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POLYCYCLIC AROMATIC
HYDROCARBONS:
SOURCES, FATE AND LEVELS IN AIR,
WATER, SOIL, SEDIMENTS, SLUDGE AND
FOOD IN ONTARIO
MAY 1992
Environment
Environnement
Ontario
ISBN 0-7729-9281-9
POLYCYCLIC AROMATIC HYDROCARBONS:
SOURCES, FATE AND LEVELS IN AIR, WATER, SOIL, SEDIMENTS,
SLUDGE AND FOOD IN ONTARIO
Report Prepared For:
Hazardous Contaminants Branch
Report Prepared By:
Concord Scientific Corporation
and
Beak Consultants
MAY 1992
o
PRINTED ON
RECYCLED PAPER
IMPRIMESUR
OU PAPIER RECYCLE
Cette publication technique
n'est disponible qu'en anglais.
Copyright: Queen's Printer for Ontario, 1992
This publication may be reproduced for non-commercial purposes
with appropriate attribution.
PIBS 1938
DISCLAIMER
This report has been approved for publication by the Hazardous Contaminants Branch of
the Ontario Nlinistry of the Environment. Approval does not signify that the contents necessarily
reflect the views and policies of the Ontario Ministry of the Environment, nor does the Ministry
warrant the information contained herein. Mention of trade names or commercial products does
not constitute endorsement or recommendation for use.
EXECLTTIVE SUMMARY
This report covers a review of literature information on polycyclic aromatic
hydrocarbons (PAH), primarily in Ontario. Specifically, this report deals
with chemical and physical properties of PAH as well as the sources,
inputs and fate of PAH to the environment.
Good, reliable data for PAH physical properties are scarce. For example,
recorded vapour pressure data frequently range over several orders of
magnitude and solubility data in solvents other than water are often non-
existent. Similarily, reliable, chemical reactivity data are equally scarce.
The report discusses the sources and input to the environment by
compiling the emission factors and profiles for point and non-point
sources in Ontario. Based on the atmospheric emissions inventory for
these compounds, it is estimated that the major PAH sources to the
atmosphere are:
o Gasoline and diesel fuelled vehicles
o Forest fires
o Woodburning stoves and fireplaces
These three sources contribute approximately 97% of the estimated 260
MT/y of total PAH atmospheric emissions in Ontario. The remaining
contributions are primarily from industrial operations (including coke
manufacturing, coal-fired thermal generation stations and incinerators)
and residential, commercial and institutional heating. The estiamted total
PAH emission is believed to be uncertain by a factor of about two,
because of data limitations.
The emission factors for PAH that are major contributors to the total PAH
from these sources are summarized in Table ES-1. Phenanthrene and
its methyl derivatives, pyrene, anthracene and its methyl derivatives and
benzo[a]pyrene are the most ubiquitous based on the number of entries
in the Table. On the other hand, benzo[k]f!uoranthene is only recorded
once; so too, are perylene, fluorene, dibenz[a]anthracene and its methyl
derivatives and benzo[e]pyrene. Whereas, it was noted that
benzo[k]fluoranthene was produced at a relatively high concentration in
coke oven emissions, the information obtained during the study provided
no guide regarding potential PAH source markers. Indeed, this review
discoverd no recorded unambiguous procedure for identifying chemical
markers that could be clearly attributed to a specific source.
The report also discusses sources and inputs to water and soil, and it is
estimated that nearly 75% of the PAH loadings to soil and water arise
from rainfall; the remaining inputs are from industrial discharges, including
water pollution control plants, and there is also a significant contribution
to water arising from urban run off (10-15%).
PAH that are most persistent and frequently detected at relatively high
concentration in water include benz[a]anthracene, benzo[k]fluoranthene
and pyrene in treated discharge from water treatment plants, based on
an extensive study carried out on 37 high flow rate plants. Similarly,
sludges produced from these facilities also showed higher levels of
acenaphthylene, phenanthrene, and pyrene, compared to the other PAH
detected.
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IV
Based on the rainfall flux data, those PAH present at relatively high
concentration compared to other PAH comprising the total are:
phenanthrene, fluorene, benzo[b]fluorene, benzo[ghi]perylene and
indeno[1,2,3,-cd]pyrene. Other studies indicate that pyrene and
benzo[a]pyrene are probably the two PAH that constitute the greatest
fraction of the total PAH loading to the Great Lakes.
It is inferred from the available data that PAH levels in soils vary widely,
with urban soils containing from 100 to 1 ,000 ug/kg benzo[a]pyrene; the
higher concentrations are associated with locations near highways,
airports, rail stations, and heavy industry. Similarly, it was inferred that
the levels of phenanthrene, fluoranthene, and pyrene were significantly
higher in urban run off than other PAH.
Information collected from the literature search indicates that nitration and
oxidation reactions occur with PAH, which alter the source PAH to a
substituted or a ring-opened form. The kinetics of these reactions are
influenced by light intensity, the size of the particle to which the PAH is
adsorbed and temperature. Reactions are very much slower in the
aqueous phase, partly because of the hydrophobic nature of these
compounds. For example, it is estimated that the \y2 ^O"" bacterial
oxidation of anthracene in water is about 15 days.
The particle-bound PAH detected in the ambient air show elevated levels
of benzo[ghi]perylene in industrialized towns with high traffic densities,
such as Hamilton, and much lower levels for towns with lower traffic
densities such as Sudbury. As expected, these levels also show a strong
seasonal dependence with significantly higher levels of PAH in the
ambient air during the Fall (October-December) for most locations. Other
studies show phenanthrene, fluoranthene, fluorene and its methyl
derivatives and pyrene are, on average, the commonest PAH in the
ambient air in Toronto.
An attempt was made to determine the exposure level for individuals
resident in Ontario. It is estimated that the PAH exposure, other than
occupational, is 1.5 - 2.1 mg/yr depending on whether the individual is
a smoker or non-smoker. Estimates are summarized in Table ES-2.
The study clearly identifies certain major gaps in terms of available and,
in many cases, reliable information on the sources and inputs of PAH
and their fate in the Ontario environment.
Thermodynamic modelling (so-called fugacity modelling) was used to
assess both the potential for partitioning of PAH into the various
environmental media as well as the sensitivity of such a tool for predicting
the uncertainty in values of the physical properties of PAH. This exercise
identified that much better data on vapour pressures are needed to
reduce the uncertainty of environmental fate modelling.
As an example of the kind of useful information produced by this
modelling exercise, phenanthrene, which has a relatively high vapour
pressure and thus might be expected to reside primarily as a vapour in
the atmosphere, is predicted to be widely distributed among all
environmental compartments. This prediction is consistent with actual
measurements, which show phenanthrene to be ubiquitous in all media.
VI
TABLE ES-2
Estimated Average PAH Exposure
per Individual in Ontario
PAH
INTAKE
(ug/yr)
Food
1299
Drinking Water
36.5
Inhalation
30
Indoors (non-smoking)
70
Indoors (smoking & fireplaces)
700
TOTAL PAH INTAKE
1.5-2.1 mg/yr,
VII
One of the most important conclusions to be drawn from the review is
that it would be beneficial for Canadian (and other) jurisdictions to agree
on a standard minimum set of PAH to monitor in all environmental media,
or compartments, regardless of whether a particular PAH is expected to
be found in a particular medium. The reason for this suggestion is that
only when the detailed distributions of selected PAH are known can it be
concluded that their environmental transformations and fate are
understood. Missing data for some media create significant uncertainty.
The data presented in the report suggest that it is not possible at present
to track mixtures emitted or discharged from a given source
unambiguously from source to receptors over any significant distance.
Near sources of emissions or discharge, that is, within the first few
kilometres, source profiles of PAH mixtures may allow direct source-
receptor analysis. Further away from sources, environmental chemistry
(especially atmospheric chemistry) and physical processes (especially
atmospheric precipitation) alter the characteristics of PAH mixtures so that
measurements at distant receptor sites contain ambiguous information
about source origins. For this reason, the report suggests that air
emission source assessment for regulatory source apportionment would
best be concentrated on chemically stable, particle-bound PAH.
Human and environmental exposure estimates by means of
measurements in the environment, depend of course, upon total PAH
measurements, rather than just particle-bound PAH.
ACKNOWLEDGEMENT
The authors wish to acknowledge Mr. Warren Stiver, doctoral candidate
in the Department of Chemical Engineering and Applied Chemistry,
University of Toronto, for carrying out the environmental fate modelling
described in Chapter 8.
TABLE OF CONTENTS
Page
Acknowledgement
Executive Summary i
1.0 INTRODUCTION 1-1
1.1 Background 1-1
1.2 Technical Program 1-2
1.3 Polycylic Aromatic Hydrocarbon (PAH) 1-3
Priority List
1.4 Addenda 1-5
2.0 PHYSICAL AND CHEMICAL PROPERTIES OF POLYCYCUC 2-1
AROMATIC HYDROCARBONS
2.1 Nomenclature and Structure 2-1
2.2 Chemical and Physical Properties 2-2
2.3 Chemical Reactivity 2-15.
2.4 Summary 2-19
3.0 SOURCES AND INPUTS OF PAH TO THE ATMOSPHERE 3-1
3.1
Introduction
3-1
3.2
Sources and Inputs
3-1
3.3
3.3.1
3.3.1.1
Atomospheric/Terrestrial Sources
and Input Data
Industrial Discharges
Coke Production
3-5
3-6
3-7
3.3.1.2
3.3.1.3
3.3.1.4
3.3.1.5
Metal Processing
Coal Liquefaction
Petroleum Refining (Catalytic Cracking)
Production of Pyrolysis Products (Coal Tar/
Creosote/Anthracene Oil/Coal Tar Pitch/
Carbon Black)
3-15
3-15
3-17
3-19
3.3.2
Power Production Sources
3-25
3.3.2.1
3.3.2.2
Coal Mining (Coal Dust)
Coal-Fired Power Plants
3-25
3-25
3.3.3
Municipal and Hazardous Waste Incinerators
3-31
TABLE OF CONTENTS (cont'd)
Pgge
3.3.4
Transportation
3-36
3.3.4.1
Gasoline and Diesel-Powered Vehicles
3-40
3.3.4.2
Tire Wear
3-44
3.3.4.3
Source Markers (Transportation)
3-44
3.3.5
Residential Heating
3-45
3.3.5.1
Emission Factors
3-45
3.3.5.2
Emissions (Oil and Gas Heating)
3-49
3.3.6
Open Burning of Biomasss
3-50
3.3.6.1
Burning of Treated Wood Waste
3-50
3.3.6.2
Forest Fires
3-52
3.3.6.3
Burning of Agricultural Waste
3-54
3.3.6.4
Source Markers
3-57
3.4
Summary of PAH Emissions to the Atomosphere
3-57
4.0 TERRESTRIAL AND AQUATIC SOURCES AND INPUTS
4.1 Wet/Dry Deposition from the Atomosphere 4-1
4.1.1 Plant Uptake 4-1
4.1.2 Biosynthesis 4-7
4.1.3 Diagenesis - 4-7
4.1.4 Wood Preservation 4-9
4.1.5 Sewage Sludge Disposal 4-13
4.1.6 Disposal of Oil Refinery Sludges 4-20
4.1.7 Coal Gasification Wastes 4-22
4.2 Direct Deposition (Wet/Dry) from the 4-24
Atmosphere to Aquatic Systems
4.2.1 Rainfall as a PAH Source 4-25
4.3 PAH Uptake in Soils near Industrial Operations 4-35
4.4 Municipal Effluents 4-38
4.5 Runoff 4-42
4.6 Oil Spills 4-46
4.6.1 Refinery Losses to Water 4-50
4.6.2 Gasoline and Diesel Oil Loss During Tank Refilling 4-50
4.6.3 Waste Oill from Transportation 4-50
4.6.4 Disposal of Dredging Spoils 4-52
4.7 Leachate from Waste Disposal Sites 4-52
4.8 Treated Wood Structures for Piers 4-52
4.9 Transboundary Movement of Polycyclic Aromatic 4-57
Hydrocarbons
4.10 Summary of PAH Emissions to Soil and Water 4-60
TABLE OF CONTENTS (cont'd)
5.0
Page
ENVIRONMENTAL PROCESSES AND FATE OF PAH
5-1
5.1
Introduction (Air)
5-1
5.1.1
Chemical Reactivity (Atmospheric Aspects)
5-2
5.1.1.1
Particle Lifetime
5-3
5.1.1.2
Reactions with Nitrating Species
5-6
5.1.1.3
Reactions with Ozone
5-11
5.1.1.4
Reactions with Sulphur Oxides
5-13
5.1.1.5
Photolysis of PAH Compounds
5-15
5.1.1.6
Concentrations of Oxy- and Nitro-PAH
in the Atmosphere
5-18
5.1.2
Long Range Transport of PAH
5-22
5.2
Soil
5-24
5.2.1
Degradation in Soils
5-25
5.2.2
Sorption
5-29
5.2.3
Volatilization
5-31
5.2.4
Photolysis
5-31
5.3
Water
5-32
5.3.1
Abiotic Processes
5-32
5.3.1.1
■ Solubility and Sorption
5-32
5.3.1.2
Volatilization
- 5-38
5.3.1.3
Photodegradation
5-42
5.3.2
Biological Processes
5-48
5.3.2.1
Bioavailability and Bioaccummulation
5-48
5.3.2.2
Biodegradation
5-60
5.3.4
Micocosm Studies
5-63
6.0 ENVIRONMENTAL LEVELS AND SOURCE CONTRIBUTIONS 6-1
6.1 Levels in Air, Soil and Water 6-1
6.2 Residue Levels in Ontario Soils 6-10
6.3 Residue Levels in Sediments 6-10
6.4 Residue Levels in Fish and Wildlife 6-13
6.5 PAH in Human Tissues and Fluids 6-21
6.6 Trends in Levels and Source Contributions 6-23
TABLE OF CONTENTS (cont'd)
Page
7.0 HUMAN EXPOSURE LEVELS (DIETARY AND 7-1
UFESTYLE SOURCES)
7.1 PAH in Ontario Food 7-2
7.1.1 Analysis of PAH in Foods 7-4
7.1.2 Total Diet Studies 7-10
7.1.3 Summary - PAH in Food Available in Ontario 7-11
7.2 PAH in Ontario Drinking Water 7-1 1
7.2.1 Summary 7-13
7.3 PAH in Pharmaceuticals, Cosmetics 7-14
7.3.1 Coal Tars 7-14
7.3.2 Soft and Liquid Paraffins, Mineral Oils 7-15
7.3.3 Summary 7-15
7.4 PAH in Indoor Air 7-16
7.4.1 Effects of Combustion Processes (for Heating) 7-16
7.4.2 Effects of Tobacco Smoke . 7-18
7.4.3 Indoor PAH from Miscellaneous Sources 7-20
7.4.4 Summary 7-21
7.5 Estimated Levels of Human Exposures 7-21
8.0 ASSESSMENT OF ENTRY, MOBILITY AND FATE 8-1
OF POLYCYCUC AROMATIC HYDROCARBONS
8.1 Physical Partitioning and Chemical 8-1
Transformation of PAH and PAH Mixtures
8.2 Modelling the Fate of PAH Compounds 8-4
8.3 Results and Discussion 8-9
8.4 Advanced Model Predictions 8-18
8.5 Emission Rate Estimates 8-25
8.5.1 Estimation of Ambient PAH Concentration 8-26
for Hamilton
8.6 Estimate of the LRT Contribution to PAH 8-31
Levels in Ontario
8.7 Regional Estimates for PAH Emissions 8-35
in Ontario
9.0 REFERENCES 9-1
APPENDIX A - Level 3 Fugacity Model Data
APPENDIX B - Maps Illustrating PAH Regional Estimates
LIST OF TABLES
Table ES-1 Summary Table of Emission Factors for
Major PAH from Predominant Sources
in Ontario
Table ES-2 Estimated Average PAH Exposure
per Individual in Ontario
Table 1-1 Polycyclic Aromatic Hydrocarbons in the
Preliminary List by the Ministry
of the Environment
Table 1-2 List of 47 PAH Species
Table 1-3 Polycyclic Aromatic Hydrocarbons on
Revised Ministry of the Environment List
That Were Not Assessed in This Study
Table 2-1 Chemical Structures of Priority
Polycyclic Aromatic Compounds
Table 2-2 Summary Table of Selected Chemical and
Physical Properties of Polycyclic Aromatic
Hydrocarbons
Table 3-1 Estimated PAH Emissions
Table 3-2 Global Estimated Inputs of Benzo[a]Pyrene
and Total PAH to the Aquatic Environment
from Various Sources
Table 3-3 Emission Factors for Coke Production
Table 3-4 Emission Profiles and Factors for
Coke Production
Table 3-5 Summary of PAH Emission Factors for
Battery Stack Emissions During
Coke Pushing Operations in Ontario
Table 3-6 Estimated Values for PAH, B[a]P and Particulate
Matter During Total Coking Operations in Ontario
Table 3-7 Emission Factors for Foundries
Page
Hi
iv
1-4
1-6
1-7
2-4
2-13
3-3
3-4
3-8
3-9
3-10
3-12
3-16
LIST OF TABLES (cont'd)
Page
Table 3-8 Mean Emission Factors for Fluid Bed 3-18
Petroleum Cracking Catalyst Regenerators
(controlled versus uncontrolled)
Table 3-9 PAH Concentration Range in the Plume 3-20
from the Three Refineries Located in Montreal
Table 3-10 Average PAH Emission Factors and Total PAH 3-21
Loading for Fluid Bed (FCC) Catalyst Regenerating
Units for Petroleum Refineries
Table 3-1 1 PAH Emission Rates from a Coal Tar 3-22
Distillation Plant in Hamilton
Table 3-12 PAH Emission Factors from an Oil- Furnace 3-24
Carbon Black Plant
Table 3-13 Average PAH Emission Factors for Differently- 3-27
Fired Coal Power Plants
Table 3-14 PAH Emission Factors for Coal-Fired Thermal - 3-28
Generating Stations
Table 3-15 Ontario Hydro Lambton and Nanticoke Thermal 3-29
Generating Stations Emission Rates of
Polycyclic Aromatic Hydrocarbons
Table 3-16 Estimated PAH, B[a]P and Total Suspended 3-30
Particulate (TSP) Emissions for Ontario Hydro's
Coal-Fired Power Generating Statins
Table 3-17 PAH Emission Data for TRICIL Hazardous 3-32
Waste Incinerator
Table 3-18 NITEP PAH Emission Data and Levels in 3-34
Incinerator Ash
Table 3-19 Group Test Averages for Stack Sampling Results 3-35
Table 3-20 Summary Table of PAH Emission Factors from MSW 3-37
Industrial Liquid Waste and Commercial
Incinerators
Table 3-21 Estimated Annual PAH Atmospheric Emissions 3-38
to the Ontario Environment from Incinerators
LIST OF TABLES (cont'd)
Table 3-22
Table 3-23
Table 3-24
Table 3-25
Table 3-26
Table 3-27
Table 3-28
Table 3-29
Table 3-30
Table 3-31
Table 4-1
Table 4-2
Table 4-3
Table 4-4
Table 4-5
Page
Polycyclic Aromatic Hydrocarbons Detected on 3-41
Gasoline and Diesel Exhaust Particles
Emission Factors for Gasoline and Diesel 3-42
Powered Mobile Sources
Estimates of Yearly PAH Emissions from 3-43
Transportation Sources in Ontario
Typical Emission Factors for Residential 3-47
Heating
Emission Profiles and Factors for PAH from 3-48
Different Fuels in a Conventional Wood Stove
Derived Emission Factors for the Burning of 3-51
Creosote Treated Railway Ties
Emission Factors of Polycyclic Aromatic 3-53
Compounds for Burning Pine Needles
Wildfire and Prescribed Burn Occurrence 3-55
in Ontario, 1984 to 1988
Average Yearly Emission Data for Polycyclic 3-56
Aromatic Hydrocarbons from Wild and Prescribed
Forest Fires in Ontario
Summary Table - Atmospheric PAH Emissions 3-58
Mean PAH Concentrations (mg/MT fresh weight) in 4-3
Lettuce Grown at Various Distances from a Highway
Mean PAH Concentrations (mg/MT fresh weight) in 4-4
Whole Rye Grains Grown at Various Distances from
a Highway
Regression Equations for Plant Uptake of Polynuclear 4-6
Aromatic Hydrocarbons
PAH Content of Creosote, Creosote Sludge and Coal Tar 4-10
Northern Wood Perservers Survey 4-1 1
Soil and Sediment
LIST OF TABLES (cont'd)
Table 4-6
Table 4-7
Table 4-8
Table 4-9
Table 4-10
Table 4-1 1
Table 4-12
Table 4-13
Table 4-14
Table 4-15
Table 4-16
Table 4-17
Table 4-18
Table 4-19
Table 4-20
Northern Wood Perservers Survey
Groundwater
Summary of MISA PAH Data for 37 Ontario Water
Pollution Control Plants
Summary of Ontario Water Pollution Control Plants
Tested, Flow Rates, PAH Concentrations
and Estimated PAH Emission Rates
Page
4-12
4-14
4-15
Loadings of PAH to the Hamilton Sewage Treatment Plant4-18
in Raw Sewage and from the Plant in Treated Effluent
Summary of Canadian and Internation Data 4-19
on PAH in Municipal Sewage Sludge
Concentrations of Base-Neutral Organics 4-23
in Oil Refinery Disposal Sludges
Data for the Correlation Between PAH Deposited 4-27
and the Amount of Rain
Particle-Bound and Dissolved PAH Concentrations in 4-28
Rainfall Collected in a Residential Area of
Portland, Oregon
Annual Rates of Total Deposition of PAH in 4-29
Forest Ecosystems
Total Deposition of PAH to the Great Lakes 4-31
PAH Fluxes to Sediments from 5 Remote Sites 4-32
in the Northeastern United States and 3 Urban Sites
PAH-Rainwater Concentration Data 4-34
Concentrations of Polycyclic Aromatic Hydrocarbons 4-36
in Soil Collected at Different Sites from the
Algoma Steel: Sault Ste. Marie Plant
Ratios of the Concentrations of PAH to Benzo[e]Pyrene 4-37
in Soil Near Roadways
Concentrations of PAH in Surface Soil at
Different Sites in the Vicinity of Birmingham, U.K.
4-39
LIST OF TABLES (cont'd)
Table 4-21 Average Concentrations of PAH Discharged
from the Hamilton WPCP
Table 4-22 Estimated Output of PAH for Hamilton
Waste Treatment Plant
Table 4-23 Concentration of PAH in Palos Verdes Shelf
and Santa Monica Bay Sediments
Table 4-24 Best Estimates of Mean Concentrations of
Polyaromatic Hydrocarbons in Urban Runoff
Table 4-25 Annual Loadings of Polyaromatic Hydrocarbons
in Urban Runoff in the Great Lakes Basin
Table 4-26 New York Storm Sewer Sediment Analysis
Table 4-27 Total Loadings of PAH to Niagara River in Runoff
Table 4-28 Polycyclic Aromatic Hydrocarbons in
Used Motor Oil
Table 4-29 Estimated PAH Losses from Ontario
Refineries in Wastewater
Table 4-30 Organic Compounds Identified in Extracts of
Runoffs from Different Coals
Table 4-31 Estimated Yearly Volumes of Coal Pile Runoff
and Leachate at Coal-Fired Generated
Stations in Ontario
Table 4-32 Estimated PAH Loadings from Ontario
Generation Stations Coal Pile Runoff
Table 4-33 Loadings of PAH to the Niagara River
Table 4-34 Annual Total Loadings of PAH in Water and
Sediment in Urban Runoff for the Niagara River
Table 4-35 Input of Particulate-Associated Contaminants
to Western Lake Superior from the St. Louis
River and Duluth Harbor Area
Page
4-40
4-41
4-43
4-44
4-45
4-47
4-48
4-49
4-51
4-53
4-55
4-56
4-58
4-59
4-61
LIST OF TABLES (cont'd)
Page
Table 4-36 PAH Emission Profile from 24 Point Source 4-62
Discharges to the St. Clair River
Table 4-37 Estimated PAH Loadings to the Soil and Water 4-63
Table 5-1 Summary of Photolytic and Electrophilic 5-4
Table 5-2 Classification of PAH Based on Electrophilic 5-8
Nitration Reactions
Table 5-3 Outdoor Chamber Studies of PAH Reactions 5-14
with O3, NO2, and hv^
Table 5-4 Half-Lives (in Hours) for the Photolysis of PAH 5-19
on Different Substrates Determined in the Rotary
Photoreactor
Table 5-5 Ambient Concentrations of Selected Nitro-PAH 5-21
Table 5-6 A Comparison Half-Lives Calcualted on the Basis 5-28
of First or Zero Order Models from Data Colelcted
by Bulman et al. (1985)
Table 5-7 PAH in Subsurface Water Samples (ng/L) of the 5-35
Detroit River
Table 5-8 PAH in Suspended Solids (ng/g, dry weight) 5-36
of the Detroit River
Table 5-9 PAH in Surficial Sediment Samples (ng/g, dry weight) 5-37
of the Detroit River
Table 5-10 Mean PAH Concentrations Measured in Lake Michigan 5-39
Sediment, Pore Water, Dissolved and Particulate
Paired Samples
Table 5-1 1 Free-Radical Oxidation of Some PAH in 5-43
Air-Saturated Water
Table 5-12 Photodegradation of PAH Under Natural Light in 5-44
Mixed Acetone-Water or Carbon Tetrachloride-
Water Solutions
Table 5-13 Photo-oxidation of Some Dissolved PAH Under 5-45
Natural Sunlight Conditions
LIST OF TABLES (cont'd)
Table 5-14 Direct Photolysis of PAH in a 5 Metre-Deep
Inland Water Body
Table 5-15 Range of Polycyclic Aromatic Hydrocarbon
Concentrations
Table 5-16 Average Polycyclic Aromatic Hydrocarbon
Concentrations
Table 5-17 PAH Bioconcentration Factors (BCF) for
Selected Species of Aquatic Organisms
Table 5-18 Observed Versus Predicted Bioconcentration
Factors for Selected PAH
Table 5-19 Degradation Rate Constants (k) and Half-Lives
for Mixed Bacterial Populations in Water and
Sediment from the Sme Stream
Table 6-1 Mean Ambient Air Levels of Polycyclic Aromatic
Hydrocarbons at Niagara-on-the-Lake
Table 6-2 Mean Ambient Air Levels of Polycyclic Aromatic
Hydrocarbons Near Niagara-on-the-Lake
Table 6-3 Particle-Bound PAH Concentration Ranges in
Ambient Air Measurements for 4 Ontario Cities
Table 6-4 Total Particle-Bound Concentrations of 10 PAH
in the Ambient Air of 4 Ontario Cities
Table 6-5 Summary of Mean Total PAH Concentration
Table 6-6 PAH Concentrations in Surface Watter
Table 6-7 PAH Data for Erie, PA
Table 6-8 PAH Concentration in Suspended Sediments
Table 6-9 Polynuclear Aromatic Hydrocarbons in Hamilton
Harbour Sediments
Page
5-47
5-49
5-50
5-54
5-59
5-62
6-3
6-4
6-6
6-7
6-8
6-11
6-12
6-14
6-15
Table 6-10 Mean, Minimum and Maximum Values of PAH
6-16
LIST OF TABLES (cont'd)
Page
Table 6-11 PAH in St. Lawrence River Sediments Samples 6-17
Table 6-12 Abundances of Polycyclic Aromatics in Lake 6-18
Ontario Sediment
Table 6-13 Highest PAH Values Recorded in Fish Collected 6-19
from the Great Lakes
Table 6-14 Polynuclear Aromatic Hydrocarbons in Great Lakes 6-20
Fish
Table 6-15 Identification of Polynuclear Aromatic 6-22
Hydrocarbons in Great Lakes Herring Gull Lipid
Table 7-1 Summary of PAH Levels in Foods Available in Ontario 7-12
Table 7-2 Estimated Average PAH Exposures per Individual 7-23
in Ontario
Table 8-1 Volumes of Environmental Segments 8-8
Table 8-2 Results for Level One Analysis of PAH Data 8-10
Table 8-3 Distribution of PAH into Air, Water and Organic 8-1 1
Components Based on Raw Property Data
Table 8-4 Sensitivity Analysis of Input Data for Level 8-13
One Analysis
Table 8-5 Distribution Changes Due to Vapour Pressure 8-14
Sensitivity
Table 8-6 Distribution Changes Due to Water Solubility 8-15
Sensitivity
Table 8-7 Distribution Changes Due to K^^ Sensitivity 8-16
Table 8-8 Point Source Emission Inventory for Hamilton 8-28
Table 8-9 Summary of Modelling Results 8-30
ISCLT Model Predictions for 1979 and 1984
for the Hamilton Area
Table 8-10 Percentage Contribution to TSP from Coke Ovens, 8-32
Traffic and Other Sources in Hamilton
UST OF FIGURES
Page
Figure 2-1 lUPAC Nomenclature, Lettering and
Labelling for Pyrene and Benzo(a)Pyrene 2-3
Figure 5-1 Sediment Concentrations of Individual PAH 5-40
in Lake Erie
Figure 5-2 PAH Concentration (log scale) 5-41
Figure 6-1 Graphical Illustration of Benzo[a] Pyrene 6-5
Concentration and Year of Study at Chippewa/
Niagara Falls
Figure 8-1 Level III Fugacity Model Results 8-22
for B[a]P in Ontario
1-1
1.0 INTRODUCTION
1 . 1 Background
The Ontario Ministry of the Environment is currently examining a number
of priority chemicals with the intent of setting multi-media standards for
their emission or discharge into the environment. This work is being
undertaken by the Standards Development and Coordination Section of
the Ministry's Hazardous Contaminants Coordination Branch and has
already led to the publication of a scientific criteria document on
polychlorinated dibenzo-p-dioxins and polychlorinated dibenzofurans
(MOE, 1985). Currently, work is underway to prepare a similar document
on polycyclic aromatic hydrocarbons (PAH), and the information base is
being developed under the following headings:
0 Environmental Characteristics: Properties, Sources, Levels and Fate;
0 Environmental Toxicology; and
o Human Toxicology.
1 .2 Technical Program
The overall objective of this project is to prepare a comprehensive
background report on the physical and chemical properties of polycyclic
aromatic hydrocarbons (PAH); their sources, uses, input and fate in the
Ontario environment and levels in air, water, soil, sediments, sludge and
food in Ontario.
1-2
The scope of the project is designed to achieve the following goals:
1 . to review the nomenclature, structure, physical and chemical proper-
ties for a minimum of thirty-four environmentally or biologically
important PAH of particular interest to the Hazardous Contaminants
Coordination Branch (HCCB) of the Ontario Ministry of the
Environment. PAH such as naphthalenes and quinones were not to
be included.
2. to determine the composition and concentrations of PAH in the
environment as single compound species and/or as environmental
profile mixtures. Sources and inputs were to be addressed in the
report.
3. to estimate the ambient levels and quantities of PAH in Ontario air,
water, fish, other aquatic life, vegetation, soil, sediments, food and
mining, smelting, municipal and industrial water, where possible, for
comparison with levels found elsewhere in Canada and other
countries.
The primary goal under this heading was the determination or
estimation of a PAH mass balance (budget) for Ontario using the
data collected from items 1-3 above.
4. to estimate the average PAH exposure levels of humans in Ontario
based on the ambient levels and modifiers arising from specific
dietary and lifestyle sources.
5. to estimate the trend in environmental levels as a means of
predicting future PAH levels in Ontario.
1-3
1 .3 Polycydic Aromatic Hydrocarbon (PAH) Priority List
A list of thirty-four priority polycyclic aromatic liydrocarbons (PAH) was
proposed by tine Hazardous Contaminants Coordination Brancli (HCCB)
for assessment (Table 1 -1 ). Selection was based on the following criteria:
0 where the quantity of a selected PAH generated or emitted into the
environment was believed to be significant;
o where a selected PAH was suspected to have adverse health effects;
and
0 where a selected PAH might serve as a source marker.
In the course of this work, several additional PAH were considered for
inclusion in the final list. Compounds were selected on the basis of the
following criteria.
0 use as a potential source marker;
0 inclusion in the Ministry of the Environment Municipal/Industrial
Strategy for Abatement (MISA) Effluent Monitoring Priority Pollutants
List (EMPPL);
0 their occurrence in soil samples in Ontario;
0 toxicity; and
0 their occurrence in source profiles of what are believed to be the
five (5) largest contributors of PAH to the Ontario environment, viz:
forest fires, residential wood and oil heating, mobile sources, oil-
fired boilers and coke ovens (OME, 1979).
1-4
TABLE 1-1
Polycyclic Aromatic Hydrocarbons in the Preliminary List
by the Ministry of the Environment
1
Acenaphthene
19
Dibenz[a,h]acridine
2
Acenaphthylene
20
Dibenz[a,h]anthracene
3
Anthanthrene
21
Dibenz[a,j]acridine
4
Anthracene
22
Dinitropyrene[1,6-]
5
Benzo[a]fluorene
23
Dinitropyrene[1,8-]
6
Benzo[a]pyrene
24
Fluoranthene
7
Benzo[b]fluoranthene
25
Fluorene
8
Benzo[e]pyrene
26
lndeno[1,2,3-cd]pyrene
9
Benzo[ghi]perylene
27
Methylanthracenes
10
Benzo|j]fiuoranthene
28
Methylphenanthrene[1-
11
Benzo[k]fluoranthene
29
Nitrofluoranthene
12
Benz[a]anthracene
30
Nitropyrene[1-]
13
Carbazole
31
Perylene
14
Chrysene
32
Phenanthrene
15
Coronene
33
Pyrene
16
Cyclopenta[cd]pyrene
34
Triphenylene
17
Dibenzothiophene
18
Dibenzo[c,g]carbazole
1-5
The list developed by using these criteria and subsequently used by
CONCORD for this report is presented in Table 1 -2. This list includes the
34 PAH identified by the Ministry of the Environment as well as an
additional 13 PAH that are believed to meet the required criteria.
1 .4 Addenda
Subsequent to compiling the list of forty-seven (47) priority polycyclic
aromatic hydrocarbons, a revised list of fifty-six (56) compounds was
proposed by the Hazardous Contaminants Coordination Branch. Addi-
tional compounds in this list that were not examined in this project
because of time and budgetary constraints are presented in Table 1-3.
They include amine- and hydroxy-substituted polycyclic aromatic
hydrocarbons and relatively complex heterocyclic compounds such as
benzo[b]naptho[2,1-d]thiophene.
1-6
TABLE 1-2
List of 47 PAH Species
1
Acenaphthene*
34
Dibenz[a,h]Acridine*
2
Acenaphthylene*
35
Dibenz[a,h] Anthracene*
3
Anthanthrene*
36
Dibenz[a,j]Acridine*
4
Anthracene*
37
Dibenz[aJ]Anthracene
5
Benzo[a]Fluorene*
38
Dimethylphenanthrene[1 ,4-]
6
Benzo[a]Pyrene*
39
Dinitropyrene[1,6-]*
7
Benzo[b] Fluoranthene*
40
Dinitropyrene[1,8-]*
8
Benzo[b]Fluorene
41
Fluoranthene*
9
Benzo[c]Chrysene
42
Fluorene*
10
Benzo[c]Fluorene
43
lndeno[1 , 2, 3-cd] Pyrene*
11
Benzo[c]Phenanthrene
44
Methylanthracenes*
12
Benzo[e]Pyrene*
45
Methylchrysene[1-]
13
Benzo[ghi] Fluoranthene
46
Methylchrysene[2-,3-,4- and 6-]
14
Benzo[ghi] Perylene*
47
Methylchrysene[5-]
15
Benzo[g]Chrysene
48
Methylfluorene[2-]
16
Benzo[j] Fluoranthene*
49
Methylfluoranthene[2-]
17
Benzo[k] Fluoranthene*
50
Methylfluoranthene[3-]
18
Benz[a]Acridine
51
Methylphenanthrene[1-]*
19
Benz[a]Anthracene*
52
Naphtho[2,3-b] Pyrene
20
Benz[c]Acridine
53
Nitroanthracene[9-]
21
Acridine
54
Nitrobenzo[a]Pyrene[6-]
22
Carbazole*
55
Nitrochrysene[6-]
23
Chrysene*
56
Nitrofluoranthene*
24
Coronene*
57
Nitropyrene[1-]*
25
Cyclopenta[cd] Pyrene*
58
Perylene*
26
Dibenzothiophene*
59
Phenanthrene*
27
Dibenzo[a,e] Fluoranthene
60
Pyrene*
28
Dibenzo[a,e] Pyrene
61
Tribenzo[aei] Pyrene
29
Dibenzo[a,h] Pyrene
62
Triphenylene*
30
Dibenzo[a,i]Pyrene
63
Dimethylbenzanthracene +
31
Dibenzo[a,l]Pyrene
64
Nitro-Acenaphthene[5-]
32
Dibenzo[c,g]Carbazole*
65
Quinoline**
33
Dibenz[a,c] Anthracene
* Ministry of the Environment List
+ Presence in Refinery Waste Identified by the MOE
"* Identified for Cigarette Smoi<e (alternative to nicotine)
1-7
TABLE 1-3
Polycyclic Aromatic Hydrocarbons on Revised
Ministry of the Environment List That
Were Not Assessed in This Study
IVIethylbenzanthracene
Dibenz[c,h]acridine (other isomers were included in primary list)
Nitrophenanthrenes
Benzo[b]naphtho[2,1-d]-thiophene
Benzo[2,3]phenanthro[4,5]-thiophene
Phenanthro[4,5-bcd]-thiophene
Aminofluoranthene
Aminophenanthrene
Aminopyrene
2-hydroxy benzo[a]pyrene
Hydroxynitro PAH
2-1
2.0 PHYSICAL AND CHEMICAL PROPERTIES OF POLYCYCUC AROMATIC
HYDROCARBONS
2.1 Nomenclature and Structure
The polycyclic aromatic hydrocarbons (PAH) referred to in this report are
characterized by three or more fused benzene rings. In addition,
compounds with this basic structure but having 5-membered rings, such
as fluorene, and heterocyclic rings containing aza- or thio-arene
substituents such as acridine and thiophene, are also included under this
heading. In view of the basic symmetry of these compounds, several
structural isomers are possible. The nomenclature proposed by the
International Union of Pure and Applied Chemistry (lUPAC) is generally
used for identification of these PAH, particularly for distinguishing between
isomers and ring substituents. The following rules help determine the
orientation for assigning ring numbering or lettering:
1. the maximum number of rings lie in a horizontal row;
2. as many rings as possible are above and to the far right of the
horizontal row; and
3. if more than one orientation meets these requirements, the one with
the minimum number of rings at the lower left is chosen.
In addition to these rules, numbering of carbons in the ring structure is
carried out in a clockwise fashion, starting with the first carbon that is not
part of another ring or part of ring fusion, which is located on the upper
right. Letters are assigned in alphabetical order to ring faces and the
face between carbons 1 and 2 is labelled "a", continuing clockwise
2-2
around the molecule. This procedure is illustrated in Figure 2-1 for
pyrene and benzo[a]pyrene.
Structures for the forty-seven (47) PAH used in this evaluation are
summarized in Table 2-1.
2.2 Chemical and Physical Properties
The chemical and physical properties of the forty-seven (47) compounds
are presented in Table 2-2. Environmentally significant properties such
as the boiling point, vapour pressure, solubility, octanol/water partition
coefficient were compiled from sources such as the lARC monographs,
1983, 1984, 1985; CRC Handbook of Physics and Chemistry, 1987;
HSDB, 1987; Lane, 1988 and EPS-Ontario, 1985.
2-3
FIGURE 2-1
lUPAC Nomenclature, Lettering and Labelling
for Pyrene and Benzo[a]Pyrene
2 rings above and to the right of
horizontal row.
Pyrene
Benzo[a] Pyrene
2-4
TABLE 2-1
Chemical Structures of Priority
Poiycyclic Aromatic Compounds
Structure
Name
Identification Mol.
CAS # Formula
6
6 5
Acenaphthene
83-32-9 C12H10
6 3
Acenaphthylene
208-96-8 C12H8
Anthanthrene or
Dibenzo[def,mno]
chrysene (lUPAC)
191-26-4 C22H12
Anthracene (I UP AC)
120-12-7 C14H10
Benzo[a]fluorene
(lUPAC)
238-84-6 Ci7H,2
2-5
TABLE 2-1 (cont'd)
ChemicaJ Staictures of Priority
Polycydic Aromatic CkDmpounds
Structure
Name
Identification Mol.
CAS # Formula
Benzo[a]pyrene
(lUPAC)
Benzo [b]f luoranthene
Benzo[e]pyrene
Benzo[ghi]perylene
50-32-8 C20H12
205-99-2 C20H12
Benzo [c]phenanthrene 195-19-7 C17H12
192-97-2 C20H12
191-24-2 C21H12
BenzoO]fluoranthene 205-82-3 C20H2
2-6
TABLE 2-1 (cont'd)
Chemical Structures of Priority
Polycyclic Aromatic Compounds
Structure
Name
Benzo[k]fluoranthene
Identification Mol.
CAS # Formula
207-08-9 C20H12
Benz[a]anthracene
56-55-3 Ci8'~'i2
Acridine
260-94-6 C13H9N
Carbazole
86-74-8 C12H9N
Chrysene
218-01-9 C18H12
Coronene
191-07-1 C24H12
2-7
TABLE 2-1 (cont'd)
Chemical Structures of Priority
Polycyclic Aromatic Compounds
Structure
Name
Cyclopenta[cd]
pyrene
Identification Mol.
CAS # Formula
27208-37-3 CigHio
:6a6
Dibenzothiophene
Dibenzo[a,e]pyrene
Dibenzo[a,i]pyrene
132-65-0 CisHgS
192-65-4 C24H14
Dibenzo[a,h]pyrene 189-64-0 C24Hi4
189-55-9 C24H1,
Dibenzo[a,l]pyrene 191-30-0 C24H14
2-8
TABLE 2-1 (cont'd)
Chemical Staictures of Priority
Potycydic Aromatic CkDmpounds
Structure
Name
Dibenzo[c,g]carbazole
Dibenz[a,h]acridine
Identification Mol.
CAS # Formula
194-59-2 C20H13N
Dlbenz[a,c]anthracene 215-58-7 C22H14
226-36-8 C,,H,,N
21' '12'
Dibenz[a,h]anthracene 53-70-3 C22H14
Dibenz[a,j]acridlne
224-42-0 C21H12N
Dinitropyrene[1,6-]
42397-64-8 Ci6HgN204
2-9
TABLE 2-1 (cont'd)
Chemical Structures of Priority
Polycyclic Aromatic Cksmpounds
Structure
Name
Dinitropyrene[1,8-]
Identification Mol.
CAS # Formula
42397-65-9 CisHgNjO^
Fluoranthene
Fluorene
206-44-0 CieHio
86-73-7 '-'i3H^Q
lndeno[1,2,3-cd]
pyrene
193-39-5 C22H12
Mettiylanthracenes
(CAS # and structure
for 1-methylanthracene
is shown)
618-48-0 C15H12
Methylchrysene[1-]
3351-28-8 C19H14
2-10
TABLE 2-1 (cont'd)
Chemical Structures of Priority
Polycydic Aromatic Compounds
9~ TlOa
c
structure
Name
^CH,
4vr>^
Methylchrysene[2-,
'H •
3-, 4-and 6-]
^
(CAS # and structure
for 2-Methylanthracene
is shown here)
5-Methylchrysene
1 -Methylphenanthrene
Identification Mol.
CAS # Formula
3351-32-4 C19H14
3697-24-3 CigH^^
832-69-9 C15H12
3-Nitrofluoranthene
1-Nitropyrene
892-21-7 C16H9NO2
5522-43-0 C16H9NO2
2-11
TABLE 2-1 (cont'd)
Chemical Staictures of Priority
Polycydic Aromatic Compounds
Structure
Name
Perylene
Phenanthrene
Identification Mol.
CAS # Formula
198-55-0 C20H12
85-01-8 C14H10
Pyrene
129-00-0 CigHio
Triphenylene
Dimethylbenz-
anthracene
(CAS # and structure
for 7,12-Dimethylbenz[a]-
anthracene is shown here)
217-59-6 C18H12
57-97-6 C20H16
8 1
'rh 1 if i^^
2-12
TABLE 2-1 (cont'd)
Chemical Staictures of Priority
Polycydic Aromatic Compounds
Structure Name Identification Mol.
CAS # Formula
5-Nitro-acenaphthene 602-87-9 CijHgNOj
6SL>M>^3 Quinoline 91-22-5 CgH^N
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2-14
The following discussion on the chemical and physical properties is a
summary of information from a recent publication on source emissions
of polycyclic organic matter (U.S. EPA, 1987). More detailed treatments
of transformation and fate in air and in condensed media appear in
Chapters 6 and 8.
Polycyclic aromatic hydrocarbons are non-polar, high melting point, high
boiling point compounds that are insoluble in water. In general, melting
points range from a minimum near 100°C for phenanthrene to near 438°C
for coronene. One exception to this general rule is
benzo[c]phenanthrene, with a melting point of 68°C.
Factors that affect melting point include out-of-plane groups or fused
rings. In general, PAH are planar compounds; however, in comparing
perylene, benzo[a]pyrene, benzo[e]pyrene and benzo[k]fluoranthene,
which are molecular isomers, perylene, the least volatile of the four (4)
PAH with the highest melting point, also has the greatest number of
vertical planes of symmetry.
The vapour pressures of PAH depend on the number of fused rings and
molecular weight of the individual compounds. For example,
phenanthrene has a vapour pressure of 9.1 x 10"^ Pa (3 rings and 14
carbons) and coronene a vapour pressure of 2.0 x 10'^° Pa (7 rings and
24 carbons). However, there are insufficient good vapour pressure data
to quantify this relationship for all PAH.
The rate of adsorption of PAH on particulate matter is dependent on the
vapour pressure of the PAH, the temperature, the surface properties and
chemical composition of the particulate matter, as well as the absolute
particle concentration in air. It is evident that these properties have a
2-15
considerable influence on the adsorption characteristics and probably the
subsequent reactivity of the PAH, as well as posing a problem during
sample collection so that the analyte faithfully represents the sampled
atmosphere.
With the exception of benzo[a]pyrene and some of the smaller molecular
weight compounds, detailed chemical and physical property data for
polycyclic aromatic hydrocarbons are scarce. Furthermore, there are
several instances where the recorded values are several orders of
magnitude different from each other for the same compound. For
example, vapour pressure values for benzo[a]pyrene range from 6.7 x
1 0"^ to 7.3 X 10''' Pa. Other examples show even greater deviations. As
a result, it appears that the published data will only allow a qualitative
assessment. For this reason, a sensitivity analysis of the data was
proposed using the Mackay Level 1 fugacity model to help illustrate these
observations and to assist in identifying those parameters that are of
importance in evaluating environmental sources and fate data. Results
from this sensitivity analysis are discussed in Chapter 8.
2.3 Chemical Reactivity
The chemical reactivity of PAH has been investigated theoretically and
experimentally. The theoretical studies have focused primarily on the
relationship between 'reactive' sites on the structure determined from
molecular orbital calculations or empirical projections and various types
of reactions. The experimental investigations have relied mainly on
laboratory studies with simulated atmospheric conditions. From a
comparative viewpoint, it is found that all PAH are more reactive than
benzene and that their reactivity to methyl radicals increases with an
increase in the number of alternating single and double bonds, that is,
2-16
with greater conjugation (U.S. EPA, 1987). In general, increased
conjugation leads to greater stability; however, it leads to greater reactivity
toward free radical addition, e.g., methyl radicals. For example,
compared to benzene, benz[a]anthracene, which has greater
conjugation, reacts with methyl radicals 468 times faster (U.S. EPA, 1987).
Similarly, electrophilic and nucleophilic reactions also occur more readily
for PAH than for benzene, since ring attachment is accompanied by
displacement of a proton to restore the stable aromatic system: this
substitution mechanism is applicable to oxidation and reduction reactions
on PAH and their facility can be explained empirically by the Le Chatelier
Principle.
The atmospheric reactivity of PAH is influenced by temperature, light,
oxygen, ozone, other chemical agents, catalysts and the surface area of
the particulate matter on which the PAH are adsorbed. The reactivity of
particle-bound PAH is limited by the lifetime of the particle in the
atmosphere. Particles are removed from the atmosphere by
sedimentation and wet scavenging processes. The effect of light on
particle-bound PAH is also significant. For example, anthracene,
phenanthrene, pyrene, benzo[a]pyrene and benz[a]anthracene show
greater reactivities on alumina or silica gel than on fly ash from coal-fired
furnaces. Furthermore, this reactivity is also influenced by the
composition of the coal from which the ash is derived. Similarly, some
early studies have shown that pyrene adsorbed on garden soil undergoes
transformations that are temperature dependent as well as being
influenced by U.V. radiation (Fatiadi, 1967). Indeed, such reactions
contribute to the uncertainty that arises during ambient air or process
sampling of PAH, since the species identified in the collected sample can
correspond to a product formed by degradation of the sampled PAH
(Brorstroem-Lunden et al., 1985), or by chemical transformation via
2-17
atmospheric reactions with oxygen, ozone, NO^ and SO^ and nitric acid
(Gibson, 1986).
The most recent reference that summarises the information on the reac-
tivity of polycyclic aromatic hydrocarbons in the atmosphere was
prepared for the United States Environmental Agency (U.S. EPA, 1987).
An earlier, but equally useful review published by lEA Coal Research is
also available (Smith, 1984). In the latter review, PAH reactivity is
evaluated for atmospheric, land and water conditions. Some of the
important conclusions were:
1. Anthracene, benz[a]anthracene, benzo[a]pyrene, benzo[e]pyrene,
coronene, dibenz[a,c]anthracene, dibenz[a,h]anthracene and pyrene
are readily oxidized on exposure to light when adsorbed on an
aluminum oxide or silica gel matrix. On the other hand, chrysene,
phenanthrene and triphenylene are unreactive under these
conditions.
2. The products of photo-oxidation of adsorbed anthracene and pyrene
are quinone derivatives, e.g.,anthraquinone or 1,6[1,8]-pyrenedione.
3. No appreciable photodegradation occurs for anthracene,
benzo [a] pyrene, fluoranthene, phenanthrene and pyrene on
exposure to light when absorbed on coal fly ash. For example, the
degradation for benzo [a] pyrene is only 15% when adsorbed on fly
ash in contrast to 50% when adsorbed on alumina after the same
exposure to light.
