are the most sensitive to
changes in depth and biomass concentration for all three types of
5-70
-------
impoundments. An assumption of no biodegradation also has the
most dramatic effect on the low volatility compound with smaller
effects observed for the higher volatility compounds. The effects
of retention time are small except for the results shown for the
disposal impoundment after 5 days. The disposal impoundment
results show that for short times, the time since disposal is an
important parameter affecting emissions.
5-71
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TABLE 5-12. RESULTS OF SENSITIVITY ANALYSIS FOR
QUIESCENT STORAGE IMPOUNDMENT
Percent emitted for given Henry1
law constant, atm»m3/mole
Key emission model inputs
Base case
1C
r7
2.9
1(
r5
58
1C
~3
59
50-percent increase from base case
Volatility
Air turbulence
Retention time
Depth
Biomass concentration
No biodegradation
4.2
4.0
3.2
2.1
2.1
10
(45)C
(38)
(10)
(-28)
(-28)
(245)
61 (5)
72 (24)
62
50
52
74
(7)
(-14)
(-10)
(28)
59 (0)
76 (29)
62
49
52
80
(5)
(-17)
(-12)
(36)
This corresponds to the model unit for storage
impoundments used in the example calculation.
Each parameter is increased individually by 50
percent from its base case value.
Q
Values in parentheses are percent change from the
base case.
Base case with no biodegradation.
5-72
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TABLE 5-13. RESULTS OF SENSITIVITY ANALYSIS FOR
MECHANICALLY AERATED IMPOUNDMENTS
Percent emitted for given Henry's
law constant, atm*m3/mole
Key emission model inputs
10
-7
10
-5
10
-3
Base case
2.7
79
99
50-percent increase from base case
No
Volatility
Air turbulence
Water turbulence
Retention time
Depth
Biomass concentration
biodegradation
3.
2
3
2
1.
1.
2
9
.8
.6
.7
8
8
0
(44)"
(4)
(33)
(0)
(-33)
(-33)
(640)
85
80
85
80
73
73
94
(8)
(1)
(8)
(1)
(-8)
(-8)
(28)
99
99
99
99
98
98
100
(0)
(0)
(0)
(0)
(-1)
(-1)
(1)
b
d
This corresponds to the model unit for mechanically aerated
impoundments used in the example calculation.
Each parameter is increased individually by 50 percent from its
base case value.
Values in parentheses are percent change from the base case.
Base case with no biodegradation.
5-73
-------
TABLE 5-14. RESULTS OF SENSITIVITY ANALYSIS FOR DISPOSAL
IMPOUNDMENTS
Percent emitted for given Henry's
law constant, atm*m3/mole
-7 -5 -3
Key emission model inputs 10 10 10
Base casea 13 93 92
50-percent increase from base case
Volatility
Air turbulence
Retention time
Depth
Biomass
No biodegradation
18
17
2.3
9
9
(38)b
(31)
(-82)
(-31)
(-31)
84 (550)
94
96
55
89
89
100
(1)
(3)
(-41)
(-4)
(-4)
(8)
92
96
72
88
89
100
(0)
(4)
(-22)
(-4)
(-3)
(9)
Based on the dimensions given in the example
calculation, 100 mg/L of the constituent in 1,000 mg/L
total organics, and a time since disposal of 12 months.
Values in parentheses are percent change from base
case.
A retention time of 5 days was selected here to show
the sensitivity to retention time soon after disposal.
Base case with no biodegradation.
5-74
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5.9 REFERENCES
1. Lunney, P. D. Characterization of Wind and Depth Effects
Upon Liquid Phase Mass Transfer Coefficients: Simulation
Studies. Master's thesis, University of Arkansas,
Fayetteville, AR. January 1983. p. 119.
2. Springer, C., P. D. Lunney, and K. T. Valsaraj. Emission
of Hazardous Chemicals from Surface and Near Surface
Impoundments to Air. U.S. Environmental Protection Agency,
Solid and Hazardous Waste Research Division. Cincinnati,
OH. Project Number 808161-02. December 1984. p. 3-4 to
3-16.
3. Reference 2, p. 3-16 to 3-19.
4. Reference 2, p. 3-18.
5. Hwang, S. T. Toxic Emissions from Land Disposal
Facilities. Environmental Progress. 1_: 46-52. February
1982.
6. Mackay, D., and A. Yeun. Mass Transfer Coefficient
Correlations for Volatilization of Organic Solutes from
Water. Environmental Science and Technology. 17:211-217.
1983.
7. Reference 6, p. 214.
8. GCA Corporation. Air Emissions for Quiescent Surface
Impoundments--Emissions Data and Model Review. Draft
Technical Note. Prepared for U.S. Environmental Protection
Agency. Contract No. 68-01-6871, Assignment 49. August
1985. p. 5-1 and 5-2.
9. Reference 8, p. 4-4.
10. Reference 5, p. 47.
11. Thibodeaux, L. J. Air Stripping of Organics from
Wastewater. A Compendium. Air/Water. p. 373. (In
publication.)
12. Westat Corporation. National Survey of Hazardous Waste
Generators and TSDF's Regulated Under RCRA in 1981.
Prepared for the U.S. Environmental Protection Agency.
Contract No. 68-01-6861. April 1984.
13. Metcalf and Eddy, Inc. Wastewater Engineering. New York,
McGraw-Hill. 1972. p. 542-554.
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14. Eckenfelder, W., M. Goronszy, and T. Quirk. The Activated
Sludge Process: State of the Art. CRC Critical Reviews in
Environmental Control. L5(2):148. 1984.
15. Beardsley, M., and J. Coffey. Bioaugmentation: Optimizing
Biological Wastewater Treatment. Pollution Engineering.
December 1985. p. 32.
16. Reference 13, p. 586.
17. Reference 15, p. 32.
18. Reference 13, p. 520-521.
19. U.S. Environmental Protection Agency. EPA Design Manual:
Municipal Wastewater Stabilization Ponds. Publication
No. EPA-625/1-83-015. October 1983. p. 3.
20. Reference 13, p. 557.
21. Englande, A. J. Performance Evaluation of the Aerated
Lagoon System at North Gulfport, Mississippi. Prepared for
U.S. Environmental Protection Agency. Publication No. EPA-
600/2-80-006. March 1980. p. 39-41.
22. Allen, C. Project Summary: Site Visits of Aerated and
Nonaerated Surface Impoundments. Prepared for U.S.
Environmental Protection Agency. Contract No. 68-03-3253.
Assignment 2-8. June 1987. p. 2.
23. Petrasek, A., B. Austern, and T. Neiheisel. Removal and
Partitioning of Volatile Organic Priority Pollutants in
Wastewater Treatment. Presented at the Ninth U.S.-Japan
Conference on Sewage Treatment Technology. Tokyo, Japan.
September 1983. p. 16.
24. Bishop, D. The Role of Municipal Wastewater Treatment in
Control of Toxics. Presented at the NATO/CCMS Meeting.
Bari, Italy. September 1982. p. 18.
25. Hannah, S., B. Austern, A. Eralp, and R. Wise. Comparative
Removal of Toxic Pollutants by Six Wastewater Treatment
Processes. Journal WPCF. 58 (1) :30. 1986.
26. Kincannon, D., and E. Stover. Fate of Organic Compounds
During Biological Treatment. Presented at ASCE
Environmental Engineering Conference. 1981. p. 6.
27. Melcer, H. Biological Removal of Organic Priority
Pollutants. Presented at Hazardous Substances in
Wastewater Seminar. Toronto, Canada. November 1982.
p. 20.
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28. Reference 19, p. 75-146.
29. Reference 13, p. 481-573.
30. Reference 14, p. 119.
31. Bailey, J. E., and D. F. Ollis. Biochemical Engineering
Fundamentals. New York, McGraw-Hill. 1977. p. 343-349.
32. Kincannon, D., and E. Stover. Determination of Activated
Sludge Biokinetic Constants for Chemical and Plastic
Industrial Wastewaters. Prepared for U.S. Environmental
Protection Agency. Publication No. EPA-600/2-83-073a.
August 1983. p. 18-20.
33. Fitter, P. Determination of Biological Degradability of
Organic Substances. Water Research. 1_0: 231-235. 1976.
34. Kincannon, D., A. Weinert, R. Padorr, and E. L. Stover.
Predicting Treatability of Multiple Organic Priority
Pollutant Wastewater from Single Treatability Studies.
Presented at the 37th Purdue Industrial Waste Conference,
West Lafayette, IN. May 1982.
35. Reference 23, p. 4-16.
36. Reference 25, p. 27-34.
37. Coburn, J., C. Allen, D. Green, and K. Leese. Site Visits
of Aerated and Nonaerated Impoundments. Revised Draft
Summary Report. Prepared for U.S. Environmental Protection
Agency. Contract No. 68-03-3253, Work Assignment No. 3-8.
April 1988. p. A-l to A-34.
38. Matter-Muller, C., W. Gujer, W. Giger, and W. Stumm. Non-
Biological Elimination Mechanisms in a Biological Sewage
Treatment Plant. Prog. Water Tech. 1_2_:305. 1980.
39. Dobbs, R., M. Jelus, and K. Chang. Partitioning of Toxic
Organic Compounds on Municipal Wastewater Treatment Plant
Solids. Proceedings of the International Conference on
Innovative Biological Treatment of Toxic Wastewaters.
Scholze, R. J., Ed. Arlington, VA. June 1986. p. 585-
601.
40. Reference 5, p. 46.
41. Reinhardt, J. R. Gas-Side Mass-Transfer Coefficient and
Interfacial Phenomena of Flat-Bladed Surface Agitators.
Ph.D. dissertation, University of Arkansas, Fayetteville,
AR. 1977. 96 p.
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42. GCA Corporation. Emissions Data and Model Review for
Wastewater Treatment Operations. Draft Technical Note.
Prepared for U.S. Environmental Protection Agency.
Contract No. 68-01-6871, Assignment 49. August 1985.
p. 4-3.
43. Reference 42, p. 4-2.
44. Reference 5, p. 47.
45. Reference 42, p. 4-3.
46. Reference 41, p. 48.
47. Reference 19, p. 3.
48. Reference 13, p. 557.
49. Reference 13, p. 519.
50. GCA Corporation. Hazardous Waste TSDF Waste Process
Sampling. Prepared for U.S. Evironmental Protection
Agency. Report No. EMB/85-HNS-3. October 1985. p. 1-11.
51. GCA Corporation. Evaluation and Selection of Models for
Estimating Air Emissions from Hazardous Waste Treatment,
Storage, and Disposal Facilities. Prepared for U. S.
Environmental Protection Agency. Publication No. EPA-
450/3-84-020. December 1984. p. 69.
52. Thibodeaux, L., and D. Parker. Desorption Limits of
Selected Gases and Liquids from Aerated Basins. AIChE
Sumposium Series. 72 (156) :424-434 . 1976.
53. Reference 51, p. 67.
54. Reference 51, p. 67.
55. Reference 13, p. 519.
56. Allen, C. C. Prediction of Air Emissions from Surface
Impoundments. Paper 31a. (Presented at 1986 Summer
Meeting of AIChE. Boston, MA. August 1986.) 26 p.
57. Branscome, M., and A. Gitelman. Sensitivity Analysis:
Emission Estimates for Surface Impoundments. Prepared for
the U.S. Environmental Protection Agency. March 1986.
67 p.
5-78
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6.0 WASTEWATER TREATMENT MODELS (WATERS)
This section describes a series of wastewater models that
can be used to estimate air emissions from miscellaneous
wastewater treatment units. Many of the models presented in this
section are not included with CHEMDAT8 due to the nature of the
calculations that are required. Section 6.1 presents an overview
of the models that are included in WATERS and some general
guidance for the use of these models. Section 6.2 presents a
discussion of trickling filters and a proposed model for
trickling filters. Section 6.3 discusses a cooling tower model.
Section 6.4 discusses a model for an API separator. Table 6-1
lists selected units and the appropriate models.
6.1 UNITS FOR MODELING EMISSIONS OF VOLATILE COMPOUNDS
Although presented as discrete units, it should be noted
that these units are present in a number of different treatment
plants, and that most treatment plants can be composed of unit
processes that fit into the broad categories of the units defined
here. For example, a trickling filtration unit could be used in
the treatment train ahead of an activated sludge unit. In this
capacity, the trickling filter operates as a roughing filter to
pretreat wastewater prior to secondary treatment and not as a
secondary treatment process.
It should be emphasized that treatment systems vary widely
depending on the nature of the wastewater, the availability of
land, prior regulatory pressure, the composition and flow rate
6-1
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TABLE 6-1. REFERENCE
TREATMENT UNITS
TABLE FOR THE LOCATION OF WASTEWATER
MODELS AND RELATED DISCUSSION.
Unit Description
Collection system
Sump
Cooling tower
Wastewater separator
Trickling filters
mix tanks
Activated sludge
Agitated impoundment
Disposal impoundment
Plug flow system
Trench
Clarif ier
Storage tank
Waterfall or weir
Pretreatment
Oil film surface
WATERS
YES
YES
YES
YES
YES
YES
YES
YES
YES
YES
YES
YES
YES
YES
YES
YES
CHEMDAT8
NO
NO
NO
NO
NO
YES
YES
YES
YES
NO
NO
NO
NO
NO
NO
YES
THIS REPORT
Section 4
Section 4
Section 6
Section 6
SECTION 6
SECTION 5
SECTION 5
SECTION 5
SECTION 5
SECTION 5
SECTION 4
SECTION 4
SECTION 9
SECTION 4
SECTION 6
SECTION 5
6-2
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that may have existed at the time the system was designed, and
many other factors. The units that follow do not represent a
"typical" system. The unit parameters can be defined as those of
a specific system component parameters to estimate the air
emissions from that specific system component.
6.1.1 Conventional Activated Sludge System
A typical configuration of an activated sludge system is
pretreatment, optional primary sedimentation (primary clarifier)
followed by the aeration process including secondary
clarification, and post treatment. The principal treatment
process is the aeration tank. This is a continuous flow,
biological treatment process characterized by the turbulent
suspension of microscopic aerobes. The turbulence promotes
mixing and induces a relatively homogenous state in which the
microbes are able to absorb and oxidize soluble and colloidal
organics. The process involves an aeration step followed by a
solid-liquid separation step in which part of the separated
sludge is recycled back to the system for mixing with the raw
influent.
There are many variations of the activated sludge process;
however, they generally can be reduced to looking at either the
loading rates in terms of BOD or the physical arrangement of the
process train. The loading is typically one of three basic
types. High rate takes advantage of the settleability of sludge
when the system is loaded at a rate of 0.80-1.15 g of BOD/g of
mixed liquor suspended solids per day. Conventional rate is of
the range 0.2 to 0.5 g BOD/g mixed liquor volatile suspended
solids per day. This rate is typical for most larger municipal
treatment plants. Extended aeration rate is the lowest range of
process loading and is used in those plants which are small in
size and do not receive 24 hour supervision. As such they are
generally conservatively designed and operate in the range of
0.05-0.15 g of BOD applied/g of MLVSS/day; industrial wastewaters
6-3
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vary widely in terms of influent concentration and
biodegradability. Thus, a wide range of loading rates are used
depending on individual circumstances.
Physical arrangements are of three types; the complete mix
activated sludge, plug flow activated sludge, and activated
sludge with reaeration. In the complete mix arrangement, the
return sludge and the wastewater are uniformly introduced into
the aeration basin through several points in order to obtain a
homogeneous mixture. In a plug flow arrangement both the
untreated wastewater and the return sludge are introduced at the
head of the plant and flow through the plant in a modified plug
flow. Such plants are often compartmentalized to maintain the
plug flow regime. Activated sludge with sludge reaeration
constitutes a rearrangement of process streams. In this instance
the sludge is compartmentalized and aerated prior to its contact
with the untreated waste.
6.1.2 Sludge handling
Sludge handling involves the stabilization of the
solid-water mixtures derived from the primary and secondary
clarifier as well as the excess biomass from the activated sludge
process and chemical reactions. These mixtures undergo
thickening, anaerobic or aerobic digestion and dewatering prior
to ultimate disposal. Anaerobic digestion is designed for
minimal air/sludge contact. Emissions from the other processes
are likely to be insignificant because the upstream processing
units will have provided extensive opportunities for
volatilization prior to the sludge handling operations.
6.1.3 Conventional Activated Sludge (Mechanical Aeration)
The principle component of the mechanical aeration system is
the aerator. There are two types in general use today, surface
aerators and turbine aerators. The surface aerator is highly
developed and widely used, particularly in the treatment of
industrial waste. The surface aerators may either float or be
6-4
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mounted on supports in the aeration basin. They enhance the
entrainment of atmospheric air in the aeration basin by producing
a region of high turbulence around the periphery of the aerator.
Oxygen transfer efficiency of these aerators increases with the
depth of submersion, as does power cost; consequently, there is a
trade off between efficiency and cost.
Since 1950, the submerged turbine has been widely used in
the chemical industry. It offers an attractive means of
upgrading existing facilities to handle increased loads. These
aerators are used because of improved oxygen transfer efficiency
and lower horsepower requirements. Oxygen transfer efficiency
for aerators, as rated in terms of mass of oxygen transferred per
energy input, is typically on the order of 1.2 Kg 02/KW-hr (2 Ib
02/hp.hr). Air and energy requirements for an aeration system
are typically on the order of 50-90 m3/kg BOD removed and 0.040
to 0.26 Kw/m3 of basin volume.1
The mechanical aerator approach is found in large open
basins particularly in those plants operating in a complete mix,
conventional activated sludge mode. The turbulence introduced by
the rotary action of the aerator blades promotes a homogeneous
mixing and enhances the overall complete mix mode of operation.
6.1.4 Conventional Activated Sludge (Diffused Air: Coarse and
Fine Bubble)
A second approach to aeration is the use of diffuser systems
which are generally used in plug flow systems and sludge
reaeration systems, the most common types of aeration systems
used in activated sludge plants. The distribution system
consists of an array of diffusers situated near the bottom of the
basin. These diffusers are designed to produce either coarse or
fine bubbles and are supplied with air by compressors. In the
period from 1950 to 1978, the fine bubble systems were in wide
use. At that time, it was felt that the increase in oxygen
transfer efficiency of the smaller bubble diameter (8 percent vs.
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5 percent for the coarse bubble) was important. Later, however,
inefficiencies such as clogging decreased the overall
attractiveness of these systems.
The most common type of fine bubble diffusers are nylon or
Dacron socks and saran wrapped tubes. Other systems include
porous ceramic plates that generate small diameter bubbles.
Coarse bubble diffusers can be tubes covered with synthetic
fabric or wound with filaments, and sprayers with multiple
openings created by drilling holes in pipes or loosely attaching
plates or discs to a supporting piece of pipe. Although the
oxygen transfer efficiency is lower, coarse bubble diffusers do
not suffer from clogging and have lower initial cost and
maintenance. Many treatment plants are reported to have switched
to the coarse bubble systems in order to take advantage of these
features.2
6.1.5 Aerated Lagoons (Mechanical Air)
Aerated lagoon systems are medium depth basins designed for
biological treatment on a continuous flow basis. They are
equipped with surface aerators and are primarily used to treat
wastes of low-medium strength in areas where land is inexpensive.
They are not as widely used as stabilization ponds, but their
feasibility has been fully demonstrated and they may represent an
upgrading of an oxidation pond.
Aerated lagoons have detention times on the order of 3-
10 days. Aerated lagoons are staged in series and are designed
to achieve partial mixing. Consequently, aerobic and anaerobic
stratification can occur. A large fraction of the incoming
solids may in fact settle out near the head of the plant.3
6.1.6 Spray Evaporation Ponds
Spray evaporation ponds are used primarily to reduce the
amount of water contained in a waste. These are basically ponds
equipped with submersible pumps attached to vertical pipes ending
in standard irrigation spray headers. Water is pumped through
6-6
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this system and dispersed in the air. As the droplets fall back
to the pond they are enriched with oxygen and subjected to
evaporative processes. These ponds occur primarily in waste
treatment systems involving large quantities of recycled water.
Evaporation ponds without spray headers are termed solar
evaporation ponds and are not intentionally aerated.
6.1.7 Dissolved Air Flotation (DAF)
Dissolved air flotation is widely used in industry to remove
suspended solids by flotation. The flotation of the particulate
is induced by microscopic air bubbles attaching to the
particulate or agglomerate and giving it buoyancy. Particles are
floated to the surface where they are removed by skimmers for
further treatment.
The DAF system generates a supersaturated solution of
wastewater and air by pressurizing either the influent wastewater
(or a side stream of the influent wastewater) and introducing
compressed air. The pressure is then released in the detention
tank generating the numerous microscopic bubbles which adhere to
particulates or are trapped by any floe which may be present.
6.1.8 Neutralization (Equalization) Process
Although neutralization and equalization units perform
different functions, i.e., pH neutralization vs. flow
equalization, these operations can be considered together as they
permit similar modes of air/water contact. Primarily, these
units are open basins or tanks with varying size depending upon
the desired retention time. Mechanical agitation by stirrers is
used to assure a homogeneous mixture. The design criteria for
these processes are dependent on the variation in influent
composition. For example, when the objective is equalization,
more erratic fluctuations in the influent composition
necessitates longer residence times.
6.1.9 Miscelaneous Physical-Chemical Treatment Systems
6-7
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Physical-chemical processes are generally defined as those
operations which effect the removal and/or destruction of
undesirable constituents in wastewater by means other than
biological degradation or biological conversion.
Physical-chemical processes include a wide array of
traditional and innovative processes. These processes include
dissolved air flotation, and mechanically agitated
equalization/neutralization basins. These processes can be used
as adjuncts to the model plant flow charts presented in this
section. If appropriate, an open agitated tank model can be used
to characterize some of the miscelaneous wastewater treatment
operations.
6.2 AIR EMISSIONS OF VOLATILE COMPOUNDS FROM TRICKLING FILTERS
The typical trickling filter plant consists of the following
units: pretreatment, primary clarifier, trickling filter,
secondary clarifier and post-treatment unit (see Figure 6-1).
The heart of the system is the trickling filter itself, which
consists of a circular basin 1-2.4 m deep packed with a bed of
either rock or plastic media, over which wastewater is sprayed.
A zoogleal slime which attaches to the media assimilates and
oxidizes the organics in the wastewater. Oxygen and organic
matter diffuse into the zoogleal mass and end products of
oxidation counter-diffuse back into the flowing liquid or to the
void spaces. The treated water and any particulates from the
filter bed are collected in an underdrain system and sent to
secondary clarifiers for sedimentation.
The packing media is typically dosed with a rotary
distributor which sprays the waste over the media. The media may
be either plastic or rock. The rock medium represents a
traditional approach; the plastic however, offers advantages such
as lower specific weights and higher void spaces and is amenable
to above ground installation.
6-8
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The performance of the unit is affected by many factors such
as hydraulic and organic loadings, depth and physical
characteristic of the media, the method of wastewater
distribution, ventilation, and characteristics of the applied
wastewater.2 Municipal wastewater and a wide variety of
industrial wastewaters are amenable to treatment in trickling
filters.
inlet water
recycle to filter
exit water
underflow from filter
Figure 6-1. Illustration of a trickling filter.
6-9
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The principal components of the trickling filter process
are:
1. The distribution system
2. The filter media
3. The underdrain
4. Final sedimentation.
The rotary distributor consists of two or more horizontal arms
mounted on a turntable assembly anchored to a center column. The
wastewater is uniformly distributed over the media through
orifices located in the arms. The principal drive mechanism for
these arms is the reaction force from the spray on the radial
arms. The arms are sized to limit velocities to 1.2 m/sec at
maximum flow. The rotation speed of the arms varies with flow
rate in the range of 0.1-2 rpm.
Ventilation is extremely important in achieving efficient
filter operation. Usually, if the underdrain is properly sized,
the differences in air and water temperature will provide a
natural driving force for ventilation. An air flow rate of
approximately 0.03 m3/m2 filter area per minute is required to
sustain aerobic conditions within the bed. When forced
ventilation systems are required, they are typically designed to
provide an air flow of 0.3 m3/m2 of filter area per minute.
Organic and hydraulic loading determines the classification
of the filters: low-rate, high-rate, or roughing-rate. Low-rate
filters are generally not equipped with recirculation an are
rarely used. High-rate filters use recirculation to dilute the
influent organic strength and to flush the media voids. This
permits higher BOD loadings per volume of media and promotes the
return of activated organisms as a seed. The high-rate filters
are generally designed to accept a continuous flow of wastewater
and may be either single stage or two staged. High-rate filters
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also have a number of modifications of the basic recirculation
scheme.
Roughing rate trickling filters provide an intermediate
stage of treatment and are used frequently to precede activated
sludge units or second stage filters. The purpose of this
operation is to reduce high organic loadings prior to further
treatment. This intermediate stage is typical for industrial
systems.
The trickling filter model is based on a design currently in
operation in a U.S. municipality.2 It represents a single train
of a multi-train high rate process. The operating conditions and
specifications fall within the range expected for industrial
waste treatment. The design parameters are given in Table 6-2.
6.3 AIR EMISSIONS OF VOLATILE COMPOUNDS FROM COOLING TOWERS
Cooling towers are used in the chemical industry and in the
pulp and paper industry to cool the wastewater before biological
treatment. Excessively high wastewater temperatures can cause
the biological treatment plant to fail to perform as designed.
Cooling towers have been used in pulp mills, even in the cooler
climate of the north central United States. Part of the
wastewater evaporates, cooling the wastewater. An illustration
of a cooling tower is presented in Figure 6-2.
Cooling towers may not be needed to cool high temperature
wastewater if aeration basins are located before the biological
units. It has been observed in several plants that only part of
the wastwater has been diverted to the cooling tower. The
overall temperature of the combined wastewater should be less
than 50 °C. A cooling tower is typically a forced air cooling
tower where the wastewater is contacted with ambient air. In the
mass transfer with the ambient air, volatile organics can
transfer to the air along with the water.
6-11
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TABLE 6-2. MODEL PLANT OPERATIONAL PARAMETERS: TRICKLING FILTER
Parameter
Plant Flow
Plant Performance
Influent BOD
Influent Suspended
Solids
Trickling Filter
Clarif iers
Value
6.4 MGD (0.28 m3/s)
85 percent BOD Removal
75 percent Suspended Solids Removal
183 mg/1
188 mg/1
Diameter
Depth
Area
Volume
Hydraulic loading
Recirculation
Diameter
Depth
Area
Weir height
Surface loading
Detention time
190 ft (58 m)
5 ft (1.5m)
28353 ft2 (2640 m2)
141764 ft3 (3960 m3)
29 MGD/acre
(1.1 m3/m2-hr)
190 percent
100 ft (30 m)
9.2 ft (2.8 m)
7854 ft2 (730 m2)
1 ft (30 cm)
1350 gal/ft2/day
(0.47 m3/m2-day)
1 . 2 hours
6-12
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Figure 6-2. Cooling tower,
6-13
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6.3.1 Cooling Tower Default Parameters
The typical operating conditions of cooling towers at pulp and
paper mills are presented in Table 6-3.3
6.3.2 Performance Data of Cooling Towers
Reference 3 indicates in the abstract that 25 percent to 30
percent BOD reduction (presumably predominantly methanol) can be
achieved by cooling tower treatment. This corresponds to
physical stripping of volatile components at a rate of 60,000 Ib
BOD/day or 11,000 tons BOD/year from a Kraft linerboard mill of
850 ton/day. The wastes treated included the pulp mill
condensates, the decker filtrate, turpentine decanter underflow,
and the condenser waters from a barometric type evaporator.
The BOD removal in a laboratory cooling tower was related to
the liquid to gas ratio. Lower liquid rates permitted a cooling
tower to remove up to 70 percent of the BOD. There was some
evidence of biodegradation contributing to the removal of BOD in
the cooling tower, up to 15 percent of the total BOD removal. It
was demonstrated that the main removal mechanism was air
stripping.
Figure 6-1 illustrates a cooling tower with recycle. Some
of the cooled water is recycled to the entrance of the cooling
tower. This permits multiple passes of part of the wastewater
being treated with the cooling tower, basins are located before
the biological units. It has been observed in several plants
that only part of the wastwater has been diverted to the cooling
tower.
6.3.3 Air Emission Modeling for Cooling Tower
The method selected for the modeling of air emissions from a
cooling tower is to model both the mass transfer of water and the
mass transfer of methanol by the same mechanism. The predicted
performance of the cooling tower would then be subject to
6-14
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TABLE 6-3 TREATMENT OF SELECTED INTERNAL KRAFT MILL WASTES
IN A COOLING TOWER.
Parameter
Air velocity
Waste loading
Inlet temperature
Exit temperature
Recycle ratio for treatment
Blowdown
Recycle ratio for cooling
Value
200-600 ft/min
1-4 gal/min-ft2
50 C
32 C
0.8 gal recycled/gal leaving
the tower
15 percent to 20 percent of
tower flow
none, assumed for current case
6-15
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verification and model parameter adjustment by temperature
measurements. The overall mass transfer of the methanol in the
cooling tower is given by the two resistance model. In the case
of water, the resistance of the liquid phase is negliable.
where, the above equation has the following variables and units:
K01 cm/s overall mass transfer coefficient;
K! cm/s liquid phase mass transfer coefficient;
Kg cm/s gas phase mass transfer coefficient;
H atm-m3/mol Henry's law constant; and
40.9 mol/atm-m3 1/RT (at 25C, R is the gas constant).
For computational purposed, the cooling tower is divided
into ten equal sections by partitioning with imaginary horizontal
planes. The number of moles transferred in each volume element
of the cooling tower is given by the following equation:
-AL- (6-2)
dt 18 ( IL K)
where, the above equation has the following variables and units:
dm mols mols transferred to the gas phase;
t sec time;
K01 cm/s overall mass transfer coefficient;
A cm2 wetted surface area in tower section;
x1L mol fraction methanol in water;
18 cm/s per mol/cm2-s-mol fraction;
K y/x Henry's law constant; and
y mol fraction methanol in the gas.
The mass transfer is calculated for each of the ten segments
in the cooling tower. The values of the temperature of the gas,
the temperature of the liquid, the equilibrium concentration of
6-16
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water vapor, the flow rate of the gas phase, and the Henry's law
value is calculated separately for each of the segments in the
cooling tower.
6.3.4 Material Balance with Recycle
The following terms describe the cooling tower recycle
concentrations:
= the inlet mol fraction of the wastewater;
vout = the outlet mol fraction of the treated wastewater;
f = the fraction of the component removed each treatment
pass;
r = the recycle fraction of the cooling tower water;
c = the ratio of xout to x±; and
F = the overall fraction removed by the cooling tower.
The overall removal may be written as follows:
C = ( 1 - F) (6-3)
F = '"' "°uv (6-4)
On a single pass the removal is a function of the inlet and
outlet concentrations:
f = l-x I 1-r] + v (r] - v ] / l-x (1-r] + v (r] ] (6-S'
-L \^i \ -L J- I ' xout \LI xout I ' \xi \ -1 Ll ' xout \ L I I \ D ° i
or,
f = (1-r + c r - c) / (1-r + c r ) (6-6;
Rearranging the previous equation,
f-fr + fcr = l-r + cr-c (6-7)
Substituting the equation for c into the above equation,
6-17
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f - fr + fr(l - F) = 1 - r + (r - 1)(1 - F)
Solving for f, an equation is obtained which relates the single
pass removal to the removal with recycle:
,
1 -r + r(l - F)
6-9,
Reference 3 reported 20 to 30 percent removal of methanol
with a recycle ratio of r = 0.8. Assuming an average value of
25 percent removal, F = 0.25. Substituting the values of r and F
into the above equation, the single pass removal fraction f is
estimated as 0.0625.
0.0625 = ~
1 - 0.8 + 0.8 (l - .25)
With 15 percent removal under the same conditions, the
single pass removal fraction f is estimated as 0.034. From the
available data from Table 6-3, it is concluded that the removal
of methanol in the wastewater treated in a cooling tower is
between 3 and 6 percent. Greater removal of methanol is expected
with cooling tower recycle.
6.4 ESTIMATION OF AIR EMISSIONS FROM API SEPARATOR UNITS
This section presents the model for the API separator and
illustrates the use of the model with a sample calculation. The
API separator model is composed of three regions: the flow
distribution region, the separation region, and the exit region
that may have flow over a weir. The total air emissions are the
sum of the air emissions from the three regions.
6.4.1 API Separator Model Elements
The API separator is modeled as the unit which separates oil
from the wastewater. If additional units are used to treat the
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wastewater before separation in the API unit, those additional
units should be modeled separately. Also, if additional units
are used to dewater the oil by heating or other methods, those
additional oil units should be modeled separately.
6.4.1.1 Region 1 flow distribution. The mass transfer from
the wastewater in the flow distribution region is characterized
by the resistance of two phases, the liquid phase resistance and
the gas phase resistance. The overall mass transfer from this
two-resistance model, K , is a combination of the gas and the
liquid mass transfer coefficients:
"
9 Kg K fp
where,
K0 = the overall mass transfer coefficient based upon the
liquid concentrations (m/s);
K! = the liquid phase mass transfer coefficient (m/s);
Kg = the gas phase mass transfer coefficient (m/s);
fp = the fraction of the compound in the water phase; and
K = the partition coefficient (atm-m3/mol) .
Kg is estimated by:
-•11 ( 0.0012 D V6?
K = 0.00482
0.000181
K! is estimated by:
K, = 2.61 10~7 —1
1 ( 100 j 0.0000085
where,
V = the wind velocity at 10 meters over the surface (cm/s);
dia = the width of the region (cm);
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Dg = the gas phase diuffusivity (cm2/s); and
D! = the liquid phase diffusivity (cm2/s).
The fraction of the compound in the water phase is used to
correct the partitioning in the gas phase. The fraction of the
compound in the water phase is estimated from the octanol-water
partition coefficient.
OWR = UVm; °lltract (6-11)
1 - oilfract
f = OWR (6-12)
1 + OWR
Fn = 1 - fn (6-13)
where,
OWPC = the octanol water partition coefficient;
oilfract = the fraction of the waste that is oil and
insoluble in water;
OWR = the ratio of the amount in oil to the amount in
the water;
f0 = the fraction of the component in the oil phase;
and
fp = the fraction of the component in the water phase,
The air losses, falri, from the two-resistance model are as
follows:
' K A'
f,, = 1 - EXP I - -2—1 (6-14!
airl
where,
K0 = the overall mass transfer coefficient based upon the
liquid concentrations (m/s);
q = the liquid flow rate (m3/s);
A = the surface area of the region (m2); and
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falrl = the liquid flow rate (m3/s),
6.4.1.2 Region 2 oil film separation. The mass transfer
from the floating oil on the wastewater surface is characterized
by a resistance of only one phase, the gas phase resistance. The
overall mass transfer from this one resistance model, K0, is
estimated as follows:
Ko = 240KgKq (6-15)
where,
K0 = the overall mass transfer coefficient based upon the
liquid concentrations (m/s);
Kg = the gas phase mass transfer coefficient (identical to
region 1, m/s);
Keq = the partition between the gas phase and the oil phase;
and
K = the partition coefficient (atm-m3/mol) .
The fraction of the compound in the water phase is used to
correct the partitioning in the gas phase. The fraction of the
compound in the water phase is estimated from the octanol-water
partition coefficient as described above.
The partition between the gas phase and the oil phase is
estimated as follows:
0.0012 V mwt
K = p- (6-16)
eq p 28.8 760
where,
Vp = the vapor pressure of the pure component at the surface
temperature (mm Hg);
mwt = the molecular weight of the compound (g/g-mol); and
• = the liquid density (g/cm3) .
6-21
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The air losses, falr2, from the one-resistance model are as
follows:
- f K A
f ^ = 1 - EXP '
air2
'6-17'
where
K0
q
fo
A
fair2 ~~
the overall mass transfer coefficient based upon the
liquid concentrations (m/s);
the liquid flow rate (m3/s);
the fraction of the compound in the oil phase;
the surface area of the region (m2); and
the liquid flow rate (m3/s).
The concentration in the oil phase on the surface is assumed
to be in equilibrium with the concentration in the aqueous phase.
The exponential form of the estimation of falr2 prevents the
possibility of estimating air emissions that are in excess of the
total amount present.
Some of the component will be removed with the oil that is
removed from the surface. The fraction in the oil is estimated
with the following equation. Components removed with the oil are
not available for contributing to air emissions in region 3.
roil
- f
- falr2
6.4.1.3 Region 3 weir overflow. The air emissions from the
weir outfall of the API separator are estimated by a modification
of the weir model presented by Nakasone.4 The equations used in
this method are presented in this section.
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In (r) = 0.0785 Z131 qa428 h0310 (6-19;
where,
r = Cs/(Cs-Co) deficit ratio, assumes that there is no
oxygen before the weir;
Cs = saturated oxygen concentration;
Co = outlet oxygen concentration;
Z = distance of fall (m), includes 1.5 times the distance
from the weir top to the critical depth above the weir;
q = flow rate per length of weir (m3/h-m); and
h = the tailwater depth (m).
It is assumed that the rate limiting step for the diffusion
of oxygen is the mass transfer in the liquid phase (oxygen is
only slightly soluble in the water). From the following
equation, a value of the liquid phase mass transfer coefficient
can be estimated for volatile organics, after correcting for the
relative diffusion coefficient of the organic and oxygen in
water.
_ q
D ^ °-66 ( v, \
_Jl ln(r) —2^ (6-20)
^ ' 3600 sj
lo
where,
= the mass transfer coefficient of the volatile organic
in water (m/s) ;
ln(r) = the natural log of the deficit ratio for oxygen in
the water flowing over the weir;
Dlv = the diffusion coefficient of the volatile organic in
water (cm2/s) ;
D10 = the diffusion coefficient of oxygen in water
(0.000024 cm2/s) ;
6-23
-------
Z = distance of fall (m) , includes 1.5 times the distance
from the weir top to the critical depth above the
weir; and
q = flow rate per length of weir (m3/h-m) .
The gas phase mass transfer coefficient of the flow at the
weir is estimated as 0.05 m/s for benzene. This is approximately
the magnitude for mechanically aerated systems. The assumption
of a gas phase mass transfer coefficient of this magnitude will
lower the estimate of the oxygen transfer from the correlation by
only a few percent. Significantly lower gas phase mass transfer
coefficients will lower the predicted oxygen transfer to an
extent not predicted by the correlation.
°-66
where,
kg = the mass transfer coefficient of the volatile organic
in air (m/s);
Dgv = the diffusion coefficient of the volatile organic in
air (cm2/s) ;
Dgo = the diffusion coefficient of the reference material in
air (0.088 cm2/s); and
0.05 = the assumed mass transfer coefficient of a turbulent
surface.
The value of the overall mass transfer coefficient is
estimated by combining the liquid and gas mass transfer
coefficients.
-i
(6-22)
Kj 40.9 Kg K
where,
6-24
-------
K0 = the overall mass transfer coefficient based upon the
liquid concentrations (m/s);
K! = the liquid phase mass transfer coefficient (m/s);
Kg = the gas phase mass transfer coefficient (m/s); and
K = the partition coefficient (atm-m3/mol) .
The fraction of the volatile components that are emitted to
the air is estimated by the following relationship:
f. = l - EXP - 36°° S6C (6-23)
hr
where,
K0 = the overall mass transfer coefficient based upon the
liquid concentrations (m/s);
q = the liquid flow rate per length of the weir (m3/h-m) ;
Z = the distance of fall (m) ; and
falr = the fraction of the volatile component emitted to
the air.
6.4.2 Example Calculation
It is assumed that an aqueous stream with 1 percent
dispersed oil is cleaned in an open API separator. The oil
stream is stored in a heated fixed roof tank to remove the water
from the oil. The heated tank is vented without a condenser to
the atmosphere. In this example the air emissions of benzene is
estimated from the separator unit.
The open API oil/water separator used for the example
calculations was characterized as having three regions, an
entrance distribution region of 20 m2, a separation region where
the floating oil was removed, and a third region with a weir.
Since the concentration of benzene in the wastewater is not
specified, the estimates of air emissions should be expressed in
fraction of inlet benzene that is lost to the atmosphere.
6-25
-------
6.4.2.1 Properties of benzene and unit specifications.
Properties of benzene as well as the constants needed for the
sample calculations are presented in Tables 6-4 and 6-5.
6.4.2.2 Region 1 Calculations. The fraction of the compound
in the water phase is used to correct the partitioning in the gas
phase. The fraction of the compound in the water phase is
estimated from the octanol-water partition coefficient:
where,
OWR
0.412
1.4268
OWPC oilfract
1 - oilfract
1 - 0.588
141.25(0.01)
1 - 0.01
OWR
0.588 =
1 + OWR
1.4268
1 + 1.4268
(6-24)
TABLE 6-4. PROPERTIES OF BENZENE USED FOR SAMPLE CALCULATIONS
Variable
Diffusivity in water
Diffusivity in air
Molecular weight
Henry's law constant
Diffusivity 02 in water
Vapor pressure benzene
Octanol water partition
coefficient
Symbol
Dl
Dv
Mwt
H
Do
vp
owpc
Number and units
0.98 x 1(T5 cm2/s
.088 cm2/s
78 g/g-mol
0.00555 atm-mVmol
2.5 x 10~5 cm2/s
95.2 mm Hg
141.25
6-26
-------
TABLE 6-5. UNIT PARAMETER NAMES AND SPECIFICATIONS
FOR EXAMPLE CALCULATIONS.
Unit Specification
Wastewater flow rate
Wind speed
Number of units
Temperature
Region 1 area
Region 2 area
Oil in waste
Density of oil
Oil molecular weight
Waterfall drop height
Waterfall width
Symbol
q
V
n
T
fo
do
mwt
h
w
Variable Name
q
V
n%
T
area . enter
area . oil
oilf ract
densoil
mwtoil
drop
widthfall
Value
0.10 m3/s
447 cm/s
1
25 deg. C
20 m2
50 m2
0.01
0.7 g/cm3
180
20 cm
4 m
6-27
-------
OWPC = the octanol water partition coefficient;
oilfract = the fraction of the waste that is oil and
insoluble in water;
OWR = the ratio of the amount in oil to the amount
in the water;
f0 = the fraction of the component in the oil
phase; and
fp = the fraction of the component in the water
phase.
fp = 1 - f
The effective diameter of the region 1 surface is estimated
with the following general equations:
dia = J — v^; (6_25)
71
20m2 (4) ( 100cm
505 cm
In region 1, the area is 20 m2 (see Table 6-5). The mass
transfer from the wastewater in the flow distribution region is
characterized by two phases, the liquid phase resistance and the
gas phase resistance. The overall mass transfer from this
two-resistance model, K, is a combination of the gas Kg and the
liquid mass transfer coefficients K^ . Kg is estimated by:
( V V78 ( diaV'11 f 0.0012 DvV67
K =0.00482 — — v
g I 100 J I 100 J 0.000181
k =000482 ilZ SOS (0.0012) (0.088)
g lOO 100 0.000181
k = 9.04 10~3 —
g sec
6-28
-------
is estimated by:
D l67
K, =2.61 10-7 1^-1 I =L
1 ' 100 j 0.0000085
where,
V = the wind velocity at 10 meters over the surface
(447 cm/s);
dia = the width of the region (505 cm);
Dg = the gas phase diuffusivity (0.088 cm2/s); and
D! = the liquid phase diffusivity (0.98 10~5 cm2/s).
100 0.0000085
5.73 10~6 —
sec
The overall mass transfer can then be written as follows:
:, 40.9 K K f
i g P
where,
K0 = the overall mass transfer coefficient based upon the
liquid concentrations (m/s);
K! = the liquid phase mass transfer coefficient
(5.736 lO'6 m/s) ;
Kg = the gas phase mass transfer coefficient (9.04 10~3 m/s);
fp = the fraction of the compound in the water phase, 0.4121;
and
K = the partition coefficient (0.0055 atm-m3/mol) .
1 1 Vl
(5.736) (10"b) (40.9) (9.04) (10~J) (0.00555) (0.412)
6-29
-------
K = 5.697 • 10~6 —
sec
The air losses, falri, from the two-resistance model are as
follows, using Equation 6-14:
where,
q
A
f airl
fat
rl
q
the overall mass transfer coefficient based upon the
liquid concentrations (5.697 10~6m/s);
the liquid flow rate (0.10 m3/s);
the surface area of the region (20 m2) ; and
the fraction emitted to the air in the entrance
region.
(5.697) (10'6) (20)
0.10
fairl = 0.001139
6.4.2.3 Region 2 calculations. Oil floats to the surface
of region 2, where it is removed. Since the refinery is assumed
to be operating without abnormal problems, the oil is a
relatively small fraction of the wastewater and the concentration
in the oil is assumed to be in equilibrium with the water. The
partition between the gas phase and the oil phase is estimated as
follows, using Equation 6-16:
where,
Vp = the vapor pressure of the pure component at the
surface temperature (95.2 mm Hg);
mwt = the molecular weight of the oil (180 g/g-mol); and
6-30
-------
= the oil density (0.7 g/cm3) .
eq
0.0012 V mwt
_ p _
p 28.8 760
= (0.0012) (95.2) (180)
eq " (0.7) (28.8) (760)
K = 0.00134
eq
The mass transfer from the thin floating oil layer on the
wastewater surface is characterized a resistance of only one
phase, the gas phase resistance. The overall mass transfer from
this one resistance model, K, is estimated as follows:
Ko = Kg Keq (6-27)
Ko = (0.00904) (0.00134)
Ko = 1.19 • 10~5 =SL
sec
where,
K0 = the overall mass transfer coefficient based upon the
liquid concentrations (m/s);
Kg = the gas phase mass transfer coefficient (identical to
region 1, m/s); and
Keq = the partition between the gas phase and the oil phase,
0.001342.
The fraction of the compound in the water phase is used to
correct the partitioning in the gas phase. The fraction of the
compound in the water phase is estimated from the octanol-water
partition coefficient as described above.
The air losses, falr2, from the one-resistance model are as
follows:
6-31
-------
[ K A owpc 1
fair2 = 1 - EXP - -2 - i_ (6-27)
I q )
1.19 • IP'5 (50) (144.2)
^ ~~ 1 HyVr
f ;„, = 0.576
where,
K0 = the overall mass transfer coefficient based upon the
liquid concentrations (1.19 10~5 m/s);
q = the liquid flow rate (0.10 m3/s) ;
A = the surface area of the region (50 m2) ; and
falr2 = the fraction emitted to the air in region 2 .
The concentration in the oil phase on the surface is assumed
to be in equilibrium with the concentration in the aqueous phase.
The exponential form of the estimation of falr prevents the
possibility of estimating air emissions that are in excess of the
total amount present.
Some of the component will be removed with the oil that is
removed from the surface. The fraction in the oil is estimated
with the following equation. Components removed with the oil are
not available for contributing to air emissions in region 3.
f0n = ( 1 - fairi ) ( 1 - fair2 ) f (6-29)
foil = ( 1 - 0.001139 ) ( 1 - 0.576 ) 0.588
f =0.25
6-32
-------
6.4.2.4 Weir calculations. In the case of benzene, the
deficit ratio is calculated using the following factors:
In (r) = 0.0785 Z131 qa428 h0310
In (r) = 0.0785 (0.2)01 (90)a428 (0.3)0310
In (r) = 0.04503
(6-30)
where the above constants are for the situation where Z < 1.2 m
and q < 235,
r =
Cs =
Co =
Z =
q =
h =
Cs/(Cs-Co) deficit ratio, assumes that there is no
oxygen before the weir,
saturated oxygen concentration,
outlet oxygen concentration,
distance of fall (0.2 m), includes 1.5 times the
distance from the weir top to the critical depth
above the weir,
flow rate per length of weir (90 m3/h-m), and
the tailwater depth (.3 m).
The natural log of the deficit ratio, ln(r), is calculated
as 0.20363. Next, the liquid phase mass transfer coefficient is
estimated:
m
_ q
D
°-67
hr
3600 s,
'6-31'
o.oos £ -*L °*
si 0.2 2.4
i 0.66
0.04503
hr
3600 s,
Next, the gas phase mass transfer coefficient of toluene is
estimated, based upon the reference mass transfer coefficient of
benzene.
6-33
-------
D ^ °-66
= 0.05 -SI I (6-32;
where,
Kg = the mass transfer coefficient of the volatile organic
in air (m/s) ;
Dgv = the diffusion coefficient of the volatile organic in
air (0.088 cm2/s) ;
Dgo = the diffusion coefficient of the reference material in
air (0.088 cm2/s); and
0.05 = the assumed mass transfer coefficient of a turbulent
surface .
0.05 = 0.05
s 0.088
The estimated gas phase mass transfer coefficient is
0.05 m/s. Next, the overall mass transfer coefficient is
calculated.
K =( -1 + I
0 [ Kj 40.9 Kg K
0.00241 = ' 1 1 '
0.003 40.9 0.05 0.00555 )
where,
K0 = the overall mass transfer coefficient based upon the
liquid concentrations (m/s);
K! = the liquid phase mass transfer coefficient
(0.00305 m/s);
Kg = the gas phase mass transfer coefficient (0.05 m/s); and
K = the partition coefficient (0.00555 atm-m3/mol) .
6-34
-------
The overall mass transfer coefficient is 0.00241 m/s. Next,
the fraction of air emissions is estimated.
f. = 1 - EXP
air
3600 sec
hr
where,
K0 = the overall mass transfer coefficient based upon the
liquid concentrations (0.000868 m/s);
q = the liquid flow rate per length of the weir
(90 mVh-m) ;
Z = the distance of fall (0.2 m) ; and
falr = the fraction of the volatile component emitted to the
air .
0019 = 1 -EXP- - °'2 36°°
90 hr
The fraction of benzene that is emitted to the air due to
the weir drop is 0.019.
J airl ~ JrEnt + ^ ~ JrEnt' J roil + ^ ~ JrEnt' ^ ~ J roil' ^ ~ J rinoil' Jrwf
fairl = 0.0012+( 1 -0.0012) .576+( 1-0.0012)( 1-0.576 )(l-.25) 0.019
falrl= 0.584
The overall fraction of benzene that is emitted from the API
separator is as follows:
frEnt = the fraction lost from the entrance region, 0.001139;
froll = the fraction lost from the separator region, 0.576;
frlnoll = the fraction removed by the recovered oil, 0.25; and
frwf = the fraction lost from the waterfall; 0.019.
6-35
-------
The overall loss of benzene from the API separator as air
emissions is estimated as 0.584. This does not include the
fraction recovered in the oil. Estimates of air emissions from
the recovered oil are not included in this unit.
6.5 MODEL FOR PRETREATMENT UNITS
In the entrance to a wastewater treatment plant, a
pretreatment unit can be used to remove solid objects, grit, or
other items that can be separated from the wastewater. The
following model is recommended for the estimation of air
emissions from a pretreatment unit.
This section provides the following:
• the equations used for the estimation, and
• an example calculation.
6.5.1 Pretreatment Equations
This section presents the model for the pretreatment unit
and illustrates the use of the model with a sample calculation.
The pretreatment unit has an agitated surface, and can contain
agitation by forced submerged air. The total air emissions are
the sum of the air emissions from the surface volatilization and
the submerged air flow.
If oil is present in the unit, some of the volatile
materials will partition into the oil and reduce the
concentrations in the water phase.
The mass transfer from the surface of the wastewater in the
pretreatment unit is characterized by the resistance of two
phases, the liquid phase resistance and the gas phase resistance.
The overall mass transfer from this two-resistance model, K, is a
combination of the gas and the liquid mass transfer coefficients
in Equation 6-10:
6-36
-------
where,
K0 = the overall mass transfer coefficient based upon the
liquid concentrations (m/s);
K! = the liquid phase mass transfer coefficient (m/s);
Kg = the gas phase mass transfer coefficient (m/s);
fp = the fraction of the compound in the water phase; and
K = the partition coefficient (atm-m3/mol) .
Kg is estimated by
v - 1UU I nnm 0.000462 US
JN. — U.UU1
g 24300 ' NSCH0-67
US = (6.1 + 0.0063 v)1
,0.5 V
100
Kj = 0.001
NSCH = —
n
where,
V = the wind velocity at 10 meters over the surface (cm/s);
and
Dg = the gas phase diffusivity (cm2/s).
The fraction of the compound in the water phase is used to
correct the partitioning in the gas phase. The fraction of the
compound in the water phase is estimated from the octanol-water
partition coefficient.
6-37
-------
OWPC oilfract
1 - oilfract
f = 1 - f
P ®
f = OWR
1 + OWR
where:
OWPC = the octanol water partition coefficient,
oilfract = the fraction of the waste that is oil and
insoluble in water,
OWR = the ratio of the amount in oil to the amount in
the water,
f0 = the fraction of the component in the oil phase,
and
fp = the fraction of the component in the water phase.
The air losses, falr, from the two-resistance model are as
follows, from Equation 6-14:
K A
.p _ i T^A/T» I o
airl
where,
K0 = the overall mass transfer coefficient based upon the
liquid concentrations (m/s);
q = the liquid flow rate (m3/s);
A = the surface area of the region (m2); and
falrl = the liquid flow rate (m3/s) .
6.5.2 Pretreatment Examples
It is assumed that an aqueous stream with no dispersed oil
is pretreated in a bar screen unit. The open-bar screen unit
6-38
-------
used for the example calculations was characterized as having a
turbulent region of 20 m2 . Since the concentration of benzene in
the wastewater is one part per million by weight. Properties of
benzene as well as the constants needed for the sample
calculations are presented in Tables 6-6 and 6-7.
The fraction of the compound in the water phase is used to
correct the partitioning in the gas phase. The fraction of the
compound in the water phase is estimated from the octanol-water
partition coefficient.
™™ _
UWK. -
OWPC oilfract
-
1 - oilfract
f =
OWR
1 + OWR
F = 0
f = 1 - f
where,
OWPC
= the octanol water partition coefficient;
oilfract = the fraction of the waste that is oil and insoluble
in water;
OWR
= the ratio of the amount in oil to the amount in
the water;
f0 = the fraction of the component in the oil phase; and
fp = the fraction of the component in the water phase.
The effective diameter of the region 1 surface is 20 m2 . The
mass transfer from the wastewater in the flow distribution region
is characterized by two phases, the liquid phase resistance and
the gas phase resistance. The overall mass transfer from this
6-39
-------
two-resistance model, K, is a combination of the gas Kg and the
liquid mass transfer coefficients K^ Kg is estimated by the
following calculations. (See TABLE 4-1) The Schmidt number, NSCH,
and the friction velocity U* are needed for the calculation of
Kg-
0.5
0.088
TABLE 6.6 PROPERTIES OF BENZENE USED FOR SAMPLE CALCULATIONS
Variable
Diffusivity in water
Diffusivity in air
Molecular weight
Henry's law constant
Diffusivity 02 in water
Vapor pressure benzene
Octanol water partition
coefficient
Symbol
Dl
Dv
Mwt
H
Do
vp
owpc
Number and units
0.98 x 10~5 cm2/s
.088 cm/s2
78 g/g-mol
0.00555 atm-m3/mol
2.5 x lO'5 cm2/s
95.2 mm Hg
141.25
6-40
-------
TABLE 6-7. UNIT PARAMETER NAMES AND SPECIFICATIONS FOR EXAMPLE
CALCULATIONS.
Unit Specification
Wastewater flow rate
Wind speed
Number of units
Temperature
Region 1 area
Oil in waste
Density of oil
Oil molecular weight
Symbol
q
V
n
T
fo
do
mwt
Variable Name
q
V
n%
T
area . enter
oilf ract
densoil
mwtoil
Value
0.10 m3/s
447 cm/s
1
25 deg. C
20 m2
0.00
0.7 g/cm3
180
6-41
-------
U* = ( 6.1 + 0.0063 447
o.s 447
100
IT = 13.35
0.000462 13.35
K = _ oooi
g 24300 i.7°-67
upon evaluation, the liquid mass transfer coefficient is
specified:
Kj = 0.001 (m/s)
Kg = 2.178 105 (g mol/cm2-s)
The overall mass transfer can then be written as follows, from
Equation 6-10:
K - (-L . l v
O ^ -rr
t 40.9 Ke Kfv
where,
K0 = the overall mass transfer coefficient based upon the
liquid concentrations (m/s);
K1 = the liquid phase mass transfer coefficient (0.001 m/s);
Kg = the gas phase mass transfer coefficient (2.187 10~3 m/s);
fp = the fraction of the compound in the water phase, 1; and
K = the partition coefficient (0.0055 atm-m3/mol) .
5.48 ID"4 — = ( -i
sec 0.001 (40.9) (2.187) (10~5) (0.00555) (1)
6-42
-------
6-43
-------
The air losses, fair, from the two-resistance model are as
follows from Equation 6-14:
K A
fairl = l
where,
K0 = the overall mass transfer coefficient based upon the
liquid concentrations (5.48 10~4m/s);
q = the liquid flow rate (0.10 m3/s);
A = the surface area of the region (20 m2) ; and
falrl = the fraction emitted to the air in the entrance
region.
(5.48) (1(T4) (20)
f _ i FXP ( .
Jain -1-W(- — )
fairl = 0.104
6.6 REFERENCES
1. GCA Corporation, Hazardous Waste TSDF Process Sampling.
EMB Report 85-HWS-3. October, 1985.
2. U.S. Environmental Protection Agency. Municipal
Operations Branch. Process Control Manual for Aerobic Biological
Wastewater Treatment Facilities. Publication No. EPA-430/9-77-
006. March 1977.
3. Treatment of Selected Internal Kraft Mill Wastes in a
Cooling Tower. Georgia Kraft Co. Rome Georgia. August 1971.
NTIS 208 217.
4. Nakasone, H. Study of Aeration at Weirs and Cascades.
Journal of Environmental Engineering, ASCE. 113:64. 1987.
6-44
-------
7.0 LAND TREATMENT
This chapter presents the approach used to estimate air
emissions from land treatment operations. A mathematical model
for diffusion in porous media with simultaneous sorption and
biodegradation is presented for estimating emissions from land
treatment operations. This model is also applicable to spills,
excavations of contaminated soils, solid waste transfer
operations, and other situations involving the diffusion of
volatile organics in porous media. Analytical models to estimate
the air emissions, representative values of land treatment model
input parameters, and example calculations are included.
7.1 NARRATIVE DESCRIPTION OF LAND TREATMENT AIR EMISSIONS
Land treatment is one of several land disposal methods used
for final disposition of hazardous wastes. At land treatment
facilities, wastes are either spread onto or injected into the
soil, after which they are normally tilled into the soil. Other
activities likely to occur at land treatment facilities include
storage of wastes in tanks or surface impoundments, loading and
unloading of wastes in vacuum trucks or dump trucks, and
dewatering of wastes using filtration devices. All of these
activities have emission points associated with them. The
following paragraphs describe analytical models used to estimate
emissions from the application, tilling, and final disposition of
hazardous waste at a land treatment disposal site. Emissions
from other land treatment activities, such as truck loading,
storage tanks, and fugitive emissions from transfer and handling
7-1
-------
operations, are estimated using procedures described in Chapter 9
of this report.
Estimating emissions from land treatment may involve one to
three independent steps depending on operating practices at a
land treatment site. If waste is applied from a vacuum truck to
the soil surface, allowed to remain on the surface for a period
of time, and then tilled into the soil, emissions are estimated
in three steps: (1) during application of waste onto the soil
from a vacuum truck, (2) after waste application and before
tilling, and (3) after tilling the waste into the soil. If waste
is applied to the soil surface and immediately tilled into the
soil, emissions are estimated in only two steps: (1) during
waste application, and (2) after tilling. If waste is applied by
subsurface injection and immediately tilled, only one step is
required to estimate emissions.
This section presents three separate analytical models that
can be used to estimate air emissions from separate land
treatment activities. Primary emphasis is given to the RTI land
treatment model that is used to calculate emissions from waste
that is mixed with the soil. This condition may exist when waste
has been applied to the soil surface and has seeped into the soil
or when waste has been injected beneath the soil surface or has
been tilled into the soil. The RTI land treatment model is
described below in Subsection 7.2.1, which includes separate
discussions of the following topics:
Subsection Topic
7.2.1.2 Biodegradation
7.2.1.4 Effective diffusivity
7.2.1.5 Waste partitioning
7.2.1.10 Model selection rationale
After waste is applied to the surface of the soil by spray
application, it seeps into the soil. While the waste is on the
surface, the concentrations at the surface can be approximated by
7-2
-------
the concentration in the waste. During this short period that
the waste covers the surface of the soil, the maximum short-term
emission rate is expected. For this special case, emissions are
estimated as the product of an overall mass transfer coefficient,
constituent concentration, and surface area of the land treatment
site. The model for estimating the mass transfer coefficient
from the surface to the wind was developed by McKay and Matsuga
and is briefly discussed in Subsection 7.2.3.