4. Adsorbed benzo [a] pyrene or anthracene on flyash appears to
degrade in a similar manner with or without exposure to light;
2-18
whereas adsorption of pyrene on flyash led to stabilization towards
photodegradation. There remained some uncertainty with regard to
this effect on the degradation rate of exposure to light.
5. The chemical lifetime of PAH adsorbed on fine particles in the
atmosphere is of the order of days, rather than hours.
In addition to the photo-oxidation processes discussed, the available
information on chemical transformations was also reviewed (Smith, 1984).
The information may be summarized as follows:
1 . In general, most PAH containing a benzylic carbon atom (a saturated
carbon atom attached to an aromatic ring), e.g., benzo[a]fluorene,
benzo[b]fluorene, or fluorene, undergo spontaneous oxidation
without light when adsorbed on coal flyash. For other compounds
without a benzylic carbon atom, this simple evaluation is no longer
applicable.
2. Polycyclic aromatic hydrocarbons will react with ozone as well as
nitrogen oxides in the atmosphere to form oxy- and nitro-PAH,
respectively. Less information is available on the reaction of PAH
with sulphur oxides; however, there is evidence that PAH
decomposition and ring splitting may occur during reactions with
sulphur trioxide.
For PAH reactions in the soil and in water, the following review
information was presented (Smith, 1984).
1. Polycyclic aromatic hydrocarbons will biodegrade by means of soil
microorganisms to form PAH compounds that are less aromatic in
2-19
character. For example, benzo[a] pyrene is converted mainly to 9, 1 0-
dihydroxydihydrobenzo[a]pyrene as an intermediate by soil bacteria.
Other PAH that are known to show similar characteristics are:
benz[a]anthracene, benzo[ghi]perylene, benzo[e]pyrene,
dibenz[a,h]anthracene, 7,12-dimethyl benz[a]anthracene, perylene
and pyrene. The extent to which biodegradation occurred appeared
to be independent of solubility.
2. PAH in solution or dispersed as sub-micron sols in water will
selectively adsorb at the water/suspended matter interface;
consequently, an increase in turbidity of waters containing PAH is
associated with an increase in concentration of particle-bound PAH.
3. Volatilization does not play an important role in removing PAH from
aquatic systems. Other processes such as adsorption, photolysis
and biodegradation are more critical. Furthermore, since the
particle-bound PAH gradually settle out and are deposited in
sediments, such PAH are less susceptible to photochemical or
biological oxidation.
4. Anthracene, benz[a]anthracene, benzo[a]pyrene and pyrene are
rapidly transformed in water by the action of sunlight; however,
chrysene, fluoranthene and phenanthrene photolyzed slowly.
2.4 Summary
The lUPAC nomenclature and structures for the forty-seven (47) priority
polycyclic aromatic hydrocarbons are presented in this section. The
chemical and physical properties are also summarized. The lack of
reliable data and their variability are emphasized. Whereas some of the
2-20
data for the physical properties have been peer-reviewed and are
consequently accepted in such publications as lARC and the CRC
Handbook of Physics and Chemistry, others are less reliable and are
quoted in this report as the most commonly cited value or as a range of
values. This range may correspond to several orders of magnitude in
some cases. To reduce the uncertainty with regard to PAH fate and
persistence in the environment which are directly dependent on these
properties, a sensitivity analysis of the data is recommended using
Mackay's level 1 fugacity model. This model is discussed in Chapter 8
of this report.
The chemical properties of PAH are also discussed with regard to their
reactivity in the atmosphere, soil, water and sediments. It is believed the
photo-oxidation and chemical transformations play a significant role for
particulate and gaseous PAH in the atmosphere, whereas biodegrada-
tion and reactions controlled by particle size of the particle-bound PAH
are more significant in soil, sediment and soil media.
3-1
3.0 SOURCES AND INPUTS OF PAH TO THE ATMOSPHERE
3. 1 Introduction
Incomplete combustion of any organic material results in the production
of a variety of chemicals, of which polycyclic aromatic hydrocarbons
(PAH) are an example. These PAH are present in combustion emissions
as either vapour-phase or particle-bound species and their relative
abundances are largely, but not exclusively, governed by the temperature
of the emission stream (Howard et al., 1984).
It is estimated that more than 90% of the particulate matter emitted from
combustion sources is below 0.2 um in diameter and that the majority of
PAH generated in these processes are adsorbed on particles 5.0 um or
less in diameter (Handa et al., 1984). As a result, sampling techniques
based on PAH capture on filters may not provide an accurate value of
total PAH because the filter capture efficiency depends on particle size
of the adsorbed PAH as well as the vapour/solid phase distribution.
3.2 Sources and Inputs
Polycyclic aromatic hydrocarbons are emitted to the atmosphere during
the combustion or pyrolysis of organic materials. These sources can be
separated into natural and man-made ones. The natural sources of PAH
arise from forest fires, volcanic activity and possible biosynthesis by
plants and microorganisms. Major man-made sources can be subdivided
into point and non-point groupings. Point sources are not considered to
be predominant contributors of PAH. These include processes such as
power generation, coke production, petroleum catalytic cracking,
aluminum production, incineration and carbon black production. Non-
3-2
point sources include mobile sources, wood burning, oil burners as well
as off road mobile sources such as tractors and farm equipment.
The total PAH emissions for the United States, Sweden and Norway are
summarized in Table 3-1. They show that mobile sources are the major
source in the United States corresponding to 35% of the total emissions.
Industrial sources are responsible in the United States for 26% of the total
PAH emissions (Bjorseth and Ramdahl, 1985).
PAH enter the aquatic environment through petroleum spills, runoff from
roads and waste storage areas, atmospheric fallout, industrial effluents
and by seepage from creosoted structures (Mix, 1984; NRCC, 1983; Neff,
1979). Neff (1979) estimated that the global aquatic PAH could be
divided as follows: biosynthesis (1%); petroleum spillage (74%); domestic
and industrial waste (2%); surface runoff (1%); and wet and dry
atmospheric fallout (22%). Since petroleum tanker spills, offshore
production leaks and natural seeps are more applicable to the world's
oceans, the "petroleum spillage" component should be dramatically
reduced with respect to Ontario.
The global estimates compiled by the National Research Council are
presented in Table 3-2. The water-based component includes sources
(e.g., tanker spills) that are not relevant to Ontario; however, the results
indicate that atmospheric deposition may be the major source of PAH to
Ontario waters.
3-3
TABLE 3-1
Estimated PAH Emissions
Source
PAH Emissions
(MT/yr)
PAH Emissions
(%)
U.S.A. NoHA/ay
Sweden
U.S.A.
Residential combustion
715
62.5
132
12.3
Industrial production
1,637
202.7
53.3
28
Power generation
401
1.3
0.5
7
Incineration
(incl. forest fires)
1,150
13.7
3.5
19.8
Mobile sources
2,170
20
47
36
TOTAL
6,073
300
236
From: Bjorseth and Ramdahl, 1985
3-4
TABLE 3-2
Global Estimated Inputs of Benzo[a]Pyrene and Total PAH
to the Aquatic Environment from Various Sources
Estimated Input
(MT/yr)
Source
Benzo[a]pyrene Total PAH*
Percentage
(Total PAH)
Atmospheric deposition
Water-based discharges
1,700 34,000
1-5 30,000-
150,000
17-42.5
37.5 - 75
Land-based discharges
Domestic and industrial
effluents
Surface runoff
Spent lubncant disposal
17
960
1
1,400
10,800
190
0.7-
5.4-
0.1 ■
■ 1.8
- 13.5
■0.2
Biosynthesis
24
2,700
1.3-
-3.4
Approximate total
2,700
80,000-
200,000
Composition of "total PAH" was not present in ohginal reference.
From: NRCC, 1983
3-5
After entry to the aquatic environment, there is rapid absorption of most
PAH to particulate matter in the water column. While some PAH are
taken up by biota, a large proportion is deposited to the sediments. More
detailed discussion of these processes will be presented in Chapter 4.
3.3 Atmospheric/Terrestrial Sources and Input Data
Air emission factors for PAH are usually reported as a rate per unit of raw
material consumed during production. These factors are then multiplied
by average production or consumption rates to obtain the overall PAH
emission rate to the environment.
Emission factors are generally based on a limited number of tests using
a variety of sampling and analytical procedures. The lack of a
standardized procedure makes it difficult to compare data from different
research groups and accounts for the confusion in the reported data.
The work on emission factors has been reviewed, NRC, 1983; U.S. EPA,
1987. The former report included a comprehensive compilation of
emission factors from both mobile and stationary sources as well as other
anthropogenic and natural sources; whereas the latter dealt with factors
for polycyclic organic matter in general, including polycyclic aromatic
hydrocarbons. PAH emissions from coal utilization have been extensively
reviewed (Smith, 1984; Energy and Environment, 1987; and Bjorseth and
Ramdahl, 1985). ORTECH International recently compiled a detailed set
of emission factors for airborne toxics, including PAH, for Ontario MOE
(ORTECH, 1988).
Polycyclic aromatic hydrocarbon emissions are characterized by specific
PAH profiles for specific mixtures. Although it is recognized that source
profiles will differ significantly from ambient levels because of atmospheric
3-6
reactions, these profiles may be useful in evaluating their environmental
impact and perhaps for the identification of sources based on ambient air
measurements.
Other methods of source identification include:
0 a comparison of binary PAH ratios; and
0 the use of source markers, including substances other than PAH
such as heavy metals, radiocarbon tracers, etc.
The method of using binary ratios requires the evaluation of characteristic
PAH pair ratios to identify emission sources. Source markers, on the
other hand, are tracers that are unique to, or almost exclusively emitted
from, a particular source. They are not restricted to PAH species, as
tracer elements such as lead and vanadium have been identified, and the
total PAH concentration is then statistically correlated to these non-PAH.
These concepts are referred to in the remaining chapters, where
information is available.
3.3.1 Industrial Discharges
Petroleum refineries, steel mills, the wood preserving industry and coal
processing operations have been documented sources of PAH in plant
discharges in Ontario. Other industries that are possible contributors are
the plastics and dyestuffs industries, gas works and lime processing
(NRCC, 1983).
3-7
3.3.1.1 Coke Production
Coke is produced by the destructive distillation of bituminous coal and its
major use is for steel production in Ontario. It is believed that in North
America as a whole, 93 percent of the coke produced is used to convert
iron ore to the metal (Radian, 1983). Iron foundries and chemical-
producing plants account for the remainder (Smith, 1984).
By-product coke production is carried out in enclosed slot-type ovens.
The major components of this process are:
0 charging the ovens with pulverized coal;
0 the coking reaction;
0 mechanical 'pushing' of the coke into quench cars after coking is
complete;
0 water quenching of the hot coke; and
0 by-product recovery.
At each stage of the process, fugitive emissions arising from leakages
occur, often through badly-fitting coke oven doors (Smith, 1984). Battery
stacks are located on the ovens to provide a natural draft of the
combustion gas used to heat the battery and oven gases leak through
the walls and are emitted via stacks. Emission factors for coke
production are presented in Tables 3-3 and 3-4 based on data by Radian
(1983); Smith (1984); ORF (1978; 1979); MOE (1980); Krugei (1979);
Adamek (1980); Lao et al. (1979); Ortech, 1988; and MOE Air Emissions
Inventory, 1985.
3-8
TABLE 3-3
Emission Factors for Coke Production
Process
Benzene-soluble Fraction
mg/MT coal
B[a]P
mg/MT coal
Charging
50,000 - 550,000
-
Door Leaks
4,200 - 260,000
2- 1,400
Topside Leaks
4-50
-
Pushing
8,000- 17,000
-
Quenching
11,000-2,800,000
0.5
Battery Stacks
1,600
2-20
Total
74,800 - 3,628,650
4.5-1,420
no data
From Radian, 1983; U.S. EPA, 1987.
3-9
TABLE 3-4
Emission Profiles and
Factors for Coke Production
Operation
PAH
Emission Factor
(mg/MT coal)
Coke Quenching
1 Anthracene & phenanthrene
0.65
Benz[a]anthracene & chrysene
0.3
Benzo[c]phenanthrene
0.2
Dimethylbenz[a]anthracene
0.3
Fluoranthene
0.4
Methyl anthracenes
0.4
Methyl chrysenes
0.02
Methyl fluoranthene
■
& methyl pyrene
0.2
Pyrene
0.3 . -
Total PAH
2.5 - 2.9
Door Leaks
Benzo [a] pyrene
1.0- 1,400
Total PAH
4,300
(control unknown)
Total PAH
0.07
(ESP controlled)
Battery Stacks*
Benzo[k]fluoranthene
31
Benzo[a]pyrene
55
Benzo[ghi]perylene
25
Fluoranthene
100
Perylene
44
* Canadian emission factors
From: Ortech, 1988
3-10
TABLE 3-5
Summary of PAH Emission Factors for
Battery Stack Emissions During
Coke Pushing Operations in Ontario
Compound
Steico
mg/MT coal
1. MCE, 1979; ORF, 1978
2. ORF, 1979b; MOE, 1980
3. ORF, 1979a
Dofasco
mg/MT coal
Particulate
2,350
11,000
Fluoranthene
2.2
-
Perylene
1.6
3
Benzo[k]fluoranthene
1.2
-
Benzo[a]pyrene
1.4
-
Benzo[ghi]perylene
0.8
0.2
Total PAH
7.2
3
Algoma
mg/MT coal
2,600
Notes:
Emissions from Steico controlled by wet electrostatic precipitator; Hamilton plant
only.
Emissions from Dofasco controlled by a venturi wet scrubber.
Emission from Algoma controlled (control system not recorded).
no data
3-11
Also reported in Table 3-5 are data compiled from stack tests on numbers
6 and 7 coke oven batteries at Steico, Hilton Works (ORF, 1978; MOE,
1979); Dofasco, Hamilton Plant (MOE, 1980; ORF, 1979b); and Algoma,
Sault Ste. Marie (ORF, 1979a). Results show that despite the high
benzene soluble fraction, the PAH levels are low since the former includes
a large contribution from naphthalenes.
The studies conducted at Steico and Dofasco indicate that the most
abundant PAH in coke oven emissions is fluoranthene.
A comparison of source profiles of coke oven emissions with mobile
sources (gasoline) and residential coal heating indicates that benzo[a]
pyrene/benzo[e]pyrene and anthanthrene/benzo[e]pyrene ratios are
higher for coke oven emissions than for residential coal heating and that
B[a]P is enriched in particulate matter for coke ovens in comparison to
emissions from the other two sources (Daisey et al., 1986). These
observations imply that in the Hamilton and Sault Ste. Marie regions for
example, ambient B[a]P levels should be high compared to other regions
in Ontario and that B[a]P/B[e]P and Anthn/B[e]P ratios are also relatively
high. A summary table of particulate matter, PAH and B[a]P emissions
from Steico, Dofasco and Algoma is presented in Table 3-6.
According to the Ontario MOE Emission Inventory for 1985, the total
particulate matter emission from coke oven operations in Ontario (four
facilities) is about 1300 MT/y. If one assumes that the total PAH emission
factor for Steico (Hamilton) from Table 3-5 applies to the provincial
emission inventory of total particulate emission from coking, the estimated
emission quantities are:
PAH: 1300 MT/y x 0.0031 = 4 MT/y
B[a]P: 1300 MT/y x 0.0006 = 0.8 MT/y
3-12
TABLE 3-6
Estimated Values for PAH, B[a]P and Particulate Matter Emissions
During Total Coking Operations in Ontario
Particulate PAH** B[a]P***
Steico (Hamilton)
Yearly emissions 281.2* 0.86 0.17
(MT/yr)
Steico (Nanticoke)
Yearly emissions 139.8* 0.43 0.08
(MT/yr)
Dofasco (Hamilton)
Yearly emissions 681.6* 0.19 0.004
(MT/yr) '
Algoma (Sault Ste. Marie)
Yearly emissions 189.8* 0.58 0.11
(MT/yr)
ESTIMATED TOTAL
EMISSIONS (MT/yr) 1,292 2.1 0.4
* From: Ontario MOE Emission Inventory (1985). Presumed to be based on full namepiate
production capacity.
** Calculated by assuming Particulate: PAH emissions ratios from Table 3-5. For Steico 326:1;
for Dofasco 3,667:1; and Algoma 326:1.
*** Calculated by assuming for Steico, Dofasco and Algoma (Table 3-5) PAH/B[a]P levels =
5.1:1.
3-13
and so on, for the other measured PAH. Emissions of any PAH listed in
Table 3-4 may be calculated similarly by using the emission factor relative
to that of B[a]P.
Another way of estimating the PAH emission is based on estimated coal
charging consumption and the emission factors shown in Tables 3-3 and
3-4. The following assumptions are used, derived from information in the
references to Table 3-5.
Coal Consumption Data
Steico (Hamilton) - Assume charge of 24.5 MT coal per oven and an
average coking time of 17h. Assume 25 ovens (one-half capacity) are
pushed during this period.
Steico (Nanticoke) - Calculated from MCE particulate emission data and
assuming an emission factor equal to that used for Steico (Hamilton).
Dofasco (Hamilton) - Assume charge of 32.7 MT coal per oven and an
average coking time of 16h. Assume 25 ovens (one-half capacity) are
charged and pushed during this period.
Algoma (Sault Ste. Marie) - Assume charge of 16.1 MT coal per oven and
an average coking time of 23h. Assume 25 ovens (one-half capacity) are
charged and pushed during this period.
These assumptions lead to a calculated total quantity of coal
consumption for coking in Ontario of about 1x10^ MT/y at 50% capacity,
a reasonable basis for emissions estimation.
3-14
The total PAH controlled emission factor resulting from the sum of
emission factors shown in Table 3-4 is about 300 mg/MT coal.
These estimates lead to a total PAH emission of 0.3 MT/y, or 0.6 MT/y
at full capacity.
Using the Bjorseth and Ramdahl (1985) total PAH emission factor of 15
g/MT coal yields an Ontario emissions estimate of 15 MT(PAH)/y. The
level of control is not specified for this emission factor, but by comparison
with Tables 3-3 to 3-5, it must refer to a largely uncontrolled emission.
This Is not the case in Ontario.
The coking emission estimates for total PAH by the various methods of
calculation are then:
Source Total PAH Emission
Ontario MOE Emission Inventory 2.1 MT/y
and Ontario source testing data
Ontario MOE E.I. and Steico 4
only source test
Coal consumption estimate and 0.3-0.6
Ortech emission factor
Coal consumption estimate and 15
Bjorseth and Ramdahl (1985) E.F.
The most likely value is suggested to be in the range of the first two
entries in the above table, i.e., 2 to 4 MT/y. B[a]P and other specific PAH
emissions may be estimated from the data in Tables 3-3 to 3-6.
3-15
3.3.1.2 Metal Processing
The 'sintering' process, in which coke is burned to agglomerate the iron
ore for feedstock to the blast furnaces, is also a source of PAH emissions
to the atmosphere. Estimated B[a]P emission factors for this process
range from 0.6 mg/MT sinter feed processed (U.S. EPA, 1987).
Iron foundries are recognized as another source of PAH emissions into
the environment. Organic binders used in the molds consist of pitches,
asphalts, oils and synthetic polymers. The organic binders decompose
to form PAH during casting. PAH emission profiles factors for ferrous
foundries and ferro alloy manufacture are presented in Table 3-7 (Ortech,
1988).
There are currently no aluminum smelters or anode plants located in
Ontario. The closest installations are in Massena, New York, across the
St. Lawrence River from Cornwall, Ontario, where both Reynold Metals
Company and the Aluminum Company of America have smelting and
casting operations. Any contributions to the overall Ontario PAH mass
balance from the aluminum plants would occur by transboundary
migration across the St. Lawrence River or short-intermediate range
transport between the two borders.
3.3.1.3 Coal Liquefaction
There are no large-scale, continuous operating coal liquefaction plants
in Ontario and no such plants are anticipated in the foreseeable future.
3-16
TABLE 3-7
Emission Factors for Foundries
Sector
PAH
Emission Factor
(mg/MT)
Ferrous foundries
(shakeout process)
Ferro alloy Manufacture"
Si/Mn production
Total PAH
Anthracene & phenanthrene
Benz[a]anthracene & chrysene
Fluoranthene
Fluorenes
Methyl anthracenes
Pyrene
Total PAH
7,700
2,000
15
230
1,400
70
210
4,000
* Canadian emission factors
Data tabulated for scrubber controlled emissions.
From; Ortech, 1988.
3-17
3.3.1.4 Petroleum Refining (Catalytic Cracking)
Crude petroleum and many petroleum products contain an extremely
complex mixture of several thousand organic compounds including many
PAH (NRCC, 1983). The production of hydrocarbon fuels and other
refined petroleum products results in the release of hydrocarbons in
effluents, even after treatment.
Bitumen and fuel oils are derived from crude oils in the petroleum refining
process. The heavy crude is distilled to produce various fractions of
residues and distillates. Selected fractions of the distillate are upgraded
to distillate fuel by catalytic cracking (breaking up of long chain
hydrocarbons); heavier fractions of the distillate crude are further
processed to produce various grades of asphalt.
Catalytic cracking takes place in the presence of a catalyst that becomes
deactivated through continual deposition of carbon, as coke, on active
sites. The catalyst is then regenerated by combusting these coke
deposits, which result in PAH emissions and thereby make catalyst
regenerators a potential PAH source.
Three types of catalytic crackers are in use in the petroleum industry:
fluid-bed, Thermofor and Houdriflow cracking units, and the latter two
are based on a moving bed design. Most operations in Ontario are
based on the fluid-bed design. Emission factors for fluid-bed units (FCC)
are presented in Table 3-8, based on information reviewed in U.S. EPA,
1987.
3-18
Compound
TABLE 3-8
Mean Emission Factors for Ruid Bed
Petroleum Cracking Catalyst Regenerators
(controlled versus uncontrolled)
PAH Emission Factors
Controlled Uncontrolled
(CO waste heat boiler)
ug/barrel oil ug/GJ ug/barrel oil ug/GJ
Benzo[a]pyrene
Pyrene
Benzo[e]pyrene
Perylene
Benzo[ghi]perylene
Anthanthrene
Coronene
Anthracene
Phenanthrene
Fluoranthene
Total
11
92
13
ND
18
ND
ND
ND
ND
59
193
1.95
169
30
16.3
9,402
1,669
2.3
1,221
216.7
ND
ND
ND
3.2
146
25.9
ND
ND
ND
ND
ND
ND
ND
690
122
ND
133,333
23,663
10.5
6,735
1,195
34.25
151,606
26,906
1 barrel = 0.159 m"^ oil
Assume average oil density = 0.83 - 0.92 MT/m^ (Avg = 0.875 MT/m^)
Calorific value: 40.5 MJ/kg = 40.5 GJ/MT
From: U.S. EPA, 1987
3-19
PAH profiles were reported in a study conducted for the U.S. Department
of Energy by Warner, 1984. Table 3-9 summarizes PAH data for samples
from an East Montreal refinery plume (PACE, 1984). Emission factors for
petroleum refineries are presented in Table 3-10 (Ortech, 1988).
Estimates for PAH loading for Ontario are presented in Table 3-10. The
estimated annual PAH loading from Ontario refineries based on the 1986
data of crude oil transformed (26,685,100 kL) is 66.7 - 648 kg depending
on the level of control.
3.3.1.5 Production of Pyrolysis Products (Coal Tar/Creosote/Anthracene
Oil/Coal Tar Pitch/Carbon Black)
Crude coal tars are by-products produced in the carbonization process
used to make coke and/or gas. The product of crude coal tar is closely
related to the steel industry because it is a coke oven product.
A detailed study has been carried out at Domtar Chemicals coal tar
distillation plant before upgrading the air pollution control system
(Hamaliuk, 1987). Emission rates are presented in Table 3-11.
Carbon blacks are manufactured by the vapour phase pyrolysis of
hydrocarbons. Production is usually based on the oil-furnace process
to form carbon black and hydrogen: other feed materials used for carbon
black production are natural gas and acetylene.
3-20
TABLE 3-9
PAH Concentration Range in the Plume
from the Three Refineries
Located in Montreal
PAH
Adsorbed PAH/
Particulate
(mg/MT)
Anthracene/Phenanthrene
60
Methylanthracenes and Methylphenanthrenes
80
Benz[a]anthracene/Benzphenanthrenes
670
Methylbenz[a]anthracene
50-70
Acenaphthene
60-160
Fluorene
60
Benzothiophene
40-90
From: PACE, 1984.
3-21
TABLE 3-10
Average PAH Emission Factors and Total PAH Loading
for Ruid Bed (FCC) Catalyst Regenerating Units
for Petroleum Refineries
Emission Factors
Estimated
PAH
Uncontrolled
Controlled
Loading*
(mg/kL)
(mg/kL)
(kg/yr)
Anthracene
5.8
N.D.
Anthanthrene
N.D.
N.D.
Benzota]pyrene
1.4
0.9
24-37
Benzo[e]pyrene
10.2
0.1
-
Benzo[ghi]perylene 1.2
0.15
Coronene
N.D.
N.D.
Fluoranthene
56.5
0.5
Perylene
N.D.
N.D.
Phenanthrene
1,120
N.D.
Pyrene
78.8
0.8
Total PAH
24.3**
2.5***
66.7-648
* Based on 26,685,100 kL total crude oil transformed to refined products in 1986
(PACE Report 81-4, 1987). Estimates cover the range of "controlled" to
"uncontrolled" emissions.
** Assume average oil density = 0.875 MT/kL; 28 mg/MT = 24.3 mg/kL
*** Calculated from sum of average values for 10 PAH.
From: Ortech, 1988
3-22
TABLE 3-1 1
PAH Emission Rates from a Coal Tar
Distillation Plant in Hamilton
Emission Rate (kg/day)
Source
PAH
iC)
Benzene-sol
Fraction
Total
Particulate(^)
Comments
Hotwell
0
3000
5780
1. Before upgrading
scrubber system.
Wash Oil
Tank
(upset condition)
5
898
2090
2. Before incinera-
tion of emissions
from the Wash Oil,
Tar Mix and
Pitch Tanks.
Tar Mix Tank
2
48
125
Pitch Tanks
9
258
570
C)
C)
PAH include: acenaphthylene, acenaphthene, fluorene, anthracene,
phenthrene, fluoranthene and pyrene.
Sampled using the Montreal Urban Community cold trap method. Benzene
soluble and total particulate fraction contain naphthalenes.
Personal Communication, Hamaliuk, G., Domtar Chemicals Ltd., 1987.
3-23
Emission factors for carbon black manufacturing in all cases are
applicable to emissions from the main process vent (U.S. EPA, 1987).
No data are available for raw product processes such as grinding, drying
or packaging. Emission factors for polycyclic aromatic hydrocarbons
produced during an oil-furnace carbon black operation are presented in
Table 3-12. The major constituent are acenaphthene (42%), pyrene
(26%) and methyl and dimethyl-substituted anthracenes/phenanthenes
(12%) (Serth et a!., 1980; U.S. EPA, 1987).
Bitumens are mainly used in hot-mix asphalt plants and roofing
manufacturing. Hot mix asphalt, used primarily for road paving, is
produced by mixing and blending of stone aggregates, sand and bitumen
at 120°C to 180°C. Asphalt roofing products generally involve the
impregnation of heavy paper felt with various types of asphalt. The major
sources of PAH emissions from asphalt roofing manufacturing are the air
blowing operation and the asphalt saturator. During air blowing, air is
bubbled through hot asphalt at 220° to 290°C. Hot asphalt at 200-230°C
is applied to the heavy felt by a spraying and/or dipping process at the
asphalt saturator.
PAH emissions during asphalt production have been reported (Ortech,
1988). The major PAH identified were anthracene and phenanthrene,
dibenzothiophene, methylanthracenes and methylphenanthrenes with
emissions factors of 4.0, 3.6 and 6.9 mg/MT asphalt produced in Hot Mix
plant operations. The total PAH emission factor was reported to be 19.7
mg/MT (Ortech, 1988). An earlier study reported particle-bound PAH
emission rates of 1 x 10"* g/hr for benzo[k]fluoranthene and 1.6 x 10'^
g/hr for chrysene (MOE, 1977).
3-24
TABLE 3-12
PAH Emission Factors from an
Oil-Furnace Carbon Black Plant
Compound Mean Emission Factor
mg/MT
Acenaphthylene 800
Anthrancene/phenanthrene 70
Benzo[c]phenanthrene <2
Benzofluoranthenes 30
Benzo[ghi]pery!ene 40
Benzopyrenes & perylene 30 -
Chrysene & benz[a]anthracene 9
Dibenzanthracenes <2
Dibenzocarbazole <2
Dibenzopyrenes <2
Dibenzothiophene 14
Fluoranthene 60
!ndeno[1,2,3-cd]pyrene <2
Methylanthracene/phenanthrenes 1 00
Methyifluoranthene/pyrene 23
Pyrene 500
Total PAH (1) 1.900
(1) includes other PAH not on the priority list.
From: U.S. EPA, 1987.
3-25
Gunkel and Bowles (1985) reported an emission factor of 1.07-0.25
ug/MT asphalt for 16 volatile organic compounds in a study of emissions
from batch mix asphalt and drum mix asphalt plants. PAH that were
detected included acenaphthylene, acenaphthene, phenanthrene, fluorene
and pyrene.
3.3.2 Power Production Sources
3.3.2.1 Coal Mining (Coal Dust)
The presence of phenanthrene, pyrene, chrysene and perylene, as well
as other PAH has been reported present in respirable coal dust (Setzer,
1979).
This result has been verified in work by other authors, in which thirteen
(13) PAH were identified in the respirable dust fraction in coal mines
(Shultz et al., 1972). As a source of PAH atmospheric emissions, coal
mining plays an insignificant role, but amounts generated by fugitive
emissions during loading from storage piles etc. are unknown.
3.3.2.2 Coai-Rred Power Plants
In 1979, it was estimated that coal accounted for 90% of the fossil fuel
consumption for the generation of electric power in Ontario (MOE, 1979).
Despite the predictions in the 1950s and 60s indicating that nuclear-
powered stations would supersede this method, Ontario Hydro operates
large coal-fired stations at Nanticoke, Lakeview and Lambton and smaller
units at Atikokan and Thunder Bay to generate about 30% of total power
production. At full load, Nanticoke is nominally rated for 4000 MW,
Lakeview at 2400 and Lambton at 2000 MW. Atikokan and Thunder Bay
3-26
are nominally rated at full load at 200 and 400 MW respectively. In
general, these units are front- or tangential-fired with wet bottom furnaces
or with a travelling grate spreader stoker.
A detailed examination of the data related to the method of firing and
consequently rate of firing for units of different design was reported by
Hangebrauck et a!., 1967. The results indicated significant amounts of
B[a]P, pyrene, B[e]P, benzo[ghi]perylene and fluorene regardless of the
method of firing, as well as some variation in detectability of other PAH
including, perylene, anthanthrene, coronene, anthracene and
phenanthrene.
Another study also provided an emission factor database that
distinguishes between vapour phase and particle-bound PAH (NRC,
1983). Data are reproduced in Table 3-13. Emission factors complied
in a more recent study for the Ontario Ministry of the Environment are
summarized in Table 3-14 (Ortech, 1988).
The emission profile, relative PAH emission rates for Ontario Hydro's
Lambton and Nanticoke thermal generating stations are presented in
Table 3-15 (ORF, 1986; Evans et al., 1985). Table 3-16 provides a
summary of the estimated PAH and B[a]P emissions for Ontario Hydro's
coal-fired generating stations.
3-27
TABLE 3-13
Average PAH Emission
Factors for Drfferentty-Rred Coal Power Plants
Unit BSF B[a]P
mg/MT mg/GJ ug/MT ug/GJ
Vertically-fired 29.6 1140 0.003 0.12
dry bottom C)
Front wall-fired 8.3 320 0.0005 0.02
dry bottom C)
Tangentially-fired 16.1 620 0.003 0.13
dry bottom C)
Opposed, 32.0 1230 0.003 0.13
base-directed burners,
wet bottom C)
Cyclone-fired 52 2000 0.009 0.35
wet bottom f )
Spreader stoker 27 1040 0.0005 0.02
travelling grate f )
BSF - benzene soluble fraction
C) pulverized coal
( ) crushed coal
From NRC, 1983.
3-28
TABLE 3-14
PAH Emission Factors for Coal-Rred
Thermal Generating Stations
Emission Factor*
(mg/MT coal)
PAH
Bituminous
Lignite
Acenaphthene
N.D. -0.73
N.D.
Acenaphthylene
N.D. - 1.62
N.D.
Anthracene
0.11 -2.22
0.3 - 0.54
Benz[a]anthracene
N.D. -3.02
N.D. - 0.29
Benzo[b]fluoranthene
N.D. - 1.73
N.D. - 0.93
Benzo[k[fluoranthene
N.D. - 1.55
N.D. -0.69
BenzG[a]pyrene
0.29 - 48
N.D. - 0.24 (2.0)**
Benzo[e]pyrene
0.13-0.28
0.69 - 0.95
Benzo[a]phenanthrene
N.D. -0.51
0.39 - 4.0
Benzo[a]fluorene
0.73 - 3.9
0.1 -0.3
Benzo[ghi]perylene
0.04 - 5.7
0.53 - 3.9 -
Chrysene
N.D. - 5.54
N.D. -0.54
Dibenz[a,h]anthracene
N.D. - 5.0
N.D. -0.36
Fluoranthene
N.D. - 15.65
0.05 - 0.66
Fluorene
N.D. - 2.68
0.15-0.18
Indeno[1 ,2,3-cd]pyrene
N.D. - 1.2
N.D. -0.59
2-methylanthracene
0.15-2.5
0.6 - 0.93
9-methylanthracene
N.D. - 14
N.D.
9, 1 0-dimethylanthracene
1.6- 12
2.2-3.5
1 -methylphenanthrene
0.54 - 7.1
0.3 - 0.44
1-nitropyrene
0.24 - 2.3
0.42 - 2.8
Phenanthrene
0.03-31.4
0.18-0.93
Pyrene
3.3- 17.0
4.3 - 6.9
Triphenylene
N.D. -0.11
N.D.
Total PAH
22-120 (28)**
26-31
* Canadian Emission Factors
** Average for an ESP controlled facility.
From: Ortech, 1988.
3-29
TABLE 3-15
Ontario Hydro Lambton and Nanticoke Thermal Generating Stations
Emission Rates of Polycyclic Aromatic Hydrocarbons
Compound
Emission Rate
(ug/s)
Lambton
Nanticoke
i,3
Naphthalene
0.3
Acenaphthylene
0.3
-
Acenaphthene
0.68
-
Fluorene
2.0
-
Phenanthrene
28.9
6.7
Anthracene
0.92
27
Fluoranthene
21.4
-
Pyrene
9.8
443
Benzo [a] anthracene
6.3
-
Chrysene
10.5
-
Benzo[b]fluoranthene
5.5
-
Benzo[k]fluoranthene
5.5
-
Benzo [a] pyrene
3.4
115
Benzo[e] pyrene
-
11.4
lndeno[123-cd]pyrene
0.66
13
Dibenzo[ah]anthracene
0.24
137
Benzo[ghi]perylene
0.78
49.3
Total
97.2
1835
1 ORF (1986)
2 Evans et al. (1985)
3 IVlean of two runs.
3-30
TABLE 3-16
Estimated PAH, B[a]P and TotaJ Suspended
Particulate (TSP) Emissions
for Ontario Hydro's Coal-Rred
Power Generating Stations
Utility
Coal
Type
Coal
Consumption
(MT/yr)*
PAH** B[a]P
(MT/yr) (MT/yr)
-rp***
(MT/yr)
Nanticoke
Bituminous
5,807,000
0.16
0.01
4,200
Lakeview
Bituminous
1,353,000
0.04
0.002
1,100
Lambton
Bituminous
2,962,000
0.08
0.006
3,200
Thunder Bay
Bituminous
1,228,000
0.03
0.002
260
Atikokan
Lignite
612,000
0.02
0.0001
140
Total Emissions
0.33
0.21
8,400
* 1987 figures
** Data from Ortech, 1988 (Table 3-14 of this report).
*** Total particulate matter from Ontario Hydro report to MOE, Jan. - Dec, 1987.
This figure includes contributions from oil consumption in power generation for
4 of the 5 stations.
Assumed B[a]P emission factor for bituminous coal = lignite coal factor = 2 mg/MT.
3-31
3.3.3 Municipal and Hazardous Waste Incinerators
Only limited data are available on the release of polycydic aromatic
hydrocarbons from facilities of this type. Ontario based data are virtually
non-existent; results from SWARU were reported to be very low in the
Ontario PAH source survey (MOE, 1977); so low as to make the data
suspect (0.5-1.5 ng/m^ of benzo[k]fluoranthene and benzo[a]pyrene).
More recent data are available from the National Incinerator Testing and
Evaluation Program (NITEP) and other municipal solid waste (MSW)
incinerator sampling programs. These data are presented in this section.
SWARU and the London Victoria Hospital EFW incinerator are the only
two operating municipal incinerators in Ontario at present.
Few data on hazardous waste incinerator emissions are available;
however, values obtained for the Tricil incinerator in Sarnia have been
reported and are summarized in Table 3-17 (ORF, 1987). The Ontario
Research Foundation report on the Tricil unit suggests that the stack flow
is approximately 17.5 mVs with 27% moisture at 216°C and 13.8% O2.
Apartment incinerators are currently prohibited in the province. They
have been largely uncontrolled and poorly maintained in the past. No
data are available for PAH from this source.
Biomedical incinerators in most facilities are similar to the batch-fed, multi-
chamber incinerators used in apartments. The units are likely better
maintained than apartment units, but the majority of the units do not meet
MOE combustion guidelines (CSC, 1987).
3-32
TABLE 3-17
PAH Emission Data
for TRICIL Hazardous Waste Incinerator
(ng/Rm^@ 11% O2)
Compound PAH Concentration (ng/Rm^)
Acenaphthylene
160
310
Acenaphthene
170
460
Anthracene
250
870
Benz[a]Anthracene
80
420
Benzo[b]Anthracene
7
20
Benzo[b]Fluoranthene +
Benzo[k]Fluoranthene
70
320
Benzo[a]Fluorene
20
120
Benzo[b]Fluorene
7
120
Benzo[ghi]Perylene
110
260
Benzo[a]Pyrene
40
110
Benzo[e]Pyrene
1.5 -
10 -
Chrysene + Triphenylene
170
740
Coronene
30
190
Dibenz[ac]Anthracene +
Dibenz[dh]Anthracene +
Picene
10
70
Dibenzo[ac]Pyrene
6
40
9,10-Dimethylanthracene
N.D.
7,12-Dimethylbenz[a]
anthracene
100
500
Fluoranthene
950
3900
Fluorene
320
990
lndeno[123-cd]Pyrene
70
230
2-Methylanthracene
1 -Methylphenanthrene
410
1200
9-Methylphenanthrene
80
Perylene
10
40
Phenanthrene
2500
7300
Quinoline
Total
6460
21010
N.D. no data
From: ORF, 1987.
3-33
Some hospitals have semi-continuous units of the starved air modular
type. Some data from a unit of this type in British Columbia indicates
total PAH levels in the range of 6.1 ug/Rm^ @ 11% Og.
Three NITEP studies have been completed at Charlottetown and Quebec
City, as well as an air pollution control (APC) test in Quebec City. PAH
data from these systems under normal operating conditions are
summarized in Tables 3-18 and 3-19 (NITEP, 1985; 1988).
It is generally thought that PAH levels will be a function of particulate
matter levels; however. Table 3-18 shows no consistent relationship
between these parameters.
Total PAH data from Quebec City (Table 3-19) illustrate the influence of
low temperatures as well as the fact that poor air distribution does not
produce a significant change in total PAH levels.
The total PAH values for PEI incinerator ash are 1.2-4.7 ug/g; for the
boiler the value is 0.03 ug/g and for the economizer 0.01-0.07 ug/g. It
is estimated that incinerator ash may total 15% of the feed excluding
glass and metal; boiler and economizer ash 0.07% of feed.
Quebec City values were 0.1-0.5 ug/g incinerator, 0.02-0.07 ug/g boiler
and 0.1-0.3 ug/g for the precipitator. Typical APC levels were 0.3-2.0
ug/g maximum of total PAH. The incinerator ash at Quebec City is
approximately 25% of the total feed including glass and metal, boiler 0.5%
and precipitator 0.8%.
3-34
TABLE 3-18
NUEP PAH Emission Data and
Levels in Incinerator Ash
PAH
Ash Concentrations (ng/g)
Emissions
Incin.
Bottom
Exit
(ng/Rm^)
Acenaphthylene
70
37
5
Acenaphthene
80
28
-
-
Fluorene
360
150
4
1
Phenanthrene
1630
541
21
21
Anthracene
590
53
1
4
Fluoranthene
390
113
3
15
Pyrene
310
146
3
8
Chrysene
10
37
-
1
Benzanthracene
130
39
5
9
Benzene, pyrene
170
51
-
-
& fluorene
Indenopyrene
20
3
-
4
Dibenzanthracene
7
4
-
4
Benzpen/lene
3
5
-
3
Total PAH
3750
1206
38
73
Total Particulate Matten50
(mg/Rm^@ 11% O2)
From: Vol. I NITEP Charlottetown Testing Program, 1985.
3-35
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3-36
A comparison of the PAH emission factors and emission profiles for
MSW, industrial liquid waste and commercial incinerators treating
hazardous waste is presented in Table 3-20.
Currently in use or contemplated {*) are the following facilities: London
Victoria Hospital, 273 tpd starved air with APC; Peel*, 364 tpd starved air
with APC; Toronto Refuse-fired Steam Plant*, 1700 tpd mass burn with
APC; SWARU 480 tpd with APC; and Trintek* 500 tpd starved air with
APC. The Commissioners Street, 273 tpd unit has been mothballed and
will probably not be recommissioned.
Other projects are being investigated in Ottawa, Guelph, Windsor and
Kingston. These units will probably have a design similar to Peel, except
for Ottawa which will have a mass burn system.
Based on the currently operating and planned incinerator facilities, the
estimated PAH loading to the atmosphere is 1.4 MT/yr; data for each
incinerator are compiled in Table 3-21.
3.3.4 Transportation
Gas-phase and particulate polycyclic aromatic hydrocarbons are major
components in the emissions from internal combustion engines. Some
nitro-substituted PAH are believed to form in the exhaust through reaction
of the PAH and nitrogen oxides. The emission rates and PAH profiles are
dependent on the temperature of the combustion chamber and exhaust
system and consequently on engine size, design, working load and
operating speed (U.S. EPA, 1987).
3-37
TABLE 3-20
Summary Table of PAH Emission Factors
from MSW Industrial Liquid Waste and Commercial Incinerators
PAH
MSW
Incinerators
Emission Factors (mg/MT feed)
Industrial
Liquid Waste Commercial
Incinerators Incinerators
Acenaphthene
12
3.6
^
Acenaplnttnylene
1,000
2.7
-
Anthracene
-
5.8
104-190
Anthanthrene
-
-
14.6-174
Benz[a]ant(nracene
3.1
3.0
-
& Chrysene
.
Benzo[a]pyrene
24
1.0
117-573
Benzo[e]pyrene
-
1.5
99.2-573
Benzo[b,j&k]fluoranthene
78
2.5
-
Benzo[ghi]perylene
14
2.1
198-1,918
Chrysene &
62
5.5
-
triphenylene
Coronene
0.2
1.3
46.3-463
Dibenz[a,h]anthracene
0.11
0.6
-
& dibenz[a,c]anthracene
Fluoranthene
140
24.8
458-8,600
Fluorene
68
6.7
-
lndeno[1 ,2,3-cd]pyrene
0.77
1.9
-
Phenanthrene
426
50.2
130-309
Perylene
0.77
0.3
6.8-132
Pyrene
168
14
706-9,261
2-methylanthracene
-
18.9
-
1 -methylphenanthrene
-
8.5
-
9-methylphenanthrene
-
0.3
-
Total PAH
1,160*
374
1,910-22,200
- no data
* Column does not add to total because of differing databases for average values
for individual PAH.
From: Ortech, 1988.
3-38
TABLE 3-21
Estimated Annual PAH Atmospheric Emissions
to the Ontario Environment from Incinerators
Facility
Feed
(MT/yr)
PAH Emission
Factor Loading
(mg/MT feed) (MT/yr)
London Victoria
91,250
1,160
0.10
Peel*
120,450
1,160
0.14
Toronto Steam Plant*
562,100
1,160
0.65
Trintek*
164,250
1,910
0.3
SWARU
159,140
1,160
0.18
Total PAH
Existing
Proposed
0.3
1.1
Proposed installations.
3-39
The PAH emissions are diluted approximately 1000-fold after being
exhausted and are cooled very rapidly. For example, the combustion of
an efficiently-operated gasoline engine operates around 3500°C with
exhaust temperatures between 400° and 600°C, whereas, for diesel
engines, the corresponding tempeatures are 2000°C and 200 to 400°C
respectively. It is estimated that within a few seconds, exhaust emissions
are rapidly dispersed and cooled to ambient tempertures (U.S. EPA,
1987).
Heavier PAH, e.g., B[a]P, are present in mobile source emissions mainly
as sub-micron particle-boud species, with the lighter PAH in the vapour
phase. Engine oil accumulates PAH, which then leaks into the
combustion chamber or exhaust system, thereby enhancing these levels
in the emissions. It is estimated for the United States that 28 to 36
percent of B[a]P, B[a]A, CHR and P in engine oil that leaks into an
automobile combustion chamber or exhaust system is emitted to the
atmosphere (Handa et al., 1979).
Internal combustion engines without oil crankcases include two-stroke
engines used for motorcycles, outboard motors, lawnmowers and
chainsaws that are operated on a mixture of oil and gas. Because of
their mode of operation, it is probable that PAH levels in the oil are in
direct relation to PAH levels in the emissions. Typical PAH identified in
such emissions are: fluorene, anthracene, pyrene, benz[a]anthracene,
benzo[a]pyrene and benzo[ghi]perylene. It is estimated for the United
States that the contribution from motorcycle emissions to the total PAH
emissions from mobile sources in 1979 were less than 0.3 percent (NRC,
1983). There is a general trend toward fewer motorcycles in the United
States and in Canada, in favour of passenger automobiles; therefore, the
PAH contribution from motorcycles and two-stroke engines in general is
3-40
probably insignificant compared to other sources and emissions from
two-stroke engines are therefore only briefly discussed in this report.
There are few data on aircraft turbine engines, specifically particulate
emissions from the combustion of kerosene fuels in gas turbines. PAH
detected in studies of these emissions were: fluorene, anthracene/
phenanthrene, methylfluorene, fluoranthene, pyrene, acenaphthylene,
benzofluoranthene, chrysene, benzopyrenes (mainly as benzo[e]pyrene)
and perylene (Robertson et al., 1980). An emission factor of 30 ug
B[a]P/kg fuel consumed has been reported by Smith, 1984.
3.3.4.1 Gasoline- and Diesel-Powered Vehicles
Because of the operating temperature and difference in engine design,
diesel engines emit from 30 to 100 times more particulate matter by mass
than gasoline engines, over 90 percent of which is below 0.1 um in
aerodynamic diameter (NRC, 1983).
A list of compounds present in particulate matter exhausted from gasoline
and diesel powered sources is given in Table 3-22. It is concluded that
a higher level of nitro-substituted PAH is present in diesel-powered
sources than in gasoline-powered sources. This is related to PAH
reactivity in the exhaust with nitrogen oxides at the exhaust temperatures
(U.S. EPA, 1987).
A detailed list of emission factors for light gasoline- and diesel-powered
engines is presented in Table 3-23. Estimates for total PAH produced
from this source are presented in Table 3-24.
3-41
TABLE 3-22
Polycyclic Aromatic Hydrocarbons Detected
on Gasoline and Diesel Exhaust Particles
Gasoline Powered (1)
Diesel Powered (2)
Phenanthrene
Anthracene
Methylphenanthrenes/
anthracenes
Fluoranthene
Pyrene
Methylpyrene
Benzo[a&b]fluorene
Benz[a]anthracene
Chrysene & Triphenylene
Benzo[b,j&k]fluoranthenes
Benzo[a]pyrene
Benzo[e] pyrene
Perylene
lndeno[1 ,2,3-cd]pyrene
Benzo[ghi]perylene
Anthanthrene
Coronene
Acenaphthene
Fluorene
Phenanthrene
Anthracene
Methylphenanthrenes
Fluoranthene
Pyrene
Methylpyrenes/fluoranthenes
Benzo[a&b]fluorene
Benz[a]anthracene
Chrysene & Triphenylene
Benzo[b,j&k]fluoranthenes
Benzo[a]pyrene
Benzo[e]pyrene
Perylene
lndeno[1 ,2,3-cd]pyrene
Benzo[ghi]perylene
Anthanthrene
Coronene
Methylfluorenes
Dibenzothiophene
Dibenzo[a,h]anthracene
Nitropyrenes/fluoranthenes
Nitrofluorenes
Dinitrofluorenes (tent.)
(1)Alsberg et al., 1985
(2) Li-Yu et al., 1981; Tong et al., 1984; and Schuetzle et al., 1981,
3-42
TABLE 3-23
Emission Factors for Gasoline
and Diesel Powered Mobile Sources
Emission Factors (mg/kL)
Gasoline
Diesel
PAH
Leaded**
Unleaded*
Unleaded* Heavy-Duty
Light-Duty
Heavy-Duty
Motoro
Anthracene
200
54
13
591
696
.
592
Phenanthrene
4.9
220
52
2,400
2,600
-
2.400
Methylphenanthrene
1.500
275
32
1,500
1,760
-
294
Ruoranthene
360
115
27
1.257
1,860
200
1.257
Pyrene
718
30
34
1.593
2,335
-
1.593
Benzofluorene
•
36
4
200
233
-
61
Benzanthracene
97
29
1
61
217
83
86
Triphenylene
-
9.3
1
50
59
-
50
Cydopentapyrene
35
120
14
653
768
-
653
Chrysene
204
23.5
5
248
391
-
248
Indenopyrene
12
9.3
0.1
0.06
149
79
O.C
Methylchrysene
185
32.5
4
177
-
-
-
1-nitropyrene
0.9
0.7
0.3
1
72
6.7
0.9
Benzofluoranthene
185
32.5
4
177
313
93
198
Bezo[e] pyrene
41
9.0
2
86
101
-
86
Benzo[a]pyrene
31
9.9
2
72
101
62
72
Perylene
5.8
1.6
0.3
7
8
-
7
Anthanthrene
•
-
-
-
-
-
-
Ditjenzanthracene
21
3.7
0.5
21
25
9.4
21
Coronene
41
22
3
133
142
-
133
Benzoperylene
-
-
-
-
235
130
191
9-nitroanthrac8ne
-
-
-
-
-
5
-
7-nitrot)en7anthracene
-
-
•
-
-
1.2
-
6-nitrochrysene
•
-
-
-
-
<0.2
•
Total PAH
1,970
233
9,290
11,960
-
7,990
TSP*"
193,478
101,346
1.726,790
Data from Ortech International, 1988.
references.
Collated from six (6) primary and review
no cataJyst
with oxidation catalyst
with 3-way catalyst
Data from U.S. EPA, 1987.
consumption rate of 7 km/L
no data
Estimated by assuming an average gas
3-43
TABLE 3-24
Estimates of Yearly PAH Emissions
from Transporation Sources in Ontario
Fuel Type
Number
(fleet)*
Gas
Consumption
(kL/yr)**
PAH Emission Total
Factor PAH
(mg/kL) (MT/yr)
Gasoline
Automobiles & Light Trucks
- leaded 444,492
- unleaded (oxid) 3,303,188
(3-way) 610,088
(heavy-duty) 881,342
5,609,100
8,504,647
1,570,780
2,269,172
4,578
1,970
233
9,290
25.7
16.8
0.37
21.1
2-stroke engines
(motorcycle,
snowmobile, moped)
407,085
ND
7,990
-
Diesel
Automobiles
Heavy-duty/Commercial
33,872
127,652
1,932,000 +
11,960
9,240
20.5
Jet Fuel
Airline jets
306,608/y
r+ +
2-10
B[a]P
(mg/min)
0.028
as B[a]P
ND
+
+ +
no data available for 2-stroke engine gas comsumption. The total gasoline
consumption includes this volume and are included in the values considered
under autombiles and light trucks.
MOT, 1987. Registrant/Plate/Vehicle Population Statistics. Environment Canada,
1989. Private communication Lavallee, F. Percentage distribution of automobiles
with 3-way, oxidation catalysts and no catalytic converters in Canada.
Statistics Caanda, 1987. Catalogue 57-004.
MOE, 1989. Private communication, Rohac, I. 1986 estimate for diesel, oil
consumption in Ontario.
Aviation Statistics Centre, 1987. Aircraft movement statistics. 1987 annual report.
Number of take-offs and landings per year.
3-44
3.3.4.2 Tire Wear
Particle-bound and vapour phase PAH are continuously released from
tires during the normal operation of a vehicle. An emission factor of 0.14
g/d per million people was similarly estimated for B[a]P (U.S. EPA, 1987).
3.3.4.3 Source Markers (Transportation)
PAH profiles from car exhaust streams were reviewed by Daisey et al.,
1986 and an attempt was made to identify characteristic PAH ratio pairs
and fingerprint compounds. The profiles indicated high levels of
phenanthrene, pyrene, chrysene and anthracene. Cyclopenta[1,2,3-
cd]pyrene was identified as a potential source marker for vehicle exhaust
since it was present in vehicle exhaust in larger amounts (Daisey et al.,
1986) compared to other sources, such as coke oven and oil burning.
Benzo[c]phenanthrene, benzo[ghi]perylene and coronene were also
observed to be enriched in samples collected in a tunnel (Daisey et al.,
1986).
Hering et al. (1984), have also identified the following as signature PAH
compounds and heavy metals for sub-1.3 um sized particulate matter
collected from a 1983 California vehicle fleet consisting of 3-6% diesels.
These were: dibenzanthracene, benzo[ghi]perylene, indenopyrene,
benzo[b&k]fluoranthene, lead, zinc and iron. Results indicated an
increased level of benzofluoranthenes for an increase in the number of
diesels, whereas benzo[ghi]erylene, indeno[1,2,3-cd]pyrene and
dibenzanthracene were independent of the diesel population.
A study conducted by Harrison and Johnston (1985), on the deposition
of PAH with lead, cadmium and copper particulate in the U.K. indicated
3-45
that allowing for temporal variations, the fluxes for these compounds and
elements, particularly lead, were elevated close to a major hghway and
decreased to background levels within 20-40 m; however, with the
reducing use of lead as anti-knock additives in gasoline, this conclusion
is no longer applicable.
From these results, it can be inferred that when emissions from mobile
sources are compared with emissions from other sources then higher
levels of heavy metals such as lead, zinc, iron, cadmium and copper are
present for mobile sources. In addition, the level of
benzo[b&k]luoranthenes increases with an increase in diesel to gasoline
vehicle traffic volume in areas of low solar intensity which tends to reduce
PAH chemical reactivity. Also, the level of nitro-substituted PAH,
particularly nitro- and dinitro-pyrenes increases with an increase in diesel
to gasoline vehicle traffic volume.
3.3.5 Residential Heating
Extensive studies on the emissions from woodburning fireplaces as well
as wood and oil-burning stoves and furnaces have been conducted (Hall
and De Angelis, 1980; Ragland et al., 1985). Detailed reviews have also
been presented (NRC, 1983; Smith, 1984; Nero & Assoc, 1984) including
a review for Health and Welfare Canada by Concord Scientific (Davis,
1987).
3.3.5.1 Emission Factors
It is estimated that the emission factor for polycyclic organic matter, of
which PAH form a component, is 5x10"^ to 0.2 g/kg, and the associated
B[a]P emission factor is 4x10'^ - 0.0025 g/kg for woodburning fireplaces
3-46
and stoves in the United States (Lipfert and Lee, 1985). The emissions
are affected by appliance type, condition and type of wood fired, the rate
of burn and measurement method, which is one of the reasons for the
large range in values.
Particulate emission factors for unvented kerosene heaters were
estimated by Ragland et al. (1985), and ranged from 11.8 to 25.3 mg/kg
fuel for particulate and 0.6 to 54 ug/kg for total PAH emissions. Studies
were conducted on both radiant and convective heaters.
Average values for emission factors for different fuels, appliances and
operating conditions have been reported (Smith, 1984). These values are
reproduced in Table 3-25. An emission source profile from both wood
and treated lumber in woodburning stoves is presented in Table 3-26.
For both northern and southern Ontario, approximately 37% of
households burned wood. This translates to an average yearly
consumption of 5.1 full chords per household (10.2 MT) in northern
Ontario and 3.7 full chords (7.4 MT) in southern Ontario (MacLaren,
1985).
Based on an emission factor of 29-40 mg PAH/kg and 0.5-0.7 mg
B[a]P/kg wood consumed for residential heating and an estimated
35,700 households in northern Ontario and 255,000 households in
southern Ontario burning wood, the estimated annual PAH emissions for
Ontario from this source are 65-90 MT and the B[a]P emissions are 1.1-
1.6 MT.
3-47
TABLE 3-25
Typical Emission Factors for Residential Heating
Fuel
Type
Total PAH
(mg/kg)
B[a]P
(mg/kg)
Oil (30 kW)
Stove
0.15
0.0022
Oil (7.5
kW)
Stove
10
-
Gas (21
Btu/hr)
0,000
Furnace
13 (ug/m^)
1 (ug/m^)
Gas
Furnace
65 (ug/m')
-
Oil
Furnace
0.15
0.0022
Oil
Furnace
0.13
-
Wood
Stove
40
0.5
Wood
Fireplace
29
0.7
From: Smith, 1984; Radian, 1983; Hangebrauck et al., 1967.
3-48
TABLE 3-26
Emission Profiles and Factors for PAH
from Different Fuels In a
Conventional Wood Stove
Emission Factors (mg/GJ)
Compound Wood Treated Lumber
Anthracene
395
1100
Benz[a]anthracene
3.9
185
Benzofluoranthene
<1
<1
Benzo[ghi]perylene
2.7
16.3
Benzo[a]pyrene
39.5
131
Benzo[e]pyrene
32.9
147
Chrysene
105
81.7
Coronene
<1
32.7
Dibenz[a,h]anthracene
<1
NA
Fluoranthene
244
687
Fluorene
308
103
Methylanthracene
144
27.2
Methyiphenanthrene
967
583
Peryiene
52.5
16.3
Phenanthrene
644
1820
Pyrene
171
545
From: KHM, 1983; Smith, 1984.
3-49
3.3.5.2 Emissions (Oil and Gas Heating)
Oil and gas furnaces are commonly used in Ontario households to heat
water and recirculating air. Gas furnaces burn a premixed mixture of gas
and air and generally emit relatively small amounts of PAH per unit of heat
input.
On the other hand, oil fired units introduce the fuel by pressure
atomization or vaporization and, in comparison, relatively larger PAH
emission rates per unit of heat input are produced. That is oil heating
produces about 4 mg (POM)/GJ compared with 1 mg (POM)/GJ for gas
(Peters, 1981).
PAH emission factors have been estimated for various types of oil and
gas fired furnaces (NRG, 1983; Radian, 1983). These values are also
listed in Table 3-25.
The 1979 MOE report identified commercial and institutional boilers
(primarily oil- and gas-fired) as the main contributors to PAH emissions
in the category of heat and power generation. Residential furnaces (oil-
and gas-fired) were estimated to contribute little to the provincial total
PAH emissions. Recent emission factor compilations (e.g.. Table 3-25
and Ortech, 1988) indicate no significant changes in emission factor
estimates relative to the data used for MOE (1979). Assuming no
increase (or decline) since 1976, the base year of that study, in the use
of coal for industrial boiler use, about 25% increase in the use of residual
and distillate oils for residential, industrial, commercial and institutional
heating and a 50% increase in gas utilization for these purposes, the total
PAH emission from heat and power generation from residential heating,
industrial, commercial and institutional boilers is estimated to be no more
3-50
than about 2 MT/y. These sources are distributed throughout Ontario.
In the context of the current assessment, these sources, then, would
appear to be insignificant. They are not addressed in detail in this report.
3.3.6 Open Burning of Biomass
The emissions from open burning of biomass, which includes sources
such as uncontrolled forest fires, prescribed refuse and agricultural waste
burning, may contribute significantly to the PAH content of the
atmosphere. However, there are few recorded data on emission factors
for these sources.
3.3.6. 1 Burning of Treated Wood Waste
Emissions from the burning of railway ties were studied during a test burn
(Becker et al., 1984). The burn was conducted under controlled
meteorological conditions with approximately 681 kg of creosote treated
wood doused with #2 fuel oil and ignited with a railroad flare to simulate
the practice of Burlington Northern Railroad. While the practice used by
Canadian National and Canadian Pacific in Ontario is unknown, no other
data for the burning of railway ties were available.
PAH detected during the test burn are presented in Table 3-27 as
fractions of the total suspended particulate concentration. An average
total suspended particulate emission rate of 3.68 kg/h was estimated for
these burns, resulting in an emission factor of 10.6 g/kg of wood burned,
assuming that all the PAH were generated during the first two hours of
the burn.
3-51
TABLE 3-27
Derived Emission Factors for the Burning
of Creosote Treated Railway Ties
Compound
[PAH/TSP] X 1000
Emission Factor
(g/MT ties)
Total suspended
particulate
.
10,600
Acenaphthylene
0.62
5.7
Acenaphthene
2.88
27.1
Phenanthrene
0.09
0.9
Anthracene
0.02
0.2
Fluorene
0.29
2.8
Pyrene
0.36
3.3
Chrysene
0.18
1.7
Benz[a]anthracene
0.88
8.3
Benzo[b]fluoranthene
0.54
5.1
Benzo [k]fluoranthene
0.16
1.5
Benzo [a] pyrene
0.59
5.5
Dibenz[a,h]anthracene
0.45
4.2
Benzo[ghi]perylene
0.09
0.9
3-52
Discussions with Canadian National environmental staff suggest that
majority of the non-usable railroad ties are now disposed in secured
landfill sites and only a small percentage is burned. Furthermore, the use
of burning as a means of disposal is to be phased out in the early 1990's.
3.3.6.2 Forest Rres
Only limited data on PAH emissions from prescribed or uncontrolled
forest fires are available. The most frequently cited work was based on
simulating forest burning conditions in the laboratory by burning various
loadings of pine needles on a metal table equipped to change slope and
to take into account wind effects. The airborne, suspended particulate
matter generated in this manner was collected on a glass fibre filter using
a modified high-volume sampler and analyzed by gas chromatography/
mass spectroscopy (McMahon and Tsoukalas, 1978).
PAH emissions from forest fires are dependent on the type of vegetation
burned; the burn conditions; e.g., back fires, in which the fire perimeter
spreads against the wind; or head fires, in which the fire perimeter moves
with the wind; fire intensity and combustion phase (smoldering or
flaming); as well as weather conditions. In general, the majority of
uncontrolled (wild) fires in Ontario are head fires with a small backing
component; however, this ratio is not known (Ward, Ministry of Natural
Resources, 1989).
Emission factors for the polycyclic aromatic hydrocarbons detected in the
simulated burn of pine needles are presented in Table 3-28.
3-53
TABLE 3-28
Emission Factors of Polycyclic Aromatic Compounds
for Burning Pine Needles
(mg/MT fuel, dry weight)
Fire Type & Fuel Loading
Backing
Fires
Heading
Fires
0.5
1.5
2.4
0.5
1.5
2.4
PAH
kg/m^
kg/m^
kg/m^
kg/m^
kg/m^
kg/m^
Anthracene/phenanthrene
12,181
2,189
584
2,525
5,242
6,768
Methylanthracene
9,400
1,147
449
1,057
4,965
7,611
Fluoranthene
14,563
2,140
687
733
974
1,051
Pyrene
20,407
3,102
1,084
1,121
979
1,133
Methyl pyrene/fiuoranthene
18,580
2,466
1,229
730
1,648
2,453
Benzo[c]phenanthrene
8,845
1,808
468
244
142
175
Chrysene/benz[a]anthracene
28,724
5,228
2,033
581
543
836
Methylchrysene
17,753
1,891
877
282
1,287
1,559
Benzofluoranthene
12,835
1,216
818
164
129
241
Benzo[a]pyrene
3,454
555
238
40
97
33
Benzo[e]pyrene
5,836
1,172
680
61
78
152
Perylene
2,128
198
134
33
24
46
Methylbenzopyrenes
6,582
963
384
65
198
665
lndenopyrene[1 ,2,3-cd]pyrene
4,282
655
169
-
-
-
Benzo[ghi]perylene
6,181
1,009
419
-
-
-
Total PAH
171,750
25,735
10,249
7,632
16,549
22,787
From U.S. EPA (1987). Data based on McMahon and Tsoukalas, 1978.
3-54
Fuel consumption is highly variable. On prescribed burn sites in Ontario,
total fuel loadings of 0-15 kg/m^ are considered light-moderate, and
greater than 15 kg/m^ as heavy. In general, the bulk of the fine fraction
of this fuel will be consumed and this rarely exceeds 3.5 kg/m^. Clearly,
the bulk of the total fuel loading occurs in the heavy and duff fuels: in the
former, the consumption levels range from 0.1 - 3.5 kg/m^; in the latter,
the consumption is 0.1 - 7.0 kg/m^, depending on the duff type. The
values are approximate (Ward, Ministry of Natural Resources, 1989).
Ontario data for the total coverage of wild and prescribed fires from 1984
to 1988 are presented in Table 3-29. There were an average 1,669 wild
fires with an average coverage of 146,655 hectares and similarly, 44.4
prescribed fires covering an average of 9,714 hectares for the five year
period.
The results from Tables 3-28 and 3-29 for emission factors and total
yearly PAH emissions to the atmosphere from wild and prescribed fires
in Ontario are summarized in Table 3-30.
3.3.6.3 Burning of Agricultural Waste
Prescribed burns also include the burning of waste consisting of leaves
and slash, which has been referred to in the preceding chapter. For
such burns the fuel loading can exceed 15 kg/m^ and the typical
emission factor is the same as for forest fires, Bjorseth and Ramdahl
(1985). No data are available for prescribed burns conducted by farmers
and householders.
3-55
Year
TABLE 3-29
Wildfire and Prescribed Burn Occurrence
in Ontario, 1984 to 1988
Wildfires Prescribed Burns
Number Hectares Number l-lectares
1988
3206
390,705
15
5,255
1987
1923
75,582
59
13,458
1986
1088
145,561
58
14,323
1985
887
1,007
43
10,635
1984
1240
120,420
47
4,901
Data from Ward, Ministry of Natural Resources, 1989.
3-56
TABLE 3-30
Average Yearly Emission Data for
Polycyciic Aromatic Hydrocarbons
from Wild and Prescribed
Forest Rres in Ontario
PAH
Emission
Factor
(g/MT)
Coverage
(ha/yr)
Total
(MT/yr)
TSP
59,000-42,050*
156,369**
221,418-157,808
Total PAH
19.0-22.8*
(20)
156,369**
71.4 -85.5
Benzo[a]pyrene
0.033 - 0.095*
(0.1)
156,369**
0.12-0.36
Notes:
0
Range calculated for 100% head fires and 70%/30% head/back fires, with an average fuel
loading of 2.4 kg/m^
five year average for wild and prescribed fires
values in parenthesis from Bjorseth and Ramdahl, 1985
1976 estimated B[a]P production from wild forest fires in Ontario was 0.0075 MT/yr (Mellon et
al., 1986).
3-57
Estimates for the total TSP, PAH and B[a]P emissions from wild and
prescribed burns, including agricultural waste, have been included in
Table 3-30.
3.3.6.4 Source Markers
Retene (1-methyl-7-ispropylphenanthrene) has been suggested as a
source marker since it is produced as a result of the thermal
transformation of resinous materials in wood, particulary softwood such
as pine and spruce (Ramdahl et al., 1984). Similarly, it is claimed that the
presence abietic acid, which is a precursor of retene, allows the
distinction between smoke from a coniferous forest fire and smoke from
grass or bush fires (Standley and Simonett, 1987). On the other hand,
retene is also found in the ambient air as a result of coal combustion.
Consequently, the recommended source-specific markers for forest fires
are soil corrected potassium salts, carbon isotopes and beta levusan,
Hornig et al. (1985).
3.4 Summary of PAH Emissions to the Atmosphere
A summary table of estimated total annual PAH emissions from some of
the major sources to the atmosphere is provided (Table 3-31). These
data are compared with estimated annual PAH emissions for the United
States.
3-58
TABLE 3-31
Summary Table -
Atmospheric PAH Emissions
Source
Ontario PAH Emissions
(MT/yr) %
U.S. PAH Emissions*
(MT/yr) %
Industrial Production
Coke Manufacturing
Petroleum Cracking
2-4
0.07 - 0.6
0.8
0.1
700
N.D.
11
Power Generation
Coal-fired Plants
Oil and gas boilers
0.3
1
0.1
0.3
1
<0.1
Incineration
•
Municipal Incineration
0.3- 1.4
0.3
50
0.8
Mobile Sources
-
Gasoline & Diesel Traffic
84.3
34
2,170
. 36
Natural Sources
Forest Fires
71.4-85.5 32
1,000
17
Residential
Oil and gas
Wood-burning
(fireplaces & stoves)
1
65-90
0.3
32
700
12**
Total
260
100%
4,620
76%***
N.D. not determined
* From: Bjorseth and Ramdahl, 1985.
** This data includes coal burning in fireplaces and stoves.
*** Other industrial sources contribute the remainder.
4-1
4.0 TERRESTRIAL AND AQUATIC SOURCES AND INPUTS
4.1 Wet/Dry Deposition from the Atmosphere
Direct deposition from the atmosphere, both wet and dry, is probably the
greatest source of PAH to soil and aquatic environments. Major sources
to these media also include runoff, while municipal and industrial effluents
also contribute to loadings to aquatic environments.
4.1.1 Plant Uptake
Plants may be exposed to PAH in the atmosphere and in soil, and thus
may accumulate these compounds from either route. Considerable
research has been carried out to determine PAH accumulation on leafy
plant parts and by plants such as mosses which have a high
bioaccumulation potential. Some information is available on PAH in plant
material in Ontario, primarily from studies by Agriculture Canada on PAH
occurrence in some food items.
The quantity of PAH accumulated by plants from the atmosphere is
largely a function of the surface area to mass ratio of the plant parts
considered. Thus, broad-leaved edible vegetables typically show the
highest PAH concentrations (MOE, 1979; Grimmer, 1983). Thomas et al.
(1984) measured concentrations of B[ghi]P, B[a]P, F and
indeno[cd]perylene in a range of vegetation from an industrial area of
Sweden, and found the highest concentrations in leaf litter, mosses and
lichens (high surface area materials), and the lowest concentrations in
conifer needles (low surface area materials). Thomas (1984) made
concurrent measurements of PAH (1,12-benzoperylene, B[a]P, F) in
atmospheric dust, precipitation and epiphytic mosses in Germany and
4-2
used multiple regression to demonstrate that both dry deposition and wet
deposition were important modes of PAH bioaccumulation.
Plants grown in an atmospheric concentration gradient of PAH have
been found to accumulate PAH in proportion to the degree of
contamination. Larsson (1985) measured the accumulation of 20 PAH
in lettuce and rye grown at varying distances from a highway and found
much greater concentrations in the lettuce than in rye, with decreasing
concentrations occurring with distance from the road (Tables 4-1 and 4-
2). In his review on PAH, Grimmer (1983) also reported that the PAH
content of plant tissues depends on atmospheric PAH content.
The only Ontario study on PAH uptake by plants grown in a suspected
pollution gradient showed plant tissues (grasses, pine needles, pear and
apple leaves) were close to or below the detection limit for B[a]P and
B[k]F based on 10 mg/MT dry weight (MOE, unpublished). The lack of
any apparent PAH gradient or accumulation in this case can be attributed
in part to the high detection limits relative to reported PAH concentra-
tions in plant tissues in PAH-polluted environments (e.g., Thomas et al.,
1984; Larsson, 1985).
Few studies have examined uptake and translocation of PAH from soils
by plants and results of this research are inconclusive. Graf and Nowak
(1986) reported root uptake of several PAH including B[a]P by tobacco,
rye and radishes, while Harms (1975) reported negligible translocation
of B[a]P from roots to shoots in wheat, and Gunther et al. (1976) reported
no translocation of PAH into plant parts after application to the orange
rind. Ellwardt (1977) reported little uptake of PAH from soils by several
crops, while Durmishidze et al. (1974) observed translocations from
leaves to roots and vice versa in several crops.
4-3
TABLE 4-1
Mean PAH Concentrations (mg/MT fresh weight)
in Lettuce Grown at Various Distances
from a Highway
PAH
PAH Concentration (mg/MT)
Distance from Highway
8 m 15 m 25 m 35 m 45 m 65 m
PHEN
4.8
4.6
2.2
2.6
3.4
2.1
A
0.2
0.1
0.1
0.1
0.1
ND
2-MPHEN
1.6
1.5
0.6
1.0
0.8
0.6
MA
0.1
ND
ND
ND
ND
ND
1-MPHEN
1.8
1.4
0.7
1.0
0.8
0.7
F
7.1
5.5
■ 3.2
4.1
3.6
3.8
P
8.6
7.0
4.2
4.6
4.5
3.8 -
B[a]FLN
2.4
1.2
0.7
0.7
0.4
0.2
B[b]FLN
3.3
0.9
ND
ND
ND
ND
1-MP
2.8
1.6
0.7
ND
ND
ND
B[a]A
1.7
1.3
0.9
0.7
0.6
0.4
CHR + TRI
5.1
3.6
2.6
2.5
1.9
1.6
BF's
3.5
2.3
1.6
1.4
1.6
1.3
B[e]P
1.6
1.2
0.9
0.6
0.6
0.6
B[a]P
0.8
0.4
0.5
0.4
0.3
0.3
PER
0.1
0.1
0.1
0.1
ND
ND
IN[1,2,3-cd]P
0.7
0.5
0.4
0.4
0.1
0.3
DBAs
ND
ND
ND
ND
ND
ND
B[ghi]PER
1.7
1.1
0.9
0.9
0.8
0.7
ANTHN
ND
ND
ND
ND
ND
ND
Total PAH
46
34
22
22
20
16
ND - not detected
From: Larsson, 1985.
4-4
TABLE 4-2
Mean PAH Concentrations (mg/MT fresh weight)
in Whole Rye Grains Grown at Various
Distances from a Highway
PAH Concentration (mg/MT)
Distance from Highway
PAH 7 m 15 m 25 m
PHEN
1.5
1.3
1.4
2-MPHEN
0.7
0.5
0.6
1-MPHEN
0.3
0.3
0.3
F
1.3
1.0
0.9
P
2.3
1.8
1.6
B[a]A
0.2
ND
ND
CHR + TRI
0.7
0.4
0.3
BFLN
0.2
0.2
ND
B[e]P
ND
ai
ND
Total PAH
7.5
6.0
5.7
ND - not detected.
From: Larsson, 1985.
4-5
Edwards et al. (1982) reported uptake and translocation of radio-labelled
anthracene from nutrient solution, with the degree of uptake proportional
to concentration in solution. Using radio-labelled A and B[a]A, Edwards
(1985) reported rapid assimilation and retention from solution in bush
bean roots, with assimilation varying directly with PAH level in solution
and rapid translocation of PAH metabolites in the plant. In the latter
study, bioconcentration factors for the parent compound were reported
as 4,613 and 2,515 for B[a]A and A in roots, respectively. No
accumulation of B[a]A occurred in stems, although some accumulation
of A in stems was noted (bioconcentration factor of 1.9). Using soils
spiked with PAH (B[a]P, B[b]F, B[k]F and DB[a,h]A), Wegmann et al.
(1987) also found that PAH accumulated from soil was retained mainly
by root tissues, with the degree of accumulation depending on exposure
concentration. Overcash et al. (1986) measured bioaccumulation of
B[a]A, A and PHEN by corn, wheat, fescue and soybean grown in soils
containing PAH concentrations of 0, 0.1, 1.00 and 10.0 ppm. Linear
regression equations were developed that predicted uptake of PAH from
soil under the experimental conditions imposed. Equations for corn,
wheat seed and soybean seed are presented in Table 4-3.
The degree of uptake consistently occurred in the order A > B[a]A >
PHEN, suggesting that smaller molecules are accumulated more readily
than larger molecules. Based on the most recent of these studies, it may
be concluded that PAH are accumulated via root uptake and that root
tissues may be expected to accumulate the highest PAH concentrations.
4-6
TABLE 4-3
Regression Equations for Plant Uptake
of Polynuclear Aromatic Hydrocarbons
Plant
PAH
Regression Equation
Corn
Anthracene
Logio[dw(ppb) + 1] = 0.108 + 1.137 [logio(rate + 1)]
r^ (correlation coefficient) = 0.95
Soybean Seed Anthracene
Logio [ciw(ppb) + 1] = 0.136 + 0.018 logio(rate)
^ = 0.94
Wheat Seed Anthracene
Logio [dw(ppb) + 1] = -0.044 + 0.012 logio (rate)
r^ = 0.99
Corn
Benz[a]anthracene
Logio[dw(ppb) + 1] = -0.164 + 1 .056 logio (rate)
^ = 0.95
Soybean Seed Benz[a]anthracene
Logio [dw(ppb) + 1] = -0.019 -t- 0.008 logio (rate)
r^ = 0.85
Wheat Seed Benz[a]anthracene
Corn Phenanthrene
rate is not significant
Logio [dw(ppb) + 1] = 0.054 -i- 0.008 logio (rate)
r^ = 0.97
Soybean Seed Phenanthrene
Logio [dw(ppb) + 1] = -0.082 + 0.319 [logio(rate + 1)]
r^ = 0.90
Wheat Seed Phenanthrene
Logio[dw(ppb) + 1] = 0.016 + 0.004 logio (rate)
r^ = 0.98
dw (ppb) - dry weight in ppb in plant/seed
"rate" as PAH concentration in soil (ppm). Note: this is a concentration term.
From: Overcash et al., 1986.
4-7
4.1.2 Biosynthesis
Evidence for the biosynthesis of PAH compounds by organisms is
inconclusive. Some organisms, including certain bacteria, fungi, plants
and some animals, have been shown to synthesize a variety of polycyclic
quinone pigments (Thompson, 1971), which may be transformed to PAH
compounds by diagenesis in the open environment (NRCC, 1983).
Neff (1979) and Harms (1975) examined evidence both for and against
complete biosynthesis of PAH. In some cases of reported biosynthesis,
contamination by PAH external to the experimental system could not be
dismissed (NRCC, 1983). While there appears to be a general
acceptance that some limited biosynthesis of PAH by microorganisms
may occur under certain environmental conditions, there is also
agreement that its significance in the overall PAH budget is very low
(Harms, 1975; Suess, 1976; Grimmer, 1983; Matzner, 1984). On this
basis, it may be concluded that biosynthesis is an insignificant source of
PAH in soils and other natural media in Ontario.
4.1.3 Diagenesis
PAH are formed naturally during carbonization processes, such as coal
and mineral oil formation. Low temperatures (typically less than 150 to
200°C) favour the formation of alkylated PAH slowly in these deposits
over periods of millions of years (Youngblood and Blumer, 1975;
Grimmer, 1983). PAH precursors are slowly transformed into extensively
alkylated and cylcloalkylated forms and unalkylated PAH occur only in
low abundances in these deposits (Blumer and Youngblood, 1975).
Thus, fossil fuels tend to show high ratios of alkylated to unalkylated
forms.
4-8
In diagenically-formed PAH, the PAH profiles typically show marked
differences from those formed during combustion (Grimmer, 1983). For
example, in mineral oil, B[e]P is the predominant benzopyrene, while
B[a]P is not abundant; conversely, in oil or gasoline combustion
products, the ratio of these isomers is about 1:1. Phenanthrene is about
50 times more abundant than anthracene in mineral oil, while the P:A ratio
in automotive combustion gases is about 4:1.
PAH may also form in marine and lake sediments. The molecular weights
and composition of PAH in these mixtures are affected by the source of
PAH precursors and by the depositional environment (Aizenshtaf, 1973).
For instance. Maxwell et al. (1971) reported that PAH may form in anoxic
sediments through dehydrogenation, dehydroxylation and aromatization
of polyhydroxy-quinone pigments. Conversion of carotenoid pigments
from marine sediments to PAH under low temperatures (65 to 200°C) has
been observed over short time frames (two months) (Ikan et al., 1975).
Other diagenetic pathways for specific PAH are noted in a recent review
NRCC (1983).
The role of diagenesis of PAH in the overall budget of PAH in the open
environment remains unresolved. Diagenesis has occurred over geologic
time scales in Ontario in oil deposits, such as in southwestern Ontario
and probably occurs over shorter periods in lake sediments. The
diagenesis of PAH in soils has apparently not been reported; although,
as in lake sediments, the formation of PAH in soils under certain
conditions cannot be discounted.
4-9
4.1.4 Wood Preservation
There are 17 wood preserving plants in Ontario, six of whicli use organic
wood preservatives (Beak, 1987). Of these six, three use creosote, which
typically contains high concentrations of PAH. PAH are also found in
fuel oils that are usually used a solvents for facilities that use only
pentachlorophenol (POP) as the preserving agent (U.S. EPA, 1986),
although creosote probably represents a greater potential source of PAH
to the environment than do POP solvents.
Creosote is a distillate of coal tar used extensively in wood preservation.
There are three wood preserving plants in Ontario that use creosote - one
at Thunder Bay, one at Trenton and one at Newcastle (presently closed).
An estimated 20,000 tonnes of creosote are used annually by the ten
wood preserving plants in Canada (K. McKellar, Department of Regional
Industrial Expansion, pers. comm.). Based on the number of wood
presen/ers in Ontario and in Canada, an estimated 20 to 30% of the total
creosote usage, or 4,000 to 6,000 tonnes, takes place in Ontario.
The detailed composition of creosote used in Canada has apparently
never been determined (Ralph, C, Agriculture Canada, pers. comm.),
although Uthe (1979) provided data on creosote, coal tar and wood
preservative sludges (Table 4-4).
Berard and Tseng (1986) reported PAH concentrations in surface soils
and groundwater at the Northern Wood Preservers plant in Thunder Bay,
as shown in Tables 4-5 and 4-6. This PAH contamination of the soil and
groundwater at the site can probably be attributed to routine operational
spillage and losses from storage areas.
4-10
TABLE 4-4
PAH Content of Creosote, Creosote
Sludge and Coal Tar
Compound
Creosote
Creosote
Creosote
Sludge
Coal Tar
Wood
Preservation
Sludge
Anthracene
43
18
.
Benz[a]anthracene
-
-
31
43
(2)
5
Benzo[b]chrysene
-
-
23
(4)
5
Benzo[j]fluoranthene
-
-
1
0.3
Benzo[k]fluoranthene
-
-
33
7
Benzo[ghi]perylene
-
-
18
(5)
-
Benzo[a]pyrene
-
13
33
3.6
Chrysene
-
-
-
2.5
Fluoranthene
5.5
-
70
-
5(3)
Perylene
-
-
-
54
26
Phenanthrene
186
125
193
3
-
Pyrene
2.6
-
64
81
(1)
15(1)
Acenaphthene
60
70
67
47
24
Fluorene
103
53
48
-
-
2-Methylanthracene
9.8
-
-
27
6
9-Methylanthracene
2.4
-
-
-
8
Benzofluorene
1.2
■
■
"
•
(1) Phenanthrene and anthracene.
(2) Contains chrysene, triphenylene and all benzanthracenes.
(3) Contains chrysene and triphenylene.
(4) Contains benzo[a]chrysene and phenylenepyrene.
(5) Contains benzo[ghi]perylene and anthanthrene.
From: Uthe, 1979.
4-11
TABLE 4-5
Northern Wood Preservers Survey
Soil and Sediment
Extractable Organics by GC/MS
Concentration
ug/g dry weight
Acenaphthene
117
Acenaphthylene
6
Anthracene
53
Phenanthrene
1292
Benz[a]anthracene/chrysene
164
Benzo[b & k]fluoranthene
71
Fluoranthene
378
Fluorene
213
Pyrene
242
Benzo[a]pyrene
-
Carbazole
40
Quinoline
69
not detected
From: Berard & Tseng, 1986.
4-12
TABLE 4-6
Northern Wood Preservers Survey
Groundwater
Extractable Organics by GC/MS
Well
#7
(ug/L)
Well
#8
(ug/L)
Well
#12
(ug/L)
Well
#15
(ug/L)
Acenaphthene
<5
<70
12
Anthracene
<3
<2
Fluoranthene
17
<5
<4
Pyrene
8
<5
No phenanthrene, benz[a]anthracene, chrysene, benzo[b&k]fluoranthene,
fluorene, acenaphthylene or quinoline were detected.
From: Berard & Tseng, 1986.
4-13
These contaminated soils and groundwaters may be expected to result
in some locally significant PAH loadings into nearby surface waters,
through erosion and groundwater flow. The Ministry of the Environment
has now issued a Control Order to Northern Wood preservers for clean
up of PAH contamination in Thunder Bay Harbour. Information on sludge
generation and PAH levels at other major wood preserving facilities in
Ontario is not readily available.
4.1.5 Sewage Sludge Disposal
Disposal of sludges generated in sewage treatment poses a disposal
problem that is generally addressed through incineration or land disposal
in Ontario. Open water disposal of sludges is not practiced in Ontario.
Incineration of sludges is expensive in terms of energy consumption, but
achieves substantial reductions in volume and thereby facilitates disposal.
Use of sewage sludge as a fertilizer on farmland is becoming an
increasingly attractive disposal option and is widely practiced in Ontario
following the Ontario Ministries of Agriculture and Food, Environment and
Health guidelines for sludge utilization (OMAF/OME/OMH, 1986).
Approximately 1 .25 x 10^ m^ of wet sludge, or nearly 20% of the 6.5 to 7
x 10 m of sludge generated annually in Ontario, is spread on agricultural
land (F. Iliffe, OME, pers. comm.).
PAH levels in sewage sludge depend on loadings into sewage treatment
systems and on system performance. PAH in Ontario sewage systems
were recently investigated in a survey of 37 water pollution control
treatment plants under the provincial Municipal/Industrial Strategy for
Abatement (MISA) program (Environment Ontario, 1988). A summary of
these data is presented in Tables 4-7 and 4-8.
4-14
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4-15
TABLE 4-8
Summary of Ontario Water Pollution Control Plants Tested,
Flow Rates, PAH Concentrations and Estimated PAH Emission Rates
Average
PAH
Estimated
Test Plants
Flow Rate
Concentration
Emission Rate
(10^kL/day)
(ug/L)
(kg/yr)
Guelph
43.4
4.3
68.1
Secondary Plants
4.8
Belle River
5.6
9.8
Brantford
52.1
91.3
Burlington
67.0
117.4
Grimsby
13.1
23
Hamilton
306.5
537
Kingston
18.0
31.5
Kitchener
70.6
123.7
London (Greenway)
110.8
194
London (Potterburg)
16.3
28.6
Mississauga (Clarkson)
74.7
130.9
Mississauga (Lakeview)
256.9
450
Moore
2.2
3.9
Niagara Falls
58.2*
102
Oakville
13.5
23.6
Paris
2.5
4.4
Peterborough
50.8
89
Pickering
176
308
Sault Ste. Marie
6.65
11.7
Sudbury
49
85.8
Toronto (Highland Creek)
170
298
Toronto (Humber)
403
706
Toronto (Main)
767
1,344
Toronto (North)
36.6
64.1
Waterloo
46.4
81.3
Wallaceburg
6.8
11.9
Whitby
3.6
6.3
Windsor
32.8
57.5
* Design flow rate: actual flow rate was not recorded.
From: Environment Ontario, 1988.
4-16
TABLE 4-8 (cont'd)
Summary of Ontario Water Pollution Control Plants Tested,
Flow Rates, PAH Concentrations and Estimated PAH Emission Rates
Average
PAH
Estimated
Test Plants
Flow Rate
Concentration
Emission Rate
(lO^kL/day)
(ug/L)
(kg/yr)
Primary Plants
7.05
Cornwall
43.7
16.7
Kingston (City)
63.5
24.3
Ottawa
400
153
Sarnia
54
20.7
Sault Ste. Marie
(East)
32
12.3
Thunder Bay
81
31
Windsor (Westerly)
123.6
47.4
Lagoons
1.08
Lindsay
14.2
5.6
Niagara-on-the-Lake
6.4
2.5
Total
3,678.5
5,316.3
* Design flow rate: actual flow rate was not recorded.
From: Environment Ontario, 1988.
4-17
The estimated annual PAH emission from the 37 plants for 1987 was
5,316 kg (Table 4-8). The estimated total flow from these plants was
3,678,000 kL/day or 73.6% of the total Ontario flow from the 412
municipal treatment facilities for the same year. A rough estimate of the
total annual PAH loading from these facilities is 5,316/0.736 = 7.2 MT.
Zukovs et al. (1984) provided quantitative data on PAH partitioning in
aqueous and solid phases of Hamilton sewage, but did not analyze PAH
in the digested sludges. PAH concentrations were, however, measured
in primary treatment solids and in waste activated sludge, which would
subsequently pass through further treatment (digestion and possibly
dewatering) followed by disposal (incineration at Hamilton, land-base
disposal at many other plants). Data on total PAH loadings to the plant
in raw sewage and to Hamilton Harbour in treated sewage are available
(Table 4-9). Assuming no biodegradation of PAH in the sewage plant, the
difference between the influent and effluent loadings in Table 4-9 would
represent PAH loadings into the solids disposal pathway.
Unpublished data from Grimmer (cited in Grimmer, 1983) and an earlier
study by Borneff and Kunte (1967) show no evidence for biodegradation
of PAH in a European sewage plant, although effective biodegradation
has been demonstrated in soil systems (Bulman et al., 1985).
Webber and Lesage (1987) analyzed PAH in sewage sludges from
several Canadian cities. These sludges were thought to represent worst-
case sludges in terms of contamination by organics, because they were
obtained from industrial centres. A summary of their data, in terms of
concentration ranges and frequencies of occurrence, and a comparison
with sludges from the U.S. are summarized in Table 4-10.
4-18
TABLE 4-9
Loadings of PAH to the Hamilton Sewage Treatment Plant
in Raw Sewage and from the Plant in Treated Effluent
Contaminant
Trace Organics
In
(kg/yr)
*
Out
(kg/yr)
Acenaphthylene
613
4.1
Fluorene
1,526
19.0
Fluoranthene
4,066
65.7
Carbazole
2,310
43.8
Pyrene
3,705
84.0
Benzo[a]pyrene
4,420
69.4
* These estimates are based on average values from 14 sampling
days, and include very high values measured on one or two days
for most contaminants. Thus, these annual averages are considered
to be high estimates.
From: Zukovs et al., 1984.
4-19
TABLE 4-10
Summary of Canadian and International Data
on PAH in Municipal Sewage Sludge^
15 Canac
Jian
40 U.S.A.
Hamilton
Sludges'
)
Sludges
Dec.
1981 -Jan. 1982
Occurrence
Occurrence
Concentration
(
voncentration
Concentration
Compound
Range
%
Range
Median
%
Range
Acenaphthene
4-6
33
t-<3
t
5
<1-115
Acenaphthylene
42-47
27
t-5
t
1
<1-8
Anthracene
141-599
20
t-32
1
48
< 1-250
Phenanthrene
53
t-36
2
53
< 1-250
Benz[a]anthracene
n
7
<0.5
27
<1-38
Chrysene
39-60^
27
t-23
1.5
31
<1-38
Benzo[b]fluoranthene
7
0.5
Benzo [k [f luoranthene
40-43^
13
0.5-9
5
-
Benzo[ghi]perylene
t-42
13
t-0.3
0.2
Benzo[a]pyrene
28-34
13
4-7.2
5.6
5
<1-12
Benzo[e]pyrene
Dibenzo [a, h] anthracene
7
13
Fluoranthene
232-334
53
t-33
2
44
< 1-250
Fluorene
98-115
60
t-3
2
6
1-32
lndeno[1 ,2,3-cd]pyrene
t-38
7
7
2
<1-2
Naphthalene
23-45
60
t-5.8
1
34
<1-130
Pyrene
171-236
67
t-29
3.5
53
<1-43
t trace
1 all concentrations in
mg/g dry sludge
2 excluding data on Hamilton sludges prior to October 1983
3 Benz[a]anthracene and chrysene combinec
4 Benzofblfluoranthene and benzofklfluroantt
1
lene com
bined
From: Webber and Lesage, 1987.
4-20
They also concluded that PAH concentrations were highest in Hamilton
among Canadian sludges, apparently due to the local steel industry, but
that differences in PAH concentrations among other sludges were
relatively small. Higher PAH concentrations were reported for PAH in
Hamilton prior to the onset of the economic recession in 1982, possibly
due to a reduction in industrial activity (Table 4-10).
The average solids content of wet sludge, as it is landspread, is about
3.5%. If PAH concentrations in Ontario sewage sludges are similar to
those reported by Webber and Lesage (1987), then an average PAH
profile for wet sludge in Ontario could be constructed. For example, the
average B[a]P content of dried Canadian sludge is 5.6 mg/kg and 1.3
mg/kg for Ontario treated sludge (from Tables 4-7 and 4-10), which is
equivalent to 0.05-0.2 mg/L on a wet sludge basis. In the 6.5 to 7 million
cubic metres of sludge generated in the province, there would be an
estimated 337 to 1350 kg of B[a]P, of which about 60-250 kg would be
applied to farmland; the remainder being incinerated or disposed of in
landfills. Application rates to farmland, following the OMAF/OME/OMH
(1986) application guideline of 135 kg of ammonium plus nitrate nitrogen
per hectare per five-year period, assuming a typical nitrogen content of
about 500 ppm for digested sludge, would thus be about 13-54 g ha"^ of
B[a]P over five years.
4.1.6 Disposal of Oil Refinery Sludges
Oil refinery sludges have been treated in a variety of ways in the past.
Landspreading of oil sludges has been practiced since the 1950s, but
came into common practice only more recently. Alternatives to
landspreading including landfilling, incineration, lagooning and
solidification. Landspreading of petroleum sludged is practiced to
4-21
immobilize the sludge mass within the upper soil layers and to allow
biodegradation of hydrocarbons while preventing releases of harmful
vapours, runoff and leachate (Brown et a!., 1980).
To maximize biodegradation, treated soils are frequently tilled to maintain
aerobic conditions. Fertilizers are also applied in many cases. Sites used
for landspreading are typically owned or controlled by the refineries
generating the oily wastes and are not used for other purposes. Reviews
of landspreading practices for oil sludges are provided by Beak (1981)
and Canviro (1983).
Various types of refinery sludges are generated at Canadian facilities.
PACE (1980) identified 10 categories for Canadian oil refinery sludges:
desalting sludge, A.P.I, sludge, flotation froth, biosludge, basin settlings,
storm silt, filter backwash, slop emulsions, cooling water tower sludge
and unleaded tank bottom sludge. These sludges vary greatly in
composition, from those containing very low hydrocarbon contents, to
those with high hydrocarbon contents.
Rates of refinery sludge application onto soils vary with site conditions
(soil characteristics, climate, etc.). CONCAWE (1980) suggested a
maximum rate of application of oily constituents of 15 kg/m^, subject to
site limitations. Beak (1 981 ) noted that reapplication is normally practiced
only when the oil content in the surface soil has decreased to 1 -4%. Data
on overall loadings of oily sludges in refinery landfarms in Ontario or
Canada have apparently not been compiled, although limited data on
individual refineries were provided in the Beak (1981) report, showing
limited data on individual refineries and showing typical total annual rates
of a few hundreds to a few thousands of tonnes of sludges per year per
refinery. Among Canadian refineries, landspreading is the second-most
4-22
important disposal method (on a dry weight basis) for refinery wastes,
with landfill disposal being the most important PACE (1980).
PAH concentrations in oil refinery waste sludges were determined by Can
Test (1982). More recent analyses of PAH levels in sludges from a
Canadian refinery having the "best practicable treatment technology" were
undertaken by Beak (1985). Data from these sources are presented in
Tables 4-11. Loadings of some PAH to landfarms could be roughly
estimated using the limited available data. Data on losses of PAH to the
open environment from landfarms through leachate, runoff or volatiliza-
tion, are unknown.
4.1.7 Coal Gasification Wastes
Intera Technologies Limited (1987) conducted a historical survey and site
reconnaissance of coal gasification solid waste disposal sites in Ontario
and identified 41 sites in 36 different municipalities province-wide. These-
wastes were generated through the manufacture of gas from coal or oil
over the period of about 1850 to 1950 for use in street lights, appliances,
furnaces and industrial engines. Sludges, tars and other solid wastes
from these operations are rich in PAH; coal tar may consist of up to 3%
of PAH by weight (Grimmer, 1985).
Many of these waste sites have been identified as presenting a high
potential for release of waste constituents into the local environment.
Data on PAH from the site identified at Port Stanley (D. Veal, OME, pers.
comm.) show that PAH are leaking from the buried coal gas waste
through the local groundwater system and into the local surface water
environment (Kettle Creek).
4-23
TABLE 4-1 1
Concentrations of Base-Neutral Organics
in Oil Refinery Disposal Sludges
Base Neutral Organics, ug/g, dry wt.
Acenaphthene
ND
Trace
Acenaptliyiene
ND
ND
Anthracene
Trace
Trace
2-Methyl Anthracene
Trace
20.2
Benz[a]anthracene
ND
ND
Benzo[k]fluoranthene
ND
ND
Benzo[a]pyrene
ND
ND
Chrysene
19.3
20.8
Fluoranthene
ND
ND
Fluorene
ND
Trace
Phenanthrene
Trace
Trace
Pyrene
ND
Trace
ND - not detected
From: Beak, 1985.
4-24
Similar wastes at Sydney, Nova Scotia are the subject of intense study
by Environment Canada, owing to liigh rates of contaminant release into
the marine environment. The importance of these buried wastes as a
PAH source to local groundwaters and other environments is yet
unknown, although the potential significance of some of these wastes as
major local PAH sources in Ontario should not be discounted.
4.2 Direct Deposition (Wet/Dry) from the Atmosphere to Aquatic Systems
The identification of PAH in sediments and aquatic biota has led many
authors to speculate that the atmosphere deposition of PAH-containing
particles to surface waters and adjacent watersheds is a significant route
of entry for these compounds into the aquatic environment and may be
responsible for much of the background concentration in the absence of
other identifiable sources (NRCC, 1983). Direct atmospheric input
appears to be the major source of PAH to the Great Lakes (Eadie, 1984).
In Lake Michigan, concentrations of total PAH in the surface microlayer
varied from 0. 1 5 to 0.45 ug/L, which represented approximately 1 0^ times
the atmospheric concentration (Strand and Andren, 1980). Analysis of
the surface film of water from the Detroit River showed PAH
concentrations were often 10^ to lO"* above subsurface water samples
(Comba et al., 1985). These results support the suggestion that aerosols
are a major source of PAH and indicate that the microlayer is a repository
until PAH are removed by adsorption and sedimentation (Strand and
Andren, 1980).
4-25
4.2.1 Rainfall as a PAH Source
Trace organics such as PAH exist in the atmosphere in both the vapour
phase and adsorbed to particulate matter. Since atmospheric fluxes of
contaminants to water are a combination of dry and wet deposition
processes, reliable data on vapour and particle-associated concentrations
are required to estimate these fluxes. Unfortunately, atmospheric
sampling methods are inadequate to differentiate between vapour and
aerosol PAH. Estimation of deposition rates to Ontario must rely on
incorporation of data from world-wide studies. It is expected however,
that wet deposition of praticle-bound PAH would dominate (Ryan and
Cohen, 1986; Mackay et al., 1986).
. Eisenreich et al. (1981) reported much lower PAH concentrations for
precipitation in the Great Lakes basin, with concentration ranges between
0.1 and 4.5 ng/Lfor individual compounds (anthracene, phenanthrene,
pyrene, benzo[a]anthracene, perylene and benzo[a]pyrene). The
dominant PAH in rain and snow samples in urban and rural samples
from southern California were PHEN, F and P with total PAH
concentrations reported as 17 to 261 ng/L for urban samples, and 27 to
80 ng/L for rural samples (Kawamura and Kaplan, 1986).