Emissions from a waste stream as it is applied onto the soil
surface from a vacuum truck, regardless of waste type, are also
calculated as the product of an overall mass transfer
coefficient, the surface area of the waste stream, and the
concentration of a specific constituent. Preliminary
calculations indicate that emissions from the spraying waste
application are relatively small and can be ignored in most
situations. Even so, a brief discussion of a model for
estimating these emissions is presented in Subsection 7.2.2, and
the model can be used if desired. Also included in this section
are Subsection 7.2.4, which discusses representative values of
input parameters for the analytical models, and Subsection 7.2.6,
which presents example calculations using each of the three
models presented.
At many existing land treatment sites, waste is applied onto
the soil from a vacuum truck and is allowed to remain for about
24 hours before being tilled into the soil. Under these
conditions, three separate calculations may be needed to estimate
air emissions. Emissions during waste application could be
estimated using the waste application model described in
Subsection 7.2.2; emissions after application but before tilling
would be estimated using the RTI land treatment model as
described in Subsection 7.2.1 (or, if a visible oil film exists
on the soil surface, the oil film surface model as presented in
Subsection 7.2.3); and emissions after tilling would be estimated
using the RTI land treatment model. At other existing sites,
7-3
-------
waste is injected into the soil using subsurface injection and is
immediately tilled. At these sites, only one calculation is
needed to estimate emissions. In this situation, the RTI land
treatment model would be used.
7.2 LAND TREATMENT
7.2.1 Land Treatment Emission Model Descriptions
7.2.1.1 Analytical Correlations. Emissions from land
treatment after waste is applied to the soil are estimated using
a model developed by Clark Allen of Research Triangle Institute
(the RTI model). This model assumes that emissions from the
surface of the soil/waste mixture are limited by the diffusion of
vapors through the pore spaces in the soil/waste mixture and
further assumes that an equilibrium concentration of organic
vapors exists at all times within the pore spaces. The model is
based on Fick's second law of diffusion applied to a flat slab as
described by Crank and includes a term to estimate biological
decay assuming a decay rate that is first order with respect to
waste loading in the soil.
The solution to the diffusion equation developed by Crank is
for diffusion out of a slab that initially has a uniform
concentration of diffusing material throughout and that has equal
concentrations of diffusing material at each surface.
The general solution to the diffusion equation for those
conditions, as presented by Crank, is:
F = ^=1_E°° 8_- exp* -D (2n+l)2.2t, , (7_1}
o n=o
where,
F = fraction of initially applied material that has
diffused out of the slab at time t;
M, = mass of material that has diffused out of the slab at
time t;
M = initial mass of material present;
7-4
-------
D = diffusion coefficient;
1 = distance from center to surface of slab; and
t = time after initial distribution of diffusing material
into the slab.
This series solution converges very slowly for small values of
2
time (i.e., Dt/1 immediately after waste application or
tilling) , Crank presented an alternative solution that is valid
during this short time. The following equation is obtained from
the alternative solution for Dt/1 <.213:
Equation (7-2) a^roxiiria/tes t\h<3 Crank solution but excludes a
y-i _ t _ I -Lsl I / "~7 O \
small error funcTTbn coJ:p:e"cJipn used by Crank. (i~^i
m0 yij \ ' /
To verify the validity of Crank's solution for short times
and to test the accuracy of an approximation for use over longer
times, the values predicted by the solution for short times and
the values obtained using the first term of the series solution
[Equation (7-3)] are compared to the values obtained using the
first three terms of the series solution. Table 7-1 presents the
results for a range of values of the dimensionless parameter,
Dt/12:
F , = 1 - exp - | ,7-3;
M T?
Table 7-1 shows that, for values of the dimensionless
parameter greater than 0.213, the first term of the series
solution, Equation (7-3), can be used to estimate total
emissions. The table also shows that the solution for short
7-5
-------
TABLE 7-1. COMPARISON OF THE ESTIMATED FRACTION
EMITTED USING THREE DIFFERENT EQUATIONS
(INTEGRATED FLUX FROM SOIL)
Time
parameter
(Dt/12)
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
.000
.025
.050
.075
.100
.125
.150
.175
.200
.213
.250
.275
.300
.325
.350
.375
.400
.425
.450
.475
.500
.525
.550
.575
.600
.625
.650
.675
.700
.725
.750
.775
.800
.825
.850
.875
.900
.925
.950
.975
Short-term
solution
Dt ^1/2
*
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
1
1
1
1
1
1
1
1
. I2*
.000
.178
.252
.309
.357
.399
.437
.472
.505
.521
.564
.592
.618
.643
.668
.691
.714
.736
.757
.778
.798
.818
.837
.856
.874
.892
.910
.927
.944
.961
.977
.993
.009
.025
.040
.056
.070
.085
.100
.114
Fi
ser
1 -
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
rst term of
•ies solutio
8 * Df
- exp*-
41
.189
.238
.284
.326
.367
.405
.440
.474
.505
.521
.562
.589
.613
.636
.658
.679
.698
.716
.733
.749
.764
.778
.791
.804
.816
.827
.837
.847
.856
.864
.873
.880
.887
.894
.900
.906
.912
.917
.922
.927
n F
thre
2 of s
* sol
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
'irst
e terms
eries
ution
.067
.179
.252
.309
.357
.399
.437
.472
.504
.520
.562
.589
.613
.636
.658
.679
.698
.716
.733
.749
.764
.778
.791
.803
.816
.827
.837
.847
.856
.865
.873
.880
.887
.894
.900
.906
.912
.917
.922
.927
7-6
-------
times, Equation (7-2), is valid for values of the dimensionless
parameter below 0.213. Equations (7-2) and (7-3) give identical
results for a parameter value of 0.213. This comparison
indicates that sufficient accuracy can be attained under all
conditions if the equation for short times is used for values of
the dimensionless parameter below 0.213 and the first term of the
general solution is used for values above 0.213. It is observed
that the fraction of material that diffuses out of the slab is
linear with respect to the square root of time up to the point
where approximately 50 percent of the diffusing material is
lost.
The conditions defined for the above solutions by Crank
are analogous to diffusion of volatile organics out of a surface
layer of a soil/waste mixture as happens in land treatment
operations. Because of the symmetry of conditions on which the
above solutions are based, an impenetrable plane could, in
theory, be inserted at the midpoint of the slab without changing
the solution. One-half of the slab with an impenetrable boundary
layer on the bottom would represent the surface layer of soil
into which waste is mixed during land treatment.
In a land treatment operation, the applied material
partitions into several phases including evaporation into a vapor
phase, adsorption onto soil particles, and absorption into oil
and water in the soil/waste mixture. Only the vapor phase is
available for diffusion out of the soil/ waste mixture.
Therefore, to apply the above equations to land treatment, the
amount of material in the vapor phase must be known. The amount
of material that partitions into the vapor phase can be estimated
by calculating equilibrium conditions within the soil/waste
mixture. This equilibrium is estimated by defining Keq, the
ratio of the amount of organics in the vapor phase to the total
amount of organics in the soil/waste mixture. The instantaneous
emission rate, E, at any time, t, can be estimated by the
following Equations (7-4) and (7-5), which are obtained by
7-7
-------
differentiating Equations (7-2) and (7-3) and adding the
equilibrium constant, Keq, and a term to account for
wastebiodegradation, (-t/t, ):
(short times) E = -— * * e b , and (7-4|
-Keq D • t* -t/t
longer times) E = M *—^* exp*—^ ^* e b. (7-5
where t, = the biological degradation time constant.
The above equations account for the removal of organic
material from the soil/waste mixture both biological degradation
and air emissions. In a land treatment operation, the primary
objective is to dispose of organic materials by biodegradation;
thus, significant quantities of waste would be expected to be
depleted from the soil by biological degradation. Other
mechanisms of removal such as leaching and photolysis also are
possible but are not accounted for in this model because of the
estimated small amount of materials lost by these processes.
7.2.1.2 Biodegradation. Biodegradation at land treatment
sites is generally considered to be a first-order process with
respect to waste concentration in the soil up to the point where
saturation is achieved. In addition to literature sources that
make such statements, comments on a draft of this document
provided by Chevron Research Company offer further evidence of
the first-order nature of biodegradation at land treatment
pj
5
4
sites. A first-order decay process is defined in the literature
as having the following form:
dM ... ._
dt = - KbM '
where,
M = mass of organic material in the soil; and
K, = biological decay constant.
7-8
-------
Integrating and using the boundary conditions M = M at t = 0
results in:
dM , ,,
log M = -Kfat + C1
or
-V
M = C2e
where C, and Cp are constants of integration. Substituting the
boundary conditions gives:
-K,t
M = M e
o
K, has units of s and can be expressed as the reciprocal
of the biological decay time constant, l/t-,. The exponential was
introduced directly into the rate relationship, Equations (7-4)
and (7-5), to reduce the amount of material available for air
emissions by the fraction of material removed by biooxidation.
7.2.1.3 Estimation of equilibrium coefficient, Keg.
Partitioning of volatile constituents in the waste is assumed to
occur between the vapor space in the soil/waste mixture,
adsorbent solids in the soil, and absorbent liquids in the soil
3
and waste. Using 1 cm of the soil/waste mixture as a basis for
calculation, the total volume of gas (i.e., void space) in the
cubic centimeter is described by the air porosity, • . Using the
3 a
ideal gas law, the number of moles of gas in 1 cm of the
soil/waste mixture is P* / (RT) , where P is the pressure of a
a *
constituent in the gas phase and is usually equal to XP (X is
*
the mole fraction of the constituent in the liquid phase and P
is the pure component vapor pressure) . The moles of constituent
3
in the gas phase in 1 cm of the soil/waste mixture is thus
XP • / (RT) . Oil loading in the soil/waste mixture in units of
a 3
grams of oil per cubic centimeter of mixture is L (g .-./cm
mixture) , and the total moles of constituent per cubic centimeter
7-9
-------
of the mixture is XL/MW .,. The equilibrium coefficient, Keq, is
defined as the moles of constituent in the gas phase per unit
volume of the soil/waste mixture divided by the total moles of
constituent per unit volume of the soil/waste mixture.
Therefore, the following equation can be written:
XP*» /(RT) P*MW . •
= a = oil a
q ~ X L/MW . n ~ RTL
oil
This equation differs from the usual equation for equilibrium
coefficient by the factor • , which is included to account for
a
the limited air space available within the soil/waste mixture.
The ratio of moles per mole and grams per gram can be used
interchangeably in this equation. The value of Keq can be
calculated from measurements, if available, of constituent con-
centrations in the pore space and in the soil/waste mixture.
In a similar manner, it can be shown that Keq can be
estimated for aqueous wastes with an assumed value of the Henry's
law constant, H :
H 106
,, c
Keq =
RT» ,
waste
where
• , = the volume fraction of the soil/waste mixture that
waste . •!! j_
is occupied by waste.
In the above equations, it is assumed that equilibrium is
controlled by Raoult's law for oily wastes and by Henry's law for
aqueous wastes. An aqueous waste is assumed to contain water and
organic constituents that are dissolved in water. An example of
an aqueous waste is a sludge containing 10 percent solids,
5 percent acetone, 1 percent methanol, 500 ppmw benzene, and the
remainder water. If the waste contains oil mixed with the water,
or the waste contains volatile constituents at concentrations
greater than the solubility in water, it is modeled as an oily
7-10
-------
waste. It is important to specify the molecular weight of this
separate organic phase for this Raoult's law calculation of Keq.
7.2.1.4 Estimation of effective diffusivity. The
diffusivity of specific compounds, as reported in the literature,
assumes that the diffusion occurs in free air. In a land
treatment operation, diffusion of vapors out of the soil must
take place within the confines of the air-filled voids within the
soil. This characteristic of soil is referred to as the air
porosity. The ratio of effective diffusivity of a constituent in
the soil to its diffusivity in air can be described by the
following equation:
D -10/3
e a
D 2
•T
where,
D = effective diffusivity of constituent in soil;
D = diffusivity of constituent in air;
a
• = air porosity of soil; and
a
•
„ = total porosity of soil.
When air porosity and total porosity are the same (i.e., for
dry soil), this equation reduces to:
D^ . 4/3
D a
a
Total porosity refers to the fraction of the land treatment
medium that is made up of nonsoil (or nonsolid) materials, i.e.,
the sum of the void space, water-filled space, and space occupied
by the oil in the applied waste.
Soil air porosity undergoes substantial changes over time as
soil dries out and when moisture is added by rainfall or by
watering. As a result, accurately accounting for soil porosity
in an analytical model is difficult. The use of average or
7-11
-------
typical values of soil porosity may be the most practical
approach.
7.2.1.5 Waste partitioning. A large percentage of wastes
that are disposed of by land treatment are refinery sludges.
These wastes are mostly sludge emulsions and consist of varying
fractions of water, oil, and inorganic solids, where oil
represents the total organic portion of the waste including
volatile compounds. A much smaller amount of land-treated wastes
are dilute aqueous solutions of water and organic compounds.
When wastes are applied to a land treatment area, volatile
materials in the soil have the potential for partitioning into
four different phases--a vapor phase, an oil phase where volatile
material is dissolved in the oil, a water phase where volatile
material is absorbed in the soil moisture, and a soil phase where
volatile material is adsorbed by organic carbon within the soil.
For oily wastes, VO compounds will preferentially dissolve in oil
rather than water so that the fraction of volatile materials in
the water phase is estimated to be very small. Partitioning of
volatile materials into the soil phase by adsorption is a
function of the amount of organic carbon in the soil. It is also
estimated to account for only a small fraction of the applied
organics because the surfaces in the soil are expected to contain
oil from the application and tilling of waste materials that
contain oil. This oil in the soil is expected to both absorb the
volatile constituents and to interfere with the relatively lower
adsorption rates on soil surfaces. For high molecular weight
constituents present in aqueous wastes, adsorption may be more
important. An equilibrium equation can be written that takes all
four phases into account in the estimation of equilibrium vapor
concentration in the soil. However, as presented here, the
equilibrium equation in the RTI model includes only two phases.
Calculations by one researcher looked at the difference in
estimated emissions using two-phase partitioning of waste into an
oil phase and vapor phase and using four-phase partitioning. The
7-12
-------
results of these comparisons are given in Table 7-2 and show for
the conditions considered that, for soils having an organic
carbon content of up to 10 percent, the estimated fraction of
applied organics emitted using four-phase partitioning is only
about 10 percent less than the estimated fraction emitted using
two-phase partitioning. In a given situation, the amount of
material adsorbed by organic carbon in the soil is relatively
constant; thus, in soils with high organic carbon content,
adsorption of materials in the soil may become more significant
if low loading rates of oil or aqueous wastes are used. One of
the products of biodegradation is organic carbon; thus, land
treatment sites that have been active for an extended time may
have elevated concentrations of organic carbon. Even so, with
the normal oil loading used in land treatment, it is
likely that a large fraction of the available adsorption sites
would be occupied by the oil itself, thus limiting the effects of
adsorption on emissions of the lighter constituents.
For oily sludges, Keq is calculated using vapor pressure and
waste loading is calculated exclusive of water content. For
dilute aqueous waste, partitioning is estimated to be in a water
phase and a vapor phase, and the parameter Keq is calculated
using Henry's law constant; waste loading is calculated using the
total waste applied. Keq may be calculated from site-specific
land treatment soil, vapor, and solids analyses if available.
Table 7-3 summarizes the equations that make up the RTI land
treatment model.
7.2.1.6 Emissions at Short Times.
When a sludge containing volatile organics is applied onto or
tilled into the soil at a land treatment site, the maximum rate
of air emissions will occur immediately after application or
tilling. Volatile organics will leave the surface and enter the
environment through wind currents. Although the RTI model is
based on the premise that emissions from land treatment are
limited by vapor diffusion through the soil, the maximum rate of
air emissions immediately
7-13
-------
TABLE 7-2. EMISSION ESTIMATES USING TWO DIFFERENT EQUATIONS
FOR THE VAPOR-SOIL PARTITION COEFFICIENT7
Organic carbon
content of soil
(fraction)
Estimated emission
fraction--two-phase
partitioning
Estimated emission
fraction--four-phase
partitioning
0
0.001
0.010
0.100
0.622
0.622
0.622
0.622
0.622
0.621
0.614
0.559
7-14
-------
TABLE 7-3. RTI MODEL FOR LAND TREATMENT EMISSIONS
Emission rate equations
Short-term solution (Kv t < .22}
ZJ1 ^
" ~T
K kr
ea (7
Long-term solution (Kv t >_ 0.22
2 K D
E = M
eq e
exp
K D 7i2 t t
eq e I
4 I2 t.
Fraction air emissions
Short-term solution (Kv t <0.22;
j—i e^ e
at ~ ~^~
i I (
Long-term solution (Kv t > 0.22^
2 ,-li
3 /.
-i
71
r, t,
d b i
1 - exp
Long-term solution (Fa <0.33 and Kvtb
exp
.1878
Very long-term solution (t -> °°) (K^t, > 0.62
0.811 K t
Fa - (Kd tb + 1^ °'187
Very long-term solution (t -> °°) (K,t,< 0.62)
[ continued)
7-15
-------
TABLE 7-3 (continued)
Keq =
H
P*MW . ..
oil a
RT
Keq = — (106;
RT
L
a
(used for oily sludges)
(used for dilute aqueous waste]
waste
k = 4.82 (10-3) U°'78 Sc
M =
o
K =
v
Kd =
D
a a
L1C
KeqD
1
2
(volatilization constant]
2
If both air porosity and total porosity are known:
D = D
e a
De = (4A/-
D = D
e a
4/3
a
(if only air porosity is known)
4.83 (107)
B
W f il W
L = — (for oily sludges); L = (for dilute aqueous waste]
Equilibrium equations are adjusted to account for volume
fractions of air and waste within the soil. (continued)
7-16
-------
TABLE 7-3 (continued)
Variable
Keq
Definition
Data source
Equilibrium coefficient of constituent Calculated
in the soil (dimensionless)
Gas-phase mass transfer
coefficient(cm/sec)
Calculated
C
Mt
MWr
MW
R
Concentration (weight fraction) of
constituent in the oil phase or (for
dilute aqueous waste)in water
Diffusion^coefficient of constituent
in air,cm /s
Effective diffusion coefficient of
constituent in the soil, cm /s
2
Emission rate of constituent, g/cm /s
Henry'.s law constant for constituent,
atm»cm /g mol
Depth to which waste is mixed in the
soil, cm
Oil or aqueous waste loading
in the soil,g/cm3
Air emissions of constituent from the
soil, g/cm
Initial loading of constituent on the
land treatment site, g/cm
Average molecular weight of the oil,
g/g mol oil
Molecular weight of constituent,
g/g mol
Pure component vapor pressure of
constituent, atm
Ideal gas constant,
82.1 atm»cm /g mol»K
Definition
Data base
Calculated
Calculated
Data base
Facility
operation
Calculated
from annual
throughput
Calculated
Calculated
Estimated
Data base
Data base
Literature
[ continued)
7-17
-------
TABLE 7-3 (continued)
Variable
K,,
-oil
u
Definition
Time constant for biological decay of
constituent, sb
Time after waste application to the
land treatment site, s
Temperature of vapor in soil, K
Volume fraction of air-filled voids
in the soil (soil air-filled porosity)
(dimensionless)
Total porosity of the soil (equivalent
to dry basis bulk density divided by
soil particle density) (dimensionless)
Volume fraction of waste in the soil
waste (dimensionless)
Biorate of constituent, mg V0/g»h
Volatilization constant for constituent,
s
Modified volatilization constant, s
-1
Biodegradation constant for constituent,
s
Fraction of constituent emitted to the
air after a long time
Fraction of constituent emitted to the
air at time t
Fraction by weight of applied waste that
is oil (organic)
Windspeed, m/s
Data source
Literature,
or site
specific
Facility
operation
Assumed
Estimated
from litera-
ture data
Estimated
Calculated
Data base
Calculated
Calculated
Calculated
Calculated
Calculated
Definition
Estimated
Time constant is the time required for (continued)
biodegradation of 63.2 percent of a pollutant.
7-18
-------
TABLE 7-3 (continued)
Variable Definition Data source
W Total waste applied to land treatment
site, g Definition
A Area of land treatment site to which
waste is applied, cm Definition
(m in calculation of de)
Sc« Schmidt number (gas phase) Calculated
(j
de Effective diameter of land treatment
area,m Calculated
ua Viscosity of air, g/cm»s Literature
3
• a Density of air, g/cm Literature
7-19
-------
after application or tilling will be limited by the gas-phase
mass transfer coefficient, k . Within a few hours after
application or tilling, the rate of air emissions from the
volatile components will be substantially less than the maximum
rate because the volatiles at the surface have been removed by
the wind and the remaining volatiles must diffuse up through a
layer of porous solids, a relatively slow process.
The land treatment model can be used for short-term emissions.
The emissions from the short-term use of the land treatment model
will be somewhat less than the oil-film model, although the
initial rate from both of these models is equivalent. The oil
film model is used to estimate maximum emission rates and the
land treatment model is used to account for surface drying during
short-term emission estimations. The equation for the emission
rate immediately after application or tilling is:
exp| - —I
(7-6)
ZJ1 ^
/
1
a _,_
kG Keq \
71 t
DeKeq
T
\ b
The basis of the above equation is a resistance in series model
where the resistance (inverse of the mass transfer coefficient)
is the sum of the resistance of the soil and the resistance at
the wind-porous solid interface. The mass transfer coefficient
of the soil is defined in
1/2
Equation (7-4) in the term: (Keg D)
U •
The resistance of the soil to mass transfer is the inverse of
the above or (Keq D / »t ) ~1/2 . The resistance at the air-soil
interface is defined by I/ Keq ]<:„ . Because Keq has previously
(j
been defined as containing a factor to account for soil porosity,
this factor (soil porosity) must be included in the above
equation to maintain a consistent definition of Keq throughout
this discussion. The revised resistance is represented by
• a/Keq ]<:„ . Summing the two resistances and substituting into
7-20
-------
Equation (7-4) gives Equation (7-6) . The gas-phase mass transfer
coefficient, ]<:„, is calculated as described in Table 4-1 for a
(j
surface impoundment.
7.2.1.7 Estimating the fraction emitted at short times.
The fraction of a constituent emitted to the air after some time,
t, can be estimated by integrating the equation for air emissions
from time 0 to time t:
I/O f-
* Keq D ' C^ / -t/t
F = * - ^ * J t 1/Z e °dt .
at • r o
The exponential term can be replaced by a series,
-t/tb 1 _ t_ , 1* t2* __ 1« t\ 1* tj
e -1 j_ + 2* t * 6* t * 24* t * * * * '
r b rb rb r b
which can be substituted into the above integral, and each of the
individual terms integrated. The results of these integrations
are :
This series solution converges with only a few terms for values
of t/t1 glesE ^rgfaare 1 . 2 yh^erefpr^el %rie f-oilfcwinql sii^mi_f icatjion can
f .7 \ t \ i n \ t \ 49 \ t \ I
rofc '^JuitTre^ 'fe(/L . e . , lj_r*tzegrate$
f / 2 .7
be used to ksfamatie the fvpactr
emissions) at short times:
1/2
* - u y. 1/0* 11-*
p = * - E. * 2t * 1 - — — * (7-7)
-L-o- dc O* ^u * Q-t-** \ ' ' J
at . 12 j tb
The resistance to emissions presented by gas-phase mass transfer
at the soil surface is only considered important for the
estimation of the emission rates immediately after application or
tilling. This resistance is omitted in the above equation with
little loss in accuracy.
The above equation is used to predict the fraction of a
constituent emitted to the air when K t is less than 0.22.
7.2.1.8 Estimating the fraction emitted for longer times.
For longer times, when most of the constituent is not present in
the soil, the short-term solution (Equations (7-4) and (7-7))
7-21
-------
will overestimate air emissions. Under these conditions,
Equation (7-5) can be integrated to estimate the fraction removed
by volatilization. Equation (7-5) can be simplified by
Keq D •2
defining the constant, K,, as ^ :
a 4 1z
M 8K,
E = °2 exp * - Kdt - t/tb * . (7-8)
Integrating from time 0 to t gives:
R 1 ~l
Fat= ^2 d+ jTt ) * 1 ~ SXP (" Kdt ~ t7tb}* + °-1878- <7~9)
In the above equation, terms after the first (n > 0) in the
series solution are replaced by the constant 0.1878. This
equation is used for estimating air emissions when K t is greater
than or equal to 0.22.
When K t is less than 0.22, the following simplification can
be used to estimate air emissions at longer times. An
exponential decay factor is established to relate the fraction
emitted at any time, t, to the fraction emitted at very long
times (i.e., t -> °°) as estimated using Equation (7-12), which
follows. The resulting equation is:
Fat = Fa [1 ~ exp ("Kdfc ~ t7tb)
where
F = fraction of constituent emitted at very long times (t -> °°) .
a
For very long times (i.e., t -> °°) , the fraction emitted can
be estimated using the following procedure. The integrated form
of the general solution without dropping terms is:
1 - exp{-(2n+?) K,t -
(2n+l)2
7-22
-------
This equation can be simplified using the following rationale:
For large values of t, the exponential terms are negligibly
small, and for large values of n, l/(t,K,) becomes negligibly
small compared to (2n+l) . If these conditions are true for all
terms where n > 0, the simplified equation is:
8 * Kd^~b
^ ^
_ _
9 * K t-
Kdtb
n
U-
The value of 0.2317 was obtained by evaluating the first 125
terms of the series for n > 0 with negligibly small values of
l/(tbKd):
125
- - - o = 0.2317 .
n=l
Combining terms and simplifying, the equation becomes:
0.81057K ,t,
Fa = ~K~t + 1 0.1878. (7-11
The assumptions used in developing Equation (7-11) are not
valid for small values of K ,t, (K^t, approximated by the
following relationship:
F =
\
(7-12'
This relationship was established by using multiple terms of the
general solution to calculate values of F for a series of input
2 a
values for the parameters KeqD /I , which is identified as the
volatilization constant, K , and t, and then using a curve-
fitting routine to derive the relationship in Equation (7-12) for
KdV
Table 7-4 presents the results of calculations of the long-
term fraction emitted (i.e., t -» °°) using 100 terms of the
general solution and inputting several values of the
7-23
-------
dimensionless ratio, K t, , designated as T. This ratio is an
indicator of the relative rates of volatilization and
degradation. Table 7-4 also shows the results if the above
approximating equations are used to calculate the long-term
fraction emitted, and it shows good agreement between these
results and the results obtained by the general solution.
Table 7-5 shows a comparison of the estimated emission
fractions for a range of values of K t and t/t, using the first
100 terms of the general solution and using the approximations
given in Equations (7-7) and (7-9) . This table shows good
agreement between the approximating equations and the rigorous
solution.
To calculate the amount of waste remaining in the soil at
any time, it is necessary to know both the amount emitted to the
air and the amount biodegraded. At very long times (i.e., t ->
°°) , all waste is assumed to disappear from the soil. Thus, the
fraction of waste emitted plus the fraction biodegraded must be
equal to 1 if other mechanisms of removal are ignored.
Therefore, at very long times:
Fa , (7-13)
where
Fb = fraction of constituent that is biodegraded after a
long time (i.e., t -> °°) .
7.2.1.9 Tilling . To apply the model to a situation where
the land treatment plot is retilled after the initial waste
application and tilling, estimates of the amount of waste emitted
7-24
-------
TABLE 7-4. ESTIMATED AIR EMISSION FRACTION AT LONG TIMES
Value of T
*T = K t *
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
2
.050
.100
.150
.200
.250
.300
.350
.400
.450
.500
.550
.600
.650
.700
.750
.800
.850
.900
.950
.000
.050
.100
.150
.200
.250
.300
.350
.400
.450
.500
.550
.600
.650
.700
.750
.800
.850
.900
.950
.000
Estimated
fraction
(rigorous equation)
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
.222
.313
.381
.435
.480
.518
.551
.579
.604
.626
.646
.664
.680
.694
.708
.720
.731
.741
.751
.760
.768
.776
.783
.789
.796
.802
.807
.813
.818
.822
.827
.831
.835
.839
.843
.846
.850
.853
.856
.859
Estimated
fraction
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
.224
.316
.387
.447
.500
.548
.592
.632
.671
.707
.742
.775
.806
.837
.866
.894
.922
.949
.975
.000
.025
.049
.072
.095
.118
.140
.162
.183
.204
.225
.245
.265
.285
.304
.323
.342
.360
.378
.396
.414
Estimated ,
fraction
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
.277
.348
.407
.456
.497
.533
.563
.590
.614
.635
.654
.672
.687
.701
.714
.725
.737
.747
.750
.765
.773
.780
.787
.794
.800
.805
.811
.816
.821
.826
.830
.834
.839
.842
.846
.849
.853
.856
.859
.862
a F - T°'5
a ba - l
, 0.81057 K, t,
F = rr-r -i + 0.187!
a K ,tn + 1
7-25
-------
TABLE 7-5. RIGOROUS VS. APPROXIMATE ESTIMATES
OF EMISSION FRACTIONS
t/t, K t
b v
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
.01
.01
.01
.01
.01
.01
.01
.01
.01
.01
.01
.01
.01
.01
.01
.01
.01
.01
.01
.01
.10
.10
.10
.10
.10
.10
.10
.10
.10
.10
.10
.10
.10
.10
.10
.10
.10
.10
.10
.10
.30
.30
.30
.30
.30
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
1
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
1
0
0
0
0
0
.05
.10
.15
.20
.25
.30
.35
.40
.45
.50
.55
.60
.65
.70
.75
.80
.85
.90
.95
.00
.05
.10
.15
.20
.25
.30
.35
.40
.45
.50
.55
.60
.65
.70
.75
.80
.85
.90
.95
.00
.05
.10
.15
.20
.25
K t,
vb
5.
10.
15.
20.
25.
30.
35.
40.
45.
50.
55.
60.
65.
70.
75.
80.
85.
90.
95.
100.
0.
1.
1.
2.
2.
3.
3.
4.
4.
5.
5.
6.
6.
7.
7.
8.
8.
9.
9.
10.
0.
0.
0.
0.
0.
00
00
00
00
00
00
00
00
00
00
00
00
00
00
00
00
00
00
00
00
50
00
50
00
50
00
50
00
50
00
50
00
50
00
50
00
50
00
50
00
17
33
50
67
83
Estimated Estimated Estimated
fraction fraction fraction ^
(rigorous) (approximated by) (approximated by)
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
.25
.35
.43
.50
.56
.61
.65
.69
.73
.76
.79
.81
.83
.85
.87
.88
.90
.91
.92
.93
.24
.34
.42
.49
.54
.59
.64
.67
.71
.74
.77
.79
.81
.83
.85
.86
.87
.89
.90
.91
.23
.32
.39
.46
.51
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
.25
.36
.44
.50
.56
.62
.67
.71
.75
.79
.24
.34
.42
.49
.55
.60
.65
.69
.73
.77
.23
.32
.39
.45
.51
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
.28
.36
.44
.50
.56
.61
.65
.69
.73
.76
.79
.81
.83
.85
.87
.88
.90
.91
.92
.93
.28
.36
.43
.49
.54
.59
.64
.68
.71
.74
.77
.79
.81
.83
.85
.86
.88
.89
.90
.91
.27
.34
.41
.46
.52
See notes at end of table. (continued)
7-26
-------
TABLE 7-5 (continued)
Estimated
fraction
t/t, K t
b v
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
a
.30
.30
.30
.30
.30
.30
.30
.30
.30
.30
.30
.30
.30
.30
.30
.00
.00
.00
.00
.00
.00
.00
.00
.00
.00
.00
.00
.00
.00
.00
.00
.00
.00
.00
.00
0
0
0
0
0
0
0
0
0
0
0
0
0
0
1
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
1
.30
.35
.40
.45
.50
.55
.60
.65
.70
.75
.80
.85
.90
.95
.00
.05
.10
.15
.20
.25
.30
.35
.40
.45
.50
.55
.60
.65
.70
.75
.80
.85
.90
.95
.00
Approximated
Approximated
K t
V
1.
1.
1.
1.
1.
1.
2.
2.
2.
2.
2.
2.
3.
3.
3.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
1.
by:
by:
, (rigorous)
00
17
33
50
67
83
00
17
33
50
67
83
00
17
33
05
10
15
20
25
30
35
40
45
50
55
60
65
70
75
80
85
90
95
00
Fat =
F - 8
Fat 2
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
0.
56
60
63
67
70
72
75
77
79
80
82
83
84
85
86
19
26
32
37
42
46
49
52
55
58
60
62
64
66
68
69
71
72
73
74
1.128 V
*1 +
)# -"-T
!
K ,t
Estimated
fraction
Estimated
fraction
o l-\
(approximated by) (approximated by)
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
'>yvt (1-
-1
- | jl- e]
"U
.56
.60
.64
.68
.72
.17
.24
.29
.34
.38
.41
.44
.48
.50
.53
1/3 t/tb)
^r># V +-
-------
to the air and the amount biodegraded are required. When
retilling occurs, the amount of material remaining in the soil at
the time of retilling is estimated using the following equation:
~, ^ -t/t,
-b , (7-14)
o cL L
where,
F = fraction of constituent remaining in the soil; and
O
F' = fraction of material emitted to the air at time t
at assuming no biodegradation (F? , can be estimated by
setting t/t, = 0 in Equation (7-7) or (7-9), whichever
is appropriate).
To continue modeling emissions after retilling occurs, M0 is set
equal to Fs M0 and t is reset to zero. If a reapplication of
waste occurs, the total waste loading is the sum of the waste
remaining in the soil and the newly applied waste:
M = F0M + M , (7-15)
o S o n ' ^ '
where Mn = amount of constituent newly applied to the land
treatment site. To continue the modeling after waste
reapplication and tilling, t is reset to zero.
7.2.1.10 Model selection.
The RTI model was selected for use in this regulatory effort
after a review of three models of land treatment emissions. The
models reviewed were the RTI model, the Thibodeaux-Hwang model,
and a model developed by EPA's Office of Research and Development
located in Ada, Oklahoma (the Ada model). The review considered
three selection criteria: technical basis, representativeness,
and availability of inputs. The Ada model is the most ambitious
of the three in attempting to account for mechanisms of pollutant
removal other than air emissions and biodegradation. However,
that model requires detailed site-specific model inputs that may
not be available or reasonably estimated. Because of these
characteristics of the Ada model, it was not considered
appropriate for use in the current effort. Both the Thibodeaux-
Hwang and the RTI models have input requirements that are
reasonably available, both have been compared with available
7-28
-------
measured data, and both have shown reasonable agreement with the
Q
measurements. Apparently, either of these two models is
satisfactory as a means of estimating emission rates at specific
times for some organic compounds. However, if the Thibodeaux-
Hwang model is used to estimate long-term, steady-state
emissions, it would predict that all of the applied volatile
organics are emitted because it does not account for biodegrada-
tion. Such a prediction would contradict data obtained from
laboratory and field studies that indicate biodegradation of some
9
organic compounds in land treatment applications. The RTI
model, in contrast, estimates biodegradation of individual
compounds based on constituent-specific biodegradation rates.
The RTI and the Thibodeaux Hwang models predict similar emission
rates for initial volatile losses in the absence of
biodegradation. Thus, the results of the RTI model show varying
levels of biodegradation when used to evaluate the fate of
different organic compounds.
In summary, the Ada model has had limited public review,
accounts for multiple waste removal mechanisms, requires numerous
detailed model inputs, and has no published comparisons of
estimated and measured emissions. The Thibodeaux-Hwang model has
been publicly reviewed, accounts for one major waste removal
mechanism (volatilization), requires reasonably available model
inputs, and there are published comparisons of measured and
estimated emissions. The RTI model has had limited public
review, accounts for the two major waste removal mechanisms
(volatilization and biodegradation), requires reasonably
available model inputs, and there are published comparisons of
measured and estimated emissions. The peer review, emission
comparisons, and data availability are compelling reasons for
using the Thibodeaux-Hwang model. However, the absence of
biodegradation in that model is considered a major shortcoming
because of the importance attached to this removal mechanism by
industry personnel and many independent researchers. Emission
comparisons and data availability for the RTI model are roughly
7-29
-------
equivalent to those of the Thibodeaux-Hwang model, and the RTI
model includes terms that account for biodegradation. The RTI
and the Thibodeaux models predict similar emission rates for
initial volatile losses in the absence of biodegradation. Thus,
the RTI model was selected for use in the standards development
process.
7.2.2 Waste Application Model
At land treatment facilities that do not use subsurface
injection with immediate tilling, emissions may occur during the
time that waste is being applied to the soil surface and while
the waste lies on the soil before it is tilled into the soil. No
existing models were identified that predict emissions during
application of an oily sludge to the soil surface. The approach
selected for this case was to calculate an overall mass transfer
coefficient of volatile material from the surface of the stream
of sludge as it falls from the end of a hose to the soil surface.
The mass transfer coefficients were calculated using an equation
presented in Section 5.0 (Table 5-1). The constant in the
equation for gas-phase resistance was increased by a factor of
two in an attempt to account for an increase in mass transfer
caused by the motion of the waste stream through the air. The
equations for making this calculation are presented in Table 7-6
along with the definitions of the variables used and the sources
of input data.
7.2.3 Oil Film Model
Emissions from waste lying on the soil surface are estimated
in either of two ways depending on the form of the waste as it
lays on the soil surface. In typical situations where the
applied waste is spread over the surface of soil, the RTI land
treatment model can be used to estimate emissions. The equation
for short-term emissions given above as Equations (7-4) and (7-7)
7-30
-------
TABLE 7-6. WASTE APPLICATION EMISSION MODEL
Emission equations
E = KCA
A = 2» rl
111
K = Keq k (used for oily sludges) ; ~~ = ~~ +
G K k Keqk
L G
(used for dilute aqueous waste);
-3 0.78 -0.67 -0.11
k = 9.64(10 )U Sc de
G G
H
Keq = (used for dilute aqueous waste)
RT
* • MW .,
Keq = (used for oily sludges)
o L air
-6 -4 *2.2 -0.5
k = 1 (10 ) + 144 (10 ) U Sc
L L
de = *—•
.5
; Sc
U
Da
* 05
U = 0.01U(6.1 + 0.63U)
Variable Definition
Sc =
L
D
w w
Data source
E
K
Keq
H
R
Emission ratefor constituent,g/s
Overall mass transfer coefficient,m/s
Equilibrium coefficient, dimensionless
Henry's law constant for constituent,
atm cm /g mol
Universal gas constant, atm cm /
g mol K
Calculated
Calculated
Calculated
Literature
Literature
T
kG
-k
P
Temperature, K
Gas-phase mass
Vapor pressure
transfer coefficient, m/s
of constituent, mm Hg
Measured
Calculated
Literature
[ continued)
7-31
-------
TABLE 7-6 (continued)
Variable
P
o
U
SCG
SCL
U
a
•
w
•
a
•
L
D
a
Definition
System pressure (atmospheric pressure) ,
mm Hg
Windspeed, m/s
Schmidt number on gas side
Schmidt number on liquid side
Viscosity of air, g/cm»s
Density of water, g/cm
Density of air, g/cm
Density of oil, g/cm
Diffusivity of constituent in air,
2 ,
cm /s
Data source
Definition
Definition
Calculated
Calculated
Literature
Literature
Literature
Estimated
Literature
A
r
1
C
de
MW
MW . n
oil
D
w
U
Surface area of cylindrical waste
stream, m
Radius of cylindrical waste stream, m
Length of cylindrical waste stream, m
Concentration of constituent in the
waste, g/cm
Effective diameter of waste stream
surface area, m
Molecular weight of air, g/g mol
Molecular weight of oil, g/g mol
Viscosity of water, g/cm»s
Diffusivity of component in water,
cm /s
Friction velocity, m/s
Liquid-phase mass transfer coefficient,
m/s
Calculated
Measured
Measured
Measured
Calculated
Literature
Estimated
Literature
Data base
Calculated
Calculated
7-32
-------
would be used for this situation. If the applied waste has a
visible oil film on top, emissions immediately after spreading
are estimated by calculating an overall mass transfer coefficient
as described in Chapter 5.0 for an oil film on a surface
impoundment. The mass transfer equation was developed by McKay
and Matsuga and is based on data obtained from liquid hydrocarbon
spills on land and water. The equations used to calculate
emissions under this situation are given in Table 7-7 along with
definitions of the variables used.
7.2.4 Model Inputs
Typical values of input parameters for the RTI model
are based primarily on a data base developed by EPA from site
visits and contacts with State, regional, and industry sources
supplemented by information from recent literature. These values
were chosen as reasonably representative of average or typical
practices currently used at land treatment operations. Oil
loading in the soil is a model input that is calculated from
several other parameters that might change independently.
Varying the value of the oil loading rate, thus, has the same
effect as varying any one or any combination of the other
parameters. Oil loading is defined by waste throughput, the
percent oil in the waste, area of the land treatment site, and
the depth to which the waste is mixed in the soil (tilling
depth). Typical values of oil loading are defined from median
values for those parameters by which it is defined. The data
base shows annual throughput varying from about 2 Mg/yr to about
400,000 Mg/yr with a median value of about 1,800 Mg/yr. The area
of land treatment sites ranges from less than 1 hectare (ha) to
about 250 ha with a median value of 5 ha. The data base shows
tilling depth varying from 15 cm to one case of 65 cm, with most
being in the range of 15 to 30 cm. The single most frequently
reported tilling depth is 20 cm, which is selected as a typical
value. This value is in line with values of 15 to 30 cm reported
7-33
-------
TABLE 7-7. OIL FILM SURFACE EMISSION MODEL
Emission rate equation
E = KCtA
Ct = CQ [exp (-Kt/D)]
K = k Keq (used for oily sludges)
i A QO /in~3N TT°-78c -°-67^ -0.11
k = 4.82 (10 ) U Sc~ de
Keq =
de =
• D
a a
P* 'a^oil
P - -MW .
o L air
(used for oily sludges);
4A
0.5
Variable
E
K
C
Definition
Emission rate for constituent, g/s
Overall mass transfer coefficient,
m/s
Concentration of constituent in the
oil phase at time t
Data source
Calculated
Calculated
Calculated
Initial concentration of constituent Definition
o
D
A
kG
in the waste
Oil film thickness, m
2
Area of land treatment, m
Gas-phase mass transfer coefficient,
m/s
Measured
Measured
Calculated
[ continued)
7-34
-------
TABLE 7-7 (continued)
Variable
U
Sc
G
"a
a
D
a
de
Keq
P*
P
o
MW . ..
oil
MW
a
* L
R
Definition
Windspeed, m/s
Schmidt number--gas phase
Viscosity of air, g/cm»s
3
Density of air, g/cm
Diffusion coefficient of constituent
in air, cm /s
Effective diameter of land treatment
area, m
Equilibrium coefficient of constituent
Calculated
Vapor pressure of constituent, mm Hg Literature
Data source
Definition
Calculated
Literature
Literature
Literature
Calculated
Atmospheric pressure, mm Hg
Definition
Molecular weight of the oil, g/g mol Definition
Molecular weight of air, g/g mol Literature
Density of oil, g/cm Estimated
Universal gas constant, atm cm /g mol K
Literature
Temperature, K
Measured
7-35
-------
in another study.12 The data base shows oil content of the waste
streams varying from about 2 to 50 percent, with a median value
of about 12 percent and a mode value of 10 percent. The 10-
percent figure is selected as typical.
Average molecular weight of the oil from which a particular
constituent evaporates is one of the determining factors in the
rate of evaporation and thus must be specified. Little data are
available as guidance for selecting a value for this parameter.
The distribution of constituents by molecular weight in land-
treated wastes is not well known. In one field measurement study
13
of land treatment emissions, a value of 282 g/g mol was used as
the average molecular weight of the oil. This value was based on
distillation of oil from a refinery sludge and identification of
the constituent corresponding to the midpoint distillation
temperature (i.e., the temperature at which 50 percent of the oil
was distilled). The value 282 g/g mol is selected for use. A
sensitivity analysis using the RTI model shows that emissions are
not highly sensitive to this parameter.
Soil air porosity and total porosity impact the effective
diffusivity of a constituent in the soil. Very little soil
porosity information has been identified. One study reported
measured values of soil porosity in a land treatment plot as
14
ranging from 43.3 to 65.1 percent, with an average value of
about 50 percent. The literature values are assumed to represent
air porosity. Total soil porosity would include the air porosity
and the space occupied by oil and water within the soil. One
field study reported measured values of both total porosity and
15
air-filled porosity. Measured values of total porosity ranged
from 54.7 to 64.8 percent, with an average value of 60.7 percent.
Measured values of air-filled porosity ranged from 27.4 to 46.9
percent, with an average of 37.2 percent. The value of 61
percent for total porosity is assumed to be a representative
value. A value of 0.5 is used in the model as a default for air
porosity.
7-36
-------
Biorate data used in the RTI model data base (CHEMDAT8)
represent measured rates in aqueous systems. In order to use the
aqueous biorate data in a land treatment process, a factor was
established for converting aqueous data to land treatment values
using measured data for benzene. A recent publication by the
American Petroleum Institute (API) reported experimentally
determined values of biological decay constants for land
treatment studies using two different soil types. Decay
constants were measured for six compounds including two
compounds, benzene and toluene, that have aqueous biorates in the
land treatment model data base (CHEMDAT8). For benzene, the
ratio of the API data, measured in units of day , and the
aqueous data, measured in units of mg VO/g, . »hr, was
calculated as 0.00179. This value is also a close approximation
of the ratio of the two data points for toluene, the other
compound for which data from both sources were available. The
other compounds for which data were reported by API did not have
referenced aqueous data in the data base. The above calculated
relationship was used to calculate equivalent aqueous data values
for those compounds. Reported and calculated values of aqueous
biorates and land treatment biological decay constants are
presented in Table 7-8. The ratio of 0.00179 is used for all
compounds to convert from aqueous biorates to decay constants
that can be used in the land treatment model. The input
parameter for the land treatment model is a biological decay time
constant, t, , in units of seconds. The equation for calculating
t, from the aqueous biorate is derived as follows.
The biological decay time constant is, by definition, equal
to the reciprocol of the biological decay constant, or
tb = —±- , (7-16)
7-37
-------
TABLE 7-8. MEASURED AND ESTIMATED BIORATES AND DECAY CONSTANTS
FOR SELECTED ORGANIC CONSTITUENTS
Aqueous
Calculated „, , ,
Measured decay
decay constant, day
Organic biorate, constant, ,
constituent mg VO/g biomass»h day Nunn Kidman
Benzene 19.0
Ethylbenzene 46. 4d
Xylene(-o) 40. 8d
Naphthalene 42. 5d
Toluene 73.5
Methyl naphthalene 24. Od
0.034 0.034 0.013
0.083 0.083 0.076
0.073 0.073 0.026
0.076 0.076 0.050
0.132 0.106 0.119
0.043 0.043 0.059
b
Reference 17.
Data obtained using a clay loam soil (Nunn soil).
c
Data obtained using a sandy loam soil (Kidman soil]
Values calculated from API experimental data.
7-38
-------
where K, = biological decay constant. The ratio, r, of decay
constant to aqueous biorate is:
K, g, . h ,
b bio T, „ , -1
r = =— ,„ , or K, = rB day
B mg VO day b •*
Substituting into Equation (7-16) gives:
tb = day .
To obtain a result in seconds, this equation must be multiplied
by 86,400 s/day. Making this conversion and inserting the value
of r (i.e., 0.00179) gives:
86,400 = 4.83 (1Q7)
b 0.00179B B
For situations in which petroleum wastes are landfarmed and
no information is known about the nature of the volatile
materials, it is possible to estimate a default biorate from the
average decay constant values reported in the API investigation,
0.07 day , which corresponds to a biorate, B, of 40 mg/g-h.
This value is between the values for benzene and toluene in the
data base. The average value of the biological rate constant in
the two soils investigated by API was not significantly
different.
In summary, parameters and selected typical values for use
in the RTI model are as follows:
Annual waste throughput = 1,800 Mg
Area of land treatment = 5 ha
Oil content of waste = 10 percent
Average molecular weight of the oil = 282 g/g mol
Soil air porosity = 0.5
Soil total porosity = 0.61
Tilling depth = 20 cm
Temperature = 25 °C.
7-39
-------
7.2.5 Estimation of Total VO Emissions
The preceding discussion has been limited to estimating
emissions of a single constituent in a hazardous waste. Using
the models presented here to estimate total VO emissions can be
accomplished using any of several approaches. The most obvious
approach, and the one that should give the most accurate results,
would be to obtain a detailed analysis of the constituents in the
waste being land treated. The emission equations could be used
to calculate emissions of each constituent, and total emissions
could be calculated by summing the emissions of individual
constituents. In many cases, a detailed analysis of the applied
waste may not be available, and other, less accurate methods may
be needed to estimate total VO emissions. An alternative to the
constituent approach could make use of a boiling curve or steam
stripping test of a sample of the waste. Experimental data
1 8
developed by Chevron Research Company indicate that a large
fraction of the constituents that boil at temperatures of 400 °F
or lower will be emitted to the atmosphere and that those
constituents with higher boiling points will tend to remain in
the soil for a sufficient time to undergo biodegradation.
Similar results are obtained by applying the RTI land
treatment model to the constituents in the CHEMDAT8 data base.
If a sample of waste were subjected to a laboratory boiling test
or other equivalent test at a temperature of 400 °F, the fraction
of oil evaporated would approximate the fraction that potentially
would be emitted to the air in a land treatment operation.
A third approach to estimating total VO emissions would
again make use of the experimental results generated by Chevron
Research. The test results showed that approximately 25 percent
of the applied oil in the land treatment test was emitted to the
air. In the absence of a detailed constituent analysis and with
no boiling or steam stripping test of the waste, a crude estimate
of total VO emissions could be made by assuming that emissions
are equal to 25 percent of the applied oil. This approximating
7-40
-------
alternative would only apply to raw oily refinery wastes that
have not undergone any pretreatment to remove VO.
7.2.6 Example Calculations
7.2.6.1 Emissions from land treatment soil. The following
calculation demonstrates the use of the RTI model to calculate
the long-term fraction of applied material emitted to the
atmosphere and to calculate the short-term and long-term emission
rates and emission fractions. The calculations are made for
benzene at a concentration of 2,000 ppm by weight in a waste
stream that is 10 percent oil.
Input values are:
Land area = 2.5 ha (half of total area of 5
ha assumed active)
Annual throughput = 1,800 Mg
Oil content of waste = 10% (by weight)
Benzene concentration in oil = 2,000 ppm (by weight) (2
mg/g oil)
Tilling depth = 20 cm
Soil air porosity = 0.5
Soil total porosity = 0.61
Average molecular weight of oil= 282 g/g mol .
a. Calculate oil loading (Equation from Table 7-3) :
(1, 800 x 106 g . ) (0.1 g . , /g
^' ^ ^
T _ _ n nq,- /
L = - 5 - ^ - = 0.036 g. -, / cm
(2.5 x 10 cm ) (20 cm)
b. Identify constituent properties of benzene:
h
B = 19.00 mg VO/g, . •
biomass
D =0.088 cm2/s
cL
P* = 95.2 mm Hg = 0.125 atm.
7-41
-------
c. Calculate the equilibrium coefficient (Table 7-3):
oil a
(0.125 atm) (282 g/g mol) (0.5 cm3/cm3)
Kecf RT L (82.05 atm»cm3 /g mol-K) (298 K) (0.036 g/cm3)
Keq = 0.02002 .
d. Calculate the biological degradation time constant (Equation
from Table 7-3) :
= 4.83(107) = 4.83 (107) = 6
L Z. . 0 J.U ,1 t> .
e. Calculate the effective diffusivity of constituent in the
soil (Equation from Table 7-3) :
• 10/3 10/3
D = D -^5 = 0.088 cm2/s (^—\ = 0.0235 cm2/s
6 a -T2 0.612
•2KeqD
f. Calculate the value of K, = ~— :
d 4 I2
R = (9.87) (0.0235) cm2/s (0.02002) = 2>g (10-6} s~l
d 4 (400) cm2
g. Calculate the fraction of constituent emitted to the air
after a long time (Equation (7-11)):
0.81 K,t,
F = TF-T -^r-^ + 0.1878 = 0.90 .
a Kdtb + 1
h. Calculate the long-term emission rate after 60 h (216,000 s]
Det 0.02002 x 0.0235 x 216,000
Keq
7-42
-------
Use Equation (7-5) (long-term equation):
2 K D
E = M
eg e
exp
exp-l
4/2
2 (0.72) (2) (0.02002) (0.0235) ,-0.02002(0.0235) (9.87) (216,000;
x e (
400
1, 600
x e[-216,000/2.54(10 )]
= 3.38(10~6) e(-0.627) e(-0.085) = 3.38(10~6) e(-0.712;
E = 3.38(10 6) (0.491) = 1.7(10 6) .
mg
2
cm s
c. Calculate the short-term emission rate after 15 min (900 s]
V = 0.02002 x 0.0235 (900) =
u.uuiu
k = 4.82 (10-3) U°'78 Sc-°'67
where
U = windspeed = 4.47 m/s
de = effective diameter of land treatment area
de =
0.5
-------
=4.82 (103
.7!
(4.47
= 0.0061 m/s ( 0.61 cm/s )
d.7i7°-67
'17
ZJ1 ^
/
1
ea
+
vt V
E =
0.72 g 2 mg
2
1 cm 20 cm g
0-5
*0.61 x 0.02002
= 0.072 (0.0004) e(-0.0004) = 2.87 (10 )
»mvvm»'
0.0235 x 0.02002
-5, mg
(-900/2.54 (10 ) )
2
cm s
Table 7-9 shows estimated emission rates and emission
fractions for various times up to 40 days (960 hours).
7.2.6.2 Emissions from Waste Application. The
following is an example calculation for the application of an
oily waste to a land treatment plot using the equations in
Table 7-6. For benzene in waste oil, the calculation is:
Input values:
r
L
a
a
U
R
T
C
0. 038 m
= 0.46 m
1.81 (10~4) g/cm»s
= 1.2 (10~3) g/cm3
4.47 m/s
= 82.05 atm»cm3/K»g mol
= 298 K
= 200 ppm =
200 ug/g
=
3
(assuming a density of 1 g/cm )
A
MW.
C
MW
oil
= 2 • rL = 2(3.14) (0.038 m) (0.457 m) = 0.11 m
i / 3
= 1 g/cm
= 29 g/g mol
= 282 g/g mol.
2
7-44
-------
a. Calculate the effective diameter of the waste stream
surface (Equation from Table 7-6):
0.5
de = *^^* = 0.37 m
TABLE 7-9. ESTIMATED EMISSION RATES AND FRACTIONS EMITTED
VERSUS TIME FOR EXAMPLE LAND TREATMENT CALCULATION
Time after
application
/tilling, h
1
2
4
8
24
48
96
192
480
960
Emission
rate, 0
i n~6 , 2
10 mg/cm
14.4
10.3
7.30
5.12
2.90
1.98
1.08
0.348
0.011
0.00004
Equation
•s used
Short term
Short term
Short term
Short term
Short term
Short term
Long term
Long term
Long term
Long term
Fraction
emitted,
fraction
0.073
0.104
0.147
0.207
0.356
0.497
0.674
0.827
0.899
0.901
Equation
used
Short term
Short term
Short term
Short term
Short term
Short term
Long term
Long term
Long term
Long term
7-45
-------
b. Calculate the Schmidt number (Equation from Table 7-6):
11 — 4
^ _ Ka _ 1.81 (10 ) g/cm*s _ , 7,
G *a Da [1.2 (10~3) g/cm3](0.088 cm2/s)
c. Calculate the equilibrium coefficient
(Equation from Table 7-6):
* • MW -3 3
P a oil = (95.2 mm Hg) [1.2 (10 )g/cm ] (282 g/gmol)
^ P • MW 3
o L a (760 mm Hg)(1 g/cm )(29 g/g mol)
Keq = 0.0015.
d. Calculate the gas-phase mass transfer coefficient
(Equation from Table 7-6) :
kr = 9.64(10-3) U°'78 Scr -°'67 de-0'11
(j (j
= 9.64(10~3) (4.47)°'78(1.71)~°'67 (0.37)-0'11
= 0.024 m/s
e. Calculate the overall mass transfer coefficient (Equation
from Table 7-6):
K = k_ Keq = (0.0015)(0.024) m/s = 0.000036 m/s .
(j
f. Calculate emissions from a unit volume of waste
(Equation from Table 7-6) :
E = KCA = 0.000036 m/s (200 g/m3) (0.11) m2
= 7.92 (10~4) g/s .
Using a calculated fall time of 0.305 s:
E = (7.92»10~4 g/s) (0.30 s) = 2.38 (10~4 ) g .
Stream volume = • r2L = (3.14) (0.038 m)2(0.46 m) = 0.002 m3
7-46
-------
= 2,073 cm3 .
Mass of constituent = (0.002 m3) (200 g/m3) = 0.4 g
g. Calculate the fraction of constituent emitted to the air:
-4
Fraction emitted = 2'3^(^° ^ = 0.00059 = 0.06 percent.
w • fi y
7.2.6.3 Emissions from an oil layer on soil prior to
tilling. An example calculation for butanol-1 in an oil layer
on the soil surface of a land treatment site is given below
using the equation from Table 7-7.
Input values:
u = 1.81 (10~4) g/cm»s
= 1.2 (10~3) g/cm3
cL
U = 4.47 m/s
MW . -, = 282 g/g mol
oil ^
• = 1 g/cm
j_i
MW = 29 g/g mol
C = 0.0002 g/cm3 = 200 g/m3
A = 25,000 m2
t = 24 h = 86,400 s
d = 0.072 m .
a. Calculate the effective diameter of the soil surface
(Equation from Table 7-7):
, * 4 A, ° ' 5 *4x25,000:);0'5 . _ 0
= *__±* = * _ r_ * =
b. Calculate the Schmidt number (Equation from Table 1-1',
_ ^a _ 1.81 (10 ) g/cm*s
G *aDa [1.2 (10~3) g/cm3] (0.080 cm2/s)
7-47
-------
c. Calculate the equilibrium coefficient
(Equation from Table 7-7):
P • MW -3 3
g oil (6.5 mm Hg) [1.2(10 Jg/cm ] (282 g/gmol)
q P • MW 3
o L a (760 mm Hg) (1 g/cm ) (29 g/g moi;
= 1.0 (10~4)
d. Calculate the gas-phase mass transfer coefficient
(Equation from Table 7-7):
k = 4.82(10 3)U°'78Sc°'67de °'l= 4.82(10 3) (4 . 47? ' 78 (1.89)°'67
(j (j
x (17870'11
= 5.7 (10~3)
e. Calculate the overall mass transfer coefficient (Equation
from Table 7-7):
K = k,, Keq = [5.7 (10~3 ) m/s] [1.0 (10~4 ) ] m/s .
-7
K =5.70 (10 ) m/s.
f. Calculate the fraction of constituent emitted to the air at
time t (Equation from Table 7-7):
f = 1 - A = 1 - e~Kt/D
o
f = 1 - exp [-5.70 (10~7) (86,400)/O.072]
= 1 - 0.50 = 0.50 .
7.2.6.4 Examples of the Use of the Land Treatment Model for
Specific Cases. To illustrate how the land treatment model is
used to estimate long-term and short-term emissions from various
waste management options, the following case studies are
presented. Each of these examples represents the land treatment
of API separator sludge/DAF float, with the exception of
example 5. The waste contains 10 percent organics and is applied
at a rate of 2,500 Mg of waste per year. The loading, benzene
concentration, and porosity are identical to the example
7-4!
-------
presented in Section 7.2.6.1. The land area is 35,000 square
meters.
Example 1. Waste is applied monthly for 9 months of the
year. No waste is applied during December, January, or February
because the soil is frozen or saturated with water during those
months. The waste is applied from a vacuum truck by spraying
onto the soil surface with a nozzle. The soil is tilled 24 hours
after application and is tilled again (no waste application)
after 2 weeks.
Two calculations of air emissions are required: after
application and after tilling. The first time period is 1 day
and the second time period is 14 days. The amount of waste
applied is 2,520 Mg/9 or 280 Mg/application. It is assumed that
3
the oil content in the soil is 0.036 grams of oil/cm . The
amount of oil applied each application is 0.036/9 or 0.004 g/cnT
The concentration of benzene in the waste is 2,000 ppmw and the
concentration of benzene added to the oil in the soil is 2,000/9
or 222 ppmw each application. The land area is 3,500 square
meters. After application, the liquid is assumed to seep into
the soil to a depth of 5 cm, and the oil loading in the waste on
the soil surface was assumed to be the same as the oil content of
the soil.
To estimate the amount of air emissions between application
and tilling, the following parameters are used in CHEMDAT8:
• Concentration of benzene: 2,000 ppmw
• Tilling depth: 5 cm
• Time of calculation: 1 day
3
• Loading (10 percent oil): 0.036 g/cm .
The fraction lost during the first day is 98 percent with 0.007
percent lost to biological decay.
The fraction lost after the first tilling is estimated by
the use of the following parameters:
• Concentration of benzene: 222 ppmw
• Tilling depth: 20 cm
7-49
-------
• Time of calculation: 14 days
• Loading: 0.036 g oil/cm .