In Rotterdam, van Noort and Wondergen (1985) reported PHEN, F,
B[a]A, B[b]F, B[a]P, DB[a,h]A, B[ghi]PER and IN[1,2,3-cd]P as the
dominant PAH in rainfall at 7 to 180 ng/L each, with P and CHR detected
only sporadically. In the Rotterdam study, it was found that the instanta-
neous rate of deposition declined with the quantity of precipitation, and
removal rate constants on a precipitation amount basis were 1 .46 to 3
mm'\ Deposition rates to the ground surface (and thus to soils or runoff)
for one event were 11.5 to 124.5 ng/m^.mm for individual PAH during a
4-26
for one event were 11.5 to 124.5 ng/m^.mm for individual PAH during a
single precipitation event (Table 4-12).
Ligocki et al. (1985a,b) reported dissolved and particulate concentrations
of several PAH in Portland, Oregon rainfall (Table 4-13), and in ambient
air, and found that, for most PAH, particle scavenging was less important
than gas scavenging during precipitation events. Limited data on PAH
in precipitation samples from Sarnia and Windsor show high concen-
trations of PHEN, P and F (700 to 900 ng/L) at some locations.
Rates of dry deposition for PAH are generally unavailable, but are
included in measurements of bulk deposition onto terrestrial
environments. Bulk deposition rates appear to be unavailable for Ontario.
In rural, forested areas of West Germany, Matzner (1984) measured
annual deposition rates of B[a]P, 1,12-benzoperylene, IN[1,2,3-cd]P and
F through canopy drip, stemflow and litterfall, and reported that
accumulation rates of 385 to 2,720 mg ha'' yr''' (Table 4-14). Much of the
PAH flux to the soil was transferred by litterfall, indicating adsorption of
PAH on leaf surfaces.
Harrison et al., (1985) measured deposition rates of six PAH at varying
distances from a highway, and reported rates of about 1 1 ug m'^ for F,
7 ug m'^wk"^ for B[b]F, and 2 to 5 ug m'^wk'^ for A, B[a]A, B[k]F and
B[a]P. Most of the PAH was deposited within 15 m of the roadway, and
background deposition rates (less than 1 ug m'^wk'^) were found at
greater distances. For reference purposes, it is noted that measurements
of soil concentrations near a highway in the U.K. indicated measureable
PAH fallout from traffic to a distance of at least 100 m from the highway
(Butler et al., 1984).
4-27
TABLE 4-12
Data for the Correlation Between PAH Deposited
and the Amount of Rain
Compound
Correlation
Coefficient
r^
PAH Concentration/
Amount of Rainfall
Ratio
ng/(m^mm)
Phenanthrene
1.000
89.2
Fluorene
0.9987
124.5
Benz[a]anthracene
0.9819
21.8
Benzo[b]fluorene
0.9985
65.5
Benzo[k]fluorene
0.9980
27.9
Benzo[a]pyrene
0.9933
27.3
Dibenz[a,h]anthracene
0.9912
11.5
Benzo[ghi]perylene
0.9962
62.2
lndeno[1 ,2,3-cd]pyrene
0.9955
89.8
From: Van Noort & Wondergen, 1985.
4-28
TABLE 4-13
Particle-Bound and Dissolved PAH Concentrations in
Rainfall Collected in a Residential Area of
Portland, Oregon
Compound
Particle-Bound
Dissolved PAH
PAH Concentration
Concentration
(ng/L)
(ug/L)
5.4
-
24
0.44
14
4.1
90
-
5.1
3.3
30
4.4
48
4.1
39
1.5
3.3
3.6
7.8
3.0
<0.37
2.8
<0.18
0.58
-
3.3
-
Acenaphttiene
Acenaphthylene
Fluoranthene
Phenanthrene
Anthracene
Methylphenanthrene
Fluorene
Pyrene
Benz[a] anthracene
Chrysene
Benzo[e]pyrene
Benzo[a]pyrene
Perylene
Coronene
Means of 7 samples.
From Ligocki et al., (1985a, b)
4-29
TABLE 4-14
Annual Rates of Total Deposition of
PAH in Forest Ecosystems
Deposition Rate (mg ha'^yr'^)
Spruce Beech
lndeno[1 ,2,3-cd]pyrene
Canopy drip
Stemflow
Litterfall
Total deposition
Benzo[ghi]perylene
Canopy drip
Stemflow
Utterfall
Total deposition
Fluoranthene
Canopy drip
Stemflow
Utterfall
Total deposition
412
± 40
426
+.110
0
60
470
+ 50
230
+ 30
882
± 64
716
±118
663
+ 140
524
± 60
0
80
520
+ 80
240
+ 30
1183
±160
844
± 67
880
+ 130
580
±100
0
90
1840
+ 200
710
+ 70
2720
+ 240
1380
+ 120
From: Matzner, 1984.
4-30
Background deposition rates of 29 to 71 ug m'^yr'^ for total PAH in the
Harrison et a!., (1985) study were slightly below background deposition
rates of 100 to 170 ugm'^yr'^ reported by Quaghebeur et al. (1983) in
Belgium, and considerably lower than the approximately 330 to 560
ugm'^yr"^ measured for four PAH by Matzner (1984) in forested German
ecosystems. Assuming an average atmospheric deposition rate of 100
ug.m'^.yr"^ for the Ontario landmass (916,734 km^), the estimated PAH
loading in 91.7 MT/yr.
Eisenreich et al. (1981) used information from a variety of sources to
estimate total PAH deposition to the Great Lakes (Table 4-15). The total
flux of PAH (A, Phen, P, B[a]A, Per, B[a]P) to all the lakes was estimated
to be 484 tonnes per year. Due to the proximity of much of Ontario's
population and atmospheric PAH sources to the Great Lakes, the
estimated aerial fluxes of each compound may not be representative of
most of the Ontario land mass. Deposition rates are much greater near
their sources (urban centres, highways, etc.) (Kawamura and Kaplan,
1986; Harrison and Johnston, 1985; Butler et al., 1984).
Based on core profiles and estimated deposition rates, PAH fluxes to
sediments in the northeastern United States were estimated for various
time periods during the 20th century (Gschwend and Hites, 1981; Hites
and Gschwend, 1982). PAH deposition rates were clearly higher near
urban centres, but the proportions of PAH derived from the atmosphere
were not compared with water-based inputs (Table 4-16).
Since PAH input to the Great Lakes area is thought to be predominantly
from atmospheric sources (Eadie, 1984), it is assumed that the sediment
deposition rates will reflect these atmospheric deposition trends.
4-31
TABLE 4-15
Total Deposition of PAH to the Great Lakes
Compoung
Deposition Rate (MT/yr)
Superior Michigan Huron Erie Ontario Total
Total PAH
163
114
118
51
38
484
Anthracene
4.8
3.4
3.5
1.5
1.1
14.3
Phenanthrene
4.8
3.4
3.5
1.5
1.1
14.3
Pyrene
8.3
5.9
6.1
2.6
1.9
24.8
Benz[a]anthracene
4.1
2.9
3.0
1.5
1.1
12.6
Perylene
4.8
3.3
3.4
1.5
1.1
14.1
Benzo [a] pyrene
7.9
5.6
5.8
2.5
1.8
23.6
Deposition based on PAH concentration in air as reported in the literature.
From: Eisenreich et al., 1981,
4-32
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4-33
Atmospheric deposition is thought to be the source of PAH found in
remote aquatic environments, so demonstrated by Gschwend and Hites
(1981) and Bailey and Howell (1983). Indeed, the chronology of
atmospheric PAH contamination appears to be preserved in the
sediments of remote lakes (Gschwend and Hites, 1981).
Very little data on PAH concentrations in precipitation have been collected
in Ontario. Preliminary data on PAH in rainwater in Sarnia and Windsor
(industrial areas) were provided by J. Marsaiek of the National Water
Research Institute (pers. comm.), as shown in Table 4-17. These data
may be used to estimate wet depositional fluxes of some PAH at these
locations. The current Ontario MOE Air Resources Branch 6-site network
for measuring wet deposition and airborne concentrations of organics,
including PAH will soon address this deficiency.
It has been assumed that the only major sources of PAH to aquatic
environments in rural Ontario are atmospheric. Background deposition
rates are not yet available for Ontario and no Canadian data were found.
A study conducted on the west coat of the U.K. indicated background
bulk deposition of A, F, B[a]A, B[a]F, B[k]F and B[a]P was about 30
ug/mVy (Harrison and Johnston, 1985). Another European study
conducted in inland Belgium resulted in higher flux background rates of
approximately 100 to 170 ug/m^ PAH/year (F, B[b]F, B[k]F, B[a]P,
B[ghi]Per and ln[1,2,3-cd]P (Quaghebeur et al., 1983). Assuming an
average PAH deposition rate to surface water within the land boundaries
for Ontario of 100 ug/mVyr, the loading to inland waters (177,390 km^)
is 17.7 MT/yr in comparison with the loading of 484 MT/yr to the Great
Lakes.
4-34
TABLE 4-17
PAH-Rainwater Concentration Data
Concentration Range (ng/L)
Sarnia - rain Windsor - rain
Acenaphthylene
Acenaphtliene
Fluorene
Phenantlirene
Fluoranthene
Pyrene
<50
<50
<50
<50
50
<50
109 - 683
143
84 - 921
322
80 - 692
577
Source:
Sarnia - at Pollution Control Plant, May - October, 1986, wet-preciptation
only samples.
Windsor- Little River Poll. Control Plant, August - November, 1985, all blanks
less than 50 ng/L; wet precipitation only samples (collector
covered during dry weather).
Detection Limit is 50 ng/L. Extractables only.
From: J. Marsaiek, pers. comm.
4-35
4.3 PAH Uptake in Soils near Industrial Operations
Data on PAH in soils resulting from atmospheric deposition in Ontario
are apparently limited to an unpublished study by the MOE on soil
concentrations near a Sault Ste. Marie steel mill and at a remote area
75 km to the north, and a study by Daisey et al. (1983) on PAH in soils
near a highway in Toronto.
In the Sault Ste. Marie study, soil was collected at two or three depths
between the soil surface and 15 cm below the surface. The three sites
sampled were within about 750 m of the Algoma Steel mill.
Concentrations of ten PAH ranged from 4 to 9,523 ug/kg in the soil
surface, with B[a]A and CHR occurring at the highest levels (Table 4-
18). Concentrations generally declined with depth at two of the three
sites. In general, PAH concentrations in the soil were one to two orders
of magnitude greater near the steel mill than at the control site. The soil
concentration in Sault Ste. Marie is generally within the range of PAH
concentrations reported for urban soils in the vicinity of Birmingham, U.K.
(Table 4-19), although B[a]A and CHR concentrations tend to be
somewhat higher and B [a] P concentrations somewhat lower at Sault Ste.
Marie.
Daisey et al. (1983) reviewed PAH soil concentrations at a site near
highway 401 in Toronto, Ontario. The concentrations of 15 PAH
expressed as ratios with respect to B[e]P were reported with fluoran-
thene and pyrene occurring at the highest levels (Table 4-19). Soil data
were also reported for samples collected at sites near a motorway in the
Midlands, U.K. The ratios reported were generally higher than those for
Toronto.
4-36
TABLE 4-18
Concentrations of Potycyclic Aromatic Hydrocarbons
in Soil Collected at Different Sites
from the Algoma Steel: Sault Ste. Marie Plant
July 7, 1980
PAH
Sample
Depth
Concentrations (\
1m 14 m
ug/kg soil)
19 m
Control
(cm)
Fluoranthene
0-5
625
655
235
9
5-10
446
1000
75
10-15
150
Pyrene
0-5
5-10
10-15
550
546
369
1762
560
235
125
13
Benz[a]anthracene
0-5
5-10
10-15
7250
4658
2087
9523
9250
2720
960
159
Chrysene
0-5
5-10
10-15
6750
4600
2330
9523
7750
2480
1200
168
Benzo [k]f luoranthene
0-5
5-10
10-15
240
134
57
238
360
68
25
4
Benzo [a] pyrene
0-5
5-10
10-15
482
313
141
310
650
138
54
5
Dibenz[a,h]anthracene
0-5
5-10
10-15
40
33
16
48
88
4
4
ND
Benzo[ghi]perylene
0-5
5-10
10-15
243
187
200
429
225
57
44
5
Anthanthrene
0-5
148
156
30
ND
5-10
89
320
13
10-15
37
ND = not detectable
From: MOE unpublished data, Air Resources Branch, Phytotoxicity Section.
4-37
TABLE 4-19
Ratios of the Concentrations of PAH
to Benzo[e]Pyrene in Soil Near Roadways
Compound
PAH/B[e]P
PAH/B[e]P
Location A
Location B
0.4
ND
0.4-1.1
ND
1.2-1.8
1 .2-3.7
0.9-1.9
0.8-2.5
0.8-2.9
0.8-2.0
0.8-2.3
1.1-1.8
0.8-1.0
ND
0.5-0.6
ND
1.0
1.0
0.9-1.1
0.7-1.4
0.3-0.8
ND
1.2-1.3
ND
0.7
ND
0.1-0.7
0.1-0.3
0.2-0.3
ND
Anthracene
Phenanthrene
Fluoranthene
Pyrene
Benz[a]anthracene
Chrysene
Benzo[b]fiuoranthene
Benzo[k]fluoranthene
Benzo[e]pyrene
Benzo[a]pyrene
Perylene
Benzo[ghi]perylene
lndeno[1 ,2,3-cd]pyrene
Coronene
Anthanthrene
ng B[e]P/g soil
95-745
363-2,293
A - Sannples collected near Highway 401 in Toronto, Canada.
B - Samples collected near Midlands motorway interchange with heavy traffic in England.
From: Daisey et al., 1983
4-38
The Midlands values also correlate well with other Midlands studies
carried out in Birmingham, U.K., presented as the last column in Table
4-20. A variety of typical urban sources, principally automotive traffic, was
implicated as the major PAH source to soils in Birmingham.
Based on a review by Grimmer (1983), B[a]P levels in soils range widely,
from less than 0.02 ug/kg in most areas of Iceland where very little fossil
fuel is burned, to 200 mg/kg near an oil refinery in the U.S.S.R. In
general, urban soils appear to have 100 to 1,000 ug/kg of B[a]P, with
higher concentrations occurring near traffic sources, airports, railroad
stations and areas of heavy industry. B[a]P concentrations in Sault Ste.
Marie soils appear to be typical of concentrations reported for urban soils
from other areas. Background soil concentrations of 5 ug/kg of B[a]P
75 km north of Sault Ste. Marie are probably representative of
concentrations in remote areas of Ontario and may reflect fallout from
natural sources such as forest fires.
4.4 Municipal Effluents
Zukovs et al. (1984) conducted an extensive evaluation of the Hamilton
STP which receives a major industrial waste component. He showed that
95 to 100% of 8 PAH measured in the influent were removed through
deposition with sewage sludges. Effluent concentrations averaged 0.04
to 0.8 ug/L (Table 4-21). Total environmental loadings of PAH to
Hamilton Harbour were estimated to be 286 kg PAH per year. The PAH
profile is presented in Table 4-22 and may be compared with the Ontario
MISA data obtained on 37 WPCPs and summarized in Tables 4-7 and 4-
8.
4-39
TABLE 4-20
Concentrations of PAH in Surface Soil
at Different Sites in the
Vicinity of Birmingham, U.K.
Compound
PAH Concentration (ug/kg surface soil)
PAH/
Control 1m 100 m 500 m 600 m 700 m 730 m 940 m 1 m
B[e]P
Pyrene 293
Fluoranthene 417
Benz[a]anthracene 290
Chrysene 566
Benzo[a]pyrene 356
Benzo[e]pyrene 363
Coronene 32
4515 1828 1057 1825 300
3734 1106 1138 2636 275
3297 1868 939 1234 459
2703 1686 497 1283 365
3196 758 442 523 165
2293 947 461 719 159
322 122 121 151 58
145
1122
2.0
200
996
1.6
169
1065
1.4
251
972
1.2
170
657
1.4
164
745
1.0
57
70
0.1
From: Butler et al., 1984
4-40
TABLE 4-21
Average Concentrations of PAH
Discharged from the Hamilton WPCP
Contaminant
Average
Concentration
(ug/L)
Concentration
Range
(ug/L)
0.04
0.0-0.54
0.19
0.0-2.05
0.61
0.05-3.13
0.41
0.0-1.07
0.80
0.0-4.96
0.62
0.0-2.70
Acenaphthylene
Fluorene
Fluoranthene
Carbazole
Pyrene
Benzo[a]pyrene
From: Zukovs et a!., 1984.
4-41
TABLE 4-22
Estimated Output of PAH for
Hamilton Waste Treatment Plant*
Contaminant Out
(kg/yr)
Acenaphthylene 4.1
Fluorene 19.0
Fluoranthene 65.7
Carbazole 43.8
Pyrene 84.0
Benzo[a]pyrene 69.4
Total PAH** 286.0
* These estimates are based on average values from 14 sampling
days and include very high values measured on one or two days for
most contaminants. Thus, these annual averages are considered
to be high estimates.
Estimate based on 6 PAH compounds.
From: Zul<ovs et a!., 1984.
4-42
Contamination of aquatic sediments with PAH has been attributed to
sewage outfalls. PAH concentrations in sediments were examined at
seven stations along a pollution gradient down current from the Los
Angeles County Sanitation District's sewage outfalls in California. Data
are summarized in Table 4-23.
4.5 Runoff
Surface runoff can contain significant quantities of PAH due to erosion of
contaminated soils and contamination by petroleum residues on
roadways (Hermann, 1981). Marsaiek and Schroeter (1984) measured
concentrations of PAH in runoff in 12 urban centres in Ontario to the
Great Lakes (Table 4-24). Runoff water levels of each PAH measured
1 .0 ug/L or less, while the sediment component contained concentrations
of 0.43 to 2.4 ug/g. Mean total PAH loading in the Canadian Great Lakes
basin from runoff was estimated at 7.7 tonnes per year (Table 4-25).
Greater loads of PAH (as mass/drainage area/year) have been measured
in runoff from highway and industrial land uses in comparison to
commercial and residential areas (Hoffman et al., 1984).
PAH measurements made in runoff from four different urban land use
areas of Rhode Island were similar in terms of percentage chemical
distribution. Fluoranthene and pyrene were most prevalent in runoff while
benz[a] anthracene was notably elevated in the industrial runoff area
(Hoffman et al., 1984).
4-43
TABLE 4-23
Concentration of PAH in Palos Verdes Shelf and
Santa Monica Bay Sediments
Compounds
PAH Concentration Range''
(ug/kg dry weight)
Acenaphthylene
Anthracene
Fluoranthene
Phenanthrene
Benzo [a] anthracene
Benzo[b]fluoranthene
Pyrene
Chrysene
14-160
35-623
92-294
290
1330
338-633
118-838
606
Base/neutral compounds.
From: Swartz et al., 1985
4-44
TABLE 4-24
Best Estimates of Mean Concentrations of
Poiyaromatic Hydrocarbons in Urban Runoff
Water
Sediment
Parameter
No. of
Samples
Freq.
%
MeanNo. of
ppb Samples
Freq.
%
Mean
ppm
Quinoline
53
4
1.0
88
17
0.530
Acenaphthylene
53
17
0.96
88
10
0.700
Acenaphthene
53
13
0.97
88
1
0.450
Fluorene
53
2
1.0
88
8
0.490
Phenanthrene
53
6
1.0
88
30
1.700
Fluoranttiene
53
13
1.0
87
37
2.400
Pyrene
48
17
1.0
86
28
2.200
From: Marsaiek and Schroeter, 1984.
4-45
TABLE 4-25
Annual Loadings of Polyaromatic Hydrocarbons in
Urban Runoff in the Great Lakes Basin
Parameter
Sub-Basin Annual Loadings (kg/yr)
Erie Huron Ontario St. Clair Superior
Whole
Basin
Loading
Quinoline 96
77
385
Acenaphthylene 94
76
378
Acenaphthene 92
75
370
Fluorene 95
77
383
Phenanthrene 108
87
437
Fluoranthene 115
93
467
Pyrene 113
91
459
56
55
55
56
63
68
66
10
9
9
10
11
12
12
624
612
601
621
706
755
741
Total PAH n 1179 954 4753
695
119
7700
C) including contributions for naphthalenes, etc.
From: Marsaiek and Schroeter, 1984.
4-46
Sixteen PAH in runoff were measured in 28 U.S. cities and 11 PAH were
identified at concentrations of 0.3 to 12 ug/L (Cole et al., 1984). Highway
runoff and combined sewer overflows were cited as major sources of
aquatic PAH by Ellis et al. (1985). Road surface runoff in Norway
contained total PAH levels of 1 .6 to 1 1 ug/L and a total of 10 g PAH per
km vehicle day was the mean runoff estimate for the whole year (Lygren
et al., 1984). Gjessing et al. (1984) showed, however, that PAH in
highway runoff were largely retained in adjacent soil surfaces and that a
nearby lake was influenced more by atmospheric deposition.
Analysis of nine storm sewer sediment samples along the Niagara River
yielded varied concentrations of PAH up to 47 ug/g (Table 4-26). Total
runoff of selected PAH to the Niagara River was estimated as 25 kg/yr
(Table 4-27).
4.6 Oil Spills
No information has yet been documented concerning PAH loadings to
Ontario waters from oil spills. About 0.5 million gallons (2300 m^) of
petroleum products were reported spilled each year in Ontario (P. Belling,
Spills Action Centre, MOE, personal communication), excluding
occasional larger spills of up to several hundred thousand gallons that
occur from time to time. This includes both land-based spills and spills
into surface waters. Data on quantities of each type of petroleum product
are not available. In addition, the content is dependent on the petroleum
product and therefore will vary considerably. An example of the PAH
profile in used motor oil is presented in Table 4-28.
4-47
TABLE 4-26
New York Storm Sewer Sediment Analysis
of PAH
(Niagara River, Dry Conditions)
Number of
Times
Identified
Number of
Samples
Maximum
Concentration
(ug/g)
Ace
6
9
ND
A
8
9
47
B[a]A
5
9
2.4
B[a]P
4
9
2.5
B[b]F
5
9
ND
B[ghi]Per
1 ■
9
ND
B[k]F
0
9
-
Chr
7
9
3.5
DB[ah]A
1
9
ND
F
4
9
24
Fin
6
9
28
l[1,2,3-cd]P
2
9
ND
Phen
7
9
ND
P
8
9
22
ND - no data
X
From: IJC,
1987a.
4-48
TABLE 4-27
Total Loadings of PAH to Niagara River
in Runoff (Water Plus Sediment)
PAH Best Estimate
(kg/yr)
Acey 0.74
Ace 0.88
Fin 0.67
Phen 3.2
F 4.3
P 15
Total 24.8
From: Niagara River Toxics Committee, 1984.
4-49
TABLE 4-28
Polycyclic Aromatic
Hydrocarbons
in Used Motor Oil
PAH
Concentration
(mg/L Oil)
Fluorene
1.5
Phenanthrene
7.8
Anthracene
0.3
Methylphenanthrenes
9.2
Fluoranthene
4.4
Pyrene
6.7
Benzofluorenes
2.8
Benzo[c]phenanthrenes
0.1
Benz[a]anthracene
1.1
Chrysene & triphenylene
i 2.5
Benzo[k]fluoranthene
1.4
Benzo[e] pyrene
1.7
Benzo [a] pyrene
0.4
Perylene
0.1
Benzo[ghi]perylene
1.7
Total PAH*
82.5
The total includes other alkyl-substituted PAH as well as PAH not included
in the priority list for this document.
From: NRC, 1983.
4-50
4.6.1 Refinery Losses to Water
Losses from refineries as effluent to the aquatic environment may also
occur. An estimate of such losses in Ontario is presented in Table 4-29.
The total aquatic loading of PAH is estimated to be 80 kg/yr for all
Ontario from refineries (PACE, 1987).
4.6.2 Gasoline and Diesel Oil Loss During Tank Refilling
All gasoline and diesel fuel oils contain PAH compounds. For example,
a commercial diesel fuel has been found to contain the following:
phenanthrene (202 mg/L); anthracene (1.9 mg/L); pyrene (50.2 mg/L);
B[b]F (13.8 mg/L); and B[a]P (1.9 mg/L) (Obuchi et al., 1984).
No data are available for gasoline or diesel oil loss during refilling of
vehicles. If approximately 0.5 mL were lost and the average tank capacity
is 40 L, then a conservative estimate of the annual loss would be
0.0005/40 of the total yearly gasoline utilization, that is, 1.25 x 10"^ x
8,526,390 = 106.6 kL/yr. For a total PAH content of 270 mg/L and a
B[a]P content of 1.9 mg/L this corresponds to a discharge to soil in
Ontario of 28.7 kg PAH/yr and 0.2 kg B[a]P/yr.
4.6.3 Waste Oil from Transportation
Although no data on oil spills from mobile sources are available for
Ontario, a rough estimate can be made based on the ad-hoc assumption
that about 1 L of oil is spilled per year by each fit and active vehicle
operating in Ontario. From this assumption, it can be inferred that for a
fleet of 5,807,720 vehicles including automobiles, light trucks and two
stroke engines, that approximately 5,800 kL are lost annually to the soil.
4-51
TABLE 4-29
Estimated PAH Losses from Ontario
Refineries in Wastewater^
PAH
Total Load'
(kg/yr)
Acenaplithene
Acenaplithylene
Anthracene
Benzo [a] anthracene
Benzo[k]fluoranthene
Benzo[a]pyrene
Fluoranthene
Fluorene
Phenanthrene
Pyrene
Chrysene
0.0
0.0
0.85
2.4
0.42
0.0
0.0
14
31
8.9
21
Total PAH
78.6
From total crude oil transformed to refined products of
26,685,100,000 L or 133,425,500 barrels in 1986.
Concentration, flow and BPSD (barrels per stream day) data from
PACE Report No. 80-4.
4-52
Assuming an average PAH concentration of 82.5mg/L and a B[a]P
concentration of 1 mg/L, it is estimated that 478 kg PAH and 5.8 kg
B[a]P are spilled annually.
4.6.4 Disposal of Dredging Spoils
It is likely that the deposition of dredging spoils in the aquatic environ-
ment results in no net PAH loading change for the province, but involves
instead the transport of PAH from one location to another. Because har-
bours and river mouths tend to be areas of contaminant deposition and
areas where considerable dredging is required, the quantities of PAH in
dredged sediments in Ontario waterways may be substantial.
4.7 Leachate from Waste Disposal Sites
Leachate contamination from localized PAH sources has been identified.
A simulation of rainfall runoff from model coal piles, for example, demon-
strated that many PAH can be released, although most estimated
individual PAH concentrations were less than 50 ug/L (Stahl, 1984).
Extreme case simulations resulted in individual PAH levels that were less
than 1 ug/L (limit of detection) to 107 ug/L (Table 4-30).
4.8 Treated Wood Structures for Piers
The use of preserved wood in piers and other harbour structures has
been shown to result in PAH contamination of the aquatic environment.
Lake et al. (1979) reported elevated sediment PAH concentrations near
piers treated with coal tar.
4-53
TABLE 4-30
Organic Compounds Identified in Extracts of
Runoffs from Different Coals
Estimated Concentration (ug/L)
Compound
Illinois #6 Coal
Kentu
Acenaphthene
1
Fluorene
5
-
Phenanthrene
65
191
Anthracene
0.6
-
Fluoranthene
3
67
Pyrene
4
-
Chrysene
1
25
Benz[a] anthracene
2
29
Benzo[k]fluoranthene
0.6
-
Benzo [a] pyrene
0.6
30
Concentration refers to that in the original runoff.
- no data
From: Stahl (1984).
4-54
In Atlantic Canada, Eaton and Zitko (1978) found 25.7, 35 and 48 ug/g
F, P and B[a]A, respectively in sediments near a creosoted wharf. No
information was found on PAH contamination of surface waters from
railway ties or utility poles, although it is probable that some PAH losses
to surface water environments also occur from these widely used items.
Total runoff/leachate volume estimates have been made for coal pile sites
at the four operating coal-fired generating stations in Ontario (Table 4-
31). Based on the concentration data presented by Stahl et al. (1984),
the total yearly PAH losses from coal piles at Ontario generating stations,
are estimated to be in the range of 3.8-46.9 kg/yr (Table 4-32).
Coal and oil gas manufacturing plants operated in 36 Ontario
communities from the mid 1880s until the mid 1950s provide gas for a
local domestic and industrial consumption. Intera Technologies Ltd.
(1987) conducted a reconnaissance of manufactured gas plant locations
to evaluate the potential for adverse environmental effects from buried
wastes from these plants.
Wastes from these plants include tars, sludges, liquors and other gas
cleaning wastes that are rich in PAH. Many sites were qualitatively
assessed as currently impacting off-site properties and water resources,
indicating that these probably represent significant PAH sources to local
surface water environments. Unpublished MOE data on PAH
concentrations in sediments and water downstream from a coal gas
waste site at Port Stanley confirm this assessment (D. Veal, OME, pers.
comm.).
4-55
TABLE 4-31
Estimated Yearly Volumes of Coal Pile Runoff
and Leachate at Coal-Fired
Generated Stations in Ontario
Station^
Coal Pile ,
Area (ha)'
Precipitation
Minus Total Runoff/
Evapotranspiration Leachate
(mm/year)
(myyo
1
2
3
4
13
200
26,000
19
200
38,000
28
200
56,000
6.4
200
12,800
Station identification confidential.
From Dearborn (1982).
Estimated from precipitation and evapotranspiration rates provided by Fisheries
and Environment Canada (1978).
4-56
TABLE 4-32
Estimated PAH Loadings from Ontario
Generation Stations Coal Pile Runoff
Estimated
Concentration
Annual PAH
PAH
in Runoff
Loading*
(ug/L)
(kg/yr)
Acenapthene
1
0.13
Fluorene
5
0.66
Phenanthrene
11-195
1.5-25.9
Anthracene
0.6
0.08
Fluoranthene
3-67
0.4-8.9
Pyrene
4
0.5
Chrysene
1-25
0.1-3.3
Benz[a]anthracene
2-29
0.3-3.8
Benzo[k]fluoranthene
0.6
0.08
Benzo[a]pyrene
0.6-30
0.08-4.0
Total PAH**
3.8-46.9
* Based on a total runoff from 4 coal-fired generating stations of 132,800 kL/yr.
** Based on 10 PAH.
From: Stahl (1984).
4-57
4.9 Transboundary Movement of Polycyclic Aromatic Hydrocarbons
PAH are not as mobile in the aquatic environment as in the atmosphere.
Rapid sorption and sedimentation prevent long range transport, resulting
in elevated concentrations near their source (Eisler, 1987). The transport
distance would be a function of physical processes, such as particle size,
currents and size of the water body, as well as physio-chemical and
biological degradation processes.
PAH occurring in the water column, either in suspended particulate or
dissolved form, may be transported downcurrent. Because PAH usually
occur in low concentrations in surface waters, and because analytical
techniques for quantification of PAH at low concentrations have been of
questionable reliability, relatively little information exists on surface water
transport of PAH.
Environment Canada has an ongoing project to monitor water quality
entering (Fort Erie) and leaving (Niagara-on-the-Lake) the Niagara River,
in order to discern the contributions of toxic substances entering the river
from Lake Erie and from sources along the river. The difference in PAH
flux between the source and mouth of the river represents the PAH
loadings from all sources (industrial, hazardous waste site seepage,
runoff, etc.) along the course of the river. Data are available for 1 1 PAH,
based on analysis of suspended sediment (concentrations were less than
detection limits in the dissolved phase). Table 4-33 shows PAH mean
loading estimates for these compounds from inputs from Lake Erie to the
river, outputs from the river to Lake Ontario, as well as sources along the
river (IJC, 1987a). Urban runoff data for the Niagara River (Table 4-34)
indicate that this source (25.5 kg PAH/yr) represents a small fraction of
the total loadings along the course of the river (14.5 MT/yr).
4-58
TABLE 4-33
Loadings of PAH to the Niagara River
Loadings (kg/day)
Fort
Erie
Niagara
on-the-
Lake
Niagara River
Loading
(by difference)
Acenaphthene
0.05
0.43
0.38
Anthracene
0.40
1.25
0.85
Benzanthracene
1.50
8.05
6.55
Benzo[a]pyrene
1.23
3.78
2.55
Benz[a,h]anthracene
0.57
2.18
0.61
Benzo[b&k]fluoranthene
1.66
9.72
8.06
Chrysene
0.81
4.60
3.79
Fluorene
3.85
11.44
7.59
Fluoranthene
0.13
0.54
0.41
Phenanthrene
2.58
6.07
3.49
Pyrene
3.07
8.58
5.51
Total PAH
15.8
55.6
39.8
Based on analysis of suspended sediment samples from water samples,
December 1984 to March 1986.
From: IJC, 1987a.
4-59
TABLE 4-34
Annual Total Loadings of PAH
in Water and Sediment in Urban Runoff
for the Niagara River
PAH Estimated Loading
(kg/yr)
Quinoline 0.74
Acenaphthylene 0.74
Acenaphthene 0.88
Fluorene 0.67
Phenanthrene 3.2
Fluoranthene 4.3
Pyrene 15.0
Total PAH 25.5
From: IJC, 1987a.
4-60
Bahnick and Markee (1 985) monitored PAH concentrations on suspended
sediment in the St. Louis River at Duluth, and estimated annual loadings
from the river to Lake Superior (Table 4-35). The authors noted that
these local loadings to the lake are very small relative to atmospheric
deposition estimated for Lake Superior by Eisenreich et al., (1981).
Comba et al. (1985) presented detailed data on PAH partitioning in
surface water (particulate and total), surficial sediment, sediment pore
water and in the water surface microlayer in the Detroit River. This
information provides considerable information on partitioning and
transport processes for PAH in the Detroit River. Data collected by the
National Water Research Institute on PAH in the St. Clair River are
presented in Table 4-36 and illustrate the emission profile of 24 industrial
point source discharges (Environment Canada/MOE, 1987); however, no
flow rate data were available to estimate loadings to the river.
4.10 Summary of PAH Emissions to Soil and Water
The annual PAH emission rates in the form of precipitation and as
municipal or industrial liquid discharges as well as in the form of sludge
application to the Ontario ecosystem are presented in Table 4-37. They
indicate that the single largest contributor of PAH to soil and water is
through atmospheric preciptation.
4-61
TABLE 4-35
Input of Particulate-Associated Contaminants
to Western Lake Superior from the
St. Louis River and Duluth Harbor Area
Parameter
Particulate
Concentration
(ug/g)
Particulate Transport
Lake Superior*
(kg/yr)
0.15
0.42
1.04
2.9
0.06
0.17
1.88
5.3
1.76
4.9
Fluorene
Phenanthrene
Methylanthracene
Fluoranthene
Pyrene
Based on a discharge rate of 30 m /sec, an average suspended solid
concentration of 10 mg/L and 30% particulates input to the lake.
From: Bahnick & Markee, 1985.
4-62
TABLE 4-36
PAH Emission Profile ft-om
24 Point Source Discharges
to the St. Clair River
PAH Concentration
(ug/L)
Acenaphthylene
4.14
Acenaphthene
1.17
Fluorene
2.80
Phenanthrene
3.39
Anthracene
0.76
FJuoranthene
0.94
Pyrene
1.26
Benz[a]anthracene
0.34
Chrysene
0.71
Benzo[b]fluoranthene
0.30
Benzo[k]fluoranthene
0.06
Benzo [a] pyrene
0.79
Indenopyrene
0.86
Dibenz[ah]anthracene
0.76
Benzo[ghi]perylene
0.94
Total PAH
19.2
From; Environment Canada/MOE, 1987.
4-63
TABLE 4-37
Estimated PAH Loadings to the Soil and Water
Source and Input
PAH
(MT/yr)
Loadings
%
Global PAH Loading*
%
Atmospheric
Precipitation
109
75
17-42.5
Soil
Land treatment
(sewage sludge)
0.3-1.3
(as B[a]P)
3.7**
0.8-2.0
Crankcase Oil
0.5
0.3
Gasoline Spills
0.03
0.03
Water
51.6-80.4
Water Pollution
Control Plants (WPCP)
7.2
5.0
-
Hamilton WPCP
0.3
0.2
Leachate (coal piles)
0.004
-0.05
0.02
Runoff (to Great Lakes)
7.7
5.3
Urban Runoff to
Niagara River
14.5
10
Petroleum Refineries
0.08
0.6
Biosynthhesis
neglig.
0
1.3-3.4
neglig. - negligibly small
* From: NRCC, 1983. (The estimate in this citation is based on a significant contribution
from marine petroleum spillage which does not apply to Ontario.)
** Estimated by assuming a PAH : B[a]P ratio of 6.7:1
5-1
5.0 ENVIRONMENTAL PROCESSES AND FATE OF PAH
5.1 Introduction (Air)
The literature on the fate of polycyclic aromatic hydrocarbons (PAH) in
the environment is limited. The pathways open to these compounds are
expected to be no different from those available to other predominantly
particulate species in the same size range. The principal reaction proucts
of atmospheric processes which are of interest are the oxy- and nitro-
PAH.
There have been no reports on studies to determine the particle size
distribution of ambient oxy- and nitro-PAH, but several studies have
established that the parent PAH compounds are predominantly in the size
fraction less than 1 um aerodynamic diameter (Pierce and Katz, 1976;
Demaio and Corn, 1966; Butler and Crossley, 1981). Currently, all
indications are that PAH and their oxy- and nitro-PAH compounds would
be expected to have similar size distributions.
The particulate oxy-and nitro-PAH may undergo dispersion over hundreds
of kilometres in the process of long range transport (LRT), deposition and
chemical transformation. The LRT of oxy- and nitro-PAH is expected to
be similar to that of the parent of PAH and such transport is expected to
account for the presence of these compounds in remote areas.
Deposition of oxy- and nitro-PAH compounds is also expected to be
similar to that of PAH. The dry deposition processes should be described
by those of 1.0 um particles which have deposition velocities, V^, of the
order of 0.02 to 0.32 cm/s based on estimates by Sehmel (1980) or
Cawse (1974), as noted by Strand and Andren (1980). Similarly, during
5-2
wet deposition, scavenging ratios for oxy- and nitro-PAH should be no
different from those of particles of 1 .0 urn or less. Because of somewhat
enhanced solubility, the highly oxidized oxy-derivatives (acids and anhy-
drides) may prove to have higher washout ratios. Information on the
presence of these oxy-PAH derivatives in ground water may provide
useful information on the deposition processes to water bodies.
5.1.1 Chemical Reactivity (Atmospheric Aspects)
PAH exist in the atmosphere in the vapour phase or bound to particles.
Their chemical reactivity in the atmosphere is determined not only by the
inherent molecular structure of the PAH but also by the physical and
chemical nature of the substrate on which PAH are adsorbed, on the
concentrations and composition of gases and on the intensity of
absorbed radiation. Chemical reactivity of PAH has been investigated
theoretically and experimentally. The former investigations have been
based on structure-reactivity relationships that have been derived from
molecular orbital calculations or empirical correlations with various types
of reactions.
The atmospheric reactivity of particle-bound PAH is limited by the lifetime
of the particle in the atmosphere. Particles are removed from the
atmosphere by diffusion, sedimentation and wet scavenging processes.
Superimposed on these processes are the photochemical and thermo-
chemical reactions with atmospheric species which transform PAH and
contribute to the determination of their atmospheric lifetimes. The
transformation reactions of PAH, which include nitration reactions,
reactions with oxygen species (ozone, oxygen atoms and excited state
molecular oxygen species), and sulphur oxides, as well as photolysis
5-3
reactions, are described in this section. Rate data are summarized
(expressed as half lives or as relative rates) for these transformations.
The determination of the rates of PAH transformation has relied on experi-
mental conditions which have not always adequately simulated natural
conditions; hence reported rates may be more appropriately viewed in
terms of the relative rates for a series of PAH. Selected studies in which
reaction rates are available for relevant reactions that pertain to the
atmospheric reactions, as well as the occurrence and fate of PAH
products are presented in Table 5-1 . The following sections focus on
those investigations which are most representative of natural atmospheric
conditions based on the half lives for relevant reactions.
5.1.1.1 Particle Lifetime
Since most of the PAH of interest occur adsorbed to particles in the
atmosphere, the lifetimes of particles represent an upper limit for the
atmospheric lifetime of particle-bound PAH. Particle lifetime is very
dependent on the aerodynamic diameter.
Particles in the 1-10 urn diameter range have been estimated to have life-
times of 100-1000 h (Esmen and Corn, 1971). These estimates assumed
a mixing height of 2 km, a monotonic non-increasing flux and a mono-
tonic non-increasing concentration-height function. The lifetime estimates
approximate to the dry deposition process. Consideration of the wet
deposition process was included in estimating the physical residence time
of PAH by Mueller (1984). Laboratory measurements were used to
estimate chemical lifetimes for pyrene (18 h), B[e]P (96 h) and B[ghi]P
(96 h). The overall mean lifetimes (expressed as the inverse of the sum
of the reciprocal physical and chemical lifetimes) were 16 h, 48 h and 6
h for pyrene, B[e]P and B[ghi]P, respectively.
5-4
TABLE 5-1
Summary of Phototytic and Electrophilic
Reactions of PAH and Nitro-PAH
Activator
PAH/Substrate
Products/Comments
Light
Light
Light
Light
Ught
PAH solutions evaporated on
thin layer peth dishes.
A (0.2), B[a]A (4.2), DB[a]A
(9.6), DB[a.c]A (9.2), P (4.2),
B[a]P (5.3), B[e]P (21.1),
B[b]F (8.7), B[k]F (74.1).
Tropospheric lifetimes
A < B[a]F < B[a]A < BO]F
< B[b]Chr < B[a]P < B[k]F <
B[b]F < F < < P measured by
photooxidation on silica gel
substrate.
Effect of heating to 400°C
presampled and clean Hi-Vol
filters. Extract with cyclo-
hexane. Half lives estimated
for PAH in collected
particulate. P (11.9 h),
B[a] (14.4 h), B[a]P(11.6h).
Photolysis of P and B[a]P on
coal ash, alumina, silica gel
and flaked graphite.
Half lives in hours in brackets.
Lane and Katz, 1977.
Photooxidation rates similar
to Lane et al., 1977.
Blau and Gustan, 1982.
Fifteen PAH adsorbed on silica
gel, alumina, flyash and carbon
black. Half lives with flyash
as adsorbent ranged from 29-49
hours.
P and B[a]P stable on glass
fibre filters. Heat enhanced
photodegradation.
Cyclohexane solvent increased
photodegradation of P.
Valeric et al., 1984.
No quantitative data. Thermal
decomposition of B[a]P and
P on substrates negligible.
Yokley etal., 1986.
Behymer and Hites, 1985.
5-5
TABLE 5-1 (cont'd)
Summary of Photolytic and Electrophilic
Reactions of PAH and Nitro-PAH
Activator
Light
HNO2, NO2, HNO
O3 (0.2 ppb)
'3'
PAH /Substrate
Products/Comments
Woodsmoke and sunlight in a Photodegradation slower at
Teflon chamber. With/without -7°C than 20°C.
<30 ppb NO2 and/or <10 ppb Og.Kamens et al., 1986.
Half lives (mins): B[a]A (54
min.), CHR + TRI (196),
B[b]F (232), B[k]F, B[j]F,
B[k]Fand BO]F (156), B[a]P
(69), B[ghi]Per (100).
No degradation of PAH with Lindskog, 1983.
0.1-1.0 ppm HNO2 during sampling.
10-nitro B[a]A, 6-nitro B[a]P,
3-nitro P found for 0.12 ppm
NO, reaction with PAH.
HNO3 leads to significant
degradation of B[a]A, B[a]P,
Per and B(ghi)Per from 20% to
55%.
N2O5
Pyrene, perylene adsorbed on
glass fibre filters. Pyrene
0.3% (12 h) nitrated product
per hour [N2O5] = 0.6 ppb.
Upper limit of 1.8%/h for
[N2O5] = 14 ppb.
No reaction with NO3. Relative
reaction rates with N2O5
different from that with HNO3
in solution. Per < P for N2O5
- converse for HNO3 in solution.
Pitts et al. (1985b).
5-6
5.1.1.2 Reactions with Nitrating Species
The nitration reactions of PAH have commanded much attention in view
of the mutagenic (Rosenkranz and Mermelstein, 1983) and carcinogenic
(Hirose et al., 1984; Ohgaki et al., 1984, 1985; Tokiwa and Ohnishi, 1986)
properties of their nitro-PAH reaction products. The nature and origin of
the nitro-PAH are important in order to assess the potential health
impacts of PAH and nitro-PAH. The nitration reactions have been
investigated using deposits of pure PAH on filters, PAH adsorbed on
various solid particle substrates and in reaction chambers.
The reactions of pure PAH (deposited on filters) with nitrating species
have been studied by several investigators (Pitts et al., 1980b; Tanner
and Fajer, 1983; Tokiwa et al., 1981; Grosjean et al., 1983a). B[a]P
deposited on clean glass fibre filters was transformed to the 1-, 3-, and
6-nitroB[a]P by NOg with traces of HNO3 (Pitts et al., 1978b; Pitts, 1979)
or by HNO3 only (Grosjean et al., 1983b). Similarly, exposures of
perylene to NOj and NO2/HNO3 mixtures yielded nitration products with
significant yields only when nitric acid was present. Earlier studies (Pitts
et al., 1978b; Jager and Hanus, 1980; Hughes et al., 1980; Butler and
Crossley, 1981; Tokiwa et al., 1981; Brorstroem et al., 1983a; Lindskog,
1983) reported reactions of PAH (present in ambient or diesel exhaust
particulates) with NOj, but the losses of PAH should properly have been
ascribed to reaction to traces of HNO3 (Grosjean et al., 1983a).
The nitration by N2O4, HNO2 and HNO3 of 25 PAH in solution was studied
to investigate the reactivity of PAH to nitrating species that are relevant
in atmospheric chemistry (Nielsen, 1984). A classification of various PAH
in which the Class I compounds are most reactive and Class VI least
reactive, was established (Table 5-2). The classification was based on
5-7
the relative rate constants for electrophilic nitration reactions in solution
(Nielsen, 1984; Dewar et al., 1956) as well as on observed correlations
between rates of nitration reactions and spectroscopic properties (the
position of the first p-band), the aromaticity index or the first ionization
potential.
The classification is consistent with several features of the occurrence of
PAH and nitro-PAH in environmental samples. For example, the PAH
compounds in Class I, the most reactive, have not been identified in
environmental samples. Similarly, the compounds in Class II have been
shown to react with NOg and HNO3 under a variety of conditions, whereas
less reactive compounds, e.g., coronene, in Class IV were more resistant
to attack by NOj. The classification, although preliminary, offers a
reasonable basis for anticipating the occurrence of some nitro-PAH in
environmental samples. However, the authors pointed out that the
relative reactivity of PAH implied by the classification may be altered when
photo-activation is involved.
The nature of the substrate (on which PAH are absorbed) alters the
reactivity of the PAH substantially (Korfmacher et al., 1980a). For
example, a different product distribution is found in photoactivated
nitration (Pitts et al., 1983), when compared to nitration in solution, thus
a different mechanism for such reactions is implied. In contrast, the
relative rates of disappearance of several PAH on (dark) reaction with
NO2 (Nielsen et al., 1983a; Tanner and Fajer, 1983) were consistent with
Nielsen's classification scheme. Wehry et al. (1984) made similar
conclusions based on reactions of NO2/HNO3 mixtures and NO2
separately with PAH adsorbed on coal fly ashes, alumina, silica or
graphite.
5-8
TABLE 5-2
Classification of PAH Based on Electrophilic
Nitration Reactions
Class I
dibenzo[a,h]pyrene
Class II
anthracene
benzo[a]pyrene
dibenzo[a,l]pyrene
dibenzo[a,i]pyrene
perylene
Class III
benz[a] anthracene
benzo[g]chrysene
benzo[ghi] perylene
dibenzo[a,e]pyrene
pyrene
Class IV
benzo[c]chrysene
benzo[c]phenanthrene
benzo[e] pyrene
chrysene
coronene
dibenzanthracenes
dibenzo[e,l]pyrene
Class V
acenaphthylene
benzofluoranthenes
fluoranthene
indeno[1 ,2,3-cd]pyrene
phenanthrene
triphenylene
From Nielsen, 1984.
5-9
Since particle-bound PAH present a complex heterogeneous physical and
chemical reaction system, several investigators have examined nitration
reactions of PAH adsorbed on various 'model' substrates. While some
of the laboratory investigations of PAH nitrations on various substrates
have yielded relative rate information that is consistent with the nitration
rates in solution, the rates and mechanisms of the particle-bound PAH
nitration reactions are different from the solution phase nitrations. The
complexity of the particle-bound nitrations make estimation of their
influence on the lifetime of particle bound PAH difficult.
Recent investigations have relied on chamber studies and in the isolation
and speciation of the nitroarene product distribution in order to deter-
mine the rates and mechanisms of nitration reactions, especially in
polluted atmospheres. The relative rates of reaction of six PAH towards
NgOgWas found to decrease in the order pyrene > fluoranthene > B[a]P
> benz[a]anthracene > perylene > chrysene, in contrast to the order
expected on the basis of the nitration reactions in solution. The
mechanism of nitration by N2O5 is therefore different from that in solution
(Pitts et al., 1985a). The nitration of PAH by N2O5 is important in view of
the presence of NgOg in the atmosphere and, therefore, its role in
influencing the atmospheric lifetime of PAH. A nitration rate for pyrene
by N2O5 at night was estimated at 1.8% h'^ (Pitts et al., 1985b).
More recent studies recognize (Pitts et al., 1986; Atkinson et al., 1986;
Pitts et al., 1985b) the importance of N2O5 and NO3 as well as the role of
OH radicals. The nitration of perylene and pyrene by NgOg rather than
by NO3 was demonstrated by Pitts et al. (1986) and earlier studies
implicating nitration by HNO3 exclusively (Grosjean, 1983) were shown to
be due to N2O5 (Pitts et al., 1986).