The calculated fraction lost during the first tilling period is
0.89 to air emissions and 0.095 to biodegradation. This fraction
is independent of the concentration of benzene and is expected to
also apply to the second tilling period.
The loss of benzene during the month is 97.8 percent during
the spreading period and (1-0.978-0.0067) x 0.89 x 100 or 1.4
percent during the tilling periods, for a total air emission loss
of 99.2 percent.
Example 2. Waste is applied weekly except when the ground
is saturated with water. Waste is applied from a vacuum truck,
and the waste is spread over the surface of the soil. The soil
is tilled on the day following the application and weekly between
applications. The waste is applied monthly to the land treatment
plot throughout the year.
As in example 1, two calculations of emissions are required:
after application and after tilling. The first time period is
1 day and the second time period is 7 days. The amount of waste
applied is 2,520 Mg/12 or 210 Mg/application. It is assumed that
the oil content in the soil is 0.036 grams of oil/cm . The
amount of oil applied each application is 0.036/12 or 0.003
g/cm . The concentration of benzene in the waste is 2,000 ppmw
and the concentration of benzene added to the oil in the soil is
2,000/12 or 167 ppmw each application. The land area is 3,500
square meters. After application, the liquid is assumed to seep
into the soil to a depth of 5 cm, and the oil loading in the
waste on the soil surface was assumed to be the same as the oil
content of the soil.
To estimate the amount of air emissions between application
and tilling, the following parameters are used in CHEMDAT8:
• Concentration of benzene: 2,000 ppmw
• Tilling depth: 5 cm
• Time of calculation: 1 day
3
• Loading (10 percent oil): 0.036 g/cm .
7-50
-------
The fraction lost during the first day is 98 percent with 0.007
percent lost to biological decay.
The fraction lost after the first tilling is estimated by
the use of the following parameters:
• Concentration of benzene: 167 ppmw
• Tilling depth: 20 cm
• Time of calculation: 7 days
• Loading: 0.036 g oil/cm .
The calculated fraction lost during the first tilling period is
0.80 to air emissions and 0.083 to biodegradation. This fraction
is independent of the concentration of benzene and is expected to
also apply to the second tilling period.
The fraction loss of benzene during four tilling periods is
(0.805) + (0.0995)(0.805) + (0.805)(0.0995)2 + 0.805 (0.0995)3 or
0.89.
The loss of benzene during the month is 97.8 percent during
the spreading period and (1-0.978-0.0067) x 0.89 x 100 or 1.4
percent during the tilling periods, for a total air emission loss
of 99.2 percent.
Example 3. The waste is dewatered prior to land treatment.
The parameters are the same as those used in example 1, except
the waste is dewatered and the filter cake is land-treated. The
oil content of the filter cake is 20 percent. The waste is
applied from a dump truck and is spread by a bulldozer. The
waste is tilled into the soil on the day following spreading. It
is assumed that the dewatering process removes 60 percent of the
oil from the waste.
As in the preceeding examples, two calculations of air
emissions are required: after spreading and after tilling. The
first time period is 1 day and the second time period is 14 days.
The amount of waste applied is 2,520 Mg/9 or 280 Mg/application.
It is assumed that the oil content in the soil is 0.036 grams of
oil/cm3. The amount of oil applied each application is 0.036/9
or 0.004 g/cm3. The concentration of benzene in the waste is
7-51
-------
2,000 ppmw and the concentration of benzene added to the oil in
the soil is 2,000/9 or 222 ppmw each application. The land area
is 3,500 square meters. After application, the liquid is assumed
to be retained in the waste, and the oil loading in the waste on
the soil surface is assumed to be the same as the oil content of
the waste.
To estimate the amount of air emissions between application
and tilling, the following parameters are used in CHEMDAT8:
• Concentration of benzene: 2,000 ppmw
• Tilling depth: 2 cm
• Time of calculation: 1 day
3
• Loading (20 percent oil) : 0.2 g/cm .
The fraction lost during the first day is 98 percent with 0.006
percent lost to biological decay.
The fraction lost after the first tilling is estimated by
the use of the following parameters:
• Concentration of benzene: 222 ppmw
• Tilling depth: 20 cm
• Time of calculation: 14 days
3
• Loading: 0.036 g oil/cm .
The calculated fraction lost during the first tilling period is
0.89 to air emissions and 0.095 to biodegradation. This fraction
is independent of the concentration of benzene and is expected to
also apply to the second tilling period.
The air emission loss of benzene during the application
period is 98.3 percent during the spreading period and (1-0.983-
0.006) x 0.89 x 100 or 1.2 percent during the tilling periods,
for a total air emission loss of 99.5 percent. The air emissions
on the basis of the untreated waste would depend on the recovery
of oil in the dewatering process and the air emissions from the
dewatering process.
Example 4. The waste is tilled as it is applied to the
soil. The tilling depth is 20 cm. The period between tillings
is 3 days. Waste is applied monthly.
7-52
-------
Only one calculation of air emissions is required: after
tilling. The amount of waste applied is 2,520 Mg/12 or 210
Mg/application. It is assumed that the oil content in the soil
is 0.036 grams of oil/cm . The amount of oil applied each
3
application is 0.036/12 or 0.003 g/cm . The concentration of
benzene in the waste is 2,000 ppmw, and the concentration of
benzene added to the oil in the soil is 2,000/12 or 167 ppmw each
application. The land area is 3,500 square meters.
The fraction lost after the first tilling is estimated by
the use of the following parameters:
• Concentration of benzene: 167 ppmw
• Tilling depth: 20 cm
• Time of calculation: 3 days
3
• Loading: 0.036 g oil/cm .
The calculated fraction lost during the first tilling period is
0.60 to air emissions and 0.055 to biodegradation. This fraction
is independent of the concentration of benzene and is expected to
also apply to the following tilling periods.
The loss of benzene during the first tilling period is 60
percent with (100-60-5.5) or 34.5 percent remaining. The loss of
benzene during the second tilling period is 0.60 (34.5) or 20.7
2
percent, with a fraction of (0.345) or 0.119 of benzene
remaining. The total loss of benzene for the month is
92 percent.
Example 5. Waste is applied monthly for 9 months of the
year. No waste is applied during December, January, or February
because the soil is frozen or saturated with water during those
months. The waste is applied from a vacuum truck by spraying
onto the soil surface with a nozzle. The soil is tilled 24 hours
after application and is tilled again (no waste application)
after 2 weeks. The waste is aqueous containing 10 percent
organics and 2,000 ppmw benzene. The waste contains 10 percent
solids by weight.
Two calculations of air emissions are required: after
application and after tilling. The first time period is 1 day
7-53
-------
and the second time period is 14 days. The amount of waste
applied is 2,520 Mg/9 or 280 Mg/application. It is assumed that
3
the water content in the soil is 0.10 grams/cm . The
concentration of benzene in the waste is 2,000 ppmw and the
concentration of benzene added to the water in the soil is 300
ppmw each application. The land area is 3,500 square meters.
After application, the liquid is assumed to seep into the soil to
a depth of 5 cm, and the water loading in the waste on the soil
surface was assumed to be the same as the water content of the
soil.
To estimate the amount of air emissions between application
and tilling, the following parameters are used in CHEMDAT8:
• Concentration of benzene: 300 ppmw
• Tilling depth: 5 cm
• Time of calculation: 1 day
3
• Loading (10 percent): 0.10 g/cm .
The fraction lost to air emissions during the first day is 99.8
percent with less than 0.01 percent lost to biological decay.
7.2.7 Assumptions and Sensitivity Analyses
The RTI model incorporates the following assumptions to
simplify development and use of the model:
• Volatilization and biodegradation are the predominant
waste removal mechanisms (i.e., other mechanisms can be
ignored).
• Waste is mixed uniformly within a surface layer of the
soil.
• Waste does not flow as a liquid within the soil.
• The adsorption isotherm of a constituent is linear
within the application surface layer and does not
change with time.
• No bulk flow of gas is induced within the soil.
• The rate of biological decay/chemical reaction is a
first-order process.
• The diffusion coefficient does not vary with either
concentration or time.
7-54
-------
• The concentration of a constituent in the gas phase at
the surface of the soil is much lower than the
concentration of that constituent in the gas phase
within the soil.
• There is no diffusion of waste into the soil beneath
the zone of incorporation.
• Liquid-vapor equilibrium is established at all times
within the soil.
The RTI model was evaluated for sensitivity to each of the
input parameters. In the analysis, each input parameter was
varied over the entire range of reasonable values. The effect on
emissions of parameter variations was noted, and the parameters
showing the highest sensitivity were identified.
Individual constituent properties were found to have the
most significant impact on land treatment emissions. These
properties include vapor pressure, diffusivity in air, and
biodegradation rate. The more volatile compounds are mostly
emitted to the atmosphere unless a volatile compound also has a
high biodegradation rate or low diffusivity. Compounds with low
vapor pressures tend also to have lower diffusivities; thus, even
if such compounds also have moderate or low biorates, they are
mostly biodegraded rather than emitted to the air.
Operating and field parameters also have an impact on
emissions but to a lesser extent than compound properties.
Tilling depth, soil porosity, and waste loading all have an
impact on air emissions, with the largest impact on the more
volatile compounds. Tilling depth can have a substantial impact
on air emissions of volatile compounds, especially if a compound
also has a relatively high biorate. As tilling depth increases,
materials remain in the soil for a longer time and thus have a
greater chance of being biodegraded.
Waste loading can be varied in two ways: by increasing the
concentration of a compound in the waste or by increasing the
amount of material applied to the soil per unit area. If the
concentration of a compound is changed, air emissions change in
direct proportion to the change in concentration (i.e., the
7-55
-------
fraction of the compound that is emitted to the air remains
constant). If total waste loading is changed, air emissions
change in the same direction as the change in loading but not in
direct proportion (i.e., the fraction emitted is lower for higher
loading rates). These results assume that a treatment site is
not overloaded to the point where biodegradation ceases to be a
first-order process.
Average molecular weight of the oil has an effect on air
emissions, but the magnitude of the effect is less than that of
the other parameters studied. As average molecular weight goes
up, the fraction emitted for a specific constituent increases;
the fraction emitted decreases if the average molecular weight is
reduced.
7.3 REFERENCES
1. Crank, J. The Mathematics of Diffusion. London, Oxford
University Press. 1970. p. 45-47.
2. Crank, J. Diffusion in Polymers. New York, Academic Press.
1968. p. 16.
3. Huddleston, R. L., C. A. Bleckman, and J. R. Wolfe. Land
Treatment Biological Decay Processes. In: Land Treatment,
a Hazardous Waste Management Alternative. Water Resources
Symposium Number Thirteen. Austin, The University of Texas
at Austin. 1985. p. 44.
4. Letter and attachment from Klett, R. J., Chevron
Corporation, to Wyatt, Susan R., U.S. Environmental
Protection Agency, July 8, 1987, p. 1-2. Comments on Draft
Emission Models Report.
5. Levenspiel, 0., Chemical Reaction Engineering. John Wiley
and Sons, Inc., New York, New York. May 1967. p. 47.
6. Millington, R. J, and J. P. Quirk. Permeability of Porous
Solids. Trans. Faraday Society. 5_7_: 1200-1207. 1961.
7. Letter and attachments from Sonenville, G. F., Chevron
Research Company, to Thorneloe, S. A., U.S. EPA. May 22,
1986. p. 19. Comments on preliminary draft BID for land
treatment.
8. GCA Corporation. Air Emissions from Land Treatment--
Emissions Data and Model Review. Draft Technical Note.
7-56
-------
Prepared for U.S. Environmental Protection Agency. Research
Triangle Park, NC. September 1985. Chapter 4.
9. Radiation Technologies, Inc. Treatability Data in Support
of a Treatment Zone Model for Petroleum Refining Land
Treatment Facilities. Prepared for American Petroleum
Institute. Washington, DC. March 1986. p. 154.
10. McKay, Donald, and Ronald S. Matsuga. Evaporation Rates of
Liquid Hydrocarbon Spills on Land and Water. The Canadian
Journal of Chemical Engineering. 51:434-439. August 1973.
11. Memorandum from Thorneloe, S., EPA/OAQPS, to Durham, J.,
EPA/OAQPS, January 31, 1986. Land treatment data base.
12. Environmental Research and Technology. Land Treatment
Practices in the Petroleum Industry. Prepared for American
Petroleum Institute. Washington, DC. June 1983. p. 1-2.
13. Radian Corporation. Field Assessment of Air Emissions and
Their Control at a Refinery Land Treatment Facility. Volume
I. Prepared for U.S. Environmental Protection Agency.
Cincinnati, OH. September 1986. p. 154 and 179.
14. Reference 13, p. 43.
15. Utah Water Research Laboratory. Evaluation of
Volatilization of Hazardous Constituents at Hazardous Waste
Land Treatment Sites. For U.S. Environmental Protection
Agency, Office of Research and Development, Ada, Oklahoma.
Undated, p. 55.
16. American Petroleum Institute. Land Treatability of Appendix
VIII Constituents Present in Petroleum Refinery Waste:
Laboratory and Modeling Studies. API Publication No. 4455.
American Petroleum Institute, Washington, DC. April 1987.
P. 3-10 and 3-12.
17. Reference 16.
18. Ricciardelli, A. J., et al. 1986. Landfarm Simulator
Program. Summary Report. Chevron Corporation, Richmond,
California. July 1987. p. 18-24.
7-57
-------
8.0 LANDFILLS AND WASTEPILES
8.1 INTRODUCTION
The main facilities in this category that constitute the
model plants are waste fixation units, open landfills, closed
landfills, and wastepiles.
All wastes that contain free liquids and that are destined
for disposal in a landfill must be treated to eliminate the
free liquids. This is often accomplished by adding a
"fixative" to the waste, such as portland cement, cement kiln
dust, or lime flue dust. These materials react with water in
the waste and set up to form a dry material that encapsulates
or binds the organic constituents of the waste. This fixation
process is most often conducted in lined open pits or open
tanks into which the liquid waste is poured. The fixative then
is added and the materials are thoroughly mixed, most often
with a backhoe. Alternatively, mechanical mixers such as
pugmills can be used to blend the waste and fixative. Emis-
sions are generated for as long as the waste remains in the
pit. Emissions from this process may be estimated by using the
open dump model.
A landfill is a facility, usually an excavated, lined pit,
into which wastes are placed for permanent disposal. Emissions
from open landfills, those still receiving wastes, can be
estimated by applying the Research Triangle Institute (RTI)
land treatment model. Emissions from closed landfills, those
filled to design capacity and with a cap (final cover)
installed, can be estimated with the RTI closed landfill model.
-------
Wastepiles are temporary accumulations of waste. They
serve a storage function and have a limited life span.
Emissions from wastepiles can be estimated by applying the RTI
2
land treatment model.
8.2 CLOSED LANDFILLS
8.2.1 Emission Model Equations
The RTI closed landfill model is used to estimate the
time-dependent behavior of emissions from waste placed in a
closed (capped) landfill that is vented to the atmosphere and
(as a special case) open-landfill waste covered with daily
earth covers. This model accounts for escape of the
constituent of interest via two primary, independent
mechanisms: diffusion through the cap and convective loss from
barometric pumping through the vent(s). It is the purpose of
this section to describe the model and its evolution in a
general way and to present all model equations and major
assumptions.
The model is based primarily upon the work of Farmer
et al.,3 who applied Fick's first law for steady-state
diffusion to closed landfills. Farmer's equation utilizes an
effective diffusion coefficient for the soil cap based on the
work of Millington and Quirk.4 A previous EPA study5 was
dedicated to the evaluation of available models for estimating
emissions from hazardous waste treatment, storage, and disposal
facilities (TSDF), including closed landfills. This study
endorsed the models of Farmer et al.6 and Thibodeaux7 for
closed landfills, apparently because of their treatment of
soil-pore diffusion. Of the two, the Farmer et al.8 model
alone has received experimental verification (although to a
limited degree) via a laboratory experiment using
hexachlorobenzene-containing waste in a simulated landfill.
The diffusion model of Farmer et al.9 was subsequently
modified by RTI to allow for convective losses of the
-------
constituent of interest from the landfill, which can occur from
barometric pumping. Furthermore, the decline in the emission
rate from closed landfills over the long term was accounted for
via the incorporation of a time-dependent decay function.
The barometric pumping emission mechanism results from
changes in atmospheric pressure—as the pressure is lowered,
gas flows out of the landfill through the vent(s) to equalize
internal pressure. The contribution to total emissions
resulting from barometric pumping equals the concentration of
the constituent of interest in the gas within the landfill
multiplied by the total flow of gas from the landfill. It is
recognized that under certain conditions (e.g., the presence of
significant biomass) biogas could be generated in a landfill.
Biogas consists of methane and carbon dioxide, which is
produced from the action of bacteria on organic material.
Because of the convective or purging action of biogas in remov-
ing the constituent of interest in vapor form, biological decay
(if it occurs) results in a net increase in the emission rate.
However, it should be noted that there is no evidence that
there is significant biomass (necessary for biogas generation)
in any chemical waste landfill. Therefore, in this analysis it
is assumed, as suggested in the literature, that the toxic
property of the waste will inhibit biological processes and
thus prevent biogas generation.10 Hence, closed landfill model
equations presented in this document account for diffusion
through the cap and barometric pumping only.
The equations inherent in the RTI closed landfill model
are as follows: Fick's first law for steady-state diffusion,
based on the work of Farmer et al.,ll for a landfill is given
as:
(C2i
where,
5-3
-------
J. = vapor flux of~the constituent through the soil
surface, g/cm »s;
2
D . = effective diffusion coefficient, cm /s;
C,-, . = concentration of constituent in the air above the
cap, g/cm air;
C . = concentration of the constituent in the vapor space
beneath the cap, g/cm ; and
1 = cap thickness, cm.
(Because the concentration of the constituent at the surface is
negligible, C~• = 0.)
Emissions associated with diffusion alone (E., ., g/s) are
obtained from the above equation by multiplying by the landfill
2
surface area (A) in cm :
Eli = Ji x A- (8-2:
The effective diffusion coefficient of the constituent in
soil, D ., is computed (using the expression developed by
61 12 13
Millington and Quirk and applied by Farmer et al. ) from the
diffusion coefficient of the constituent in air, D ., as:
a i
Dei = Dai (.3-33 /.2 } (8_3;
where,
2
D . = vapor diffusion coefficient in air, cm /s;
a i
• = soil cap air-filled porosity, cm /cm (the actual
air-filled porosity of the moist soil); and
• = total porosity of the soil cap.
The concentration of the constituent of interest in the
vapor space beneath the cap is computed using the ideal gas law
as follows:
C . = P.MW./RT' = P.MW./R(T + 273) (8-4;
si i i 11
where,
8-4
-------
P. = equilibrium partial pressure of constituent, atm;
MW . = molecular weight of constituent, g/g mol;
1 3 .
T-, 4- 4- no nr cm— • atm
R = gas constant, 82.05 -^
T" = absolute temperature in the landfill, K; and
T = temperature in the landfill, °C.
Calculation of the equilibrium partial pressure, P., depends on
the type of waste liquid as follows:
a. For dilute aqueous solutions (where Henry's law applies),
the equilibrium partial pressure of constituent within the
landfill (P., atm) is computed as:
H .• . . , X. , 3
ci liquid i 6 cm_
i MWn . . , X 1U 3 tb ^
liquid m
where,
H . = Henry's law constant, m »atm/mol;
O _L
• = density of waste liquid, g/cm
(1 g/cm is generally a good estimate for this
parameter) ;
X. = mole fraction of constituent i in waste liquid;
where,
X. = (C./MW. )/[Cw /18 + C./MW. ];
1 11 n~(J 1 1
where,
C. = weight fraction of constituent i in the
original waste liquid;
CH = weight fraction of water in the
2 original waste liquid; and
MW-, . . , = average molecular weight
liqUld jr j. -i • • i / n
^ of waste liquid, g/mol
b. For two-phase (water + organic liquid) or organic liquid
waste, the equilibrium partial pressure of the constituent
8-5
-------
of interest within the landfill (P., a tin) is computed using
Raoult's law:
P. = X.P* (8-6)
where,
X. = mole fraction of constituent in the
organic liquid phase;
where,
Xi = (Ci/MWi)/[Ci/MWi + Coil/MWoil];
where,
C. = weight fraction of constituent
the original waste liquid;
C ., = weight fraction of oil carrier-
liquid in the original waste liquid;
MW ., = molecular weight of oil carrier-
oil -. . . -, / -. ,
liquid, g/g mol; and
*
P. = pure component vapor pressure of the constituent
of interest, atm.
Emissions from barometric pumping are computed as:
E2i = Q x Csi x A (8-7)
where,
Ep. = emissions from barometric pumping, g/s;
Q = flow rate^of gas through the vent, expressed as a
flux, cm /cm landfill area's;
C . = concentration of constituent in the gas within the
si ^
landfill, g/cm gas (see Equation 8-4; and
2
A = surface area of the landfill, cm .
The gas flow rate, Q, is estimated using the following
procedure.
a. Compute volume of gas available for barometric pumping,
assuming the entire void-volume of the waste is available:
5-6
-------
V = D x A x • f (8-8)
c fw ^ '
where,
V = volume of void space, cm ;
O
D = thickness of waste bed within landfill, cm; and
2
A = surface area of the landfill, cm .
• = air porosity fraction of fixed waste
(dimensionless).
b. Compute the total volume of gas (cm ) exiting the vent of
the landfill due to changes in barometric pressure and/or
temperature within the landfill:
+ P T + ?73
Tt *c — ^f* * j-i T^'J# #>
v = v * * — c-*-* * - ± - * _i *
"•R v p * * ** #-"-#
where,
* * P * *T _ + 273* *
1 ref
V = total volume of gas exiting landfill, cm ;
JD
P f = initial (reference) barometric pressure, mm Hg;
PI = final barometric pressure, mm Hg;
TI = final landfill temperature, °C; and
T f = initial (reference) landfill temperature, °C.
For cases in which P- > P f and/or TI V may be negative
(indicating a condition of gas flow into the landfill and
because this condition results in no emissions associated
with barometric pumping}, V_ should be set equal to zero to
JD
avoid calculational errors in the following steps.
c. Compute the average flow rate of gas from the landfill over
the time interval of interest:
5-7
-------
where,
Q = average flow rate of gas from the vent due to
barometric pumping, cm /cm landfill area's;
1t = time interval over which the change in pres-
sure and/or temperature occurred, s; and
2
A = landfill area, cm .
In an average day, barometric pressure drops 4 mbar from a
typical value of 1,013 mbar. Landfill temperature is assumed to
remain constant. Hence, under these conditions, P f = 1,013
mbar, P, = 1,009 mbar, T f = T, = 15 °C, and »t = 8.64 x 104 s.
1 ref 1
Having computed the instantaneous emissions associated with
diffusion through the cap and barometric pumping, the total
initial emission rate at the time of landfill closure,
*
E. (g/s), is computed as the sum:
E/ = EI:L + E2i. (8-11)
The total instantaneous emission rate at any time then is
computed via an exponential decay function:
(3,600 s/h) (24 h/d) (365.25 d/yr)E*
E. (t) = £ exp (-t)
1 °
E. (t) = 31.56 E* exp (-«t) (8-12;
where,
E.(t) = total time-dependent emission rate, Mg/yr;
*
E. = initial emission rate, at time of landfill
closure, g/s;
t = time since landfill closure, months; and
• = "decay" constant, months
The "decay" constant, •, is computed as follows:
-------
(3,600 s/h) x (24 h/d) x 365.25 d/yr) x E*
12 mo/yr x M .
•* 01
• = 2.63 x 106 E*/M . (8-13)
i 01
where M . is the total mass of the constituent of interest in
01
the landfill (g) . This parameter can be computed from the
weight fraction of the constituent in the original waste liquid
(C. ) , the mass of original waste liquid in a unit volume of
fixed waste (W) , the landfill surface area (A) , and the
thickness of the waste layer within the landfill (D) :
M . = C. W A D . (8-14)
01 i
The average emission rate from a closed, vented landfill
over the time since landfill closure is equal to the integral of
the emission rate equation over the time period divided by the
time period, which yields the following expression:
(3,600 s/h) (24 h/d) (365.25 d/yr) E*
~*
(10° g/Mg)'t
31.56E*
EM(t) = — ^ - - [1 - e *r] (8-15)
where
E . (t) = average emission rate over the time since landfill
closure, Mg/yr
t = time since landfill closure, mo.
Table 8-1 summarizes the equations necessary to apply the
RTI closed landfill model.
The model is highly sensitive to the air porosity of the
clay cap (• ) , which largely determines the diffusion rate
a
through the cap. The model is sensitive to the properties of
the constituent of interest, particularly the vapor pressure
*
(P .), Henry's law constant (H .), and mole fraction in the
8-9
-------
TABLE 8-1. RTI CLOSED LANDFILL MODEL
Ei(t) = 31.56 E£ exp (-•t)
31-56 E^ „
EAi(t) = _r^1 *1 - e ^
• = 2.63 x 106 E*/M .
i 01
M . = C. W A D
01 i
E* = E,. + E0.
i li 2i
ET . = J. X A
ll 1
J. = -D . (C0. - C .)/I
i ei 2i si
Dei = Dai Ca3-33/'T2>
c2i = o
C . = P. MW./R (T + 273)
si i i
H . • . . , X. r 3
_ ci liquid i In6 cm ,,.-,, ,
P. = —TT^ x 10 _ (dilute aqueous waste
_L 11 i/v -i • • -i O -i • • -i \
liquid m liquids)
X. = (C./MW. )/((:„ ~/18 + C./MW.) (dilute aqueous
1 11 JTl^O 1 1 j_ n • • T \
2 waste liquids)
*
P. = X. P. (two-phase liquid or organic
liquid waste)
Xi = (Ci/MWi)/(Ci/MWi + Coil/MWoil) (two-phase
liquid or organic liquid waste)
E0. = Q C . A
2i si
Q = VB/(«t A)
3 T -I- 9 7 "3
ref $ $ Tl + 273
'C _ * p1 * * Tref +
„ = D A •
C fw
(continued)
5-10
-------
TABLE 8-1 (continued)
Variable
Definition
A
C
si
C.
C
H2°
C
oil
D
D
ai
D
ei
EAi(t:
Ei(t;
E.
2
Landfill surface area, cm
Concentration of constituent i
in the gas within the landfill,
g/cm gas
Concentration of constituent i
in air above the cap, g/cm
Weight fraction of constituent i
in the original waste liquid
(dimensionless)
Weight fraction of water in the
original waste liquid (dimension-
less)
Weight fraction of oil carrier-
liquid in the original waste
liquid (dimensionless)
Thickness of waste bed within
landfill, cm
Diffusivity of constituent i in
air, cm /s
Effective diffusion coefficient
of^constituent i in clay cap,
cm /s
Average emission rate of con-
stituent i over time t since
landfill closure, Mg/yr
Total instantaneous emission rate
of constituent i at time t since
landfill closure, Mg/yr
Total initial emission rate of
constituent i at time of landfill
closure, g/s
Initial emission rate of constit-
uent i at landfill closure due to
diffusion alone, g/s
Data source
Westat surveyc
Calculated
Assumed
Definition
Definition
Definition
Westat survey
Literature
Calculated
Calculated
Calculated
Calculated
Calculated
b
Reference 14.
Reference 15.
(continued)
5-11
-------
TABLE 8-1 (continued)
Variable
E2i
H .
ci
Ji
M .
01
MW.
MWn . . ,
liquid
MW . n
oil
P .
P.
Definition
Initial emission rate of constit-
uent i at landfill closure due to
barometric pumping alone, g/s
Henry's law constant for constit-
uent i, m »atm/mol
Initial diffusion flux of con-
stituent i, g/cm »s
Landfill cap thickness, cm
Initial mass of constituent i
in the landfill, g
Molecular weight of constituent i,
g/g mol
Average molecular weight of the
dilute aqueous waste liquid,
g/g mol (assumed to be 18 g/g mol)
Molecular weight of the oil
carrier-liquid, g/g mol
Pure component vapor pressure of
constituent i, atm
Equilibrium partial pressure of
constituent i in the vapor space,
atm
Data source
Calculated
Literature
Calculated
Literature
Definition or
calculated
Literature
Estimated
Definition
or estimated
Literature
Calculated
P
ref
P,
Q
R
Initial (reference) barometric
pressure, mm Hg
Final barometric pressure after
•t, mm Hg
Average flow rate of gas from
landfill vent(s) due to baro-
metric pumping, cm /cm landfill
area's
3
Ideal gas constant, 82.05 cm »atm/
g mol»K
Time since landfill closure, mo
Meteorological
information
Meteorological
information
Calculated
Literature
Definition
(continued;
5-12
-------
TABLE 8-1 (continued)
Variable
Definition
Data source
Time interval used to determine
average barometric pumping rate, s
Temperature within landfill, °C
Definition
Estimated from
literature data
T
ref
Tl
Initial (reference) landfill
temperature, °C
Final landfill temperature after
• t, °C
Estimated from
literature data
Estimated
Total volume of gas exiting land-
fill in •t, cm
Calculated
V
X.
Total volume of void space within
waste, cm
Mass of original waste liquid in
a unit volume of fixed waste,
g/cm
Mole fraction of constituent i in
the aqueous liquid (for dilute
aqueous waste) or in the organic
phase (for two-phase or organic
liquid waste) (dimensionless)
Air porosity of the clay cap
(dimensionless)
Calculated
Definition or
estimated
Definition
Estimated from
clay property
data
Total porosity of the clay cap
(dimensionless)
Estimated from
clay property
data
f Air porosity of the fixed waste
(dimensionless)
Density of dilute aqueous waste-,
liquid (generally equals 1 g/cm )
Estimated from
fixed waste
property data
Definition or
liquid, g/cm
Exponential decay constant, mo
-1
Calculated
5-13
-------
waste liquid (X.). Because temperature affects volatility, the
model is sensitive to temperature. Other parameters to which the
model is sensitive include the depth of the fixed waste (D), the
air porosity of the fixed waste (• f ), the landfill surface area
(A), and the barometric pressure change (P ^ - P, ) . This latter
group of parameters is significant in that it impacts the
barometric pumping rate or the volume of gas available for
pumping. In contrast to these parameters, the model exhibits
relatively low sensitivity to the diffusivity of the constituent
in air (D .), the cap thickness (1), and the total mass of the
a i
constituent in the landfill (M .).
The major assumptions associated with the RTI closed
landfill model are as follows:
• The liquid waste containing the constituent of interest
is assumed to be bound in the fixed waste within the
landfill.
• The constituent of interest in the gas within the
landfill is assumed to be in equilibrium with the
liquid in the waste.
• Adsorption of the constituent of interest onto the soil
of the cap is assumed to be negligible.
• The fraction of air-filled space in the landfill cap
(air porosity) is assumed to remain relatively constant
over the long term.
• The effective diffusion coefficient of the cap is
assumed not to vary with either the concentration of
the constituent of interest or time.
• The concentration of the constituent of interest in air
at the top of the landfill cap is assumed to be
effectively 0.
• No biodegradation (with concurrent production of
biogas) is assumed to occur due to the suppression of
biological processes by the toxic waste.
• The landfill is assumed to be vented to the atmosphere.
The volume of gas available for barometric pumping is
assumed to consist of the total void-volume of the
waste bed.
5-14
-------
• No transport of the constituent of interest in moving
water is assumed to occur.
8.2.2 Model Plant Parameters for Closed Landfills
The characteristics of a model closed landfill facility are
discussed here. This model facility is used as the basis for an
example calculation in Section 8.2.3.
The model facility for closed landfills has an area of
8 9
1.417 x 10 cm (3.5 acres). This value represents an
approximately midrange value from the Westat survey. A
reasonable value of landfill depth, also selected from the Westat
survey, is 458 cm (15 ft). The landfill cap is assumed to be
composed of compacted clay. The cap thickness value of 107 cm
(3.5 ft) represents the average of extremes in thickness of clay
1 8
caps reported in site studies (2 ft to 6 ft). The value used
for air porosity of the clay cap is 0.08 (8 percent), while the
total porosity is 0.41 (41 percent). These values were computed
based on reasonable physical properties and level of compaction
19
for compacted clay. The landfill is assumed to be vented to
the atmosphere. The temperature beneath the landfill cap is
estimated at 15 °C, which represents the temperature of shallow
20
ground water at a midlatitude U.S. location. This temperature
is assumed to remain constant. The landfill is assumed to be
exposed to a nominal barometric pressure of 1,013 mbar, which
represents an estimate of the annual average atmospheric pressure
21
in the United States. Barometric pumping is estimated for the
landfill using a daily pressure drop from the nominal value of 4
mbar. The 4 mbar value represents an estimate of the annual
22
average diurnal pressure drop.
The model closed landfill facility is assumed to contain
fixed waste. The waste liquid (before fixation) selected for the
facility is assumed to be a two-phase aqueous/organic containing
20 percent chloroform, 20 percent low-volatility organic, {for
modeling purposes, this component of the waste liquid represents
the oil carrier-liquid} and 60 percent water (by weight). This
8-15
-------
liquid has an average density of 1.16 g/cm3. The fixation
industry indicates that waste liquid, when combined with
fixative, may in actuality increase in volume by as much as 50
23 24
percent. ' The volume change, which is a function of the
specific waste being fixed and the specific formulation of the
fixative, can only be determined experimentally. In view of the
inherent variability in the fixation process and the lack of real
data, for the purpose of this calculation the assumption is made
that the fixation process does not change the waste volume. This
assumption is environmentally conservative and may result in an
overestimation of the landfill emissions. Actual volume changes
that may take place as a result of fixation can easily be
accounted for because the change in the calculated emissions is
inversely proportional to the change in waste volume. One
industry contact indicated that, for the purposes of estimating
emissions, the assumption of no volume change during fixation was
25
reasonable. Based on the waste liquid density and the
assumption of no volume increase from fixation, the mass of waste
3
liquid in a unit volume of fixed waste is 1.16 g/cm . The air
porosity of the fixed waste (used to estimate the total volume of
gas available for barometric pumping) is taken to be 0.25 (25
percent). This value was inferred from measurements of total
O £.
porosity and moisture content of various fixed wastes, and, for
the purposes of this analysis, is assumed to pertain to waste
within the landfill as opposed to waste immediately following
fixation. As discussed previously, there is no evidence for
significant biomass in any chemical waste landfill. Therefore,
in this analysis it is assumed, as suggested in the literature,
that the toxic property of the waste will inhibit biological
m.
3
27
processes and thus prevent biogas generation. Hence, the waste
biomass concentration is taken to be 0 g/cm"
The properties of chloroform that are pertinent to this
analysis include the molecular weight (119.4 g/g mol), pure
component vapor pressure at 15 °C (0.162 atm), diffusivity in air
5-16
-------
at 15 °C (0.10 cm2/s) and density (1.49 g/cm3) . The low-
volatility organic liquid present in the waste has a molecular
3
weight of 147 g/g mol and a density of 1.31 g/cm .
Table 8-2 smmarizes the model facility parameters for closed
landfills used in the example calculation in Section 8.2.3. For
facilities that accept more than one type of waste, the weighted
average constituent concentrations may be used.
8.2.3 Example Calculation for Closed Landfill
This section presents a step-by-step calculation of
emissions from a closed landfill that is vented to the
atmosphere. The equations discussed in Section 8.2.1 and
summarized in Table 8-1 are used with the model unit parameters
in Section 8.2.2 to estimate emissions from a fixed, two-phase
aqueous/organic waste containing chloroform:
• Waste liquid 20 percent chloroform, 20
(before fixation): percent low volatility organic
liquid, 60 percent water (by
weight)
• Liquid/fixative: 1 unit volume liquid + dry
fixative = 1 unit volume fixed
waste
• Waste biomass concentration: 0 g/cm
8 2
• Landfill area: 1.417 x 10 cm (3.5 acres)
• Waste bed thickness: 457 cm (15 ft)
• Cap thickness: 107 cm (3.5 ft)
• Type landfill: vented
• Temperature beneath cap: 15 °C
• Time period for emission calculation: 1 yr.
5-17
-------
TABLE 8-2. INPUT PARAMETERS—CLOSED LANDFILL
Parameter
Value
Area
Waste bed thickness
Cap thickness
Cap air porosity
Cap total porosity
Type landfill
Temperature beneath cap
Typical barometric pressure
Daily barometric pressure drop
Waste liquid (before fixation)
Liquid composition
Liquid/fixative
2
'3.5 acres'
1.417 x 10 cm"
457 cm (15 ft)
107 cm (3.5 ft)
0.08 (8%)
0.41 (41%)
Vented
15 °C
1,013 mbar
4 mbar
Two-phase aqueous/organic
20% chloroform, 20% low-
volatility organic (oil),
water (by weight)
1 unit volume liquid + dry
fixative = 1 unit volume fixed
60%
Liquid in fixed waste
Air porosity fixed waste
Biomass concentration
Chloroform properties
Molecular weight
Vapor pressure (15 °C)
Diffusivity in air (15 °C)
Density
Low-volatility organic3 properties
Molecular weight
Density
waste
1.16 g/cm
0.25 (^5%
0 g/cm
3
119.4 g/g mol
0.162 a tin (123 mm Hg)
0.10 cm /s
1.49 g/cm
147 g/g mol
1.31 g/cm
3
Also referred to as oil "carrier-liquid."
5-18
-------
2
a. Compute the effective diffusion coefficient, D . (cm /s)
(Equation (8-3)): ei
D . = D . ('3'33/' 2)
ei ai a T
Dei = (0.10 cm2/s) (0.08)3'33/(0.41)2
D . = 1.32 x 10~4 cm2/s .
ei
b. Compute the equilibrium partial pressure of chloroform in
the vapor space, P1 (atm):
The waste before fixation was a two-phase liquid. Hence,
Raoult's law applies. The mole fraction for this case is
computed as:
Xi = (Ci/MWi)/(Ci/MWi + Coil/MWoil)
X± = (0.20/119.4 g/g mol) +
[0.20/119.4 g/g mol + 0.20/147 g/g mol]
X± =0.55
From Raoult's law (Equation (8-6)):
P. = X. P*.
i 11
P. = (0.55)(0.162 atm)
P. = 8.91 x 10~2 atm .
c. Compute the concentration of chloroform in the vapor space
beneath the cap, C . (g/cm void space) (Equation (8-4)):
o -L
C . = P.MW./R(T + 273)
si i i
(8.91 x 10~2 atm)(119.4 g/g mol)
\^ • Q
31 (82.05 cm •atm/g mol-K) (15 + 273)
C . = 4.50 x 10~4 g/cm3 .
si ^
d. Compute initial chloroform emission flux resulting from
2
diffusion through the cap only, J. (g/cm »s) (Equation 8-1) :
8-19
-------
J. = -D .(C0. - C
i ei 2i si1
J. = -(1.32 x 10 4 cm2/s)(0 g/cm3 - 4.50 x 10 4 g/cm3)
1 /107 cm
J. = 5.55 x 10 g/cm »s
e. Compute initial chloroform emissions resulting from
diffusion through the cap only, E., . (g/s) (Equation (8-2)) :
EI . = J. x A
EI:L = (5.55 x 10~10 g/cm2»s) (1.417 x 108 cm2)
EI:L = 7.86 x 10~2 g/s .
f. Estimate the barometric pumping-induced gas flow rate
through the vent(s):
1. Compute the volume of gas available for barometric
3
pumping, V (cm ) (Equation (8-8)):
V = D x A x •
c fw
V = (457 cm)(1.417 x 108 cm2)(0.25)
V =1.62xl010 cm3 .
c
2. Compute volume of gas exiting the vent due to barometric
pressure change, V (cm ) (Equation 8-9):
8-20
-------
+ P T + 273
IE xZ--f J-i T ^ ' J
= v * *
v
# # * -1-*-
* * P * *T ,+ 273* *
1 ref
For this case, T., = T ,- = 15 °C, and barometric
pressure drops by 4 mbar from the nominal value
of 1,013 mbar:
_ 1 co in10 3 *!' 013 mbar »» 15 °C + 273 K
VB 1>bZ 1U cm 1 1,009 mbar 15 °C + 273 K
¥„ = 6.42 x 107 cm3 .
3. Compute the average flow rate of gas over the time
inte
10) :
3 2
interval, Q, (cm /cm landfill area • s) (Equation 8-
The average diurnal pressure drop of 4 mbar occurs
within a 24-h period. For convenience, the gas flow
from this pressure change is averaged over 24 h (equals
8.64 x 104 s).
Q= VB
6.42 x 107cm3
i.64 x 104 s) (1.417 x 108 cm2)
Q = 5.25 x 10~6 cm3/cm2»s .
4. Compute the barometric pumping-induced emission rate,
(g/s) (Equation (8-7)):
E0 . = Q x C . x A
2i si
= (5.25 x 10~6 cm3/cm2»s) (4.50 x 10~4 g/cm3) (1.417 x 10 8 cm)2
8-21
-------
E2i = 0.335 g/s
*
g. Compute the total initial emission rate, E. (g/s)
(Equation 8-11):
E* = E,. + E0.
i li 2i
E* = 7.86 x 10~2 + 0.335
E* = 0.413 g/s
h. Compute the time-dependent instantaneous emission rate:
1. Compute total mass of constituent i in landfill, M .:
First compute W, the mass of original waste liquid in a
unit volume of fixed waste. Assuming one unit volume of
waste liquid results in one unit volume of fixed waste,
this parameter can be computed using the densities of
the waste liquid components and their weight fractions
as follows:
W = [(1.49 g/cm3)(0.2)+(1.31 g/cm3)(0.2)+(1 g/cm
3 3
x 1 cm liquid/cm fixed waste
= 1.16 g/cm
M . is then computed as:
01
M . = C. W A D
01 i
_ 20 g chloroform 1.16 g liquid - .-_ - Q8 2
oi ~ 100 g liquid X 3 ,. , . X 1-41/ X 1U Cm
^ ^ cm fixed waste
x 457 cm = 1.50 x 10 g chloroform
2. Compute the decay constant, • (mo ) (Equation (8-13)):
• = 2.63 x 106 E*/M .
i 01
• = (2.63 x 106) (0.413 g/s)/1.50 x 1010 g
-------
•=7.25xl05mo1.
3. Compute the instantaneous emission rate, E., in Mg/yr,
after 1 yr (Equation (8-12)): 1
E.(t) = 31.56 E* exp(-•t)
E .= (31.56)(0.413 g/s) exp(-7.25 x 10 ~5mo ~1x 12 mo;
E. = 13.0 Mg/yr
i. Compute the average emission rate in the first year, E .,
in Mg/yr (Equation (8-15)):
31.56 E*
EAi(t) = .t 1 *1 - e^*
E,. 01.56) (0.413 g/s) t^p,.^ mo x 7.25 x 105 mo1-
1 (7.25x10 mo ) (12 mo)'
EAi = 13-° Mg/yr
8.3 FIXATION PITS
This section is currently under review, (pages 25-34 not shown)
-------
8.4 OPEN LANDFILLS AND WASTEPILES
8.4.1 Emission Model Equations
34
The RTI land treatment model (discussed in Section 5.2)
is used to estimate the air emission rate of the constituent of
interest from open (active) landfills and wastepiles. This model
is based on the theory of diffusion out of an infinite flat slab
and was intended originally for use in estimating emissions from
land treatment operations. The intent of this section is to
discuss use of the model with regard to the estimation of
emissions from open landfills and wastepiles; a detailed
description of the model relevant to land treatment operations
and the theoretical basis for the model are presented in
Section 5.0 of this report and will not be repeated here.
A land-treatment-type model was selected for estimating
emissions from open landfills and wastepiles because (1) no
adequate models exist for these sources, and (2) there are a
number of similarities in physical characteristics of open
landfills, wastepiles, and land treatment operations. A previous
35
EPA study dedicated to the evaluation of models for estimating
emissions from hazardous waste TSDF identified only one model for
open waste dumps such as landfills and wastepiles--the open dump
model. A serious limitation of the model for this application,
however, is that it does not account for depletion of the
volatilizing chemical from the waste surface. Hence, the open
dump model is judged unsuitable for the estimation of emissions
from landfills and wastepiles over the time period of interest
(months or longer).
The similarity in physical characteristics among open
landfills, wastepiles, and land treatment operations is apparent
upon close examination--in all three, the waste liquid is
ultimately mixed homogeneously with a "carrier" matrix (soil in
the case of land treatment; dry fixative in the case of active
landfills; and soil, fixative, or some other solid matrix in the
case of wastepiles). In all cases, the matrix is porous and
3-35
-------
permeable, allowing the diffusion of the constituent of interest
through the matrix and into the air. Hence, in all cases,
diffusion theory can be used to model the emission rate. The
notable difference between land treatment operations and open
landfills/wastepiles is the presence of an additional mechanism
affecting emissions in the case of land treatment—biological
decay of the constituent. Because biodegradation is not thought
to occur. There is no evidence that there is significant biomass
(necessary for biological decay) in any chemical waste landfill.
It is assumed that the toxic property of the waste will inhibit
Q £.
biological processes. In open landfills/wastepiles, however,
its effect is not accounted for in the modeling of air emissions.
The RTI land treatment model, which was selected for
estimating emissions from open landfills and wastepiles, has the
following characteristics: a sound basis in scientific theory,
limited validation against measured emissions from land treatment
37
operations, and reasonably available input data. The model
considers effects such as evaporation of the constituent of
interest from interstitial surfaces of the carrier matrix and
diffusion of material through air-filled pore spaces.
The equations necessary to apply the land treatment model to
open landfills and wastepiles are summarized in Table 7-3. These
equations, explained in Chapter 7.0, can be used to estimate the
fraction of the constituent emitted (F, ) and the instantaneous
emission rate (E). It should be noted that the absence of
biodegradation represents a special case that allows some
simplification of several of the Chapter 7.0 equations, e.g.,
Equations (7-4) and (7-5). (The absence of biomass implies that
biomass concentration equals 0. Hence, t, , the time constant for
biological decay, equals infinity. Consequently, the exponential
term e b becomes unity.) Also, the absence of biodegradation
implies that the fraction of the constituent emitted after a long
time, F , would equal unity.
a
-------
Because the land treatment model was derived originally for
land treatment operations, model input parameters are not
3-37
-------
necessarily in the most convenient units and terminology for open
landfills and wastepiles. Hence, several points should be noted:
• Fixed waste is analogous (for modeling purposes) to the
waste-laden soil in land treatment.
2
• M , the area-loading of the constituent in g/cm , is
geared toward land treatment operations. For open
landfills and wastepiles, the landfill equivalent
should be computed.
• No "tilling" (as discussed in Chapter 7.0) is performed
in open landfills or wastepiles.
• Waste liquid is "applied" or mixed with fixative only
once. Hence, waste "reapplication" (used in the sense
discussed in Chapter 7.0) does not occur in open
landfills and wastepiles.
• The waste bed depth in open landfills and wastepiles is
analogous to the "depth to which waste is mixed" in
land treatment, as discussed in Chapter 7.0.
The approach required to estimate emissions from open
landfills or wastepiles is as follows, based on equations in
Table 7-3:
1. Compute the loading of waste liquid (L) in the fixative
or soil, using the known waste composition. (For two-
phase aqueous organics or organic liquid wastes, L
should be computed as grams organic phase per cubic
centimeter solid material. For dilute aqueous waste
liquids, L equals grams aqueous liquid per cubic
centimeter solid material.)
2. Compute the effective diffusion coefficient (D ).
3. Compute the partition coefficient (Keq).
4. Use the appropriate emission equation to compute the
fraction of constituent emitted (F,) and/or the
instantaneous emission rate (E). For wastepile
calculations, the time input to these equations should
be no greater than the life of the wastepile (retention
time).
1-31
-------
The sensitivity of the land treatment model to some
parameters differs in its application to open landfills and
wastepiles from that in land treatment operations because of the
difference (in some cases) in the expected range of the
parameters. In general, it can be stated that, for application
to open landfills and wastepiles, the model is sensitive to the
air porosity of the solid waste, the liquid loading in the solid
waste, the waste depth, the concentration of the constituent in
the waste, and the volatility of the constituent under
consideration. In contrast, the model exhibits a relatively low
sensitivity to the diffusion coefficient of the constituent in
air.
The following major assumptions are associated with the RTI
land treatment model and its application to open landfills and
wastepiles:
• The waste liquid is mixed uniformly with the carrier
matrix (either fixative, soil, or some other granular
solid material) before placement in an open landfill or
wastepile.
• The liquid waste containing the constituent of interest
is assumed to be bound in the waste after fixation and
placement in the open landfill or wastepile.
• The waste liquid does not flow within the carrier
matrix.
• The adsorption isotherm of the constituent of interest
is linear within the depth of the waste and does not
change with time.
• No bulk flow of gas is induced within the waste matrix.
• The diffusion coefficient does not vary with either
concentration or time.
• The concentration of the constituent of interest in the
gas phase at the surface of the open landfill/wastepile
is much lower than the concentration of the constituent
of interest in the gas phase within the waste matrix.
3-39
-------
• No diffusion of the waste liquid into depths below the
waste layer is assumed.
• Liquid-vapor equilibrium is established at all times
within the waste matrix.
• For the case of fixed waste in the landfill or
wastepile, the fixed waste mixture behaves as a soil
with regard to diffusion of the constituent of
interest.
• No biodegradation of the constituent of interest occurs
in open landfills or wastepiles.
8.4.2 Model Plant Parameters for Open Landfills and Wastepiles
The characteristics of model facilities for open landfills
and wastepiles are discussed here. The model open landfill
facility is used as the basis for an example calculation using
the model.
8.4.2.1 Parameters for open landfills. The model facility
8 9
for open landfills has a surface area of 1.42 x 10 cm (3.5
acres). This value represents an approximately midrange value
O Q
from the Westat survey. A reasonable value of landfill depth
39
from the Westat survey was 458 cm (15 ft). The model open
landfill is assumed to be half full, and hence has a waste depth
of 229 cm (7.5 ft). The landfill is assumed to contain fixed
waste. A standard temperature of 25 °C is assumed to apply.
The waste liquid (before fixation) selected for this model
facility is assumed to be a two-phase aqueous/organic containing
20 percent chloroform, 20 percent low-volatility organic, and 60
percent water (by weight). This liquid has an average density of
1.16 g/cm . The fixation industry indicates that waste liquid,
when combined with fixative, may increase in volume by up to 50
40
percent, depending on the specific combination of waste and
fixative. In view of the inherent variability in the fixation
process and the lack of real data on volume changes, for purposes
of this report, the assumption is made that the waste volume does
41
not change during fixation. Measurements performed on various
types of fixed waste yielded a broad range of total porosities.
8-40
-------
Fifty percent, as used in this study, is a reasonable estimate of
this parameter. These porosity values are assumed to be
representative of waste in an open landfill, rather than waste
that has recently undergone fixation. A 25-percent air porosity
appears to be a reasonable value; this value was inferred from
42
measurements of total porosity and moisture content. As
discussed previously, there is no evidence of significant biomass
in any chemical waste landfill. Therefore, in this analysis it
is assumed, as suggested in the literature, that the toxic
property of the waste will inhibit biological processes and thus
prevent biogas generation.43 Hence, the waste biomass
concentration is taken to be 0 g/cm .
The properties of chloroform that are pertinent to this
analysis include the molecular weight (119.4 g/g mol), pure
component vapor pressure (208 mm Hg), and diffusivity in air
2
(0.104 cm /s). The low-volatility organic liquid present in the
waste has a molecular weight of 147 g/g mol.
Table 8-7 summarizes the model facility parameters for open
landfills used in the example calculation in Section 8.4.3.
8.4.2.2 Parameters for wastepiles. A review of information
44
in the Westat survey led to the selection of an approximately
/~ O
midrange value for basal area of 4.65 x 10 cm . For modeling
purposes, the pile is assumed to be flat. A uniform height of
100 cm was inferred, using the Westat information and engineering
judgment. All waste ultimately disposed of in the landfill is
assumed to be stored initially in the wastepile. The open
landfill described previously (Section 8.4.2.1) is assumed to be
filled to capacity in 1 yr. Based on the filled landfill volume
Q O 103
(1.42 x 10 cm x 458 cm depth = 6.50 x 10 cm ), the wastepile
volume (4.65 x 10 cm x 100 cm =4.65 x 10 cm), and the
filling time of 1 yr, it can be concluded that the wastepile
undergoes a turnover rate of 140 turnovers/yr. Hence, the
wastepile retention time is 2.6 d/turnover.
5-41
-------
TABLE 8-7. INPUT PARAMETERS OPEN LANDFILL
Area
Waste depth (half full)
Volume
Temperature
Waste liquid (before fixation)
Liquid composition
Liquid density (average]
Liquid/fixative
1.42°108 cm2 (3.5 acres)
229 cm (7.5 ft)
3.25 1010 cm3
25 degrees C.
Two-phase awueous/organic
20 percent chloroform, twenty
percent low-volatility organic,
and 60 percent water.
1.16 g/cm3
1 unit volume liquid + dry
fixative = 1 unit volume fixed
waste .
Air porosity fixed waste
Total porosity fixed waste
Biomass concentration
0.25 (25 percent)
0.50 (50 percent)
0 g/cm3
Chloroform properties
Molecular weight
Vapor pressure
Diffusivity in air
119.4 g/g mol
208 mm Hg
0.104 cm2/s
Low-volatility organic properties
Molecular weight 147 g/g mol
3-42
-------
The properties of the waste liquid and the resulting fixed
waste accommodated by the model wastepile are identical to those
for the open landfill (Section 8.4.2.1) and will not be repeated
here. Table 8-8 summarizes the model facility parameters used
for wastepiles.
8.4.3 Example Calculation for Open Landfill
This section presents a step-by-step calculation of emissions
from an open landfill. The equations identified in Table 7-3 are
used with the model unit parameters in Section 8.4.2 to estimate
emissions from a fixed, two-phase aqueous/organic waste
containing chloroform; the same equations would be applied to the
estimation of emissions from wastepiles:
• Waste liquid (before fixation): 20 percent chloroform,
20 percent low-
volatility organic
liquid, 60 percent
water
• Liquid/fixative: 1 unit volume liquid + dry fixative =
1 unit volume fixed waste
• Waste biomass concentration: 0 g/cm
8 2
• Landfill area: 1.42 x 10 cm (3.5 acres)
• Landfill depth: 229 cm (7.5 ft)
• Temperature: 25 °C
• Time period for emission calculation: 3.15 x 10 s (1
yr) .
a. Compute waste loading, L:
Liquid density before fixation = 1.16 g/cm
1 cm liquid waste + fixative = 1 cm fixed waste
L = g organic phase/cm fixed waste
= (0.20 + 0.20) x 1.16 g/cm3 = 0.46 g/cm3 .
(Note that weight fraction of chloroform in the oil
phase [C] = 0.2/(0.2 + 0.2) = 0.50.)
3-43
-------
TABLE 8-8. INPUT PARAMETERS WASTEPILES
Surface Area
Average height
Turnover rate
Retention time
Temperature
Windspeed
Waste liquid
Liquid composition
Liquid density (average]
Liquid/fixative
Air porosity fixed waste
Total porosity fixed waste
Biomass concentration
Chloroform properties
Molecular weight
Vapor pressure
Diffusivity in air
cm
C.
4.65°10f
100 cm
139/yr
2.6 days
25 degrees
4.92 m/s
Fixed waste
20 percent chloroform, 20 percent
low-volatility organic, and 60
percent water.
1.16 g/cm3
1 unit volume liquid + dry
fixative = 1 unit volume fixed
waste.
0.25 (25 percent)
0.50 (50 percent)
0 g/cm3
119.4 g/g mol
208 mm Hg
0.104 cm2/s
Low-volatility organic properties
Molecular weight
147 g/g mol
3-44
-------
b. Compute effective diffusion coefficient for fixed waste:
3.33
D = D
where
a . 2
T
• = air porosity fixed waste = 0.25
a
• = total porosity fixed waste = 0.50.
Then
D = diffusivity of chloroform in air =
cL
D = 0.104 cm2/s
cL
D = (0.104 cm2/s) (°'25) =-
6 (0.50)Z
D = 4.11 x 10~3 cm2/s .
e
(Note: D /D = 3.96 x 10~2.)
e a
c. Compute "partition" coefficient, Keq (ratio of gas-
phase chloroform to total chloroform in the waste):
For oily waste,
P*MW . -, •
T, o 11 a
Keq = R T j-
where,
*
P = pure component vapor pressure of chloroform =
(208 mm Hg)/ (760 mm Hg/atm) = 0.274 atm
MW . , = molecular weight low-volatility organic = 147
g/g mol
R = ideal gas constant = 82.05 cm »atm/g mol»K
T = temperature within solid waste, K
T = 273 K + 25 °C = 298 K
3-45
-------
Keq= (0.274 atm) (147 g/g mol) (0.25)
(82.05 cm3»atm/g mol»K) (298 K) (0.46 g/cm
Keq = 8.95 x 10~4
d. Compute fraction of total chloroform emitted, F, , after
1 year: First, determine which solution applies by
computing from the equations in Table 7-3.
KeqiDe 8.95 x 10~4 x 4.11 x 10~3 cm2/s
I2 (229 cm)2
= 7.01 x 10"11 s"1
Therefore,
Keq D t _ _
= 7.01 x 10 s x 3.15 x 10 s
=
= 2.21 x 103
Keq D t 2 _
p • — 3
K t = _ ^— - = 5 45 (in }
L\ , l_ ,p , *J . "~t *J \ -L U I
2
Because Keq D t/1 is less than 0.25, the first equation of
Table 7-3 applies, and
Ft = 0.72 (Kdt)
Ft = 0.72 (5.45 x 10~3) 1/2
Ft = 0.053 .
e. Compute instantaneous emission rate, E, after 1 yr :
1. Compute initial mass of chloroform in landfill:
M . = 1 L C
01
where
1 = waste depth = 229 cm
3-46
-------
3 3
L = g organic/cm fixed waste = 0.46 g/cm
C = weight fraction chloroform in oil = 0.50.
Then
M = (229 cm) (0.46 g/cm3) (0.50)
2
M =52.7 g/cm
o
3. Compute instantaneous emission rate, E.. Because Keq
2
D t/1 0.213,
emission rate:
2
D t/1 0.213, use the following equation to compute the
= Mo _ 1
* kQ Keq + Keq Dg *
kp = 4.82 (10-3) U°'78 Sc -°'67 de-0'11
(j (j
U = windspeed = 4.92 m/s
47V °'5
de = effective diameter of landfill area =*— * = 134 m
Sc = ug
bCG • D
a a
where: ug = viscosity of air = 1.8 (10 ) g/cm/s.
-3 3
• = Density of air = 1.2 (10 ) g/cm .
a „
D = 0.104 cm /s.
cL
Scp = ' - = 1.45
1.2 (10 ) (0.104)
= 4.82 (10~3) (4.92)0'78 (1.45)~°'67 (134)"0'11
= 0.0076 m/s = 0.76 cm/s
3-47
-------
E =
52.7
229
4.43
1
0.25
(^8^cM?-;s.\
F ~ 0 230
i
368 + 5.18106
3.14 x 3.15 107
8.95 10~4 x 4.11 10~3
-------
8.5 REFERENCES
1. GCA Corporation. Air Emissions from Land Treatment-
Emissions Data and Model Review. Draft Technical Note.
Prepared for U.S. Environmental Protection Agency. Research
Triangle Park, NC. August 1985. 120 pp.
2. Reference 1.
3. Farmer, W. J., M. S. Yang, J. Letey, W. F. Spencer, and M.
H. Roulier. Land Disposal of Hexachlorobenzene Wastes:
Controlling Vapor Movement in Soils. Fourth Annual Research
Symposium. U.S. Environmental Protection Agency.
Publication No. EPA-600/9-78-016. August 1978.
4. Millington, R. J., and J. P. Quirk. Permeability of Porous
Solids. Trans. Faraday Society. 5_7_: 1200-1207. 1961.
5. U.S. Environmental Protection Agency. Evaluation and
Selection of Models for Estimating Emissions from Hazardous
Waste Treatment, Storage, and Disposal Facilities. Office
of Air Quality Planning and Standards, Research Triangle
Park, NC. Publication No. EPA-450/3-84-020. December 1984.
6. Reference 3.
7. Thibodeaux, L. J. Estimating the Air Emissions of Chemicals
from Hazardous Waste Landfills. Journal of Hazardous
Materials. 4:235-244. 1981.
8. Reference 3.
9. Reference 3.
10. Shen, T. T. Estimating Hazardous Air Emissions from
Disposal Sites. Pollution Engineering. 31-34. August
1981.
11. Reference 3.
12. Reference 4.
13. Reference 3.
14. Westat, Inc. National Survey of Hazardous Waste Generators
and TSD Facilities Regulated Under RCRA in 1981. Prepared
for U.S. Environmental Protection Agency. Contract No. 68-
61-6861. 1981.