5-10
Nitro-PAH found in ambient air include 1- and 2-nitronaphthalene, 1-
nitropyrene, 2-nitropyrene, 2-nitrofluoranthene, 3-nitrobiphenyl and 4-
nitrobiphenyl. The two most abundant particulate nitro-PAH in ambient
samples are 2-nitrofluoranthene and 2-nitropyrene, and their presence
has been ascribed to atmospheric nitration reactions (Arey et al., 1987)
rather than to direct emissions from combustion sources (Tokiwa and
Ohnishi, 1986). The nitro-PAH isomers found in the atmosphere are
consistent with the their formation through reactions of parent PAH with
OH radicals in the presence of NO^ (Atkinson et al., 1987; Arey et al.,
1987). The most abundant nitro-PAH found in ambient samples have
been shown to be the more volatile, vapour-phase nitro-PAH such as 2-
nitronaphthalene and 3-nitrobiphenyl (Arey et al., 1987). These isomers
are not those expected from electrophilic nitration reactions of the parent
PAH, but rather, as a result of the atmospheric transformation by
atmospheric nitrating species (OH in the presence of NO^ and NgOg).
Measurements of nitro-PaH at industrial and remote sites showed that 1-
nitropyrene concentrations were higher at the most remote site
(Bermuda) than at sites near industrial sources in Michigan (Gibson,
1986). The ratio of B[a]P to TSP at sites near the sources were consider-
ably higher than B[a]P/TSP ratios at the distant sources which indicates
the importance of the transformation reactions of B[a]P during transport.
The formation of 1-nitropyrene during transport was suggested.
The more volatile PAH are abundant in the atmosphere and their lifetimes
with respect to OH and nitration reactions are expected to be of the
order of 9 h for phenanthrene and 2 h for anthracene (Bierman et al.,
1985; Atkinson, 1986). The formation of nitro-PAH by two pathways, one
involving OH radicals in the presence of NO^, and the other due to
reaction with NjOg, have been shown to be important for 2-
nitrofluoranthene and 2-nitropyrene (Arey et al., 1986). Estimates of the
5-11
half lives of fluoranthene with respect to the OH/NO^ and the NjOg
reactions (under southern California conditions) were approximately 6 h
and 7 weeks respectively. In Ontario urban atmospheres, NO, NO^ and
O3 concentrations of 400, >500 and 90 ppb respectively have been
observed - similar to those in Southern California and ambient concen-
trations of naphthalene (vs 2800 ng/m^), fluoranthene (vs 9.7 ng/m^) and
pyrene (vs 12 ng/m^) for example are similar (see Arey et al., 1987).
Thus the nitroarene concentrations and the half lives of the parent PAH
with respect to these reactions in Ontario air could be similar to those
determined for more southerly latitudes, under certain conditions.
The fate of the nitroarenes - especially those formed in atmospheric
nitration reactions, needs to be considered in the overall assessment of
PAH. The limited information on the photochemistry of the nitroarenes
indicates that quinone products are formed. The photolysis of 9,10-
nitroanthracene (absorbed on silica gel) forms the 9,10-anthraquinone,
while 6-nitroB[a]P forms the 1,6-, 3,6-, and 6,12- isomers of B[a]P
quinones. Analogous photoxidation products are expected for other
nitroarenes. Finlayson-Pitts and Pitts (1986) proposed a scheme to
predict the relative photolysis rates for nitroarenes, and the relatively rapid
photolyses of the 6-nitroB[a]P and 1-nitropyrene - both of which are
present in polluted atmospheres, are consistent with the predictions.
Further work is needed to establish the fate of nitroarenes - especially
those formed in atmospheric nitration reactions.
5.1.1.3 Reactions with Ozone
The solution phase reactions of B[a]P with ozone provide a simple model
system on which some of the reactions of ozone with PAH may be based.
Ozonolysis of B[a]P in solution (methylene chloride, 3:1 methylene
5-12
chloride-methanol) forms the 3,6- and 1,6-diones of B[a]P. With excess
ozone, more highly oxidized products 7H-benz[d,e]anthracen-7-one 3,4-
dicarboxylic and 1, 2-anthraquinonedicarboxylic acid are formed (Moriconi
et al., 1961). Confirmation of the mechanism was not feasible but one
of three schemes involving a two-step electrophilic attack by ozone at the
most reactive centres (i.e., carbons with the lowest carbon localization
energies) to give a sigma complex followed by nucleophilic 1-4 addition
to give a primary ozonide and thence to p-quinones was thought to be
applicable.
Several investigators have exposed PAH (from previously collected
ambient or source particulates or pure compounds deposited on filters
or evaporated onto glass surfaces) to ozone (Lane and Katz, 1977; Pitts,
1979; Pitts et al., 1980a; Peters and Siefert, 1980: Rajagapolan et al.,
1983; Brorstroem et al., 1983a; Grosjean et al., 1983b; Lindskog et al.,
1983). Most studies reported the disappearance of the reagent PAH, but
few identified reaction products. B[a]P quinones (1,6-, 3,6- and 6,-12
isomers) were identified by Rajagapolan et al. (1983), B[a]P dihydrols,
B[a]P diphenols and B[a]P-phenol and B[a]P-quinones (van
Cauwenberghe et al., 1979; Pitts, 1979) and ring-opened compounds
including dialdehydes, dicarboxylic acids, ketocarboxylic acids and
benzo[a]pyrene-4,5 oxide (Pitts et al., 1980a) were also identified. In
contrast, Grosjean et al. (1983b) found no evidence of reaction, but the
exposure conditions used (3 hours, 100 ppb ozone) were milder than
those in previous studies.
The relative rates of disappearance of several PAH in diesel particulates
exposed to ozone were found to be consistent with electron density
calculations (van Vaeck and van Cauwenberghe, 1984). No products
were identified and the data were consistent with earlier work by Lane
5-13
and Katz (1977). The presence of several oxy-PAH in ambient samples
suggests the oxidation of PAH by ozone could be important in
determining the nature of compounds found in ambient samples.
The most relevant investigations of the atmospheric stability of PAH have
been derived from outdoor reaction chamber studies in which PAH from
woodsmoke were reacted with ozone and nitrogen oxides under natural
daytime solar radiation and in the dark (Table 5-3). The half lives of PAH
on woodsmoke particles exposed to ozone (540 ppb), to NOg in the dark
and to 300 ppb O3, ranged from 30 to 60 minutes. The reaction of B[a]A
was found to be very temperature dependent - the half life decreasing by
a factor of four to ten for temperatures decreasing from 20 to -7°C
(Kamens et al., 1986). PAH decay at lower temperatures and solar
intensities, such as would apply in Ontario, therefore, would extend the
half lives to several hours. The more complete characterization of the
ozone reactions with a wider range of PAH is needed.
5.1.1.4 Reactions with Sulphur Oxides
The reactions of PAH with sulphur oxides have been investigated by
Jager and Rakovic (1974), Hughes et al. (1980), Butler and Crossley
(1981), Tebbens et al. (1966) and Grosjean et al. (1983a).
PAH absorbed on fly ash or alumina reacted with SO2 to form several
sulphur-containing compounds including pyrene-1-sulphonic acid, pyrene-
disulphonic acid, B[a]P-sulphonic acid (Jager and Rakovic, 1974).
5-14
TABLE 5-3
Outdoor Chamber Studies of PAH Reactions
with O3, NO2 and hv^
PAH ^ Half Life (minutes)
c^nn onK K\r^
hv^ 200 ppb 03^ 500 ppb NO/
Pyrene
42.5
271
B[a]A
45
186
??4
Chr/Tri'*
61.9
335
224
B[b]F
84.5
B[k]F
48.8
173
??7
B[a]P
37.1
235
??4
B[e]P
106
^ Kamens et al., 1985
^ Average total solar radiation intensity 1.2 cal cm'^ min"^
^ Reaction in the dark. Ozone half lives estimated based on rate
constants derived at 570 ppb O3.
'* Chrysene/triphenylene
5-15
In contrast, Hughes et al. (1980) using PAH adsorbed on coal fly ash,
alumina, silica and activated charcoal, found no reaction due to SOj, but
SO3 produced unidentified products. Similar studies by Butler and
Crossley (1981), but with PAH on soot, showed no effect due to SOj,
but Tebbens et al. (1966) did report degradation of B[a]P (products not
identified) by SOj.
The solution phase reaction of anthracene with SO2 is photocatalyzed
yielding anthracene-9-sulphonic acid (Nagai et al., 1986). Pyrene also
reacts with concentrated H2SO4 to give a mixture of sulphonic acids
(Valkman et al., 1937). These reactions suggest that it is theoretically
possible for PAH to react with SO2 or H2SO4 in the environment to form
sulphonic acids. These water soluble compounds have not been
identified in ambient samples but this may be due to the use of
inappropriate solvents for the extraction of ambient particulates (Nielsen
et al., 1983b). If any sulphonic acids formed react further to form
sulphonates, extraction methods used in the above studies to isolate
sulphonic acids would miss the sulphonates.
5.1.1.5 Photolysis of PAH Compounds
The photochemistry of PAH has been studied for many years. The low
temperature solid state photochemistry of condensed hydrocarbons has
provided a wealth of information on the spectroscopic properties of
molecules and the photophysical processes subsequent to absorption of
photons. Of importance to the photochemistry of PAH in the environment
is information on the accessibility and stability of excited photochemical
states. PAH in environmental matrices are adsorbed to the fine particu-
late matter and effects of the sorbent on the photochemical properties of
PAH are likely to be important.
5-16
The solid state photolysis of the pure PAH as well as their solution phase
photochemistry are limited, but the photochemistry of PAH adsorbed on
various real-environment and model sorbents has been studied by several
investigators.
The photolysis of pure PAH (B[a]P, benzo[b]fluoranthene and
benzo[k]fluoranthene) deposited on petri dishes was reported by Lane
and Katz (1977). They pointed out that surface reactions (photolysis
and reaction with ozone, for example) are likely to be important. The
effect of the nature of the particles onto which PAH are absorbed is
therefore likely to affect the photoreactivity. The spectral distribution and
intensity of the light source used were similar to sunlight, but the PAH
substrate (solutions of PAH evaporated in petri dishes) was not the same
as present in natural conditions.
Sorbents used have been soot (Thomas et al., 1968; Tebbens et a!.,
1971), particles on glass fibre filters (Fox and Olive, 1979; Peters and
Siefert, 1980; Pitts et al., 1980a), coal fly ash (Jager and Rakovic, 1974;
Jager and Hanus, 1980; Korfmacher et al., 1980b; Wehry et al., 1984;
Hughes et al., 1980; Blau and Gusten, 1982), silica gel, alumina, carbon
microneedles (Barofsky and Baum, 1976), diesel particulates and
chromosorb (Eisenberg et al., 1983) and soil (Fatiadi, 1967).
Oxidation products were identified in the photolysis of anthracene, B[a]A,
B[a]P, pyrene, perylene and fluoranthene on soot (Tebbens et al., 1971)
and on carbon needles (Barofsky and Baum, 1976). Chrysene and
coronene photolyzed under similar conditions were stable (Barofsky and
Baum, 1976).
5-17
In contrast, later work (McCoy and Rosenkranz, 1980) on the photolysis
of chrysene (as well as 3-methylcholanthrene) yielded unidentified
products that were shown to have increased mutagenetic activities. It
was postulated that the mechanism of the photooxidations involved the
triplet state of the PAH and singlet molecular oxygen.
A photooxidation mechanism involving singlet molecular oxygen was
suggested also (Fox and Olive, 1 979) for photolyses of anthracene spiked
onto previously collected ambient particulates. The products implicated
included anthraquinone, anthrone, bianthryl or the anthracene
photodimer. Pitts (1979) also suggested electrophilic attack by singlet
molecular oxygen was likely to be an important mechanism in
photooxidation of PAH.
Eisenberg et al. (1983) presented evidence consistent with the oxidation
of PAH by singlet oxygen according to the following mechanism (where
the * indicates an electronically excited species):
O2
PAH + h.--> PAH* — > PAH + O2* --> oxy-PAH
Several PAH as well as diesel particulates were found to be efficient
singlet oxygen sensitizers. Model compounds (9,10-diphenylanthracene
and chrysene) absorbed on Chromosorb 102 reacted in high yield with
singlet oxygen to form oxy-PAH products. Similar reactions are likely to
occur in the atmosphere. Fatiadi (1967) also postulated the reaction of
photoexcited pyrene molecules with adsorbed oxygen.
The reduced photosensitivity, especially of B[a]P, pyrene and
anthracenes adsorbed on coal fly ash particles compared to alumina,
silica gel or the pure solid (or even PAH in solution), was reported by
5-18
Korfmacher et al. (1980a). Taking into account this reduced
photosensitivity, presumably resulting from the energetics of surface
adsorption, the presence of PAH on coal fly ash particles in the
environment would imply long lifetimes. Thus, the persistence of PAH
after long range transport of submicron particles is feasible (Blau and
Gusten, 1982). Illustrative half-life data for different substrates are
presented in Table 5-4.
5.1.1.6 Concentrations of Oxy- and NItro-PAH in the Atmosphere
The previous discussion has indicated that both oxy- and nitro-PAH may
be emitted directly to the atmosphere or may be formed from PAH by
reactions in the open atmosphere. There is some evidence that the
predominant nitro-PAH are not those related to direct emissions, but,
instead, are the result of atmospheric nitration reactions. Observed
concentrations in ambient air, then, will reflect contributions from many
sources and processes.
Only recently have sampling and analytical methods for PAH derivatives
achieved levels of reliability which would allow ambient monitoring data
to be accepted with confidence. Thus, historical results ought to be
interpreted as qualitative illustrations of compound identifications and
atmospheric processes, rather than as quantitative data for exposure
estimation purposes. See Davis etal. (1986) and Finlayson-Pitts and Pitts
(1986) for discussions of this point.
5-19
TABLE 5-4
Half-Lives (in Hours) for the Photolysis of PAH
on Different Substrates Determined in the Rotary
Photoreactor (Approximately 25 ug of Each PAH/g
of Substrate, Except for the Carbon Black)
Silica
Carbon
PAH
Gel
Alumina
Fly Ash
Black
acenaphthylene
0.7
2.2
44
170
acenaphthene
2.0
2.2
44
a
fluorene
110
62
37
>1000
phenanthrene
150
45
49
>1000
anthracene
2.9
0.5
48
310
fluoranthene
74
23
44
>1000
pyrene
21
31
46
>1000
benz[a]anthracene
4.0
2.0
38
650
chrysene
100
78
38
690
benzo[e]pyrene
70
110
35
>1000
benzo [a] pyrene
4.7
1.4
31
570
perlyene
3.9
1.2
33
870
benzo[ghi]perylene
7
22
29
>1000
* Acenaphthene was not originally present on the carbon black studied.
From: Blau and Gusten, 1982.
5-20
Both oxy- and nitro-PAH have been identified in the air of Ontario cities
(D'Agostino, 1983; Nielsen, 1983; Nielsen et ai., 1983a; Ramdahl et al.,
1982; Pierce and Katz, 1976; Davis et al-, 1986). Systematic studies,
however, have not been carried out to allow conclusions about typical,
average or peak concentrations.
Table 5-5 shows ranges of measured ambient air concentrations of
selected nitro-PAH, for illustrative purposes. It appears from available
data that some nitro-PAH are present in concentrations similar to those
of many unsubstituted PAH.
The concentrations of oxy-PAH may be comparable to those of B[a]P in
highly polluted areas (Konig et al., 1983a). Relationships between the
concentrations of PAH and their respective oxidation products, e.g., for
B[a]P, B[a]A and their respective quinones, indicate that atmospheric
oxidation takes place, especially in summer (Pierce and Katz, 1976).
In other cases, the similar profiles of oxy-PAH in ambient and in source
samples show that the sources, in particular diesel exhausts, can account
for the presence of oxy-PAH in ambient samples. The absence of oxy-
PAH in rural samples, while present in urban samples (Tanner and Fajer,
1983), lends support to urban sources, especially automobile emissions,
as major contributors to ambient levels of oxy-PAH.
5-21
TABLE 5-5
Ambient Concentrations of Selected Nitro-PAH
Compound^ Location
Concentration
(ng/m^)
Reference
1-NP
R
0.02
Nielsen eta!. (1983)
U
0.016
Gibson (1982)
U''
0.031-0.1
D'AgostIno (1983)
R
0.009 + 0.005
Nielsen etal. (1984)
U
0.03 - 0.04
Arey etal. (1987)
U
0.008 - 0.03
Pitts etal. (1985c)
Re
0.01
Gibson (1986)
Ru
0.013
ibid.
S
0.015-0.022
ibid.
U
0.03
ibid.
1
0.029 - 0.057
ibid.
2-NP
U
0.03 - 0.04
Areyetal. (1987)
U
0.003 - 0.02
Pitts etal. (1985c)
2-NF
U
0.03 - 0.04
Areyetal. (1987)
U
0.07 - 0.3
Pitts el al. (1985c)
9-NA
U
0.04
Nielsen etal. (1983)
R
0.03 + 0.01
Nielsen etal. (1984)
U*^
0.008 - 0.034
D'Agostino (1983)
U
0.05-0.1
Areyetal. (1987)
10-NB[a]A
R
0.01
Nielsen etal. (1983)
R
0.014 + 0.007
Nielsen etal. (1984)
Notes:
a)
b)
1-NP
2-NP
2-NF
9-NA
= 1-nitropyrene
= 2-nitropyrene
= 2-nitrofluoranthene
= 9-nitroanthracene
10-NB[a]A =10-nitrobenz[a]anthracene
Re = remote
Ru = rural
U = urban
S = suburban
I = industrial
c) May include 8-nitrofluoranthene.
d) May include nitrophenanthrene.
5-22
The temporal variability of oxy- and nitro-PAH is likely to be similar to that
for the parent PAH. The higher levels seem likely in winter months (due
to increased emissions, reduced dispersion and lower chance of thermal
and photochemical degradation), but lower levels are likely in summer
(although higher ratios of oxy- and nitro-PAH to parent PAH may obtain).
Other than data by Pierce and Katz (1976), there are no other data to
support the above hypothesis.
Pitts et al. (1982) found that there is a diurnal variation of ambient
particulate mutagenicity that is similar to that of primary pollutants, as
evidenced by the high correlation of mutagenicity of 3-hour average
samples with CO, NO^ and Pb concentrations. The short-term (3 hour
average) peak mutagen activities of particles were much higher than the
24-hour average values, but there was agreement between the average
of the 3-hour samples and the 24 hour samples. The diurnal variation
was not observed in a subsequent study (Pitts et al. 1985c).
5. 1 .2 Long Range Transport of PAH
The dispersion of pollutants by long range transport is well known for
inorganic species, e.g., sulphates, nitrates, and also for organic pollutants
such as PAH. Given the common sources of PAH and oxy-PAH, it is
expected that long range transport will also be an important factor in
determining the distribution PAH and of their oxy- and nitro-PAH products
formed during transport.
5-23
Mesoscale and long range transport of PAH and oxy- and nitro-PAH have
been demonstrated by:
0 the presence of PAH at remote sites; and
o profiles (relative abundances) of PAH.
Nielson et al. (1983) showed that mesoscale transport of nitro-and parent
PAH were indicated in Riso, Denmark. This was based on comparisons
of PAH profiles of ambient samples with those for known sources.
Transformation of 1 -nitropyrene was thought to be insignificant. In
contrast, measurements of B[a]P, 1 -nitropyrene and marker inorganic
(Pb, Se) and organic (elemental carbon (EC)) species, at near-source and
remote sites, in conjunction with back trajectory determination, indicated
the formation of 1 -nitropyrene during LRT. In fact,the 1 -nitropyrene
concentrations at the remote site were higher than B[a]P levels at the
remote site. Also, the ratio of B[a]P to TSP at the industrial site (near-
source) was up to 250 times higher than the B[a]P/TSP ratio at the
remote site. The lower B[a]P/marker ratio at the remote site- as
compared to the near-source sites, is consistent with the reaction of
B[a]P during transport, but the disappearance of B[a]P appeared to
stabilise in aged aerosols (Gibson and Wolff, 1985). The changing ratio
of particle-bound B[a]P/TSP was, undoubtedly, also influenced by the
enhancement of vapour-phase B[a]P as TSP concentration decreased
due to dispersion.
Daisey and Kreip (1979) also suggested that the Long Range Transport
(LRT) of PAH into New York City may be important. The use of B[a]P as
a reference compound showed behaviour contrary to the expectation of
its rapid degradation by O3 and HNO3; thus the ratio of B[a]P to other
PAH in ambient samples increased instead of decreased relative to
5-24
source measurements. The increased ratio was the basis for concluding
that LRT under certain meteorological conditions was likely.
Bjorseth et al. (1979) indicated LRT of PAH in Europe, since the
concentrations of PAH varied significantly with origin of air masses, and
there were correlations of peak levels of PAH with other pollutants (SO4,
soot). The higher winter PAH levels were assumed to be due to less
dispersion (lower mixing heights) and a reduced likelihood of
photochemical degradation.
Lindskog and Brorstroem (1981) also asserted that the presence of high
concentrations of B[a]P was indicative of little chemical degradation, but
also used the absence of methylated PAH, together with low
concentrations of B[a]P, to support LRT. Similarly, the high correlation
of B[a]P concentrations with soot levels together with back trajectories
was used to distinguish between local and distant sources.
5.2 Soil
The available information shows that PAH occur in Ontario soils at higher
concentrations near a steel mill source (Sault Ste. Marie), near a Toronto
highway, as well as in soils and sediments around a coal gasification
waste disposal site at Port Stanley, as described elsewhere in this report.
Atmospheric deposition through PAH-contaminated rainfall has also been
measured in southwestern Ontario. Undoubtedly, PAH occur widely in
Ontario soils, and processes of adsorption, biodegradation and possibly
volatilization and photolysis influence the fate of PAH in the soil.
5-25
5.2.1 Degradation in Soils
Soil microorganisms are capable of metabolizing PAH to varying degrees,
as documented in reviews by Radding et al. (1976), Neff (1979), Sims and
Overcash (1983) and Bulman et al. (1985). Biodegradation studies have
included both measurements of total and specific PAH disappearance
rates in soils. The former type of study provides insight into the species
responsible for PAH metabolism, while the latter type provides a better
indication of "fate" under real world conditions.
Bulman et al. (1985) reviewed the metabolic pathways of PAH
degradation. Monooxygenases and dioxygenases are enzymes which
catalyze the incorporation of one or two oxygen atoms into the PAH ring
structure. The intermediates formed in these pathways undergo further
dihydroxylation steps that lead to the eventual cleavage of the aromatic
ring and degradation of the resulting phenols and carboxylic acids.
Culture studies have also demonstrated that some PAH resistant to
biodegradation may be readily degraded in the presence of other PAH
which support microbial growth, indicating that these more resistant forms
may be cooxidized in PAH mixtures occurring in soils, sediments and
sludges.
Several investigators have measured degradation (or perhaps, more
accurately, disappearance) rates of PAH in soils, although variations in
experimental conditions and analytical techniques have caused problems
in defining degradation rates in soil environments. Measured removal
rates may also be influenced by the superposition of slow solubilization
and desorption rates (i.e., competing physical rate processes). Sims and
Overcash (1983) reported half-lives of 3.3 to 175 days for A, 2.5 to 26
days for PHEN, 3 to 35 days for P, 44 to 182 days for F and 4 to 6,250
5-26
days for B[a]A. Studies with CHR and B[a]P have reported half-lives as
low as 5.5 and 2 days, respectively while other studies have indicated no
degradation (Bulman et a!., 1985). Gardner et al. (1979) reported
degradation half-lives in marsh sediment of 108 to 175 days for A and 105
to 182 days for F. Herbes (1981) reported half-lives of 1.8 and 8.8 days
for A and B[a]A, respectively, but no degradation of B[a]P, in pre-
exposed sediment downstream from a coking plant discharge. Lee et al.
(1984) observed biodegradation of FLN in subsoil near a creosoting plant
of 20 to 30% per week. In a study of degradation rates in sediments from
a contaminated stream, Herbes and Schwall (1978) measured rate
constants (h"^) of 2.5 x 10'^, 1 x 10"^ an L 3 x 10"^ from A, B[a]A and
B[a]P, respectively, while corresponding rates for sediments from an
uncontaminated stream were 2.5 x 10"^, 4 x 10"^ and L3 x 10"^. This study
indicates that biodegradation proceeds more quickly in acclimated
systems, and that larger PAH (4- and 5-ring compounds) are much more
resistant to breakdown.
In soil incubation experiments, Bossert and Bartha (1986) found that the
biodegradation rate was inversely affected by the number of aromatic
rings and directly correlated with water solubility. In this study, 3-ring
compounds (A, PHEN and ACEY) were mostly or entirely degraded over
four to 16 months, while most of the 4- and 5-ring PAH remained after
16 months of incubation. P which is quite water soluble was an
exception; this compound was 97% decomposed after 16 months. Two
of the 5-ring compounds, PER and 1,2,5,6-dibenzoanthracene, showed
no degradation after 16 months.
While most studies of PAH degradation in soil have assumed that first
order kinetics could be used to describe PAH disappearance from soils,
Bulman et al. (1985) found that either a model other than first order, or
5-27
a combination of two different models was required to describe ttie loss
of 99% of PAH from previously uncontaminated soil. PHEN, A, P and F
initially disappeared rapidly over 200 days or less until 94 to 98% loss
occurred, and rate constants for removal were about the same for both
5 and 50 mg kg'^ concentrations for all compounds except A. Loss of
PAH was probably due to complexation with soil particulates since
adsorbed fractions reduced the 'free' compound available for analysis.
Following the initial loss, the remaining 2 to 6% was lost at a much slower
rate. For B[a]A, CHR and B[a]P, only 22 to 88% was degraded over 400
days, and only one kinetic stage was identified for each compound and
concentration. Zero order kinetics was appropriate for describing the loss
of CHR and B[a]P, leading the authors to conclude that the assumption
of first order kinetics in modelling biodegradation of these compounds
could seriously underestimate their persistence in soil. Table 5-6
summarizes the half-lives for biodegradation in soil reported by Bulman
et al. (1985), along with comparative degradation half-lives compiled from
Sims and Overcash (1983).
Sims and Overcash (1983) listed several factors that affect the rate of
PAH biodegradation. These include temperature, pH, soil aeration,
moisture content, PAH concentration and previous exposure.
Degradation tends to increase under conditions ideal for microbial activity.
Degradation rate is generally greater at higher concentrations, and is
enhanced by previous exposure to the PAH.
5-28
TABLE 5-6
A Comparison Half-Lives Calculated on the Basis of
Rrst or Zero Order Models from Data Collected
by Bulman et al. (1985) with those Reviewed by
Sims and Overcash (1983)
PAH
Half Life
5 g/MT
added
(days)
Half Life
50 g/MT
added
(days)
Range of Half
Lives from
Sims and Overcash
(days)
First Order
Phenanthrene
9.7
14
2.5 to 26
Anthracene
17
45
3.3 to 175
Fluoranthene
39
34
44 to 182
Pyrene
58
48
3 to 35
Benzo[a]anthracene
240
130
4 to 6250
Zero Order
Chrysene
328
224
5.5 +
Benzo[a]pyrene
347
218
2 +
5-29
5.2.2 Sorption
Aqueous concentrations of hydropliobic compounds such as PAH in
soils and sediments depend on adsorptive/desorptive equilibria with
sorbents (solid particles) within the systems. In soils, liquid-solid
partitioning plays a significant role in retarding the migration of PAH in
groundwater.
Adsorption may also play a role in PAH degradation through surface-
associated chemical and biological processes (McCartyetal., 1981). The
affinity of all PAH for soil particles is high, and PAH-sorbent associations
are thought to occur primarily through van der Waals forces (Lyman et
a!., 1982).
The Freundlich adsorption model has been generally applied in evaluating
adsorption characteristics of PAH in soil/sediment-water systems. This
model is presented as follows:
X/M = KC
1/n
where: X = mass of compound adsorbed from a given mass of
solution (ug);
M = mass of solid adsorbent (g);
C = equilibria concentration in the liquid phase (ug/L); and
K,n = empirical constants.
K is a measure of adsorption strength or capacity and ^'^ is an indicator
of intensity, i.e., whether adsorption remains proportional to concentration
(n = 1), or changes with changing adsorbate concentration (1 ^ n <. 3).
When n = 1, a linear isotherm results and:
5-30
S = X/M = KC
where: S = concentration in the solid phase (ug/g).
Several studies on adsorption of hydrophobic compounds including PAH
have shown that linear adsorption isotherms are appropriate for
describing PAH adsorption (e.g., Karickhoff and Brown, 1979; Means et
al., 1980; Dzombak and Luthy, 1984). Adsorption coefficients have been
reported for many PAH for a variety of soils and sediments. The most
important soil and sediment property affecting adsorption is the organic
matter content, and adsorption coefficients are frequently reported as K^g
values which are normalized for organic matter content. K^g is calculated
as K/OC, where K is defined as indicated previously, and OC is the
fractional mass of organic carbon in a soil. The implication here is that
the degree of adsorption will vary directly with the organic content of the
soil. Kqc is independent of soil or sediment type, and is a constant for a
given chemical. It has been shown that K^g values can be estimated for
PAH and many other hydrophobic compounds using the octanol-water
partition coefficient (K^Jof the compound based on equations developed
by Karickhoff et al. (1979) (log K^^ = log K,,^ -0.21) and Hassett et al.
(1980) (log Kqc = log K^^ -0.317). Dzombak and Luthy (1984) examined
the application of K^^ in prediction of K^^ of PAH, and concluded that
relationships such as those noted above appear reliable for lower
molecular weight PAH, although more experimental K^^and K^gdata are
required to verify the reliability and scope of these relationships for higher
molecular weight PAH.
Typical values for octanol-water partition coefficients of PAH are listed
in Chapter 3 of this report. In general, log K^^ values range between
5-31
four and six for PAH, with values tending to increase with molecular
weight.
5.2.3 Volatilization
In soils, volatilization is complicated by adsorption and diffusivity.
Because most PAH are associated with the solid phase rather than in
the aqueous phase in soil, relatively little is available for volatilization.
The small fraction for the soil-water interface will have a relatively low
potential for volatilization owing to the low Henry's Law constants of PAH.
5.2.4 Photolysis
Little is known of the photoreactivity of sorbed PAH, although there is
evidence that adsorption both enhances and inhibits the photochemical
breakdown of PAH (Bulman et al., 1985). While it seems reasonable to
assume that some photolysis of PAH occurs at the soil surface, this
process would be prevented in the subsurface due to the blockage of
light penetration. No information was found in the literature that provided
insight on the significance of photolysis in the overall fate of PAH in soil
systems.
5-32
5.3 Water
5.3.1 Abiotic Processes
5.3. 1 . 1 Solubility and Sorption
Due to their low water solubilities, PAH compounds are generally
considered to occur in particulate form in lakes and rivers (Herbes, 1977;
NRCC, 1983). An equilibrium occurs between the adsorbed and
dissolved fractions, and while adsorption is usually favoured, dilute
solutions may contain a significant quantity of the dissolved form (NRCC,
1983).
Solubility of PAH in natural waters is affected by several factors.
Acenaphthene and pyrene were 24 and 31% less soluble in seawater
than in distilled water (Rossi and Thomas, 1981), but salinity effects on
PAH solubility are not relevant to most Ontario waters. Solubility tends
to increase with temperature over the normal environmental range (May
et al., 1978). The NRCC (1983) reviewed molecular properties of PAH
that affect solubility. Solubility decreases with increasing molecular
weight. Linear PAH molecules tend to be less soluble than angular
molecules, and alkyl substitution generally decreases solubility.
A number of different organic compounds have been found to increase
PAH solubility in water (NRCC, 1983). They include purines (Weil-
Malherbe, 1946), butyric and lactic acids (Ekwall and Sjoblom, 1952),
nitrogen-containing organic compounds (Eisenbrand, 1971), acetone,
ethanol and dioxane (Suess, 1972).
5-33
A recent study by Whitehouse (1985) concluded that "partitioning of PAH
into the dissolved phase is significantly influenced by the presence of
naturally occurring DOM (dissolved organic material); however, the PAH-
DOM interactions are specific with respect to the type of DOM and the
compound". Generally, the less soluble PAH were more interactive with
DOM, particularly with DOM with a higher molecular weight distribution.
McCarthy and Jimenez (1985a, b) similarly found a direct relationship
between the hydrophobicity of PAH and the affinity for binding to
dissolved humic material (DHM). Gjessing and Berglind (1981) also
demonstrated an increase in B[a]P solubility in humic acid-rich water,
although Boehm and Quinn (1973) found no effect on phenanthrene and
anthracene solubility in sea-water by humic-like organic matter.
Whitehouse (1985) and McCarthy and Jimenez (1985a, b) reported rapid
sorption (within minutes) of PAH with organic particles. Karickhoff and
Morris (1985), however, showed that more hydrophobic PAH sorbed
more strongly, but also more slowly than less hydrophobic compounds;
some PAH required days to weeks to reach sorption equilibrium. It may
be concluded that adsorption rates may vary considerably with
environmental conditions.
PAH also interact with particulate aqueous components, particularly
organics. Herbes (1977) demonstrated that a constant fraction (0.45
+.0.01) of total anthracene was adsorbed to 35 mg/L yeast cells over
an anthracene concentration range of 0.02 to 31 ug/L The partitioning
of anthracene between dissolved and adsorbed phases was dramatically
affected by the concentration of yeast cells with adsorption from 0 to a
maximum of 72% at 250 mg particles per litre. Increased temperature
resulted in decreased adsorption. The author concluded that 15 to 65%
of anthracene would be associated with detrital and living organic matter
5-34
in natural waters containing moderate levels of suspended organic solids.
He also suggested that the role of suspended mineral particulate material
may be far less significant in adsorption of PAH than is the role of
suspended organic matter, since only 1 to 5% available anthracene
adsorbed to montmorillonite clay. Results from Meyers and Quinn (1973)
support the last statement, as only 22% of available anthracene adsorbed
to bentonite clay. The observed sorption was attributed to van der Waals
forces. Sorption of PAH by organic particulates can be predicted using
positive correlations between the organic sorption partition coefficient, K^^,
and the octanol-water partition coefficient, K^^.
Suspended organic and inorganic particles associated with PAH gradually
settle out of the water column. This process is probably the most
significant route of PAH removal from water columns (Neff, 1979; Knap
and Williams, 1982; Neff, 1985). Suspended sedimenttransport, however,
has been found to be important in the downstream export of PAH in lakes
and rivers (e.g., Bahnick and Markee, 1985). Once in bottom sediments,
PAH are less subject to a degradative (e.g., photochemical or biological
oxidation) or physical (e.g., volatilization) processes. Sediments,
therefore, tend to accumulate PAH concentrations by a factor of 1 ,000 or
more relative to the overlying water and can serve as useful indices of the
rates of PAH inputs to the aquatic environment (Neff, 1979).
A study conducted by Comba et al. (1985) demonstrated the partitioning
of PAH between water and sediment phases in the Detroit River. Table
5-7 shows that low molecular weight hydrocarbons were detected in the
water column more frequently than high molecular weight compounds.
5-35
TABLE 5-7
PAH in Subsurface Water Samples (ng/L)
of the Detroit River (1983)
PAH
Cone. Range
(ng/L)
Frequency
(20 Stations)
ACEY
5.5-130
3
ACE
3.4-45
7
FLN
8.5-100
9
PHEN
4.7-180
13
A
1.5
1
F
9.5-91
4
P
4.5-11
3
CHR
-
0
BO or k]F
B[b]F
B[a]A
B[b]CHR
B[e]P
B[a]P
PER
8
4.8
8
3.4-17
1
0
0
0
1
1
2
IN[1,2,3,-cd]P
49
1
From: Comba et al. (1985).
5-36
TABLE 5-8
PAH in Suspended Solids (ng/g, dry weight)
of the Detroit River (1983)
Cone. Range
Frequency
PAH
(ng/g)
(13 Stations)
ACEY
120-1800
7
ACE
120-420
3
F
120-2000
4
PHEN
230-1300
2
A
92-770
2
F
180-2800
5
P
180-1900
3
CHR
160
1
BGork]F
44
1
B[b]F
44-57
2
B[a]A
76-680
4
B[b]CHR
210-1900
5
B[e]P
22-1400
4
B[a]P
40-4600
6
PER
53-290
2
B[ghi]PER
1.3-5100
8
From: Comba et al. (1985).
5-37
TABLE 5-9
PAH in Surficial Sediment Samples (ng/g, dry weight)
of the Detroit River (1983)
Cone. Range
Frequency
PAH
(ng/L)
(16 Stations)
ACEY
750
• 1
ACE
330-1100
2
FLN
450-520
2
PHEN
500-1800
9
A
370-1300
6
F
1000-4000
10
P
550-4000
10
CHR
1000-5000
6
B[j or k]F
1700-3000
4
B[b]F
560-3000
6
B[a]A
330-2700
7
B[b]CHR
-
0
B[e]P
310-1700
6
B[a]P
600-4900
6
PER
280-370
2
B[ghi]PER
370-810
4
From: Comba et al. (1985)
5-38
Conversely, high molecular weight PAH were common in suspended
bottom sediment samples (Tables 5-8 and 5-9). Low molecular weight
PAH were present in suspended but not bottom sediments. Pore waters
from bottom sediments rarely contained any PAH.
Concentrations of aquatic PAH reported by Eadie et al. (1983) for Lakes
Erie and Michigan follow a similar pattern of partitioning (Table 5-10 and
Figures 5-1 and 5-2) Niagara River data also showed that PAH are
generally transported in association with suspended particulates rather
than in a dissolved form in the Great Lakes.
5.3.1.2 Volatilization
Low molecular weight PAH with high water solubilities and low partition
coefficients, such as naphthalene, phenanthrene and anthracene, are not
as highly associated with suspended particles as are higher molecular
weight PAH, but are subject to other removal processes such as
volatilization (Readman etal., 1984). Henry's law constant, an equilibrium
coefficient which describes the distribution between gaseous and
aqueous phases, decreases with increasing molecular weight of PAH.
Volatilization is also highly dependent on mixing rates within the air and
water columns (Southworth, 1979a,b). High molecular weight PAH, such
as B[a]P and B[a]A are less prone to volatilization, and were not as
sensitive to changes in air and water currents. Southworth (1979a,b)
concluded that the rate of vaporization of PAH with four or more rings
is relatively insignificant under all environmental conditions. Volatilization
losses of lower molecular weight PAH may only be significant compared
to other processes in clear, turbulent waters.
5-39
TABLE 5-10
Mean PAH Concentrations (ng g'^ or ng mL'^)
Measured in Lake Michigan Sediment, Pore Water,
Dissolved and Particulate Paired Samples
PHEN A F P CHR
B[a]P
B[ghi]P
Nondepositional Sediments^ (n = 5)
90.0 18.4 185. 148. 76.1
82.9 2.7 132. 104. 49.8
62.1
49.9
65.7
61.1
Pore Waters From Nondepositional Sediments (n = 5)
0.73 0.12 2.02 1.68 2.95
0.46 0.036 2.42 1.81 3.52
2.32
1.26
1.75
Depositional Sediments^ (n = 9)
836. 195. 1,162. 999. 720.
252. 89.0 357. 319. 338.
462.
266.
369.
362. .
Pore Waters From Depositional Sedmiments (n = 9)
0.43 0.21 0.83 0.82 0.39
0.53 0.30 1.29 1.09 0.23
0.85
1.26
2.67
Lake Michigan Suspended Particulate Matter (n = 5)
2,405. 56.7 4,378. 3,890. 3,678.
2,411. - 1,768. 1,576. 2,156.
2,253.
1,954.
1081.
Lake Michigan Filtered Water (n = 5)
0.024 0.006 0.015 0.014 0.015
0.025 - 0.009 0.006 0.010
0.014
0.008
-
1. Represents near shore sites or non-depositional regions.
2. Was selcted as an area of recent sediment accumulation of deposition.
From: Eadie et al., 1983.
5-40
FIGURE 5-1
Sediment concentrations of individual PAH in Lake Erie. (A) Mouth of River Raisin
1 km up river of power station, 5 km and 10 km north. (B) Surficial sediments from
the remainder of the study area. The bar represents 1 standard deviation from one
composite sample and the individual samples. (C) Concentrations in oligochaete
worms from 1 and 10 km. (D) Concentration in chironomid midges from 1, 5, and
10 km. Bars represent 1 standard deviation, multiple analyses of a single extract.
(B)
(C)
(D)
800p
500 -
400 -
300 -
200 -
100 -
0-
200 r-
Sedi merits
S
f
(f
Hill
I
nil
I
150
100
I
50
Sediments
I
iit
0^
.., ill iii ili Hi Hi
iiir
i 200p womw
ISO
O
too
so
200 r-
I
n
a
ii.ii
I ■ 8
^"°r M:dQ«S
ili.JJiil.ili!Liii
From: Eadie et at., 1983.
5-41
FIGURE 5-2
PAH concentrations (log scale) in Pontoporeia hoyi (top panel), sediments (middle
panel), and pore water (bottom panel) from three stations. Solid bars are data from
the 24-m station; open bars are data from the 45-m station. Hatched bars are data
from the 60-m deep station in southeast Lake Michigan with high sedimentation
rates.
*ooo
C3000
«
J 2000
ca
^»
?100G
800
600
c
o
2 400
c
«
u
c
o
O
200
100
Pontoporeia
HiSed.
- 45 m ^
_24m
1000
800
600
= HiSed.
400-
200
c
100
o
80
S
60
c
«
40
o
c
o
(3
20
10
^^1^
£
10
■ —
8
9
C
6
-■-»
c
4
o
(S
h*
c
2
o
<J
Z 45m
8.5
Sediments
i NO
0
J£l
It
1
.JIU_
Pore Water
45m
24m|HiSe<l
Ph
An
From: Eadie et al., 1983.
5-42
Evaporation and sublimation were considered to be major factors in the
disappearance of a fluorene from 11 experimental ponds (Boyle et al.,
1984). The half-life of fluorene via volatilization from a 1.0-m column was
estimated at 100 hours.
5.3. 1 .3 Photodegradation
The chemical reactions of photo-induced oxidation of aqueous PAH by
singlet oxygen, ozone, HO radical, and other oxidizing agents are similar
to those involved in photo-oxidation of atmospheric PAH (Neff, 1979,
1985). The formation of endoperoxides is the most common oxidation
reaction, with subsequent photolysis or pyrolysis by a free-radical
mechanism to form a variety of products (Neff, 1979; cf. Table 5-11).
Early photo-oxygenation reactions were often examined in solutions,
including a solubilizer and under high oxygen conditions (Nagata and
Kondo, 1977; Neff, 1985) (Table 5-12), and may not, therefore, be
relevant to dilute aqueous conditions.
Zepp and Schlotzhauer (1979) showed that PAH in pure freshwater or
seawater are more likely to undergo direct photolysis than photo-
oxygenation. They reported half-lives of 30 to 40 minutes for B[a]P,
B[a]A and P, and 21 hours for fluoranthene. Smith et al. (1978) and
Southworth (1979a) reported similar half-lives for A, B[a]A and B[a]P of
approximately 30 to 35 minutes (Table 5-13). Picel et al. (1985) reported
photolysis rate constants of 0.82, 1.0 and 1.4 h'^ for P, B[a]A and B[a]P,
respectively, in pure water, but rates were six- to nine-fold less in an
aqueous coal matrix. Fluoranthene was an exception, with basically no
change in photolysis rate between the two solutions.
5-43
TABLE 5-1 1
Free-Radical Oxidation of Some PAH
in Air-Saturated Water
Compound
Oxidation Rate
(mole'' sec"')
Half-Life
Benz[a]anthracene
5.0 X 10^
1 .6 days
Benzo[a]pyrene
1.9 X 10^
4.3 days
Quinoline
2.8
8 years
Carbazole
29
280 years
Dibenzothiophene
<7.5
>3.5 years
Assumes [ROj] = 10 M.
From: Smith et al., 1977; 1978.
5-44
TABLE 5-12
Photodegradation of PAH Under NaturaJ Light
in Mixed Acetone-Water or Carbon
Tetrachloride-Water Solutions
Compound
5 hoL
Anthracene
52.9
Phenanthrene
57.0
Benz[a]anthracene
45.5
Chrysene
96.0
Fluorene
94.3
Pyrene
94.6
Benzo[a]pyrene
93.6
Degree of Photodegradation
(% compound remaining) at
10 hours
32.6
35.9
0.0
94.0
91.9
89.1
90.5
From: Nagata et al., 1977; Neff, 1985.
5-45
TABLE 5-13
Photo-oxidation of Some Dissolved PAH
Under Natural Sunlight Conditions
Compound
Rate
(sec"^)
Half- Life
(hours)
Anthracene
3.3x10-^*
0.6
Benz[a] anthracene
3.3 X 10^**
0.6 ■
Benzo[a]pyrene
3.6 X 10"***
0.5
Carbazole
1.9x10"***
1.0
Dibenzothiophene
1.5 X 10"^***
128
* Anthracene in distilled water exposed to mid-day sunlight in mid-
summer_at a latitude of 35°N (24-h photolysis for mid-summer =
1.2 X 10 sec and for mid-winter = 4.0 x 10'^ sec"Y
** Instantaneous rate constant for mid-day, mid-summer at a latitude
of 40°N.
*** 24-hour rate constant for mid-summer at a latitude of 40°N.
From: Smith et al., 1978; Southworth, 1979a.
5-46
Zepp and Schlotzhauer (1979) also reported unusual photochemical
behaviour for fluoranthene, which showed an unusually long half-live
(above).
Direct photolysis is most important among higher molecular weight,
compact PAH, (Zepp and Schlotzhauer, 1979; Neff, 1985) (Tables 5-12
and 5-13). Since light attenuates with depth in the water column,
photolysis rates of PAH also decrease. The presence of other materials
in the water column (e.g., dissolved or particulate organic substances)
can also affect the amount of light reaching PAH by absorbing or reflec-
ting specific wavelengths, or by affecting the mechanism of the
photochemical reaction itself (Oliver et al., 1979; Picel et al., 1985a, b, c).
Sorption on bottom sediments further decreases photolysis rates (Zepp
and Schlotzhauer, 1979; Neff, 1985).
In a field experiment, the half-lives of nine PAH ranged from less than a
day to 200 days (Table 5-14), with partitioning between suspended and
bottom sediments causing increased persistence (Zepp and
Schlotzhauer, 1979). Anthracene disappearance in outdoor channel
microcosms was largely due to photolysis with some volatilization losses
(Giesey et al., 1983; Bowling et al., 1984). Landrum et al. (1984) reported
similar results for anthracene in a stream microcosm,; photolytic
degradation to anthraquinone demonstrated to a half-life of 43 minutes.
5-47
TABLE 5-14
Direct Photolysis of PAH in a 5 Metre-Deep
Inland Water Body^
Sorption Partition
Coefficient^
Half-Life^
fdays^
Compound
No
With
Partitioning''
Partitioning"*'^
Phenanthrene
180
59
69
Anthracene
160
4.5
5.2
9-methylanthracene
550
0.8
1.2
Fluoranthene
280
160
200
Pyrene
400
4.2
5.9
Benz[a]anthracene
1,500
3.7
9.2
Chrysene
4,200
13
68
Naphthacene
3,800
0.2
1.0
Benzo [a] pyrene
3,100
3.2
13
Suspended sediment concentration, 20 mgL'''; diffuse attenuation coefficient, 1 1
m"\
Kp computed from octanoi-water partition coefficients.
Integrated over full summer day, latitude 40°N.
No partitioning assumes that PAH are completely in the water column and the
photolysis rates are affected only by light attenuation.
With partitioning describes rapid exchange between the top centimetre of the
bottom sediment and the water column. This exchange could involve continual
sedimentation and resuspension of the top layer of the bottom sediment.
From: Zepp et al., 1979.
5-48
5.3.2 Biological Processes
5.3.2.1 Bioavailability and Bioaccumulation
The presence of PAH in the tissues of various aquatic organisms
indicates that organisms are able to accumulate PAH at low
concentrations from ambient media and food (Neff, 1985). Fish,
invertebrates, insects and algae located near large sources of PAH have
demonstrated tissue levels in the ng/g to ug/g range (cf. Eadie et al.,
1982a, b; Knutzen and Sortland, 1982; Pruell et al., 1984), while biota
from remote or relatively unpolluted areas contain non-detectable levels
to concentrations in the low ng/g range as presented in Tables 5-15 and
5-16 (Brown and Pancirov, 1979; Murray et al., 1981). The accumulation
of PAH in aquatic organisms is due to the highly hydrohobic/lipophilic
nature of most PAH causing the chemicals to partition into lipid stores
in the organism (Neff, 1985). Accumulation is the net of the processes
of uptake, metabolism and depuration or excretion.
The partitioning of PAH in the aquatic environment (e.g., adsorbed,
dissolved, complexed with DOM) may affect the extent to which biota
take up ambient PAH. McCarthy et al. (1985) tested the bioavailability
of B[a]P, B[a]A and A in the presence of dissolved humic material (DHM).
High hydrophobicity was positively correlated with sorption to DHM and
resulted in reduced availability for uptake by Daphnia magna. B[a]P
uptake by oysters was reduced in the presence of DOM (Fortner and
Sick, 1985).
5-49
TABLE 5-15
Range of Polycyclic Aromatic Hydrocarbon
Concentrations (ug/kg, wet weight) in Bivalve
Shellfish from Different Oregon Bays
Bay
Site
Species
PAH
Concentration
Degree of
Industrialization
Tillamook
TIM
M. edulis
40-60
Relatively pristine.
Tillamook
TSS
M. arenaria
30-60
Relatively pristine.
Tillamook
TBC
Q. gigas
35-45
Relatively pristine.
Yaquina
Y140
C. gigas
30-45
Relatively pristine.
Coos
C3S
M. arenaria
70-90
Relatively pristine near
highway.
Coos
C11G
T. capax
30-110
Light; nearby marinas; fish
processing plant.
Coos
CSS
M. arenaria
480-650
Heavy; shipping docks;
wood products industry,
marinas.
Yaquina
YIM
M. edulis
140-440
Light shipping docks.
Yaquina
Y2M
M. edulis
675-1,325
Heavy; marinas, fish
processing plants,
recreational development.
From: Mix,
1984.
5-50
TABLE 5-16
Average Polycydic Aromatic Hydrocarbon
Concentrations (ug/kg., wet weight) in Mussels
Tvoe of Site fnumber of sites^
Clean (8)
Urban (12)
Industrial (6)
PAH
Mean
Range
Mean
Range
Mean
Range
Phenanthrene
17.2
7-33
17.7
4-43
214.0
28-621
Fluoranthene
22.4
2-85
46.1
6-198
215.3
8-476
Pyrene
15.8
2-78
43.8
5-158
199.7
7-540
Benzo[a]pyrene
7.1
1-13
38.3
2-236
101.2
5-329
From: Mackie et al. (1979).
5-51
Leversee et al. (1983) reported decreased accumulation of B[a]P by D,
magna in water containing humics, but anthracene and
dibenzoanthracene uptake were unaffected. The results were confirmed
for B[a]P in the same study in the presence of natural organics from
surface waters of South Carolina. Uptake of B[a]P by bluegills (Lepomis
macrochirus) was reduced by 90% in water with DHM. Naphthalene
uptake was not changed by DHM, but naphthalene is a low molecular
weight PAH with a low binding affinity for DHM. Anthracene and
phenanthrene uptake by the amphipod Pontoporeia hoyi were not altered
by DHM, but B[a]P, B[a]A and P were less bioavailable (Landrum et al.,
1985).
Neff (1 984a) demonstrated that PAH sorbed to sediments and suspended
particles were less bioavailable to aquatic organisms. Availability was
directly related to the compound solubility and sediment grain size, while
inversely related to organic carbon concentration and animal size. Since
sediments may contain high concentrations of adsorbed organic
pollutants, he concluded that they represent an important source of
contaminants to organisms despite their lower bioavailability. Varanasi
et al. (1985) also concluded that not all B[a]P bound to sediments was
bioavailable to a variety of estuarine organisms.
Only limited data are available on PAH concentrations in aquatic biota
occurring in the open environment in Ontario. A survey of contaminant
concentrations in Great Lakes sport fish was conducted by Zenon (1985).
Seven to twenty (20) fish per lake were tested for ten PAH levels. While
most PAH were non-detectable in the majority of samples, residues were
measured as high as 1 10 ng/g (phenanthrene in brown bullhead from the
St. Lawrence River). Most other maximum reported residues were less
than 50 ng/g. Corresponding water concentrations were not tested, so
5-52
the degree of bioaccumulation cannot be estimated. Konasewich (1978)
also reported PAH in Great Lakes fish, but the levels were not quantified.
Lake trout from Lake Superior and burbot from Lake Huron contained
phenanthrene and alkylated phenanthrene above the detection limit of
approximately 0.01 to 0.5 ug/g. PER, B[k]F, B[a]P and COR were
measured in carp and pike fillets from Hamilton Harbour contained the
greatest concentrations, particularly of B[a]P and COR (up to 0.4 ug/g),
while Detroit River fish had levels that were generally below 0.05 ug/g and
were often undetectable.
Fish do not appear to accumulate PAH to the extent of aquatic
invertebrates, probably due to the ability of the former group to
metabolize hydrocarbons (Eisler, 1987). Elevated levels of a hepatic
enzyme related to hydrocarbon metabolism, aryl hydrocarbon
hydroxylase (AHH), have been measured in lake trout inhabiting the
industrialized area of western Lake Ontario (Luxon et al., 1987). Roubal
et al. (1978) suspected that higher BCFs for PAH in flounder than in
coho salmon was related to differences in AHH activity rather than to
differences in lipid content.
PAH are metabolized relatively rapidly by fish. Under and Bergman
(1984) reported a BCF of 200 for rainbow trout relative to a 36 ug/L
exposure concentration of A after an 18-hour exposure. In the following
96 hours, A was rapidly metabolized and eliminated, but metabolism
proceeded quicker during the 8-hour dark phase of a photo cycle,
indicating that PAH metabolism may vary diurnally. Radio-labelled B[a]P
was taken up by northern pike through the gut and gill, and possibly
through the skin, was metabolized in the liver and subsequently excreted
in bile and urine (Balk et al., 1984). Eight and one-half days after the
initial exposure, most of the radioactivity was in the form of PAH
5-53
metabolites. Spacie et al. (1983) reported half-lives A and B[a]P of 17
and 67 hours respectively, in bluegills. The BCFs relative to water were
900 and 4,900 for A and B[a]P, respectively, and were lower than
predicted, due to rapid metabolism of the compounds.
A table of bioconcentration factors (BCFs) measured under controlled
conditions for a number of PAH was compiled by Eisler (1987) (Table
5-17). Values differ widely between species, and seem to be dependent
on a number of factors (discussed below). Generally, algae, molluscs
and other species which are incapable of metabolizing PAH show greater
accumulation. Increases in the molecular weight of PAH, K^^, exposure
time and the lipid content of the organism are some factors which tend
to encourage bioaccumulation (Eisler, 1987).
PAH uptake by two benthic invertebrates fPontoporeia hoyi and Mysis
rgl'Cta) were examined by Frez and Landrum (1985). Both represent an
important food source to some Great Lakes fish, but demonstrated
significant differences in uptake and depuration of B[a]P, A and Phen.
In addition, seasonal variation occurred. Increased water temperature
and decreased PAH solubility enhanced uptake in M. relicta. while uptake
and depuration were unaffected by PAH solubility in P. hoyi. The PAH
half-life in P. hoyi is approximately four times greater than that in M.
relicta. Jovanovich and Marion (1985) similarly demonstrated that
anthracene uptake and depuration in clams increased with temperature,
but concluded nutritional status and reproductive stage play a minor role
in anthracene accumulation. Varanasi et al. (1985) concluded that factors
such as feeding strategy and excretion rates probably account for higher
B[a]P bioaccumulation in amphipods than in clams.
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5-55
TABLE 5-17 (cont'd)
PAH Bioconcentration Factors (BCF)
for Selected Species of Aquatic Organisms
PAH compound, organisms Exposure BCF
and other variables period
PERYLENE
Cladoceran, Daplinia
pulex 24 h 7,191
PHENANTHRENE
Clam, Rangia
cuneata 24 h 32
Cladoceran, Daphnia
pulex 24 h 325
PYRENE
Cladoceran, Daphnia
pulex 24 h 2,702
Rainbow trout, liver 21 d 69
a) m = minutes, h = hours, d = days
Reproduced from: Eisler, 1987.
5-56
Uptake of B[a]P, A, PHEN and P by an oligochaete, Stylodrilus
heringianus, were measured in water and sediment from Lake Michigan
(Frank et al., 1985). Half-lives in the organisms were generally less than
two days. Uptake rate constants were similar to those of P. hoyi, but
depuration rates were more comparable to M. relicta. The authors
concluded that PAH fate in S. heringianus was more dependent on
metabolic rate than on any particular property of the PAH.
Pittinger et al. (1985) conducted an in situ experiment in Virginia to
measure bioaccumulation of Phen, F, P, B[a]A+Chr, BF and B[a]P by
oysters relocated from a non-impacted site to an urban/industrial area.
PAH increased from 0 to as much as 11.7 ppm dry weight within three
days, then stabilized. Depuration occurred to non-detectable levels after
oysters were transferred to pristine waters. Levels in indigenous oysters
corresponded with degree of industrial and urban development and
shipping traffic in the habitats.
Water solubility was related to depuration rates of PAH in contaminated
lobsters from Sydney, Nova Scotia (Uthe and Musial, 1986).
Approximately 85% of F was lost from the digestive glands over one year,
while as little as 6% of Chr (less water soluble) was lost. In an earlier
study, Uthe et al. (1984) measured more rapid depuration of PAH from
digestive glands of lobsters (31 to 77% over five weeks), but uptake of
the PAH had also occurred over a shorter period of time. Lobsters
exposed to a diesel oil spill in Newfoundland for less than ten hours
contained significantly higher concentrations of PHEN and P over controls
(Williams et al., 1985).
Gerould et al., (1983) measured bioconcentration of anthracene in the
midge Chironomus riparius. and found that the BCF was more strongly
5-57
affected by differences in biotransformation rate due to temperature than
by differences in uptake rate.
Landrum and Scavia (1983) investigated the influence of sediment on
anthracene uptake, depuration and biotransformation by the amphipod
Hyallela azteca. The mean uptake rate constant for waterborne A was
the same in the presence or absence of sediment. Sediment-associated
A (i.e., sorbed and in pore water) was estimated to contribute 77% of the
steady-state equilibrium burden of H. azteca. The role of sediment B[a]P
in the uptake by P. hoyi was variable, but deemed to be important to the
body burden when in high concentrations (Landrum et al., 1983).
The prediction of BCFs for compounds in mixtures may be difficult using
single compound kinetics. For example, tissue accumulations of
radiolabelled B[a]P in the oyster, Crassostrea virginica, were not affected
by the simultaneous presence of naphthalene and PCBs (Fortner and
Sick, 1985), while exposure of rainbow trout to anthracene alone resulted
in higher BCFs than when they were exposed to A in oil shale retort water
(Under et al., 1985). The retort water was believed to either decrease the
bioavailability of A to the fish, or limit the transport of contaminants from
uptake sites to storage and processing sites. The authors concluded that
prediction of BCFs for complex mixtures may be difficult based on single
compound kinetics. In addition, contaminants in mixtures may act as
inhibitors to PAH metabolism, thereby affecting extent of bioaccumulation.
PAH are also accumulated by fish eggs and larvae. Solbakken et al.
(1984) observed uptake of Phen and B[a]P in coastal cod, Gadus
morhua. eggs and larvae after 24 hours of exposure. The degree of
uptake appeared dependent on the molecular weight and the lipiophilic
characteristics of the contaminants.
5-58
The uptake of PAH by fish can be related to sediment concentrations,
particularly for bottom-dwelling species (Connor, 1984). Fish/sediment
ratios for a specific compound in different areas were correlated with the
residence time of the water in that area (e.g., lakes had higher ratios than
well-flushed coastal areas). The author proposed that predictions of
BCFs from sediment concentrations would reduce variability stemming
from water concentration data, but employing the model for non-bottom-
dwelling fish would also result in a ten-fold increase in variability. A similar
experiment by Landrum and Scavia (1983) resulted in overestimates of
BCFs for benthic organisms, based on the water concentration of
anthracene. Since 77% of the body burden was derived from sediment
contaminants, sediment-associated anthracene must have been less bio-
available. They also proposed that BCF estimations be based on
sediment concentrations for benthic organisms.
For several years, researchers have attempted to relate contaminant
uptake by organisms to various physical properties of the compound,
such as solubility or the octanol-water partition coefficient (K^J. A
number of these relationships have been developed, based on a wide
variety of organic contaminants including PAH. Table 5-18 presents some
observed BCFs for various PAH against their predicted values.
5-59
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5-60
5.3.2.2 Biodegradation
Bacteria initially metabolize PAH by incorporating the oxygen atoms of
molecular oxygen into the PAH structure to form a cis-dihydrodiol.
Oxidation reaction occur with further enzymatic action to form catechols
and eventually COj and water. Mammalian oxidation enzymes
(monooxygenases) metabolize PAH by forming reacting arene oxides
then a trans-dihydrodiol. It appears that some fungi also metabolize
PAH to trans-dihyrodiols with an enzyme system similar to mammalian
monooxygenases (Gibson et al., 1975; Cerniglia et al., 1982; Neff, 1985).
The evaluation of the importance of biodegradation of PAH in the open
environment is difficult due to the number of factors which influence
degradation rates.
Variations in the type of PAH present, the structure of the microbial
community, environmental conditions, and the method of assessment
have led to extreme variations in reported biodegradation rates (Oudot,
1984). Reviews of the subject generally conclude that PAH are degraded
more rapidly in aerobic than anaerobic conditions.
Also, the higher molecular weight PAH are more resistant to
biodegradation. Microorganisms which were previously adapted to PAH
are able to metabolize PAH more readily than unexposed organisms.
PAH which are resistant to degradation may be metabolized more readily
in the presence of other, more easily degraded PAH (NRCC, 1983; Neff,
1985).
Oudot (1984) examined biodegradation of bulk Arabian light crude oil by
a microbial culture over 60 days. The aromatic hydrocarbons degraded
5-61
approximately 50%. Degradation rates for each compound were related
to the number of rings in the molecule, decreasing in the order 1 > 3 >
2 > 4 > 5 rings. In an experiment using oil in water and sediments from
the North Sea, Massie et al. (1985) also showed that microorganisms
have the potential to degrade smaller PAH rapidly in the water column
and in surface sediments. B[a]P mineralization was tested in sediments
only, and was found to be minimal.
Biodegradation of PAH in activated sludge was investigated by Freitag
et al. (1985). The conversion of PAH to CO2 was 0.3% and 39.6% for A
and Phen, but less than 1% of the 4- and 5-ring PAH (B[a]P, Per, B[a]A
and DB[a,h]A) were mineralized to CO2.
There is controversy in the literature concerning the effect of suspended
solids on degradation rates. Hall et al. (1986) found that no significant
microbial degradation of ACE occurred when no suspended solids were
present. Degradation was significantly more rapid in the highest
concentration of one suspended sediment sample tested over the next
lowest concentration after seven days. In the presence of a different
sample of suspended sediment with a higher organic content, significant
degradation of A also occurred, but was not related to the concentration
of the suspended sediment. It should be noted that the behaviour of
naturally occurring suspended solids may be different from that of
experimentally re-suspended sediments.
Herbes (1981) suggested that the larger 5-ring PAH were less
biodegradable in sediments due to their strong sorption to sediment
particles and resultant reduced availability. Comparative half-lives of six
PAH in water and sediment in a stream are presented in Table 5-19.
5-62
TABLE 5-19
Degradation Rate Constants (k) and Half-Lives (t 1/2)
for Mixed Bacterial Populations in Water and
Sediment from the Same Stream
Sediment
k(h-^)
t 1/2
Wa
k(h-^)
Anthracene
1.6x10'^
43 h
2.0 X 10
Benz[a]anthracene
3.3 X 10"^
208 h
ND
Benzo[a]pyrene
3.4 X 10"^
83 h
ND
Dibenz[a,h]anthracene
1.2 X lO'^(l)
166 h
ND
t 1/2
350 h
ND = not detectable
(1) single determination
From: Herbes, 1981.
5-63
Other studies of biodegradation in soils and sediments tiave shown that
while biodegradation does occur in the aquatic environment, the rate of
metabolism is dependent on many environmental factors as well as the
composition of the microbial community and the PAH composition.
5.3.4 Microcosm Studies
The intentional contamination of artificial ecosystems, or microcosms,
has been widely practiced in studies of contaminant transport and
fate. Microcosm studies of some PAH have been conducted to
determine the fate of each compound and the relative importance of
various environmental compartments and fate.
Fluorene was applied to experimental pond ecosystems at concen-
trations of 0.12 to 10.0 mg/L (Boyle et al., 1984). Fluorene at
concentrations above the limit of solubility appeared to sublime from
the pond surface. Fluorene disappearance was rapid and attributed
largely to evaporation and sublimation as well as sedimentation and
degradation. Seven days after treatment, only a small fraction of
applied fluorene was accounted for (7 to 18%). The half-life of Fin in
the 0.12-mg/L pond was 6.7 days, while at the highest application
concentration (10.0 mg/L) the half-life was 27.4 days.
Anthracene added to an artificial stream microcosm also rapidly
disappeared (Landrum et al., 1984). Anthracene was rapidly
photolyzed to anthraquinone (half-life 43 minutes) which was
subsequently rapidly photolyzed. The organic sediment acted as a
major repository, absorbing 0.2% of the 14-day input dose. Periphyton
took up 0.04% of the applied dose, and all other compartments (water
and biota) contained relatively minor amounts.
5-64
Another channel experiment with anthracene provided similar results
(Bowling et al., 1984). Anthracene was input at 15 ug/L continuously
for 36 days. Downstream dissipation was rapid via photolysis, during
daylight. Some loss was attributed to volatilization. Aufwuchs
achieved maximum concentrations within 4 days (BCF = 1260) but the
concentration relative to the total input was only 0.02%. After
discontinuing A input, background concentrations were achieved in
water and aufwuchs within 24 hours and 72 hours, respectively.
In a closed model aquatic ecosystem, B[a]P was applied to water at
0.002 mg/L (Lu et al., 1977). Bioaccumulation was observed for three
days in organisms from several trophic levels. BCFs for fish, alga,
mosquito, snail and daphnia were 930, 5,258, 11,536, 82,231 and
134,248 respectively.
Labelled B[a]A was introduced to a large-scale marine microcosm
resembling shallow coastal waters of the northeastern U.S. (Hinga et
a!., 1980). All B[a]A and breakdown products were removed from the
water with a half-life of 52 hours. Most B[a]A was rapidly transferred
to sediments. Some ^""C activity was observed in particulates and
plankton with an initial half-life of 35 hours. After 230 days, 29% of the
applied radioactivity had been respired to COj, while the remaining
activity was evenly divided between parent compound and intermediate
metabolic products.
6-1
6.0 ENVIRONMENTAL LEVELS AND SOURCE CONTRIBUTIONS
The purpose of this chapter is to integrate information from several
chapters to describe typical concentrations of PAH which may be found
in Ontario in media through which environmental exposure to PAH may
occur. Emission profiles and emission rates are used to estimate the
relative contributions of selected source categories to the atmospheric
loading of PAH in Ontario. Finally, summary conclusions regarding an
assessment of the information contained in the report are offered.
6.1 Levels in Air, Soil and Water
Ambient air PAH levels near the Niagara River were measured in a study
by the Atmospheric Environment Service (Hoff and Chan, 1987). Hi-vol
samples were collected at three (3) sites during 1982 and 1983.
Sampling sites were at the following locations:
0 Niagara-on-the-Lake;
0 Fort Erie; and
0 Niagara Falls.
The sampling configuration was a filter followed by a polyurethane foam
(PUF) plug. The two fractions represent nominally, the particle and
vapour portions. Because of blow off and volatilization processes, the
particle fraction may be distorted for intermediate molecular weight PAH.
The sum of the two fractions is more accurate.
Average particulate and gaseous PAH concentrations from the three sites
are summarized in Tables 6-1 and 6-2 for tests conducted in September
1982 and January 1983 respectively. The results showed a strong, local
6-2
influence of heavier weight particulate PAH from the winter sampling
period (January, 1983).
For example, in ambient air, the PAH concentration was highest at Fort
Erie, lower at Niagara Falls and lowest at Niagara-on-the Lake, which is
directly related to the inputs from industry and mobile sources (Hoff,
1987). An examination of wind flow direction during the sampling period
indicated a strong influence on the PAH level arising from emissions from
the urban areas of Buffalo, Niagara Falls, NY, and Niagara Falls, Ontario,
for easterly winds. In addition, the increase in PAH emissions at Niagara-
on-the-Lake for north northwesterly winds was possibly due to transport
from the Toronto-Hamilton urban corridor.
Ontario Ministry of the Environment annual PAH results for the period
1973 to 1983 at the Niagara Falls sites are illustrated in Figure 6-1 along
with results of the AES study (Hoff and Chan, 1987).
A definitive study was carried out by Katz et al., 1978, on the PAH
distribution in the ambient air of four (4) Ontario cities, i.e., Toronto,
Hamilton, Sarnia and Sudbury. Two sites in Toronto were reported.
Selected results from the various sampling sites are summarized in Tables
6-3 and 6-4.
The study shows that the highest PAH levels occurred at the Hamilton
site, followed by Toronto, Sarnia and Sudbury. B[ghi]PER was reported
at significant levels at all the sites. The likely source was postulated to
be from the exhaust gases of motor vehicles.
A recent survey of ambient air PAH has also been carried out in Toronto
from 1984-1986 (Dann, 1988). Data are presented in Table 6-5.
6-3
TABLE 6-1
Mean Ambient Air Levels of Polycyclic
Aromatic Hydrocarbons at Niagara-on-the-Lake,
Niagara Falls and Fort Erie
in September, 1982
Compound
Filter
concn
n
pgm'^
PUF
concn.
n
pgnT^
Phenanthrene
Pyrene
5
3
180 ±140
65 ± 53
5
5
4800 ±1100
300 ± 350
n = number of samples
From: Hoff and Chan, 1987.
6-4
TABLE 6-2
Mean Ambient Air Levels of Polycyclic
Aromatic Hydrocarbons Near Niagara-on-the-Lake,
Niagara Falls and Fort Erie
in January, 1983
Filter
PUF
concn.
concn.
Compound
n
pg m"-"
n
pgm"^
Phenanthrene
19
830 + 1000
16
13000 +
5900
Anthracene
19
45 i 57
9
990 +
960
Fluoranthene
19
1400 ± 1900
9
3700 +
2200
Pyrene
19
1200 + 1800
8
3000 +
2100
Benz[a]
19
2800 ± 5600
bdl
anthracene
BenzoO] +
18
1100 ± 1500
bdl
benzo[k]
fluoranthene
Benzo[e]pyrene
17
230 + 440
bdl
Perylene
11
23 + 52
bdl
Benzo[ghi]
12
530 + 1500
bdl
perylene
n = number of samples
bdl = below detection level
From: Hoff and Chan, 1987.
6-5
FIGURE 6-1
Graphical Illustration of Benzo[a]pyrene
Concentration and Year of Study
at Chippewa/Niagara Falls
in the Particulate Phase
10,000
1,000
n
E
I
z
o
F
<
UJ
o
z
o
o
100
10
_L
BENZO[a]PYRENE
• CHIPPAWA (MOE)
A HOFF AND CHAN, 1987
■ NIAGARA FALLS (MOE)
_L
_L
_L
_L
_L
_L
_L
_L
73 74 75 76 77 78 79 80 81 82 83
YEAR
From: Hoff and Chan, 1987.
6-6
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6-8
TABLE 6-5
Summary of Mean Total PAH
Concentrations (ng/m^) in Toronto
(October 1984 - July 1986 - 42 Sampling Days)
Compounds
Mean C
oncem
Acenaphthylene*
4.12 +
5.98
Acenaphthene*
2.24 +
2.04
Fluorene*
5.51 +
4.37
2-Methyl-Fluorene
3.66 +
2.34
Phenanthrene
15.64 +
8.68
Anthracene
1.60 +
2.85
Fluoranthene
4.91 +
2.70
Pyrene
3.87 +
2.88
Benzo[a] Fluorene
0.53 +
0.48
Benzo[b]Fluorene
0.27 +
0.27
1-Methyl-Pyrene
0.18 i
0.21
Benzo[ghi] Fluoranthene
0.49 +
0.69
Benzo[a] Anthracene
0.40 +
0.92
Chrysene & Thphenylene
0.73 +
0.58
7-Methyl-Benzo [a] Anthracene
0.00 +
0.00
Benzo[b&k] Fluoranthene
1.26 i
1.45
Benzo[e] Pyrene
0.48 +
0.60
Benzo[a] Pyrene
0.30 i
0.52
Perylene
0.04 +
0.11
lndeno[1 , 2, 3-cd] Pyrene
0.46 +
0.62
Dibenzo[a,c]&[a,h]Anthracene
0.04 ±
0.09
Benzo[b]Chrysene
0.04 +
0.13
Benzo[ghi] Perylene
0.07 +
1.16
Anthanthrene
0.02 ±
0.05
Total PAH
47.57 ±
27.26
Collection efficiency poor and dependent on ambient temperature (see
text).
From: Dann, 1988.
6-9
Although useful for general information, the data presented in the
preceding tables do not provide a link between source emissions and
ambient air data. Outside of Ontario, there are some examples which
associate source emissions with ambient air data: Daisey et al. (1981;
1986); Daisey (1985); Thrane and Wikstrom (1983).
Whereas all the authors admit that any attempt to conduct source
apportionment receptor modelling for PAH will probably leave many
unanswered questions because of the lack of accurate information and
because of the reactivities of PAH, Daisey and Kneip (1981) have
attempted to group the emissions from six (6) sources and report in
terms of the B[a]P/B[ghi]PER ratios. Her group have also attempted
source apportionment receptor modelling for PAH (Daisey, 1985).
It has also been noted in various sections of the report that B{a]P/B[e]P,
Anthn/B[e]P, B[ghi]PER/B[e]P ratios as well as B[k]F, coronene,
nitropyrene or retene levels may be indicative of certain sources. No
unambiguous picture emerges from the data, however, because of the
severe temporal and spatial varaiability of relative concentrations of PAH
in ambient air samples. In the absence of unique source PAH markers,
it is not possible to apply receptor modelling to obtain quantitative source
apportionment except in specific local situations, where long-term studies
might establish an adequate database to resolve PAH source
contributions. Inorganic (elemental) source and receptor sample profiles
will undoubtedly provide better data than PAH or other organic
composition with which to attempt source apportionment, for the near
future.
6-10
6.2 Residue Levels in Ontario Soils
The results from soil evaluation in the Port Credit and Oakville/Burlington
areas, which formed part of a site decommissioning program undertaken
by Texaco and Shell, showed anthracene present in one sample, and
phenanthrene, benz[a] anthracene and chrysene, benzo[b&k]fluoranthene
in 7 to 8 samples (Golder Associates, 1987).
The following PAH were detected in a large number of the soil samples
analyzed.
0 Pyrene (15 samples), mean concentration of 0.13 ppm with a
standard deviation of 0.15 ppm;
0 Fluoranthene (16 samples), mean concentration of 0.15 ppm with a
standard deviation of 0.20 ppm; and
0 Benzo [a] pyrene (28 samples), mean concentration of 0.024 ppm
with a standard deviation of 0.034 ppm.
6.3 Residue Levels in Sediments
PAH levels in surface waters has been reported by the Great Lake
Environmental Research Laboratory on all Great Lakes (Table 6-6). A
compilation of PAH data of domestic water sources is available in Table
6-7 for the city of Erie, based on samples collected in 1976 (Eadie et al.,
1982; IJC, 1978). Concentrations in raw lake water are typically below
1 ug/L, and often below detection limits. The data indicate that average
concentrations of PAH in filtered lake water are typically below 0.1 ug/L.
6-11
TABLE 6-6
PAH Concentrations in Surface Water (Rltered)
SurficiaJ Sediments, Sedimerrt Porewater
and Benthos from the Great Lakes
Phenanthrene Anthracene Ruoranthene Pyrene Chrysene Benzo[a]pyrene
Wafer
n = 6
Mean
s
0.024
0.025
Surficial Sediment (ppb dry):
range/n
Superior
(n = 1)
Michigan
(n = 10)
Huron (n = 3)
Erie (n = 4)
Ontario (n = 5)
34
6-1,268
11-272
18-431
40-205
Surficial Porewater (ppb):
n = 9
Mean 0.43
s 0.53
Bulk Sediment 84
(ppb dry)
60 um and Rner 192
Sediment (ppb dry)
and Oiigochaetes 185
(ppb wet)
(Lake Erie near shore)
0.006
0.006
0.21
0.30
0.015
0.009
88
9-1,664
33-487
65-285
210-1,000
0.83
1.29
30
440
190
0.014
0.006
53
8-1,430
36-256
57-287
56-1,182
0.82
1.09
22
343
250
0.014
0.010
0.39
0.23
14
322
130
0.012
0.008
28
4-944
23-294
56-173
76-306
0.85
1.26
69
242
20
1
Filtered lake water.
From: Eadie et al., 1982a.
6-12
TABLE 6-7
PAH Data for Erie, PA (Rnished Water)
Concentration
(ng/L)
Detection Limit
(ng/L)
Fluoranthene
ND
10
Benzo[k]fluoranthene
A
Benzo[b]fluoranthene
ND
30
Benzo[ghi]perylene
ND
50
Benzo[a]pyrene
ND
30
lndeno[1 ,2,3-cd]pyrene
ND
50
ND - not detected
A - analysis was not attempted
From: IJC, 1978.
6-13
Additional data on suspended sediment concentrations for the Great
Lakes are available for the Niagara River (Table 6-8). Concentrations in
suspended sediments in the Niagara River range between 4 to 16 ppb
for ACEY to about 100 to 1,500 ppb for A/PHEN, F, P, CHR/B[a]A,
B[b,k]Fand B[a,e]P.
Additional data on PAH levels in surficial sediments in the Great Lakes
basin are presented for all Great Lakes (Table 6-6), Hamilton Harbour
(Table 6-9), and the St. Lawrence River (Tables 6-10 and 6-11). Data
from Eadie et al. (1982a; Table 6-6) indicate that most PAH in surficial
sediments are associated with the solid phase rather than pore water,
and that PAH are preferentially associated with fine fractions.
Sediment core data for PAH in the Great Lakes demonstrate the presence
of greater levels of contamination in the surficial layer than in deeper
deposits. Table 6-12 provides a list of PAH profiles found in sediment
core off Toronto (IJC, 1976). Core profiles of PAH and alkylated PAH at
varying distances from the Niagara River mouth further demonstrate this
surface contamination phenomenon, and illustrate the direction of PAH
transport from the Niagara River (Onuska et al., 1983).
6.4 Residue Levels in Fish and Wildlife
Data on PAH concentrations in tissues of fish from the Great Lakes basin
are provided in Tables 6-13 and 6-14. Concentrations in fish flesh for
individual PAH are typically in the ppb range.
6-14
TABLE 6-8
PAH Concentration in Suspended Sediments: Niagara-on-the-l_ake
PAH Concentration (ppm)
Acey
0.004 to 0.016
Ace
0.008 to 0.038
Fin
0.010 to 0.042
A/Phen
0.166 to 1.58
F
0.173 to 0.942
P
0.141 to 0.824
Chr/B[a]A
0.105 to 1.51
B[b and
k]F
0.193 to 1.08
B[a and
e]P
0.190 to 1.10
From: Canada Ontario Review Board, 1981.
6-15
TABLE 6-9
Polynuclear Aromatic Hydrocarbons
in Hamilton Harbour Sediments 1982
Compound
Range (ug/g)*
Fluoranthene
1.9-4.3
Perylene
1.2-9.7
Benzo[k]fluoranthene
1.1 -9.0
Benzo[a]pyrene
1.2- 11.1
Benzo[ghi] perylene
1.6-8.6
lndeno[1,2,3-cd]pyrene 1.1-9.7
* Number of samples = 6
From: MOE, 1985.
6-16
TABLE 6-10
Mean, Minimum and Maximum Values of PAH (ng/g)
in Sediments from 30 Stations in the St. Lawrence River, 1981
PAH Concentration
Mean Minimum Maximum
PAH Total
551.5
31
1,883
Anthracene + Phenanthrene
32.0
2
960
Fluoranthene
120.0
11
360
Benz[a] anthracene
53.0
4
120
Benzo[b]fluoranthene
130.0
7
770
Benzo[k]fluoranthene
46
3
130
Benzo[g,h,i]perylene
70
2
190
lndeno[1 ,2,3-cd]pyrene
68
2
900
From: IJC, 1987a.
6-17
TABLE 6-1 1
PAH in St. Lawrence River Sediment Samples
Collected in the Vicinity of the General Motors
Facility at Massena, New York
Number of Samples
Concentrations Out of 8 in Which
(mg/kg, dry weight)Compound Detected
Acenaphthene
BMDL
2
Anthracene
BMDLto 1.01
2
Ben2[a]anthracene
BMDL to 4.00
5
Benzo[a]pyrene
4.32 to 6.55
2
Benzo[b]fluoranthene
1.72 to 7.92
5
Benzo[ghi]perylene
BMDL to 2.02
3
Benzo[l]fluoranthene
2.94 to 4.37
2
Chrysene
BMDL to 4.57
5
Dibenz[a,h]anthracene
BMDL
2
Fluoranthene
BMDL to 3.64
7
Fluorene
BMDL
5
lndeno[1 ,2,3-cd]pyrene
BMDL to 3.74
3
Naphthalene
BMDL
1
Phenanthrene
BMDL to 4.16
6
Pyrene
BMDL to 2.78
6
BMDL - below minimum detection limit
From: IJC, 1987a.
6-18
TABLE 6-12
Abundances of Polycyclic Aromatics in
Lake Ontario Sediment
(Latitude 43°39', Longitude 78°12')
(ug/g dry sediment)
- no data
From: IJC, 1976.
0-5 10-15 20-25 30-35 55-60 70-75
cm cm cm cm cm cm
Biphenyl
0.014
0.007
0.009
0.004
0.004
Tetrahydropyrene
0.056
0.029
-
-
-
-
Fluoranthene
0.281
0.058
-
-
-
-
Pyrene
0.056
0.029
-
-
-
Chrysene/
0.225
0.088
0.052
-
-
-
Triphenylene
Dimethyl chrysene
0.112
-
-
-
-
0.018
Benzo[b]fluoranthene
0.450
0.029
0.017
0.034
0.010
0.009
Methyl benzofluoranthene
0.056
-
-
-
-
-
Benzpyrenes
0.337
-
0.017
0.034
0.010
0.009
Perylene
0.056
0.029
0.017
0.034
0.30
0.046
Methyl benzpyrene
0.056
-
-
-
-
-
20-methyl cholanthrene
0.337
-
-
-
-
0.018
Benzperylene
0.225
-
-
-
-
-
Coronene
0.562
-
-
-
-
-
Total
2.935
0.269
0.112
0.089
0.084
0.131
6-19
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6-20
TABLE 6-14
Polynuclear Aromatic Hydrocarbons in
Great Lakes Rsh Identified by
Mass Spectrometry
Hamilton Harbor
Detroit River
PAH
Carp
Pike
Carp
Pike
acenaphthene
X
X
fluorene
X
X
anthracene
X
X
X
phenanthrene
X
X
X
1 -methyl phenanthrene
X
X
X
1 -methyl anthracene
X
X
X
2-methyl anthracene
X
X
X
2-methyl phenanthrene
X
X
X
9-methyl anthracene
X
fluoranthrene
X
X
X
pyrene
X
X
X
1,2-benzofluorene
X
X
2,3-benzofluorene
X
X
chrysene
X
X
X
benzo [a] pyrene
X
X
perylene
X
X
dibenz[a,h]anthracene
X
X
X
coronene
X
X
X
X detected
other compounds scanned for but not found inlude 1 -methyl pyrene,
benzo[e]pyrene, anthantrene, benzo[ghi]perylene and dibenz pyrenes.
From: IJC, 1978.
6-21
The only recorded data on PAH concentrations in wildlife tissues (other
than fish tissues) available for the Great Lakes area are provided by
Hallett et al. (1977), who reported several PAH in herring gull lipids from
samples collected in the Lake Ontario basin (Table 6-15). Because PAH
concentrations were generally lower than those reported in fish, it was
concluded that food chain bioconcentration did not take place. Since that
study, no further analyses have been reported on Great Lakes gull tissues
(Canadian Wildlife Service, pers. comm.).
6.5 PAH in Human Tissues and Ruids
PAH have been demonstrated to accumulate in human tissue.
Concentrations of 9 target PAH were determined in human fat and liver
samples from 10 normal people with uncharacterized smoking habits,
occupation and residence (Cobana et al., 1981). Benz[a]anthracene and
dibenz[a,h]anthracene were not detected in either tissue type. Total (of
7 remaining target) PAH concentrations average 1 100 pg/g in fat and 380
pg/g in liver. Concentrations of benzo [a] pyrene ranged from < 5-59 pg/g
in fat, from 10-32 pg/g in liver; concentrations of benzo[b]fluoranthene
ranged from 56-260 pg/g in fat, from 33-88 pg/g in liver.
Concentrations of 4 target PAH determined in human bronchial carcinoma
tissue demonstrated benzo[a]pyrene to be the most significant PAH, with
values ranging from 0.3-15,000 ng/g (Tomingas et al., 1976). Several
PAHs (acenaphthene, fluorene, phenanthrene, anthracene and pyrene)
have also been detected in human atherosclerotic aortas, in
concentrations of 8-30 ng/g (Ferrario, 1985).
6-22
TABLE 6-15
Identification of Polynuclear Aromatic
Hydrocarbons in Great Lakes Herring Gull Lipid
Concentration (ug/kg)
Pigeon
^^-f
Mass Spectra
Compounds
Island
Kingston
Confirmation
acenaphthene
0.038
0.007
fluorene
0.044
0.003
-
anthracene
0.152
0.024
+
phenanthrene
nd
0.002
-H
2-methyl phenanthrene
0.021
0.007
-1-
1 -methyl phenanthrene
0.010
0.015
-1-
9-methyl anthracene
0.011
0.025
-(-
fluoranthrene
0.082
0.017
-1-
pyrene
0.076
0.015
-t-
1,2-benzofluorene
a
a
-1-
2,3-benzofluorene
a
a
+
chrysene
0.053
a
-(-
benz[e]pyrene
0.026
0.021
+
benzo[a]pyrene
0.038
0.030
-h
perylene
0.053
0.026
-1-
From: Hallett et al., 1977
a = PCB interference; b = standards of compounds unavailable, compounds
identified by mass spectra; nd = not detected.
Other compounds scanned for but not found inlude 1 -methyl pyrene,
benzo[e]pyrene, anthantrene, benzo[ghi]perylene and dibenzopyrenes.
6-23
The milk of nursing mothers who smoke has been found to contain
benzo[a]pyrene in concentrations of 7.6-387 mg/L (average 1 29.5 mg/L)
(Health & Welfare Canada, 1979). There were no comparable data on
non-smoking nursing mothers.
6.6 Trends in Levels and Source Contributions
Although specific estimates of trends in PAH emissions in Ontario are not
available, projections for surrogate pollutants may be used to estimate
trends in total (but not individual) PAH emissions from some sources.
For example, tighter controls on industrial point source emissions of
particulate matter, which are being instituted now and which are projected
to become more stringent under the proposed Ontario Clean Air Program
(CAP), will reduce emissions of particle-bound PAH. Projections of the
provincial emission inventory of particulate matter from combustion
sources, then, should predict the total PAH emission trend for the future.
Based on such a qualitative analysis, the major PAH emission sources
in Ontario identified in Chapter 3 are likely to behave as follows.
It is unlikely that the incidence of forest fires in Ontario will change
significantly over the next few yars. Therefore, it is reasonable to assume
that PAH emissions from this source will remain constant for a ten to
twenty year planning period.
Vehicle emissions of PAH may be assumed to decrease significantly as
more of the fleet is equipped with 3-way converters. Although vehicle
emission control standards are becoming more stringent, projected
increases in the number of vehicles and annual distance travelled per
vehicle will probably counterbalance the decreasing unit emissions. A
6-24
decrease is projected over the next ten years, but an increase will occur
after that, which leads to the conclusion that PAH emissions could be at
present levels at the beginning of the next century. This trend is
predicted for vehicle NO^ emissions (R. Salmon, Environment Canada,
conference presentation, February 1989).
Estimates of residential fuel wood future consumption in Ontario are not
available. It seems reasonable to assume, however, that the increased
use of fuel wood which would be implied by increased population in the
province will not materialize, for several reasons. The escalating price of
fuel wood because of scarcity of supply near urban areas, and the
apparent trend toward fewer detached homes being built in urban areas
would imply a lower per capita fuel wood utilization rate, at least in
southern Ontario. Emission factors may be further reduced by tighter
specifications on residential woodburning equipment or regulation of the
use of such equipment (as now occurs in New England and the Pacific
Northwest of the U.S.) during air pollution episodes. No specific
information is available about regulatory plans in either of these areas in
Ontario. It may be assumed, then, that PAH emissions from residential
woodburning in Ontario will decrease from estimated current levels, but
there is no rational basis for estimating this trend quantitatively.
Control programs for coke oven emissions in Ontario and improved
process technology at newer facilities have reduced particulate (therefore
PAH) emissions from existing plants significantly over the past fifteen
years. PAH emissions from coke ovens would be expected to decrease
in conjunction with particulate emissions. If the best available technology
and best available control technology have already been applied to coke
ovens in Ontario, then it is reasonable to expect that only modest
emission reductions will be achieved in the near futue. In this case,
6-25
emissions from coke ovens in Ontario will be related to future production
levels. In the absence of quantitative estimates of future production or
controls, it is reasonable to assume continuation of coke oven emissions
at apporoximately their current levels, for the purposes of regulatory
assessment of PAH.
The qualitative picture which emerges from the foregoing is a relative
decrease in importance of residential woodburning as a source of PAH
in Ontario, the other major sources identified remaining constant.
Two of the minor sources of PAH identified - municipal incineration and
coal-fired power plants - are and will be subject to particulate matter
emission control programs and, therefore, are not expected to increase
and may decrease in importance relative to other types of sources. The
recent banning in Ontario of incinerators in apartment buildings has
eliminated that source of PAH emission.
Air monitoring data indicate qualitative trends which have occurred
between the mid-1 970's and mid-1 980's, at selected sites in Ontario. The
data presented in Tables 6-3 to 6-7 and 6-10 and in Figure 6-1 show that
typical B[a]P levels in Toronto and Niagara, for example, have decreased
from about 1 ng/m^ annual average to about 0.1 to 0.3 ng/m^ over
similar ten-year monitoring periods. The Toronto data are for two
separate periods about 10 years apart. Comparing the Toronto data of
Katz et al (1978) with those of Environment Canada (Dann, 1988),
average concentrations of other PAH have also decreased between
1975/6 and 1984/6. For example, the annual mean concentrations of
B[e]P, a chemically stable PAH present almost exclusively in particulate
matter, decreased from about 0.75 to about 0.5 ng/m^ over that period
of time. The seasonal maximum (in fall) concentration of B[e]P
6-26
decreased from 1.3 ng/m^ in 1975/6 to 0.6 ng/m^ in 1984/67. B[ghi]
Per decreased dramatically from an annual average of about 7 ng/m^ in
1975/6 to 0.8 ng/m^ in 1984/6. Thus, various PAH have exhibited
reductions in concentration with time to differing degrees, but all major
PAH have decreased, according to the very limited data available. The
two studies cited above as the basis for comparisons in Toronto used
different sampling, extraction and analytical techniques, but there is no
reason to believe that the results are not comparable. These results are
consistent with downward trends observed elsewhere (Dann, 1988), but
data are so fragmentary that calling the existing evidence a trend is not
appropriate.
7-1
7.0 HUMAN EXPOSURE LEVELS (DIETARY AND LIFESTYLE SOURCES)
PAH are widespread in the environment, and various PAH have been
demonstrated to be carcinogenic. Therefore, attempts have been made
to quantify human exposures to PAH, and to assess the relative
importance of sources and exposure pathways. One reviewer
(Santodonato et al., 1981) estimated human exposure to total PAH in
North America to average 1.8-16.2 ug/day, from the inhalation of ambient
air (0.207 ug/day), the ingestion of drinking water (0.027 ug/day) and the
ingestion of food (1.6-16 ug/day).
B[a]P was estimated, by the same reviewer, to constitute approximately
1 0% of the total PAH including naphthalenes; estimates of the percentage
contribution of other carcinogenic PAH were incomplete, due to lack of
data. It can be seen from these estimates that food is considered a
major source of PAH exposure, that there is a high degree of variability
in the exposure estimate, and that the role of carcinogenic PAH, other
than B[a]P, in total human exposure has not been well characterized.
Note also that lifestyle factors which contribute to total PAH exposures,
such as smoking or the use of pharmaceuticals/cosmetics containing
PAH were not included in these estimates.
The following section addresses these issues. The objective of this
section is to review the available data, as well as recent and current work
which contribute to an understanding of human exposures to PAH in
Ontario. Where data specific to Ontario are lacking, other relevant data
are cited.
7-2
7.1 PAH in Ontario Food
The occurrence of PAH in unprocessed foods may be the result of
external deposition, uptake, or biosynthesis. PAH contamination of
unprocessed seafoods is the result of rapid uptake from water, and
accumulation in crustaceans and bivalves (Uthe and Musial, 1986). A
recent review of PAH in the terrestrial environment (Edwards, 1983)
indicated that most PAH contamination of vegetation is by direct
deposition from the atmosphere; that surface-to-mass ratio (relative leaf
area) is a dominant factor in PAH accumulation; that some plants in some
growing conditions can take up PAH and may translocate and/or
concentrate specific PAH in other plant parts; and that concentrations of
PAH are generally greater on plant surfaces, such as peelings, than in
internal tissues. For example, the accumulation of 20 PAH in lettuce
showed a positive gradient with proximity to a busy highway, whereas the
levels in rye did not (Larsson, 1985a). Vegetables such as kale, with
larger surface areas and a longer growing season, accumulate higher
levels of PAH than other foodstuffs (Vaessenetal., 1984) (Grimmer, 1982;
Edwards, 1983).
Bush beans grown in laboratory conditions were shown to assimilate and
translocate PAH from roots to leaves and stems, as had soybeans,
ryegrass, alfalfa, chick pea, and cucumbers in earlier work (Edwards,
1985; Edwards, 1983); but the effects of variables, such as PAH
molecular weight, solubility in water, physio-chemical form in the
substrate, and plant species differences are not yet clear. Nonetheless,
PAH concentrations in peeled onions, beets, oranges and apples have
been shown to be much lower (an order of magnitude) than concentra-
tions in the peels (Edwards, 1983; Grimmer, 1982).
7-3
The occurrence of PAH in processed or cooked foods may be the result
of combustion fumes, formed in hot smoking processes and in the
generation of heat during ghlling processes, reaching the surface of the
food. Endogenous formation of PAH on the surface of food subjected
to high temperatures (e.g., frying or electrical broiling) does not appear
to cause appreciable increases in PAH concentrations (Larsson et al.,
1983). Factors affecting PAH levels in hot smoked foods include the
smoking process (type of fuel, airflow, combustion temperature), smoke
generation techniques, smoke treatment (cleaning) techniques, and
smoking time. Factors affecting PAH levels in grilled foods include the
type of fuel, distance between food and heat source, cooking
temperature, cooking time, and fat content of food being grilled (Larsson,
1985b). For example, beef patties containing 30% fat and grilled over
hickory wood were found to contain 16 PAH in total concentration of 68
ug/kg of meat; the same beef cooked over mesquite wood (from the
southwest U.S.) was found to contain 24 PAH in total concentration of
549 ug/kg of meat (Maga, 1986). Similarly beef containing 10% fat had
half the total PAH concentration of the 30% fat beef. Barbecue briquets
(coal-based and wood charcoal) available to the Canadian consumer in
the mid-1980s were found to contain total PAH concentrations of 2.5 to
13 ug/g of briquet, and contained 30-53 individual PAH in various
samples (Kushwaha et al., 1985). There were no data describing PAH
compositions of combustion fumes, or of PAH concentrations in food
cooked over these briquets.
PAH in smoked foods (and likely in grilled foods as well) are concentrated
in the outer layer. For example, skin from smoked herring had total PAH
concentrations 4-24 times higher than those in the flesh; as well, the
projection of high molecular weight PAH was higher in the skin (Larsson,
7-4
1985b). Therefore, higher concentrations will likely be found in products
or portions with higher surface-area-to-weight ratio.
7.1.1 Analysis of PAH in Foods
Although this is probably a minor source of PAH in food, the potential
for PAH migration from food packaging material is not known. However,
PAH have been observed in some petroleum and synthetic waxes used
in food packaging (Sheraishi et a!., 1975; Howard et al., 1965).
Concentrations of total PAH exceeded 0. 1 ppm in 2 of 32 waxes analyzed
(Howard et al., 1965); B[a]P was observed in average concentrations of
2.9 and 4.6 ppb in petroleum and synthetic wax respectively (Sheraishi
etal., 1975).
There have been several surveys to determine numerous PAH in
Canadian foods. Those involving a variety of food types have been
carried out by the Food Research Division, Health Protection Branch,
Health and Welfare Canada (Lawrence and Das, 1986; Lawrence and
Weber, 1984a, b; Panalaks, 1976). Foods were purchased from local
retail and fast food outlets in Toronto and Ottawa. Samples were
digested with alcoholic KOH, followed by partitioning into solvents such
as cyclohexane or isooctane; interferences were removed by solvent
partitioning and column chromatography on silica gel, Florisil or Sephadex
LH-20. Analyses were done by high pressure liquid chromato-graphy
(HPLC) - fluorescence detector (Lawrence and Weber, 1984b), or UV and
fluorescence detectors (Panalaks, 1976) with some confirmation by gas
chromatography - mass spectrometry/selected ion monitoring (GC-
MS/SIM) (Lawrence and Das, 1986; Lawrence and Weber, 1984a).
Recoveries typically averaged 60-75%, although there were ingredients
of much lower recoveries. Results were not corrected for recoveries.
7-5
The results are summarized below. Where concentrations of total
carcinogenic PAH are given, the totals refer to the sum of individual PAH
analyzed which are considered to be carcinogenic (U.S. EPA, 1987).
1. Leafy vegetables included swiss chard, cabbage, romaine lettuce,
local spinach, and imported spinach (Lawrence and Das, 1986).
They contained the lowest levels of PAH in the foods analyzed. No
PAH were detected in any chard, cabbage or lettuce. Only
fluoranthene was found (32 ng/g) in one spinach (3 samples) in
concentrations of 0.1-0.5 ug/kg.
2. There are no Canadian data, and little other data on PAH in fruits.
One review cited B[a]P concentrations of 0.5-30 ug/kg in fruits
(lARC Vol. 32).
3. Charbroiled hamburgers from four different retail outlets contained
total (of 8 target) PAH in concentrations of 1.8-37.7 ug/kg
(Lawrence and Das, 1986). Concentrations of carcinogenic (6
target) PAH were 1-21 ug/kg, and constituted 47-69% total PAH.
Dimethylbenzanthracene, B[a]P and DB[ah]A constituted over half
the total carcinogenic PAH.
Comparisons of PAH types and concentrations in charbroiled and
fried hamburgers from commercial outlets (Lawrence and Weber,
1984a), demonstrated the effect of the cooking process - of 18 target
PAH, 5 or fewer were observed in fried hamburgers, with total PAH
of up to 1.3 ug/kg; whereas 10-14 were observed in charbroiled
hamburgers, with total PAH of approximately 35 ug/kg.
Concentrations of carcinogenic PAH also increased from 0.3 ug/kg
to 14 ug/kg.
7-6
There are almost no other data on PAH levels in other (unsmoked)
meats cooked by conventional oven or stove-top methods. Other
charcoal broiled meats (steak, pork chops, chicken) were observed
to contain up to 6 (of 9 reported target) PAH, with total PAH
concentrations in the range of 5-55 ug/kg (Panalaks, 1976).
4. Smoked meat products, including bologna, frankfurters, salami,
sausages, bacon, hams, beef and pork were found to contain up to
5 (of 11 target) PAH in some samples of each food type, at
maximum levels of 0.2-8 ug/kg (Panalaks, 1976). Of the 5 most
frequently observed PAH in this group of samples, 4 are
carcinogenic.