15. Reference 14.
3-49
-------
16. Reference 14.
17. Reference 14.
18. Ely, R. L., G. L. Kingsbury, M. R. Branscome, L. J. Goldman,
C. M. Northeim, J. H. Turner, and F. 0. Mixon, Jr.
Performance of Clay Caps and Liners for Disposal Facilities.
Research Triangle Institute, Research Triangle Park, NC.
Prepared for U.S. Environmental Protection Agency,
Cincinnati, OH. EPA Contract No. 68-03-3149. March 1983.
19. Telecon. Goldman, Leonard, Research Triangle Institute,
with Borden, Roy, Department of Civil Engineering, North
Carolina State University, Raleigh, NC. August 13, 1986.
20. Gerachty, J. J., D. W. Miller, F. Vander Leeden, and F. L.
Troise. Water Atlas of the United States. Water
Information Center, Inc., Port Washington, NY. 1973. Plate
30.
21. Telecon. Goldman, Leonard, Research Triangle Institute,
with Hughes, John, National Climatic Center, Asheville, NC.
May 15, 1986.
22. Reference 21.
23. Telecon. Goldman, Leonard, Research Triangle Institute,
with Boyenga, Dave, MBI Corporation, Dayton, OH, November
20, 1985.
24. Telecon. Goldman, Leonard, Research Triangle Institute,
with Webber, Emlyn, VFL Technology Corporation, Malvern, PA.
November 12, 1985.
25. Telecon. Massoglia, Martin, Research Triangle Institute,
with Webber, Emlyn, VFL Technology Corporation, Malvern, PA.
January 13, 1987.
26. Telecon. Goldman, Leonard, Research Triangle Institute,
with Hannak, Peter, Alberta Environmental Center. April 4,
1986.
27. Reference 10.
28. Reference 5.
29. Arnold, J. H. Studies in Diffusion: III. Unsteady-State
Vaporization and Absorption. Transactions of the American
Institute of Chemical Engineers. ^0:361-379. 1944.
30. Reference 29.
3-50
-------
31. Reference 29.
32. Letter and attachment from Shen, T., New York State
Department of Environmental Conservation, to McDonald, R.,
EPA/OAQPS. Modification of Arnold's equation. February 16,
1986. (See also Reference 10.)
33. Trip Report. Goldman, Leonard, Research Triangle Institute,
with Chemical Waste Management, Sulphur, Louisiana.
February 25, 1986.
34. Reference 1.
35. Reference 5.
36. Reference 10.
37. Memorandum and attachment from M. Wright, Research Triangle
Institute, to S. Thorneloe, EPA/OAQPS. May 30, 1986.
Selection of an emissions model for land treatment.
38. Reference 14.
39. Reference 14.
40. References 23, 24, and 25.
41. Reference 26.
42. Reference 26.
43. Reference 10.
44. Reference 14.
5-51
-------
9.0 TRANSFER, STORAGE, AND HANDLING OPERATIONS
9.1 NARRATIVE DESCRIPTION OF MODEL PLANTS AND EMISSIONS
This chapter presents models for estimating VO emissions of
hazardous wastes from container loading, storage, and cleaning;
stationary tank loading and storage; spills; fugitive sources;
and vacuum truck loading.
9.2 CONTAINER LOADING
This section addresses emission-estimating practices for
hazardous waste loading into tank trucks, railroad tank cars,
marine vessels, and 55-gal (0.208-m ) drums.
9.2.1 Emission Model for Container Loading
The process of loading containers with volatile hazardous
wastes generates emissions. If containers were assumed to be
clean when loaded, only those vapors generated by the loaded
waste could be emitted. If containers hold residues of previous
volatile wastes, additional emissions from the residues will
also be generated.
To calculate container loading losses, the AP-42 equation
for loading petroleum liquids is applied. This equation was
derived for tanks, cars, and marine vessels. It is also applied
to 55-gal drums in this case because the loading principles are
similar and because no equation has been developed exclusively
for small containers such as drums. The loading equation is as
follows:
12.46 SMP*
where,
9-1
-------
3
L = loading loss, lb/10 gal of liquid loaded;
j_i
M = molecular weight of vapors, Ib/lb mol;
*
P = true vapor pressure of liquid loaded, psia; and
T = bulk temperature of liquid loaded, °R (°F + 460) .
S = saturation factor, dimensionless (see Table 9-1).
Equation (9-1) for estimating emissions from containers is not
applicable to open dumpsters because they are designed with no
tops, unlike drums that have limited venting through bungs.
9.2.2 Model Parameters
Containers are considered to be splash-loaded (as opposed to
submerged-loaded) for emission-estimating purposes. This loading
method creates larger quantities of VO vapors and increases the
saturation factor, S, of each volatile compound within the
container. A saturation factor is a dimensionless quantity that
"represents the expelled vapors' fractional approach to
saturation and accounts for the variations observed in emission
2
rates from the different unloading and loading methods. A
saturation factor of 1.45 was selected for these emission
estimates, based on previous documentation of splash-loading
petroleum liquids. '
Typical capacities for containers are assumed to be as
follows:
• Drums: 55 gal (0.208 m3)
• Tank trucks: 7,000 gal (26.5 m )
• Railroad tank cars: 30,000 gal (114 m3)
• Marine vessels: 20,000 tons.
It is assumed that 55-gal drums and tank trucks are the
predominant containers used in the waste management industry.
Bulk liquid hazardous waste is shipped predominantly by highway;
therefore, hazardous waste tank truck models are used for
estimating emissions.
9-2
-------
TABLE 9-1. FACTORS FOR CALCULATING PETROLEUM LOADING LOSSES
Cargo carrier
Mode of operation
s factor
Tank trucks and tank cars
Marine vessels5
Submerged loading of a
clean cargo tank 0.50
Splash loading of a clean 1.45
cargo tank
Submerged loading: normal 0.60
dedicated service
Splash loading: normal
dedicated service 1.45
Submerged loading:
dedicated 1.00
vapor balance service
Splash loading: dedicated 1.00
vapor balance service
Submerged loading: ships 0.2
Submerged loading: barges 0.5
aTo be used for products other than gasoline.
9-3
-------
Molecular weight and vapor pressure are functions of the waste
loaded, and 25 °C is considered an annual average ambient
operating temperature.
9.2.3 Sample Calculation for Tank Loading
The following sample calculation may be used to estimate VO
emissions from tank truck loading of an organic liquid. Waste
stream compounds and properties for the sample calculation are as
follows. The same waste stream is employed in each sample
calculation in this section; only the type of emission source is
varied.
Constituent
Benzene
Naphthalene
Phenol
Weight
fraction
0.3
0.3
0.4
Molecular
weight
(Ib/mol)
78
128
94
Vapor
pressure
(psia)
1.84
0.0044
0.0066
Mole
fraction
0.368
0.224
0.408
The input parameters for the truck loading model are as follows
• True vapor pressure of loading liquid, psia: 0.68
(calculated in a., below)
• Molecular weight of vapor, Ib/mol: 78.23
(calculated in b., below)
• Saturation factor, dimensionless: 1.45 (splash
loading)
• Bulk temperature of liquid loaded, °R: 537
• Annual throughput, gal/yr: 28,000
*
a. Calculate P , true vapor pressure of liquid, by Raoult's
Law:
9-4
-------
p* = (PI • x ^ + (?2 • x ^ + (P 3» x ^
where, ^
P = true vapor pressure, psia;
P, , Pp, and P~, = vapor pressures of pure components;
X1, Xp, and X., = mole fractions of VO components in liquid;
P* = (1.84 psia x 0.368) + (0.0044 psia x 0.224;
+ (0.0066 psia x 0.408); and
P = 0. 68 (psia) .
b. Calculate M, molecular weight of vapors:
I P «Y 'I I P «Y 'I I P «Y 'I
\ r-, "A -,; ( c p"A A ( c .5A i
M = , • M1 + , • M 4- j • M _
^ X ")
P P P
where
M = molecular weight of vapor
MI, Mp, and M^ = molecular weight of each component
«1.84 x 0.368 ^ ^ 0.0044 x 0.224,
0.68 X 0.68
i OQ _L =,0.0066 x 0.408* n.
x 128 + * :r—^o * x 94
= 78.23 (Ib/mol).
c. Calculate emissions from truck loading:
T 12.46 SMP*
LL = T
12.46 x 1.45 x 78.23 x 0.68
537 °R
1.79 lb/1,000 gal
, . . T 1.79 x 10~3 Ib/gal x 28,000 gal/yr
Annual emissions, LT = ^—^-^—n -, ,,,
L 2,205 Ib/Mg
0.023 Mg/yr
9-5
-------
9.3 CONTAINER STORAGE
This section addresses storage emissions from tank trucks,
railroad tank cars, 55-gal drums, marine vessels, and open
dumpsters.
9.3.1 Emission Model for 55-Gal Drums, Tank Trucks, and Railroad
Tank Cars
With regard to 55-gal drums, container storage is considered
a location where multiple drums are most likely to accumulate and
be stored for more than 90 days. Because drums are designed to
be stored with a sealed lid and bung, the potential for breathing
losses is minimal. Therefore, breathing loss is assumed to be
negligible. However, the potential does exist for a drum to
rupture or become damaged and leak during storage. Thus, the
emissions from drum storage may be estimated using the same spill
-4
fraction used for drum handling--10 (to be discussed in more
detail in Section 9.7, Spills). The following equation is used
to estimate emissions from drum storage:
E = 10~4 x I x W. x V. (9-2)
11
where,
E = emission from drum storage, Mg/yr;
I = throughput, Mg/yr;
W. = VO weight fraction; and
V. = volatilization fraction.
Emission-estimating methodologies have not been developed
for storage in tank trucks and railroad tank cars. Only loading
information was available in the literature for these containers.
The assumed same emission estimates principle for drum storage is
: c
6
-5
applied with an emission factor of 10 (to be discussed in more
detail in Section 9.7, Spills;
9-6
-------
9.3.2 Model Parameters for Drum Storage
It is assumed that 50 percent of the VO storage loss from
drum spill or rupture will be volatilized. The remaining
volatiles will be captured with RCRA spill response measures
taken at the facility.
9.3.3 Sample Calculations for Drum Storage
Input parameters:
Waste stream: organic liquid
(See Section 9.2.3 for constituents.)
3
Waste density: 1.04 Mg/m
Drum storage capacity: 182 drums (0.208 m /drum)
Turnovers per year: 12
Spill fraction: 10~4
Weight fraction: 1
Volatilization fraction: 0.5.
a. Calculate annual throughput, I, Mg/yr:
I = 182 x 0.208 m3 x 12 x 1.04 Mg/m3
= 472 Mg/yr
b. Calculate air emissions:
E = 10~4 x I x W. x V.
i i
= 10~4 x 472 Mg/yr x 1 x 0.5
= 0.024 Mg/yr
9.3.4 Emission Model for Open Dumpsters
Open dumpsters are used to contain waste materials.
Volatile organics can diffuse from this waste material to the
surface of the waste, where they can be emitted to the
atmosphere. Wastes held in dumpsters may range from sludges to
contaminated filters. Because an open dumpster is similar to
open landfills and wastepiles, the land treatment model is used
to estimate the air emission rates of the constituent of interest
from open landfills, wastepiles, and open dumpsters. Chapter 7.0
9-7
-------
describes the use of the land treatment model for open landfills
and wastepiles. The theoretical basis for the land treatment
model is presented in Chapter 5.0 of this report and will not be
repeated here.
A land-treatment-type model was selected for estimating
emissions from open dumpsters because (1) no adequate models
exist for this source, and (2) there are a number of similarities
in physical characteristics of the wastes in open dumpsters and
land treatment operations. A previous EPA study dedicated to
the evaluation of models for estimating emissions from hazardous
waste TSDF identified only one model for open waste dumps such as
landfills and wastepiles--the open dump model. A serious
limitation of the open dump model for this application, however,
is that it does not account for depletion of the volatilizing
chemical from the waste surface. Hence, the open dump model is
judged unsuitable for the estimation of emissions from open
dumpsters over the time period of interest (months or longer).
The similarity in physical characteristics between open
dumpsters and land treatment operations is apparent upon close
examination--in both the waste liquid is ultimately mixed
homogeneously with a "carrier" matrix (soil in the case of land
treatment; waste in the case of open dumpsters). In both cases,
the matrix is porous and permeable, allowing the diffusion of the
constituent of interest through the matrix and into the air.
Hence, in both cases, diffusion theory can be used to model the
emission rate. The notable difference between land treatment
operations and open dumpsters is the presence of an additional
mechanism affecting emissions in the case of land treatment--
biological decay of the constituent. Because biodegradation is
not thought to occur in open dumpsters, however, its effect is
not accounted for in the modeling of air emissions.
The RTI land treatment model, which was selected for
estimating emissions from open dumpsters, has the following
characteristics: a sound basis in scientific theory, limited
-------
validation against measured emissions from land treatment
operations, and reasonably available input data.8 The model
considers effects such as evaporation of the constituent of
interest from interstitial surfaces of the carrier matrix and
diffusion of material through air-filled pore spaces.
The equations necessary to apply the land treatment model to
open dumpsters are summarized in Table 7-3. The equations,
presented in Chapter 7.0, can be used to estimate the fraction of
the constituent emitted (F, ) and the instantaneous emission rate
(E). It should be noted that the absence of biodegradation
represents a special case that allows some simplification of
several of the Chapter 7.0 land treatment equations, e.g.,
Equations (7-4) and (7-5). (The absence of biomass implies that
biomass concentration equals 0. Hence, t, , the time constant for
biological decay, equals infinity. Consequently, the exponential
term e b becomes unity.) Also, the absence of biodegradation
implies that the fraction of the constituent emitted after a long
time, F , would equal unity.
a
Because the land treatment model was derived originally for
land treatment operations, model input parameters are not
necessarily in the most convenient units and terminology for open
dumpsters. Hence, several points should be noted:
• The dumpster waste is analogous (for modeling purposes)
to the waste-laden soil in land treatment.
2
• M , the area-loading of the constituent in g/cm , is
compatible with land treatment operations. For open
dumpsters, it should be computed from the concentration
in the waste, M = 1 LC.
• No "tilling" (as discussed in Chapter 7.0) is performed
in open dumpsters, but the initial mixing upon dumpster
loading is equivalent to a tilling.
• Waste liquid is generally not "applied" or mixed with
the solids in the dumpster. Hence, waste
"reapplication" (used in the sense discussed in
Chapter 7.0) does not occur in open dumpsters. If
liquid wastes are applied to the surface of the waste
9-9
-------
in the open dumpster, then the land treatment
application model would be used.
• The waste bed depth in open dumpsters is analogous to
the "depth to which waste is mixed" in land treatment,
as discussed in Chapter 7.0.
The approach required to estimate emissions from open
dumpsters is as follows, based on equations in Table 7-3:
1. Compute the loading of waste liquid (L) in waste, using
the known waste composition. (For two-phase aqueous
organics or organic liquid wastes, L should be computed
as grams organic phase per cubic centimeter solid
material. For dilute aqueous waste liquids, L equals
grams aqueous liquid per cubic centimeter waste.)
2. Compute the effective diffusion coefficient (D ). If
the sludge is not porous, the effective diffusion
coefficient (D ) equals the product of the total
porosity (or volume fraction liquid) and the liquid
diffusion coefficient (D ).
j_i
3. Compute the partition coefficient (Keq).
4. Use the appropriate emission equation to compute the
fraction of constituent emitted (F,) and/or the instan-
taneous emission rate (E). For dumpster calculations,
the time input to these equations should be no greater
than the holding time of the waste in the dumpster.
The sensitivity of the land treatment model to some
parameters differs in its application to open dumpsters from that
in land treatment operations because of the difference (in some
cases) in the expected range of the parameters. In general, it
can be stated that, for application to open dumpsters, the model
is sensitive to the air porosity of the solid waste, the liquid
loading in the solid waste, the waste depth, the concentration of
the constituent in the waste, and the volatility of the
constituent under consideration. In contrast, the model exhibits
a relatively low sensitivity to the diffusion coefficient of the
constituent in air.
The following major assumptions are associated with the RTI
land treatment model and its application to open dumpsters:
9-10
-------
• The waste liquid is mixed uniformly in the waste solids
in an open dumpster and does not drain to the bottom of
the dumpster.
• The absorption of the volatile constituent into the
waste liquids is linear within the depth of the waste
and does not change with time.
• No bulk flow of gas is induced within the waste matrix.
• The diffusion coefficient does not vary with either
concentration or time.
• The concentration of the constituent of interest in the
gas phase at the surface of the open dumpster is much
lower than the concentration of the constituent of
interest in the gas phase within the waste matrix.
• The waste liquid does not migrate.
• Liquid-vapor equilibrium is established at all times
within the waste matrix.
• No biodegradation of the constituent of interest occurs
in dumpsters.
9.3.5 Model Parameters for Open Dumster Storage
The input parameters required for the model are divided into
three groups:
• Meteorological conditions. An average annual ambient
temperature of 25 °C and an average windspeed of 447
cm/s were used.
3
• Size of the dumpster. A 4-yd uncovered dumpster with
the following dimensions was used:
Length = 1.85 m (73 in)
Width = 1.45 m (57 in)
Height = 1.22 m (48 in)
• Properties of waste stored. These properties include
molecular weight, vapor pressure, and diffusivity in
air. The properties of specific compounds were
obtained from literature sources.
9-11
-------
9.3.6 Sample Calculation for Open Dumpster Storage
This section presents a step-by-step calculation of
emissions from an open dumpster. The equations identified in
Table 7-3 are used with the dumpster input parameters to estimate
emissions from an organic waste containing benzene, naphthalene,
and phenol.
Input parameters:
Waste stream: organic liquid
VO constituents = benzene, naphthalene, and phenol
Concentration of each constituent in oil, ppmw = 50,000
Initial VO amount, kg = 122.7 (total for three components)
U, windspeed, cm/s = 447
1, waste depth in the dumpster, cm = 122
2
area of the dumpster, m =2.68
molecular weight of oil = 282
Total porosity of waste = 0.5
Air porosity of waste = 0.25
T, ambient temperature, K = 298
Dumpster turnovers per year = 2
a. Compute waste loading, L:
The waste is 50 percent by volume solids, 25 percent by
volume air, and 25 percent by volume organic oil. At an
estimated density of 1 g/cm of oil, the waste loading
is 0.25 g oil/cm .
b. Compute effective diffusion coefficient for fixed waste:
D = D
3.33
a
e a . 2
* T
where
= air porosity fixed waste = 0.25
T
a
= total porosity fixed waste = 0.50.
9-12
-------
Then
2
D = diffusivity of benzene in air = 0.088 cm /s
3.
9 Id OR1)
D = (0.088 cmZ/s) ( ' '
(0.50)
D = 3.48 x 10~3 cm2/s. (Note: D /D = 3.96 x 10~2.)
e e 3.
c. Compute "partition" coefficient, Keq (ratio of
gas-phase benzene to total benzene in the waste) :
For oily waste,
P*MW . .. •
o 11 a
Keq = ITf L
where
P = pure component vapor pressure of chloroform
= (95.2 mm Hg)/ (760 mm Hg/atm) = 0.125 atm
MW ., = molecular weight low-volatility organic
= 282 g/g mol
3
R = ideal gas constant = 82.05 cm »atm/gmol»K
T = temperature within solid waste, K
T = 273 K + 25 °C = 298 K
Keq = (0.125 atm) (282 g/g mol) (0.25)
(82.05 cm3»atm/g mol»K) (298 K) (0.25 g/cm3 )
Keq = 0.00144
d. Compute fraction of benzene emitted, F,, after
1/2 year:
Keq D t
Determine which solution applies by computing —
2
9-13
-------
;Table 7-3):
KeqiDe 0.00144 x 0.00348 cm2/s
I2 (122 cm)2
= 3.37 • 10 10 s
Therefore,
Keq D t _ _
=-^- = 3.37 • 10 s x 1.57 x 10 s
1 = 0.0053 .
Keq D t 2
Kdt = —^- '— = 0.013 .
2
Because Keq D t/1 is less than 0.25, the third
equation of Table 7-3 applies, and
Ft = 0.72 (Kdt)
Ft = 0.72 (0.013)1/2
Ft = 0.082 .
e. Estimate annual emissions, Mg/yr:
kg
Benzene
Naphthalene
Phenol
Total
4
4
4
0
0
0
122
.9
.9
.9
.7
Init
0
0
0
0
Fraction
. VO,
to air
.082
.003
.005
.090
Annual en
kg/turnover
3
0
0
3
.37
.13
.19
.69
missions
kg/yr
6
0
0
7
.74
.26
.38
.38
9.4 CONTAINER CLEANING
9.4.1 Emission Model for Container Cleaning
9
An AP-42 document on tank truck cleaning is used as the
primary source for container-cleaning emission estimates. AP-42
states that tank truck cleaning typically involves washing the
9-14
-------
truck interior with agents such as water, steam, detergents, or
other chemicals. The document also provides emission factors
that are a function of vapor pressure and viscosity. These
factors have been applied to emission estimates for cleaning all
types of containers, as follows:
Tank truck Emission factor
residue to be removed g/truck (Ib/truck)
High vapor pressure, low viscosity 215 (0.474)
Medium vapor pressure, medium viscosity 32.4 (0.071)
Low vapor pressure, low viscosity 5.5 (0.012)
The following equation is used to estimate emissions for
container cleaning:
E=F xN x W. x 10~6 (9-3)
c i
where
E = cleaning loss, Mg/yr
F = emission factor for cleaning, g/container
N = number of cleanings per year
W. = VO weight fraction.
9.4.2 Model Parameters
In all containers, the primary input parameter for
estimating cleaning emissions is the cleaning emission factor,
which is determined from (1) residue vapor pressure and
viscosity (functions of waste handled), and (2) container
volume.
10 3
Based on AP-42, a typical tank truck volume of 26.5 m
(7,000 gal) is assumed.
Because no data are currently available for drum cleaning,
the emission factors for tank trucks were used to calculate
cleaning emissions from drums by comparing the proportion of
drum volume (55 gal) to that of the tank truck (7,000 gal) .
9-15
-------
[-[-„-,, Emission factor
55-Gal drum
residue to be removed g/drum (Ib/drum)
High vapor pressure, low viscosity 1.69 (0.0037)
Medium vapor pressure, medium viscosity 0.25 (0.0006)
Low vapor pressure, low viscosity 0.04 (0.00009)
Emissions from marine vessels have not been addressed because
of the low usage of such vessels in the waste management
industry.
9.4.3 Sample Calculation for Tank Truck Cleaning
The general assumptions for truck cleaning are as follows:
• Residue: pure organic liquid (benzene)
• Number of truck cleanings per year: 4
• Truck capacity: typical truck
• Weight fraction: 1.
a. Determine the cleaning emission factor, F :
(215 g/truck was used because of high vapor pressure and
low viscosity of pure benzene residue).
b. Calculate cleaning emissions:
E=F xNxW. x 10~6
c i
= 215 g x 4 x 1 x 10~6 Mg/g
= 8.6 x 10~4 Mg/yr
9.5 STATIONARY TANK LOADING
9.5.1 Emission Model for Stationary Tank Model
Stationary tank working losses are those emissions from
waste loading and unloading operations. AP-42's "Storage of
Organic Liquids" provides an equation to estimate loading and
9-16
-------
unloading emissions from storage tanks. The equation was
developed for handling VO liquid in the following industries:
• Petroleum producing/refining
• Petrochemical and chemical manufacturing
• Bulk storage and transfer operations
• Other industries consuming or producing organic
liquids.
Because hazardous wastes have the potential to contain VO
compounds, as do organic liquids, and because they are most
commonly stored in the same fashion as these liquid products,
the following equation was selected from AP-42:
j = 2.40 x!05M • P* • V • N • K • K (9-4'
w v n c
where
L = working losses, Ib/yr
w
M = molecular weight of vapor in storage tank, Ib/lb
mol
*
P = true vapor pressure at bulk liquid conditions,
psia
N = number of turnovers per year (dimensionless)
„ _ total throughput per year (gal)
tank capacity, V (gal)
V = tank capacity, gal
K = turnover factor, dimensionless (for turnovers
n 180 + N
for turnovers > 36, K = rvr )
n 6N
K = product factor, dimensionless for crude oil, K
0.84; for all other organic liquids, K = 1).
9.5.2 Model Parameters
It is assumed that all stationary tanks are fixed-roof.
12
According to responses to the 1982 Westat Mail Survey, which
9-17
-------
13
were examined by the GCA Corporation, there are four sizes of
tanks that best represent the waste management industry:
5.3 m3 (1,500 gal)
30.3 m3 (8,000 gal)
75.0 m3 (20,000 gal)
795 m3 (210, 000 gal) .
Table 9-2 lists typical input parameters for these model tanks.
Turnovers per year were selected based on volume of waste
processed in waste management scenarios recorded in various
documents. Molecular weight and vapor pressure are a function
of the waste loaded.
9.5.3 Sample Calculation for Tank Loading Emission Model
Input parameters:
Waste stream: organic liquid (see Section 9.2.3 for
constituents)
M , molecular weight of vapor, Ib/lb mol: 78.23
P , true vapor pressure of loading liquid, psia: 0.68
K , product factor for working loss: 1
V, fixed-roof tank capacity, gal: 20,000
N, turnovers per year: 44
K , turnover factor, dimensionless: 0.848.
n
a. Calculate M , molecular weight of vapor:
(see Section 9.2.3 for calculation).
*
b. Calculate P , true vapor pressure of loading liquid:
(see Section 9.2.3 for calculation).
c. Calculate K , turnover factor: because N = 44,
K = ^±^ = 0.848
n 6N
9-1!
-------
TABLE 9-2. PERTINENT FIXED-ROOF TANK SPECIFICATIONS
14,15,16
Specifications
Model Model
A B
Model
C
Model
D
Capacity, m3
(gal)
Tank height, m
Tank diameter, m
Average vapor space
height, m
Adjustment for small
diameter
(dimensionless)
Average diurnal temp.
change, "C ("F)
Paint factor
(dimensionless)
5.3 30.3 75.7
(1,500) (8,000) (20,000)
2.4
1.7
1.2
2.4
4
1.2
0.26 0.65
11
(20;
i
11
(20;
i
2.7
5.8
1.4
0.86
11
(20;
i
795
(210, ooo;
12.2
9.1
6.1
1
11
(20;
i
Relation of tank to
ground
Product factor
Above
1
Above
1
Above
1
Above
1
9-19
-------
d. Calculate air emissions:
L = 2.40 x 10~5 xM • P* • V • N • K • K
w v n c
= 2.40 x 10~5 x 78.23 x 0.68 x 20,000 x 44 x 0.848 x 1
= 953 Ib/yr
= 0.43 Mg/yr.
9.6 S TATIONARY TANK S T ORAGE
9.6.1 Model Description
Fixed-roof tank storage of hazardous wastes results in VO
"breathing" emissions through vents as ambient temperature and
barometric pressure fluctuate. Emissions occur in the absence of
17
any liquid level change in the tank. An existing AP-42
equation was used to estimate VO breathing losses from hazardous
waste storage tanks as follows:
-9 P* 0.68 17-3 051 05
L, = 2.26 10 M * -^t- * . D . H . • T • F • C • K (9-5)
b v *p7N-P * p c
_T\
where
L, = fixed-roof breathing loss, Ib/yr
M = molecular weight of vapor in tank, Ib/lb mol
*
P = true vapor pressure at bulk liquid conditions, psia
P = average atmospheric pressure at tank location, psia
£\.
D = tank diameter, ft
H = average vapor space height, ft (assumed to be one-half
of tank height)
•T = average ambient diurnal temperature change, °F (20 °F
assumed as a typical value)
F = paint factor, dimensionless (see Table 9-3)
C = adjustment factor for small diameter tanks,
dimensionless (for diameter > 30 ft, C = 1; for
diameter < 30 ft,
9-20
-------
C = 0.0771 D - 0.0013 D2 - 0.1334)
K = product factor, dimensionless (for crude oil, K =0.65,
for all other organic liquids, K = 1.0).
The above equation requires an estimation of the true vapor
pressure using the liquid concentration. For very volatile
constituents, the liquid concentration depends on the amount lost
as air emissions. To correct for the loss to the air in
estimating the liquid concentration, the following equation may
be used:
fraction lost to air =
Lb + Lt '
where Lt is the tank input of the volatile constituent in pounds
per year.
9.6.2 Model Parameters
Table 9-3 identifies the model parameters for estimating
tank breathing losses. Molecular weight and vapor pressure are
functions of the waste stored.
9.6.3 Sample Calculation for Tank Storage Emission
Model Input parameters:
Waste stream, organic liquid (see Section 9.2.3 for
constituents)
M , molecular weight of vapor, Ib/lb mol: 78.23
P , true vapor pressure of loading liquid, psia: 0.68
K , product factor for breathing loss: 1
v, fixed-roof tank capacity, gal: 20,000
D, tank diameter, ft: 19
H, average vapor space height, ft: 4.5
•T, diurnal temperature change, °F: 20
F , paint factor, dimensionless: 1
C, adjustment factor for small tanks: 0.86 (calculate in
c., below)
9-21
-------
TABLE 7-3. PAINT FACTORS FOR FIXED-ROOF TANKS
18
Tank color
Roof
Shell
Paint factors (FP)
Paint condition
Good Poor
White
Aluminum (specular)
White
Aluminum (specular)
White
Aluminum (diffuse)
White
Light gray
Medium gray
White 1.00
White 1.04
Aluminum (specular) 1.16
Aluminum (specular) 1.20
Aluminum (diffuse) 1.30
Aluminum (diffuse) 1.39
Gray 1.30
Light gray 1.33
Medium gray 1.40
1.15
1.18
1.24
1.29
1.38
1.46
1.38
1.44a
1.58a
Estimated from the ratios of the seven preceding paint factors
9-22
-------
a. Calculate molecular weight of vapor:
(see Section 9.2.3 for calculation).
b. Calculate true vapor pressure of loading liquid:
(see Section 9.2.3 for calculation).
c. Calculate adjustment factor for small tanks:
C = 0.0771 x 19 - 0.0013(19)2 - 0.1334
= 0.86 .
d. Calculate air emissions:
-2 P 173 051 05
L, =2.26x10 M * — * . D»H»«T»F»C»K
**
A
_ p c
=2.26x10 x78.23x * . 'Q g * x (19) 'x (4.5) '
x (20) °'5 x ( 0.86 )
= 300 Ib/yr
=0.14 Mg/yr
9.7 SPILLS
9.7.1 Model Description
19
An ICF study of truck transport to and from TSDF and truck
emissions at TSDF terminals provided the background information
necessary to estimate spillage losses during TSDF and trucking
-4
operations. As a result of this study, spill fractions of 10
-5
and 10 were assumed for drum movement of wastes and all other
remaining waste movement, respectively. Thus, for every 10,000
Mg of drummed hazardous waste moved, 1 Mg is assumed to be
spilled. The following equation is used to estimate spill
emissions :
E=F x I x W. x V. (9-6)
s i i
9-23
-------
where
E = spill emissions, Mg/yr
-4 -5
F = emission fraction, 10 or 10
o
I = annual throughput, Mg/yr
W. = VO weight fraction
V. = fraction for volatilization.
9.7.2 Model Parameters
In both cases of spills, it is assumed that 50 percent of
the volatiles in the waste are lost. The remaining 50 percent
are recovered by RCRA spill plan response. Therefore, most
spills would be mitigated before 100 percent of VO is lost to the
atmosphere.
It is assumed that spills do not occur during the transfer
of waste into a stationary tank if loading is automated through
fixed piping.
9.7.3 Sample Calculation for Drum Storage
Model Input parameters:
Waste stream: organic liquid (see Section 9.2.3 for
constituents)
3
Waste density: 1.04 Mg/m
-4
Emission fraction: 10
Weight fraction: 1
Volatilization fraction: 0.5
3
Number of drums handled: 2,184 (0.208 m /drum).
a. Calculate annual throughput, I, Mg/yr:
I = 2,184 x 0.208 m3 x 1.04 Mg/m3
= 472 Mg/yr
b. Calculate air emissions:
E = 10~4 x 472 Mg/yr x 1 x 0.5
9-24
-------
= 0.024 Mg/yr
9-25
-------
9.8 FUGITIVE EMISSIONS
9.8.1 Emission Model for Fugitives
Waste transfer operations often involve pumping wastes
through pipelines into a variety of containment units. Such
pumping creates the potential for fugitive emission losses from
pumps, valves, and flanges. Table 9-4 presents the Synthetic
Organic Chemical Manufacturing Industries (SOCMI) emission
20
factors that had been developed to estimate VO that leak from
pump seals, valves, and flanges. These factors are independent
of the throughput, type, or size of the process unit.
TABLE 9-4. SOCMI EMISSION FACTORS FOR FUGITIVE LOSSES
Equipment
Pump seals
Valves
Flanges
Type of Emission factor
service (kg/h-source)
Light liquid 4
Light liquid 7
8
.94
.10
.30
E-2
E-3
E-4
The following equation is used to estimate fugitive
emissions:
E = • (Ff x Ni) x h x 10~3 (9-7)
where,
E = fugitive emissions, Mg/yr;
Ff = emission factor per source-type, kg/h-source (see Table
9-4);
N. = number of sources per source-type; and
h = residence time in the equipment (assume = 8,760 h/yr).
9.8.2 Model Parameters
The major input parameters required for the emission model
are emission factor, number of sources, and residence time. It
is assumed that waste remains in the transfer equipment 24 h/d,
9-26
-------
365 d/yr; therefore, VO are continuously being leaked to the
atmosphere.
Minimal information has been compiled on typical quantities
of pumps, valves, and flanges at waste management facilities.
Therefore, previous contractors have turned to data collected
from the petroleum refining industry and SOCMI. GCA recommended
that "for any hazardous waste filling operation, transfer
operation, or handling operation involving pumps, the estimate of
two pumps, 35 valves, and 80 flanges be used. This includes tank
21
filling, tank truck or car filling, and drum filling." Because
the relationship 2:35:80 appears to be too high for pumping waste
into a single drum, one pump, three valves, and eight flanges are
used for estimating emissions. GCA recommended that smaller
quantities of pumps, valves, and flanges identified by SOCMI be
applied for transfer operations to injection wells and
22
incinerators, i.e., 1 pump, 18 valves, and 40 flanges.
9.8.3 Sample Calculation for Fugitive Emission Model
Estimate the annual fugitive emissions from a set of piping
lines that connect to a storage tank, given the following
information.
Input parameters:
Assume 2 pumps, 35 valves, and 80 flanges associated with
the piping equipment.
Assume the waste stream is organic liquid.
Assume waste remains in piping line 24 h/d, 365 d/yr.
a. The emission factor for light liquids was used because of
the high VO content.
b. Calculate fugitive emissions:
E = (0.0494 kg/h x 2 + 0.0071 kg/h x 35 + 8.3 x 10~4kg/h x 8)
x 8,760 h/yr x 10~3 Mg/kg = 3.62 Mg/yr.
9-27
-------
9.9 VACUUM TRUCK LOADING
9.9.1 Emission Model for Vacuum Truck Loading
Emissions from vacuum truck loading are estimated by
calculating an equilibrium concentration of organic vapors in the
vacuum truck at its operating conditions and assuming that a
total volume of gas equal to the vacuum truck volume is emitted
to the atmosphere for each loading episode. Equations for making
the calculations are presented as follows:
E. = N x Y. x MW.
i v i i
*
X. P
Y. = —5 (for oily waste)
1 t
N = V
v [P ¥„ (T/273)]/P.
Ob "C
where,
E. = air emissions of compound i, g/truckload;
N = total moles of vapor discharged, g mol;
Y. = mole fraction of compound i in vapor phase;
X. = mole fraction of compound i in liquid phase;
MW. = molecular weight of compound i, g/g mol;
*
P = vapor pressure of compound i, mm Hg;
P, = total system operating pressure, mm Hg;
P = atmospheric pressure, mm Hg;
V = vacuum truck volume, m ;
¥„ = volume of 1 g mol of gas at STP, 0.0224 m3/g mol; and
(j
T = operating temperature, K.
9.9.2 Model Parameters
Based on information obtained during site visits to
refineries using land treatment, vacuum trucks have a capacity of
3
about 21 m (5,500 gal) and operate at a pressure of
9-28
-------
approximately 303 mm Hg. These values are used in estimating
vacuum truck emissions.
Molecular weight and vapor pressure are functions of waste
loaded, and 25 °C is considered a standard operating temperature,
9.9.3 Sample Calculation
The following is a sample calculation of benzene emissions
during loading of a vacuum truck with organic liquid.
Input parameters:
Waste stream: organic liquid
(see Section 9.2.3 for constituents)
VO constituent: benzene
MW., molecular weight, g/g mol: 78
* -1
P , pure compound vapor pressure: 95.2
P., system operating pressure, mm Hg: 303
P , atmospheric pressure, mm Hg: 760
X., mole fraction in liquid: 0.368
1 3
V, vacuum truck volume, m : 21
¥„, volume of 1 g mol of gas at STP, m3/g mol: 0.0224
(j
T, operating temperature, K: 298
N, turnovers per year, truckload/yr: 10.
a. Calculate total moles of vapor discharged, g mol:
V
Nv [P ¥„ (T/273)]/P.
a (j L.
21 m3
(760 mm Hg x 0.0224 m3/g mol x 298 K/273 K)/303 mm Hg
= 342.41 g mol/truckload .
b. Calculate mole fraction of benzene in vapor phase, Y.:
*
P *i 95.2 0.368 _ 11[;c
Yi = -^T- = 303 = °'1156 '
c. Calculate air emissions per truckload, g/truckload:
9-29
-------
E. = N x Y. x MW.
i v i i
= (342.41 g mol/truckload) (0.1156) (78 g/g mol)
= 3,087 g/truckload .
d. Calculate annual emissions for benzene, Mg/yr:
Annual emission = E. x N
= 3,087 g/truckload x 10 truckload/yr
= 30,870 g/yr
= 0.031 Mg/yr
e. Repeat the above procedures, a through d., to compute
emissions for each constituent as follows:
Constituent E.,g/truckload Annual emissions, Mg/yr
Benzene
Naphthalene
Phenol
Total emissions
3,
3,
087
7
14
108
0
0
0
0
.031
.00007
.00014
.0312
9.10 REFERENCES
1. U.S. Environmental Protection Agency. Transportation
and Marketing of Petroleum Liquids. In: AP-42.
Compilation of Air Pollutant Emission Factors. Third
Edition, Supplement 12, Section 4.4. Research Triangle
Park, NC. Office of Air Quality Planning and Standards
July 1979. 13 pp.
2. GCA Corporation. Air Emission Estimation Methods for
Transfer, Storage, and Handling Operations. Draft
Technical Note. Prepared for U.S. Environmental
Protection Agency, Office of Air Quality Planning and
Standards. Research Triangle Park, NC. Contract No.
68-01-6871. August 1985.
3 . Reference 1.
4 . Reference 2.
9-30
-------
5. U.S. Environmental Protection Agency. Assessing the
Release and Costs Associated with Truck Transport of
Hazardous Wastes. PB 84-224-468 (Contract No. 68-01-
0021). Washington, DC. January 1984. 151 p.
6. Reference 5.
7. Shen, T. T. Estimating Hazardous Air Emissions from
Disposal Sites. Pollution Engineering. 31-34. August
1981.
8. GCA Corporation. Air Emissions of VOC from Waste Piles
at Hazardous Waste Treatment, Storage, and Disposal
Facilities. Prepared for U.S. Environmental Protection
Agency, Office of Air Quality Planning and Standards.
Research Triangle Park, NC. Contract No. 68-01-6871.
August 1985.
9. U.S. Environmental Protection Agency. Tank and Drum
Cleaning. In: AP-42. Compilation of Air Pollutant
Emission Factors. Fourth Edition, Section 4.8.
Research Triangle Park, NC. Office of Air Quality
Planning and Standards. September 1985. 4 pp.
10. Reference 9.
11. U.S. Environmental Protection Agency. Storage of
Organic Liquids. In: AP-42. Compilation of Air
Pollutant Emission Factors. Fourth Edition, Section
4.3. Research Triangle Park, NC. Office of Air Quality
Planning and Standards. September 1985. 25 pp.
12. Westat, Inc. National Survey of Hazardous Waste
Generators and Treatment, Storage, and Disposal
Facilities Regulated Under RCRA in 1981. Prepared for
U.S. Environmental Protection Agency. Office of Solid
Waste, Washington, DC. April 1984.
13. Addendum to Memorandum dated September 6, 1985, from
Eichinger, Jeanne, GCA Corporation, to Hustvedt, K. C.,
EPA/OAQPS. September 12, 1985. TSDF model source
parameters and operating practices data base.
14. Reference 11.
15. Reference 13.
16. Graver Standard Cone-Roof, Flat-Bottom Tanks. Sizes and
Capacities.
9-31
-------
17. Reference 11.
18. TRW Environmental, Inc. Background Documentation for
Storage of Organic Liquids. Prepared for U.S.
Environmental Protection Agency. Research Triangle
Park, NC. Contract No. 68-02-3174. May 1981.
19. Reference 5.
20. U.S. Environmental Protection Agency. Control of
Volatile Organic Compound Leaks from Synthetic Organic
Chemical and Polymer Manufacturing Equipment. Research
Triangle Park, NC. Publication No. EPA-450/ 3-83-006.
March 1984.
21. Reference 2.
22. Reference 2.
9-32
-------
10.0 COMPARISON OF MODEL RESULTS WITH FIELD TEST DATA
10.1 INTRODUCTION
Predictions from TSDF emission models are compared with
field test data in this chapter. In general, considering the
uncertainty of field emission measurements, agreement between
measured and predicted values is considered reasonable. Measured
and predicted emissions generally agree within an order of
magnitude.
The following caveats must be considered in any evaluation
of the comparison results presented in the following sections:
1. The field test data did not always include all of the
input parameters required to use the emission models.
In such cases, parameter values representative of
field operations were used as defaults.
2. The emission models use average influent and effluent
concentrations to estimate annual emissions.
Variations in concentrations and constituents are not
reflected.
3. Field test data provide information on a limited
number of hazardous constituents. Extrapolation of
comparisons on limited constituents to all
constituents of interest may not always be possible.
4. The method of measuring emissions (e.g., flux chambers
and other enclosure systems) could alter the real-
world system being tested and affect the
representativeness of the measured emissions.
10.2 SURFACE IMPOUNDMENTS AND OPEN TANKS
10.2.1 Summary
Emission test data were available from tests of five
quiescent surface impoundments. The overall mass transfer
10-1
-------
coefficients determined in these tests agreed generally within an
order of magnitude with the overall coefficient predicted by the
mass transfer correlations described in Chapter 5.0. Predicted
values were higher than measured values in some cases and lower
in others.
The emission models used for impoundments also were applied
to open tanks. The comparison of measured and predicted values
for the overall mass transfer coefficient for open wastewater
treatment tanks yielded mixed results. For tanks with quiescent
surfaces (e.g., clarifiers and equalization basins), the model
predictions were generally lower than measured values but agreed
within an order of magnitude. For the aerated systems, the model
predictions agreed well with material balance and ambient air
measurements for an open aerated system.
10.2.2 Details of Comparisons
The approach to the comparison of predicted and measured
values is to estimate the overall mass transfer coefficient from
the correlations given in Chapter 5.0 and to compare this value
to the overall mass transfer coefficient from the test data. The
overall mass transfer coefficient from the test data is
calculated from a measured emission rate and a measured or
estimated bulk concentration in the liquid phase. Note that
errors in either the measured emission rate or liquid-phase
concentration have a direct effect on the errors in the
calculated mass transfer coefficient.
Most of the measured emission data were obtained by flux
chamber measurements. At a few sources, ambient air monitoring
and material balances were used to determine the emission rate
for calculation of the overall mass transfer coefficient.
GCA Corporation performed an analysis of data from
impoundments. The results are given in Tables 10-1 through 10-4
for four ponds at two different sites. Site 5 is a commercial
hazardous waste facility with a wastewater treatment system
onsite. The reducing lagoon receives wastes classified as
10-2
-------
TABLE 10-1. COMPARISON OF RESULTS FOR REDUCING LAGOON 1
AT SITE 51'2
Mass transfer coefficient (x 10 m/s
Model predictions
s
Average flux (for 5 to 10 m/
Constituent
Benzene
Toluene
Ethylbenzene
Naphthalene
Methylene chloride
Chloroform
1, 1, 1-Trichloroethane
Carbon tetrachloride
p-Dichlorobenzene
Styrene
chamber measurement
4.9
5.0
5.5
2.6
12
5.7
7.6
11
2.6
5.7
windspeed)
4.2-17
3.9-15
3.6-14
3.5-14
4.7-19
4.3-17
3.9-15
3.9-16
3.6-14
3.7-15
b
Calculated from reported emission rate and corresponding
liquid-phase concentration.
Windspeed during the test ranged from 5 to 10 m/s.
10-3
-------
TABLE 10-2.
COMPARISON OF RESULTS, FDR HOLDING POND 6
AT SITE 5 '
Mass transfer coefficient (x 10 m/s
Average flux
Model predictions
(for 5 to 10 m/s
Constituent
Benzene
Toluene
Ethylbenzene
Naphthalene
Methylene chloride
Chloroform
1, 1, 1-Trichloroethane
Chlorobenzene
p-Dichlorobenzene
Acetaldehyde
chamber measurement
2.7
2.3
2.6
1.6
3.1
2.2
3.9
<.039
4.3
3.4
windspeed)
5.3-21
4.9-19
4.6-18
4.4-18
6.0-24
5.4-21
4.9-19
4.9-20
4.6-18
5.7-19
a
b
Calculated from reported emission rate and corresponding
liquid-phase concentration.
Windspeed during the test ranged from 5 to 10 m/s.
10-4
-------
TABLE 10-3. COMPARISON OF RESULTS FQR^OXIDIZING LAGOON 2
AT SITE 5 '
Mass transfer coefficient (x 10 m/s]
Model predictions
Average flux (for 5 to 10 m/s
Constituent chamber measurement windspeed)
Toluene 0.38 3.8-15
Ethylbenzene 0.037 3.6-14
1,1,1-Trichloroethane 35 3.9-15
Calculated from reported emission rate and corresponding
, liquid-phase concentration.
Windspeed during the test ranged from 5 to 10 m/s.
TABLE 10-4. COMPARISON OF RESULTSoFOR SURFACE IMPOUNDMENT
AT SITE 4 '
Mass transfer coefficient (x 10 m/s)
Flux ,, , , , . , .
, , ,a Model predictions
chamber measurement /f r , in /
(for 5 to 10 m/s
Constituent Average Range windspeed)
Toluene 2.4 1.9-2.7 6.3-25.1
Ethylbenzene 1.0 0.46-1.4 5.9-23.5
Methylene chloride 8.4 5.6-10.0 7.7-30.5
1,1,1-Trichloroethane 2.6 1.1-3.6 6.3-24.7
Chloroform 12.0 5.4-15.0 7.0-27.6
p-Dichlorobenzene 0.44 0.079-0.75 5.9-23.1
aResults for June 22, 1984.
Windspeed during the test ranged from 5 to 10 m/s.
10-5
-------
reducing agents from tank trucks. The lagoon is operated on a
batch basis and was observed to contain a zone of solids and a
surface with a floating oil film. The holding pond receives
aqueous wastes from the water treatment system and is filled (and
discharged) on a monthly basis. The oxidizing lagoon receives
oxidizing agents including halogens and other organic compounds.
The accumulation of solids and oil film also was observed on this
lagoon. Site 4 also is a commercial hazardous waste facility,
and its impoundment is used to contain aqueous wastes.
Table 10-5 presents a comparison of results for Site 3, which is
a chemical manufacturing plant that produces primarily nitrated
aromatics and aromatic amines. This impoundment is a wastewater
holding pond for the wastewater treatment system at the plant.
Two wastewater streams that enter the treatment system are
distillation bottoms from aniline production (K083) and the
nitrobenzene production wastewater (K104).
The results in Tables 10-1 through 10-5 show a reasonable
agreement between measured and predicted values of the overall
mass transfer coefficient. The measured results for the
impoundment in Table 10-3 may have been affected by an oil film
observed on the surface or from poor mixing in the impoundment,
which can complicate representative sampling of the liquid-phase
concentration. Table 10-5 shows good agreement of results for
toluene and benzene, which were present in the liquid phase at
2.6 and 17 mg/L, respectively. The liquid-phase concentrations
of the other four compounds in Table 10-5 ranged from 0.029 to
0.15 mg/L. The differences in measured and predicted values for
these four compounds may have been affected by the accuracy of
sampling and analysis of the liquid. The compounds listed in
Tables 10-1 through 10-5 are controlled by the liquid-phase mass
transfer. Consequently, the results are most dependent on
Springer's correlation for k (the liquid-phase mass transfer
j_i
coefficient) and suggest that Springer's model is probably
accurate within an order of magnitude.
10-6
-------
TABLE 10-5. COMPARISON OF RESULTS FOR WASTEWATER HOLDING LAGOON
AT SITE 39
Mass transfer coefficient (x 10 m/s)
Flux chamber
Constituent measurement Predicted
Cyclohexane 0.39 3.8
Tetrachloroethylene 0.10 3.7
Toluene 9.0 3.8
Benzene 3.7 4.1
n-Undecane 0.21 2.8
Methylchloride 35.0 3.1
Based on an average measured windspeed of 3.7 m/s and an average
temperature of 22 °C.
10-7
-------
TABLE 10-6. COMPARISON OF RESULTS FOR PRIMARY CLARIFIERS
AT SITE 811
Mass transfer coefficient (x 10 m/s)
Material
Constituent
Ambient
balance monitors
Model
prediction
Tetralin
2-Ethyl hexanol 96.0
2-Ethyl hexyl acrylate
Naphthalene 179.0
1,2-Dichloroethane 58.0
Benzene 5.4
Toluene 35.0
Ethyl benzene 156.0
227.0
42.0
123.0
92.0
2.9
18.0
50.0
39.0
2.0
2.7
3.4
4.0
4.1
3.8
3.5
-------
TABLE 10-7. COMPARISON OF RESULTS FOR EQUALIZATION BASIN
AT SITE 812
Mass transfer coefficient (x 10 m/s)
Material Ambient Model
Constituent balance monitors prediction
1,2-Dichloroethane 20 19.0 5.0
Benzene 20 8.9 5.1
Toluene 25 42.0 4.7
Ethyl benzene 25 5.4 4.4
10-9
-------
TABLE 10-8. COMPARISON OF RESULTS FOR AERATED STABILIZATION
BASINS AT SITE 813
Mass transfer coefficient (x 10 m/s)
Material Ambient Model
Constituent balance monitors prediction
2-Ethyl hexanol 0.05 0.01 0.17
2-Ethyl hexyl acrylate 4.8 8.3 2.9
1,2-Dichloroethane 2.0 0.52 5.7
Benzene 12.4 1.1 10.6
Toluene 5.0 5.8 10.1
Ethyl benzene 2.9 0.55 9.9
10-10
-------
TABLE 10-9. COMPARISON OF RESULTS FOR COVERED AERATED LAGOON
AT SITE 715'16
4
Mass transfer coefficient (x 10 m/s)
Vent rate
Constituent measurement Predicted
1,2-Dichloroethane 0.05 7.2
Benzene 0.30 8.9
Toluene 0.95 8.8
Based on an estimated windspeed (not measured) of 5 m/s and an
1 8
estimated turbulent area of about 50 percent.
10-11
-------
GCA, in a separate document, examined measured and predicted
mass transfer coefficients for open tanks that are part of
wastewater treatment systems. The results for Site 8, which is
an industrial wastewater treatment operation, included a primary
clarifier, an equalization basin, and aerated stabilization
basins. The various influent and effluent liquid streams were
analyzed, and air emissions around the unit were monitored.
Overall mass transfer coefficients were calculated from material
balance data and from ambient air monitoring. These values are
listed in Tables 10-6 through 10-8 along with the predicted
values from the mass transfer correlations given in Chapter 5.0.
The primary clarifier, equalization basin, and the quiescent
portion of the stabilization basin were modeled as quiescent
surfaces, and the correlations of Springer and MacKay/Matasugu
were used. The turbulent portion of the stabilization basins was
modeled using the correlations of Thibodeaux and Reinhardt.
Useful conclusions from the comparison of measured and
predicted values are difficult because of the lack of consistent
results from air monitoring, probably due to very short sampling
periods, changes in the windspeed and direction, and the
contribution to emissions from sources near the mentioned source.
In addition, material balance calculations are subject to error
from changes in influent/effluent flow rates and concentrations
of specific compounds. In general, the model predictions for the
primary clarifier and equalization basin are lower than the
measured values. For the aerated stabilization basin, most of
the predicted mass transfer coefficients are higher than the
measured values; however, the agreement is within an order of
magnitude. The measured values for the primary clarifier may
have been higher than the predicted values because of the
contribution from nearby sources to measured air concentrations
or because of the contribution from the falling film at the
clarifier.
10-12
-------
GCA also performed an analysis on an aerated lagoon at
14
Site 7. This lagoon was covered and was purged with air at a
3 3
rate of 1.4 m /s (3,000 ft /min). Predicted and calculated
values for the mass transfer coefficients are given in Table 10-9
and show that predicted values are 1 to 2 orders of magnitude
higher. The basis of the predicted values includes Thibodeaux
and Reinhardt's correlations for aerated systems. No strong
conclusions on the model's validity can be drawn from these data
because the dome enclosure over the system and its evacuation
rate probably have a direct effect on emissions. In addition,
difficulties with air measurements and determination of the
appropriate driving-force concentration (assumed to be the bulk
liquid concentration) can lead to errors in the calculated values
of the overall mass transfer coefficient.
The results of the biodegradation model were also compared
to available data from biodegradation measurements. The most
desirable comparison would be for a system in which the air
emission rate and biodegradation rate were measured directly.
However, the extent of biodegradation from studies of real
systems has usually been determined by difference from a material
balance (what enters the system minus what leaves with the ef-
fluent and with air emissions).
Petrasek et al. performed such a study on a large pilot-
19
scale activated sludge system with diffused air aeration. The
activated sludge unit was enclosed, and the diffused air that was
removed was sampled (for flow rate and concentration) to
determine air emissions. This system was designed for a flow
rate of 2.2 L/s (35 gal/min) with an air purge rate of 57 L/s. A
summary of the operating parameters is given in Table 10-10. The
study used a synthetic wastewater that contained individual
volatile compounds at levels of 32 to 300 ppb. The biomass
concentration was 2 g/L, and the resultant food-to-microorganism
(F/M) ratio of 0.5 is well within the recommended design range of
0.2 to 0.6.
10-13
-------
Petrasek reported the percent of each compound entering the
activated sludge unit that was emitted with the diffused air; the
fraction biodegraded could be determined by difference-assuming
all unrecovered material was biodegraded. The results are
summarized in Table 10-11 and show a range of measured values
from 5 percent for chlorobenzene to 62 percent for 1,1,1-
trichloroethane. The predictions of the biodegradation model
discussed in Chapter 5.0 are also presented in Table 10-11 for
comparison. The comparison shows that the model predictions
agree well with the Petrasek measurements for nearly every
compound.
Another type of comparison between measurements and
predictions involves effluent concentrations for well-defined
22
wastewater treatment systems. Namkung and Rittman reported
influent and effluent concentrations of volatile organics for two
Chicago wastewater treatment plants that receive large shares of
industrial discharges. The measurements were made for two large
activated sludge units aerated by diffused air. In addition, the
system's operational parameters were defined (Table 10-12) and
provided the necessary inputs for the mathematical model that
includes air emissions (diffused air system) and biodegradation.
The results of measured and predicted effluent
concentrations are summarized in Table 10-13. The most
convincing comparison is the close match for both plants for
tetrachloroethylene, which the authors stated was not
biodegradable in these systems. Therefore, a biorate equal to
zero was used in the model for this compound. The close
agreement between measured and predicted effluent concentrations
suggests that this compound is almost entirely removed by air
stripping, and the quantity predicted to be air stripped by the
model is reasonably accurate.
10-14
-------
TABLE 10-10. DESCRIPTION OF PETRASEK'S ACTIVATED SLUDGE SYSTEM20
Parameter Value
Flow rate (L/s) 2.2
Volume (m3) 59.8
Residence time (h) 7.5
Air rate (L/s) 57
Fraction of surface agitated 0
Biomass concentration (g/L) 2.0
Concentration range for organics (ppm) 0.032 - 0.30
F/Ma 0 . 5
F/M = Food to microorganism ratio (Ib/lb biomass • day)
based on chemical oxygen demand.
10-15
-------
TABLE 10-11. COMPARISON OF PETRASEK'S MEASUREMENTS AND MODEL
PREDICTIONS
Reported Predicted Fraction
Compound
Benzene
Carbon tetrachloride
Chlorobenzene
Chloroform
Dichloropropane (1,2)
Ethyl benzene
Tertachloroe thane
and -ethene
Toluene
Trichloroethane (1,1,1)
Trichloroethane (1,1,2)
Trichloroethene
fraction fraction
emitted emitted
0
0
0
0
0
0
0
0
0
0
0
.15
.59
.05
.34
.32
.21
.27
.20
.62
.25
.41
0
0
0
0
0
0
0
0
0
0
0
.19
.54
.02
.20
.09
.15
.37°
.15
.57
.06d
.37
assumed
biodeg.
0
0
0
0
0
0
0
0
0
0
0
.80
.41
.95
.66
.68
.79
.73
.80
.38
.75
.59
0
0
0
0
0
0
0
0
0
0
0
Predicted
fraction
biodeg.
.78
.44
.97
.75
.88
.82
.58°
.84
.40
.87d
.59
b
d
21
Data from Petrasek et al.; the fraction biodegraded is
assumed to be the fraction unaccounted for based on the
analyses of the sludge, the air, and the effluent.
Model predictions based on the equations presented in
Chapter 5.0 assuming influent VO concentrations of 0.10 mg/L
and operating parameters as provided in Table 10-10.
Arithmetic average for the removal fractions calculated for
1,1,2,2-tetrachloroethane and tetrachloroethene.
Employed 1,1,1-trichloroethane's biodegradation rate
constants.
10-16
-------
TABLE 10-12. DESCRIPTION OF TWO CHICAGO ACTIVATED SLUDGE UNITS23
Operating parameters
3
Volume (m )
Depth (m)
3
Wastewater flow (m /s)
3
Air rate (m /s)
Residence time (h)
Total organics (mg/L)
Biomass (g/L)
Fraction surface agitated
Concentrations (ppb)
Chloroform, in
out
Ethylbenzene, in
out
Methylene chloride, in
out
Tetrachloroethylene, in
out
Toluene, in
out
1 , 1 , 1-Trichloroethane, in
out
Trichloroethylene, in
out
Calumet
184,500
1.8
10.0
55
5.1
115
2.2
0
4.0
7.1
18
0.5
9.8
11
16
2.1
85
6.2
13
2.9
9.7
0.5
West- southwest
802,300
1.8
36.6
193
6.1
180
2.0
0
4.4
2.4
10
BDL
48
11
12
1.6
22
BDL
15
2.2
22
2.1
BDL = Below detection limit.
10-17
-------
TABLE 10-13. COMPARISON OF MEASURED AND PREDICTED EFFLUENT
CONCENTRATIONS FOR CHICAGO WASTEWATER TREATMENT PLANTS
24
Calumet effluent West-southwest effluent
concentrations, ppb concentrations, ppb
Compound Measured
Chloroform
Ethyl benzene
Methylene chloride
Tetrachloroethylene
Toluene
1, 1, 1-Trichloroethane
Trichloroethylene
b
0.5
b
2.1
6.2
2.9
0.5
Predicted Measured
b 2.4
0.68 c
b 11
1.0 1.6
2.9 c
1.0 2.2
0.75 2.1
Predicted
2
0
7
0
0
1
1
.3
.16
.1
.77
.69
.1
.6
Based on the equations presented in Chapter 5.0.
b
No comparison possible because measured concentration in
effluent was greater than measured concentration in influent.
c
'Measured effluent concentration was below detection limit.
10-18
-------
The results in Table 10-13 also indicate that 1,1,1-
trichloroethane and trichloroethylene are biodegraded. The model
predictions used a biorate for these two compounds that was
derived from Petrasek's data in Table 10-11. Both Petrasek's
data and the comparison in Table 10-13 indicate that these
compounds are biodegraded to some extent; otherwise, the measured
effluent concentrations in Table 10-13 would have been higher
than those predicted by the model with biodegradation included.
25
Tabak et al. conducted an extensive study of the
biodegradability of numerous toxic compounds. They found that,
when the microbial culture is properly acclimated, almost all
nonpesticide compounds could be, at least partially, biodegraded.