More recent work done on smoked cheese, poultry, pork and beef
products in the U.S. (Joe et al., 1984) demonstrated the presence
of 8 PAH (of which 5 are carcinogens) of 1 2 target PAH in most of
the food types analyzed. Total PAH concentrations were 2-7 ug/kg
in the poultry, pork and beef products, and 6-28 ug/kg in the
cheeses. Total (6 target) carcinogenic PAH concentrations were 0.2-
3.4 ug/kg. Sausages and frankfurters were found to have lower
PAH concentrations, with total (of 1 2 target) PAH at concentrations
of 2.4-3.5 ug/kg and total (6 target) carcinogenic PAH at
concentrations of < 0.1 -0.7 ug/kg.
5. Fresh Lake Ontario fish (edible portion) were found to contain 2.1-
7.9 ug/kg total (of 1 1 target) PAH. Concentrations of (5 target)
carcinogenic PAH ranged from 0.3-4 ug/kg, with concentrations in
excess of 0.5 ug/kg found only in smelt and eel.
7-7
Another source of data for PAH concentrations in Ontario fish is the
Ministry of Environment's program to monitor Great Lakes Sport
Fish (Zenon Environmental Inc., 1985). The results from this
program have demonstrated the presence of PAH in numerous
species in all monitored lakes. For example, total (of 10 target) PAH
concentrations in Lake Ontario fish ranged from 2.2-240 ug/kg, and
were found in lake trout (1 of 4 samples), bass (1 of 1 sample),
brown bullheads, red horse and white suckers (all samples).
However, these analyses have been done on whole fish, rather than
on the basis of edible portion. Therefore, their usefulness in evalua-
tion of human exposure through diet is limited.
6. Smoked fish, fresh and canned, were found to contain similar levels
of PAH related to the degree of smoking (Lawrence and Weber,
1984a). Lightly smoked fish, such as smoked haddock, cod, arctic
char and canned sardines, had total (of 9 target) PAH in concen-
trations of 10-15 ug/kg; concentrations of (3 target) carcinogenic
PAH ranged from 0.2-2 ug/kg. In contrast, more heavily smoked
fish, such as smoked herring and digby chix, had total (of 9 target)
PAH concentrations of 30-500 ug/kg, and carcinogenic PAH
concentrations of 0.2-45 ug/kg. Similar results for smoked herring
were found in another study (Alfheim, 1984), where total (of 8 target)
PAH concentrations were approximately 200 ug/kg, and
carcinogenic (6 target) PAH concentrations were approximately 1 5
ug/kg.
The packing oils in the canned products had PAH approximately 5
times more concentrated than in the food itself.
7-8
7. Canned smoked mussels and oysters (imported) were found to
contain total (of 13 target) PAH in concentrations of 65-203 ug/kg,
and carcinogenic (6 target) PAH in concentrations of 4-50 ug/kg
(Lawrence and Weber, 1984a). There was no apparent difference
between mussels and oysters in PAH concentrations. In another
study (Lawrence and Das, 1986) canned oysters (imported) had
similarly high levels of (8 target) PAH (35-112 ug/kg), as well as
carcinogenic (6 target) PAH (3-63 ug/kg).
The packing oils had PAH approximately 7 times more concentrated
than in the food itself (Lawrence and Weber, 1984a).
8. Canned lobster products (lobster spread and lobster meat) were
found to contain total (of 12 target) PAH in concentrations of 8-367
ug/kg, and carcinogenic (6 target) PAH in concentrations of 2-166
ug/kg (Lawrence and Weber, 1984a). Similar variability was
observed in another study (Lawrence and Das, 1986), and it has
been attributed to the use of lobster hepato-pancreas in some
products. This digestive organ has been found to selectively
accumulate PAH by about 10 times compared to tail meat (Dunn et
a!., 1979; Dunn & Fee, 1979). Elevated PAH may be attributable to
creosote contamination during impoundment or to an industrial
source of marine contamination (Uthe, 1986; Williams et al., 1985).
9. Frozen and canned shrimp were found to contain generally lower
levels (of 8 target) PAH - less than 1 ug/kg, except for one sample
of canned shrimp, which contained 8.6 ug/kg (Lawrence and Das,
1986). In this same sample, carcinogenic PAH (B[a]A, B[a]P and
B[b]F) constituted 86% of the total PAH. There was no discussion
as to the reason for the variability.
7-9
10. Milled wheat fractions from wheat grain in southern Ontario were
observed to contain total (of 6) PAH concentrations of 5-12 ug/kg
in flour (Lawrence and Weber, 1984b). Benzo[a]anthracene and
benzo[a]pyrene constituted approximately 5% and 1% respectively
of the totals. PAH concentrations in the bran fraction were about 5-
10 times higher than in other milled fractions, suggesting
contamination was concentrated on the outer portions of the grains.
1 1 . Breakfast cereals included wheat, corn, oats, rice and bran cereals
(Lawrence and Weber, 1984b; Lawrence and Das, 1986). Total (of
11 target) PAH concentrations observed in the earlier work
(Lawrence and Weber, 1984b) ranged from 6-60 ug/kg.
Carcinogenic PAH (of 5 target) concentrations ranged from 0.3-13.3
ug/kg. Concentrations of both total and carcinogenic PAH were
higher in wheat products, and substantially higher in wheat bran
cereal than other cereals examined. In the more recent work
(Lawrence, 1986), results were lower (total of 8 PAH 0.7-3.4 ug/g,
carcinogenic PAH (6 target) 0.02-0.28 ug/kg).
12. Dried dairy products, such as powdered milk, were found to have
total (of 8 target) PAH concentrations of approximately 1 ug/kg and
carcinogenic (of 6 target) PAH constituting 3% (Lawrence and Das,
1986). Similar samples analyzed earlier (Lawrence and Weber,
1984b) were found to vary from below detection limits for all 13
target PAH to total PAH of 8 ug/kg and total carcinogenic (of 6
target) PAH of 2.7 ug/g. The variability was attributed to the drying
process used.
13. Cooking oils were observed to contain total (of 14 target) PAH
concentrations of 0.6-14 ug/kg and carcinogenic (of 6 target) PAH
7-10
concentrations of 0.1-4.5 ug/kg (Lawrence and Weber, 1984b).
These results are similar to more extensive PAH characterizations
done on margarines, butter and vegetable oils elsewhere (Hopia et
al., 1986). It has been noted that PAH concentrations in vegetable
oils are reduced in the oil refining processes, probably due to the
steam deodorizing or filtering through activated charcoal (Larsson
et al., 1987).
14. Tea leaves commonly used in Canada were observed to contain
high residues of benzo[a]anthracene (7.7-11.3 ug/kg) and
benzo[a]pyrene (3.3-4.2 ug/kg) (Lawrence and Weber, 1984b).
However, PAH concentrations in tea infusion are about 1% those of
tea leaves (Vaessen et al., 1984).
There are no Canadian data currently available on PAH
concentrations in coffee or other foodstuffs. However, one review
(lARC Vol. 32) has cited concentrations of B[a]A, B[a]P and
chrysene in roasted coffee as being approximately 0.5-25 ug/kg
each. PAH concentrations in coffee infusion are not given, but can
be expected to be much lower.
7.1.2 Total Diet Studies
A total diet study is currently underway in the Food and Drug Directorate
of Health and Welfare Canada (Koniker, pers. comm.). For the study,
average intakes of various food types were estimated from the Nutrition
Canada Survey (Health & Welfare Canada, undated, ca. 1977) and from
food purchase data (Stats Can, 1982; Family Food Expenditure in
Canada). Foods were purchased locally in Ottawa, were cooked
"normally" (i.e., generally roasted, steamed or fried), were composited
7-11
and analyzed. Preliminary results are similar to what might be expected
from earlier analyses (Lawrence and Weber, 1984a, b; Lawrence and
Das, 1986) of food types (Lawrence, J.F., pers. comm.), but no specific
data are available.
7.1 .3 Summary - PAH in Food Available in Ontario
Observed concentrations of total PAH in foods available in Ontario are
presented in Table 7-1. Results show a large variability in PAH
concentrations among samples of similar foods, as a result of:
0 geographic location of foodstuff origin;
0 method of foodstuff processing;
0 method of foodstuff cooking; or
0 personal food consumption pattern.
Although total diet studies will improve the accuracy of estimating the
average person's exposure to PAH, the potential for large variability in
actual intake is still great, with only moderate variations in dietary
preferences.
7.2 PAH in Ontario Drinking Water
Drinking water supplies of five eastern Ontario municipalities and sixteen
Great Lakes municipalities have been analyzed for various PAH by the
Environmental Health Directorate of Health and Welfare Canada over the
last ten years (Benoit et al., 1979; Williams et al., 1982; LeBel et al.,
1987). Drinking water was sampled using XAD-2 resin sampling
cartridges to extract and concentrate organics. Adsorbed organics were
eluted with acetone/hexane and the eluates concentrated.
7-12
TABLE 7-1
Summary of PAH Levels in
Foods Available in Ontario
Food
PAH
Concentration
(ug/kg)
Fruits & Vegetables
Fresh fish, shrimp, meat (not
charbroiled or barbecued),
flour and dried milk
Oils & Fats
Smoked fish, oyster, lobstres,
lobsters and charbroiled
meats
0.1 - 1
1 - 10
1 -30
10 - 500
7-13
The concentrated eluates were analyzed by gas chromatography - mass
spectroscopy, with compound identification using a mass spectra library.
The (9) compounds monitored were selected on the basis of their
prominence in earlier analyses of Ottawa tap water, and included fluorene
phenanthrene, anthracene, fluoranthene, pyrene (all locations),
acenaphthalene, o-methyl-phenanthrene, benz[a]anthracene and
chrysene (eastern Ontario locations) and methylanthracene (Great Lakes
locations).
Concentrations of individual PAH were generally in the range of 0.1-5
ng/L, with occasional excursions (for example, Sault Ste. Marie -
phenanthrene/ anthracene in summer sample 571 ng/L, in winter sample
1269 ng/L, other PAH were also elevated although not to the same
extent, for example, St, Catharines winter sample, 4 PAH which ranged
from 25-80 ng/L).
7.2.1 Summary
Exposure to potentially hazardous PAH may occur through ingestion of
Ontario drinking water. Substantial fluctuations in concentrations of
individual PAH at a single location have been demonstrated.
Concentrations of individual PAH (of 9 selected for monitoring) were
generally observed in the ppt range.
7-14
7.3 PAH in Pharmaceuticals, Cosmetics
7.3.1 Coal Tars
Coal tar has been in the British Pharmacopoeia since 1893, and is in
three grades. It is also available in the U.S. Pharmacopoeia (USP) (lARC
No. 35). Analyses of pharmaceutical grade coal tars conducted 25 years
ago (lARC No. 35 and/or Lijinsky et al., 1 963) demonstrated the presence
of 15 PAH in concentrations ranging from 0.23-1 7.5g/kg, several of
which are known carcinogens, cocarcinogens, or tumor initiators. These
coal tar ointments have been used for many years for the treatment of
various dermatoses. One study has shown evidence of absorption
through healthy adult skin after application of these ointments
(Steinegger, 1984). After repeated application of 2% coal tar product
(containing 12 PAH in concentrations of 40-650 ug/mL) blood levels of
acetnaphthene, fluorene, phenanthrene, anthracene, fluoranthene, and
pyrene after coal tar product application were elevated from less than .04
to 11 ng/mL; 6 other PAH found in the ointment were not found in the
blood samples. There was no apparent correlation between quantity of
coal tar product used and PAH absorbed, although the small number of
study participants and the uncontrolled application and exposure
procedure may have been confounding factors. It has been noted that
abraded or diseased skin may result in increased percutaneous
absorption. It has also been noted that the use of a coal tar containing
shampoo resulted in induction and/or enhancement of the enzyme
activity considered critical to cancer induction (Merk et. a!., 1987).
However, there are few other data characterizing exposure - absorption
potential.
7-15
7.3.2 Soft and Liquid Paraffins, Mineral Oils
White soft paraffins and liquid paraffins from petroleum are widely used
for external medicine and cosmetic purposes, as ointment bases for
suntan oils, creams, baby toiletries, and others. Analyses of several (2)
paraffin samples of each type demonstrated considerable variability in the
number and concentrations of single PAH, even in the same type of
product (Monarca and Fagioli, 1981). For example, the total (of 7) PAH
was 6.1-82.6 ng/g (white soft paraffin) and 30.1-30.5 ng/g (liquid
paraffin); carcinogenic PAH (2) concentrations were 0.1-1 1.6 ng/g (white
soft paraffin) and 3.1-10.6 ng/g (liquid paraffin). All samples contained
B[a]P.
Analyses of four commercial samples of suntan oils (available in Italy),
based on mineral or vegetable oils, showed that concentrations of total
(5) PAH ranged from 89-189 ng/g, and concentrations of B[a]P (found
in all samples) ranged from 1.5-45.7 ng/g (Monarca et al., 1982).
Anthanthrene (1.2 ng/g) was observed in the one sample based on
mixed vegetable oils, which was also the sample with the highest B[a]P
and total PAH concentrations. Other carcinogenic PAH, such as B[a]A,
chrysene, were not detected.
7.3.3 Summary
Exposures to potentially hazardous PAH may occur through the use of
pharmaceuticals based on coal tar, and personal care products such as
suntan oils and presumably other cosmetics with similar bases. Coal tar
pharmaceuticals have PAH concentrations in the ppm range as well as
medicinal paraffins. Use of these products may have effects both directly
on skin and systemically (through absorption).
7-16
7.4 PAH in Indoor Air
Sources of PAH in indoor air include:
0 outside (ambient) air, infiltrating the building envelope;
0 combustion processes (such as wood stoves and kerosene heaters) ;
and
0 tobacco smoke.
7.4.1 Effects of Combustion Processes (for Heating)
It is clear that wood smoke particulate matter contains PAH, and that the
emission of PAH type and amount during wood burning depends upon
the kind of wood burned, the moisture content and the burning regime.
PAH emission factors have been determined for fireplaces, baffled and
non-baffled stoves burning two different wood types (Peters et al., 1981;
Travis et al., 1985). However, these data were determined by measuring
flue gas constituents; they are therefore not indicative of PAH concentra-
tions inside houses were wood is burned.
The effects of the use of wood burning stoves on indoor-outdoor PAH
concentrations have been carried out in Whitehorse by the Monitoring
and Criteria Division, Environmental Health Directorate, Health and
Welfare Canada (R. Otson, pers. comm.). Analyses were done for over
100 aromatic compounds, including many PAH. As many homes in the
Whitehorse area are heated by wood, the study is expected to
demonstrate the effects of the use of wood stoves on indoor air quality.
Preliminary results indicate that PAH were largely seen in indoor air
samples (detection limits estimated to be approximately 5 ng/m^ for
higher molecular weight PAH approximately 3 ng/m^ for lower molecular
7-17
weight PAH). PAH concentrations indoors were also generally lower
than those outdoors, indicating that neither outdoor air nor backdrafting
from the wood stoves was contributing significantly to long-term average
indoor PAH concentrations.
An American study of 24 homes in Vermont (Sexton et al., 1984) also
demonstrated that outdoor concentrations of PAH frequently exceed
those in homes with wood stoves. Total (of 8 target) PAH indoors ranged
from 1.3-20.4 ng/m^, those outdoors ranged from 3.0-35.3 ng/m^. Total
(of 5) carcinogenic PAH indoors ranged from 0.9-15.3 ng/m^.
A Norwegian study (Alfheim, 1984) demonstrated that burning wood in
an "airtight" stove caused small increases in indoor total PAH
concentrations. Total (of 30 target) PAH indoors ranged from < 1 ng/m^
in an electrically heated house to 16 ng/m^ in one with a wood stove
operating normally. However, when wood was burned in an open
fireplace, indoor PAH concentrations increased substantially to 150-206
ng/m^ (B[a]P concentrations 13-18 ng/m^).
PAH are also emitted during kerosene combustion, 18 PAH have been
identified in kerosene soot (Kaden et al., 1979). Portable kerosene space
heaters which are common in the U.S. were studied recently, in well-
tuned and badly-tuned operating conditions (Traynor et al., 1986).
Phenanthrene and fluoranthene were observed in all tests, with source
strengths of 1.9-16 ug/h (phenanthrene) and 0.1-1 .8 ug/h (fluoranthene).
Anthracene, chrysene, gnd indeno[1,2,3,-cd]pyrene were also observed
in individual tests, with source strengths of 2.27, 0.05 and 0.12 ug/h
respectively. Another study demonstrated the presence of benzo[a]-
pyrene, benzo[b]fluoranthene, and benzo[k]fluoranthene in kerosene
heater emission particulate matter (Tokiwa et al., 1985). However, there
7-18
was not enough information given to determine PAH emission rates.
There were no data on PAH concentrations in air resulting from the use
of kerosene space heaters.
7.4.2 Effects of Tobacco Smoke
An eight home pilot study of indoor-outdoor concentrations of PAH and
PAH derivatives was also carried out recently by the U.S. EPA (Wilson
and Chuang, 1987). Its objective was to compare concentrations in
homes with and without smokers. Preliminary results indicate indoor
concentrations of PAH (14 target compounds of interest) in homes
without smokers were generally slightly higher than outdoor
concentrations (although the differences may not have been analytically
or statistically significant), and that all target compounds had higher
concentrations in homes with smokers. For example, concentrations of
individual PAH for a non-smoking house were 0.18-29 ng/m^ (outdoors),
0.18-59 ng/m^ (indoors), total (of 14) PAH 81-97 ng/m^; for the smoking
house, they were 0.34-54 ng/m^ (outdoors), 0.64-210 ng/m^ (indoors),
total (of 14) PAH 121-245 ng/m^
Concentrations of (13 target) PAH were monitored in a 36 m^ room with
a single air change per hour, under conditions of no smoking and
smoking approximately 40 cigarettes over an 8-hour period (lARC No.
29). Total PAH in non-smoking conditions was 134 ng/m^, in smoking
conditions averaged 429 ng/m^. Concentrations of individual PAH ranged
from <2-50 ng/m^ (no smoking) and from <2-116 ng/m^ (smoking);
concentrations of B[a]P increased with smoking from <3 ng/m^ to 22
ng/m^
7-19
Concentrations of (6 target) PAH were also monitored in a 38 m^ closed
room with natural ventilation, under no smoking and smoking 15-30
cigarettes (Grimmier et al., 1977). Total PAH in non-smoking conditions
was 32 ng/m^ in smoking conditions averaged 214 ng/m^
Concentrations of individual PAH ranged from < 1 -75 ng/m^ (no smoking)
and from 27-214 ng/m^ (smoking); concentrations of B[a]P increased
with smoking from 5 ng/m^ to 88 ng/m^.
Numerous (38) analytical studies have shown the presence of 37 PAH
in tobacco smoke, and quantitative data on their occurrence in cigarette
mainstream and sidestream smoke, cigar and pipe smoke are
summarized in one review (lARC Monograph #38, 1983). For example,
mainstream cigarette smoke contains the following amounts of
carcinogenic PAH:
Benz[a]anthracene
Ben2o[b]fluoranthene
Benzo[f|fluoranthene
Benzo[c]phenanthrene
Benzo[a]pyrene
Chrysene
Dibenz [a, c] anthracene
Dibenz [a,h]anthracene
Dibenz [a,j] anthracene
Dibenzo[a,e]pyrene
Dibenzo[a,h]pyrene
Dibenzo[a,i]pyrene
indeno[1,2,3-cd]pyrene
UQ/1QQ cigarRttp?^
0.4-7.6
0.4-2.2
0.6-2.1
present
0.5-7.8
0.6-9.6
present
0.4
1.1
present
0.17-0.32
0.4-2.0
7-20
Constituents in sidestream smoke may vary somewhat in relative
proportion to those in mainstream smoke, but are present to the same
order of magnitude.
Total PAH in mainstream cigarette smoke range from 0.31-2.1
ug/cigarette (sum of individual PAH concentrations found by different
authors, generally using non-filter cigarettes smoked under standardized
lab conditions). It should be noted that most of the data are based upon
smoking patterns of 30 years ago, which are not the same today.
Conventional filter types can be expected to reduce compounds found
in the particulate phase (such as PAH); however, reduced nicotine
delivery induces the smoker to puff more frequently and inhale more
deeply; also, by obstructing the holes in perforated filter tips, the smoker
can inhale more smoke than would be expected. Therefore, the above
data may be subject to greater variability than indicated by the ranges
given.
7.4.3 Indoor PAH from Miscellaneous Sources
PAH result from the combustion or pyrolysis of carbonaceous matter.
Therefore, one potential domestic source for PAH is pyrolysis during
cooking. No studies were found concerning PAH concentrations in air
during certain domestic high temperature cooking processes (e.g., deep
fat frying). However, the studies of airborne PAH in homes cited earlier
(Otson, pers. comm.; Wilson & Chuang, 1987) involved long-term
monitoring in several locations (including kitchens), and no cooking
effects were noted.
Emissions arising from pyrolysis of insecticidal coils have been found to
contain 30 PAH in the air-entrained particulate, at concentrations of 500-
7-21
700 ng per mg particulate (Lazaridis, 1987). Carcinogenic PAH
(benzo[a]pyrene, benz[a]anthracene, chrysene, and indeno (1,2,3-cd)
pyrene were in concentrations of 131-169 ng/mg. Assuming a tiypotlie-
tical use of one coil in a 250 m^ summer house with 3 air changes per
hour, good mixing, and a 10 hour burning time, total PAH concentration
is estimated to be 15 ng/m^.
7.4.4 Summary
Indoor PAH concentrations are generally less than, and sometimes similar
to those in ambient air, for homes without combustion devices and
smokers. The use of air-tight wood burning appliances generally
increases indoor PAH concentrations only slightly or not at all; the use
of open fireplaces increases total PAH concentrations substantially,
concentrations of 200 ng/m^ having been observed. The use of
kerosene space heaters is expected to elevate total PAH concentrations,
but the effect has not been characterized. Tobacco smoking in the home
increases PAH concentrations substantially - total (of 14) PAH
concentrations of 121-245 ng/m^ have been observed in homes.
Tobacco smoking appears to be the most significant single source of
PAH exposure, with the mainstream smoke of one cigarette estimated to
contain 0.31-2.1 ug total PAH and 0.05-0.33 ug carcinogenic PAH.
7.5 Estimated Levels of Human Exposures
The major routes of exposures to PAH appear to be ingestion of food,
inhalation of tobacco smoke, inhalation of ambient air, and ingestion of
drinking water. Average exposures can be estimated, based upon
observed PAH concentrations in the various media and on average rates
of ingestion/inhalation.
7-22
A summary of estimated average exposures has been compiled in Table
7-2, based upon data given in the previous sections. The assumptions
made in deriving the estimate are also in the table. The purpose of the
estimate is to illustrate the relative importance of exposure
routes/consumption patterns in PAH exposures.
There are several weaknesses in this estimation process. The data for
"total PAH" and "carcinogenic PAH" are not entirely comparable, as
different studies selected different target PAH. Therefore, reported levels
may be biased low, because they do not include PAH not looked for.
Similarly, most authors' results have not been corrected for analytical
recovery rates, which vary substantially with individual PAH, but may
average 60-90%.
7-23
TABLE 7-2
Estimated Average PAH Exposures
per MMdual In Ontario
1. Ingestion of Food
.6
Fresh fruits
Vegetables
Meat, fresh & canned,
beef, pork
Chicken and other
Meat, cured/prep 'd
Rsh, fresh, fry, canned
Fish, cured
Shellfish
Cheese
Oils, fats
Cereals, Grains
Rice
Rour and mixes
Baked goods, pasta"'
Milk, dried
Milk, fresh fluid
TotsU Food
2. Ingestion of Drinking Water
3. Inhalation
5000 m-'/y (WHO) (ng/m^
Ambient (0.3)
Indoor (0.7)
- non-smoking
- smoking/fireplaces
Total Air
TOTAL
Smoker
Estimated
Intake
kg/yr
SO
40
73
15
A£
0.1
0.4
6.0
8.0
5.6
2.9
8.2
93
0.3
93L
2 L/day
20
Total PAH
Carcinogenic PAH
Mid-Range
Mid-Range
Cone.
Intake
Cone.
Intake
ug/kg
ug/yr
ug/kg
ug/yr
0.3
261
0.3
26
1.3
52
0.3
52
18
126
10
70
5
75
1.6
24
38
16
0.4
2
10
1
1.0
0.1
172
69
23
9
11
7
1.6
10
6.7
54
2.2
18
2.0
11
no data
0.1
0.6
8.5
70
0.5
4
8.5
791
0.5
47
3.5
1
no data
1.4
0.4
1299
263
(ug/yr)
(ug/yr)
750
36.5
2-^%
1.5
ng/L
ng/L'
ug/yr
30
22
20
200
70
700
100-730
ug/yr
1.5-2.1
mg/yr
2
20
7
70
10-73
ug/yr
0.45-09.51
mg/yr
10 cig/day
0.9
ug/cig
3.3 mg/yr
0.14
ug/cig
0.51 mg/yr
Footnotes:
Intakes estimated from reference (Statistics Canada - Family Food Expenditure In Canada 1982, Table 13, based
on 2.74 persons per Ontario feunily reported In Table 1).
Assumed fried, steamed, roasted conventionally.
Assumed charbroiled/barbecued.
Assumed [PAH] of canned, smoked oysters.
Assumed (PAH) from more recent data.
Assumed [PAH] of flour.
No Canadian data. Assumed U.S. average (Santodonato et al., 1981).
7-24
Perhaps the major weakness in the estimation process is the large
variabiity in the observed PAH concentrations in foods, even similar food
types sampled and analyzed at the same time. However, with this in
mind, the estimates do suggest the following:
1 . That mainstream tobacco smoke is the single largest source of PAH
and carcinogenic PAH exposure.
2. That sidestream tobacco smoke is a significant contributor to total
PAH exposures of non-smokers.
3. That for non-smokers, ingestion of food constitutes the major
exposure pathway.
4. That a moderate change in food consumption patterns can result in
a significant variation in PAH intake.
8-1
8.0 ASSESSMENT OF ENTRY, MOBILIPf AND FATE OF POLYCYCUC
AROMATIC HYDROCARBONS
This section attempts to integrate and interpret the data presented in the
previous sections by means of rudimentary modelling results and evalua-
tion of modifying factors which may be important in estimating a PAH
budget for Ontario.
Several important aspects are discussed.
o Implications of physical partitioning of PAH into environmental media.
o Implications of chemical transformations (differing for vapours and
particles).
o Estimates of total and speciated PAH emissions in Ontario.
o Evaluative models for PAHs and PAH mixtures, as they relate to
estimation of exposure pathways.
0 Source - receptor relationships, especially for mixtures.
8.1 Physical Partitioning and Chemical Transformation of PAH and PAH
Mixtures
It is important to determine whether PAH mixtures can be traced from
source to receptor, more to the point, from multiple sources to a receptor
through multiple pathways (source apportionment). The potential for
determining characteristic source profiles to be used in source
apportionment has been discussed in Chapter 4. The relative con-
8-2
centrations of PAH in a mixture emitted or discharged from a source
begin to change upon entering the environment, by both physical and
chemical processes. This section describes how these processes
influence the "traceability" of PAH from source to receptor.
The vapour- and particulate-phase components of emissions to the
atmosphere will partition differently. Coarse particles will deposit near the
source. Particles smaller than about 20 um aerodynamic diameter, which
are dispersed in the atmosphere and behave as passively transported
contaminants, will behave differently from vapours or gases, primarily
because of their greater susceptibility to precipitation removal (rainout or
washout). PAH vapours are hydrophobic (see Table 2-2), so that their
removal in rain or snow will be minimal, and their lifetimes in the
atmosphere are very likely to be somewhat longer than those for PAH
predominantly bound to particles, at least with respect to physical
processes.
Vapour-particle partitioning of PAH will change as a contaminated air
mass disperses, since the fraction of a given PAH that is bound to
particles depends not only upon its vapour pressure (or related sorptive
properties) but also upon the absolute particle concentration (Cupitt,
1980; Yamasaki et al., 1982; Mackay et al., 1986). Thus, PAH will be
partitioned more toward the vapour-phase in conditions of lower particle
concentration, such as in rural areas away from sources.
Chemical transformation of vapour-phase PAH and particle-bound PAH
differ (Korfmacher et al., 1980). Vapour-phase chemical reactions will
depend, of course, on the presence of co-pollutants in the air mass into
which the PAH are emitted and the intensity of solar actinic (i.e.,
photochemically active) radiation. The rate of nitration of pyrene, for
8-3
example, depends upon the available concentrations of oxidant, nitrogen
oxides and water vapour. See Chapter 5 and a review of the mechanism
by Finlayson-Pitts and Pitts (1986). Furthermore, atmospheric reactions,
nitration, for example, do not stop with the formation of the initial reaction
product. Nitro-PAH have been shown to photolyse readily to produce
keto-derivatives (quinones). That is, each reaction is one step toward
complete oxidation of the hydrocarbon. Oxidation eventually leads to
cleavage of aromatic rings to form aldehydes and acids. See Finlayson-
Pitts and Pitts (1986, p.935) and the discussion in Chapter 5 on the
analogous aquatic and terrestrial fate data.
The foregoing implies that, at any time after release of emitted PAH
mixtures, both physical and chemical separation processes will have
changed the composition and properties of the mixture. A possible
exception to this may be emissions into a very cold, dry, and othenwise
unpolluted air mass experiencing low levels of solar actinic radiation.
Under the latter circumstances, which may obtain during a portion of
Ontario's winter, airborne mixtures may be transported significant
distances with relatively little modification, except for settling of larger
particles. Evidence for this hypothesis is lacking, but the relatively slow
transformation rates of PAH in LRTfrom Europe to Scandanavia observed
by Nielsen et al. (1983) indicate that under certain atmospheric
conditions, little transformation may take place.
Thus, at a receptor site, the composition and characteristics of the
contributions of various sources to total PAH burden will be intermediate
between their initial characteristics and their ultimate state. The degree
of modification depends on the time between emission and reception.
8-4
Near emission sources, the particle-piiase PAH will have changed
composition least and will likely be a relatively stable indicator of source
origin. Away from dominant emission sources, where both direct source-
receptor and indirect pathways (influenced by multi-media partitioning and
transformation) contribute significantly, source information in the chemical
composition of the PAH fraction may be unresolvably masked. Source
information is also confused by mixing with emissions from multiple
sources.
The foregoing suggests that the only reliable indicator within the PAH
fraction of PAH burden in the atmosphere is that set of PAH which
predominates in the particulate phase. Simple fugacity modelling to
demonstrate which of the priority PAH behave in that manner is described
in the next section. This argument also suggests that constituents other
than PAH of media through which exposure may occur, such as chemical
elements or stable compounds, ought to be considered in assessing PAH
source contributions to exposure.
Nielsen (1984) has described a reactivity classification system for PAH
(Section 5.1.1), which includes explicitly 25 different PAH. Only those
PAH which are at least as stable as, for example, the benzopyrenes
(Nielsen's Classes II and III) are likely to be useful to characterize PAH
mixtures for source apportionment purposes.
8.2 Modelling the Fate of PAH Compounds
The fate of polycyclic aromatic compounds is dependent on their
physical, chemical and biological properties, as well as the environmen-
tal conditions such as temperature, transport parameters and biological
activity. It is possible to model this fate at various levels of sophistication
8-5
and accuracy dependent on the availability and reliability of data. Mackay
et al. in a series of papers have presented a structured format for
performing these fate assessments at various levels of sophistication.
This work will apply the "fugacity" models of Mackay to the extent that is
justified by the available data.
As an introduction to the application of fugacity models, it is appropriate
to provide a brief review of the concepts. Mackay and Paterson (1982)
have described four levels of environmental transport and fate model
calculations based on the fugacity approach (Level I to Level IV).
Fugacity is a thermodynamic quantity related to chemical potential (or
activity), which characterizes the escaping tendency (viz., the Latin root
of fugacity) of a chemical from one phase to another. By appropriately
defining environmental compartments and their components as confined
phases, the model calculations may be used to characterize the
movement, depletion and accumulation of chemicals among such
compartments.
The Level I fugacity model treats the apportionment of a chemical among
environmental compartments at equilibrium steady state, without mass
transfer flow between compartments. Levels II, III and IV treat
progressively more complex situations, culminating in the simulation of
a nonequilibrium, nonsteady-state flow system at Level IV. The latter
model is necessasry to simulate the behavior of the real environment,
which has time-dependent emissions and discharges. The choice of any
of the fugacity models depends on the amount and quality of data
available for a given chemical.
The principal objective of this section is to characterize the tendencies of
the priority PAH to move into the various environmental media
8-6
compartments. A Level I model suffices for this purpose. In a later
section, an example Level III (non-equilibrium, steady-state, with flow)
calculation for B[a]P is presented.
Additional information about this approach to modelling environmental
behavior of chemicals may be found in Cohen (1986); Ryan and Cohen
(1986); Mackay, Joy and Paterson (1983); Mackay, Paterson and
Schroeder (1986); IJC (1987, 1988) and Cohen and Ryan (1985).
The next section describes the Level I calculations for the priority PAH,
with the objective of evaluating their potential for partitioning into and
accumulating in selected environmental compartments.
Level One
This is the simplest of the models and easiest to apply. The necessary
compound data are molecular weight, water solubility, vapour pressure
and octanol-water partition coefficient. The compound is then distributed
in the environment based on these properties and the assumption that
equilibrium has been achieved. The assumption of equilibrium holds for
real partitioning processes in the absence of competing processes, given
enough time. The equilibrium partitioning indicates the tendency of a
chemical to accumulate in any of the media or phases included. This
distribition may represent the steady-state distribution in more complex
(i.e. realistic) situations. In a later section, the results of the simple
calculations are compared with more realistic model results to indicate the
level of confidence which can be placed in the Level I model.
It is not necessary at this level to know the emission rate of the com-
pound because equilibrium is assumed and because no statements will
8-7
be made with respect to the life expectancy of the compound. The result
of this level is simply the distribution of the compound expected in a
typical environment and at equilibrium. The volumes of the various
environmental segments are given in Table 8-1. These values were
chosen to be representative of Southern Ontario. As can be seen, the
environment is divided into six compartments (air, water, soil, sediment,
suspended sediment and biota). A number of other assumptions are
included in this application of the model. The partitioning between solid
and water is described by a partition coefficient, Kp, which can be
obtained from the organic carbon-water partition coefficient, Koc, using
the fraction organic carbon in the soil (foe).
The soil is assumed to have a fraction organic carbon of 2%. The value
of organic carbon-water partition coefficient can be obtained from the
octanol-water partition coefficient by the use of the following correlation:
Koc = 0.41 K,,
A similar process is used for the sediment and the suspended sediment,
except the fraction organic carbon is 4%. For biota, the partitioning is
handled by a bioconcentration factor which can be obtained from the
octanol-water partition coefficient using the following correlation:
BCF = 0.048 K^^
"Air" refers only to the vapour phase in this model. Later on, in Section
8-4, a model including airborne particles is described. In any case, the
small mass and short lifetime of airborne particles mean that the capacity
of this sub-compartment is very small compared with others.
8-8
TABLE 8-1
Volumes of Environmental Segments
Fugacity - Level 1
UNIT WORLD INFORMATION
Volume
Foe*
Density
(g/m')
Air
6.0E + 09
1
Water
7000000
1000
Soil
45000
0.02
1500
Sediment
21000
0.04
1500
Suspended Sediment
35
0.04
1500
Biota
7
1000
Temperature (°C)
20
* Fraction Organic Carbon
8-9
8.3 Results and Discussion
The property data for the various PAHs have been presented elsewhere
in this report (see Table 2-2). Although only a small amount of proper-
ty data is necessary to perform the "level one fugacity model", the data
are unavailable for a number of the PAH in this study. Shown in Table
8-2 are the level one results for 22 of the 47 PAH in the study. Of the
remaining 25 compounds, 20 are not included because the vapour
pressure had not been measured nor estimated. The remaining 5 were
not included because more than one of the required data was
unavailable.
In Table 8-2 the mole percentage distribution is given for each of the
PAH, as well as the property data that were used to perform the
calculation. The 4 compartments: soil, sediment, suspended sediment
and biota are combined into a compartment called 'organic'. The
justification for this is that four compartments are all represented by the
affinity of the compound for the organic phase.
To assist in the interpretation of these results, Table 8-3 has been
prepared to illustrate the distribution of the compounds. Table 8-3 is a
matrix grouping the compounds of similar distribution together. The
classes are dominant, significant, slight and trace for each compartment.
Pyrene results show that the air, water and organic distribution is 1 .8, 5.6
and 92%, respectively. Thus, pyrene can be found in the dominant
category for the organic phase, slight category for the water phase and
slight category for the air phase.
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8-11
TABLE 8-3
Distribution of PAH into Air, Water and
Organic Components BasecJ on Raw Property Data
AIR
WATER
ORGANIC*
Dominant
ACE
QUINOL
A
DB[AC]A
(>50 mol%)
ACR
F
TRI
CARB
B[A]A
B[A]P
B[E]P
B[GHI]PER
B[K]F
CHR
COR
INP
MPHEN
PHEN
P
DB[AH]A
Significant
ACEY
A
ACEY
(<50 mol%
FLN
FLN
FLN
>10 mot%)
PHEN
PHEN
CARB
Slight
A
ACE
F
ACE •
{<10 mol%
P
ACEY
INP
ACR
>0.1 mol%)
QUINOL
ACR
B[A]A
B[A]P
B[E]P
CHR
MPHEN
P
F
QUINOL
Trace
B[A]A
COR
B[GHI]PER
TRI
TRI
(<0.1 mol%
B[A]P
B[E]P
B[GHI]PER
B[K]F
CHR
DB[AC]A
INP
MPHEN
DB[AH]A
CARB
B[K]F
COR
DB[AC]A
DB[AH]A
* Includes soil, sediment, suspended sediment and biota.
See Table 2-2 text for full name of compounds.
Table is based on a Fugacity Level I calculation.
8-12
These distribution data can be of assistance in setting priorities for
obtaining additional data. That is, for a compound such as thphenylene,
the important reaction rates will be those in the air phase because the
majority of the compound will reside in that phase. The reaction rates in
the soil phase will be of secondary importance and therefore do not need
to be known as accurately.
Table 8-4 is a summary of the results of a sensitivity study for each of the
compounds. The table includes a range of vapour pressure, water
solubility and K^^ values from the literature. The Level I distribution in the
three phases is given for each extreme of one property holding fixed the
base, or preferred, values of the properties (Table 8-2) for the other para-
meters. As an example, pyrene tests the sensitivity of the results to
vapour pressure by using 7.0E-5 and 700 E-5 Pa in conjunction with a
water solubility of 0.16 g/m^ and log K^^ of 5.03.
The result of this sensitivity test is a distribution range of 0.2-15, 5.7-4.9
and 94-80 mole percent for the air, water and organic phases,
respectively. A similar analysis was performed for the water solubility and
Kq^ and also on each of the compounds. The choice of the range of data
to test was based on one of two criteria, whichever produced the larger
range, as follows. The first is vapour pressure, a factor of 5 in each
direction, water solubility, a factor of 2, and log K^^, plus or minus 0.2.
These are based on a subjective assessment of the data in the database.
The second criterion was based on the premise that if data for a
particular compound suggest a wider range is necessary, then that range
was used.
8-13
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8-14
TABLE 8-5
Distribution Changes Due to Vapour Pressure Sensitivity
AIR
WATER
ORGANIC*
Dominant
(>50 nnoi%)
ACR
F
acey
fin
phen
acr
PHEN
acey
f
fin
Significant
(<50 mol%
>10 mol%)
ACEY
FLN
PHEN
a
P
FLN
PHEN
ace
acey
ACEY
FLN
ace
acr
phen
Slight
(<10 nnol%
>0.1 mol%)
A
P
acr
f
fin
phen
carb
ACE
ACEY1
ACR
F
fin
phen
ACE
ACR
F
fin
Trace
(<0.1 mol%)
CARB
f
* Includes soil, sediment, suspended sediment and biota.
See text for full name of compounds.
Capital letters refer to positions consistent with Table 8-3 while lower case letters refer to potential
classification changes due to errors in the value of the vapour pressure of the compound.
8-15
TABLE 8-6
Distribution Changes Due to Water Solubility Sensitivity
AIR
WATER
ORGANiC^
Dominant
(>50 nnoi%)
acey
fin
phen
QUINOL
Significant
(<50 mol%
>10 moi%)
ACEY
FLN
PHEN
a
quinoi
FLN
PHEN
acey
quinoi
Slight
(<10 nnol%
>0.1 mol%)
A
QUINOL
b[a]p
ACEY
fin
phen
Trace
(<0.1 mol%)
B[A]P
PHEN
ACEY
ACE
acey
ace
phen
* Includes soil, sediment, suspended sediment and biota.
See text for full name of compounds.
Capital letters refer to positions consistent with Table 8-3 while lower case letters refer to potential
classification changes due to errors in the value of the water solubility of the compound.
8-16
TABLE 8-7
Distribution Changes Due to Kq,^ Sensitivity
AIR
WATER
ORGANIC*
Dominant
(>50 mol%)
acey
fin
PHEN acey
Significant
(<50 mol%
>10 mol%)
ACEY
FLN
a
ANTHRACE
PHEN
ACEY ace
phen
Slight
{<10 mol%
>0.1 mol%).
A
b[k]f
phen
ACE
Trace
(<0.1 nnol%)
B[K]F
* Includes soil, sediment, suspended sediment and biota.
See text for full name of compounds.
Capital letters refer to positions consistent with Table 8-3 while lower case letters refer to potential
classification changes due to errors in the value of the Ko^of the compound.
8-17
To assist in the interpretation of these sensitivity tests, Tables 8-5, 8-6 and
8-7 were prepared. In each table, the potential classification changes with
respect to Table 8-3 are given.
For example, in Table 8-5, the base location of pyrene is given in the left
hand column of each compartment pair. As a result of realistic changes
in vapour pressure, the classification of pyrene in the air phase may
move from slight to significant (right hand column). Compounds which
do not change classification are not shown.
Although the classifications are subjective, some interesting conclusions
can be made. It is evident that uncertainty associated with the vapour
pressure of some compounds can lead to significant changes in the
predicted distribution of that compound.
Combining this with the fact that vapour pressure data are unavailable for
20 compounds, it is evident that a principal weakness in PAH data is
vapour pressure. The uncertainty in the water solubility and K^^ is of
secondary importance.
Some of the results in Tables 8-2 and 8-3 may be counter-intuitive for
some PAH. For example, pyrene might be expected to be found primarily
in air because of its relatively high vapour pressure. The results of the
sensitivity analysis shown in Table 8-4 indicate that the range of expected
partitioning of pyrene may be from 0.2 to 15% into the air compartment,
and that for the entire range of current values of its properties, pyrene is
expected to be found predominantly associated with the "organic"
compartments. It is an experimental fact that airborne pyrene does
appear predominantly in the vapour phase (e.g., Dann, 1988),
undoubtedly due to its moderately high vapour pressure. The Level I
8-18
model is suggesting, however, that because of pyrene's high water
solubility and moderately high affimity for organic carbon (KqJ, it will tend
to end up in the long term in the "organic" compartments, like many of
the heavier PAH with much lower vapour pressure.
Table 8-4 indicates the significant ranges of predictions that are possible
based on available physical property data. Tables 8-5 to 8-7 reflect those
ranges calculated in the sensitivity tests.
8.4 Advanced Model Predictions
The level one model results are shown in Table 8-2 and the results of
sensitivity tests are shown in Table 8-4. The distribution data should be
used to decide on priorities for the obtaining of secondary property data.
Vapour pressure data uncertainty and unavailability is the greatest
shortcoming of the available PAH database.
As described by the model of Mackay et al. (1986), certain PAH may
cycle many times between aqueous and atmospheric media before
reaching their ultimate residence. Based on the foregoing analyses, the
ultimate media will be sediments and soils for most PAH. Recognizing
that atmospheric transformation processes are oxidative, it is probable
that sequential oxidative steps will make the PAH less hydrophobic
(phenolic and acidic moieties being added) and with decreasing vapour
pressures, so that the probability of incorporation into surface soil or
suspended sediment and then bottom sediment increases.
Until reaching buried sediments or sub-surface soils, where anaerobic
(reductive) chemical and microbiological processes dominate, it must be
recognized that PAH are not as persistent as many of the environmentally
8-19
longer-lived organic compounds, such as the organochlorine pesticides
and related substances. See Chapter 5. This means that one should
attempt to characterize PAH mixtures only by their most stable, particle-
bound components. These components include, of course, non-PAH
compounds, and in particular, elements and inorganic compounds.
The model for B[a]P distribution in the Ohio River ecosystem proposed
by Ryan and Cohen (1986) is a convincing example of the
appropriateness of concentrating on the stable, particle-bound portion of
the PAH 'spectrum'. This one-dimensional, dynamic, multi-media
partitioning model presumes that transport of B[a]P in gaseous or
dissolved form is negligible. Its results are in good agreement with
observation.
Using actual regional emission and discharge data, Ryan and Cohen's
model predicted that by the twelfth year of constant release rates of
B[a]P, a "pseudo-steady" state had been reached. More than 99% of the
B[a]P mass was predicted to reside in surface soil. Sediment was not
a modelled medium, but the Level I fugacity results presented earlier in
this section and the Level ill calculations presented below suggest that
B[a]P would likely partition to both surface soil and sediment.
The excellent agreement between the predictions of the Ryan and Cohen
model and reported local measurements of the various media suggests
that the long-term partitioning of B[a]P in similar multi-media
environments, such as the Great Lakes Basin, would strongly favour soil
and sediments. The same fate would be expected for PAH with physical
and chemical properties similar to those of B[a]P.
8-20
The relatively simple static input-output mass balance model of Strachan
and Eisenrich (IJC 1987, 1988) also focusses on B[a]P as representative
of PAH and also presumes that vapour-phase B[a]P is not important.
Since this model uses observed environmental concentrations as inputs,
it cannot be calibrated against these data.
Cupitt (1980) also presumes that only particulate B[a]P (again,
representing PAH, or more generally, polycyclic organic matter) is
important in determining fate and atmospheric residence time.
Mackay and Paterson (1988) have carried out advanced Level III fugacity
modelling for B[a]P in an environment defined by the conditions and
dimensions of southern Ontario. This model incorporates detailed
advection, degradation and intermedia transfer rates, as well as estimated
emission (discharge) rates to air, water and soil. It also predicts that the
predominant steady-state partitioning is to soil and sediments. That is,
the equilibrium partitioning predicted by the lower-level fugacity models
is confirmed by the higher-level models incorporating kinetics.
For example, the Level I and Level III models for B[a]P predict the
following distributions (%):
Air
Water
Soil
Sediment
Level 1
0
0.2
52
48
Level III
0.1
5
72
23
That is, the predictions of the different levels of fugacity modelling are
similar. This general agreement lends credibility to the use of the Level
I calculations for comparing the environmental behavior of the other PAH.
Mackay, Paterson and Schroeder (1986) show how the detailed
8-21
processes of air-water interchange of contaminants can be incorporated
into a fugacity-based model, to account for a physically realistic detailed
partitioning to airborne particles and precipitation, for example. Mackay
et al. (1986) address substances with properties similar to PAH. Mackay
and Paterson (1988) include these detailed processes, as well as
intermedia exchanges involving soil and sediment, as shown in Figure 8-
1. This figure is explained below.
In summary, equilibrium and dynamic modelling exercises which have
been carried out for B[a]P for this study and elsewhere (as described
above) consistently predict that this compound resides ultimately in soils
and (buried) sediments. B[a]P is intermediate in properties which
determine environmental fate in the PAH family, and may, therefore, be
considered typical. In the latter medium, this typical PAH, presumably,
undergoes mineralization or biotransformation over a very long period of
time. The atmosphere and aquatic media, then, serve as vectors for
transport of PAH to soils and sediments and do not act as cumulative
resen/oirs for PAH.
Figure 8-1, which is based on the work by Mackay and Paterson (1988)
referred to above, shows a calculated mass balance for B[a]P in the
Great Lakes Basin (southern Ontario) from advanced Level III fugacity
modelling. The key to the figure indicates the definitions of the processes
and quantities for "emissions" (includes discharges to air, water and soil),
amounts advected, transferred to various media or ultimate residences
(sinks), and amounts degraded by reaction. The report by Mackay and
Paterson (1988 and W. Stiver, private communication, 1988) should be
consulted for detailed explanation.
8-22
FIGURE 8-1
Level III Fugacity Model
Results for B[a]P in Ontario
EMISSIONS ait
lOO
soil
•water
O-oZ-t
'5"o
h
^.9-
0043 ,^ 27
'•--..
O. 13
l-Sx.o—
\Z* lO'"
AIR
..-•"
A^oo
39.
^^@/^..
w |q\ \I0 \Oois
0.O88\ \ \ \®
'■••-...
IZ
2Z^io"*
SOIL
.»■•''
2 t«IO^
91
»-3
O-U
"^" --.^ 4.9
4.^x.o-«|^ATER
11 X lO"''
'■--,,
PROCESS Key
CD diffusion
(2) wet part.dep.
(3) dry part. dep.
® sed. resuspenaion
®sed. deposition
(j)to higher altitude
(g)to groundT«iter
(9) sed. buhal
—^►transfer ^ uo i-rs;
. .^,reqction Vwot/k
fc^odveclJon J
t-3
2.7
(S.
Flow and R^n time-
2 Cy
Box Key
Perststence= 4.3
L
f=fugacity(Pa)
c=conc.(nr\ol/rv\j)
m=amounl(mol)
%=%age of total arr*!^
■-...... %
' f
" ■■ -.J
c
..■■■'" \r\
Source: Mackay and Paterson, 1988.
8-23
The figure is used here to indicate that accounting for all of the major
environmental processes produces a picture of B[a]P transformation and
fate - ultimate movement toward soil and sediment - that is consistent with
other evidence presented.
Some additional explanation may help to clarify the information presented
in Figure 8-1. The key to the data boxes (calculated results) for each
compartment (air, soil, water and sediment) appears in the lower right
hand corner. The key to the process arrows is in the lower left hand
corner. The circled numbers written next to each process arrow identify
the process (see Key). The other numbers written next to each process
arrow are the calculated rates (mol/h) of chemical removal from the
compartment by the indicated process.