Although biodegradation rate constants were not determined, the
percent of compound biodegraded was shown to be dependent on the
acclimation of the culture, and (although to a lesser extent)
dependent on the concentration of the compound used. For every
compound tested, the percent biodegraded by the third subculture
(presumably the most acclimated) always decreased when the
concentration was doubled (unless both cultures were either 100
percent or 0 percent biodegraded), and this decrease was rarely a
decrease of a factor of two or more.26 If biodegradation were
strictly a first-order process, the percent biodegraded would be
independent of the concentration. If biodegradation were
strictly a zero-order process, the percent biodegraded would
decrease by a factor of two (for those compounds not biodegraded,
100 percent) when the concentration was doubled. Because an
intermediate effect was generally realized, Tabak's results
suggest Monod-type biodegradation rate kinetics are appropriate.
Another comparison that can be made is based on a series of
27
field studies, as reported by Coburn et al., in which batch,
biodegradation rate studies were performed while controlling air
emissions. The experimental first-order biodegradation rate
constant and the predicted, apparent first order rate constant
based on the Monod model can be compared in the last two columns
10-19
-------
of Table 10-14. Note for compounds whose log mean concentrations
are near or are greater than the appropriate half-saturation
constant (e.g., formaldehyde or methanol), the predicted first-
order rate constant according to the Monod model provides a
better estimate of the observed biodegradation rate than would be
provided assuming simple first-order kinetics (i.e., using KI
straight from the data base as provided, for comparison, in Table
10-14). Additionally, using the recommended biodegradation rate
constants and modeling approach, the predicted biodegradation
rates presented in Table 10-14 agree well with the reported
biodegradation rates for nearly every experimental run.
A separate study was conducted for EPA to evaluate measured
29
and predicted emissions for aerated waste treatment systems.
The correlations of Thibodeaux and Reinhardt were used (as
recommended in Chapter 5.0) to estimate the mass transfer
coefficients of the turbulent zone. The results showed an
agreement between measured and predicted values that were within
an order of magnitude. The report concluded that, when adequate
descriptions of plant operating parameters are available,
reliable emission estimates can be obtained from the models
(within the accuracy that results from variations in sampling and
chemical analysis). When plant operating parameters are known,
the major limitations in the models result from a lack of
accurate biooxidation rates and vapor/liquid equilibrium data for
.f. ^ 30
specific compounds.
10.2.3 Recommendations for Additional Data
The estimate of emissions from open liquid surfaces is
provided by the product of the mass transfer coefficient,
driving-force concentration, and surface area. Surface area can
be determined with reasonable accuracy. The previous comparison
of mass transfer coefficients indicated that they can be
estimated within an order of magnitude. Probably the greatest
source of uncertainty is in the estimate of the appropriate
driving force for mass transfer. The concentration is likely to
10-20
-------
TABLE 10-14. COMPARISON OF MEASURED AND PREDICTED BIODEGRADATION
RATES
Rate
Log mean
b K
cone . , max,
Compound mg/L mg/g/h
Acetone
Benzene
Chloroform
Dimethyl-
phthalate
Ethanol
Ethylbenzene
Ethylene
oxide
(oxirane)
Formaldehyde
Methanol
Methyl
ethyl
ketone
Methylene
chloride
1.35 1.3
2.56
0.005 19
0.10
0.008 2.94
0.002
4.2 2.2
4.9 8.8
0.005 6.8
1.7 4.2
3.2
3.9
8.0 5.0
62.
250. 18.
480.
490.
495.
0.10 2.0
0.27
0.37
0.80
0.028 22
0.031
0.053
0.15
0.23
constants First-order rate
constant, L/g/h
Y Y
IV IV -,
s' 1' b
mg/L L/g/h Experiment
1.1 1.15 1.15
0.34
13.6 1.4 0.36
2.1
3.7 0.79 0.36
0.29
0.71 3.1 0.36
9.8 0.90 0.70
3.2 2.1 0.36
4.6 0.91 0.81
0.75
0.37
20. 0.25 0.13
0.077
90. 0.200 .067
0.018
0.040
0.023
10 0.20 0.24
0.18
0.19
0.16
55 0.40 0.11
0.11
0.36
0.20
0.57
Q
Predicted
0.53
0.36
1.4
1.4
0.79
0.79
0.45
0.60
2.1
0.67
0.54
0.49
0.17
0.057
0.053
0.032
0.031
0.031
0.20
0.19
0.19
0.18
0.40
0.40
0.40
0.40
0.40
(continued)
10-21
-------
TABLE 10-14 (continued)
Rate constants
Log mean
cone . ,
Compound mg/L
2-Propanol
Thiobisme thane
Toluene
1, 1, 1-Trichloro-
ethane
Trichloroethene
Total xylenes
Recommended
2
6
1
0.
0.
0.
0
0.
0
0.
0.
b K K K ,
max, s, 1,
mg/g/h mg/L L/g/h
.9 15 200 0.75
.2
.07 0.16 0.17 0.93
014 3.5 30.6 2.4
016
081
.14
040 3.5 4.73 0.74
.16
004 3.9 4.43 0.88
097 40.8 22.7 1.8
rate constants from Appendix D,
2 8
First-order rate
constant, L/g/h
Exper .
0.069
0.085
0.13
0.28
0.34
0.63
1.9
0.38
0.88
0.41
>2.2d
Table
b c
Predict.
0.074
0.073
0.11
2.4
2.4
2.4
2.4
0.73
0.72
0.88
1.8
D-l.
concentration was calculated as follows:
- Cf)/ln(Ci/Cf)
where,
C. = initial concentration, mg/L; and
Cf = final concentration, mg/L.
Calculated as the apparent first-order rate constant using
the Monod model (Equation 5-13) based upon the log mean
concentration as follows:
d
K, = K / (K + CT,J .
1 max s LM
Final concentration was below detection limit. The final
concentration was assumed to be at the detection limit to
calculate the first-order rate constant. The actual rate
constant should be greater than the reported value.
10-22
-------
vary with time and location in the impoundment. The type of flow
system and extent of mixing in the liquid also will affect this
concentration.
For the less volatile compounds that may be controlled by
gas-phase mass transfer, the collection of equilibrium data may
be useful to compare with the estimated values used in the
models. The comparisons presented in this section primarily
address compounds with high volatility in water (high Henry's law
constant). Because semivolatile compounds also can be emitted to
a significant extent, air emission measurements for these less
volatile compounds would be useful for comparison with model
predictions.
10.3 LAND TREATMENT
Field data from four test sites and one laboratory
simulation were used as a basis for comparing measured emissions
with estimated emissions using the RTI land treatment model.
Table 10-15 summarizes the tests evaluated. Generally, estimated
emissions are within an order of magnitude of measured values.
Values of estimated emissions varied both above and below
measured values.
Comparisons of estimated and measured emission flux rates
are presented graphically in this section. Comparisons of
estimated and measured emissions by weight percent of applied
material are presented in the next section.
Considering the potential for error in measuring or
estimating values for parameters that are input to the model and
the potential for error in measuring emissions, differences in
the range of an order of magnitude are not unexpected. In making
the comparisons, values for all model inputs sometimes were not
available in the emission test reports. In these cases, values
were estimated using averages of field data or values identified
previously as typical or representative of actual land treatment
practices.
10-23
-------
TABLE 10-15. SUMMARY OF LAND TREATMENT TESTING AND TEST RESULTS
Test results
Site Test Test Test Test
No. Test site location description year sponsor procedures
12 West Coast corporate Laboratory 1986 - Private Run 1
research facility simulation 1987 corporation (raw waste)
Run 2
(raw waste)
Run 2
(treated
waste)
13 Southwest research Laboratory 1986 EPA Run 1
facility simulation (API
separator
sludge)
Box #1
Box #2
Box #3
Box #4
Run 2a
(IAF float)
Box #1
Box #2
Box #3
Box #4
14 Midwestern refinery Flux chamber 1985 ORD Plot A
Plot B
Test
duration
2 . 5 months
22 days
22 days
31 days
31 days
8 days
8 days
Waste constituent
Oil
Oil
Oil
Oil
Oil
Oil
Oil
Oil
Oil
Oil
Oil
Benzene
Toluene
Ethylbenzene
p-Xylene
m-Xylene
o-Xylene
Naphthalene
Benzene
Toluene
Ethylbenzene
p-Xylene
m-Xylene
Emission,
wt . %
35
11
1
5.2
NA
6.5
6.7
15
NA
18
19
81
41
195
16
39
28
1
110
66
402
21
83
10-24
-------
TABLE 10-15 (Continued)
Test results
Site
No.
Test site location
Test
description
Test
year
Test
sponsor
Test
procedures
Test
duration
Waste constituent
Emission,
wt. %
Plot C
days
Plot D
days
Plot E
8 days
Plot F
days
o-Xylene
Naphthalene
Benzene
Toluene
Ethylbenzene
p-Xylene
m-Xylene
o-Xylene
Naphthalene
Benzene
Toluene
Ethylbenzene
p-Xylene
m-Xylene
o-Xylene
Naphthalene
Benzene
Toluene
Ethylbenzene
p-Xylene
m-Xylene
o-Xylene
Naphthalene
Benzene
Toluene
Ethylbenzene
p-Xylene
m-Xylene
38
2
39
17
140
25
25
17
142
86
353
55
79
52
2
107
63
345
43
52
39
1
84
47
208
13
28
10-25
-------
TABLE 10-15 (Continued)
Site
No.
Test site location
Test
description
Test
year
Test
sponsor
Test
procedures
Test
duration
Test results
Waste constituent
Emission,
wt. %
o-Xylene
Naphthalene
24
1
15 West Coast refinery
Flux chamber 1984
ORD
Surface
application
5 weeks
Subsurface
5 weeks
n-heptane 60
Methylcyclohexane 61
3-Methyl-heptane 52
n-Nonane 56
1-Methylcyclohexene 49
1-Octene 50
• -Pinene 17
Limonene 22
Toluene 37
p-, m-Xylene 35
1,3,5- 21
TrimethyIbenzene
o-Ethyl-toluene 32
Total VO 30
Total oil 1.2
n-heptane 94
Methylcyclohexane 88
3-Methyl-heptane 77
n-Nonane 80
1-Methylcyclohexene 76
1-Octene 74
•-Pinene 21
Limonene 26
Toluene 56
p-, m-Xylene 48
10-26
-------
TABLE 10-15 (Continued)
Site Test Test Test Test
No. Test site location description year sponsor procedures
Test
results
Test Emission,
duration Waste constituent wt . %
1,3,5- 27
T rime thy Ibenzene
o-Ethyl-toluene 42
Total VO 36
16 Southwest research Laboratory 1983 API/EPA Run
facility simulation
Run
Run
Run
Run
Run
Run
Run
Run
Run
run
Run
Run
16
(con. ) Run
Run
Run
Run
Run
Run
Run
Run
no.
no.
no.
no.
no.
no.
no.
no.
no.
no.
no.
no.
no.
no.
no.
no.
no.
no.
no.
no.
no.
18
21
24
27
28
32
33
34
35
36
37
40
41
44
45
46
47
48
49
50
51
Total oil
8 hours'3 Oil
Oil
Oil
Oil
Oil
Oil
Oil
Oil
Oil
Oil
Oil
Oil
Oil
Oil
Oil
Oil
Oil
Oil
Oil
Oil
Oil
1.
9.
4.
0.
0.
0.
3.
2.
0.
0.
78
9.
0.
2.
4.
49
7.
6.
5.
9.
1.
0.
,4
,1
,4
02
,6
,1
,0
, 6
01
,9
.8
,9
,7
,8
,9
.9
,7
,9
,0
,7
,1
47
10-27
-------
TABLE 10-15 (Continued)
Test results
Site Test Test Test Test
No. Test site location description year sponsor procedures
10 Gulf Coast commercial Flux chamber 1983 ORD Single test0
TSDF
Test
duration
69
50
hours
hours
Emission,
Waste constituent wt . %
Total VO
Benzene
0.
3.
77
91
17 Midwestern refinery Flux chamber 1979 API Centrifuge
sludge
Test no.
Test no.
API
separator
sludge01
Test no.
Test no.
Test no.
5
6
7
8
9
19.9
307
619
122
520
hours
hours
hours
hours
hours
Oil
Oil
Oil
Oil
Oil
0,
2,
13
1.
13
.1
.5
.5
.1
.5
API = American Petroleum Institute.
IAF = Inducted air flotation.
ORD = Office of Research and Development.
aSludge applied to Box #1 and Box #3 as duplicate test; sludge treated with mercuric chloride to eliminate (or reduce)
bioactivity applied to Box #4 and no sludge applied to Box #2, which served as a control.
bEach run for which results are reported was 8 hours.
°Test was conducted using aged wastes.
dAllowed to weather for 14 days in open 5-gal buckets in an outdoor open shelter prior to application.
10-28
-------
In the 1985 test at a Midwest petroleum refinery (Case I),31
emission measurements were made at sample locations in six test
plots. For each plot, emission measurements were made after
waste application but before the plot was tilled, again after the
waste was tilled, and for another period after a second tilling.
All measurements were made using a flux chamber and tenax traps.
Emission rates were measured for six specific organic
constituents: benzene, toluene, p-xylene, o-xylene, m-xylene,
and naphthalene. Benzene and toluene were selected as a basis
for comparing measured and estimated emissions in this test. The
comparison was made for test plot A after the waste was tilled
for the first time. Estimated emissions for each compound are
higher than the measured values but generally are within a factor
of 10. Estimated and measured values are shown graphically for
benzene and toluene in Figures 10-1 and 10-2, respectively.
32
At the West Coast refinery (Case 2), emission tests were
made using three adjacent plots marked off in the land treatment
site. The center plot was used as a control and had no waste
applied while waste was applied to the other two plots. One plot
had waste applied to the soil surface and the other had waste
applied by subsurface injection. Flux chambers were situated on
each test plot and emission measurements were made during three
different test periods each lasting 4 days. Canister air
samples, sludge samples, and liquid samples were analyzed by gas
chromatography (GC). Emissions of both total VO and selected
specific constituents were measured during the test. For
comparing measured and estimated emissions, total VO and toluene
emissions from the surface application plot were used. Estimated
emission rates for both toluene and total VO agree reasonably
well with measured rates but range from higher to lower than
measured rates at different times. Estimated cumulative
emissions over the entire test period agree reasonably well with
the measured values. For both toluene and total VO, estimates
covered a 4-day period with a till occurring after 2 days.
10-29
-------
Estimated and measured values over the 4-day period for which the
comparison is made are generally within an order of magnitude, as
can be seen in Figures 10-3 and 10-4. Measured values were
reported as half-day average emission rates.
For the test at the commercial hazardous waste site in 1983,
(Case 3), waste was applied to a single test plot and tilled
into the soil. Air emission measurements were made over a 3-day
period using a flux chamber and gas canisters. Sampling
locations were selected randomly, with a control point used to
provide a common sampling position each day. Sample analyses
were made by GC. Emission comparisons of measured and estimated
emissions were made for total nonmethane hydrocarbon (NMHC) emis-
sions using data generated by GCA in a separate study of the data
34
from this test. As with previous tests, estimated emission
flux rates were greater than measured values but mostly were
within a factor of 10 or less of the measurements. Estimated
cumulative emissions also were substantially higher than measured
values. Estimated and measured values of instantaneous emission
flux rates are shown in Figure 10-5.
In the 1979 test at the Midwest petroleum refinery
35
(Case 4), three test plots were laid out. One plot was used as
a control and had no waste applied, one plot had an API separator
sludge applied, and the other plot had a centrifuge sludge
2
applied. A 1-ft collector box was placed on the test plot and
continuously purged with fresh air. The outlet from the box was
analyzed for total VO (as methane and NMHC) using a continuous
hydrocarbon analyzer. For one test run, total VO emissions were
estimated with the land treatment model for comparison with the
measured values. Measured and estimated values are shown
graphically in Figure 10-6. As can be seen, the estimated and
measured values agree quite well for this test. Total cumulative
emissions for each test were also estimated using the model and
compared with measured values. The estimated values were
generally higher than measured values for these emissions.
10-30
-------
10-31
-------
Benzene emissions
Legend
D Predicted D Measured
Tine (hours)
Figure 10-1. Estimated vs. measured benzene emission flux rates-
Case 1.
10-32
-------
Toluene emissions
Legend
Estimated \~\ Measured
Figure 10-2. Estimated vs. measured toluene emission flux rates-
Case 1.
10-33
-------
Toluene emissions
| | Estimated | | Measured
30 42
55 67
Time (hours)
78 90
Figure 10-3. Estimated vs. measured toluene emission flux rates-
Case 2 (data for 4 days only).
10-34
-------
VO emissions
Legend
Estimated Measured
30 42 55
Time (hours)
67
78
Fi
gure 10-4. Estimated vs. measured VO emission flux rates-Case 2.
10-35
-------
VO emissions
Legend
Estimated Measured
Figure 10-5. Estimated vs. measured VO emission flux rates-
Case 3.
10-36
-------
VO emissions
700
400
100
Legend
Estimated Measured
20 30
40 45
Time (hours)
82 120 140 165
Figure 10-6. Estimated vs. measured VO emission flux rates-
Case 4.
10-37
-------
10.3.1 Midwest Refinery—1985 (Case 1)
Table 10-16 presents the model input values used to compare
estimated and measured emissions for plot A of the Case 1 test
data. The information in Table 10-16 represents data for plot A
as reported in the test report. Similar information was reported
for plots B through F and those data were used as appropriate for
input to the model. Table 10-17 shows measured emissions of six
constituents made during the test. In this test, the waste was
allowed to stay on top of the soil for 24 hours before it was
tilled into the soil. Measured emissions during the first
24 hours were combined with measured emissions after tilling to
get total emissions. Table 10-17 shows variations in measured
emissions among the different test plots and shows emissions
greater than applied material for some plots and some waste
constituents. In Table 10-17, weight fraction represents the
fraction of applied material that is emitted to the air. For
ethylbenzene, all plots have measured emissions in excess of the
amount applied. To compare measured and estimated emissions, the
RTI land treatment model can be used for estimating emissions
both before and after tilling. Estimated cumulative emissions
for benzene and toluene for all plots are shown in Table 10-17
and show reasonable agreement with measured values.
10.3.2 West Coast Refinery (Case 2)
The data in Table 10-18 were used to estimate emissions of
toluene and total VO from the surface application plot at the
Case 2 land treatment facility. Estimated and measured
cumulative emissions are compared in Table 10-19. The
comparisons were made for total VO (as determined by purge and
trap) and for toluene.
10-38
-------
TABLE 10-16. INPUT PARAMETERS FOR RTI LAND TREATMENT MODEL3
Parameter
Value
Source
Organic loading
Tilling depth
Soil air porosity
Soil total porosity
Benzene concentration
Toluene concentration
3
b
b
Benzene diffusivity
Toluene diffusivity
Benzene vapor pressure
Toluene vapor pressure
Benzene biorate
Toluene biorate
0.0236 g/cm
20 cm
0.40
0.61
0.000249
0.000632
3.80 E-02 cm2/s
3.70 E-02 cm2/s
95.2 mm Hg
30.0 mm Hg
19.0 mg V0/g»h
73.0 mg V0/g»h
Molecular weight of oil 282 g/g mol
Calculated from field
data
Field data
Field data
Field data
Calculated from field
data
Calculated from field
data
Data base
Data base
Data base
Data base
Data base
Data base
Assumed
b
Source of field data:
conditions in plot A.
Weight fraction of oil.
Reference 36. Data represent
10-39
-------
10.3.3 Commercial Waste Disposal Test (Case 3)
Table 10-20 shows the inputs used to estimate emissions from
the Case 3 land treatment operation. No specific constituent
data were available so emissions were estimated using average
characteristics of the total organic phase. Results are shown in
Table 10-21. The comparison is made for the estimated versus
measured cumulative weight percent of applied oil that is emitted
after 24 hours and after 68 hours, which is the duration of the
entire test.
10.3.4 Midwest Refinery—1979 (Case 4)
The information in Table 10-22 was used to estimate emissions
from the Case 4 facility test. No specific constituent data were
available; emissions were estimated for total organics using
average parameter values. Results are presented in Table 10-23.
The comparisons are for the cumulative weight percent of applied
oil that was emitted over the entire period of each test.
10.4 LANDFILLS AND WASTEPILES
Emission testing has been performed on at least one active
(open) landfill at each of five sites. Only three of these sites
have closed or inactive landfills at which emission measurements
were performed. No emission test data are available for
wastepiles.
Meaningful comparisons can be performed of emission test data
with mathematical model predictions provided that all key model
input parameters are available from the tests. A review of
documentation from the emission tests indicates that generally
more than half of the needed model input parameters (other than
chemical property data) are unknown, despite the fact that
several emission tests were performed with the stated intention
of validating emission models. Examples of key model input
10-40
-------
TABLE 10-17. MEASURED AND ESTIMATED EMISSIONS-CASE 1
Measured cumulative emissions"
Benzene Toluene Ethylbenzene
Test
location ug/cm2 wt. frac. ug/cm2 wt. frac. ug/cm2
A 271.81 0.81 348.71 0.41 57.97
B 299.86 1.10 454.28 0.66 96.46
C 188.35 0.39 209.96 0.17 59.27
D 459.42 1.42 703.08 0.86 101.05
E 382.23 1.07 576.10 0.63 109.31
F 324.88 0.84 464.97 0.47 71.55
wt. frac.
1.95
4.02
1.40
3.53
3.45
2.08
Estimated
p-Xylene
ug/cm2
7.39
7.50
15.83
23.92
20.74
6.87
cumulative
Benzene
Test
location
All
ug/cm2
--
wt. frac.
0.83
wt. frac.
0.16
0.21
0.25
0.55
0.43
0.13
emissions
m-Xylene
ug/cm2 wt. frac.
96.40 0.39
163.84 0.83
87.17 0.25
185.32 0.79
136.39 0.52
78.04 0.28
o-Xylene Naphthalene
ug/cm2 wt. frac. ug/cm2 wt. frac.
21.11 0.28 2.15 0.01
23.18 0.38 2.31 0.02
18.76 0.17 3.08 0.01
38.02 0.52 3.35 0.02
31.56 0.39 2.46 0.01
21.39 0.24 2.44 0.01
Toluene
ug/cm2
--
wt. frac.
0.53
10-41
-------
TABLE 10-18. INPUT PARAMETERS FOR RTI LAND TREATMENT MODEL'
Parameter
Value
Source
3
Organic (oil) loading 0.0328 g/cm
Tilling depth 20 cm
Soil porosity 0.5
Molecular weight of oil 282 g/g mol
Toluene 0.00157 (wt. frac-
concentration tion of oil)
Toluene diffusivity 8.70 E-02 cm /s
Toluene vapor pressure 30.0 mm Hg
Toluene biorate
VO concentration
VO diffusivity
VO vapor pressure
VO biorate
73.0 mg V0/g»h
0.04 (wt. fraction
of oil)
6.60E-02 cm2/s
14.6 mm Hg
23. 68 mg V0/g»h
Estimated from field
data
Field data
Field data
Field data
Calculated from field
data
Data base
Data base
Data base
Calculated from field
data
Average from field
data
Average from field
data
Average from field
data
Source of field data: Reference 37.
10-42
-------
TABLE 10-19. ESTIMATED VS. MEASURED EMISSIONS—CASE 2
Toluene
Total VO
Total oil
Time
after
tilling, day/h
33/793
33/793
33/793
Estimated
emissions,
wt. %
31
32
1.3
Measured
emissions,
wt. %
37
30
1.2
TABLE 10-20. INPUT PARAMETERS FOR RTI LAND TREATMENT MODEL'
Parameter
Value
Source
Organic loading
Tilling depth
Soil porosity
Oil Molecular weight
Vapor pressure
Diffusivity in air
Biorate
0.0406 g/cm3
19.6 cm
0.5
282 g/g mol
0.57 mm Hg
2.70 E-02 cmVs
23. 68 mg V0/g»h
2
Calculated from field
data
Field data
Assumed
Assumed
Calculated by GCAb
Average from field data
Average from data base
b
Source of field data: Reference 38.
Reference 39.
10-43
-------
TABLE 10-21. ESTIMATED VS. MEASURED TOTAL VO EMISSIONS—CASE 3
Time
after
tilling, h
Estimated
emissions,
wt. % total
applied oil
Measured
emissions,
wt. % total
applied oil
68.00
4.5
0.77
TABLE 10-22. INPUT PARAMETERS FOR RTI LAND TREATMENT MODEL'
Parameter
Value
Source
Organic loading
Tilling depth
Soil porosity
Molecular weight of oil
Diffusivity in air
Vapor pressure
Biorate
0.002125 g/cm
20 cm
0.5
282 g/g mol
9.12 E-02
0.76 mm Hg
23. 68 mg V0/g»h
3
Estimated from
field data
Assumed
Assumed
Assumed
Average from data
base
Calculated by GCA
Average from data
base
b
b
Source of field data: Reference 40.
Reference 41.
10-44
-------
parameters that are generally unknown or poorly defined include
waste porosities (air and total), average waste bed temperature
(for active and closed landfills), waste composition at depths
greater than the surface layer, barometric pressures, clay cap
porosities (air and total), clay cap thickness, waste bed depth,
and (for active landfills particularly) time between core sam-
pling and air emissions determination. To apply the models,
representative default values have been used where necessary.
Because of the necessity to estimate key input parameters, the
comparisons that follow are of extremely limited value for model
validation. To achieve validation of emission models, additional
field tests or laboratory experiments are needed for active and
closed landfills and wastepiles.
Field data from two sites were used for comparison with the
land treatment model as applied to active landfills. These sites
(5 and 8) were chosen because of similarity in constituency of
selected chemicals and relative availability of model input
parameters. However, it should be noted that at each of the
sites more than half of the needed model input parameters were
not available from the tests and thus required estimation.
Information on the waste composition within closed landfills
was insufficient to allow use of the closed landfill model. At
two of the three closed/inactive landfill sites (4 and 5), no
solid samples of waste were taken; at the remaining site
(Site 9), a single soil core was apparently extracted from the 3-
ft clay cover, providing no information about the composition of
the waste below the cover. However, it should be noted that
42
Farmer et al. (who developed the precursor to the RTI closed
land-fill model, it accounts for diffusion through the clay cap
only, not barometric pumping) mentioned that their model has
received experimental verification via a laboratory experiment
using hexachlorobenzene-containing waste in a simulated landfill.
Following are the results of the comparison for active
4344 45
landfills at Sites 5 ' and 8 . Table 10-24 presents model
10-45
-------
input parameters used in the application of the RTI land
treatment model to an active landfill at Site 5. Table 10-25
presents a comparison of measured and predicted emission rates
for the Site 5 landfill.
Model predictions for the chemicals assessed here are higher
than field data values by a factor ranging from 13 to 441. This
discrepancy may be largely a result of the presence of daily
earth covers (6 in. thickness) and layers of drums within the
waste bed--neither of which are accounted for by the model.
Other contributing factors may include the estimation of key
model input parameters (e.g., air porosity of waste, temperature
of the constituent within soil) and the nonrepresentative nature
of the waste sample (obtained at the surface) for describing the
waste composition at depth.
Table 10-26 presents model input parameters used in the
application of the RTI land treatment model to an active landfill
at Site 8. Table 10-27 presents a comparison of measured and
predicted emission rates for the Site 8 landfill.
Model predictions of the emissions at Site 8 are, in
general, closer to field data than were the predictions made for
Site 5. Better overall agreement may be due to the absence of
drums and daily earth covers in this landfill.
10.5 TRANSFER, STORAGE, AND HANDLING OPERATIONS
10.5.1 Models Documented in AP-42
Emission methods for the following models were taken from
AP-42; they have been developed from the field data in the
petroleum industry and should be applicable to TSDF:
52
• Container loading (from AP-42, Section 4.4)
53
• Container cleaning (from AP-42, Section 4.8)
54
• Stationary tank loading (from AP-42, Section 4.3)
55
• Stationary tank storage (from AP-42, Section 4.3).
10.5.2 Fugitive Emissions
10-46
-------
Fugitive emission sources have been studied extensively for
the petroleum and Synthetic Organic Chemical Manufacturing
Industries (SOCMI) facilities. These SOCMI emission factors
are assumed to be applicable to similar operations at TSDF.
10.5.3 Spillage An ICF study of truck transport to and from
TSDF and truck emissions at TSDF terminals provided the
information necessary to estimate spillage losses during TSDF and
trucking operation. However, no field test data is available for
comparison.
10.5.4 Open Dumpster Storage Emissions No field data were
available for comparison.
10-47
-------
TABLE 10-23. ESTIMATED VS. MEASURED EMISSIONS—CASE 4
Test
Elapsed
time, day/h
Estimated
emissions,
wt. % total
applied oil
Measured
emissions,
wt. % total
applied oil
5
6
7
1/20
13/307
26/619
5/122
22/520
5.0
14.0
16.0
14.0
28.0
0.14
2.5
13.5
1.1
13.4
10-48
-------
TABLE 10-24. MODEL INPUT PARAMETERS USED IN APPLICATION OF THE
LAND TREATMENT MODEL TO AN ACTIVE LANDFILL AT SITE 5&
Parameter
Value
Data source
L, total organic
loading in soil
C., weight fraction of
constituent i in
organic phase
T, temperature of
constituent vapor
in soil
1, depth of waste in
landfill
• „, total porosity of
waste
• a, air porosity of
waste
Sb, soil biomass
concentration
MWoll, molecular weight
of organic carrier
liquid
t, time between soil
sampling and air
emission measurement
2.65 x 10 3 g/cm3
Inferred from
Xylene: 0.178
field data (solid
sample analysis)
assuming soil
density = 2.3 g/cm3
Field data
Methylene chloride: (solid sample
8.48 x 10~4 analysis)
Tetrachloroethylene:
1.37 x 10
25 °C
229 cm (7.5 ft)
0.50 (50%)
0.25 (25%)
0 g/cm
150 g/g mol
900 s (15 min)
Default value
Default value
Default value
Default value
Default value
Default value
Engineering judgment
10-49
-------
aLandfill 10, General Organic Cell.46'47
10-50
-------
TABLE 10-25. COMPARISON OF MEASURED AND PREDICTED EMISSION RATES
SITE 5 ACTIVE LANDFILL3
Field data result, Model
prediction, „ „
Chemical ug/m »s ug/m »s
Xylene 32.8
Methylene chloride 0.734
Tetrachloroethylene 0.0111
440.0
14.0
4.9
aLandfill 10, General Organic Cell.48'49
10-51
-------
TABLE 10-26 MODEL INPUT PARAMETERS USED IN APPLICATION OF THE
RTI LAND TREATMENT MODEL TO AN ACTIVE LANDFILL AT SITE
,50
Parameter
Value
Source
S~ Q
L, total organic loading 1.71 x 10 g/cm
C., weight fraction of
VO. in organic phase
T, temperature of VO
vapor in soil
1, depth of waste in
landfill
• „, total porosity of
waste
• , air porosity of
awaste
S, , soil biomass
concentration
MW .,, molecular weight
of organic carrier
liquid
t, time between soil
sampling and air
emission measurement
Xylene: 0.012
1,1,1-TCE: 0.19
Tetrachloroethylene
0.096
25 °Ca
229 cm (7.5 ft;
b
b
0.50 (50%)
0.25 (25%)
0 g/cm
150 g/g mol
900 s (15 min)
Field data
in soil
Field
data
(solid sample
analysis)
Default value
Default value
Default value
Default value
Default value
Default value
Engineering
judgment
Soil surface temperatures at this site were reported at 26 to
36 °C. The model unit default value of 25 °C is applied to
the constituent within the soil in this analysis.
b
A single "porosity" value of 31.7 percent was reported for
this site, based on one soil sample. Because this value is
not defined explicitly, and may not be representative of
typical waste in the landfill, the model unit default values
of air porosity and total porosity were applied in this
analysis.
10-52
-------
TABLE 10-27. COMPARISON OF MEASURED AND PREDICTED EMISSION RATES
FOR THE SITE 8 ACTIVE LANDFILL51
Field data result, Model prediction,
2 2
Chemical ug/m »s ug/m »s
Total xylene 6.21 0.23
1,1,1-Trichloroethane 3.57 3.8
Tetrachloroethylene 6.31 1.9
10-53
-------
10.6 REFERENCES
1. GCA Corporation. Air Emissions for Quiescent Surface
Impoundments-- Emissions Data and Model Review, Draft
Technical Note. Prepared for U.S. Environmental Protection
Agency. Contract No. 68-01-6871, Assignment 49. August
1985.
2. Radian Corporation. Hazardous Waste Treatment, Storage, and
Disposal Facility Area Sources: VOC Air Emissions.
Prepared for U.S. Environmental Protection Agency, Emissions
Standards and Engineering Division. Research Triangle Park,
NC. Contract No. 68-02-3850. January 25, 1985.
3. Reference 1.
4. Reference 2.
5. Reference 1.
6. Reference 2.
7. Reference 1.
8. Radian Corporation. Hazardous Waste Treatment, Storage, and
Disposal Facility Area Sources: VOC Air Emissions at
Chemical Waste Management, Inc., Kettleman Hills Facility.
Volume 1. Prepared for U.S. Environmental Protection
Agency, Emissions Standards and Engineering Division.
Research Triangle Park, NC. EMB Report 85-HW5-2. December
1984.
9. GCA Corporation. First Chemical Corporation Wastewater
Holding Lagoon Field Study. Prepared for U.S. Environmental
Protection Agency. Contract No. 68-02-3851 (WA10). 143 p.
August 1986.
10. GCA Corporation. Emissions Data and Model Review for
Wastewater Treatment Operations. Draft Technical Note.
Prepared for U.S. Environmental Protection Agency. Contract
No. 68-01-6871, Assignment 49. August 1985.
11. Reference 10.
12. Reference 10.
13. Reference 10.
14. Reference 10.
10-54
-------
15. Reference 10.
16. Nelson, Thomas P., et al. (Radian). Field Assessment of Air
Emissions and Their Control at Hazardous Waste Facilities
(Draft). Prepared for U.S. Environmental Protection Agency,
Industrial Environmental Research Laboratory. Cincinnati,
Ohio. December 1984. 77 p.
17. Reference 10.
18. Reference 16.
19. Petrasek, A., B. Austern, and T. Neiheisel. Removal and
Partitioning of Volatile Organic Priority Pollutants in
Wastewater Treatment. Presented at the Ninth U.S.-Japan
Conference on Sewage Treatment Technology. Tokyo, Japan.
September 1983. 31 p.
20. Reference 19, p. 2-4.
21. Reference 19, p. 16.
22. Namkung, E., and B. Rittman. Estimating Volatile Organic
Compound Emissions from Publicly Owned Treatment Works.
Journal WPCF. 59 (7) :677.
23. Reference 22, p. 671-672.
24. Reference 22, p. 672.
25. Tabak, H., S. Quave, C. Mashni, and E. Earth.
Biodegradability Studies with Priority Pollutant Organic
Compounds. Staff Report. Wastewater Research Division.
U.S. Environmental Protection Agency. Cincinnati, Ohio.
1980.
26. Reference 25.
27. Coburn, J., C. Allen, D. Green and K. Leese. Site Visits of
Aerated and Nonaerated Impoundments Revised Draft Summary
Report. Prepared for U.S. Environmental Protection Agency.
Contract No. 68-03-3253, Work Assignment No. 3-8. April
1988. p. A-l to A-34.
28. Reference 27.
29. Allen, C. C., et al. Preliminary Assessment of Aerated
Waste Treatment Systems at Hazardous Waste TSDFs (Draft).
Prepared for U.S. Environmental Protection Agency. Contract
10-55
-------
No. 68-02-3992. December 1985. 108 p.
30. Reference 29.
31. Utah State University. Evaluation of Volatilization of
Hazardous Constituents at Hazardous Waste Land Treatment
Sites. Prepared for U.S. Environmental Protection Agency.
Ada, Oklahoma. Undated. 157 p.
32. Radian Corporation. Field Assessment of Air Emissions and
Their Control at a Refinery Land Treatment Facility.
Volumes I and II. Prepared for U.S. Environmental
Protection Agency. Cincinnati, OH. September 1986. 699 p,
33. Radian Corporation. Hazardous Waste Treatment, Storage, and
Disposal Facility Area Sources--VOC Air Emissions. Prepared
for U.S. Environmental Protection Agency. Research Triangle
Park, NC. January 1985. 141 p.
34. GCA Corporation. Air Emissions from Land Treatment--
Emissions Data and Model Review. Draft Technical Note.
Prepared for U.S. Environmental Protection Agency. August
1985. p. 4-36.
35. Suntech, Inc. Atmospheric Emissions from Oily Waste Land
Spreading. Final Report SWM-8(563). Prepared for American
Petroleum Institute. Washington, DC. Undated. 63 p.
36. Reference 31.
37. Reference 32.
38. Reference 33.
39. Reference 34.
40. Reference 35.
41. Reference 34.
42. Farmer, W. J., M. S. Yang, J. Letey, W. F. Spencer, and M.
H. Roulier. Land Disposal of hexachlorobenzene Wastes:
Controlling Vapor Movement in Soils. Fourth Annual Research
Symposium. U.S. Environmental Protection Agency.
Publication No. EPA-600/9-78-016. August 1978.
43. Radian Corporation. Hazardous Waste Treatment, Storage and
Disposal Facility Area Sources: VOC Air Emissions.
10-56
-------
Prepared for U.S. Environmental Protection Agency. Research
Triangle Park, NC. DCN 85-222-078-17-09. January 25, 1985.
141 p.
44. Radian Corporation. Evaluation of Air Emissions from
Hazardous Waste Treatment, Storage and Disposal Facilities
in Support of the RCRA Air Emission Regulatory Impact
Analysis (RIA): Data Volume for Site 4 and Site 5.
Prepared for U.S. Environmental Protection Agency. Cincin-
nati, OH. DCN 83-203-001-63-19. January 11, 1984. 454 p.
45. Radian Corporation. Hazardous Waste Treatment, Storage and
Disposal Facility Area Sources--VOC Air Emissions. Prepared
for U.S. Environmental Protection Agency. Research Triangle
Park, NC. EMB Report 85-HWS-l. May 1985. 54 p.
46. Reference 43.
47. Reference 44.
48. Reference 43.
49. Reference 44.
50. Reference 45.
51. Reference 45.
52. U.S. Environmental Protection Agency. Transportation and
Marketing of Petroleum Liquids. In: AP-42. Compilation of
Air Pollutant Emission Factors. Third Edition,
Supplement 12, Section 4.4. Office of Air Quality Planning
and Standards. Research Triangle Park, NC. July 1979.
13 p.
53. U.S. Environmental Protection Agency. Tank and Drum
Cleaning. In: AP-42. Compilation of Air Pollutant
Emission Factors. Third Edition, Supplement 12, Section
4.8. Office of Air Quality Planning and Standards.
Research Triangle Park, NC. February 1980. 4 p.
54. U.S. Environmental Protection Agency. Storage of Organic
Liquids. In: AP-42. Compilation of Air Pollutant Emission
Factors. Third Edition, Supplement 12, Section 4.3. Office
of Air Quality Planning and Standards. Research Triangle
Park, NC. April 1981. 25 p.
55. Reference 47.
10-57
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56. U.S. Environmental Protection Agency. Control of Volatile
Organic Compound Leaks from Synthetic Organic Chemical and
Polymer Manufacturing Equipment. Research Triangle Park,
NC. Publication No. EPA-450/ 3-86-006. March 1984.
57. U.S. Environmental Protection Agency. Assessing the Release
and Costs Associated with Truck Transport of Hazardous
Wastes. PB 84-224-468 (Contract No. 68-01-0021).
Washington, DC. January 1984. 151 p.
10-58
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11.0 TECHNICAL SUPPORT FOR THE IDENTIFICATION OF
COLLECTION SYSTEMS AS MAJOR EMISSION SOURCES
11.1 INTRODUCTION
This chapter provides a summary of information that is
currently available relevant to the identification of the
potential significance of air emissions from collection systems.
Most of the relevant investigations are included, including two
recent investigations sponsored by the Chemical Manufacturers
Association (CMA). Information obtained from these CMA sponsored
investigations were used by OAQPS to revise the air emission
models for collection systems. Section 11.1 lists the data
sources that are referenced in this chapter, followed by a
summary table for each data source. Although the data are
variable in the type of source tested, the methods used for of
data collection, and types of compounds, the conclusions from the
data analysis in the different studies are generally similar.
These studies support the potential for release of most of a
higher volatility organic material from the wastewater to the
atmosphere during the path of the wastewater from the point of
origin to the wastewater treatment system.
Predictions from theoretical emission models are compared
with field test data in this chapter. In general, considering
the uncertainty of field emission measurements, agreement between
measured and predicted values is considered reasonable. Those
data sets of measured air emissions that were used as the basis
of collection system model parameter selection and data
correlation generally agreed with the models with less than 20
percent difference.
11-1
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Several conclusions are apparent and seem to be supported by
the data. These conclusions are listed as separate sections in
this document, and the data supporting each conclusion is
summarized in each section. The sections in this document that
present the major conclusions are listed below.
Section number
Conclusion illustrated
Section 11-2. page 11-2
Section 11-3. page 11-15
Section 11-4. page 11-19
Information is available which
supports the identification of
collection systems as significant
emission sources.
Organic compounds will volatilize
in the headspace of the collection
system.
Uncontrolled wastewater collection
systems can have significant
discharges of headspace to the
atmosphere.
The fraction of organic compounds
that is lost in uncontrolled
collection systems can be high,
greater than 40 percent.
11.2 SUMMARY OF REFERENCES FOR AIR EMISSIONS FROM COLLECTION
SYSTEMS
Section 11-5. page 11-25
CMA FIELD TEST:
DU PONT
OLD HICKORY
24 -48 percent of organic materials
were volatilized from one run of a
collection main. There were low flow
and high ventilation conditions
during the field test.
11-2
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CMA LABORATORY
TEST: ENVIROMEGA
The air loss from open uncontrolled
drain hubs were tested both with and
without waste flow into the drain hub.
In the case of the model predictions
of drain losses, the data for the open
drain without waste flow demonstrated
6.6 percent loss of toluene from a
drain (run 7). This agrees favorably
with the model predictions of 6.1
percent (BACT\LAER). The fraction
lost from the system was even greater
when waste was discharged into the
drain hub.
50 percent equilibrium was observed
for low air flow in sealed junction
boxes.
Jensen, ET. AL.
Research Journal,
WPCF
30+ percent of a volatile material
(krypton) was volatilized and lost
from one run of a collection main.
SHELL TESTS
Significant concentrations of organic
compounds were detected in the
headspace of a petrochemical
collection system at the Shell
facility. The magnitude of the
headspace concentrations was greater
than the equilibrium values from the
liquid concentrations, (see also
Fingas)
Fingas, et al.
Lab scale collection system.
Investigated the effects of gasoline
spills on the headspace concentrations
in the collection system. Soon after
entering the system, the higher
volatility constituents in the fuel
were lost to the headspace. The
headspace after the initial discharge
contained the lesser volatility
chemicals due to the high loss of the
more volatile chemicals.
11-3
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EPA TRACER
INVESTIGATION:
ROHM AND HAAS
EPA TRACER
INVESTIGATION:
COASTAL EAGLE
POINT REFINERY
Greater than 95% loss of a series of
organic compounds was observed from
discharge into a trench and
conventional sewer. Hexane
disappeared from both the headspace
and the liquid during the flow through
the collection system.
40 -54% loss of organic compounds was
observed from discharge into a series
of open junction boxes with water
seals.
R. L. Corsi,
PH.D. DISSERTATION
Measured concentrations in collection
main which were at equilibrium, Run 1.
Other runs were not at equilibrium.
Review of literature.
R. L. Corsi,
ENVIROMEGA, AND
WASTEWATER
TECHNOLOGY CENTRE
BP OIL TRACER
INVESTIGATION:
LIMA REFINERY
AMOCO REFINERY
27-40% loss of duterated chloroform
and 20-24% loss of ethylene dibromide
was reported in a model of a
collection system drop structure.
30-56% loss of chloroform in a
collection main. The loss was a
function of the sewer temperature.
The fraction lost to the atmosphere
could not be estimated due to lack of
information of organic compound
partition into or out of an oil layer.
17 -60% loss of organic compounds from
the wastewater in a collection main
was estimated from field tests that
measured wastewater concentrations and
flow rates.
11-4
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AMOCO REFINERY The measured collection system losses
(continued) were used to estimate the overall
losses. Greater than 99% loss of
benzene was estimated for the overall
collection system by extrapolating the
measured rates on the basis of opening
areas.
11-5
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TABLE 11-1. LITERATURE REVIEW, I
TITLE: Method for measurement of reaeration in
gravity sewers using radiotracers
AUTHORS: Niels Aagaard Jensen, Thorkild Hvitved-
Jacobsen
SOURCE: Research Journal WPCF, Volume 63, Number 5,
pages 758-767, July/August 1991.
SUMMARY: 30+ percent of a volatile material (krypton)
was volatilized and lost from a reach of a
collection main.
The article states that tracer gas buildup in
the headspace might be significant for sewers.
METHODS: The water flow was 0.02 M3/s in a 0.6 m conduit
with a 1358 m run. To avoid headspace
equilibrium effects, the tracer was injected
into the water. The mass transfer was
calculated from the changes in the ratio of a
volatile tracer (krypton 85) to a non-volatile
tracer (tritium).
COMMENTS: The study appeared to be carefully done, with
excellent reproducibility among three runs. This
study indicates that significant loss of highly
volatile materials may occur even in closed runs of
ventilated conduits, even with compounds dissolved
in water.
11-6
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TABLE 11-2. LITERATURE REVIEW, II
TITLE: Field Evaluation of Wastewater Drain System
Emissions: Rohm and Haas Bristol Facility,
Bristol, Pa., April 1993.
AUTHORS: Radian Corporation 155 Corporate Woods, Suite
100, Rochester, N.Y.
SOURCE: Submerged drain into an open trench. The
wastewater fell into a conventional collection
system.
SUMMARY: Compounds were dissolved into water and added
to wastewater. The loss of compounds from
part of the unit collection system was as
follows: hexane, 98%; acenaphthene 71%,
chlorobenzene, 96%; and 111 trichloroethane,
87%.
METHODS: Loss was measured by the ratio of loss of
volatile constituent to the loss of non-
volatile constituents. The volatilization of
organics was confirmed by headspace
concentrations (not quantified by headspace).
COMMENTS: This investigation indicates the potential for
significant air emissions. The high losses of the
organic compounds that were observed are consistent
with the laboratory data of Fingas.
11-7
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TABLE 11-3. LITERATURE REVIEW, III
TITLE: Industrial Wastewater Field Evaluation of
Wastewater Drain System: Emission Test
Report, Coastal Eagle Point Oil Company,
Westville, N.J. September 1993.
AUTHORS: Radian Corporation 155 Corporate Woods, Suite
100, Rochester, N.Y.
SOURCE: Open drain into water sealed hub. Open
junction boxes with submerged entrance.
SUMMARY: Compounds were dissolved into water and added
to wastewater. The loss of compounds from
part of the unit collection system was as
follows: nitrobenzene, 40%; chlorobenzene,
33%; and 111 trichloroethane, 54%. Covering
the sumps reduced the air emissions to 29%,
33%, and 46%.
METHODS: Loss was measured by the ratio of loss of
volatile constituent to the loss of non-
volatile constituents. The volatilization of
organics was confirmed by headspace
concentrations (not quantified by headspace).
COMMENTS: This investigation indicates the potential for
significant air emissions, even for relatively
controlled systems.
-------
TABLE 11-4. LITERATURE REVIEW, IV
TITLE: Estimation of BTEX Emissions from BP Oil
Refinery Waste Water Collection System. BP
Oil Refining Environmental. Presented at the
Air Toxic Workshop, San Diego, CA, March 31,
1993.
AUTHORS: D. E. Isaacson, BP Oil Refining Environmental.
SOURCE: Collection main that served multiple refinery
units.
SUMMARY: Compounds were dissolved into water and added
to wastewater. The loss of compounds from the
collection system was as follows: chloroform
in the hot zone of collection system, 56%
average; and chloroform in the normal
temperature zone, 30%, with an overall loss
greater than 80%..
METHODS: Loss was measured by the ratio of loss of
volatile constituent to the loss of non-
volatile constituents. Lithium chloride was
used to define liquid flow rates.
COMMENTS: This investigation indicates the potential for
significant air emissions from collection mains.
11-9
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TABLE 11-5. LITERATURE REVIEW, V
TITLE: Measurement of Hazardous Air Pollutant
Emissions from Wastewater Collection System
Components.
AUTHORS: Enviromega Ltd. PO.Box 1249, Burlington, Ont.
L7R 4LB, April, 1993. CMA sponsored research.
SOURCE: Drop structures, process drains, open
discharge into a water sealed hub, open
discharge into an unsealed drain hub.
SUMMARY: The air loss from open uncontrolled drain hubs
were tested both with and without waste flow
into the drain hub. In the case of the model
predictions of drain losses, the data for the
open drain without waste flow demonstrated 6.6
percent loss of toluene from a drain (run 7).
This agrees favorably with the model
predictions of 6.1 percent (BACT\LAER).
The fraction lost from the system was even
greater when waste was discharged into a water
sealed drain hub. The air emissions from the
open discharge of waste into a water sealed
drain ranged from 10% to 60%, depending on the
flow rate and compound volatility.
Methanol was significantly volatilized under
the test conditions.
METHODS: Toluene, tetrachloroethylene, trichloroethene,
and 1,4 dichlorobenzene were added to a sewer
reach. The concentrations added to the sewer,
the concentrations in the gas, and the
ventilation rates were measured.
COMMENTS: The data from this investigation were used to
revise the air emission models for specific
collection elements.
11-10
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TABLE 11-6. LITERATURE REVIEW, VI
TITLE: Volatile and Semi-Volatile Organic Compound
Emissions from Sewer Drop Structures.
AUTHORS: R. L. Corsi et al. University of Guelph; H.
Monteith, Enviromega Ltd. PO.Box 1249,
Burlington, Ont.; and H. Melcer, Wastewater
Technology Center, Burlington, Ontario.
SOURCE: Drop structures.
SUMMARY: 27-40% loss of duterated chloroform and 20-24%
loss of ethylene dibromide was reported in a
model of a collection system drop structure.
Methanol was significantly volatilized under
the test conditions.
METHODS: The air loss from a drop structure was tested
by measuring the concentrations of ethylene
dibromide and duterated chloroform in the
inlet and exit of the drop structure. The
concentrations added to the sewer, the
concentrations in the gas, and the ventilation
rates were measured.
COMMENTS: The authors conclude that data from this
investigation indicate that sewer drop structures
may be significant contributors to overall semi-VOC
and VOC emissions from sewers.
11-11
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TABLE 11-7. LITERATURE REVIEW, VII
TITLE: Sampling for Fugitive HAP Emissions From A
Wastewater Collection Sewer Reach at the
DuPont Old Hickory Site.
AUTHORS: Enviromega Ltd. PO.Box 1249, Burlington, Ont.
L7R 4LB, September, 1993. CMA sponsored
research.
SOURCE: 165 m (450 ft) reach of a sewer. The
wastewater depth was very shallow, 2 to 5
inches.
SUMMARY: "Using the measured dose solution
concentration the percentage VOC emissions
ranged from 24% to 48%. Using the target
concentration of VOCs in the dose solution,
the percentage emissions ranged from 16% to
29%." (p.45)
METHODS: Toluene, tetrachloroethylene, 1,1,2,2
tetrachloroethane, and 1,4 dichlorobenzene
were added to a sewer reach. The
concentrations added to the sewer, the
concentrations in the sewer, and the mass air
emissions were measured.
COMMENTS: "The consistency and magnitude of the decreases [in
wastewater concentration] suggests substantial
losses of VOCs to the atmosphere." (p.36)
11-12
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TABLE 11-8. LITERATURE REVIEW, VIII
TITLE: Ph.D. Dissertation of R. L. Corsi, University
of California, Davis, 1989.
AUTHORS: R. L. Corsi
SOURCE: Collection main
SUMMARY: Compounds were added to wastewater. The
concentrations of organic compounds in the
headspace of the collection main were
approximately the same as equilibrium values
for Run 1. The concentrations were less than
equilibrium for other runs.
METHODS: Loss was measured by the concentrations of
organic constituents and the flow rate. The
volatilization of organics was quantified by
headspace concentrations and flow rates.
COMMENTS: This thesis indicates the potential for significant
air emissions. A literature review is included.
11-13
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TABLE 11-9. LITERATURE REVIEW, IX
TITLE: Fuels in Sewers: Behavior and countermeasures,
Journal of Hazardous Materials 19 (1988) 289-
302.
AUTHORS: M. F. Fingas, K. A. Hughes, and A. M. Bobra
SOURCE: Laboratory collection system.
SUMMARY: Vapors in sewers have two distinct origins,
entry vaporization and mass transfer during
transport. The wastewater loses the more
volatile components during entry.
METHODS: Gasoline was spilled in one end of the
collection system. The volatilization of
organics was confirmed by headspace
concentrations and liquid concentrations.
COMMENTS: This investigation indicates the potential for
significant air emissions, even for relatively
controlled systems.
11-14
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TABLE 11-10. LITERATURE REVIEW, X
TITLE: Method 25D Development and Testing at the
Shell, Deer Park Industrial Wastewater
Facilities, June 1991.
AUTHORS: Radian Corporation 850 Mo-Pac Boulevard,
Austin, Texas
SOURCE: Collection system in a large petrochemical
complex. Sources included drains, sumps, and
manhole covers.
SUMMARY: Significant concentrations of organic
compounds were detected in the headspace of a
petrochemical collection system. The
magnitude of the headspace concentrations was
as great or greater than the equilibrium
values from the liquid concentrations.
Screening tests of headspace in equilibrium
with the liquid confirmed that the headspace
concentrations were greater than equilibrium.
METHODS: The volatilization of organics was confirmed
by headspace concentrations, measured
headspace flow rates into the air, and
concentrations in the liquid.
COMMENTS: This investigation indicates the potential for
significant air emissions, because of high
headspace gas concentrations.
11-15
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TABLE 11-11. LITERATURE REVIEW, XI
TITLE: An Air Quality Evaluation of the Amoco
Yorktown Refinery as Part of the Amoco-USEPA
Pollution Prevention Project., September 13,
1991
AUTHORS: Amoco Corporation Environmental Affairs and
Safety Department.
SOURCE: Collection system in a large oil refinery
petrochemical complex. Sources included
drains, sumps, and manhole covers.
SUMMARY: Three sewer openings were tested, using an
enclosure and a vent opening of 0.27 ft2. An
average of 0.2 Mg benzene /yr was reported
from the small vent opening, based upon three
sets of tests. The refinery collection system
openings were estimated as 1200 ft2, suggesting
very large amounts of benzene lost from the
collection system if the reported losses of
benzene were extrapolated on the basis of area
of openings. The vent rate from the small
vent opening was 200 ft/min, consistent with
vent rates for manhole covers reported in
Section 3 of this document.
Material balances suggested organic compound
loses of up to 60 percent in the collection
main. The fraction lost to the atmosphere
could not be determined because the
concentrations in the oil layer were not
determined. Organic compounds could partition
into or out of the oil layer.
METHODS: Existing operation of the refinery was tested
with concentrations in the vented gas,
concentrations in the water, and gas flowrate
measured.
COMMENTS: This investigation indicates the potential for
significant air emissions, because of high
headspace gas concentrations. These data seem to
be consistent with the data of Fingas and the EPA
vent rates.
11.3. ORGANIC COMPOUNDS WILL VOLATILIZE IN THE HEADSPACE OF THE
11-16
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COLLECTION SYSTEM
11.3.1 Importance of Gas Concentrations.
The gas concentrations in the collection system headspace
are important because they determine the air emissions. The rate
of air emissions is the product of the concentration in the
headspace and the flow rate of the headspace out of the
collection system and into the air.
It should be noted that the discharge of volatile material
into existing wastewater collection systems can lead to dangerous
and explosive conditions in the collection system headspace.
None of the comments presented in this section should be
interpreted that such a potentially dangerous condition could not
exist. The discharge of organics into a collection system is not
intrinsically safe, and any explosion potential should be
carefully evaluated and properly handled on an individual basis.
11.3.2 Industry Comments
Open wastewater drains at some companies are no longer used
when benzene concentrations are 50 ppm or more1. The widespread
replacement or covering of open drains for environmental reasons
would reduce the rates of benzene loss from that section of the
collection system. Many other commenters have suggested that
open drains are common at other companies.
For safety reasons junction boxes and sumps are sealed and
purged with an inert gas atmosphere. This reduces or eliminates
air flow through open units2. The use of these techniques would
eliminate air flow directly to the atmosphere from these units
(when the seals work as designed). Opening the sumps to the air
would not assure that explosive concentrations would not be
present.
11.3.3 Fingas, et al.
Fingas3 modeled a collection system to investigate the
effects of gasoline spills on the headspace concentrations in the
11-17
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collection system. Soon after entering the system, the higher
volatility constituents in the fuel were lost to the headspace
(explosion hazard) . Downstream of the point of introduction of
the fuel, the headspace of the collection system contained lower
boiling compounds. These data suggest that much of the highly
volatile compounds can be lost to the headspace early in the path
through the collection system. This also suggests that the
concentration of organic compounds in the headspace can be much
greater than expected from the liquid composition in the
wastewater.
11.3.4 Shell Petrochemical Facility
A comparison of measured and predicted estimates of air flow
rates from the wastewater collection system was developed from
data collected at that facility4.
The gas concentrations were measured and calculated by
several different methods for a number of different locations in
the Shell petrochemical facility wastewater collection system.
This paragraph briefly describes the different types of data that
were taken for evaluating the fate of organic compounds in the
Shell petrochemical facility collection system.
11.3.4.1 TLV sniffer. The TLV sniffer was used to measure
concentrations in the gas phase in the collection system
headspace and in the "jar test" of equilibrium headspace.
11.3.4.2 HNu. The HNu sniffer was used to measure
concentrations in the gas phase in the collection system
headspace and in the "jar test" of equilibrium headspace.
11.3.4.3 Canister. A test probe was introduced into
the headspace of the collection system conduits and a sample of
the headspace gas was withdrawn into an evacuated canister. The
concentrations in the canister were analyzed on a compound
specific basis, so that total organic values could be estimated
by summing the individual compound values.
11.3.4.4 Equilibrium headspace. A sample of the liquid in
11-18
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the wastewater collection system was withdrawn by a bailer and
used to fill a jar half full of liquid. The jar was capped and
was shook vigorously for a few seconds. The cap was opened
enough to briefly measure the headspace concentration in the jar
by the TLV sniffer and by the HNu sniffer.
11.3.4.5 Liquid sample analysis. A sample of the liquid in
the wastewater collection system was withdrawn by a bailer and
used to fill 40 ml VGA vials. The concentration of compounds in
the wastewater was measured by Method 8240. The concentration of
the compounds that would be in equilibrium with the liquid
concentration was calculated by multiplying the liquid
concentrations by the Henry's law constants. The Henry's law
constants were obtained from the pure component data base of
ChemdatS.
11.3.4.6 Results of tests. Ratios of headspace and liquid
equilibrium concentrations indicate that in general, the
headspace concentration was much greater than the corresponding
equilibrium value with the liquid. The following list
illustrates that the organic content of the headspace was greater
than the equilibrium value with the liquid. The headspace ratios
presented in the list are the value of organics measured in the
canister to the theoretical values calculated from the measured
liquid concentrations.
• location Ib, ratio: 13
• location 5, ratio: 36
• location 5c, ratio: 33
• location 6, ratio: 61
• location 10, VCM plant discharge, ratio: 18
• location Ib, ratio: 13
CMA representatives indicated that the collection system was
likely warmer than the 25 C assumed for the Chemdat? data base.
They suggested that the Shell headspace concentrations would be
11-19
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more appropriately represented by equilibrium, or a ratio of 1.
The concentrations in the headspace of the collection main
are much greater than expected. This suggests that the
wastewater may have lost significant quantities of organics early
in the path of the waste through the wastewater collection
system. The higher concentrations in the headspace were not
removed by the flow of wastewater under the headspace. These data
could be interpreted to suggest that there is a limited rate of
mass transfer between the headspace and the waste.
The predicted and observed rates of headspace flow in the
petrochemical wastewater collection system is presented in Table
12. These data indicate that, although the flow velocity was
highly variable, the magnitude of the velocity was comparable to
the predicted value.
11.3.5 Ph.D. Dissertation of R. L. Corsi
The data obtained by Corsi in municipal wastewater
collection systems clearly demonstrate that equilibrium can be
established in collection system mains. The measured
concentrations of chloroform in the headspace of Run 1 was almost
exactly equal to the theoretical equilibrium value of the
headspace concentration. The data also indicated that the
assumption of equilibrium may not be true for high waste flow in
very large collection mains. Corsi also presents data supporting
the importance of equilibrium in determining air emissions from
collection systems. In Run 1, the following data were obtained
by Corsi:
11-20
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TABLE 12. COLLECTION SYSTEM EQUILIBRIUM DATA
Measured parameter
Concentration in gas (outlet,
Table 6-13)
Concentration in liquid
(measured in bag, calculated for
inlet, Table 6-9)
units
mg/m3
mg/m3
value
1.3
9.6
Calculated parameter
Henry's law constant
(ratio of above concentrations)
Henry's law constant
(theoretical, Table 6-4)
units
mg/m3 gas per
mg/m3 liquid
mg/m3 gas per
mg/m3 liquid
value
0.135
0.12
11-21
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These data demonstrate that the assumption of equilibrium in
Run 1 could be valid. Other runs do not necessarily illustrate
equilibrium.
Page 170: Mass transfer models were used to estimate the
loss of organic compounds from waste collection system
components. Rapid saturation of the wet well atmosphere was
observed for a lower volatility organic material. A greater
removal of vinyl chloride (18%) was predicted from the same
model. This theoretical calculation suggests that equilibrium
can be established in collection mains.
Page 176: Both the Henry's law constant and the mass
transfer coefficient can increase with temperature.
Theoretically, the chloroform standard loss value from
residential sewers increased from 28% to 46% at higher
temperatures. This theoretical calculation suggests that
equilibrium values can be important in determining air emission
losses from collection systems.
Page 181: Conclusion 2. Discharge of organic compounds to
smaller interceptors located five kilometers or more from a
treatment plant can lead to emissions comparable to those at the
treatment facility. This is particularly true during periods of
low wastewater flow, high ventilation flow rates, or for VOCs
with high Henry's law constant. This conclusion states that
equilibrium values are important in determining air emission
losses from collection mains.
Page 182: Conclusion 3. High gas flows in combined
sanitary/storm sewers should approach infinite ventilation
conditions, and much higher relative emissions than would be
expected in separate sanitary sewers. Under conditions of high
ventilation the mass transfer is not restricted by the
limitations of equilibrium. The converse is that equilibrium is
important at lower ventilation conditions.
Page 182: Conclusion 7. Rapid organic compound
11-22
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accumulation in sewer atmospheres leads to low organic compound
losses from drops, unless high ventilation rates are present,
e.g. forced ventilation. This statement indicates that
equilibrium is rapidly achieved in collection systems, and this
equilibrium can be the rate controlling factor for air emissions
(equilibrium can limit air emissions).
Page 183: Conclusion 9. Elevated wastewater temperature,
e.g. 40°C as opposed to 20°C, significantly increases organic
compound emissions by increasing the mass transfer coefficient,
Henry's law constant, and buoyancy-driven ventilation. This
conclusion indicates that equilibrium partitioning is a
significant factor in determining the magnitude of air emissions
from collection systems.
Page 184: Conclusion 4. Sensitivity of emissions to
organic compound characteristics increases at lower ventilation
rates, where Henry's law constant is the dominant physico-
chemical property. This conclusion indicates that equilibrium
partitioning is a significant factor in determining the magnitude
of air emissions from collection systems at lower ventilation
rates.
11.4 UNCONTROLLED WASTEWATER COLLECTION SYSTEMS CAN HAVE
SIGNIFICANT DISCHARGES OF HEADSPACE.
The rates of loss of organic materials from collection
systems depend on both the ventilation rates of the collection
system and the concentrations in the headspace of the system.
The data presented in this section indicate that the discharge of
headspace from collection systems are at least as great as the
model predictions.
CMA indicated that a realistic flow of air from an open
drain is 1.9 ft/sec (114 ft/min). The current collection system
emission model and our field data indicates a slightly lower flow
velocity than the CMA value. Other data, including industry
reported rates at Amoco, indicate higher air flow rates than this
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estimate.
11.4.1 Shell facility
This petrochemical facility was not ideally suited for the
verification of the velocity predictions because the system
generally had sealed manhole covers. The system had elevated
vents which were inaccessible for testing. These vents are
expected to divert air flow away from surface openings in the
collection system. The weather was overcast and raining on one of
the sampling days. The wind direction was variable, the wind
speeds were low and variable, and the wind direction also
depended on the location relative to buildings. The measured
ventilation rates at the Shell facility are presented in Tables
11-13, 11-14, 11-15, and 11-16.
11.4.2 Velocity screening at Pulp Mills
A series of 5 pulp mills were vistied and ventilation rates
were measured at collection system openings. The data were
reported in Table 11-13.
11.4.3 Rohm and Haas Chemical Plant Collection System
There was a strong relationship between the wind speed and the
air flow in the waste collection system. The data was taken at
random times on the first day of testing to learn about the flow
patterns of gasses in the collection system and to identify
appropriate sampling conditions. The collection system vent
rates were into the system or out of the system, depending on
wind speed and direction. The wind speed changes the flow and
the flow direction of the air in the waste collection system. At
low wind speeds, the flow of the water in the system produces a
suction on the surrounding air. At high wind speeds, the
pressure of the air causes flow out of the system. When wind
flows in the direction of the opening of a conduit, the wind
pressure can cause a flow of air out of the collection system,
removing organic compounds from the headspace of the collection
system. In the system with the absence of wind, the flow of the
11-24
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water induces air to flow along with the waste, carrying air into
the system and discharging the air at some point downstream.
The concentrations and velocities were measured in the
headspace venting during the tracer tests at Rohm and Haas. This
confirmed the potential for air emissions.
11-25
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11.4.4 Velocity screening at industrial plants
The wastewater collection systems of several industrial
plants were inspected. Screening measurements were made at these
sites with an Alnor velometer (low velocity probe) to evaluate
the magnitude of collection system emissions air flow rates. The
primary concern was to determine the magnitude of the measured
air flow velocities and to compare these measured velocities with
the predicted velocities from the air emission models.
Based upon the results of the velocity screening
measurements, the following observations can be stated:
• The velocities from the openings in the wastewater
collection systems were variable, but the ranges of
velocities from the different sources were not highly
variable.
• The preliminary results of using the model for site specific
conditions was favorable, with reasonable agreement (factor
of 2) between the predicted and measured velocities.
• To improve the agreement between the model results and the
observed collection system velocities, model plant
parameters that were more representative were used in the
theoretical estimations. Examples of sources that were
modified include open drains under grates (higher emissions
than predicted) and sealed drain system (lower emissions
than predicted).
11-5. THE FRACTION OF ORGANIC COMPOUNDS LOST IN UNCONTROLLED
COLLECTION SYSTEMS
11-5.1 Shell facility
A formal material balance was not possible in the Shell facility
wastewater collection system because the liquid flow rates and
11-26
-------
TABLE 11-13. SCREENING VALUES FOR AIR VELOCITIES
AT COLLECTION SYSTEM OPENINGS
Location Velocity at opening a
Chemical collection system, Shell
Drain grate, Shell
Chemical collection system, Shell
Open drain, Shell
Manhole (1 in dia. opening), Shell
Open drain, Shell
Closed drain, Shell
Opening, Shell
Sump opening, Shell
Sample point drain, Shell
Sample point drain, Shell
Chemical collection system Sump, Shell
Drain opening, Shell
Horizontal flow in collection system,
PULP MILL 2
Manhole cover, Pulp mill 2
Manhole cover, Pulp mill 2
PULP MILL 3
Grate over collection system
Grate
Lift station opening
Lift station opening
PULP MILL 4
Floor drain (2x1 ft)
Floor trench, (1-2 ft/s)
Floor trench
Main drains between process units
Small vent on main collection conduit
Grate on main collection conduit
Grate on main collection conduit
Grate open drain
PULP MILL 5
Grate at end of trench
f t/min
80
65
200
40
60
120
150
110
50
300
100-150
50-100
50
40-50
60
50
50
0-50
100
>300
150
160-170
50
100
ft/sec
3
2
1.3
1.1
3.3
0.66
1.0
2
0-2
2.5
0 (slow fumes
out)
1.8
0-1
0.83
5.0
1.7-2.5
0.83-1.7
0.83
0.67-0.83
1
0.83
0.83
0-1
1.7
5
2.5
2.7-2.9
0.83
1.7
Measured with an Alnor velometer, low velocity probe. Ft/sec
reported without corresponding ft/min obtained by visual
inspection of plume rise.
11-27
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TABLE 11-14. A COMPARISON OF PREDICTED AND MEASURED
AIR VELOCITIES AT WASTEWATER COLLECTION SYSTEM OPENINGS
Location
Velocity at opening (ft/min)
measured
predicted
Drain, Shell
Manhole cover, Pulp mill 2
Manhole cover, Pulp mill 2
65,40,60,120
300
100-150
122 (A2
124 (B3
191 (B2
b
Average drain velocities 67
Average manhole cover opening
velocities 198
Average junction opening velocities 88
12!
66
Measured with an Alnor Velometer, low velocity probe.
Predicted values based upon a wind velocity of 3.5 MPH (300
ft/min).
b
Wind effect (3.5 MPH) probably overestimates emissions
because actual wind velocity was lower (tall process units,
wind direction different from stack emissions at high
elevation)
11-2!
-------
TABLE 11-15. SCREENING OF AIR FLOWS AND GAS CONCENTRATIONS
FROM SHELL CHEMICAL COLLECTION SYSTEM SOURCES: MAIN PLANT a
Velocity
Location time (ft/min)
6
6
7
5
5A
10
5C
11
IB
14:22 150
14:43 40
15:08 -150
15:12 65 to 75
15:26 low +
15:28 230
210
<0
0 to 60
Q
Concentration (ppm) Headspace test (ppm)
HNu
120
130
1
80
60
60
130
2
150
TLV HNu TLV
3100 25-30 120
3000
20 4 60
1200 6 0
5000 20 0X100
5900 1 35
120
6900 20 300
b
Measurements 2/6/90, main process areas leading to wastewater
treatment. Included are tank farms, alcohol distillation,
vinyl chloride monomer production, plant E, plant BFA,
spillage, and flows from other process areas. Test points are
locations at Shell.
Measured with an Alnor velometer, low velocity probe.
Positive values indicate flow out of the collection system.
Flow rates less than the detection limit of the instrument (20
ft/min) are indicated as low if they could be detected by
visual inspection.
A sample of wastewater was taken from the sample point and
placed in a glass jar. The jar was capped and shaken
vigorously. The sample probe of the test instrument was
inserted into the headspace of the partially open jar. The
reading was taken when reaching steady state, usually 5-10
seconds.
11-29
-------
TABLE 11-16. COMPARISON OF MEASURED AND PREDICTED AIR FLOWS
FROM SHELL CHEMICAL COLLECTION SYSTEM SOURCES: PHENOL ACETONE3
Measured Velocity Predicted
Velocity
Location time (ft/min)b (ft/min)c
sump(water) 9:50
sump(oil) 50 124 (B3;
A drain 120-150 122 (A2;
D manhole
B drain 30 122 (A2;
C manhole 150 124 (B3)
sump(water) 17:23
sump(oil) 30-40 124 (B3;
A 17:37 200 128
(average)
a Measurements 2/6/90, phenol acetone units. The major area
flows were sample point A, dephenolator extraction bottoms 80
gpm, and sample point B, cumene decanting wastewater 20 gpm.
Reference Figure 3. The weather was overcast and raining.
b Measured with an Alnor velometer, low velocity probe.
Positive values indicate flow out of collection system. A
range of velocities indicates that the meter reading was
rapidly cycling between the limits of the range.
c The flow rates of air out of the unit are estimated by the
protocol described in this document, with a wind velocity of
3.5 MPH Actual wind velocity at the site were variable and
generally less than 5 MPH Weather conditions were overcast
with some rain. The case used to represent field conditions
is presented in parenthesis.
11-30
-------
concentrations of all the branch streams were not characterized.
From estimates of liquid flows in the various points in the main
collection branch it is possible to establish a mass flow rate of
each constituent by multiplying the concentration of the
constituent by the liquid flow. This estimate of the material
balance is of limited significance for some compounds due to
these factors, but a more certain conclusion can be stated for
the fate of chloroform.
The simplified material balance is presented for 1,2
dichloroethane and chloroform in Tables 11-4 and 11-5.
1,2 dichloroethane and chloroform were not significantly
present in the collection main before the vinylchloride monomer
(VCM) plant discharge. After the point of discharge from the VCM
plant, the 1, 2 dichloroethane in the liquid was retained at
roughly the same mass flow rate down the collection main.
Chloroform disappeared from the collection main shortly after
discharge from the VCM plant. The total organics in the
headspace at the discharge point from the VCM plant contained
much higher concentrations of organics than expected from the
equilibrium values in the liquid.
These material balances suggest that relatively little of the
low volatility materials are removed from the liquid in the main
collection system flow:
• acetone (y/x) = 1.4
• 1,2 dichloroethane (y/x) = 66
The addition of unknown side streams containing dichloroethane
could possibly interfere with the material balance and cause an
underestimate of the losses of dichloroethane.
11-31
-------
Table 11-17. A SIMPLIFIED MATERIAL BALANCE FOR 1,2
DICHLOROETHANE IN THE COLLECTION MAIN, SHELL.
Location
5
10
5C
IB
Flow
(gpm)
534
711
962
1217
Concentration
(ppmw)
0
1.84
1.04
1.14, 1.22
Flow rate of
compound
(g/min)
0
4.9
3.6
5.2,5.6
This table demonstrates that the flow rate of 1,2
dichloroethane did not substantially change in the
collection main from location 10 to location IB.
TABLE 11-1!
A SIMPLIFIED MATERIAL BALANCE FOR CHLOROFORM IN
THE COLLECTION MAIN, SHELL.
Location
5
10
5C
IB
Flow
(gpm)
534
711
962
1217
Concentration
( ppmw )
not present
.11
not present
not present
Flow rate of
compound
(g/min)
0
0.29
0
0
This table demonstrates that the flow rate of chloroform
was significantly lowered in the collection main from
location 10 to location IB.
The concentration of chloroform in the gas phase was 73.2
ppmv at location 10. This was among the top 12 gas phase
constituents. There was a steam discharge from location
10 and the location received wastes from vinyl chloride
manufacturing.
11-32
-------
The material balances suggest that higher volatility materials
such as chloroform are significantly and possibly entirely lost
in the main collection system flow:
• chloroform (y/x) = 188
The addition of side streams to the main collection system
flow was not of great enough volumetric flow rate to account for
the disappearance of chloroform from the system by dilution;
therefore, the loss of chloroform is assumed to be due to
volatilization.
11.5.2 Rohm and Haas Tracer Investigation
Tracers are compounds that are introduced into a wastewater
stream to measure the rate of release of the organic compounds
due to air emissions from the collection system. To measure the
losses of organic compounds in wastewater collection systems the
concentration of the tracers are measured in the water phase and
in the air leaving the system.
Four organic tracers were continuously added to the collection
system at one of the points of waste discharge into the
collection system. The tracers were selected from a combination
of volatile material, semivolatile material, and non-volatile
materials. The concentrations of the tracers were measured down-
stream of the tracer addition point. The ratio of the
concentrations of the tracers that were measured were then
compared to the initial ratio, so that the loss from the system
could be calculated, as well as the relative loss from the
system.
The concentration of volatile tracer and the flow rate in the
collection system permit a calculation of the flow rate of
tracer. A comparison of this flow rate to the rate of addition
of tracer permits an independent assessment of the fraction of
the volatile constituent that is lost as air emissions.
The tracer was added at location A. Location A was a trench
outside a process building. Water flowed in the trench at a
11-33
-------
fairly high rate. The drain pipes from a process centrifuge
discharged into the trench. The plant has reasonably steady flow
of water into the collection system at Location A. The flow into
the wastewater accumulation tank (Location D) was measured by the
height increase in the tanks when there was no wastewater being
pumped out of the tanks.
Three separate peristaltic pumps were used to deliver the
organic materials to the trench stream. The non-volatile metals
were dissolved in water, chlorobenzene and chloroethane were
combined in a tank, and hexane and acetonaphthene were combined
in a third tank. The solutions were dripping from the exit tubes
(subsurface) at a rate of approximately 1.5 drops per second.
The hexane drop rose to the surface of the flowing water below
the exit tubes, and rapidly flashed from the surface. As the
hexane was removed from the surface, the acetonaphthene was
deposited on the surface (very thin light refractive layer). By
the end of the channel, the circulation patterns in the water had
collected the acetonaphthene toward the center of the water's
surface.
The four sample points in the wastewater collection system
were designated A, B, C, and D. The concentrations in the
wastewater were monitored during the test period. A summary of
the results of the tracer investigation is presented below.
There were greater than expected loss of each volatile material
in the collection system. Hexane was lost to a significant
extent in the drain channel before it entered the subsurface
collection system, as indicated by visual and olfactory
observations. Hexane was apparently lost quickly by flashing
early in the path through the collection, but the residual hexane
also was lost significantly between points B and C and between
points C and D.
11-34
-------
Compound Average fraction lost Henry's Law(y/x)
hexane 0.977 4270
acenaphthene 0.713 6.3
chlorobenzene 0.959 209
111 trichloroethane 0.866 967
The data suggest that acenaphthene may have been significantly
removed as air emissions, but there was the potential for
sampling and analytical difficulties for this compound. The
canister headspace sampling program did not include acenaphthene
analysis.
Chlorobenzene was lost significantly as the tracer flowed
through the collection system. The concentrations detected at
location A were greater than expected from the rate of release of
chlorobenzene into the collection system. The reason for the
high concentrations measured at location A may be related to the
location of the sampling line. The overall loss of chlorobenzene
in the collection system was much greater than expected from the
air emission models.
As in the case of chlorobenzene, 1,1,1-trichloroethane was
lost significantly as the tracer flowed through the collection
system. The concentrations detected at location A were greater
than expected from the rate of release of trichloroethane into
the collection system. The reason for the high concentrations
measured at location A may be related to the location of the
sampling line. The overall loss of 1,1,1-trichloroethane in the
collection system was much greater than expected from the air
emission models. Chlorobenzene and 1,1,1-trichloroethane were
lost significantly between points B and C and between points C
and D.
11-35
-------
TABLE 11-19. THE MEASURED RELEASE OF TRACER CONSTITUENTS FROM
THE COLLECTION SYSTEM HEADSPACE BY CANISTER MEASUREMENT AND
AIR FLOW RATE MEASUREMENT.
Location
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
total
released
1,1,1
trichloroe thane
(mg/min)
689
265
25.9
70.9
4.48
0.28
28.4
0.036
0.036
83.9
0.062
0.31
102
2.19
166
1438.5
1367.3
chloro-
benzene
(mg/min)
77.9
34.9
9
12.8
4.8
0.3
3
0.028
0.028
24
27.6
12.4
32
238.8
1118.5
hexane
(mg/min)
31.7
12.8
1.8
1.77
0.096
0.006
0.14
0.002
0.002
0.41
0.45
0.169
0.32
49.7
775.2
This table illustrates that significant rates of loss of
tracer materials were detected leaving selected collection
system openings. This table is not to be interpreted as a
material balance, because all sources of air release were
not monitored, and because these sources are variable in
flow (results based on one point sampling) .
11-36
-------
The loss of the non-volatile tracers (copper and zinc) through
the collection system was much less than the loss of organic
compounds. These data confirm that the ratio of the
concentrations of the higher volatility organic compounds to non-
volatiles decrease during the flow through the collection system,
indicating a substantial loss of organic materials. The estimate
of zinc loss was 21 % and the copper loss was 38%. The
appearance of a loss for the tracer metals was likely due to
errors in the wastewater flow measurement.
The results of the canister sampling of the headspace of the
collection system, together with the flow rates of the collection
system headspace gas leaving the system, indicate that
volatilization of the tracer constituents into the headspace is a
significant pathway for tracer loss in the collection system.
The following table presents a summary of some of the measured
releases of tracer constituents into the environment, based on
limited sampling at selected openings in the wastewater
collection system. It was observed that tracer organic compound
loss was observed from openings in the collection system that
were distant from the tracer wastewater flow path.
In summary, the tracer testing at the chemical plant indicated
high loss of organic compounds. Significant concentrations of
organic compounds were detected in the headspace. The magnitude
of the measured emission rates in the gas were consistent with
the loss from the liquid. No material balance was possible due
to variable flow and many vent locations. If the absence of
hexane in the liquid was due to hexane floating on the surface,
the headspace concentrations would contain hexane. The headspace
analysis showed very little hexane, but contained more of the
lesser volatile components, indicating almost total loss of
hexane.
11.5.3 Amoco refinery material balance
11-37
-------
The waste concentrations of benzene, toluene, and several
other compounds were measured at various points in the waste
collection system of the Amoco refinery between the production
units and the wastewater treatment plant. Based on those
measurements, a material balance was carried out to identify the
potential for the release of air emissions.
The material balance results indicated a disappearance from
the wastewater but the high loss of organic compounds could have
been only apparent because of sampling problems. Partitioning
into an oil phase that was present in the wastewater was not
accounted for. The mechanism for a high loss in a collection
main could involve the organic compound partitioning into an oil
phase, and the oil phase floating to the top of the waste stream.
The air emission models for the waste collection system
predict a few percent loss in the part of the system
investigated. The CMA estimates with a trench model that
approximately 1 percent organics are lost per collection system
run (between manholes). With the CMA approach the maximum loss
would be expected to be less than 5% between the unit exits and
the API separator inlet. The collection system model predicts
much higher air emissions near the waste entrance (not measured)
than downstream in a collection main (where the data were taken).
The loss of organic compounds in the entire system is therefore
theoretically expected to be substantially greater than the loss
of organic compounds in the section of the collection system
investigated by the Amoco refinery.
The data suggests the potential for significant fractions of
the organic compounds that may be lost to the air in the
collection mains, much greater than the collection system trench
models estimate. If the data are correct, they would suggest
that the CMA estimates and the collection system model estimates
using the limits of mass transfer from the bulk of the liquid
waste may omit significant factors that control the release of
11-38
-------
organic compounds in the collection system. Table 11-20
summarizes the results of the material balances.
11.5.4 Amoco Refinery Vent Measurements
This refinery was tested for ventilation rates by enclosing
and placing a vent over selected openings in the collection
system. The vent rate was 200 ft/min. This measured vent rate
was consistent with the average manhole cover vent rate measured
in the EPA screening program. The air models suggest that
increasing the area of the opening will increase the total flow
of air from the system, and therefore increase the air emissions.
The average benzene loss for this vent was 0.2 Mg benzene per
year. One way to estimate the overall collection system
ventilation rate is to extrapolate on the basis of vent rate and
the opening area. If the vent rate were reduced to the average
value for large openings in the collection system (88 ft/min)
TABLE 11-20. ESTIMATED DISAPPEARANCE OF VOLATILES FROM THE
WASTEWATER OF THE COLLECTION MAIN AT THE AMOCO REFINERY.
Location of loss
Unit exit to F04a
F04 to API separator
total loss to
separator inlet3
API separator inlet to
separator outlet
API outlet to
activated sludge inlet
benzene
1.9
16
17.6
39.3
13.5
toluene
0.7
27.4
27.9
29.6
17.5
ethyl
benzen
e
1.9
59.6
60.3
0
71.4
xylen
e
1.2
32.7
33.5
50
6.6
includes contributions estimated from tank farm (not reported). These
concentrations estimated from the tank farm are less than the concentrations
reported for the Ultraformer (greatest contribution to total flow). The
concentrations from the tank farm were selected to produce estimated emissions
from the unit exit to F04 that are comparable to the CMA estimate.
11-39
-------
and the open area was increased to the reported area of the total
collection system (1225 ft2), the estimated loss from the
collection system would equal over 99 percent. The actual
percent loss is expected to be somewhat less than this value due
to potential under reporting the concentrations leaving the
collection system.
11.5.5 Tracer testing at the Coastal Eagle Point Refinery
This tracer test indicated that there were significant losses
of organic compounds from a refinery wastewater collection
system. The collection system had some air emission controls
(water seals at the junction box inlets).
The air emission losses appeared to be significantly less than
from previous testing at a more open collection system at Rohm
and Haas (all of the higher volatility organic compounds seemed
to be lost at the previous test).
The data quality was compromised to some extent by the
variable flows in the tracer feeds. Analysis of the
concentration ratios of different organic compounds fed
simultaneously tended to confirm the calculations based on the
ratios of the volatile tracers to the cobalt tracer.
The predicted losses from the simple collection system air
emission models (Section 4.2.1) were similar to the results of
the tracer testing.
Two non-volatile reference tracers were used, cobalt and
copper. These two tracers did not have identical response in the
collection system, even though the recovery of the compounds was
excellent in the analytical tests. The results of the analysis
of the loss of organic compounds with both metal tracers
indicated significant loss of higher volatility organic compounds
in the collection system. These results are summarized in the
Table 11-21.
11-40
-------
TABLE 11-21. A COMPARISON OF THE OVERALL LOSS OF VOLATILES
BASED UPON TWO DIFFERENT NON-VOLATILE TRACERS.
Compound
111 trichlorobenzene
chlorobenzene
nitrobenzene
Overall fraction
lost with cobalt
ratios
.51
.38
.60
Overall fraction
lost with copper
ratios
.41
.26
.27
If the volatile and the non-volatile had identical
concentration ratios in all of the collection system, there
would be no air emission losses.
If compounds differ in Henry's law, a measure of volatility,
the relative amount of the compounds would be expected to change
if compounds were removed from the collection system by means of
a mass transfer mechanism that depended on volatility. In the
case of chlorobenzene and 111 trichloroethane, the ratio of the
concentration of the less volatile compound to the more volatile
compound increased in each sampling period, with an estimated
range of chlorobenzene loss of 2 to 9 percent from the drain.
These two compounds were injected as tracers together and were
sampled and analyzed together. The use of this concentration
ratio results were much less reproducible for the nitrobenzene,
perhaps because a different analytical method was used for
nitrobenzene and the other organic compounds. Even with the lack
of reproducibility the ratios of chlorobenzene to nitrobenzene
showed a decrease on average, as expected.
This method of estimating the loss from the drains results in
similar results to the predictions of case A3, air flow due to
density (Section 4.3.3), even though the model assumptions did
not match the physical conditions of the refinery collection
system. The results are presented in Table 11-22.
11-41
-------
TABLE 11-22. A COMPARISON OF THE MEASURED LOSS OF VOLATILES
FROM A DRAIN TO THE PREDICTED LOSS OF VOLATILES FROM THE
DRAIN.
Compound
111 trichlorobenzene
chlorobenzene
Fraction lost from
measured
concentration
ratios
.053
.21
Fraction lost
from case A3
model prediction
.032
.126
If two compounds with different volatilities had identical
concentration ratios in the collection system, there would be
no air emission losses.
11.5.5.1 Comparison of an Estimation of volatilization loss to
the measured loss based upon concentrations relative to the metal
tracer. It is assumed that the metal tracer will not volatilize
or be sorbed in the collection system. The tracer test was
carried out over a period of several days and there did not seem
to be evidence of sorption in the system, either for the metals
or the organics. If the concentration of the organic compounds
relative to the metal tracer decreases as the compounds travel
through the collection system, the fraction loss can be
attributed to volatilization. The following tables present the
tracer results and compare the calculated fraction loss due to
volatilization to the predicted values.
11-42
-------
TABLE 11-23. A COMPARISON OF CHLOROBENZENE TRACER RESULTS AND
PREDICTED VALUES.
Location of
Emissions
drain
Junction box
2 junction boxes
2 junction boxes
Test results
.09
.05
.38
Cumulative
Prediction
.032
.12
.29
.42
This table compares the results of the tracer analysis with
cobalt ratios at the Coastal Eagle Point Refinery to the
previously predicted air emissions (chlorobenzene case A3 for
drain, .098 junction box emission factor, Section 4.3.3).
TABLE 11-24. A COMPARISON OF 111 TRICHLOROETHANE TRACER
RESULTS AND PREDICTED VALUES.
Location of
Emissions
drain
Junction box
2 junction boxes
2 junction boxes
Test results
.239
.285
.507
Cumulative
Prediction
.12
.23
.40
.54
This table compares the results of the tracer analysis with
cobalt ratios at the Coastal Eagle Point Refinery to the
previously predicted air emissions (case A3 for drain, .12
junction box emission factor, Section 4.3.3) .
11-43
-------
11.5.6 Ph.D. dissertation of R. L. Corsi
Page 11: In the four city study Levins et al. (1979) observed
the reduction in both occurrence and concentration of
trihalomethanes from tap water to wastewater (Corsi suggested
loss from activities such as showers). This suggests that
organic compounds can be lost before the waste reaches the
collection mains.
Page 11: Another observation is the occurrence of highly
volatile carbon tetrachloride in industrial sewers, and the
absence of this compound in treatment plant influent, suggesting
significant volatile emissions from combined sewer systems. This
suggests that organic compounds can be lost, even in the
collection mains.
Page 17: The U. S. EPA investigated methodologies for
estimating the risk of human exposure to toxic contaminants.
Computational modeling significantly underestimated measured
ambient levels of chloroform and carbon tetrachloride, and it was
suggested that the difference might be due to area sources, e.g.
sewers. Volatilization from sewers was observed to be a major
loss mechanism for dichloropropane and dichloroethane. This
statement also suggests that organic compounds can be lost, even
in the collection mains.
Page 19: The laboratory investigation of Fingas et al. (1988)
demonstrated that the low boiling point fraction (higher
volatility organic compounds) were lost into the headspace during
system entry and may be distributed across the gas flow cross-
section. The high boiling point fraction (less volatile) were
retained in the wastewater and were lost to the headspace during
the flow of the waste. This investigation also suggests that
high volatility organic compounds can be lost before the waste
reaches the collection mains.
11-44
-------
Page 182: Conclusion 4. If organic compounds are discharged
well upstream of wastewater treatment facilities and traverse
building laterals and many smaller reaches with steep channel
slopes prior to reaching an interceptor, cumulative emissions of
organic compounds are likely to be higher than those which occur
at an associated treatment facility.
Page 182: Conclusion 5. Extensive relative removal (greater
than 50%) of organic compounds is likely following potable water
discharge to building laterals leading to street sewers.
11.6 REFERENCES
1. Benzene docket, Sterling Chemical
2. Ibidem, Sterling Chemical.
3. The Behavior of Dispersed and Nondispersed Fuels in a Sewer
System, ASTM STP 1018, 1989, pp.274-289.
4. EMB Report 88-1WW-08, September 1991
11-45
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APPENDIX A
A GUIDE THROUGH THE LITERATURE
-------
APPENDIX A
A GUIDE THROUGH THE LITERATURE
A.I INTRODUCTION
There is concern that volatilization of organic compounds
from hazardous waste treatment, storage, and disposal facilities
(TSDF) poses a public health problem. These organic emissions
may adversely affect ambient air quality. However, there are
other competing mechanisms or pathways through which organic
compounds can leave hazardous waste facilities. These include
adsorption, migration, runoff, biological or chemical
degradation, hydrolysis, oxidation, and hydroxyl radical
reaction. Consequently, the potential hazard of volatilization
should be assessed in relation to the potential importance of
these other pathways.
Sources of potential organic losses include, among others,
surface impoundments, landfarms, landfills, and wastewater
collection and treatment units. The important competing pathways
have been identified in earlier sections of this report. This
evaluation has been based on field and laboratory measurements as
well as predictive or mathematical models of these pathways.
This appendix supplements the body of the report; it serves to
direct the reader through the literature concerning organic
pathways.
For the convenience of the reader, a comprehensive source
list is presented in Appendix B of this report. In addition to
the references cited in this appendix and in the individual
sections of the report, this bibliography lists other literature
of interest for organic pathways and emission models.
A-2
-------
A. 2 SURFACE IMPOUNDMENTS
A.2.1 Volatilization Direct measurement of volatilization
rates from surface impoundments is extremely complicated. Hwang
and Thibodeaux reviewed the concentration profile and plume
mapping technique and proposed a new method requiring fewer
concentration measurements. This latter technique has yet to
2
gain popularity. Thibodeaux et al. used the concentration
profile technique to measure the rate at which selected organic
compounds were emitted to the air from basins in the pulp and
paper industry. The ranges of the average flux for methanol and
2,
3
2 2
acetone were 1.4 to 3.8 ng/cm »s and 0.028 to 0.10 ng/cm »s,
respectively, which were higher than background. Radian^
obtained emission rates from four different hazardous waste sites
containing surface impoundments as well as landfills and
landfarms. They used the concentration profile, transect,
materials balance, and vent sampling approaches. These
4 5
methods ' are also applicable to other nonpoint source hazardous
waste facilities such as landfills and landfarms.
Volatilization rates from surface impoundments can be
estimated via mathematical models. Mackay and Leinonen
predicted air emissions from nonaerated surface impoundments
receiving influent irregularly (unsteady state) . The liquid and
gas mass transfer coefficients in this model were modified by
7 8
Mackay and Yeun. Thibodeaux et al. developed predictive models
for both aerated and nonaerated steady-state surface
9
impoundments. DeWolf and Wetherold critiqued these models and
presented a protocol for their proper use. Shen modified the
nonaerated model of Thibodeaux et al. In an extensive review
12
of these and other predictive models, GCA judged the
13 14
theoretical work of Thibodeaux et al. and Mackay and Leinonen
as most suitable for predicting air emissions from surface
impoundments.
The use of these mathematical models for predicting volatile
emissions is less expensive and faster than actual field
A-3
-------
measurements. However, to be cost-effective, these mathematical
models must provide accurate estimates of volatilization rates.
It is disappointing to note that relatively few validation
studies are reported in the literature. A description of these
follows.
15
Hwang compared predicted and measured emission rates of
toluene and 1,1,1-trichloroethane from an evaporation pond. The
transect method was used for field measurements, and the models
-i /~
summarized in Hwang provided the predicted rates. For each
organic compound, the predicted result was within the confidence
17
limits of the average measured result. Balfour et al. used the
1 8
Thibodeaux et al. model to predict emissions from these surface
impoundments. Emission rates were measured via the flux chamber
in all three ponds as well as via the concentration profile in
the third pond. In this latter pond, the emission rate of most
compounds as determined using the flux chamber was statistically
greater than that determined using the concentration profile.
Furthermore, results of a comparison of measured versus predicted
19
emission rates were inconclusive. Vaught used the Springer et
al. and Mackay and Yeun approaches to predict air emissions
from quiescent surface impoundments. He concluded that volatil-
ization rates predicted from the Mackay and Yeun model were
comparable to rates measured via the flux chamber. In contrast,
the predicted rates from Springer et al. exceeded the measured
rates.
A.2.2 Other Pathways The role of other pathways in the
removal of organic compounds from surface impoundments has not
been addressed extensively in the literature. However,
biological removal mechanisms associated with stabilization ponds
22
and lagoons will be applicable where conditions of pH,
temperature, and nutrient levels are suitable for biological
growth.
A.3 LAND TREATMENT
A-4
-------
For the past 25 years, the petroleum industry has operated land
treatment, sludge farming, and land disposal facilities. The
pharmaceutical and organic chemical manufacturers recently have
considered this method of hazardous waste disposal because of its
comparatively reasonable cost, simplicity, and use of natural
processes. How does a land treatment effectively and safely
treat and dispose of organic compounds? The purpose of land
treatment is to exploit the microbiological actions of the upper
soil layer to degrade toxic organic material at a controlled
rate. Although photo-degradation takes place in a land treatment
23
facility, the short time that the materials are exposed to
direct sunlight and the screening effect of the oil in which
hazardous materials are suspended make this pathway negligible.
Several other pathways may exist: volatilization, runoff, and
24 25
leaching. ' However, these latter mechanisms do not occur at
a properly sited, operated, and maintained RCRA-permitted land
treatment facility.
A.3.1 Degradation The chemical structure of the hazardous
waste, application and mixing techniques, and soil
characteristics (texture, temperature, moisture content, oxygen
level, nutrient level, pH, and the kind and number of microbes)
O £. O "~7
affect biodegradation. ' Although biodegradation is purported
to be the principal mechanism for removal of organic compounds by
land treatment, only a few experiments have attempted to quantify
the resulting removal. A laboratory simulation of land treatment
of oily sludge revealed that 85 percent of the polynuclear
O Q O Q
aromatics degraded. Results from Snyder et al. are
comparable: oil removal on fertilized plots approached
80 percent at 1 year postapplication.
Mathematical models for degradation could not be found in
the literature.
A.3.2 Volatilization Techniques for direct measurement of
volatilization at landfarms ' were discussed previously.
A-5
-------
Exogenous factors affecting volatilization in land treatment
include properties of the soil, waste application techniques,
32 33 34
mixing schedules, and atmospheric conditions. ' '
35
Farmer and Letey proposed five gradient models for
pesticide volatilization rates from the soil based on diffusion
laws. The models accommodate soil-incorporated pesticides with
and without significant mobility in flowing water. With
nonincorporated pesticides, vapor density relationships and air
flow rate rather than movement in the soil control the
volatilization rate. These approaches do not, however,
Q £.
accommodate subsurface injection. Thibodeaux and Hwang
developed a gradientless model of air emissions from petroluem
waste landfarms. Their approach accurately predicted the
volatilization of dieldrin reported in Farmer and Letey and is
considered most suitable for estimating air emissions from land
treatment.
A.3.3 Migration and Runoff Migration and runoff of organic
compounds from a landfarm may occur after improper application or
treatment of the hazardous waste. A description of factors
affecting these two pathways appeared in Reference 38. Results
39
from a laboratory study of refinery and petrochemical sludge
suggested that the presence of hazardous waste in runoff
decreases with time after application. In addition, leachate
water collected 1.5 meters below the subsurface was essentially
free of toxic components.
However, as previously mentioned, migration and runoff do
not occur at a properly sited, operated, and maintained RCRA-
permitted land treatment facility. This paragraph is included
for the sake of completeness.
A.4 LANDFILLS
A.4.1 Volatilization Direct measurements of organic compound
emissions from landfills are possible. During field tests
conducted for EPA's Office of Air Quality Planning and Standards
40
(OAQPS), Radian measured air emissions from landfills at three
A-6
-------
41
of the four monitored hazardous waste TSDF. Markle et al.
collected air samples from three landfills representative of
those used by the polyvinyl chloride industry for health hazard
evaluations. To compare the efficiencies of water and soil
42
coverings in reducing volatilization, Farmer et al. measured
emission rates from simulated landfills.
Numerous equations also have been developed to model organic
compound emissions from hazardous waste landfills. The procedure
43
of Farmer et al. based on Fick's law for steady-state
diffusion, estimates emission from covered or buried landfills.
44 45
This was later modified by Shen. Thibodeaux's emission
models differentiate covered landfills by the presence or absence
4 6
of internal gas generation. Another approach incorporates
time-varying atmospheric pressure into the emission model.
Volatilization rates from landfills with no covering, i.e., open
47 48
dumps, were modeled by Shen. DeWolf and Wetherold recommend
49
Shen's emission model for covered landfills. GCA, in their
excellent comprehensive review of these and other emission
models, prefers the work of Farmer et al. and Thibodeaux.
Field validation of these mathematical models has not been
52
reported in the literature. Despite this, Baker and Mackay
53
employed Shen's model in their protocol to evaluate toxic air
pollution downwind of hazardous waste landfills.
A.4.2 Migration Several scientists have investigated the
potential problem of migration of toxic contaminants from
54
landfills. Rovers and Farquhar suggested that the production
of leachate within a landfill is not uncommon. However, the
migration of harmful compounds through adjacent soils is not
significant. Shen and Tofflemire reported that annual losses
of PCB to migration from uncovered landfills in the Hudson River
Basin (New York) were substantially less than losses to
volatilization.
A.4.3 Other Pathways The impact of other pathways is not
discussed quantitatively in the literature.
A-7
-------
A.5 WASTEWATER TREATMENT PLANT EFFLUENT
A description of the pathways operating in a wastewater
treatment plant is complicated by the number of different
treatment systems. There are closed tanks and open tanks (with
and without mixing). Air emissions from closed tanks occur
during venting.
E. C. Jordan and Burns and Roe ' examined the fate of
priority toxic pollutants in publicly owned treatment plants.
They observed a decrease in organic compound concentrations
across the activated sludge process and a lack of pollutant
accumulation in the waste-activated sludge. This suggests that
organic compounds are substantially air-stripped or biodegraded
during secondary treatment. Results from the controlled
59
laboratory experiments of Roberts et al. imply that organic
solutes more likely volatilize during wastewater treatment with
surface aeration than with bubble aeration. Lurker et al.
examined how aeration rate, suspended particle concentration, and
detergent concentration influence aerial organic chemical release
from an activated sewage treatment process.
The nonaerated open tank system is similar to the nonaerated
surface impoundment discussed previously; see Section A.2.1 for a
discussion of the corresponding emission rate models. Similarly,
open tank wastewater treatment processes with mixing can be
C" "1 C" O
estimated from Thibodeaux et al. Hwang went a step further
in his activated sludge surface aeration model. His approach was
to estimate pollutant removal by degradation, adsorption, and air
stripping via a mass balance equation. Like Hwang and Thibodeaux
et al., Freeman considered air stripping losses at the air-
water interface. Unlike Hwang, however, he viewed the adsorption
pathway as insignificant and, thus, ignored it. In an entirely
different approach, Freeman ' modeled the mass transfer of a
toxic compound into the bubbles of the aerated system (diffused
air [subsurface] activated sludge model). The structure of these
and other models was critiqued in GCA.
A-8
-------
Allen et al.67 presented models of organic compound losses
at each process encountered in wastewater treatment systems. The
models include a methodology for estimating the relative
importance of competing pathways. Additionally, these
investigators compared the loss of volatiles obtained from field
tests at several treatment facilities and from these
mathematical models. The models predict organic compound losses
due to biodegradation or volatilization in close agreement with
the field data. Results from other validation studies are not as
consistent. Predicted emission rates from aerated surface
impoundments at two wastewater treatment plants were generally
69
larger than measured values. The difference between measured
and predicted emission rates in Cox et al. appears to be a
function of the type of compound and the presence of aerators.
A-9
-------
A.6 REFERENCES
1. Hwang, S. T., and L. J. Thibodeaux. Measuring Volatile
Chemical Emission Rates from Large Waste Disposal
Facilities. Environmental Progress. 2_: 81-86. 1983.
2. Thibodeaux, L. J., D. G. Parker, and H. H. Heck (University
of Arkansas). Measurement of Volatile Chemical Emissions
from Wastewater Basins. Prepared for U.S. Environmental
Protection Agency. Washington, DC. Contract No. R805534.
December 1981.
3. Radian Corporation. Hazardous Waste Treatment, Storage, and
Disposal Facility Area Sources: VOC Air Emissions.
Prepared for U.S. Environmental Protection Agency. EPA
Contract No. 68-02-3850. January 1985.
4. Reference 1.
5. Reference 3.
6. Mackay, D., and P. J. Leinonen. Rate of Evaporation of Low-
Solubility Contaminants from Water Bodies to Atmosphere.
Environmental Science and Technology. 1^3:1178-1180. 1975.
7. Mackay, D., and A. T. K. Yeun. Mass Transfer Coefficient
Correlations for Volatilization of Organic Solutes from
Water. Environmental Science and Technology. 17 : 211-217.
1983.
8. Reference 2.
9. DeWolf, G. B., and R. G. Wetherold (Radian Corporation).
Protocols for Calculating VOC Emissions from Surface
Impoundments Using Emission Models, Technical Note.
Prepared for U.S. Environmental Protection Agency.
Washington, DC. Contract No. 68-02-3850. September 1984.
10. Shen, T. T. Estimation of Organic Compound Emissions from
Waste Lagoons. Journal of the Air Pollution Control
Association. 32. 1982.
11. Reference 2.
12. GCA Corp. Evaluation and Selection of Models for Estimating
Air Emissions from Hazardous Waste Treatment, Storage, and
Disposal Facilities. Prepared for U.S. Environmental
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December 1984.
A-10
-------
13. Reference 2.
14. Reference 6.
15. Hwang, S. T. Model Prediction of Volatile Emissions.
Environmental Progress. 4_:141-144. 1985.
16. Hwang, S. T. Toxic Emissions From Land Disposal Facilities.
Environmental Progress. 1_:46-52. 1982.
17. Balfour, W. D., C. E. Schmidt, R. G. Wetherold, D. L. Lewis,
J. I. Steinmetz, and R. C. Hanish. Field Verification of
Air Emission Models for Hazardous Waste Disposal Facilities.
In: Proceedings of the Tenth Annual Research Symposium.
Publication No. EPA-600/9-84/007. Fort Mitchell, Kentucky.
April 1984.
18. Reference 2.
19. Vaught, C. C. (GCA). Air Emissions for Quiescent Surface
Impoundments--Emissions Data and Model Review, Draft
Technical Note. Prepared for U.S. Environmental Protection
Agency. Washington, DC. Contract No. 68-01-6871. August
1985.
20. Springer, S., P. D. Lunney, K. T. Valsaraj, and L. J.
Thibodeaux (University of Arkansas and Louisiana State
University.) Emissions of Hazardous Chemicals from Surface
and Near Surface Impoundments to Air, Draft Final Report.
Prepared for U.S. Environmental Protection Agency.
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21. Reference 7.
22. Metcalf and Eddy, Inc. Wastewater Engineering: Collection,
Treatment, Disposal. New York, McGraw-Hill. 1972.
23. Kaufman, D. D. Fate of Organic Compounds in Land-Applied
Wastes. In: Land Treatment of Hazardous Wastes, Parr, J.
F., P. B. Marsh, and J. M. Kla (eds.). 1983. p. 77-151.
24. U.S. Environmental Protection Agency. Hazardous Waste Land
Treatment, Technical Resource Document. EPA Contract Nos.
68-03-2940 and 68-03-2943. April 1983.
25. Brown, K. W. Chapter 36, Land Treatment of Hazardous
Wastes. In: Proceedings of the Fourth Life Sciences
Symposium, Environment and Solid Wastes. Gatlinburg, TN.
October 4-8, 1981. p. 449-482.
26. Reference 24.
A-11
-------
27. Reference 23.
28. Bossert, I., W. M. Rachel, and R. Earth. Fate of
Hydrocarbons During Oily Sludge Disposal in Soil. Applied
and Environmental Microbiology. 47:763-767 . 1984.
29. Snyder, H. J., G. B. Rice, and J. J. Skujins. Residual
Management by Land Disposal. In: Proceedings of the
Hazardous Waste Research Symposium, Fuller, W. H., (ed.)
Publication No. EPA-600/9-76-015. July 1976.
30. Reference 1.
31. Reference 3.
32. Wetherold, R. G., J. L. Randall, and K. R. Williams (Radian
Corporation). Laboratory Assessment of Potential
Hydrocarbon Emissions from Land Treatment of Refinery Oily
Sludges. Prepared for U.S. Environmental Protection Agency.
Washington, DC. Publication No. EPA-600/2-84-108. June
1984.
33. Reference 24.
34. Farmer, W. J., and J. Letey. Volatilization Losses of
Pesticides from Soil. Prepared for U.S. Environmental
Protection Agency. Publication No. EPA-660/2-74-054.
August 1974.
35. Reference 34.
36. Thibodeaux, L. J., and S. T. Hwang. Land Farming of
Petroleum Wastes--Modeling the Air Emissions Problem.
Environmental Progress. 1^:42-46. 1982.
37. Reference 34.
38. Reference 23.
39. Reference 25.
40. Reference 3.
41. Markle, R. A., R. B. Iden, and F. A. Sliemers (Battelle). A
Preliminary Examination of Vinyl Chloride Emissions from
Polymerization Sludges during Handling and Land Disposal.
Prepared for U.S. Environmental Protection Agency.
Washington, DC. Publication No. EPA-660/2-74-054.
February 1976.
42. Farmer, W. J., M. Yang, J. Letey, and W. F. Spencer.
A-12
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Problems Associated with the Land Disposal of an Organic
Industrial Hazardous Waste Containing HCB. In: Residual
Management by Land Disposal, Proceedings of the Hazardous
Waste Research Symposium, Fuller, W. H. (ed.). Publication
No. EPA-600/9-76-015. July 1976.
43. Farmer, W. J., M. S. Yang, and J. Letey. Land Disposal of
Hazardous Wastes: Controlling Vapor Movement in Soils. In:
Fourth Annual Research Symposium. Publication No. EPA-
600/9-78-016. August 1978.
44. Shen, T. T. Estimating Hazardous Air Emissions from
Disposal Sites. Pollution Engineering. 31-34. August
1981.
45. Thibodeaux, L. J. Estimating the Air Emissions of Chemicals
from Hazardous Waste Landfills. Journal of Hazardous
Materials. _4:235-244. 1981.
46. Thibodeaux, L. J., C. Springer, and L. M. Riley. Models of
Mechanisms for the Vapor Phase Emission of Hazardous
Chemicals from Landfills. Journal of Hazardous Materials.
^:63-74. 1982.
47. Reference 44.
48. DeWolf, G. B., and R. G. Wetherold (Radian Corporation).
Protocols for Calculating VOC Emissions from Land
Applications Using Emission Models, Technical Note.
Prepared for U.S. Environmental Protection Agency.
Washington, DC. Contract No. 68-02-3850. December 1984.
49. Reference 44.
50. Reference 43.
51. Reference 2.
52. Baker, L. W., and K. P. Mackay. Hazardous Waste Management:
Screening Models for Estimating Toxic Air Pollution Near a
Hazardous Waste Landfill. Journal of the Air Pollution
Control Association. 3J5: 1190-1195. 1985.
53. Reference 44.
54. Rovers, F. A., and G. J. Farquhar. Evaluating Contaminant
Attenuation in the Soil to Improve Landfill Selection and
Design. In: Proceedings of the International Conference on
Land for Waste Management. 1974. p. 161-173.
55. Shen, T. T., and T. J. Tofflemire. Air Pollution Aspects of
A-13
-------
Land Disposal of Toxic Wastes. Environmental Engineering
Division Journal. 106:211-226. 1980.
56. E. C. Jordan Co. Fate of Priority Toxic Pollutants in
Publicly Owned Treatment Works, 30 Day Study. Prepared for
U.S. Environmental Protection Agency. Publication No. EPA-
440/1-82-302. August 1982.
57. Burns and Roe Industrial Services Corp. Fate of Priority
Pollutants in Publicly Owned Treatment Works, Final Report,
Volume II. Prepared for U.S. Environmental Protection
Agency. Publication No. EPA-440/1-82-303. July 1982.
58. Burns and Roe Industrial Services Corp. Fate of Priority
Pollutants in Publicly Owned Treatment Works, Final Report,
Volume I. Prepared for U.S. Environmental Protection
Agency. Publication No. EPA-440/1-82-303. September 1982.
59. Roberts, P. V., C. Munz, P. Dandliker, and C. Matter-Muller.
Volatilization of Organic Pollutants in Wastewater
Treatment-Model Studies, Project Summary. Prepared for U.S.
Environmental Protection Agency. Publication No. EPA-
600/52-84-047. April 1984.
60. Lurker, P. A., C. S. Clark, V. J. Elia, P. S. Gartside, and
R. N. Kinman. Aerial Organic Chemical Release from
Activated Sludge. Water Research. 1_8_: 489-494. 1984.
61. Reference 2.
62. Hwang, S. T. Treatability and Pathways of Priority
Pollutants in Biological Wastewater Treatment. (Presented
at the American Institute of Chemical Engineers Symposium.
Chicago, Illinois. November 1980.)
63. Freeman, R. A. Stripping of Hazardous Chemicals from
Surface Aerated Waste Treatment Basins. In: APCA/WPCF
Speciality Conference on Control of Specific Toxic
Pollutants. Gainesville, Florida. February 13-16, 1979.
64. Freeman, R. A. Comparison of Secondary Emissions From
Aerated Treatment Systems. (Presented at AlChe Meeting.
Paper 5c. Orlando, Florida. February 1982.)
65. Freeman, R. A. Secondary Emissions from Subsurface Aerated
Treatment Systems. Environmental Progress. 1_: 117-119.
1982.
66. Reference 12.
67. Allen, C. C., S. Simpson, and G. Brant (RTI and Associated
A-14
-------
Technologies, Inc.). Field Evaluations of Hazardous Waste
Pretreatment as an Air Pollution Control Technique.
Prepared for U.S. Environmental Protection Agency. EPA
Contract No. 68-02-3992. April 1985.
68. Alsop, G. M., R. L. Berglund, T. W. Siegrist, G. M. Whipple,
and B. E. Wilkes. Fate of Specific Organics in an
Industrial Biological Wastewater Treatment Plant, Draft
Report. Prepared for U.S. Environmental Protection Agency,
Industrial Environmental Research Laboratory. Research
Triangle Park, NC. June 29, 1984.
69. Meisenheimer, D. C. (GCA). Emissions Data and Model Review
for Wastewater Treatment Operations, Draft Technical Note.
Prepared for U.S. Environmental Protection Agency.
Washington, DC. Contract No. 68-01-6871. August 1985.
70. Cox, R. D., Lewis, R. G. Wetherold, and J. I. Steinmetz
(Radian Corporation). Evaluation of VOC Emissions from
Wastewater Systems (Secondary Emissions). Prepared for U.S.
Environmental Protection Agency. Washington, DC. Project
No. 68-03-3038. July 1983.
A-15
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APPENDIX B
COMPREHENSIVE SOURCE LIST
-------
APPENDIX B
COMPREHENSIVE SOURCE LIST
Addendum to memorandum dated September 6, 1985, from
Eichinger, Jeanne, GCA Corporation, to Hustvedt, K. C.,
EPA/OAQPS. September 12, 1985. TSDF model source
parameters and operating practices data base.
Allen, C. C. Prediction of Air Emissions from Surface
Impoundments, Paper 31a. (Presented at 1986 Summer
Meeting of AIChE. Boston, MA. August 1986.) 26 p.
Allen, C. C., and D. A. Green (Research Triangle
Institute). Review of VOC Pathway Models, Draft Report.
Prepared for U.S. Environmental Protection Agency. EPA
Contract No. 68-01-6826. 1985.
Allen, C. C., D. A. Green, and J. B. White (Research
Triangle Insitute). Preliminary Assessment of Aerated
Waste Treatment Systems at TSDFs - Phase I, Draft Final
Report. Prepared for U.S. Environmental Protection
Agency. EPA Contract No. 68-03-3149. May 1985.
Allen, C. C., D. A. Green, and J. B. White (Research
Triangle Institute). Preliminary Assessment of Aerated
Waste Treatment Systems at TSDFs—Phase I. Draft.
Prepared for U.S. Environmental Protection Agency.
Research Triangle Park, NC. EPA Contract No. 68-03-3149,
Task 54-01F. 1985.
Allen, C. C., S. Simpson, and G. Brant (Research Triangle
Institute and Associated Technologies, Inc.). Field
Evaluations of Hazardous Waste Pretreatment as an Air
Pollution Control Technique. Prepared for U.S.
Environmental Protection Agency. EPA Contract No. 68-02-
3992. April 1985.
Alsop, G. M., R. L. Berglund, T. W. Siegrist, G. M.
Whipple, and B. E. Wilkes. Fate of Specific Organics in
an Industrial Biological Wastewater Treatment Plant,
Draft Report. Prepared for U.S. Environmental Protection
Agency, Industrial Environmental Research Laboratory.
June 29, 1984.
-------
Armstrong, N. E., E. F. Gloyna, and 0. Wyss. Biological
Countermeasures for the Control of Hazardous Material
Spills, Project Summary. Publication No. EPA-600/S2-84-
071. March 1984.
Arnold, J. H. Studies in Diffusion: III. Unsteady-
State Vaporization and Absorption. Transactions of the
American Institute of Chemical Engineers. 40:361-379.
1944.
Bailey, J. E., and D. F. Ollis. Biochemical Engineering
Fundamentals. New York, McGraw-Hill. 1977. p. 343-349,
Baker, L. W., and K. P. Mackay. Hazardous Waste
Management: Screening Models for Estimating Toxic Air
Pollution Near a Hazardous Waste Landfill. Journal of
the Air Pollution Control Association. 35: 1190-1195.
1985.
Balfour, W. D., C. E. Schmidt, R. G. Wetherold, D. L.
Lewis, J. I. Steinmetz, and R. C. Hanish. Field
Verification of Air Emission Models for Hazardous Waste
Disposal Facilities. In: Proceedings of the Tenth
Annual Research Symposium. Publication No. EPA-600/9-
84/007. Fort Mitchell, Kentucky. April 1984.
Bossert, I., W. M. Rachel, and R. Earth. Fate of
Hydrocarbons during Oily Sludge Disposal in Soil.
Applied and Environmental Microbiology. 47: 763-767.
1984.
Branscome, M., and A. Gitelman. Sensitivity Analysis:
Emission Estimates for Surface Impoundments. Prepared
for the U.S. Environmental Protection Agency. (Docket
Number F-91-CESP-FFFFF). March 1986. 67 p.
Brown, K. W. Chapter 36, Land Treatment of Hazardous
Wastes. In: Proceedings of the Fourth Life Sciences
Symposium, Environment and Solid Wastes. Gatlinburg, TN.
October 4-8, 1981. p. 449-482.
Burns and Roe Industrial Services Corp. Fate of Priority
Pollutants in Publicly Owned Treatment Works, Final
Report, Volume II. Prepared for U.S. Environmental
Protection Agency. Publication No. EPA-440/1-82-303.
July 1982.
Burns and Roe Industrial Services Corp. Fate of Priority
Pollutants in Publicly Owned Treatment Works, Final
Report, Volume I. Prepared for U.S. Environmental
-------
Protection Agency.
September 1982.
Publication No. EPA-440/1-82-303 .
Canter, L., and R. C. Knox . Evaluation of Septic Tank
System Effects on Ground Water Quality. Publication No.
EPA-600/S2-84-107. September 1984.
Clark, J. W., W. Viessman, Jr., and M. J. Hammer.
Supply and Pollution Control. Scranton, PA,
International Textbook Company. 1971.
Water
Cox, R. D., D. L. Lewis, R. G. Wetherold, and J. I.
Steinmetz (Radian Corporation) . Evaluation of VOC
Emissions from Wastewater Systems (Secondary Emissions) .
Prepared for U.S. Environmental Protection Agency.
Washington, DC. EPA Project No. 68-03-3038. July 1983.
Cox, R. D., J. L. Steinmetz, and D. L. Lewis (Radian
Corporation) . Evaluation of VOC Emissions from
Wastewater Systems (Secondary Emission), Volume II,
Appendices. Prepared for U.S. Environmental Protection
Agency. Washington, DC. EPA Project No. 68-03-3038.
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-------
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B-14
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APPENDIX C
PROPERTIES FOR COMPOUNDS OF INTEREST
note: these compound properties may be updated. Please
check the EPA distribution site for current property sets.
-------
APPENDIX C
PROPERTIES FOR COMPOUNDS OF INTEREST
This appendix contains compound-specific properties of
about 500 chemicals, most of which are not included in
CHEMDAT8. These data, presented as a source of
information, can be easily incorporated into CHEMDAT8.
These compounds are stored in a file named "datatwo.wkl".
These compound properties greatly increase the utility of
CHEMDAT8. These chemicals represent those chemicals that
could be encountered in industrial facilities and that are
useful for calculating emission rates for the facilities
modeled in the body of this report. The list was extracted
from the GCA Physical/Chemical Database, WET model stream
compositions, and the Industrial Studies Database.
Specific chemical properties were updated as additional
information became available, especially the Henry's law
constants for hazardous volatile chemicals. Missing
compound properties were estimated.
The compounds listed in this appendix were not all
included in CHEMDAT8 because their inclusion would
seriously slow the execution time of the program, and the
memory requirements would prevent the program from being
run on many machines. Compounds included in the CHEMDAT8
program were selected on the basis of the estimated
frequency with which they are found in hazardous wastes and
on their position in prioritized lists of pollutants.
It is recognized that biodegradation rates can vary
widely from site to site. Therefore, the following
priority schedule is provided as guidance in determining
the appropriate biodegradation rate constants to be
employed in the emission models:
• Use site-specific biodegradation rate data where
c-2
-------
available.
• Use the rate constants suggested in the following
table.
• Estimate the biodegradation rate constants using
the following methodology:
Approximate K from available data for
K for compounds of similar structure
max- jr 4. • n i
and/or functional groups; and
Approximate KI either by using the following
correlation:
K! (L/h/g) = 0.135 Kow 0.38
where Kow = octanol-water partitioning
coefficient
or by using the default (average) value for
KI, which is KI =1 L/h/g, and then
calculate KS as: KS = K^/f^.
The following properties are given for each chemical
listed by name in Table C-l:
• Compound type code
• Molecular weight
• Density
• Vapor pressure at 25 °C
• Solubility
• Henry's law constant
• Diffusion coefficient in water
• Diffusion coefficient in air
• Boiling point
The following properties are given for each chemical
listed by name in Table C-2:
• Coefficients for the Antoine equation for
estimating vapor pressure at temperatures other
than 25 °C.
c-3
-------
• LN(OW)
• Kmax biorate (mg/g-hr)
Kl biorate (L/g-hr)
• rate of hydrolysis (per second)
To estimate vapor pressures at temperatures other than
25 °C, the Antoine equation coefficients are used with the
following equation:
ID
log Vapor Pressure (mm Hg) = A - nci\
T + C ( '
where
A, B, and C = the Antoine equation coefficients
T = temperature in °C.
Two approaches may be used to introduce a new compound
and its properties into CHEMDAT8. First, the data for one
compound in CHEMDAT8 may be replaced with data for the
compound of interest in the columns specified above. With
this approach, the number of compounds in CHEMDAT8 remains
constant. The second approach involves appending the new
compound and its properties to the existing list of chemi-
cals in CHEMDAT8. All the equations/calculations must then
be copied from one of the existing rows via Lotus 1,2,3
into the appropriate cells in the new row of the
spreadsheet. As mentioned above, the inclusion in CHEMDAT8
of all or a large part of the chemicals listed in this
appendix could result in increasing the time required to
exercise CHEMDAT8 and could prevent its use on some
microcomputers.
The properties of interest listed above mimic those in
columns B-Q of the CHEMDAT8 spreadsheet.
c-4
-------
TABLE C-1. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS, I
COMPOUND NAME
2,4 D
2,4,5 T
50% PEG
ACENAPHTHENE
ACENAPHTHYLENE
AC ETA L
ACETALDEHYDE
ACETALDOL
ACETAMIDE
ACETIC ACID
ACETONE
ACETONITRILE
ACETOPHENONE
ACETYL CHLORIDE
ACETYLAMINOFLUORENE.2-
ACETYLMETHYLPHTHALATE 4
ACETYL-2-THIOUREA.1-
ACIFLUORFEN
ACROLEIN
ACRYLAMIDE
ACRYLIC ACID
ACRYLONITRILE
ADENINE
ADIPONITRILE
ALDICARB
ALDRIN
ALLYL ALCOHOL
ALLYL CHLORIDE
ALLYL ETHER, diallyl ether
ALPHA METHYL STYRENE
AMETRYN
AMINOBIPHENYL.4-
AMINOPHENOL(-o)
AMINOPHENOL(-p)
AMINOPYRIDINE.4-
AMYL ACETATE(-n)
ANILINE
ANISIDINE.o-
ANTHRACENE
ANTHRAQUINONE
AZIRIDINEethylene imine
BENZAL CHLORIDE
BENZALDEHYDE
BENZENE
BENZIDINE
BENZOFURAN 2,3
BENZOICACID
C
C
H
A
A
H
O
H
N
O
O
N
O
C
A
A
N
A
O
N
O
N
N
H
P
P
O
C
H
A
H
N
N
N
N
O
N
A
A
O
H
C
O
A
A
A
O
MWT
221
255.48
115
154.21
152.21
118.17
45.06
81.11
73.1
60.05
58
41.03
120.16
78.5
223.27
252.22
118.15
361
56.1
71.09
72.07
53.1
135.13
108.14
190.29
364.93
58.1
76.53
98.14
118.18
227.35
169.23
109.12
109.12
94.12
130.18
93.1
123.15
178.23
210.24
43.1
127
106.13
78.1
184.23
118.14
122.13
DENS.
g/cc
1.41
1.41
1
1.07
1.02
0.8254
0.788
1
1.15
1.05
0.79
0.78
1.03
1.11
1.02
1.02
1.15
1
0.84
1.12
0.97
0.81
1.15
0.951
1.18
1.18
0.85
0.94
0.805
1.02
1
1.15
1.15
1.15
1.15
0.88
1.02
1.096
1.25
1.43
1
1.26
0.97
0.87
1.02
1.072
1.27
VP@25C
mmHg
0.0828648
5.250e-09
0
0.005
0.022952
41 .76573
870
0.1450557
86.06205
15.4
266
90
0.297
288
0.0012
0
0
0.0022605
244
0.012
5.2
114
0.1191207
0.0018275
0.0001
0.000006
23.3
368
58.94663
0.076
0.0758461
0.0004959
0.511
0.893
0.002
5.42
1
0.0551848
0.0000013
3.000e-08
160
0.07
1
95.2
0.0000004
2.009953
0.00704
HL@25C
atm-m3/mol
0.0000227
6.348e-12
0.0003
0.00771
0.000114
0.000353
0.0000877
3.350e-08
1 .800e-09
0.0000056
0.000025
0.0000199
0.0000092
0.000199
0.0000013
0.011
0.1933
0.00151
0.0000822
2.610e-10
0.0000002
0.000098
8.850e-1 1
1.310e-08
3.168e-08
0.000496
0.000018
0.00927
0.0022624
0.00591
1 .880e-1 1
0.0000003
0.0000037
0.0000197
0.1933
0.000464
0.0000018
0.0000017
0.0675
3.200e-09
0.000454
0.00741
0.0000423
0.0055502
1 .360e-1 1
0.000237
1 .820e-08
Dl
cm2/s
0.0000073
0.0000067
0.0000088
0.0000077
0.0000075
0.0000077
0.0000141
0.0000108
0.0000125
0.000012
0.0000114
0.0000166
0.0000087
0.0000115
0.000006
0.0000056
0.0000094
0.0000044
0.0000122
0.0000106
0.0000106
0.0000134
0.0000087
0.0000089
0.0000072
0.0000049
0.0000114
0.0000108
0.0000085
0.0000114
0.0000059
0.0000076
0.0000086
0.0000024
0.0000108
0.0000012
0.0000083
0.0000089
0.0000077
0.0000076
0.0000158
0.0000095
0.0000091
0.0000098
0.000015
0.000009
0.000008
Dv
cm2/s
0.0231
0.0192
0.0645
0.0421
0.04386
0.0677
0.124
0.10487
0.1148
0.113
0.124
0.128
0.06
0.099
0.02654
0.0226
0.0585
0.0145
0.105
0.097
0.098
0.122
0.0487
0.0718
0.0305
0.0132
0.114
0.11653
0.0882
0.264
0.0261
0.0361
0.0774
0.0774
0.0802
0.064
0.07
0.0565
0.0324
0.0245
0.2646
0.051
0.073
0.088
0.034
0.0603
0.0536
Bpt
deg. C
240.1
276
0
278
265
102.7
21
185
80.1
118
56.2
81.6
202.3
52
300
0
0
306.0001
53
125
141.6
77.4
220
295
287
398
97
45
94
166
228.4
302
174
164
273
148.4
184
225
0
380
56
207
178.9
80.1
400
174
122.12
C-5
-------
TABLE C-1. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS, I
COMPOUND NAME
BENZONITRILE
BENZOPHENONE
BENZOTRICHLORIDE
BENZOYL CHLORIDE
BENZO(A)ANTHRACENE
BENZO(A)PYRENE
BENZO(B)FLUORANTHENE
BENZO(k)FLUORANTHENE
BENZYL ALCOHOL
BENZYL CHLORIDE
BIPHENYL
BISPHENOL(A)
BIS(1 ,1 ,2,2-TETRACHLOROPROPYL)
BIS(2-CHLOROETHYL)ETHER
BIS(2-ETHYLHEXYL)PHTHALATE
BIS(CHLOROMETHYL)ETHER
BROMACIL
BROMOBENZENE
BROMOCHLOROMETHANE
BROMODICHLOROMETHANE
BROMOFORM
BROMOMETHANE
BROMOTOLUENE4
BROMOXYNIL
BUTADIENE-(1,3)
BUTANE
BUTANEDINITRILE
BUTANOL ISO
BUTANOL(S)
BUTANOL-1
BUTENE
BUTYL ACETATE(-n)
BUTYL ACRYLATE
BUTYL BENZENE
BUTYL CARBITOL
BUTYL CELLOSOLVE
BUTYLAMINE
BUTYLENE GLYCOL-(1 ,3)
BUTYLISOBUTYRATE
BUTYRALDEHYDE
BUTYRALDEHYDE ISO
BUTYRIC ACID
c10 linear
CAPROLACTAM
N
O
C
C
A
A
A
A
O
C
B
O
MWT
103.07
182.23
195.47
113.57
228.3
252.3
252.32
252.32
108.15
126.6
154.2
228.28
DENS.
g/cc
1.15
0.97
1.38
1.41
1.11
1.11
1.02
1.02
0.97
1.1
1.18
0.97
VP@25C
mmHg
1
0.0001763
0.2
214
0.0000002
0.000568
1562.767
9.590e-1 1
0.15
1.21
1
0.0759058
HL@25C
atm-m3/mol
0.0000136
0.0091 1
0.000981
0.000188
1 .380e-09
1 .380e-09
0.0000201
0.011
0.0000006
0.000319
0.000408
0.00228
Dl
cm2/s
0.0000102
0.0000066
0.0000078
0.0000109
0.000009
0.000009
0.0000056
0.0000056
0.000009
0.0000078
0.0000082
0.0000057
Dv
cm2/s deg
0.0706
0.0353
0.0275
0.0567
0.051
0.043
0.0226
0.0226
0.0712
0.075
0.0404
0.0264
Bpt
. C
0
0
221
0
435
312
4
480
0
179.4
254
232
ETHER
C
C
H
C
H
C
C
C
C
C
C
A
H
H
N
O
O
O
H
O
O
A
H
H
N
O
H
O
O
O
H
N
377.7
143
390.56
115
261.11
157.02
129.39
163.8
252.77
94.95
171
276.92
54.09
58.12
80.09
74
74.14
74.1
56.1
116
128.2
134.22
162.23
118.2
73.14
90.14
102.13
72.11
72.12
88.1
168
113.16
1.41
1.22
1
1.32
1
1 .4952
1.41
1.97
2.89
1.41
1 .3959
1
0.76
0.76
1.15
0.79
0.97
0.81
0.6255
0.88
0.8986
1.02
0.96
0.9
0.7327
1.004
0.891
0.97
0.794
0.97
0.75
1.02
875
1.4
0.0000002
30
0.0877172
4.1344
146.3781
59.2
5.6
1590
1 .0868
0.0017175
2100
1522.124
6
10
10
6.5
2218.377
15
5.8
1
0.00468
1.61
72
0.06
68.94406
104.2284
170
0.84
2.009953
0.0072722
43.5
0.000013
0.0000003
0.0000903
137.4758
0.00486
25.9
0.2050041
0.000532
0.0068591
0.00241
2.140e-08
0.0713
0.2910089
0.0000049
0.0000022
0.0000127
0.0000089
0.041 1
0.000164
0.00061 1
0.0883
0.0811
0.0000003
0.0000165
0.0000036
0.00719
0.000258
0.000147
0.0000017
0.0512
3.600e-09
0.0000053
0.0000075
0.0000037
0.0000094
0.0000054
0.0000093
0.00001
0.0000106
0.0000103
0.0000121
0.0000085
0.0000052
0.0000108
0.0000112
0.0000118
0.0000093
0.0000112
0.0000093
0.0000102
0.0000081
0.0000077
0.0000081
0.000007
0.0000081
0.0000096
0.0000102
0.0000088
0.0000114
0.0000101
0.0000101
0.0000059
0.000009
0.0116
0.0692
0.0351
0.0573
0.021917
0.0355
0.0474
0.0298
0.0149
0.0728
0.0326
0.0203
0.249
0.1896
0.1008
0.086
0.1207
0.08
0.2166
0.0675
0.0584
0.0519
0.0414
0.0651
0.1385
0.0902
0.0799
0.1255
0.1365
0.0945
0.0443
0.06545
21.3
178
384
104
239
0
0
90
149.5
4.6
0
281
-4.4
-0.5
265
0
0
0
-6.47
0
145
0
0
170
77.9
207.5
90
74.8
63
163.5
174
270
C-6
-------
TABLE C-1. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS, I
COMPOUND NAME
CAPTAN
CARBARYL sevin
CARBENDAZIM
CARBON DISULFIDE
CARBON TETRACHLORIDE
CARBONYLSULFIDE
CATECHOL
CHLORACETOPHENONE.2-
CHLORAL
CHLORAMBEN
CHLORAMBUCIL
CHLORDANE
CHLORNAPHAZINE
CHLORO2BUTENE.1 trans
CHLOROACETALDEHYDE
CHLOROACETICACID
CHLOROALLYL ALCOHOL 2
CHLOROANILINE(2)
CHLOROANILINE(3)
CHLOROANILINE.p-
CHLOROAZOBENZENE
CHLOROBENZENE
CHLOROBENZILATE
CHLOROBENZOPHENONE (PARA)
CHLOROBENZOTRIFLUORIDE, P
CHLOROBENZYL ALCOHOL -(m)
CHLOROBENZYLATE
CHLOROBUTADIENE.1
CHLOROETHANE (ethyl chloride)
CHLOROETHYLENE
CHLOROHYDRIN, a 3 CHLORO 1 ,2
CHLOROMETHYL METHYL ETHER
CHLORONAPHTHALENE.2-
CHLORONITROBENZENE(-o)
CHLORONITROBENZENE, p
CHLOROPHENOL-2
CHLOROPHENOL-3
CHLOROPHENOL-4
CHLOROPRENE
CHLOROPROPANE-1
CHLOROPROPANE-2
A
N
H
S
C
O
A
A
O
C
C
P
C
H
C
C
C
C
C
C
C
C
C
C
H
C
O
H
C
C
MWT
300.57
201 .22
191.18
76.1
153.8
60
110
154.59
147.4
171.58
304.23
409.8
268.2
90.55
78.5
94.5
92.53
127.6
127.57
127.57
216.7
112.6
325.2
216.67
120.53
142.59
325.2
88.54
64.51
63.5
DENS.
g/cc
1
1.232
1
1.26
1.59
1
1.02
1.32
1.51
1.41
1.41
1.11
1.41
0.9295
1.11
1.41
1.41
1.41
1.213
1.41
1.41
1.11
1.41
1.41
1
1.41
0.97
0.96
1.41
1.41
VP@25C
mmHg
0.00006
0.00004
0.3060232
366
113
3062.283
0.0153211
0.044
50
3.418682
0
0.00001
0
87.02964
60
0.13832
1.5
1
1
0.015
0.0757537
11.8
0.0000022
0.0000026
103.0833
0.0050423
2.664e-09
160.0874
1200
2946.029
HL@25C
atm-m3/mol
0.0000468
0.0000324
0.0000025
0.019152
0.0302007
0.0000991
4.020e-1 1
0.0000009
0.00097
0.0000341
10.6
0.0000367
10.6
0.00187
0.000026
6.480e-08
0.0000183
0.0168
0.0168
10.6
0.0108
0.003762
9.414e-08
3.400e-09
0.0652549
0.0000029
5.080e-10
0.010102
0.0120961
0.00544
Dl
cm2/s
0.0000049
0.0000071
0.0000065
0.00001
0.0000088
0.000013
0.0000092
0.0000087
0.0000097
0.0000085
0.000006
0.0000044
0.0000065
0.0000097
0.0000115
0.0000121
0.0000123
0.0000101
0.0000093
0.0000101
0.0000074
0.0000087
0.0000058
0.0000074
0.0000086
0.0000095
0.0000046
0.00001
0.0000115
0.0000154
Dv
cm2/s
0.0183
0.0278
0.0327
0.104
0.078
0.1617
0.068
0.0383
0.0385
0.0323
0.0153
0.0118
0.018
0.0926
0.099
0.0733
0.0755
0.0483
0.0515
0.0483
0.0237
0.073
0.0141
0.0237
0.0605
0.0415
0.0168
0.0943
0.271
0.1299
Bpt
deg. C
312.111
274
214
46.3
76.8
-13.11
245.5
244
97.6
147
0
391
0
84
85
189
146
0
209
232
0
132
400
330
79.6001
237
415
68
12.4
0
PROPANEDIOL
H
C
C
C
A
C
C
C
C
C
C
110.54
80.51
162.62
157.56
157.57
128.6
128.6
128.6
88.5
78.54
78.54
1.345
1.41
1.41
1.52
1.534
1.26
1.24
1.31
0.958
0.89
0.87
0.024749
223.079
0.017
0.0760185
0.1022261
1.4
0.5
0.18
273
350
523
5.350e-09
0.0000864
0.0182
0.00788
0.0000914
0.0000083
0.0000033
0.000001 1
0.0009293
0.013
0.017
0.0000107
1 .34E-5
0.0000088
0.0000094
0.0000094
0.0000095
0.0000094
0.0000097
0.00001
0.0000103
0.0000101
0.06
0.092
0.0347
0.0351
0.0349
0.0501
0.0505
0.0493
0.104
0.1153
0.1164
213
59.1
256
242
236
175.6
214
217
0
0
0
C-7
-------
TABLE C-1. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS, I
COMPOUND NAME
CHLOROPROPENE3
CHLOROPROPIONITRILE.3-
CHLOROPROPYLENE-2
CHLOROTOLUENE-4
CHRYSENE
CRESOL
CRESOL(-m)
CRESOL(-o)
CRESOL(-p)
CROTONALDEHYDE
CUMENE HYDROPEROXIDE
CUMENE (isopropylbenzene)
CYANOGEN
CYCLOHEXANE
CYCLOHEXANOL
CYCLOHEXANONE
CYCLOHEXYLAMINE
CYCLOPENTADIENE
CYMENE.para
DAZOMET
DDE,p,p'-
DDT
DIAZINON
DIAZOMETHANE
DIBENZOFURANS
DIBENZOPYRENE1, 2,7,8
DIBROMOCHLOROMETHANE
DIBROMOETHANE-1,2
DIBROMOMETHANE
DIBROMO-3-CHLOROPROPANE.1 ,2
DIBROMO-4-HYDROXYBENZONITRILE,
DIBUTYLPHTHALATE
DICHLORO 2-PROPANOL 1 ,3
DICHLORO PROPANOL 2,3
DICHLOROANILINE2.3
DICHLOROANILINE(2,3)
DICHLOROBENZENE(1 ,2) (-0)
DICHLOROBENZENE(1,4) (-p)
DICHLOROBENZIDINE,3,3'-
DICHLOROBENZONITRILE,2,6-
DICHLOROBENZOPHENONE P,P
C
C
C
C
A
O
O
O
O
O
H
A
N
H
O
O
N
H
H
H
C
P
H
N
A
A
C
C
C
C
3,5
A
O
C
C
A
A
C
C
C
P
MWT
75.6
89.53
76.53
126.6
228.2
108
108.1
108.1
108.1
70.09
152.2
120.2
52.04
84.2
100.2
98.2
99.17
66.1
134.22
162.27
318.03
354.49
304.36
42.04
222
302.4
208.29
187.88
173.85
236.36
290.93
278.3
129
129
162
162.03
147
147
253.13
172.01
DENS.
g/cc
1.41
1.41
1.41
1.07
1.11
1.03
1.03
1.05
1.03
0.85
1.03
0.86
0.95
0.78
0.95
0.95
1.15
0.82
0.86
1
1.41
1.18
1
1.15
1.02
1.02
2.451
1.41
1.41
1.41
1
1.043
1.35
1.35
1
1
1.31
1.29
1.41
1.18
VP@25C
mmHg
361
2.5
0
2.8
5.760e-10
0.3
0.08
0.24
0.11
30
0.733844
4.6
3980
100
1.22
4.8
8.664001
4
1 .75365
0.2642225
0.0000065
0.0000002
0.0000118
381 1 .993
0.006698
0
15
11.78
48
0.8
0.0167509
0.0001
0.27
7
0.0033803
0.0114387
1.5
1.2
0.0013037
0.0005
HL@25C
atm-m3/mol
0.359
0.0000051
0.3589956
0.00466
1.180e-09
0.0000016
0.0000007
0.0000016
0.0000007
0.0000154
0.0000312
0.0131
0.00496
0.0137
0.0000045
0.0000041
0.0000423
19.3
0.0183
0.0000021
0.000068
0.114
2.150e-08
0.000013
0.00399
0.011
0.000783
0.0109
0.000998
0.0000236
0.0000005
0.0000003
0.00046
0.0000234
0.0000005
0.0000018
0.00194
0.00317
0.0000027
0.0000063
Dl
cm2/s
0.0000139
0.0000125
0.0000138
0.0000087
0.0000062
0.0000093
0.00001
0.0000083
0.00001
0.0000102
0.0000076
0.0000071
0.0000137
0.0000091
0.0000083
0.0000086
0.0000104
0.0000109
0.0000073
0.0000072
0.0000059
0.000005
0.0000049
0.0000175
0.000006
0.000005
0.0000105
0.0000081
0.0000084
0.000007
0.000005
0.0000079
0.0000098
0.0000098
0.0000072
0.0000072
0.0000079
0.0000079
0.0000067
0.0000076
Dv
cm2/s deg
0.1008
0.0791
0.099
0.055
0.0248
0.0694
0.074
0.074
0.074
0.0903
0.0436
0.065
0.20355
0.0839
0.214
0.0784
0.0745
0.1525
0.056
0.0406
0.0144
0.0137
0.018
0.260532
0.0267
0.018
0.0196
0.0287
0.0318
0.0212
0.0191
0.0438
0.0484
0.0484
0.0407
0.0407
0.069
0.069
0.0194
0.0349
Bpt
. C
0
0
0
162
488
195
202
190.8
203
99
158
153
-21.2
81
161
157
0
40
177
217
348.111
260
369.2
-23
287
0
120.2
0
0
196
244.0
340
174
182
272
252
179
173.4
287.111
270
C 251.11 1.41 0.0000029 0.0000047 6.8E-6 0.0196
DICHLOROETHANE(1,1)ethylidenedichloride
C 98.96 1.17 591 0.0056 0.0000105 0.0742
353
C-E
-------
TABLE C-1. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS, I
COMPOUND NAME MWT
DICHLOROETHANE(1,2) C 99
DICHLOROETHENE 1,2 trans C 96.94
DICHLOROETHENE(1,1) vinylidene chloride
C 96.97
DICHLOROETHYL ETHER C 143.02
DICHLOROETHYLENE(1,2)cis C 96.95
DICHLOROMONOFLUOROMETHANE
C 102.92
DICHLOROPHENOL A 163.01
DICHLOROPHENOL2.5 A 163.01
DICHLOROPHENOL(2,4) C 163.01
DICHLOROPHENOL(2,6) C 163
DICHLOROPHENOXYACETIC ACID(2,4)
C 221
DICHLOROPROPANE 1,2 C 112.99
DICHLOROPROPENE(1,3) C 111
DICHLOROPROPYLENE.1 ,2- (cis)
C 110.97
DICHLOROPROPYLENE.1 ,2-(trans)
C 110.97
DICHLORO-2-BUTENE 1,2 C 124.9
DICHLORO-2-BUTENE(1,4) C 125
DICHLORO-2-BUTENE, 1,4 H 125
DICHLORVOS A 220.98
DIELDRIN P 380.93
DIETHANOLAMINE N 105.14
DIETHYLAMINE N 73.14
DIETHYL ETHER O 74.14
DIETHYL SULFATE S 154.19
DIETHYL (N,N) ANILINE N 149.23
DIETHYLBENZENE P A 134.22
DIETHYLENE GLYCOL DIETHYL ETHER
H 162.23
DIETHYLENE GLYCOL DIMETHYL ETHER
H 162.23
DIETHYLENE GLYCOL MONOBUTYL ETHER
H 162.23
DIETHYLENE GLYCOL MONOETHYL ETHER
H 134.18
DIETHYLENE GLYCOL MONOETHYL ETHER ACETATE
H 192.23
DIETHYLENE GLYCOL MONOMETHYL ETHER
H 120.15
DIETHYLHYDRAZINE N,N H 88.2
DENS.
g/cc
1.26
1.41
1.213
1.41
1.28
1.41
1
1
1.41
1.41
1.41
1.156
1.2
1.41
1.41
1.41
1.19
1.188
1.415
1.18
1 .0881
1.15
0.97
1.177
0.93
1.02
0.909
0.937
0.967
0.999
1
1.01
1
VP@25C
mmHg
80
331
630
8.230494
200
1360
3.600194
0.1120569
0.12
0.034
290
40
43
375.841
113.8602
0
2.87
18.3067
0.012
0.0000002
0.0005463
224.959
534.1944
0.0342
0.00283
0.9956
1.214403
3.443058
7.556084
0.0501
0.113302
8.298628
0
HL@25C
atm-m3/mol
0.0011769
0.0644767
0.0259005
0.0000205
0.0154982
921
0.0000501
0.0000016
0.0000048
0.0000048
0.0621
0.002862
0.0035499
0.00898
0.011
10.6
0.000259
0.00165
0.0000003
0.0000584
1.800e-12
0.00731
0.000265
0.0000061
5.740e-08
0.00671
0.0000021
0.0000015
2.160e-08
4.860e-08
0.0000006
4.140e-08
0.0003
Dl
cm2/s
0.0000099
0.0000119
0.0000104
0.0000095
0.0000113
0.0000115
0.0000071
0.0000071
0.0000088
0.0000088
0.0000065
0.0000087
0.00001
0.00001 1
0.00001 1
0.0000103
0.0000081
0.0000093
0.0000073
0.0000047
0.0000098
0.0000125
0.0000086
0.0000081
0.0000059
0.0000081
0.0000068
0.0000069
0.000007
0.000008
0.0000065
0.0000086
0.0000103
Dv
cm2/s
0.104
0.0707
0.09
0.0413
0.0736
0.0650
0.0404
0.0404
0.0346
0.0347
0.0588
0.0782
0.0626
0.0586
0.0586
0.0498
0.0725
0.0534
0.02315
0.0125
0.0704
0.1147
0.0782
0.0409
0.0513
0.0519
0.0424
0.0418
0.0412
0.0524
0.0325
0.0605
0.0932
Bpt
deg. C
0
47.7
31.9
142
60.7
9
145
211
210
220
47.2
95.5
112
92.5
77
0
158
123
227.111
0
268.8
0
245
208
0
0
185
162
231
202
189
194
0
DIETHYLTHIOPHOSPHATEBENZO M ETHYL PETHER
H 253.31
0.0000095
2.120e-08
0.0000055
0.0227
392.001
C-9
-------
TABLE C-1. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS, I
COMPOUND NAME
DIISOBUTYLENE
DIISODECYL PHTHALATE
DIISOPROPYL BENZENE (PARA)
DIISOPROPYL KETONE
DIISOPROPYLAMINE
DIMETHOXY-(3,3')-BENZIDINE
DIMETHYL AMINE
DIMETHYL BENZYLAMINE N,N
DIMETHYL BENZ(A)ANT7,12
DIMETHYL CARBAMOYL CHLORIDE
DIMETHYL DISULFIDE
DIMETHYL FORMAMIDE
DIMETHYL HYDRAZINE(1,1)
DIMETHYL PHTHALATE
DIMETHYL SULFATE
DIMETHYL SULFIDE
DIMETHYL TRISULFIDE
DIMETHYLACETAMIDE
DIMETHYLAMINOAZOBENZENE.4-
dimethylaniline N,N
DIMETHYLBENZIDINE 3,3
DIMETHYLBENZ(A)ANTHRACENE(7,
DIMETHYLETHYLAMINE
DIMETHYLPHENOL(2,4)
DIMETHYLPHENOL(3,4)
DIMETHYLSULFONE
DIMETHYLSULFOXIDE
DIMETHYLSULFOXIDE
DINITROBENZENEM
DINITROPHENOL2.4
DINITROTOLUENE2.6
DINITROTOLUENE(2,4)
DINITRO-o-CRESOL(4,6)
DIOXANE(1,4)
DIPHENYLHYDRAZINE(1 ,2)
DIPHENYLMETHANE
DIPROPYLAMINE
DIPROPYLBUTRAL
DIPROPYLENE GLYCOL
DI-n-OCTYL PHTHALATE
H
H
A
H
H
O
N
N
A
C
S
N
H
A
S
S
H
H
N
N
A
12)
A
N
O
O
H
H
H
N
N
N
N
N
O
B
B
N
H
H
H
MWT
112.22
446.7
162.28
114.19
101.19
244.32
45.09
135.23
212.28
107.54
94.2
73.09
60.1
194.2
126.14
62.12
126
45.082
212
128
212.28
256.33
73.19
122.16
122.17
94.33
58.08
78.13
168.1
184
182.14
182.1
198
88.2
184.23
168.23
101.22
156
146
390.56
DENS.
g/cc
0.708
1
1.02
0.806
0.722
0.97
1.15
1.15
1
1.168
1.046
0.9445
0.791
1.19
1.26
1.26
1
0.688
1.15
0.956
1.02
1.02
1.11
1.04
0.97
1
1
1
1.56
1.68
1.15
1.31
1.15
1.03
1.19
1.19
1.15
1
1.023
0.76
VP@25C
mmHg
44.69303
0.0000001
1
12.99374
73.17715
0.0758996
1520
0.07587
0
1 .757571
29.488
4
157
0.0002
0.1
420
2.200398
1483.699
0.0006401
0.708
0.0000014
1 .620e-09
20
0.0573
0
0.0923178
0
1 .008534
0.05
0.014
6
0.0051
0.018
37
0.000052
0.0003899
189.9606
0.2041487
0.0388421
4.849e-09
HL@25C
atm-m3/mol
0.1177248
0.000408
0.107
0.000568
0.000307
0.00244
0.0000052
0.00135
0.0003
0.0000004
0.0000015
0.0000002
0.0000016
0.000001
0.000004
0.00545
3.034249
0.0000102
7.320e-08
0.0000139
1 .352e-09
2.730e-10
0.000385
0.000921
0.0072
2.270e-08
0.0003
0.0000005
0.000022
0.0000051
0.0000091
0.0000072
0.0000014
0.0000055
0.0000002
0.0000363
0.000253
0.0000867
1 .590e-08
0.137
Dl
cm2/s
0.0000073
0.0000039
0.0000072
0.0000078
0.0000078
0.0000055
0.0000167
0.0000087
0.0000061
0.00001
0.0000101
0.0000103
0.0000109
0.0000063
0.0000096
0.0000146
0.0000083
0.0000123
0.0000066
0.0000141
0.0000062
0.000005
0.0000122
0.0000087
0.0000083
0.0000099
0.0000132
0.00001 1 1
0.0000076
0.0000091
0.0000073
0.0000071
0.0000069
0.0000102
0.0000074
0.0000078
0.0000103
0.0000073
0.0000077
0.0000036
Dv
cm2/s deg
0.0778
0.0111
0.0402
0.0717
0.0887
0.0242
0.2342
0.0487
0.0285
0.0662
0.0834
0.0939
0.106
0.0568
0.0514
0.140
0.057
0.286963
0.0268
0.152982
0.0283
0.0461
0.1163
0.0584
0.0602
0.0849
0.1696
0.111
0.279
0.0273
0.0327
0.203
0.0293
0.229
0.0317
0.0358
0.0724
0.0428
0.0463
0.0151
Bpt
. C
101
451 .01
0
124
84
311
6.9
0
0
167
109
153
63.9
283.8
188
36.2
172
7
298.111
193
384
477
0
211.5
0
238
0
189
0
250
0
0
254.7
0
220
0
0
209.8
206
490
C-10
-------
TABLE C-1. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS, I
COMPOUND NAME
dodecane
ENDOSULFAN
ENDRIN
EPICHLOROHYDRIN
EPOXYBUTANE1.2
ETHANE
ETHANOL
ETHANOLAMINE(mono-)
ETHOXYETHANOL-2
ETHYL ACRYLATE
ETHYL CARBAMATE
ETHYL ETHER
ETHYL MORPHOLINE, ethyl diethylene
ETHYL TOLUENE, 4
ETHYL VINYL ETHER
ETHYLACETATE
ETHYLAMINE
ETHYLBENZENE
ETHYLENE DIAMINE
ETHYLENE DIBROMIDE
ETHYLENE GLYCOL
H
P
P
C
C
H
O
N
O
O
C
O
oxime
H
H
H
O
N
A
N
C
O
MWT
170
406.95
380.93
92.5
72.11
30
46.1
61.09
90
100
89.09
74.12
115.18
120.2
72.11
88.1
45
106.2
60.11
187.88
62.08
DENS.
g/cc
0.75
1.18
1.18
1.18
0.826
0.76
0.79
1.02
0.9
0.92
1.41
0.71
1
0.861
0.754
0.9
1.15
0.87
1.15
2.7
1.11
VP@25C
mmHg
0.2747982
0.00001
0.0000002
17
207.912
2970
50
0.4
5.4
40
0.6218955
520
0.4095041
3.44306
573.6588
100
1057.349
10
10.18362
14
0.126
HL@25C
atm-m3/mol
0.0664
0.0000191
0.0000004
0.0000335
0.000461
0.0494
0.0000303
0.0000003
0.0000064
0.000254
1.170e-08
0.00068
0.0000002
0.0128
0.00213
0.000128
0.0000256
0.0078806
0.0000085
0.00065
1 .800e-09
Dl
cm2/s
0.0000059
0.0000046
0.0000047
0.0000098
0.0000103
0.0000167
0.000013
0.0000114
0.0000096
0.0000086
0.0000126
0.0000093
0.0000088
0.0000078
0.0000098
0.0000097
0.0000168
0.0000078
0.0000141
0.0000119
0.0000122
Dv
cm2/s deg
0.0436
0.0115
0.0125
0.086
0.1343
0.5139
0.123
0.107
0.0947
0.077
0.0796
0.074
0.0644
0.0649
0.1396
0.0732
0.2349
0.075
0.152523
0.0217
0.108
Bpt
. C
216.2
390
445
117
61
0
0
172
135
100
183
34.5
208
162
33
77
0
136.2
126
131.6
0
ETHYLENE GLYCOL DIMETHYL ETHER
H
90.12
0.867
83.73
0.000035
0.0000094
0.0961
85
ETHYLENE GLYCOL MONOBUTYL ETHER
H
118.18
0.903
0.339572
0.0000005
0.0000082
0.065
171
ETHYLENE GLYCOL MONOBUTYL ETHER ACETATE
H
176
1
3.416916
0.0000049
0.0000068
0.0365
208
ETHYLENE GLYCOL MONOETHYL ETHER ACETATE
H
148.17
1
1 .746256
0.0000018
0.0000076
0.0458
143
ETHYLENE GLYCOL MONOETHYL ETHER Cellosol
ETHYLENE GLYCOL MONOMETHYL
ETHYLENE GLYCOL MONOMETHYL
ETHYLENE GLYCOL MONOPHENYL
ETHYLENE GLYCOL MONOPROPYL
ETHYLENE OXIDE
ETHYLENE THIOUREA
ETHYLHEXYLACRYLATE.2-
ETHYLPHENOL.3-
ETHYL(2) HEXANOL
H
ETHER
H
ETHER
H
ETHER
H
ETHER
H
O
N
H
P
O
90.12
76.1
ACETATE
134.14
138.17
104.16
44
102.17
184.28
122.17
130.22
0.93
1
1
1.102
1
0.87
1.15
0.88
1.18
0.8344
2.746598
4.928497
5.066
5.042345
3.072762
1250
0.3716955
1.271859
1
0.36
0.000001 1
0.0000008
0.0000022
6.840e-08
0.0000009
0.0002381
1.520e-10
0.00294
0.0000001
0.0000617
0.0000098
0.0000112
0.000008
0.0000084
0.0000093
0.0000145
0.0000102
0.0000061
0.0000094
0.0000073
0.0932
0.114805
0.0529
0.0482
0.0739
0.104
0.0715
0.0364
0.0553
0.0592
135
124.5
124
237
133
0
198.222
184
0
184
C-ll
-------
TABLE C-1. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS, I
COMPOUND NAME
FLUORANTHENE
FORMALDEHYDE
FORMAMIDE
FORMIC ACID
FREON 11, fluorotrichloromethane
A
O
O
O
C
MWT
202
30
45
46.03
137.37
DENS.
g/cc
1.02
0.97
0.97
1.22
1.494
VP@25C
mmHg
0.0177
3500
342.0861
42
795.2637
HL@25C
atm-m3/mol
0.067
0.0000003
0.0000012
0.0000007
0.0527
Dl
cm2/s
0.0000064
0.0000198
0.0000151
0.0000014
0.00001
Dv
cm2/s deg
0.0302
0.178
0.2509
0.079
0.0426
Bpt
. C
250
-14
0
100.7
23.7
FREON12 DICHLORODIFLUOROMETHANE
FREON 12, dichlorodifluoromethane
FUMARICACID
FURAN
FURFURAL
Generic Organic material
GLUTARICACID
GLYCERIN (GLYCEROL)
GLYOXAL
GLYPHOSATE
GUTHION
GYLCIDOL
HEPTACHLOR
HEPTANE ISO
HEPTANE(-n)
HEXACHLOROBENZENE
HEXACHLOROBUTADIENE
C
C
O
O
O
H
O
O
O
H
H
O
C
H
H
C
C
120.92
120.91
116.07
68.08
96.09
100
132.13
92.09
58.04
169.07
317.34
74.08
373.35
100.21
100.02
284.8
260.8
1.41
1
0.97
0.94
1.16
1
0.97
1.25
1.27
1
1
1.11
1.57
0.76
0.68
2.04
1.67
5000
6768.665
0.0760844
596
2
25
0.000291 1
0.00016
221.1501
0.1363822
0.0002913
0.925
0.0003
66
46
0.0048
0.15
0.401
0.7809416
0.0000017
0.00534
0.000081 1
0.000018
2.060e-08
1 .300e-08
0.00001 1
7.800e-09
1 .680e-09
0.0000009
0.0023
4.35
2.01998
0.0017001
0.0103002
0.0000105
0.0000085
0.0000086
0.0000122
0.0000104
0.0000096
0.000008
0.0000115
0.0000153
0.000007
0.0000048
0.0000122
0.0000057
0.0000071
0.0000076
0.0000059
0.0000062
0.052
0.0603
0.0646
0.104
0.0872
0.0782
0.0542
0.0798
0.1544
0.0385
0.0171
0.114272
0.0112
0.187
0.0926
0.0542
0.0561
-29.8
-41 .001
0
31.4
161.7
125
340
216
51
186.001
340.001
162
339
0
0
322
215
HEXACHLOROCYCLOHEXANE (GAMMA ISOMER)
HEXACHLOROCYCLOPENTADIENE
HEXACHLOROETHANE
HEXAFLUOROACETONE
HEXAMETHYLENE 1,6 DIISOCYANATE
HEXAMETHYLPHOSPHORAMIDE
HEXANE(-n)
HEXANOICACID
HEXANOL-1
HEXEN-2-ONE 5
HYDRAZINE
HYDROGEN SULFIDE
HYDROQUINONE
HYDROXYACETIC ACID
INDENO(1 ,2,3-cd)-PYRENE
ISOBUTYLENE
ISODECANOL
ISOPHORONE
LINDANE hexachlorocyclohexane
C
C
C
O
N
N
H
O
O
O
N
C
O
O
N
H
H
H
C
290.83
272.77
237
166.02
168.27
179.2
86.2
116.06
102.18
98.16
32.06
34.1
110.11
76.05
276.34
56.11
152.2
138.21
290.85
1.87
1.7
2.09
1
1.04
1.03
0.66
0.9265
0.82
0.847
1 .0081
1.41
1
0.97
1.15
0.626
1
0.92
1.41
0.0000109
0.081
0.65
57353.46
0.038201 1
0.1190317
150
0.0763529
0.812
10.85701
14.4
15200
0.0006957
189.8651
1 .OOOe-09
2250.243
0.0264044
0.439
0.0309093
0.0000078
10.6
0.0083501
162.7182
0.0000003
1.280e-13
0.7680137
0.000001 1
0.0000182
0.00008
0.0000007
0.023
1 .440e-09
0.00019
5.070e-13
0.0366
0.0000053
0.0000066
0.0021
0.0000073
0.0000072
0.0000068
0.0000071
0.0000072
6.9E-6
0.0000078
0.0000084
0.0000075
0.0000088
0.000019
0.0000161
0.000009
0.00001 1
0.0000057
0.0000102
0.0000074
0.0000068
0.0000062
0.0142
0.0161
0.0025
0.0394
0.038
0.0352
0.2
0.0659
0.059
0.0862
0.4164
0.176
0.06853
0.11638
0.019
0.216458
0.0442
0.0623
0.0162
390
239
186
-101
255
233
69
205.7
158
128
113.5
-60.2
286
216.001
536
-6.9
212
215
259
C-12
-------
TABLE C-1. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS, I
COMPOUND NAME
MALEICACID
MALEIC ANHYDRIDE
METHACRYLICACID
METHANOL
METHOMYL
METHOXYCHLOR
METHYL 1-PENTENE2
METHYL ACETATE
METHYL ACRYLATE
METHYL AMINE
METHYL AZIRIDINE 2
METHYL BENZYL ALCOHOL 4
METHYL CHLORIDE
METHYL CHOLANTHRENE 3
METHYL CYCLOHEXANE
METHYL ETHERdimethyl ether
METHYL ETHYL KETONE, 2 butanone
METHYL FORMATE
METHYL HYDRAZINE
METHYL IODIDE
METHYL ISOBUTYL KETONE
METHYL ISOCYANATE
METHYL ISOPROPYL KETONE
METHYL MERCAPTAN
METHYL METHACRYLATE
METHYL MORPHOLINE
METHYL NAPTHALENE(I-)
METHYL NAPTHALENE(2-)
METHYL PARATHION
O
O
O
O
H
O
H
O
O
N
N
O
C
A
H
O
O
O
N
M
O
N
H
S
O
H
A
A
P
MWT
116.07
98.1
86.1
32
162.2
345.65
84
74.1
86.1
31.06
57.1
122.17
50.5
268.36
92.2
46.08
72.1
60.05
46.09
141.94
100.2
57.06
86.15
48.1
100.1
101.15
142.19
142.19
263.23
DENS.
g/cc
1.59
0.93
1.0153
0.79
1
0.97
0.76
0.934
0.97
1.15
1.15
1.015
0.95
1.02
0.76
0.97
0.82
0.97
0.866
1.65
0.8
1.15
0.805
0.999
0.95
0.92
1.02
1.02
1.18
VP@25C
mmHg
0.075846
0.0013406
0.1
114
0.00005
0.0001552
270.7512
235
88
1520
0
0.0864688
3830
0.000759
43
3980
100
500
49.6
400
15.7
461 .743
189.8966
1728.82
39
17.7398
0.0539
0.06772
0.0000098
HL@25C
atm-m3/mol
1 .470e-08
0.0000002
0.0000113
0.0000052
0.0000008
0.000259
10.5
0.000102
0.0000001
0.00538
0.1933
0.0000139
0.0088201
0.000134
0.979
0.00318
0.00013
0.13
0.0000004
0.00253
0.0003901
0.0105
0.000458
0.00363
0.0001409
0.0000033
0.00071
0.000058
6.789e-08
Dl
cm2/s
0.0000115
0.00001 1 1
0.0000105
0.0000164
0.0000072
0.0000045
0.000009
0.00001
0.0000102
0.000021
0.0000145
0.0000086
0.0000065
0.0000054
0.0000085
0.0000149
0.0000098
0.0000127
0.0000139
0.0000104
0.0000078
0.0000145
0.0000092
0.0000148
0.0000086
0.000009
0.0000078
0.0000078
0.0000059
Dv
cm2/s deg
0.0523
0.095
0.0958
0.15
0.0407
0.0156
0.113
0.104
0.0976
0.417
0.164502
0.059
0.126
0.0209
0.0986
0.242
0.0808
0.1635
0.253053
0.039
0.075
0.1647
0.1057
0.2244
0.077
0.0798
0.048
0.048
0.02
Bpt
. C
260.001
200
163
0
144.0001
350
0
56.9
77
0
0
215
-24
0
0
-23.6
79.6
0
87
42.4
115.8
39.1
94
1
101
117
0
0
0
METHYLENE CHLORIDE, dichloromethane
C
85
1.34
438
0.003
0.0000117
0.101
39.8
METHYLENE DIPHENYL DIISOCYANATE
METHYLENE DIPHENYLAMINE (MDA)
METHYLENEDIANILINE 4,4
METHYLENE-BIS (2-CHLOROANILINE)
METHYLSTYRENE (-4)
METHYL-TERTIARY-BUTYL ETHER
MONOMETHYL FORMANIDE
MORPHOLINE
NABAM
NAPHTHALENE
N
H
N
,4,4'-
N
A
O
O
N
H
A
236
198
172
267.15
118.19
83.1
59.07
87.12
256.35
128.2
1.15
1
1.15
1.15
1.02
0.97
1.011
1
1
1.14
0.0006439
0.000351
0.0004037
0.0000694
1.6188
185.949
0.7626558
10
0.0000004
0.23
4.860e-08
2.810e-08
0.0000005
0.0000003
0.00591
0.0005551
9.670e-10
0.0000573
2.500e-16
0.0004831
0.0000062
0.0000064
0.0000075
0.0000058
0.0000088
0.0000105
0.0000132
0.0000096
0.0000054
0.0000075
0.0233
0.0313
0.0353
0.0199
0.0616
0.1024
0.1647
0.091
0.0224
0.059
327.111
337
305.111
331
0
64
184
129
416.0001
218
C-13
-------
TABLE C-1. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS, I
COMPOUND NAME
NAPHTHOL.alpha-
NAPHTHOL.beta-
NAPHTHYLAMINE.alpha-
NAPHTHYLAMINE.beta-
NEOPENTYL GLYCOL
NITROmXYLENE, 2
NITROANILINE P
NITROBENZENE
NITROBIPHENYL.4-
NITROGLYCERIN
NITROMETHANE
NITROPHENOL.2-
NITROPHENOL.4-
NITROPROPANE2
NITROSODIMETHYLAMINE N
NITROSOMORPHOLINE
NITROSO-N-METHYLUREA N
NITROTOLUENE (-p)
NITROTOLUENE, m
NITROTOLUENE, o
NITROTOLUENE, o
NONANOL, n
OCTANE
OCTANOL 1
OCTANOL 2
OCTANOL 3
OCTANOL 4
OIL
OXALIC ACID
PARALDEHYDE
PARATHION
PCB'S (Aroclors)
PENTACHLOROBENZENE
PENTACHLOROETHANE
PENTACHLORONITROBENZENE
PENTACHLOROPHENOL
PENTADIENE 1,2
PENTAERYTHRITOL
PHENOL
PHENYLCYCLOHEXANONE 4
PHENYLENE DIAMINE(-m)
PHENYLENE DIAMINE(-o)
PHENYLENE DIAMINE(-p)
PHENYLPHENOLP
PHOSGENE (decomposes)
O
O
N
N
O
A
N
N
B
N
N
N
N
N
N
N
N
N
A
A
A
H
H
O
O
O
O
H
O
O
P
B
C
C
C
C
H
H
O
O
N
N
N
O
O
MWT
144.18
144.17
143.19
143.19
104
151.17
138.14
123.1
199.21
227.09
61.05
139.11
139.11
89.09
74.08
116.14
103.1
137.13
137.13
137.14
137.13
144.26
114.3
130.3
130.3
130.3
130.3
170
90.04
132.3
291.3
292
250.34
202.3
295.36
266.4
67.1
136.15
94.1
175
108.14
108.14
108.14
170.2
98.92
DENS.
g/cc
0.97
0.97
1.15
1.15
0.97
1.117
1.15
1.2
1.19
1.6
1.15
1.15
1.4
0.9876
1.15
1.15
1.15
1.15
1.063
1.163
1.163
0.827
0.7
0.8259
0.8207
0.8216
0.8192
0.76
1.65
0.99
1.26
1.19
1.61
1.67
1.41
1.98
0.76
0.659
1.07
0.97
1.14
1.15
1.15
0.97
1
VP@25C
mmHg
0.002204
10
0.0005366
0.0003781
0.0341692
0.1869559
0.0015
0.3
0.0002913
0.0036
27.8
0.0013423
0.0001
101.517
8
0.0990759
0.8041194
1
0.1384812
0.1778758
0.1778758
0.021735
17
0.124
0.15808
0.09424
0.12464
0
0.0006077
25.3
0.0000378
0.0041
0.0046
4.4
0.0000991
0.005
367.458
0.0098698
0.341
1 .4668
0.028
0.008
0.0046
10
1390
HL@25C
atm-m3/mol
2.540e-08
1 .390e-08
5.000e-08
3.520e-08
1 .600e-08
0.000427
0.000227
0.0000239
0.0000073
6.000e-19
0.0235
0.0000001
1.314e-09
0.000119
0.0000008
8.540e-08
2.590e-08
0.000408
0.0000684
0.0000879
0.0000879
0.0000045
3.87
0.0000434
0.0000118
0.0000071
0.0000094
3.5
3.600e-09
0.0000367
0.0000006
0.000864
0.0073
0.021
0.000385
0.0000882
0.0119
3.700e-10
0.0000013
0.00875
1 .080e-08
1 .080e-08
1 .260e-09
0.0032
0.0140401
Dl
cm2/s
0.0000076
0.0000076
0.0000084
0.0000084
0.0000092
0.000008
0.0000086
0.0000086
0.000007
0.0000078
0.000014
0.0000085
0.0000096
0.0000101
0.0000124
0.00001
0.0000102
0.0000086
0.0000082
0.0000087
0.0000087
0.0000069
0.0000071
0.0000073
0.0000073
0.0000073
0.0000073
0.0000059
0.0000137
0.000008
0.0000058
0.000008
0.0000063
0.0000073
0.0000061
0.0000061
0.0000103
0.0000062
0.0000091
0.0000067
0.0000099
0.0000099
0.0000099
0.0000068
0.000001 1
Dv
cm2/s
0.0482
0.0482
0.0451
0.0451
0.0751
0.0425
0.0473
0.076
0.0286
0.0211
0.1491
0.0469
0.043
0.0923
0.1126
0.059
0.0706
0.0478
0.0495
0.0475
0.0476
0.0518
0.0763
0.0593
0.0595
0.0595
0.0595
0.0433
0.0736
0.0536
0.017
0.0175
0.057
0.066
0.0159
0.056
0.154
0.0621
0.082
0.0373
0.0663
0.066
0.06615
0.0387
0.108
Bpt
deg. C
0
285
300.8
306.1
208
224
0
0
340
260
0
216
279
120.3
152
225
178
0
230
225
225
215
125.7
0
0
0
0
0
155
125
375
306.111
277
162
357
310
0
227
182
0
284
258
267
0
8.1
C-14
-------
TABLE C-1. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS, I
COMPOUND NAME
PHOSPHINE
PHTHALICACID
PHTHALIC ANHYDRIDE
PHTHALIMIDE
PINENE(alpha-)
PIPERAZINE
PROPANE
PROPANE SULTONE, 1,3-
PROPANOICACID
PROPANOL
PROPANOL ISO
PROPENE
PROPIOLACTONE b
PROPIONALDEHYDE
PROPIONICACID
PROPORUR (Baygon)
PROPYL ACETATE ISO
PROPYLAMINE ISO
PROPYL ETHER ISO
PROPYLENE
PROPYLENE CHLOROHYDRIN
PROPYLENE GLYCOL
PROPYLENE OXIDE
PROPYLENIMINE 1 ,22 methyl aziridine
PROPYL(-n) ACETATE
PROPYL(-n) BENZENE
PROPYN-1-OL 2(PROPARLGYL)
PYRIDINE
QUINOLINE
QUINONE
RESORCINOL
SILVEX
SODIUM FORMATE
STYRENE
STYRENE OXIDE
SUCCINICACID
TAMARON (METHAMIDIPHOS)
TEREPHTHALICACID
TERPINEOL, ALPHA
TETRACHLOROBENZENE(1 ,2,3,4)
S
O
O
N
H
H
H
H
O
O
O
H
O
O
O
N
O
N
O
H
H
O
O
N
O
A
O
O
A
N
O
P
M
A
O
O
H
A
H
MWT
34
166.14
148.1
147.1
136.2
86.14
44.09
122.14
74.08
60.1
60.09
42.08
72.1
58.08
74.09
209.24
102.13
59.08
102.18
42.12
94.54
76.11
58.1
57.09
102.13
120.19
56.06
79.1
129.16
108.09
110.12
269.51
68.01
104.2
120.15
118.09
141.12
166.13
154.24
DENS.
g/cc
1
1.59
1.33
1.15
0.86
1
0.76
1.392
0.97
0.8037
0.79
0.61
0.97
0.81
0.97
1.15
0.87
0.694
0.97
0.76
1.103
1.04
0.83
1.15
0.89
0.86
0.97
0.98
1.02
1.318
0.97
1.18
1.65
0.9
1.054
0.97
1
1
1
VP@25C
mmHg
2000
121
0.0015
118
5
3.72
760
0.0011886
8.553405
20.824
42.8
8168.507
3.4
300
10
0.0171961
73.954
460
150
7600
3.072759
0.3
525
153.257
35
2.5
11.6
20
0.095
172.1972
0.00026
0
0.0007599
7.3
0.798015
0.0007615
127.3185
0.0039109
2.071076
HL@25C
atm-m3/mol
0.2270003
0.0132
0.0000008
0.0114
0.0262
0.0000422
0.022
9.010e-09
0.0000018
0.0000019
0.000015
0.0427
0.0000001
0.0000598
0.0000487
5.770e-08
0.000317
0.000358
0.00224
2.11
0.0000024
0.0000015
0.0003559
0.0000094
0.000294
0.00659
0.0000086
0.0000236
0.0000003
0.000001
1 .880e-08
4420
1 .700e-09
0.0026048
0.0000893
1 .740e-09
0.0000045
1 .040e-08
0.000516
Dl
cm2/s
0.0000182
0.0000068
0.0000086
0.0000083
0.0000073
0.0000104
0.0000132
0.0000103
0.0000112
0.0000114
0.0000104
0.0000119
0.0000114
0.0000114
0.0000112
0.0000067
0.0000087
0.0000105
0.0000093
0.0000136
0.0000105
0.0000102
0.00001
0.0000145
0.0000088
0.0000078
0.0000133
0.0000076
0.0000083
0.0000107
0.0000087
0.0000058
0.0000163
0.000008
0.0000089
0.0000085
0.0000078
0.0000071
0.0000074
Dv
cm2/s deg
0.381
0.064
0.071
0.0435
0.0549
0.091
0.285
0.0515
0.120812
0.176
0.098
0.334106
0.125
0.102
0.1208
0.0273
0.0807
0.192
0.0769
0.305
0.0811
0.093
0.104
0.1645
0.0799
0.065
0.181
0.091
0.0546
0.0624
0.078
0.0194
0.11052
0.071
0.0594
0.0631
0.0489
0.0394
0.0434
Bpt
. C
-87.78
72.6
284
69.5
156
146
0
300
141.1
97.2
82.4
-47
162.3
49.5
0
255.111
88.2
33
68.4
0
133
188
33.9
66
101.6
0
113.6
115.5
237.7
63.111
280
0
0
145
194
0
69.0001
296.5
140
TETRACHLOROBENZENE(1,2,3,5)
215.9
215.9
1.41
1.41
0.019
0.03
0.0027
0.00426
0.0000074
0.0000074
0.0239
0.0239
254
246
C-15
-------
TABLE C-1. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS, I
COMPOUND NAME
TETRACHLOROBENZENE(1 ,2,4,5)
TETRACHLORODIBENZO-p-DIOXIN(2,
TETRACHLOROETHANE(1 ,1 ,2,2)
TETRACHLOROETHENE
TETRACHLOROPHENOL(2,3,4,6)
TETRACHLOROPHENOL(2,3,5,6)
TETRAETHYL LEAD
C
3,7,8)
C
C
C
C
C
M
MWT
215.9
321 .96
168
165.83
231.9
231.9
323.45
DENS.
g/cc
1.86
1.41
1.59
1.624
1.41
1.41
1.653
VP@25C
mmHg
0.03
0.0000012
6.5
19
0.89
0.01
0.35
HL@25C
atm-m3/mol
0.00426
0.0000398
0.0002495
0.0177003
0.0000045
110.9999
0.0809
Dl
cm2/s
0.0000088
0.0000058
7.9e-6
0.0000082
0.0000071
0.0000071
0.0000064
Dv
cm2/s deg
0.0211
0.0143
0.071
0.072
0.0217
0.0217
0.0132
Bpt
. C
246
421
146.2
121.4
164
0
200
TETRAETHYLDITHIOPYROPHOSPHATE
TETRAETHYLENE PENTAMINE
TETRAFLUOROMETHANE
TETRAHYDROFURAN
TETRALIN
TETRANITROMETHANE
THIOUREA
THIOUREA,1-(o-CHLOROPHENYL)-
TOLUENE
TOLUENE DIAMINE(2,4)
TOLUENE DIISOCYANATE(2,4)
TOLUENEDIAMINE(2,6)
TOLUENEDIAMINE(3,4)
TOLUENESULFONYL CHLORIDE
TOLUIC ACID (para-)
TOLUIC ALDEHYDE
TOLUIDINEm
TOLUIDINE P
TOLUIDINE(-O)
TOXAPHENE
TRIBUTYL PHOSPHOROTRITHIOATE
TRI BUTYL TIN ACETATE
TRIBUTYLPHOSPHATE
TRICHLOROBENZENE 1,2,3
TRICHLOROBENZENE 1,2,4
TRICHLOROBENZENE 1,3,5
S
H
C
O
H
N
N
N
A
N
N
N
N
A
O
O
A
A
A
H
SSS
O
H
O
C
C
C
322.34
189.31
104.46
72.12
132.22
196.03
76.12
186.66
92.4
122
174.16
122.17
122.17
190.65
136.16
120.14
107.16
107.17
107.17
414
314.54
230.4
266.32
181.46
181.5
181.5
1.26
1
1
0.88
1
1.15
1.41
1.15
0.87
1.11
1.2
1.15
1.15
1
0.97
1.03
0.999
1.046
0.989
1.11
0.97
1
0.97
1.41
1.41
1.41
0.0000004
0.0521293
382429.8
72.1
0.40318
8
145
0.000002
30
1 .589889
0.08
0
0.0052068
0.0042605
0.00031
0.16
0.1877656
1
0.242
0.3
0.0007587
0.026341 1
127
0.2169527
0.18
0.23
7.200e-09
1.500e-13
60.40877
0.000049
0.00188
0.1933
0.00016
0.1933
0.0064201
1 .260e-09
0.0000002
0.1933
3.280e-09
0.0000022
5.600e-09
0.000253
0.0000016
0.0000191
0.0000024
0.00489
0.000157
0.00696
0.007543
0.00787
0.0019201
0.0209
0.0000055
6.5e-6
0.0000093
0.0000105
0.0000081
0.000007
0.0000138
0.0000072
0.0000086
0.0000091
0.0000062
0.0000092
0.0000092
0.0000065
0.0000078
0.0000087
0.0000092
0.0000094
0.0000091
0.0000043
0.0000047
0.0000058
0.0000052
0.0000082
0.0000082
0.0000082
0.015
0.0331
0.0736
0.098
0.0534
0.0297
0.107
0.0317
0.087
0.0569
0.061
0.0559
0.0559
0.0328
0.052
0.06
0.0711
0.069758
0.0714
0.0116
0.0175
0.0257
0.0216
0.03
0.03
0.03
335
229.
-176.0
67
0
126
182
323
110.7
284
251
0
265
295
275
204
203
200.6
200.4
384
470.
262.
289
221
213
208.5
TRICHLOROETHANE 1 ,1 ,1 methyl chloroform
TRICHLOROETHANE 1 ,1 ,2
TRICHLOROETHYLENE
TRICHLOROFLUOROMETHANE
TRICHLOROPHENOL 2,4,5
TRICHLOROPHENOL 2,4,6
TRICHLOROPROPANE 1,1,1
C
C
C
C
P
O
C
133.4
133.4
131.4
137.37
197.45
197.46
147.43
1.33
1.3
1.4
1.49
1.18
1
1.41
123
25
75
796
0.0399045
0.0073
3.1
0.0174002
0.0008239
0.0102002
0.0583
0.0000087
0.0000177
0.029
0.0000088
0.0000088
0.0000091
0.0000097
0.000007
0.0000064
0.0000079
0.078
0.078
0.079
0.087
0.0291
0.0314
0.071
75
74
87
23.8
249
244.5
107
TABLE C-1. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS, I
C-16
-------
COMPOUND NAME
TRICHLOROPROPANE(1 ,1 ,2)
TRICHLOROPROPANE(1 ,2,2)
TRICHLOROPROPANE(1 ,2,3)
C
C
C
TRICHLORO-1 ,2,2-TRIFLUOROETHANE,1 ,1 ,
TRIETHANOLAMINE
TRIETHYLAMINE
C
N
N
MWT
147.43
147.38
147.4
2-
187.38
149.19
101.22
DENS.
g/cc
1.41
1.41
1.41
1.41
1.15
0.7326
VP@25C
mmHg
6.64
1.37
3
270
0.01
400
HL@25C
atm-m3/mol
0.029
0.029
0.028
0.3915875
1.450e-10
0.000125
Dl
cm2/s
0.0000093
0.0000093
0.0000079
0.0000081
0.0000082
0.0000079
Dv
cm2/s
0.0396
0.0397
0.071
0.0288
0.0427
0.0881
Bpt
deg. C
140
124
156
47.7
335.4
89.6
TRIETHYLENE GLYCOL DIMETHYL ETHER
TRIFLUOROETHANE(1,1,1)
TRIFLURALIN
TRIISOBUTYLENE
TRIISOPROPYLAMINE
TRIMETHYL BENZENE, 123
TRIMETHYLBENZENE (1,3,5)
TRIMETHYLPENTANE 2,2,4
TRINITROTOLUENE(2,4,6)
TRIPROPYLENE GLYCOL
UREA
URETHANE
VINYL ACETATE
VINYL ACETYLENE
VINYL BROMIDE
VINYL CHLORIDE
VINYLIDENE CHLORIDE see 1
WARFARIN
XYLENE
XYLENE(-m)
XYLENE(-o)
XYLENE(-p)
XYLENOL(3,4)
XYLIDINEdimethylaniline
H
C
N
H
H
A
A
H
N
H
N
N
O
H
C
C
,1dichloroethen
C
P
A
A
A
A
O
A
178.17
84
335.29
170.32
185.34
120.2
120.2
114.22
227.1
160.26
60.06
89.09
86.09
52.08
106.96
62.5
96.97
308.33
106.2
106.16
106.2
106.16
122.17
121.18
0.986
1.41
1.15
1
1
0.89
1.02
0.76
1.15
1
1.34
1.15
0.93
1
1.517
0.91
1.213
1.18
1.02
0.86
0.88
0.86
0.97
0.98
0.277505
9240
0.0014133
0.4095041
0.216942
1 .920766
2.4
40.6
0.046
0.0047115
6.69
10
115
14664.62
969.1209
2660
630
0.00001
8.5
8
7
9.5
0.019
0.0819225
4.680e-08
84
0.00016
0.0917
0.00343
0.0112
0.147
3.338133
0.0000137
9.530e-08
0.000264
0.0000586
0.0005078
0.0962
0.00674
0.0264963
0.0259005
4420
0.00525
0.0074341
0.004878
0.0074402
3.940e-08
0.0000028
0.0000067
0.000013
0.000005
0.000007
0.0000066
0.000008
0.0000087
0.0000075
0.0000064
0.0000072
0.0000137
0.0000106
0.0000092
0.0000141
0.0000118
0.0000123
0.0000104
0.0000054
0.0000093
0.0000078
0.00001
0.0000084
0.0000083
0.0000084
0.0361
0.0866
0.0149
0.0381
0.0341
0.064
0.0602
0.0733
0.0245
0.0413
0.122
0.0866
0.085
0.1992
0.0598
0.106
0.09
0.0163
0.0714
0.07
0.087
0.0769
0.0602
0.0606
216
-47.3
314
208
221 .001
175
0
99.2
0
238
133
-40.3
73
-66
18
-13.9
31.9
356
140
139
144.4
138.4
225
218
C-17
-------
TABLE C-2. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS,
COMPOUND NAME
2,4 D
2,4,5 T
50% PEG
ACENAPHTHENE
ACENAPHTHYLENE
AC ETA L
ACETALDEHYDE
ACETALDOL
ACETAMIDE
ACETIC ACID
ACETONE
ACETONITRILE
ACETOPHENONE
ACETYL CHLORIDE
ACETYLAMINOFLUORENE.2-
ACETYLMETHYLPHTHALATE 4
ACETYL-2-THIOUREA.1-
ACIFLUORFEN
ACROLEIN
ACRYLAMIDE
ACRYLIC ACID
ACRYLONITRILE
ADENINE
ADIPONITRILE
ALDICARB
ALDRIN
ALLYL ALCOHOL
ALLYL CHLORIDE
ALLYL ETHER, diallyl ether
ALPHA METHYL STYRENE
AMETRYN
AMINOBIPHENYL.4-
AMINOPHENOL(-o)
AMINOPHENOL(-p)
AMINOPYRIDINE.4-
AMYL ACETATE(-n)
ANILINE
ANISIDINE.o-
ANTHRACENE
ANTHRAQUINONE
AZIRIDINE ethylene imine
BENZAL CHLORIDE
BENZALDEHYDE
BENZENE
BENZIDINE
BENZOFURAN 2,3
A
6.908157
16.13832
0
7.728
7.72819
6.849664
8.005
8.206501
7.156695
7.387
7.117
7.119
7.385889
6.948
7.25633
0
0
6.934575
7.212867
11.29315
5.652
7.11
7.221354
7.213714
6.927094
9.357671
9.148151
7.576274
6.84591 1
6.92366
7.225121
7.500307
9.228947
-3.357
9.589092
8.06791
6.9502
7.464861
8.91
7.289244
9.389235
9.49223
8.461938
6.905
7.5424
6.88042
B
1746.881
7280.503
0
2534.234
2534.234
1279.732
1600.017
2072.293
1300.449
1533.313
1210.595
1314.4
1891.5
1115.954
2110.188
0
0
1974.716
1297.327
3939.877
648.629
1335.674
1812.02
2072.084
1908.791
4347.02
2319.925
1493.914
1250.578
1486.88
1843.155
2235.337
2838.639
699.157
3663.803
2186.68
1467.242
1932.244
3761
241 1 .734
2142.318
3174.547
2523.01 1
1211.033
2625.8
1520.652
C
193.6385
273.16
0
245.576
245.576
219.7466
291 .809
204.1096
224.0406
222.309
229.664
230
217.885
223.554
182.2596
0
0
181.1196
246.6905
273.16
154.683
238.207
197.4596
183.2096
184.7296
273.16
273.16
273.16
221 .3996
202.4
195.8636
181.8796
273.16
-331 .343
273.16
273.16
177.115
196.5096
273.16
167.0596
273.16
273.16
273.16
220.79
163.256
206.1996
LN(OW)
3.228066
2.854958
-4.232786
3.92
4.07
2.287809
0.43
0.1735975
-0.775268
-0.31
-0.24
-0.34
1.58
-0.308331
3.482134
2.224537
-0.643182
5.0533
-0.09
0.8008422
0.31
-0.92
-0.16
1.675912
2.174818
5.669268
0.17
0.95454
2.51897
3.463532
1 .204255
3.605957
0.5815316
0.5815316
0.28
1 .708488
0.9
1 .605756
4.45
3.729464
0.1757197
2.152145
1.48
2.15
1.81
2.625319
Km ax
mg/g-hr
10.76
10.76
15.3
31.1
31.1
15.3
82.42
15.3
9.7
14
1.3
9.7
17.56
10.76
31.1
31.1
9.7
15.3
7.8
9.7
17.56
18
9.7
15.3
15.3
15.3
17.56
10.76
15.3
31.1
15.3
9.7
9.7
9.7
9.7
17.56
7.1
15.3
31.1
17.56
15.3
10.76
17.56
19
31.1
15.3
k1 RT HYD
L/g-hr 1/s
2.275134
1.64144
0.00332
4.16815
2.7
0.999325
0.1966685
0.1571452
0.0685
0.98
1.15
0.1002612
0.5379447
0.1030783
2.841536
0.9455041
0.0768
1 1 .2357
0.36
0.2720544
0.1770657
0.75
0.1173636
0.5850384
0.9052528
19.26059
0.1566514
0.3112157
1 .223344
2.795661
0.3872162
3.166695
0.2245521
0.2245521
0.1724783
0.6019543
21
0.5502056
2.2
3.528086
0.1574373
0.8874708
0.4928757
1.4
0.6578671
1 .342646
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
C-18
-------
TABLE C-2. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS,
COMPOUND NAME
BENZOICACID
BENZONITRILE
BENZOPHENONE
BENZOTRICHLORIDE
BENZOYL CHLORIDE
BENZO(A)ANTHRACENE
BENZO(A)PYRENE
BENZO(B)FLUORANTHENE
BENZO(k)FLUORANTHENE
BENZYL ALCOHOL
BENZYL CHLORIDE
BIPHENYL
BISPHENOL(A)
BIS(1 ,1 ,2,2-TETRACHLOROPROPYL) ETHER
BIS(2-CHLOROETHYL)ETHER
BIS(2-ETHYLHEXYL)PHTHALATE
BIS(CHLOROMETHYL)ETHER
BROMACIL
BROMOBENZENE
BROMOCHLOROMETHANE
BROMODICHLOROMETHANE
BROMOFORM
BROMOMETHANE
BROMOTOLUENE4
BROMOXYNIL
BUTADIENE-(1,3)
BUTANE
BUTANEDINITRILE
BUTANOL ISO
BUTANOL(S)
BUTANOL-1
BUTENE
BUTYL ACETATE(-n)
BUTYL ACRYLATE
BUTYL BENZENE
BUTYL CARBITOL
BUTYL CELLOSOLVE
BUTYLAMINE
BUTYLENE GLYCOL-(1 ,3)
BUTYLISOBUTYRATE
BUTYRALDEHYDE
BUTYRALDEHYDE ISO
BUTYRIC ACID
c10 linear
CAPROLACTAM
A
9.033
6.74631
7.34966
8.326414
7.9245
6.9824
9.245506
6.809279
11.10133
7.19817
7.53029
7.6317
8.64308
7.812044
8.210002
10.83704
8.178545
6.907701
6.86064
6.49606
7.9655
7.988101
7.566313
7.00762
7.531281
7.216854
3.18243
8.860728
7.32705
7.4768
7.4768
6.805603
7.127
8.141759
6.98317
7.74114
6.956
8.649454
9.583491
6.84419
6.358544
7.983729
8.064204
6.88042
7.243459
B
3333.3
1436.72
2331 .4
2691 .007
2372.1
2426.6
3724.363
952.6317
6191.376
1632.593
1923.019
2167.862
2910.876
1452.055
2404.325
5228.522
1998.099
1743.059
1438.817
942.267
1846.561
2158.654
1301.449
1612.35
2171.211
1144.753
0
3218.161
1248.48
1362.39
1362.39
918.4548
1430.418
2199.925
1577.965
2056.904
1399.903
2025.139
3221.718
1237.194
913.59
1715.402
2263.387
1520.652
1997.954
C
273
181
185
273.16
273.16
156.6
273.16
238.4996
273.16
172.79
234.3478
207.61
273.16
273.16
273.16
273.16
273.16
193.8496
205.441
192.587
273.16
273.16
273.16
206.36
185.8696
269.0367
0
273.16
172.92
178.77
178.77
240.4889
210.745
273.16
201 .378
195.655
172.154
273.16
273.16
222.1596
185.48
273.16
273.16
206.1996
187.9596
LN(OW)
1.86
2.046477
4.525765
2.92
1.959114
5.61
5.98
6.843004
6.84
1 .675986
2.3
4.266275
4.651325
4.322158
1.58
5.3
-0.38
8.145402
3.229374
0.9726127
1.88
2.3
1.1
3.536277
1 .989251
1.871113
1 .859488
0.35
0.75
0.75
0.83
1.865313
1 .79263
2.105592
4.028965
2.202231
1 .554456
0.88
-0.04555
2.53453
1.125761
1.126484
0.7236805
4.463722
0.8215773
Km ax
mg/g-hr
17.56
9.7
17.56
10.76
10.76
31.1
31.1
31.1
31.1
17.56
17.75
19
17.56
10.76
10.76
0.77
10.76
15.3
10.76
10.76
10.76
10.76
10.76
10.76
15.3
15.3
15.3
9.7
7.8
7.8
7.8
15.3
17.56
17.56
31.1
15.3
15.3
9.7
17.56
15.3
17.56
17.56
17.56
15.3
9.7
k1 RT HYD
L/g-hr 1/s
0.6872873
0.8090959
7.08167
1 .737565
0.7495522
18.28721
25.27833
53.78881
53.64763
0.5850764
1.010042
5.643243
7.904033
5.926039
0.5379447
0.35
0.0968
168.115
2.277738
0.3161761
0.6994205
1.010042
0.3534569
2.979392
0.7695807
0.6940029
0.6869795
0.1833727
0.11
0.11
0.11
0.6904899
0.6479443
0.8520479
4.585114
0.9272288
0.5260546
0.2915654
0.1297253
1.240113
0.3615145
0.3617431
0.2542928
6.707476
0.2770353
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
C-19
-------
TABLE C-2. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS,
COMPOUND NAME
CAPTAN
CARBARYL sevin
CARBENDAZIM
CARBON DISULFIDE
CARBON TETRACHLORIDE
CARBONYLSULFIDE
CATECHOL
CHLORACETOPHENONE.2-
CHLORAL
CHLORAMBEN
CHLORAMBUCIL
CHLORDANE
CHLORNAPHAZINE
CHLORO2BUTENE.1 trans
CHLOROACETALDEHYDE
CHLOROACETICACID
CHLOROALLYL ALCOHOL 2
CHLOROANILINE(2)
CHLOROANILINE(3)
CHLOROANILINE.p-
CHLOROAZOBENZENE
CHLOROBENZENE
CHLOROBENZILATE
CHLOROBENZOPHENONE (PARA)
CHLOROBENZOTRIFLUORIDE, P
CHLOROBENZYL ALCOHOL -(m)
CHLOROBENZYLATE
CHLOROBUTADIENE.1
CHLOROETHANE (ethyl chloride)
CHLOROETHYLENE
CHLOROHYDRIN, a 3 CHLORO 1 ,2
CHLOROMETHYL METHYL ETHER
CHLORONAPHTHALENE.2-
CHLORONITROBENZENE(-o)
CHLORONITROBENZENE, p
CHLOROPHENOL-2
CHLOROPHENOL-3
CHLOROPHENOL-4
CHLOROPRENE
CHLOROPROPANE-1
CHLOROPROPANE-2
CHLOROPROPENE3
CHLOROPROPIONITRILE.3-
CHLOROPROPYLENE-2
CHLOROTOLUENE-4
A
10.25674
1 1 .59653
6.897335
6.942
6.9339
7.117189
7.514789
8.6451
7.734462
7.427135
0
9.300818
0
6.841615
8.360256
7.55016
9.545546
7.56265
7.55939
9.386585
-1.120589
6.978
9.669569
11.15672
6.839732
8.235456
8.326468
6.8348
6.986
6.89117
PROPANEDIOL
8.222207
6.831054
8.657456
8.376652
6.906466
6.877
7.900331
8.510922
6.161
6.92648
7.771
5.29716
7.32973
0
8.177249
B
4316.929
4768.904
1657.229
1169.11
1242.43
968.5864
2030.245
2981.1
1799.546
1629.082
0
4263.922
0
1217.14
1962.524
1723.365
2793.597
1998.6
2073.75
3286.465
0
1431.05
4569.931
4991 .71
1202.451
2309.119
3133.646
1163.8
1030.01
905.01
2199.559
1134.219
3056.778
2831 .246
1732.741
1471.61
2445.317
2759.663
783.45
1110.19
1582
418.375
1732.55
0
2304.805
C
273.16
273.16
198.5996
241 .59
230
241 .7505
192.6146
273.16
273.16
21 1 .3296
0
273.16
0
223.2996
273.16
179.98
273.16
220
215
273.16
0
217.55
273.16
273.16
224.1356
194.2296
160.4096
226.3396
238.61
239.48
198.7896
228.0306
273.16
273.16
194.4196
193.17
273.16
273.16
179.7
227.94
288
128.168
21 1 .79
0
273.16
LN(OW)
2.92452
2.36
3.384036
2.16
2.72
-0.275866
0.94
2.594686
1 .490009
1 .593383
3.274369
2.78
3.950025
1 .349343
-0.219135
-0.621223
0.3672189
0
0
1.83
5.026951
2.5
6.169221
4.209192
2.560925
2.436677
5.37252
2.454152
1.43
0.6267082
-0.956336
-0.21
4.12
2.603606
2.603606
2.15
1 .403233
2.480466
0.57
0.9487147
1 .048029
0.9545268
0.422814
1 .985027
3.536277
Km ax
mg/g-hr
15.3
9.7
15.3
15.3
1.5
15.3
31.1
15.3
15.3
10.76
10.76
15.3
10.76
15.3
10.76
10.76
10.76
0.27
0.27
0.27
10.76
0.39
10.76
10.76
15.3
10.76
17.56
15.3
10.76
10.76
15.3
10.76
10.76
10.76
15.3
15
10.76
6.5
10.76
10.76
10.76
10.76
10.76
10.76
10.76
k1 RT HYD
L/g-hr 1/s
1 .74445
1 .064485
2.607808
0.8935917
1.5
0.1060484
0.3072813
1.307137
0.4972112
0.5442809
2.3692
1 .537236
4.279104
0.43963
0.1114454
0.0783
0.1861564
0.86
0.86
0.86
10.97962
10
29.82998
5.368304
1 .269087
1.138353
14.8561
1.155892
0.472
0.2336062
0.0584
0.1123397
4.965281
1.317378
1.317378
0.8858069
0.4608566
1.182815
0.2222978
0.3096333
0.3377439
0.311212
0.1954358
0.7667412
2.979392
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
C-20
-------
TABLE C-2. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS,
COMPOUND NAME
CHRYSENE
CRESOL
CRESOL(-m)
CRESOL(-o)
CRESOL(-p)
CROTONALDEHYDE
CUMENE HYDROPEROXIDE
CUMENE (isopropylbenzene)
CYANOGEN
CYCLOHEXANE
CYCLOHEXANOL
CYCLOHEXANONE
CYCLOHEXYLAMINE
CYCLOPENTADIENE
CYMENE.para
DAZOMET
DDE,p,p'-
DDT
DIAZINON
DIAZOMETHANE
DIBENZOFURANS
DIBENZOPYRENE 1,2,7,8
DIBROMOCHLOROMETHANE
DIBROMOETHANE-1,2
DIBROMOMETHANE
DIBROMO-3-CHLOROPROPANE.1 ,2
DIBROMO-4-HYDROXYBENZONITRILE, 3,5
DIBUTYLPHTHALATE
DICHLORO 2-PROPANOL 1 ,3
DICHLORO PROPANOL 2,3
DICHLOROANILINE2.3
DICHLOROANILINE(2,3)
DICHLOROBENZENE(1 ,2) (-0)
DICHLOROBENZENE(1,4) (-p)
DICHLOROBENZIDINE,3,3'-
DICHLOROBENZONITRILE,2,6-
DICHLOROBENZOPHENONE P,P
DICHLOROETHANE(1,1)ethylidenedichloride
DICHLOROETHANE(1,2)
DICHLOROETHENE 1,2 trans
A
10.68596
8.850432
7.508
7.426974
7.035
8.536501
8.191144
6.963
3.59986
6.841
6.255
7.84918
6.68954
6.823
6.881701
6.898588
10.32563
15.19374
7.284863
6.800292
6.927094
0
7.288803
6.72148
1.68123
7.886061
7.514083
9.386544
9.783355
6.746815
7.48674
7.47752
6.882553
7.199
7.493614
10.40392
10.5374
6.992756
7.068385
6.9651
B
5940.981
2794.746
1856.36
1744.32
1511.08
2104.827
1950.172
1460.793
0
1201.53
912.87
2137.192
1229.42
1071
1530.862
1667.51
4625.263
6564.769
2370.844
864.7323
1908.791
0
1733.934
1280.82
0
2348.271
2024.306
3955.114
3086.549
1759.657
2116.834
2038.121
1537.672
1690.291
2176.46
4086.263
4794.257
1176.864
1292.54
1141.9
C
273.16
273.16
199.07
194.444
161.85
273.16
209.2396
207.78
0
222.65
109.13
273.16
188.8
271.6
205.6296
198.0296
273.16
273.16
169.1116
243.6296
184.7296
0
273.16
201 .75
0
273.16
192.8996
273.16
273.16
273.16
187.5796
191.3796
205.2496
218.09
184.7085
273.16
273.16
228.838
225
231.9
LN(OW)
5.61
0
0.97
1.98
1.94
1 .09231 1
2.308645
3.497709
0.8077466
2.529003
1 .576838
0.81
1 .997921
2.028058
4.158295
3.518476
5.69
6.19
3.821204
-0.124468
5.695732
8.199224
2.09
1.137044
0.6958743
0.2066454
2.294361
5.2
-0.480189
-0.269502
3.024556
3.024556
3.38
3.39
3.51
3.75325
6.046987
1.79
1.45
1.48
Km ax
mg/g-hr
31.1
23
23.21
22.78
23.21
17.56
15.3
31.1
9.7
15.3
17.56
11.49
9.7
15.3
15.3
15.3
10.76
15.3
15.3
9.7
31.1
31.1
10.76
10.76
10.76
10.76
15.3
0.4
10.76
10.76
15.3
15.3
2.5
6.4
10.76
15.3
10.76
10.76
2.1
10.76
k1 RT HYD
L/g-hr 1/s
1.4
17
17
17
17
0.3510868
1.017711
2.880526
0.2737029
1.234131
0.5364586
0.2742431
0.7754406
0.7961605
5.134476
2.933347
19.61316
30.37728
3.822968
0.1210698
19.71178
176.2214
0.035
0.3651012
0.2481805
0.1617557
1 .005071
1
0.0886872
0.1066405
1 .904024
1 .904024
0.58
2.3
2.911671
3.602285
26.80424
2.3
0.98
0.4928757
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
DICHLOROETHENE(1,1) vinylidene chloride
6.9722
1099.4
237.2
10.76
0.9
C-21
-------
TABLE C-2. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS,
COMPOUND NAME
DICHLOROETHYL ETHER
DICHLOROETHYLENE(1 ,2) cis
DICHLOROMONOFLUOROMETHANE
DICHLOROPHENOL
DICHLOROPHENOL2.5
DICHLOROPHENOL(2,4)
DICHLOROPHENOL(2,6)
DICHLOROPHENOXYACETIC ACID(2,4)
DICHLOROPROPANE 1,2
DICHLOROPROPENE(1 ,3)
DICHLOROPROPYLENE.1 ,2- (cis)
DICHLOROPROPYLENE.1 ,2-(trans)
DICHLORO-2-BUTENE 1,2
DICHLORO-2-BUTENE(1 ,4)
DICHLORO-2-BUTENE, 1,4
DICHLORVOS
DIELDRIN
DIETHANOLAMINE
DIETHYLAMINE
DIETHYL ETHER
DIETHYL SULFATE
DIETHYL (N,N) ANILINE
DIETHYLBENZENE P
DIETHYLENE GLYCOL DIETHYL ETHER
DIETHYLENE GLYCOL DIMETHYL ETHER
DIETHYLENE GLYCOL MONOBUTYL ETHER
DIETHYLENE GLYCOL MONOETHYL ETHER
DIETHYLENE GLYCOL MONOETHYL ETHER
A
7.6924
7.0223
7.590301
7.466021
7.498359
7.497876
6.899838
8.500344
6.98
6.80731
6.845289
6.838623
0
8.312499
6.858446
9.964257
-6.744684
8.252672
5.8016
6.92032
7.838081
7.466
6.9982
7.86919
7.083
8.232163
8.216066
ACETATE
8.208758
B
1990.8
1205.4
1328.834
1635.588
1893.996
1890.059
1677.797
1800.279
1380.1
1327.64
1245.558
1193.779
0
2341 .934
1347.971
3543.651
0
2454.928
583.3
1064.07
2173.876
1993.57
1588.31
2094.43
1556.26
2281 .689
2149.489
2090.436
C
235.347
230.6
273.16
21 1 .7096
199.1696
199.3596
197.4596
273.16
222.8
230.1337
221 .6846
224.6296
0
273.16
215.8896
273.16
0
188.1876
144.1
228.8
230.36
218.5
201 .97
231 .887
210.37
195.3696
200.8796
203.3496
LN(OW)
0.685801
2.202489
1 .83253
3.241122
3.241122
2.75
3.241122
2.792008
2.28
1.98
2.735832
2.568767
0.8726078
0.8726078
0.8726076
0.9334952
6.906134
-0.66325
1.242518
1 .639246
-0.29
3.562567
4.158206
2.934824
1.781736
2.202231
1.33416
1 .2671 1 1
Km ax
mg/g-hr
10.76
10.76
10.76
15.3
15.3
25
10.76
10.76
17
10.76
10.76
10.76
10.76
10.76
15.3
31.1
15.3
9.7
9.7
17.56
15.3
9.7
31.1
15.3
15.3
15.3
15.3
15.3
k1 RT HYD
L/g-hr 1/s
0.2460026
0.9274379
0.6709645
2.301272
2.301272
3.3
2.301272
1 .553473
1.4
0.7633762
1 .478961
1 .277825
0.2896856
0.2896856
0.2896855
0.3055373
56.84358
0.0755
0.4003997
0.5665672
0.1048994
3.048722
5.134077
1 .760248
0.641797
0.9272288
0.4338283
0.4091089
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
DIETHYLENE GLYCOL MONOMETHYL ETHER
DIETHYLHYDRAZINE N,N
DIETHYLTHIOPHOSPHATEBENZOM ETHYL
DIISOBUTYLENE
DIISODECYL PHTHALATE
DIISOPROPYL BENZENE (PARA)
DIISOPROPYL KETONE
DIISOPROPYLAMINE
DIMETHOXY-(3,3')-BENZIDINE
8.211577
0
PETHER
6.96702
6.84893
6.987966
6.9933
7.177119
7.158473
7.051436
2113.133
0
2275.218
1274.031
2483.235
1663.88
1459.448
1314.511
2436.322
202.3996
0
164.7794
220.0696
153.5677
194.41
215.6996
223.2996
273.16
0.9001058
0.6255218
4.326276
3.671378
9.650949
4.928876
2.171245
2.210661
3.150456
15.3
15.3
15.3
15.3
15.3
31.1
15.3
15.3
17.56
0.2967401
0.2333639
5.947427
3.353253
627.6503
10.07671
0.9024271
0.9340937
2.125762
0
0
0
0
0
0
0
0
0
-------
TABLE C-2. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS,
COMPOUND NAME
DIMETHYL AMINE
DIMETHYL BENZYLAMINE N,N
DIMETHYL BENZ(A)ANT7,12
DIMETHYL CARBAMOYL CHLORIDE
DIMETHYL DISULFIDE
DIMETHYL FORMAMIDE
DIMETHYL HYDRAZINE(1,1)
DIMETHYL PHTHALATE
DIMETHYL SULFATE
DIMETHYL SULFIDE
DIMETHYL TRISULFIDE
DIMETHYLACETAMIDE
DIMETHYLAMINOAZOBENZENE.4-
dimethylaniline N,N
DIMETHYLBENZIDINE 3,3
DIMETHYLBENZ(A)ANTHRACENE(7,1 2)
DIMETHYLETHYLAMINE
DIMETHYLPHENOL(2,4)
DIMETHYLPHENOL(3,4)
DIMETHYLSULFONE
DIMETHYLSULFOXIDE
DIMETHYLSULFOXIDE
DINITROBENZENEM
DINITROPHENOL2.4
DINITROTOLUENE2.6
DINITROTOLUENE(2,4)
DINITRO-o-CRESOL(4,6)
DIOXANE(1,4)
DIPHENYLHYDRAZINE(1 ,2)
DIPHENYLMETHANE
DIPROPYLAMINE
DIPROPYLBUTRAL
DIPROPYLENE GLYCOL
DI-n-OCTYL PHTHALATE
dodecane
ENDOSULFAN
ENDRIN
EPICHLOROHYDRIN
EPOXYBUTANE1.2
ETHANE
ETHANOL
ETHANOLAMINE(mono-)
ETHOXYETHANOL-2
A
7.08212
-1.119923
0
7.197148
6.8257
6.928
7.408
4.52232
7.4455
7.1509
6.879565
7.124729
7.498566
7.72649
7.535748
10.57967
7.08212
8.926483
0
6.90729
0
6.886806
4.337
7.516903
4.372
7.981089
8.884898
7.350545
13.8359
6.291
2.278649
7.216752
8.218305
7.001507
6.898254
9.318406
9.681486
8.22943
6.83185
6.82915
8.321
7.456
7.874
B
960.242
0
0
1616.599
1303.5
1400.87
1305.91
700.31
1843.343
1195.58
1513.849
1039.441
2219.947
2110.009
2561 .72
5775.389
960.242
2930.103
0
1739.619
0
1571.757
229.2
2048.072
380
3074.44
3169.326
1517.53
5402.621
1261
0
1774.272
2167.685
2621 .431
1664.768
4269.167
4883.984
2086.816
1140.529
663.72
1718.21
1577.67
1843.5
C
221 .67
0
0
207.5296
218.4
196.43
225.53
51.42
217.055
242
206.5796
237.9296
182.6185
242.88
166.2996
273.16
221 .67
273.16
0
194.0396
0
203.3496
-137
191.7596
-43.6
280.23
273.16
238.065
273.16
105
0
199.3976
200.1196
146.1596
198.1816
273.16
273.16
273.16
227.6696
256.68
237.52
173.37
234.2
LN(OW)
0.05483
2.526956
7.456762
1 .044858
3.410681
-0.041408
0.595338
1.87
0.318303
1 .70649
5.829383
0.05483
3.521143
-0.88
4.169629
7.456761
1.126405
2.42
0
-0.264333
0.0361875
0.3634962
1 .52229
1.54
2.05
2.01
2.85
1 .22037
3.03
4.798269
2.110636
3.731519
0.1654824
9.2
5.331716
3.55
5.6
0.03
1.441351
0.9913252
-0.32
-0.773026
0.6863381
Km ax
mg/g-hr
9.7
9.7
15.3
10.76
15.3
9.7
15.3
2.2
15.3
15.3
15.3
15.3
9.7
9.7
31.1
31.1
9.7
10.7
5.5
15.3
15.3
15.3
9.7
8
9.7
9.7
9.7
17.56
19
19
9.7
15.3
15.3
15.3
15.3
15.3
15.3
10.76
10.76
15.3
8.8
9.7
17.56
k1 RT HYD
L/g-hr 1/s
0.1416346
1.231922
92.02836
0.3368081
2.669322
0.1301963
0.2272812
3.1
0.1783568
0.6009024
22.15714
0.1416346
2.940198
0.0624
5.185646
92.02824
0.3617182
4.7
1.05
0.107124
0.139343
0.185551
0.5114555
0.62
0.8115939
0.7836799
1 .634334
0.392715
1.913115
8.988508
0.8558166
3.534435
0.1560334
423.0154
14.33503
3.015383
18.1279
0.1385906
0.4764868
0.3213955
0.9
0.0686
0.2461183
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
-------
TABLE C-2. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS,
COMPOUND NAME
ETHYL ACRYLATE
ETHYL CARBAMATE
ETHYL ETHER
ETHYL MORPHOLINE, ethyl diethylene oxime
ETHYL TOLUENE, 4
ETHYL VINYL ETHER
ETHYLACETATE
ETHYLAMINE
ETHYLBENZENE
ETHYLENE DIAMINE
ETHYLENE DIBROMIDE
ETHYLENE GLYCOL
ETHYLENE GLYCOL DIMETHYL ETHER
ETHYLENE GLYCOL MONOBUTYL ETHER
A
7.964455
7.364662
6.92
6.894822
6.875279
6.820354
7.101
7.05413
6.975
7.337209
7.34485
8.0908
7.232956
8.19856
B
1897.011
1737.454
1064.07
1636.683
1479.869
1047.859
1244.95
987.31
1424.255
1521.051
1675.301
2088.9
1377.29
2008.895
C
273.16
204.4896
228.8
199.7396
208.4796
232.9896
217.88
220
213.21
215.3196
244.82
203.5
232.43
206.7696
LN(OW)
1.5014
-0.447976
1 .639246
1 .924971
3.772945
1 .645062
0.9244851
0.17986
3.15
-0.269471
1.6
-0.914444
1.134076
1 .554456
Km ax
mg/g-hr
17.56
10.76
17.56
15.3
15.3
15.3
17.58
9.7
6.8
9.7
10.76
17.56
15.3
15.3
k1 RT HYD
L/g-hr 1/s
0.5021918
0.0912
0.5665672
0.7274908
3.664899
0.5694574
0.303138
0.1580087
2.1
0.1066435
0.5474415
0.0606
0.3641545
0.5260546
0
0
0
0
0
0
0
0
0
0
0
0
0
0
ETHYLENE GLYCOL MONOBUTYL ETHER ACETATE
7.2159
1659.2
191.339
1 .48746
15.3
0.4961033
0
ETHYLENE GLYCOL MONOETHYL ETHER ACETATE
7.377117
1641.85
209.2748
0.619354
15.3
0.2321078
0
ETHYLENE GLYCOL MONOETHYL ETHER Cellosol
ETHYLENE GLYCOL MONOMETHYL ETHER
ETHYLENE GLYCOL MONOMETHYL ETHER
ETHYLENE GLYCOL MONOPHENYL ETHER
ETHYLENE GLYCOL MONOPROPYL ETHER
ETHYLENE OXIDE
ETHYLENE THIOUREA
ETHYLHEXYLACRYLATE.2-
ETHYLPHENOL.3-
ETHYL(2) HEXANOL
FLUORANTHENE
FORMALDEHYDE
FORMAMIDE
FORMIC ACID
FREON 11, fluorotrichloromethane
FREON12 DICHLORODIFLUOROMETHANE
FREON 12, dichlorodifluoromethane
FUMARICACID
8.177948
8.171905
ACETATE
7.52829
8.235456
8.176798
7.128
7.211493
6.884682
7.468
9.114945
6.373
7.195
2.602059
7.581
6.816687
3.698946
6.795418
-1.118697
1846.634
1799.523
1641.851
2309.119
1837.652
1054.54
1731.503
1554.707
1856
2850.004
1756
970.6
19.62512
1699.2
1017.234
0
806.5787
0
213.6096
215.6046
209.2747
194.2296
213.9896
237.76
201 .5974
204.2996
187
273.16
118
244.1
264.1977
260.7
234.7566
0
247.0498
0
0.2502244
0.2522626
0.185323
1 .889739
1.120402
-0.3
-0.641652
3.773185
2.4
2.643529
5.33
1.94
-0.553912
-0.924103
2.576381
2.16
2.402047
-0.445571
15.3
15.3
15.3
15.3
15.3
4.2
9.7
15.3
15.3
17.56
31.1
5
17.56
17.56
15.3
10.76
15.3
17.56
0.1680427
0.1683427
0.1587658
0.705406
0.3598234
0.91
0.077
3.665669
1.102401
1 .36421 1
1.5
0.25
0.0831
0.0601
1 .286367
0.8935916
1.104378
0.0914
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
C-24
-------
TABLE C-2. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS,
COMPOUND NAME
FURAN
FURFURAL
Generic Organic material
GLUTARICACID
GLYCERIN (GLYCEROL)
GLYOXAL
GLYPHOSATE
GUTHION
GYLCIDOL
HEPTACHLOR
HEPTANE ISO
HEPTANE(-n)
HEXACHLOROBENZENE
HEXACHLOROBUTADIENE
HEXACHLOROCYCLOHEXANE (GAMMA
HEXACHLOROCYCLOPENTADIENE
HEXACHLOROETHANE
HEXAFLUOROACETONE
HEXAMETHYLENE 1,6 DIISOCYANATE
HEXAMETHYLPHOSPHORAMIDE
HEXANE(-n)
HEXANOICACID
HEXANOL-1
HEXEN-2-ONE 5
HYDRAZINE
HYDROGEN SULFIDE
HYDROQUINONE
HYDROXYACETIC ACID
INDENO(1 ,2,3-cd)-PYRENE
ISOBUTYLENE
ISODECANOL
ISOPHORONE
LINDANE hexachlorocyclohexane
MALEICACID
MALEIC ANHYDRIDE
METHACRYLICACID
METHANOL
METHOMYL
METHOXYCHLOR
METHYL 1-PENTENE2
METHYL ACETATE
METHYL ACRYLATE
METHYL AMINE
A
6.975
6.575
6.859312
6.947832
6.165
8.130124
8.207067
6.94768
9.224102
8.961409
6.8391
6.89677
9.55388
7.485238
ISOMER)
6.966292
8.41529
7.2284
7.105094
6.914242
6.90523
6.876
9.477544
7.86
7.178988
8.683764
6.99392
9.778832
3.82111
9.813007
6.805457
8.22165
7.962584
6.915866
7.956923
12.6834
1 1 .26555
7.897
20.8751
6.951463
6.83529
7.065
6.838646
7.3369
B
1060.87
1198.7
1354.71
2093.118
1036
1472.782
2076.833
2093.121
2760.353
3722.308
1335.41
1264.9
3248.572
1956.415
2268.189
2834.547
1347.978
665.0909
1798.169
1722.428
1171.17
3158.919
1761.26
1474.01
2243.776
768.13
3857.106
459.9652
5609.269
917.0536
2195.003
2480.726
1811.968
2706.393
4638.2
3657.094
1474.08
7506.508
2128.043
1121.3
1157.63
1193.779
1011.5
C
227.74
162.8
215.5096
174.6596
28
229.5696
203.9194
174.6594
273.16
273.16
237.2
216.54
203.07
215.333
165.1596
273.16
132.91
258.4496
190.8096
194.9896
224.41
273.16
196.66
214.9396
273.16
249.09
273.16
273.16
273.16
240.5706
198.9796
273.16
190.0496
273.16
273.16
273.16
229.13
273.16
172.7596
229.687
219.73
224.6296
233.3
LN(OW)
1 .853527
1.578185
0
0.0219418
-1 .432496
-1.11614
-0.494987
2.708063
-0.381763
5.05
3.162377
3.161653
5.47
3.74
0.936952
3.99
3.456234
4.214555
1 .27331
-0.000432
2.727623
1.591822
1 .774724
1 .344229
-1.37
1.173106
0.59
-1 .097389
7.66
1 .796808
2.920139
1.7
0.936952
-0.445571
0
0.6610121
-0.7
-0.301311
6.571831
2.664926
-0.09
0.9331476
-0.530324
Km ax
mg/g-hr
17.56
17.56
19
17.56
17.56
15.3
15.3
15.3
17.56
10.76
15.3
15.3
10.76
10.76
10.76
10.76
10.76
15.3
9.7
9.7
15.3
17.56
17.56
17.56
9.7
10.76
17.56
17.56
9.7
15.3
15.3
15.3
10.76
17.56
4.08
17.56
18
15.3
17.56
15.3
19.87
17.56
9.7
k1 RT HYD
L/g-hr 1/s
0.6834056
0.5370913
1.1
0.1376169
0.0385
0.0508
0.0875
1 .443459
0.0966636
1 1 .20331
2.148051
2.146692
16.17883
3.560762
0.3064629
4.431426
2.777865
5.393552
0.4113342
0.1349489
1 .468376
0.5435379
0.6378716
0.4376675
0.0407
0.3768052
0.2262222
0.0516
109.9392
0.6503171
1 .737775
0.5975001
0.3064629
0.0914
1
0.2407243
0.2
0.1037134
42.42731
1 .389991
0.1247767
4.3
0.0848
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
C-25
-------
TABLE C-2. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS,
COMPOUND NAME
METHYL AZIRIDINE 2
METHYL BENZYL ALCOHOL 4
METHYL CHLORIDE
METHYL CHOLANTHRENE 3
METHYL CYCLOHEXANE
METHYL ETHER dimethyl ether
METHYL ETHYL KETONE, 2 butanone
METHYL FORMATE
METHYL HYDRAZINE
METHYL IODIDE
METHYL ISOBUTYL KETONE
METHYL ISOCYANATE
METHYL ISOPROPYL KETONE
METHYL MERCAPTAN
METHYL METHACRYLATE
METHYL MORPHOLINE
METHYL NAPTHALENE(I-)
METHYL NAPTHALENE(2-)
METHYL PARATHION
METHYLENE CHLORIDE, dichloromethane
METHYLENE DIPHENYL DIISOCYANATE
METHYLENE DIPHENYLAMINE (MDA)
METHYLENEDIANILINE 4,4
METHYLENE-BIS(2-CHLOROANILINE),4,4'-
METHYLSTYRENE (-4)
METHYL-TERTIARY-BUTYL ETHER
MONOMETHYL FORMANIDE
MORPHOLINE
NABAM
NAPHTHALENE
NAPHTHOL.alpha-
NAPHTHOL.beta-
NAPHTHYLAMINE.alpha-
NAPHTHYLAMINE.beta-
NEOPENTYL GLYCOL
NITROmXYLENE, 2
NITROANILINE P
NITROBENZENE
NITROBIPHENYL.4-
A
0
9.069858
7.093
-3.119754
6.823
3.59986
7.11243
3.027
6.5762
7.657383
6.672
6.822807
5.483397
6.808206
6.5168
7.173846
7.03592
7.0685
-5.008742
6.96841 1
6.942754
6.946537
7.501696
7.513116
7.0112
6.85249
7.204987
7.71813
7.303247
7.3729
7.28421
7.347
7.49977
7.502136
8.219421
6.901502
9.5595
7.115
6.947679
B
0
3021 .253
948.58
0
1270.76
0
1305.006
3.02
1007.5
1507.3
1168.4
1067.989
955.5712
942.8276
1052.176
1433.993
1826.948
1840.264
0
1074.291
2048.169
2082.649
2247.655
2350.365
1535.1
1103.737
1679.083
1745.8
2548.42
1968.36
2077.56
2135
2230.587
2251 .572
2176.788
1691.519
4039.73
1746.6
2093.118
C
0
273.16
249.34
0
221 .42
0
229.2666
-11.9
181.4
273.16
191.9
231 .8306
273.16
239.0696
188.37
217.0296
195
198.4
0
222.995
177.1085
175.2296
181.2885
176.3696
200.7
222.72
204.2996
235
160.2196
222.61
184
183
182.1076
181.1006
199.7396
196.6996
273.16
201.8
174.6596
LN(OW)
0.6105353
2.385993
1.92
7.430058
2.963777
0.6286275
0.28
0.0940167
-0.478592
1.69
1.38
2.84454
0.9050759
1.173106
1 .295875
1 .357002
4.132105
4.132105
2.04
1.25
3.308513
3.379612
2.633374
4.900831
3.463532
1.90147
-0.678949
-1.08
-1.646165
3.37
3.076284
2.84
2.22
2.25
0.3583827
3.067034
1.182549
1.84
3.945698
Km ax
mg/g-hr
9.7
17.56
10.76
31.1
15.3
17.56
2
17.56
9.7
0
0.74
9.7
15.3
15.3
17.56
15.3
24.03
31.1
15.3
22
9.7
15.3
9.7
9.7
31.1
17.56
15.3
9.7
15.3
42.47
17.56
17.56
9.7
9.7
17.56
15.3
9.7
11
19
k1 RT HYD
L/g-hr 1/s
0.2303237
1 .088972
0.7243333
89.90299
1.805412
0.2339989
0.2
0.1465752
0.0888
0.5922948
0.4515826
1 .626545
0.2980333
0.3768052
4.3
0.4425861
5.018151
5.018151
0.8045235
0.4
2.441051
2.597734
1.352142
9.832447
2.795661
0.712684
0.0745
0.0524
0.0319
1
1.992183
1 .620096
0.9417578
0.9668059
0.1847226
1.976124
0.3799313
2.3
4.262934
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
C-26
-------
TABLE C-2. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS,
COMPOUND NAME
NITROGLYCERIN
NITROMETHANE
NITROPHENOL.2-
NITROPHENOL.4-
NITROPROPANE2
NITROSODIMETHYLAMINE N
NITROSOMORPHOLINE
NITROSO-N-METHYLUREA N
NITROTOLUENE (-p)
NITROTOLUENE, m
NITROTOLUENE, o
NITROTOLUENE, o
NONANOL, n
OCTANE
OCTANOL 1
OCTANOL 2
OCTANOL 3
OCTANOL 4
OIL
OXALIC ACID
PARALDEHYDE
PARATHION
PCB'S (Aroclors)
PENTACHLOROBENZENE
PENTACHLOROETHANE
PENTACHLORONITROBENZENE
PENTACHLOROPHENOL
PENTADIENE 1,2
PENTAERYTHRITOL
PHENOL
PHENYLCYCLOHEXANONE 4
PHENYLENE DIAMINE(-m)
PHENYLENE DIAMINE(-o)
PHENYLENE DIAMINE(-p)
PHENYLPHENOLP
PHOSGENE (decomposes)
PHOSPHINE
PHTHALICACID
PHTHALIC ANHYDRIDE
PHTHALIMIDE
PINENE(alpha-)
PIPERAZINE
PROPANE
PROPANE SULTONE, 1,3-
PROPANOICACID
A
9.63631 1
7.28166
11.86137
10.95784
7.271697
7.5239
7.183389
7.362276
6.9948
6.903989
6.901917
6.901917
8.223319
6.918
12.0701
6.3888
5.2215
5.7396
0
16.86465
7.286656
9.10235
-6.785018
9.054626
6.74
6.954095
8.119069
6.9182
8.229961
7.133
-0.708
7.984773
9.250531
9.309744
8.6575
6.842
3.301009
7.879505
8.022
8.300783
6.8525
8.57357
6.80338
7.256331
6.866276
B
3601 .771
1446.94
4392.938
4459.823
1530.544
1974.062
1813.598
1718.378
1720.39
1712.12
1694.951
1694.951
2208.672
1351.99
4506.8
1060.4
560.3
760.5
0
5987.326
1754.238
4032.563
-5599.125
3396.595
1378
2152.518
3054.752
1104.991
2263.417
1516.79
106.4
2843.732
3383.349
3472.661
3022.8
941 .25
0
1728.354
2868.5
1857.213
1446.4
2386.184
804
2110.189
1409.059
C
273.16
227.6
273.16
273.16
228.62
273.16
196.5096
205.4396
184.9
195.5596
196.5096
196.5096
198.4096
209.15
319.9
122.5
64.7
89.5
0
273.16
273.16
273.16
273.16
273.16
197
171.4296
273.16
228.85
196.1296
174.95
-146.6
273.16
273.16
273.16
216.1
230
0
273.16
273.16
273.16
208
273.16
247.04
182.2596
212.4506
LN(OW)
0.6459801
0.0639644
1.79
1.91
1 .283951
-0.47
-0.44
0.185811
2.45499
2.45499
2.45499
2.45499
3.076832
3.595688
2.642804
2.643529
2.643529
2.643529
5.331716
-1 .280351
2.303315
3.055472
7.308553
5.96656
2.718297
5.645749
5.01
2.068878
-1 .547323
1.46
3.692797
0.8427868
0.8427868
0.8427868
3.822525
1.179
2.344077
0.8225744
-0.62
0.8427868
3.937396
0.4268957
1.425412
-2.82
0.2895945
Km ax
mg/g-hr
9.7
9.7
9.7
9.7
9.7
9.7
9.7
9.7
9.7
15.3
15.3
15.3
15.3
15.3
17.56
17.56
17.56
17.56
15.3
17.56
17.56
15.3
19
10.76
10.76
10.76
130
15.3
15.3
97
17.56
9.7
9.7
9.7
17.56
15.3
15.3
17.56
17.56
9.7
15.3
15.3
15.3
15.3
17.56
k1 RT HYD
L/g-hr 1/s
0.2375789
0.1427711
0.6464548
0.7180231
0.4151816
0.0894
0.0918
0.1588336
1.156741
1.156741
1.156741
1.156741
1.993139
3.138369
1 .363345
1 .36421 1
1 .36421 1
1 .36421 1
14.33503
0.044
1.012976
1 .956233
80.83541
24.98281
1 .456443
18.86828
3.4
0.8251115
0.0348
13
3.41669
0.2822246
0.2822246
0.2822246
3.827388
0.3787535
1 .049757
0.2772771
0.0784
0.2822246
4.232081
0.196135
0.4698876
0.0114
0.1739324
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0.000555
0
0
0
0
0
0
C-27
-------
TABLE C-2. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS,
COMPOUND NAME
PROPANOL
PROPANOL ISO
PROPENE
PROPIOLACTONE b
PROPIONALDEHYDE
PROPIONICACID
PROPORUR (Baygon)
PROPYL ACETATE ISO
PROPYLAMINE ISO
PROPYL ETHER ISO
PROPYLENE
PROPYLENE CHLOROHYDRIN
PROPYLENE GLYCOL
PROPYLENE OXIDE
PROPYLENIMINE 1,2 2 methyl aziridine
PROPYL(-n) ACETATE
PROPYL(-n) BENZENE
PROPYN-1-OL 2(PROPARLGYL)
PYRIDINE
QUINOLINE
QUINONE
RESORCINOL
SILVEX
SODIUM FORMATE
STYRENE
STYRENE OXIDE
SUCCINICACID
TAMARON (METHAMIDIPHOS)
TEREPHTHALICACID
TERPINEOL, ALPHA
TETRACHLOROBENZENE(1 ,2,3,4)
TETRACHLOROBENZENE(1 ,2,3,5)
TETRACHLOROBENZENE(1 ,2,4,5)
TETRACHLORODIBENZO-p-DIOXIN(2,3,7,8)
TETRACHLOROETHANE(1 ,1 ,2,2)
TETRACHLOROETHENE
TETRACHLOROPHENOL(2,3,4,6)
TETRACHLOROPHENOL(2,3,5,6)
TETRAETHYL LEAD
A
7.84767
8.117
6.794065
7.982566
7.2088
6.403
7.236959
6.843416
1 1 .00768
6.834991
6.7781 1
8.176797
8.208243
7.0671
7.1502
7.016
6.95142
8.993223
7.041
7.3838
7.148885
6.924308
0
-3.119232
6.945357
6.888923
-3.118326
7.309743
6.930848
8.180823
8.872679
8.821956
8.821956
6.977444
6.893793
6.976
9.168764
-1 .999987
8.56581
B
1499.21
1580.92
787.2845
2221.617
1235.771
950.2
1942.436
1231.175
2488.127
1165.131
770.85
1837.652
2085.9
1133.267
1249.718
1282.28
1491.297
2364.043
1373.8
2300.405
1239.346
1884.547
0
0
1437.432
1588.822
0
1307.183
1941.731
1869.102
3158.681
3084.413
3084.413
2377.268
1354.506
1386.92
2748.848
0
2689.922
C
204.64
219.61
248.1896
273.16
237.47
130.3
190.7885
222.5016
273.16
226.2636
245.51
213.9896
203.5396
236.1054
226.7196
208.6
207.14
273.16
214.98
273.16
227.2685
186.0596
0
0
208.38
202.3996
0
226.1496
182.9246
212.6596
273.16
273.16
273.16
159.2696
192.43
217.53
273.16
0
273.16
LN(OW)
0.4913113
-0.16
1.43124
0.4909006
0.691664
0.2895945
2.322595
1 .359286
0.894336
2.653113
1.43124
-0.003456
-0.479632
0.235614
0.6105354
1 .358564
3.594934
0.2602102
0.65
2.03
1.016473
0.8
4.072398
-0.43941 1
3.16
2.644778
-0.412145
0.5316042
0.8225744
3.182377
5.206064
5.206064
4.51
6.64
2.56
2.6
4.762268
4.762268
2.392857
Km ax
mg/g-hr
17.56
15
15.3
17.56
17.56
17.56
9.7
17.56
9.7
17.56
15.3
15.3
17.56
17.56
9.7
17.56
31.1
17.56
35.03
31.1
9.7
17.56
15.3
0
31.1
17.56
17.56
15.3
15.3
15.3
10.76
10.76
10.76
10.76
6.2
6.2
10.76
10.76
0
k1 RT HYD
L/g-hr 1/s
0.2075073
0.075
0.4722896
0.2074327
0.2472678
0.1739324
1 .030209
0.4434716
0.2952457
1 .375699
0.4722896
0.1345924
0.0887
0.1659081
0.2303237
0.4431913
3.1363
0.1695174
0.238416
0.7975147
0.3285457
0.271854
4.76272
0.0919
0.11
1 .365701
0.0941278
0.2149536
0.2772771
2.185972
12.84254
12.84254
6.984653
45.03498
0.68
0.68
8.709778
8.709778
1 .095533
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0.42
0
0
0
0.35
0
0
TETRAETHYLDITHIOPYROPHOSPHATE
11.84911
5454.172 273.16
1.656811
15.3
0.5753421
C-28
-------
TABLE C-2. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS,
COMPOUND NAME
TETRAETHYLENE PENTAMINE
TETRAFLUOROMETHANE
TETRAHYDROFURAN
TETRALIN
TETRANITROMETHANE
THIOUREA
THIOUREA,1-(o-CHLOROPHENYL)-
TOLUENE
TOLUENE DIAMINE(2,4)
TOLUENE DIISOCYANATE(2,4)
TOLUENEDIAMINE(2,6)
TOLUENEDIAMINE(3,4)
TOLUENESULFONYL CHLORIDE
TOLUIC ACID (para-)
TOLUIC ALDEHYDE
TOLUIDINEm
TOLUIDINE P
TOLUIDINE(-O)
TOXAPHENE
TRIBUTYL PHOSPHOROTRITHIOATE SSS
TRI BUTYL TIN ACETATE
TRIBUTYLPHOSPHATE
TRICHLOROBENZENE 1,2,3
TRICHLOROBENZENE 1,2,4
TRICHLOROBENZENE 1,3,5
TRICHLOROETHANE 1 ,1 ,1 methyl chloroform
TRICHLOROETHANE 1 ,1 ,2
TRICHLOROETHYLENE
TRICHLOROFLUOROMETHANE
TRICHLOROPHENOL 2,4,5
TRICHLOROPHENOL 2,4,6
TRICHLOROPROPANE 1,1,1
TRICHLOROPROPANE(1 ,1 ,2)
TRICHLOROPROPANE(1 ,2,2)
TRICHLOROPROPANE(1 ,2,3)
TRICHLORO-1 ,2,2-TRIFLUOROETHANE,1 ,1 ,2-
TRIETHANOLAMINE
TRIETHYLAMINE
TRIETHYLENE GLYCOL DIMETHYL ETHER
TRIFLUOROETHANE(1,1,1)
A
7.386298
6.882466
6.995
7.07055
8.350471
4.247091
11.46513
6.954
7.492203
6.912613
0
7.483529
6.930257
10.50106
9.005018
7.454377
7.77229
7.19724
5.707643
6.901398
6.91708
6.0038
6.900254
7.7056
8.598745
6.827401
7.1921
6.518
6.884
8.577468
9.696309
1 1 .56902
8.218203
11.14517
6.903
7.637469
8.287391
6.958867
6.898169
3.965647
B
1913.731
386.9427
1202.29
1741.3
2183.276
621 .8834
5117.637
1344.8
2164.173
1784.378
0
2089.257
1936.527
4177.125
2922.234
1846.317
2317.386
1682.94
1857.691
2987.955
1822.323
1755.628
1681.226
2242.67
2754.108
1147.14
1480.319
1018.6
1043.004
2974.575
3528.119
3302.916
2205.203
3282.281
788.2
1526.226
2762.499
1271.96
1664.082
0
C
195.7494
272.6996
226.25
208.26
273.16
273.16
273.16
219.48
185.2996
191.5696
0
188.9096
183.2096
273.16
273.16
200.6896
273.16
191.138
273.16
273.16
189.4796
273.16
197.2696
252.836
273.16
218.5387
229.0943
192.7
236.88
273.16
273.16
273.16
273.16
273.16
243.23
273.16
175.5336
223.262
198.2196
0
LN(OW)
0.1216321
2.89165
1 .440628
3.763707
0.4822038
0.8612069
2.166816
2.69
1 .454882
6.803719
1 .454882
1 .454882
2.188507
2.105137
2.285832
2.115259
1.39
2.115272
3.3
4.803134
5.463342
5.430991
4.445513
3.98
4.445513
2.49
2.17
2.29
2.53
4.001704
3.69
0.2066454
1 .654597
1.426414
0.2066454
3.456217
-1.75
2.359257
2.429506
2.703412
Km ax
mg/g-hr
15.3
15.3
17.56
15.3
9.7
9.7
9.7
73.48
9.7
9.7
9.7
9.7
15.3
17.56
17.56
15.3
31.1
31.1
15.3
17.56
15.3
17.56
10.76
10.76
10.76
3.5
3.5
3.9
10.76
15.3
15.3
10.76
10.76
10.76
10.76
10.76
9.7
9.7
15.3
10.76
k1 RT HYD
L/g-hr 1/s
0.15016
1 .694992
0.4761852
3.635396
0.2058602
0.2868102
0.898937
2.4
0.4821615
51.97128
0.4821615
0.4821615
0.9161608
0.8517082
0.9975978
0.8592852
0.4555512
0.8592948
2.422935
9.026857
16.08485
15.63593
6.601452
4.392822
6.601452
0.74
0.74
0.88
1 .235208
4.477043
0.26
0.1617557
0.5742285
0.4702994
0.1617557
2.777823
0.0291
1 .063793
1.131232
1 .437597
0
0
0.604
0
0
0.46
0
0
0.07
0
0
0
0
0.81
0
0
0
0
0
0
0
0
0
0.55
0
0.39
0
0.12
0
0
0
0.07
0.009
0
0
0
0
0
0
0
C-29
-------
TABLE C-2. COMPOUND PROPERTIES FOR ORGANIC COMPOUNDS,
COMPOUND NAME
TRIFLURALIN
TRIISOBUTYLENE
TRIISOPROPYLAMINE
TRIMETHYL BENZENE, 123
TRIMETHYLBENZENE (1,3,5)
TRIMETHYLPENTANE 2,2,4
TRINITROTOLUENE(2,4,6)
TRIPROPYLENE GLYCOL
UREA
URETHANE
VINYL ACETATE
VINYL ACETYLENE
VINYL BROMIDE
VINYL CHLORIDE
VINYLIDENE CHLORIDE see 1,1dichloroethen
WARFARIN
XYLENE
XYLENE(-m)
XYLENE(-o)
XYLENE(-p)
XYLENOL(3,4)
XYLIDINE dimethylaniline
A
6.937691
6.894822
6.900255
6.880847
7.07436
6.797857
-1 .337234
8.236003
8.555151
7.421
7.21
6.790945
6.814487
6.9907
6.972
9.979672
7.940135
7.009
6.998
6.99
7.07919
7.461541
B
2002.526
1636.683
1681.23
1524.055
1569.22
1249.485
0
2313.693
2304.696
1758.21
1296.13
726.4801
998.5032
969.0518
1099.4
4466.33
2090.317
1462.266
1474.679
1453.43
1621.45
1904.87
C
179.5996
199.7396
197.2694
206.0096
209.58
219.7695
0
194.0396
273.16
205
226.66
251 .7996
235.8396
250.5856
237.2
273.16
273.16
215.11
213.69
215.31
159.26
197.8396
LN(OW)
5.355146
5.303749
4.965421
3.999725
3.999725
3.567022
0
2.126133
-0.469833
0.4310978
0.930263
1 .652999
1 .066461
0.06
0
5.387373
3.15
3.2
2.95
3.15
0.1
2.727331
Km ax
mg/g-hr
9.7
15.3
15.3
15.3
31.1
15.3
4.4
15.3
9.7
9.7
17.56
15.3
10.76
10.76
10.76
15.3
40.8
31.1
40.79
31.1
17.56
15.3
k1 RT HYD
L/g-hr 1/s
14.63195
13.98851
10.40413
4.469296
4.469296
3.06063
0.45
0.8675004
0.0894
0.1968575
0.3046745
0.5734261
0.343235
0.1422767
0.9039771
15.05042
1.8
2.219942
1.8
2.124915
0.1473445
1.468
0
0
0
0
0
0
0
0
0.15
0.51
0
0
0
0.88
0
0
0.091
0.07
0.12
0
0
0
C-30
-------
TECHNICAL REPORT DATA
(Please read Instructions on reverse before completing)
1. REPORT NO.
EPA-453/R-94-080A
3. RECIPIENT'S ACCESSION NO.
4. TITLE AND SUBTITLE
Air Emissions Models for Waste and Wastewater
5. REPORT DATE
November 1994
6. PERFORMING ORGANIZATION CODE
7. AUTHOR(S)
8. PERFORMING ORGANIZATION REPORT NO.
9. PERFORMING ORGANIZATION NAME AND ADDRESS
Emission Standards Division (MD-13)
Office of Air Quality Planning and Standards
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
10. PROGRAM ELEMENT NO.
11. CONTRACT/GRANT NO.
68-D1-0118
12. SPONSORING AGENCY NAME AND ADDRESS
Office of Air Quality Planning and Standards
U.S. Environmental Protection Agency
Research Triangle Park, NC 27711
13. TYPE OF REPORT AND PERIOD COVERED
14. SPONSORING AGENCY CODE
EPA/200/04
15. SUPPLEMENTARY NOTES
16. ABSTRACT
Analytical models are presented for estimating air emissions from waste and wastewater collection and
treatment units. Air emission models have been developed for collection and treatment units, aerated and
non-aerated surface impoundments, land treatment facilities, landfills, and waste piles. Emission model
predictions are compared to available field data.
This report is the documentation for the computer models ChemdatS (EPA-453/C-94-080B) and WaterS
(EPA-453/C-94-080C).
Appendices include a list of physical-chemical properties for approximately 500 compounds and a
comprehensive source list of pertinent literature in addition to that cited in the report.
17.
KEY WORDS AND DOCUMENT ANALYSIS
DESCRIPTORS
b. IDENTIFIERS/OPEN ENDED TERMS
c. COSATT Field/Group
Emission Models
Air Emission Models
Air Pollution
Air Emissions from Waste/Wastewater
Waste
Wastewater
Air Pollution control
18. DISTRIBUTION STATEMENT
Release Unlimited, available from the OAQPS
TTN bulletin board and NTIS
19. SECURITY CLASS (Report)
Unclassified
20. SECURITY CLASS (Page)
Unclassified
21. NO. OF PAGES
22. PRICE
C-31
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EPA Form 2220-1 (Rev. 4-77)PREVIOUS EDITION IS OBSOLETE
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