Note that chemical reaction rates of removal from both air and sediment
are taken to be zero in this model. This assumption implies that the rates
of other removal processes are much more significant than chemical
reaction in these compartments. Based on current knowledge, the
assumption seems reasonable. With reference to the atmospheric
compartment, zero chemical raction rate means that the PAH emitted in
Ontario are normally transported (advected) out of Ontario before
significant chemical transformation occurs.
To summarize, Figure 8-1 shows that for the 100 mol/h of B[a]P emitted
to the air in Ontario, approximately 70 mol/h (70%) are deposited to water
and soil in Ontario and 30 mol/h (30%) remain in the air to be added to
the "background" influx of 10 mol/h, producing a net flux out of the
Province of about 40 mol/h. The out flux, of course, impinges on
downwind areas. The results shown in Figure 8-1 imply that Ontario
emissions (which include discharges) produce an approximately four-
8-24
fold increase in airborne flux of B[a]P (and by extrapolation, of PAH) and
an approximately three-fold increase in airborne concentration (1 ng
m'^to 3ng m'^for B[a]P). The latter value is high compared with current
monitoring data.
The atmosphere, thus, does not appear to be a reservoir for B[a]P, and
by analogy, for other PAH with similar physical and chemical properties.
The 'emission' quantities used by Mackay and Paterson (1988) are not
entirely consistent with quantities estimated in the present work; for
example, the air emission corresponds to 200 MT/y B [a] P compared with
Concord's estimate of 250 MT/y for total PAH from major sources.
These numbers are considered to be comparable, within the current
quality of data. The Mackay and Paterson (1988) estimate of 200 MT/y
B [a] P emissions to air predict ambient air concentrations which are higher
than observed; therefore, a better estimate would probably be a factor
of at least three lower. That is, an estimated emission of 50 MT (B[a]P)/y
would bring the predictions of the Level III model more in line with
observation. Such a number would be more consistent with the
estimated total PAH emission for Ontario shown in Table 3-31. The soil
emission rate shown in Figure 8-1 is estimated to account for spillage of
fuels and lubricants, which contain PAH both as used and as a result of
aging proceses in use. The supporting data for the Level III calculation
are provided in detail in Appendix A.
Thus, the information presented in previous chapters, implies that, within
present uncertainties of physical and chemical properties and
environmental concentrations, it is wise to focus on PAH which are
predominantly associated with airborne particulate matter and soils or
sediments for the purposes of evaluating and tracing source impacts.
Within that subset of PAH, those with the lowest vapour pressure (highest
8-25
affinity for airborne particulate matter) and least chemical reactivity are
likely to be the most unambiguously traceable in the environment.
Toxicologically, the higher molecular weight PAH predominating in
particulate matter include most of the demonstrated or suspect
carcinogens (see Table 2-2), and particle-bound PAH will have greater
absorptivity in the human respiratory system, the hydrophobic vapour-
phase PAH having less affinity for epithelial tissue (and, therefore, less
tendency to be available for absorption).
For the above reasons, the estimates of emission rates and atmospheric
concentrations which are described in the following sections focussed on
particulate emissions and particle-bound PAH.
8.5 Emission Rate Estimates
Province-wide PAH emissions were estimated based on the Ministry of
the Environment's current emission inventory (1984, 1985 data) of
particulate matter from combustion processes and other PAH sources,
by applying best available estimates of PAH emission factors for a
number of individual PAH.
Estimates of background and source-influenced ambient PAH
concentrations in Ontario will assist in the establishment of regulatory
approaches for PAH. In the context of this report, background levels will
distinguish between the 'clean air' levels and those attributable to long
range mesoscale transport (LRT). Knowledge of the relative magnitudes
of the LRT-influenced and the local source-influenced PAH levels will
provide guidance for selecting regulatory approaches and, if necessary,
control strategies, in order to obtain these estimates, a preliminary emis-
8-26
sions inventory for Ontario has been developed, and the contributions of
PAH sources to ambient TSP levels in a selected area of Ontario
(Hamilton) were estimated. In addition, estimates of the background
levels of PAH as a result of long range transport were made. The
assessment of these estimates in terms of their reliability, the need for
their refinement and their implications for policy regarding PAH in Ontario
are discussed.
8.5.1 Estimation of Ambient PAH Concentration for Hamilton
Rough dispersion model estimates of the annual mean ambient PAH
concentrations in the Hamilton area were made. The dispersion model
also provided estimates of the proportions of the total TSP concentration
contributed by coke ovens as well as all steel operations. The model
estimates for the contributions of the steel operation sources were
compared with similar estimates obtained by dispersion and receptor
modelling based on 1979 emissions data (ORF et al., 1982). Estimates
of the particulate PAH concentration in Hamilton were derived by
assuming PAH emission factors for certain compounds. These data
provide the basis for estimating ambient PAH concentrations.
The Industrial Source Complex Long Term (ISCLT) model was used.
Model inputs included emissions from point and area sources in the
Hamilton area, and meteorological data from Toronto Airport were used
as model inputs. The point and area sources used were the same as
those in the Hamilton road dust study (ORF et al., 1982). Two sets of
emission rates from these sources were considered: 1979 particulate
emission rates which are identical to those used in the road dust study,
and 1984 particulate emission rates based on the most recent TSP
8-27
emissions inventory. For the 1984 data, since detailed coke oven
particulate data were not available, the following assumptions were made:
o the total emissions from sources in the area modelled were set to
2400 tonnes (in the 1984 El, the grid total for particulate emissions
is 2477 MT);
0 since some coke oven and other major point sources in the grid had
been reduced between 1 974 and 1 979, those sources which had not
been reduced between 1974 and 1979 were arbitrarily reduced in
order to agree with the total emissions as given in the 1984 El: the
latter were 2400 T, with 1000 T from coke ovens and 1400 T from
other sources;
0 the line sources used in the Hamilton road dust study were replaced
with a traffic area source whose strength was based on the 1984 El
data for gasoline and diesel sources in the grid. The emissions in
the model domain were prorated by area according to the following
relationship: traffic emissions in model domain = grid area traffic
source x (model domain area / grid area).
Estimates of the total TSP concentration for all sources, as well as the
separate TSP concentration attributable to the coke ovens, traffic and the
remaining sources were calculated. The estimates of TSP from the coke
oven and automobile sources together with emission factors of PAH
compounds from these sources will allow the estimation of ambient PAH
concentrations. The source emissions data are presented in Table 8-8.
8-28
TABLE 8-8
Point Source Emission Inventory
for Hamilton
Source
Description
TSP (MT/yr)
1974 1979
1984
1
Canron Ltd.
119.0
108
50
2
Steico Steam Plant
92.9
20
6
3
Steico Blast Furance
637.4
496
150
4
Steico Coke Quench
145.7
132
70*
5
Steico Sinter Plant
1372.2
1131
375
6
Steico Coke Ovens
930.6
803
450*
7
Steico Coal Handling
445.1
418
92
8
Steico B.O.F.
450.3
495
109
9
Steico Open Hearth
803.3
568
150
10
Proctor & Gamble
234.9
351
150
11
Dofasco B.O.F.
1125.8
459
150
12
Dofasco Coke Oven
659.9
557
300*
13
Dofasco Blast Furnace
448.3
368
100
14
Dofasco Coke Oven
560.8
181
180*
15
National Steele Can.
89.6
81
18
16
Interflow Systems
181.4
181
50
Total
8297.2
6369
2400
* Coke total 1000 MT
Grid # 295
8-29
Ambient annual TSP concentrations were estimated at four receptor
locations (the same as those used in the Hamilton road dust study). The
model predictions for the total annual mean TSP and the portion of the
annual mean TSP attributed to the coke ovens, traffic and other sources
are listed in Table 8-9.
The model predictions of the annual mean TSP concentration may be
compared with observation at monitoring sites in Hamilton during similar
years. The 1979 annual mean TSP concentrations at the modelled
receptor sites were in the range 70 to 93 ug m'^.
The model estimates determined in this study did not include any
contribution from background sources. This background level would be
equivalent to TSP measurements made at remote Ontario sites.
The Hamilton study included a background site and TSP measurements,
and estimates for the background site were of the order of 40 ug/m^
(ORF et. al. 1982). The 1984 prediction of the annual mean can be
compared only with the MOE monitoring data for Barton/Sanford (MOE
station 29025) for which the annual mean in 1984 was 81 ug/m^ The
model predictions (42 to 48 ug/m^) plus an assumed level of 40 ug/m^
for the background, produce agreement with the observations to within
25%. Such agreement is as good as can be achieved for these types of
data and the inherent uncertainties of modelling area sources.
8-30
TABLE 8-9
Summary of Modelling Results
ISCLT Model Predictions for 1979 and 1984
For the Hamilton Area
Annual Mean TSP concentrations
(ug/m^)
Coke
Traffic
Other
Total
Observed
Year
Receptor
Ovens
Diesel
Gasoline
Sources
f)
1979
E
2.0
0.4
2.6
40
45
65-81
J
2.1
0.5
3.3
36
42
93
C
2.7
0.5
3.3
38
44
70-84
0
3.0
0.4
3.2
41
48
89
1984
E
1.13
0.36
2.59
38
41
J
1.19
0.47
3.34
36
38
C
1.52
0.46
3.28
33
44
0
1.67
0.44
3.16
39
44
A
ORF,
1982. Ra
inge of observed values at the different sites are
annually adji
usted averages.
8-31
The percentage contribution of the coke ovens to TSP levels is estimated
at between 3 and 4% by mass based on the current model runs (Table
8-10). This percentage may be compared with receptor model and
dispersion model predictions for the contributions of the iron and steel
sources in Hamilton from ORF et al., 1982. These contributions were
estimated at 8 and 6%, respectively, but include all point sources and not
just the coke oven sources. The agreement is reasonable given the
assumptions made.
8.6 Estimate of the LRT contribution to PAH levels in Ontario
The dispersion modelling or receptor modelling estimates provide
preliminary estimates of the ambient levels of PAH due to local sources.
Previous receptor model estimates of the background TSP levels are
reasonable, and may be taken as an upper limit for the contribution from
distant, i.e., LRT sources.
In order to estimate the PAH concentration contributed by LRT sources,
the constancy of the PAH/TSP ratio for each source will be assumed to
apply to the background measurements. However, it must be noted that
chemical transformation of PAH may render the use of the same
PAH/TSP ratio in sources as in ambient measurements to be strictly
inappropriate. For example, the B[a]P/TSP ratios near PAH industrial
and urban sources were between 5 and 66 ug/g, while at remote sites,
the ratio was 0.3 to 0.6 ug/g (Gibson, 1986). Analogous ratios for 1-
nitropyrene were lower at urban/industrial sites (0.2 to 0.6 ug/g) than at
the remote sites (0.5) (Gitson, 1986).
8-32
TABLE 8-10
Percentage Contribution to TSP from
Coke Ovens, Traffic and Other Sources in Hamilton
Coke
Traffic
Other
Year
Receptor
Ovens
Diesel
Gasoline
Sources
1979
E
4.3
0.8
5.8
89.1
J
4.9
1.1
8.0
86.0
0
6.0
1.0
7.4
85.6
0
6.2
0.9
6.6
86.2
1984
E
2.8
0.9
6.3
92.7
J
3.1
1.2
8.7
94.1
C
3.4
1.0
7.4
74.5
0
3.8
1.0
7.2
88.6
8-33
If it is assumed that the PAH/TSP ratio, in particular, the B[a]P/TSP and
the total PAH/TSP ratios for Ontario emissions of 2 x 10'^ and 0.015
respectively, are similar to the ratios in LRT sources, then the B[a]P and
total PAH concentrations in background TSP would be of the order of 0.8
and 600 ng/m^' respectively, assuming a background TSP level of 40
ug/m^. The B[a]P estimates do not take into account the degradation
of B[a]P during transport. The B[a]P/TSP ratio can be up to 240 times
greater at near-source sites than at remote sites (Gibson 1986). A more
appropriate background B[a]P level applicable to remote Ontario
locations would therefore be about 4 pg m'^ assuming a factor of 200 for
the B[a]P/TSP ratio near sources relative to remote sites.
Background measurements of PAH in Ontario are not available since
most PAH sampling has been done in urban locations. The most
representative urban sites at which ambient PAH data are available are
for Sudbury and Niagara-on-the-lake where B[a]P levels were 0.1 to 0.4
ng m'^ (Katz et a!., 1978) and 0.4 ng m'^ (Hoff and Chan, 1987),
respectively. Measurements made at a remote site in Bermuda (Gibson,
1986; Gibson and Wolff, 1985) may be used to provide some indication
of the impact of LRT from U.S. sources on Ontario. The distance
between U.S. emission sources and Ontario is similar to the distance
between the east coast of the U.S. and Bermuda, but the absence of
additional sources in the ocean trajectory undoubtedly alters the
chemistry and concentrations relative to the two receptor areas (i.e.,
Bermuda and Ontario). The nitro-PAH levels in Bermuda were in fact
higher than B[a]P when there was LRT from U.S. sources. This reflects
the atmospheric transformation of B [a] P, the formation of nitroPAH during
transport and possibly the stability of the nitroPAH, although they are
somewhat reactive photochemically. B[a]P and 1-nitropyrene
concentrations at the remote Bermuda site were, respectively, about 5
8-34
and 10 pg/m^ with TSP levels of about 20 ug/m^ (Gibson, 1986). The
B[a]P measurements are similar to those estimated above. If similar levels
apply to air masses entering Ontario, then more emphasis must be
placed on establishing the concentrations of the stable PAH, e.g., B[e]P,
as well as the secondary PAH compounds, namely nitro- and oxy- PAH
formed during transport.
Air masses entering Ontario may be considered as either "background"
or "polluted". The former would include air masses such as those
crossing the western border with Manitoba and those segments of the
southern border with the U.S. that include Lakes Superior, Huron, and
Erie. For purposes of estimating the flux into Ontario, it is convenient to
consider a straight line strictly from Buffalo to Thunder Bay across which
air masses move. That segment of the line from Buffalo to the shore of
Lake Huron (300 km) may be considered to bear higher B[a]P levels
(approximately 0.8 ng/m^) in air masses which would enter southern
Ontario. The remaining segment from the shore of Lake Huron to
Thunder Bay (approximately 850 km) would bear background B[a]P
levels (4 pg/m\ The fluxes crossing these segments are calculated
from the following equation (see Galloway and Whelpdale, 1980):
F = cu H Lf
where c is the mean PAH concentration, u the mean annual wind speed
(7 ms'^), H is the mean annual mixing height (900 m) and L is the length
of the boundary bearing the flux and f is the fraction of the time winds
blown across the boundary (assumed to be 1 .0). The annual amount of
B[a]P entering southern Ontario is estimated at 1.5 g/s or 47.7 tonne/y.
Estimates for the remainder of the boundary are 0.68 tonnes/year.
8-35
The estimated B[a]P flux entering southern Ontario is significant,
corresponding to approximately 50% of the total Ontario B[a] P emissions.
A more realistic estimate of this flux will consider the frequency with which
winds blow across the 300 km segment. Including this frequency
(approximately 60% of the time) will reduce the B[a]P entering Ontario to
approximately 29 tonnes/year - an amount which is still significant,
relative to the total B[a]P emissions in Ontario. That is, for southern
Ontario, imported PAH appear to dominate emissions within the Povince;
whereas, for northern Ontario, the opposite obtains.
Amount entering southern Ontario:
0.8x7.0x900x300x10^
10^ X 10^ X 10^
x (60 X 60 X 24 X 365) = 47.7 tonnes
Amount entering the remainder of Ontario.
0.004 X 7 X 850 X 10^ x 900 x 60 x 60 x 24 x 365
10^x 10^ = 0.67 tonnes
8.7 Regional Estimates for PAH Emissions in Ontario
A regional PAH emissions inventory was developed for Ontario by using
the total suspended particulate (TSP) emissions inventory developed by
the MOE which includes the six major types of PAH sources in Ontario.
These include the following point sources:
8-36
o forest fires;
o gasoline fuelled vehicles;
o diesel fuelled vehicles;
o residential wood fireplaces and stoves;
0 coke ovens; and
0 coal fired thermal generating stations.
The PAH composition for each of the source types includes 22 of the 47
PAH compounds that were considered in this report. Selection was
contingent on a reported PAH emission factor for at least 3 of the 6
sources. A collective estimate was used for each group of isomers and
were identified under generic names. For example, under the generic
name of benzofluoranthenes would be included the benzo[b] and
benzo[k]fluoranthene isomers. Emission factors, F^p, for the 22
compounds are presented in Table 8-14. The total PAH emission in each
grid was taken as the sum of the emissions of the twenty-two compounds
for all six sources (equation 8.2).
8.2
Also included in the emission inventory are the estimated contributions
for sources discharging to water and to soil in southwestern Ontario.
These data are provided as Maps labelled B1-1 to B1-4 in Appendix B of
this report.
n = 6, p = 22
Eg =
/ . npg
n = 1,p = 1
9-1
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9-48
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9-49
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9-50
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APPENDIX A
Level 3 Fugacity Model Data
iitf? 03-22-'^S5
h«e; ^■>'':02!
lE'^EL 3 ?U5AC!TV fiQDEL - P)[a]P
'its O' BsniO^'i)?-''?'"!; 1' '•■Giiv'c''" 0"UriC
r.ciecijlar weight
jrMSQijs ;ol'jbl i ! t •
•■ipour presiure
herrv s const j't
'c'.inol -wstBr part ccBtf 'logi
teliperat'ir;
252.30 g/tol
3.3CO0E-''T ^■«>3 or l.iOSSE-'j joi »i3
;'.3(.00£-07 pa or ".2(-4:£-12 it» :'■
4.3*43E-02 ?3 *3'j":i
25.0 deq C or 2=3.2 f
buif cciapsrt'^e' ■.
! buii- 3ir
2 bulk xater
3 bulk roil
4 buU' iedi.nent
•■oluae
•iEight.'
jroa
?
dBn=;t :
'!3
depth («)
a2
soi.'«3.P3
kg/?3
J..:.0:>Etl4
2.yC0E+03
2.000£+n
A.i72E-02
1.1*
4.000E+12
5.000E+0!
3.00'i£*-10
4.101E«-(n
iOOO.O!
1.200£-!y
l.wOOE-Ol
1.200E+11
1.541E+05
15^X1.24
3,0i:ii)E+03
l.OOOE-02
S.OOOEMO
2.332E*n5
K20.00
total area 1*2) 2.000E+1:
=i]l)CDspart«ent
1
1
air
1,
air particles
1
T
^
Xitcr
2,
■^
water particles
1
4
biota
-■ 1
5cil air
•^
1
501 ! "ater
^
7^
•;o;I solids
4
1
pore »;a-er
• Ql'jae
(ii3
^
iiol/«3.Pa
densit)
tg/.3
4.M00E+1'
3.00')E+fi3
1.034E-01
3.31iE+0'5
2400.0(1
4.000E+12
2,»0£*iv
4.000E'06
2.063E+01
3.387Ef0i
=!.4!.gE+05
1000.00
2400. 0>^
1000.00
2,400E'-0^
3.ii")E+v9
■i.Oi)OE'-'''°
4.034E-()4
2.0o;E+01
3.397EHI5
1.1''
1000.00
2400.00
5,i0vE-03
2,i-,0Et-03
2.0b3E+0i
'.77AE+05
1000.00
2400.00
density sass fraction
.Tte
organic content •'■action
ri.jn
1.
OOE+00
i.1
OOE-II
!,
, OOE+OO
c
, OOE-06
1,
, OOE-06
0
,00£-01
-
,OnE-01
c
,00£-01
.OOE-O!
Reaction finititirs
buU ccspi'tientE rats constant hslt-lifp D .alij?
h-1 1 jol'pa.'^
i GuU air O.CvOOE +(■;■' 'i.Ovi-'^E-vv O.^^OOOEtOO
: bdt -ate' 3.300'>E-05 1.^300E^04 5.741lE*-v'
3 bulk 501 ! 3.^'X!i'iE-05 l.'Sf^OE+OS ?.ii32E+10
J tuU' souiseot 0.00(-'-)E*')(i (•'.■■}00i)E*-0(i 0,OOOOE»-00
iubcGJip artier- '.a
;, 1 dir O.C-iXiOE+00 O.OOOOE+O? i).OCOOE+00
1, 3 ai-- part::i5a 0,?00OE*0O 0.0i)0OE+0O O.OOOOE^OO
2, 2 ^ater C;.)0OOE+O0 '■;.0;-)OE+00 0.0000E+'')0
2, 3 "at?" particles 'O.OOOOE^-OO O.O'iOOE+OO O.OOOOE'-OO
2, 4 biota O.OOOOE+OO O.)00OE+'>0 0,COCOE+00
3, 1 soil air O.OOOOE'OO 0,OOOOE*-00 0.00:)0£+00
3, 2 501 1 water O.OOOOE+vO O.-H-vvtn'y O.OOOOE+OO
3, 3 5oi! solica O.OOOOE>00 O.OOOOE+OO O.OOOOEKiO
4, 2 pore »atBr O.OOOOE+OO O.OOOOE+OO O.OOOOE+OO
1, 3 sea. aoiids O.OOOOE+OO O.OOOOE+OO O.OOOOE+OO
Advective Paraueters
oapartient
i 1 n«
inMow concn
rate constant
3 value
resiae-ce ti«e
•3/h
tol;i3
h-1
■ol/pa.h
h
1 bulk air
3.30E+12
3.00E-12
3.25E-03
2.20E+11
1.21E+02
2 buU ^ater
3.3!"iE+0g
O.OOE+00
9.25E-05
1.35E+10
1.2iEt04
3 bulk soil
O.OOE+00
O.OOE+00
O.OOE+00
O.OOE+00
infinity
4 bulk aednent
O.OOE+00
O.OOE+00
0,COE+00
O.OOE+00
inHnity
Transfer to higher altitude, aediaent burial and leaching froi soil to groundwater
process velocity velocity flow rate constant D value residence tues
t'ansfef to higher ait 5.00E+01
leaching 'roi aoil
sediieot burial
velocity
velocity
«;h
flow
(j3/h
rate const
h-1
ant
D value
sol .'pa. h
re
h
5.00E+01
3.40E-O1
3.00e-04
1.03E-02
3.38E-)S
3.42E-09
2.05E+O'
4.46E+0b
2.74E+03
5.!4E-06
1.14E-05
l,3"E+03
'.tlE+07
2.13E+n?
! . ?;e+05
5. ^5=^-4
^r^r.z'^r -.;r^*.ci Dr 5 UB':^^i(' CQW^'^' t 'sS-'i* :
CC'iTiDirtUSnt
tili'I.ji
^TQ'l!
}:Efl! Z.33!E+'M !.;5rE^02
'Off
-d,25<?E'--':|
,J"5Etli) '.ivvE^v-. ',:2?iE':3
, i^^E*--^ -^.oOvE '••■•') i,977E+03
1. i-'i-CfvO
*rns
fr-,,1
trr^^ 3 ^■1
:i'J8T, ^l0'i<5 ind '-'ei -DC 1 1 1 =
D
dry deposit iGD
Gl-f'J510n
o;-tu=:on
■"st isposition
jr. jcnnjifinn
ion
(anl/htra! (a3/h)
''.ii'E+v'
l.i02E+08 7.^63E+('6
!-^30c+il '.l''!E-^:'-!
;.":OEMO !.72SEtOi
9,o.33E+i)7
!.15iE+i;7
2.402E+i)3 l.!?iE'-v7
l.^UEfl! 4.i53E.in
Ll^iE+O
ai^^'JDion
7.527Et:
'5
:--it^on
! . 42-)E-'
V
r.; ^*u5i0n
-,32-Et
■8
^= = ..:r.Rr:=;r.r.
T -. - .-.C^
;n
.iO^E *•!)'?
,592E+0!
' p^Ff-;:'
n . n f- *■ ; .-.
:i3/y
^llFrv
1,;-20E+!; ''.7()3E-05
4.08(iE+05 3.3S1E-10
2.2"lE-05 !.0e(i£+01
■■5!0C1'./
o:e-)5
5
c -
lIP-;
.1
31E-1;:
J
1;
;':? -
A
80E+0!
-;
.',:.
0E-:
.T
U
3. jO?E-iM
. t -*v",'u '.'0
,'';n;P-:".J
Dii rioK 1.045E+0'' 2.74;)Eti:'3
3..'.C'E-v- l.M2E-:3 !.."'^'E-^'^4
4.03^':E-;- 3,531E-!i5 3.4.:OE-01
2.100E+':!" 2.233E-'-3 2.0:)?E-04
■.■i'ii: •i'.zr-
Air-So;;
^;t5'--:?dj«ent
•.s h:
.■•.'■-••.if'^
■*a -r- r 1 J t
;, .;-.".it-i.-^
;«:;i
sadU'eiit
•'i!u55 'Of Ditfusi-% Fio" !'• Soil Air and *'5tir
.i^^QU- v.vPOy «
irrqth 0.0050 n
loil 31 r diffysion
".2j6E*05
1.438E*07
Bulk Coiipsrt^ent;
CDmpaTtiiiEni
alSO'jrit
percent
copcentriticos
9Qi
tQ\'»Z
iiccg-^
! tfuli i\r
!."25E»03
" . 1 Z'\'
l.lBlE-ll
2.514E-%
2 bulk na^.er
l."3Et05
4.925
4.470E-09
M'SE-;!
: buii =oii
2.iO5E^06
'l."4
2.17l£-v4
- .e;r_..*
i bulk 55"; ;*'=■■' f
S.410E+05
23.i:o
1.^51£-j:
;.2-i;E-"i
Total
i.i-jOE^Ot
100. ODO
Subcospsft/sents
coapart*5r.i
anount
pprrcnt
c-r-centraticr;
1 iir
3 air particles
2 water
3 aate' pa''ti:ie;
4 bicia
1 ;cii air
2 =01 1 nater
3 soil solids
2 po'"B later
3 sed. solids
id
2.857E+^M
H.;i;4E+04
3.472Et04
4.123E+03
1.081E-03
3.297E+01
2.605E+0.i
5.208E+01
8.410Et05
9o! 'iZ
0
001
0
129
1
J7B
334
;)
114
0
000
0
002
eicroq'q
I.?80E-;':3
1. l23E'-0t
2.45:E-v5
«i:ro!;.- s:
0. Oi.il
23.168
M42E-14
1.520E-06
1.3!:-2E-05
5,370E-01
C.171E+0I
i.isiE^oa
2.248E-08
3.u"E-:i6
5.673E->-00
4.23oE-03
4.433E-0!
1.0i.9E*0:
1.031E-03
2.60CE-:m
2.600E+05
4.506E-13
'.S'lE-OB
I.;37E-'!4
2.305E-08
5.B15E-0a
5.81jE+00
4.342E-:'4
4.541E-02
LO'^Eni^
''.299E-08
2.346E-05
2.346E+01
3.504E-03
3.C.34E-01
B.a-lOE^O?
p -
,"vE-10
, 0'0£-v'
,ll7£-0<;
."07E-0^
f-jgacitr
l."OE-
!0
1.770E-
10
1.0?0E-
09
l.O^OE-
c
l.O^OE-
09
i.irE-
".'■*
i.n7E-
09
:,ii7E-o9
4,507E-::''
4.507E-0''
Si^iiisary n+ 4 bult cospartsent uass balances f<»ol/hi
eaissions inMot<
reaction
outflow
net fluS Out
to otbe'' coipts
bulk air
bulk water
bulk soil
1.000£*02
l.OOOE+00
5.000E+01
O.OOOE+OC
9.900E+00
O.OOOE+00
O.OOOE+OO
0. 000 £■''00
O.OOOE+00
i.253E+00
9.11S£H!l
O.OOOE+00
S.S'SE+Ol
i.475E+01
O.OOOE+OO
0...vOE+00
7.090E+01
-2.001E+01
-i.l29E+01
-^.iOOEn.^!
Tot 2I
1.5I0E+02
'.'OOE+OO
= ,'HE+OL
5.373Ei-01
total irrput ■eaissions i'4 in + low)
total output 'reactions and outfloul
residerce time ihoursi 22558. '7
ida-si 93'?.c)57i
l.ii-^E+02 uol/h
l.iO^E+02 «ol/h
persistence 3"250.86
persistence 1552. ll'
^'■i'i^er jnd Trinsforsation "-ates :9ol''hl
9ii;5ion5
Jd'^fctive outflow
'■5 act ion
transfer to higher iititude
leaching fro* son
sednent burial
t''an5fer to air f''os
transfer to water fron
by diffusion air-water
by diffusion nate'^-air
net di'^usion
by rain
by wet deposition
by dry deposition
by water runoff
by soil ''unoff
transfer to soil i-'aa
by diffusion air-soil
by diffusion soil-air
net diffusion
by rain
by wet deposition
by dry deposition
transfer to sednent fi-oi
by diffusion water-sedisent
by diffusion seduent -water
net diffusion
bv :ed;sEMt depasition
h; 59a;ie'' '9=u;pBns;cn
Bul'r ai"-
l.00OE*-O2
;,'?0OE+0O
3.89gE+ul
0,OOOE*-00
2.a:7E-v2
O.OOOE+00
2.331E+01
1.7I1E-02
-1.05;E-01
-9.820E-02
2.835E-02
1.823E*01
1.014E+01
4.259E+0!
2.038E-03
-1.286E-02
-1.082E-02
4.253E-02
2.734E+01
1.522E+01
O.OOOE+00
Buik water
l.OOOE+00
0.000E*-00
l.rSE+Ol
-2.831E+01
O.OOOE+OO
-1.2<''E'00
''.600E+0f^
3.S30E-')1
-3.528E*O0
-2.a75E+00
l.!47E+01
Soil
5.000E+01
v.OOOEfOO
0.i)OOE+00
^.USE^O!
;.073E-0l
-4.25^Et01
1.257E+00
1.073E-01
1.190E+00
O.OOOE+00
O.OOOE+00
Sediment
O.OOOE+00
O.0OOE+';O
(■.OCO£»00
O.OOOE+00
:-tai
''.tOOE+OO
O.OOOE+00
-''.400E+00
O.OOOE+00
O.OOOE+OO
''.744E+
2.427E-
I.073E-
'.dOOE+i
-7.090E+<
2.001E+'
1.12=;
9.e00E+C
APPENDIX B
Maps Illustrating PAH RegionaJ Estimates
Concord Scientific Corporation
B1-1
1.0 PROCEDURE FOR ESTIMATING REGIONAL PAH EMISSIONS
PAH emissions were estimated as follows:
a) Only air emissions were considered from the following sources:
0 gasoline and diesel-powered mobile sources;
o residential wood burning stoves and fireplaces; and
o forest fires.
b) The Air Resources Branch (MOE) total suspended particulate (TSP)
emissions inventory for each of the 3 sources was used to provide
a reference database (MOE Emissions Inventory, 1985).
c) The TSP/PAH ratio was assumed to be approximately constant for
each of the three sources.
A summary of the TSP/PAH ratios used for these estimates is presented
in Table 81 -1. The data correspond to values in Tables 3-23, 3-26 and
3-30. ••
1 .1 Rationale for Source Selection
The three sources used in this analysis contribute approximately 97% of
the total annual PAH emissions in Ontario. Thus, although coke oven and
coal fired power generating station emissions, for example, will add to
certain localized areas, particularly in southwestern Ontario, these latter
sources are excluded from the analysis. In support of this approach is
the fact that sources discharging to water and soil will also contribute to
localized areas in a similar manner. Therefore, as a preliminary analysis,
Concord Scientific CorfX)ration
B1-2
TABLE BM
PAH, TSP Emissions and PAH/TSP Ratios
for Forest Rres, Mobile and
ResidentiaJ Wood Heating Sources
Unleaded
Gasoline
3 way &
oxid. cat.*
(mg/kL)
Diesel
Ught &
Heavy Duty**
(mg/kL)
Wood
Burning
Stoves***
(mg/GJ)
Forest
Fires****
(mg/mT)
PAH
3,113
20,900
TSP
823,529
50,525,000
PAH/TSP
0.0023
0.0069
0.0038
0.0004
* From Table 3-23. (Assume ratio for 3 way catalyst is also applicable for
oxidation catalyst.)
** From Table 3-23. (Assume ratio for light duty is the same as heavy duty
diesel.)
*** From Table 3-26. Total PAH taken as the sum of entries for emissions factors
for wood in the Table. TSP data from U.S. EPA PB83-250720, pp. 1.9-3 to
1.10-5.
**** From Table 3-30. Average values quoted.
Concord Scientific C()r[_x)rjition
B1-3
we have assumed that an assessment of the regional concentrations to
the total PAH inventory from forest fires, mobile and residential
woodburning stoves and fireplaces will provide an acceptable database:
the alternative would be a detailed evaluation for all major sources, which
could not be undertaken at this time.
1.2 Illustrative Regional Maps of PAH Emissions
The estimated PAH emissions and densities are presented in Table B1-
2 and illustrated in Figures B1-1 to 81 -4. The code used for identification
of levels is presented in Table 81 -3.
Concord Scientitic Corporation
B1-4
TABLE Bl-2
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSION
GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y
GRID
UTM
AREA
GRID
(Ha)
10000
1
10000
2
10000
3
10000
4
10000
5
10000
6
10000
7
10000
8
10000
9
10000
10
10000
11
10000
12
10000
13
10000
14
10000
15
10000
16
10000
17
10000
18
10000
19
10000
20
0.
0012
0.
0060
0.
0000
0.
01
0.
72;
0.
0424
0.
0242
0.
0000
0.
07
6.
66^
0.
0168
0.
0249
0.
0000
0.
04
4.
18
0
0495
0.
0412
0
0000
0.
09
9
07
0
0461
0
0506
0
0000
0
10
9
67
0
0012
0
0034
0
0000
0
00
0
46
0.
1396
0
1534
0
0000
0
29
29
3o!
0
1318
0
0835
0
0000
0
22
21
53
0
0297
0
0548
0
0000
0
08
8
45
0
"0649
0
1032
0
0000
0
17
16
81
0
1806
0
1931
0
0000
0
37
37
37
0
0502
0
0442
0
0000
0
09
9
44
0
.0223
0
0775
0
0000
0
10
9
98
0
.0104
0
0370
0
0000
0
05
4
74
0
.1175
0
.1327
0
0000
0
25
25
01
0
.0111
0
.0427
0
.0000
0
05
5
38
0
.0104
0
.0389
0
.0000
0
.05
4
.93
0
.0099
0
.0317
0
.0000
0
.04
4
.17
0
.0060
0
.0147
0
.0000
0
.02
2
.07
0
.0041
0
.0094
0
.0000
0
.01
1
.36
Concord Scientitk Cor[X)rcition
B1-5
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Elmission Density
from Three Major Sources to the Atmosphere in Ontario
UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSION
GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y)
GRID
UTM
AREA
GRID
(Ha)
2500
21
2500
22
2500
23
2500
24
2500
25
2500
26
2500
27
2500
28
2500
29
2500
30
2500
31
2500
32
2500
33
2500
34
2500
35
2500
36
2500
37
2500
38
2500
39
2500
40
0.
0074
0.
0249
0.
0000
0.03
12.!
33
0
.0889
0
. 1342
0,
.0000
0.22
89
.21
0
.0453
0
.0491
0
.0000
0.09
37
.79
0
.0366
0
.0087
0
.0000
0.05
18
. 11
0
.0276
0
.0144
0
.0000
0.04
16
.79
0
.0283
0,
.0155
0,
.0000
0.04
17
.52
0
.0272
0,
,0110
0.
,0000
0.04
15
.25
0
.0272
0,
,0106
0,
.0000
0.04
15
. 10
0,
.0361
0,
.0076
0,
.0000
0.04
17,
,48
0.
,-04 6 3
0.
,0057
0,
.0000
0.05
20,
.77
0,
.0124
0,
.0079
0.
,0000
0.02
8,
,15
0,
.0790
0.
.0654
0,
,0000
0.14
57.
,76
0.
,0014
0.
.0030
0.
.0000
0.00
1.
,76
0.
.0840
0,
, 1175
0.
,0000
0.20
80.
,63
0,
.4068
0.
,5371
0.
.0000
0.94
377,
,54
0.
,2385
0.
,2906
0.
.0000
0.53
211.
,65
0,
.0140
0,
,0215
0,
,0000
0.04
14,
,23
0
.0046
0,
.0140
0,
.0000
0.02
7,
,44
0
.0053
0
.0144
0,
.0000
0.02
7.
,86
0
.0847
0,
.0582
0.
,0000
0.14
57,
, 17
Concord Scientit'ic Cor[X)ration
B1-6
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Elmission Density
from Three Major Sources to the Atmosphere in Ontario
GRID UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSION
AREA GRID GRID AND SOURCE (MT/YEAR) PAE DENSITY
(Ha) MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y)
2500 41
2500 42
2500 43
2500 44
2500 45
2500 46
2500 47
2500 48
2500 49
2500 50
2500 51
2500 52
2500 53
10000 54
10000 55
10000 56
10000 57
10000 58
10000 59
10000 60
0
0094
0
0091
0
0000
0
02
7.40
0
0094
0
0091
0
0000
0
02
7.40
0
0035
0
0087
0
0000
0
01
4.86
0
0018
0
0087
0
0000
0
01
4.21
0
0127
0
0064
0
0000
0
.02
7.63
0
0186
0
0030
0
0000
0
02
8.67
0
0325
0
0420
0
0000
0
07
29.75
0
2431
0
3190
0
0000
0
56
224.84
0
4784
0
6096
0
0000
1
09
435.20
0
~L197
0
1187
0
0000
0
24
95.35
0
0253
0
0382
0
0000
0
06
25.40
0
0016
0
0042
0
0000
0
01
2.31
0
0028
0
0060
0
0000
0
01
3.52
0
0601
0
0197
0
0000
0
08
7.97
0
0088
0
0212
0
0000
0
03
2.99
0
0069
0
0174
0
0000
0
02
2.43
0
0923
0
0310
0
0000
0
12
12.33
0
.2711
0
2400
0
0000
0
51
51.10
0
1319
0
0877
0
0000
0
22
21.96
0
.0083
0
0231
0
0000
0
03
3.13
Concord Scientitic Corpor.ition
B1-7
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSION
GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y)
GRID
UTM
AREA
GRID
(Ha)
2500
61
2500
62
2500
63
2500
64
10000
65
10000
66
10000
67
10000
68
10000
69
2500
70
2500
71
2500
72
2500
73
10000
74
10000
75
10000
76
10000
77
10000
78
10000
79
2500
80
0.
.0012
0.
.0023
0.
.0000
0,
.00
1.
.37
0.
.0016
0,
.0045
0.
.0000
0.
.01
2,
.46
0.
.0005
0.
.0015
0,
.0000
0,
.00
0.
.79
0.
.0016
0.
.0045
0.
.0000
0,
.01
2,
.46
0.
.0088
0.
.0223
0,
.0000
0.
.03
3,
. 11
0,
,2368
0.
.2328
0.
,0000
0,
.47
46,
.96
0,
.0219
0.
.0223
0,
.0000
0.
.04
4,
.42
0.
.0576
0.
.0673
0.
.0000
0.
.12
12.
.48
0.
.0028
0,
.0068
0,
.0000
0,
.01
0.
.96
0,
."0005
0,
.0015
0,
.0000
0.
.00
0,
.79
0.
.0016
0,
.0045
0.
.0000
0,
.01
2.
.46
0.
.0014
0,
.0026
0.
.0000
0,
.00
1,
.61
0,
.0016
0.
.0045
0.
,0000
0,
.01
2,
.46
0.
.0083
0.
.0204
0,
.0000
0,
.03
2.
.37
0.
.0088
0,
.0231
0.
.0000
0,
.03
3,
.18
0.
.0417
0.
.0370
0.
.0000
0.
.08
7,
.87
0
.0428
0
.0106
0,
.0000
0
.05
5,
.34
0
.0371
0
.0163
0
.0000
0
.05
5
.33
0
.0041
0
.0076
0
.0000
0
.01
1
. 17
0
.0083
0
.0147
0
.0000
0
.02
9
.21
Concord Scientific Corporation
B1-8
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSION
GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y)
GRID
UTM
AREA
GRID
(Ha)
2500
81
2500
82
10000
83
10000
84
10000
85
10000
86
10000
87
10000
88
10000
89
10000
90
2500
91
2500
92
2500
93
10000
94
10000
95
10000
96
2500
97
2500
98
2500
99
10000
100
0
0005
0
0015
0
0000
0
00
0.79
0
0025
0
0053
0
0000
0
01
3.13
0
0147
0
0155
0
0000
0
03
3.02
0
0382
0
0106
0
0000
0
05
4.88
0
0693
0
0272
0
0000
0
10
9.65
0
0417
0
0299
0
0000
0
07
7.15
0
0018
0
0034
0
0000
0
01
0.52
0
0060
0
0079
0
0000
0
01
1.39
0
0060
0
0076
0
0000
0
01
1.35
0
"0060
0
0076
0
0000
0
01
1.35
0
0012
0
0019
0
0000
0
00
1.22
0
0272
0
0397
0
0000
0
07
26.74
0
1020
0
1417
0
0000
0
24
97.49
0
0083
0
0204
0
0000
0
03
2.87
0
0687
0
0525
0
0000
0
12
12.12
0
0055
0
0129
0
0000
0
02
1.84
0
.0014
0
0030
0
0000
0
00
1.76
0
.0016
0
.0042
0
0000
0
01
2.31
0
.0016
0
.0042
0
.0000
0
.01
2.31
0
. 1145
0
.0257
0
0000
0
14
14.02
Concord Scientitic Corfioration
B1-9
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSION
GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y)
3.59
3.03
8.33
6.44
2.21
8.03
3.73
2.72
5.12
3.74
4.04
2.69
0.00
1.22
2.61
2.61
2.87
1.06
1.22
3.28
Concord Scientitic Corporation
GRID
UTM
AREA
GRID
(Ha)
10000
101
10000
102
10000
103
10000
104
10000
105
10000
106
10000
107
10000
108
10000
109
10000
110
10000
111
10000
112
10000
113
2500
114
2500
115
2500
116
10000
117
10000
118
10000
119
2500
120
0.
0147
0.
0212
0.
0000
0.
04
0.
0182
0.
0121
0.
0000
0.
03
0.
0573
0.
0265
0.
0000
0.
08
0.
0534
0.
0110
0.
0000
0.
06
0.
0085
0.
0136
0.
,0000
0.
,02
0.
0286
0.
0518
0.
,0000
0.
08
0.
0143
0.
,0231
0.
,0000
0.
,04
0,
.0106
0.
.0166
0.
,0000
0.
,03
0.
.0180
0.
.0333
0.
,0000
0,
.05
0.
.0159
0.
,0215
0.
.0000
0.
.04
0.
.0177
0,
.0227
0,
.0000
0.
.04
0.
.0117
0.
.0151
0.
.0000
0.
.03
0
.0000
0,
.0000
0,
.0000
0,
.00
0
.0012
0
.0019
0
.0000
0
.00
0
.0016
0
.0049
0
.0000
0
.01
0
.0016
0
.0049
0
.0000
0
.01
0
.0083
0
.0204
0
.0000
0
.03
0
.0030
0
.0076
0
.0000
0
.01
0
.0039
0
.0083
0
.0000
0
.01
0
.0025
0
.0057
0
.0000
0
.01
B1-10
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSION
GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y)
0.0025 0.0057 0.0000 0.01 3,28
0.0044 0.0091 0.0000 0.01 1.34
0.0044 0.0076 0.0000 0.01 1.19
0.0322 0.0393 0.0000 0.07 7.15
0.0090 0.0147 0.0000 0.02 2.37
0.0732 0.0257 0.0000 0.10 9.89
0.0253 0.0159 0.0000 0.04 4.12
0.3567 0.3738 0.0000 0.73 73.05
0.0161 0.0268 0.0000 0.04 4.3C
0.''l691 0.1005 0.0000 0.27 26.97
0.0117 0.0200 0.0000 0.03 3.18
0.0159 0.0215 0.0000 0.04 3.74
0.0177 0.0227 0.0000 0.04 4.04
0.0000 0.0000 0.0000 0.00 O.OC
0.0071 0.0000 0.0000 0.01 0.71
0.2064 0.2532 0.0000 0.46 45. 9€
0.0041 0.0094 0.0000 0.01 5.44
0.0041 0.0094 0.0000 0.01 5.44
0.0173 0.0412 0.0000 0.06 5 . Ql
0.0092 0.0200 0.0000 0.03 2.9
Concord Scientific Corporation
GRID
UTM
AREA
GRID
(Ha)
2500
121
10000
122
10000
123
10000
124
10000
125
10000
126
10000
127
10000
128
10000
129
10000
130
10000
131
10000
132
10000
133
10000
135
10000
136
10000
137
2500
138
2500
139
10000
140
10000
141
B1-11
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
GRID
UTM
TOTAL PAH INVENTORY
FOR EACH
TOTAL
EMISSION
AREA
GRID
GRID AND SOURCE (MT/YEAR)
PAH
DENSITY
(Ha)
MOBILE
RES. WOOD
FOR. FIRE
(MT/YEAR)
(g/ha/y)
10000
142
0.0046
0.0102
0.0000
0.01
1.48
2500
143
0.0044
0.0106
0.0000
0.01
5.98
2500
144
0.0044
0.0106
0.0000
0.01
5.98
10000
145
0.0081
0.0193
0.0000
0.03
2.73
10000
146
0.0044
0.0072
0.0000
0.01
1.16
10000
147
0.0074
0.0121
0.0000
0.02
1.95
10000
148
0.0308
0.0223
0.0000
0.05
5.31
10000
149
0.0189
0.0287
0.0000
0.05
4.76
10000
150
0.0948
0.0231
0.0000
0.12
11.79
10000
151
0.'bl80
0.0302
0.0000
0.05
4.82
10000
152
0.0226
0.0393
0.0000
0.06
6.19
10000
153
0.0147
0.0246
0.0000
0.04
3.93
10000
154
0.0046
0.0079
0.0000
0.01
1.25
10000
155
0.1808
0. 1999
0.0000
0.38
38.07
10000
156
0.0263
0.0378
0.0000
0.06
6.41
10000
157
0.1071
0.1988
0.0000
0.31
30.59
10000
158
0.2853
0.2169
0.0000
0.50
50.23
10000
160
0.0094
0.0000
0.0000
0.01
0.94
10000
161
0.0030
0.0083
0.0000
0.01
1.13
10000
162
0.0035
0.0102
0.0000
0.01
1.37
Concord Scientific Corporation
B1-12
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSION
GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y)
GRID
UTM
AREA
GRID
(Ha)
2500
163
2500
164
10000
165
10000
166
10000
167
2500
168
2500
169
10000
170
10000
171
10000
172
10000
173
10000
174
10000
175
10000
176
10000
177
10000
178
10000
180
10000
181
10000
182
10000
183
0.
0041
0.
0094
0
0000
0
01
5
44
0.
0044
0
0106
0
0000
0
01
5
98
0
0237
0
0809
0
0000
0
10
10
46
0
0735
0
0752
0
0000
0
15
14
87
0
0058
0
0129
0
0000
0
02
1
86
0
0129
0
0537
0
0000
0
07
26
63
0
0088
0
0329
0
0000
0
04
16
65
0
0405
0
0299
0
0000
0
07
7
04
0
0359
0
0094
0
0000
0
05
4
54
0
1921
0
1262
0
0000
0
32
31
84
0
0633
0
0344
0
0000
0
10
9
77
0
0401
0
0276
0
0000
0
07
6
76
0
7822
0
7185
0
0000
1
50
150
06
1
2158
1
0730
0
0000
2
29
228
88
0
1008
0
0359
0
0000
0
14
13
67
0
.1206
0
0283
0
0000
0
15
14
89
0
.0194
0
.0329
0
.0000
0
.05
5
.22
0
.0191
0
.0325
0
.0000
0
.05
5
. 16
0
.0000
0
.0000
0
.0000
0
.00
0
.00
0
.0000
0
.0000
0
.0000
0
.00
0
.00
Concord Scientitic: Corfjoration
B1-13
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
GRID
UTM
TOTAL PAH INVENTORY
FOR EACH
TOTAL
EMISSION
AREA
GRID
GRID AND SOURCE (MT/YEAR)
PAH
DENSITY
(Ha)
MOBILE
RES. WOOD
FOR. FIRE
(MT/YEAR)
(g/ha/y)
10000
184
0.0000
0.0000
0.0000
0.00
0.00
10000
185
0.0369
0.0011
0.0000
0.04
3.80
10000
186
0.0094
0.0268
0.0000
0.04
3.63
10000
187
0.0094
0.0268
0.0000
0.04
3.63
10000
188
0.0064
0.0185
0.0000
0.02
2.50
10000
189
0.1269
0.1618
0.0000
0.29
28.87
10000
190
0.0000
0.0000
0.0000
0.00
0.00
10000
191
0.0000
0.0000
0.0000
0.00
0.00
10000
192
0. 1271
0.1183
0.0000
0.25
24.54
10000
193
0.0000
0.0000
0.0000
0.00
0.00
10000
194
0.2966
0.2649
0.0000
0.56
56.15
10000
195
0.0000
0.0000
0.0000
0.00
0.00
10000
196
0.1716
0.2729
0.0000
0.44
44.44
2500
197
0.0120
0.0624
0.0000
0.07
29.74
2500
198
0.0965
0.3738
0.0000
0.47
188.11
10000
199
0.0698
0.1107
0.0000
0.18
18.05
10000
200
0.0529
0.0442
0.0000
0.10
9.72
10000
201
0.0410
0.0189
0.0000
0.06
5.99
2500
202
0.0650
0.0654
0.0000
0.13
52. 14
2500
203
0.0334
0.1296
0.0000
0.16
65.21
Concord Scientific Corporation
B1-14
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSION
GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y
0.0104 0.0098 0.0000 0.02 2.02
0.0180 0.0178 0.0000 0.04 3.57
0.0067 0.0098 0.0000 0.02 1.6
0.0122 0.0185 0.0000 0.03 3.07
0.0154 0.0223 0.0000 0.04 3.77
0.1372 0.1380 0.0000 0.28 27.51
0.5078 0.4872 0.0000 0.99 99. 5C
0.0315 0.0265 0.0000 0.06 5.1
0.1672 0.1429 0.0000 ■ 0.31 31. OC
0Tl846 0.0378 0.0000 0.22 22.24
0.0194 0.0325 0.0000 0.05 5 . IS
0.0166 0.0306 0.0000 0.05 4.7
0.0104 0.0231 0.0000 0.03 3.3^
0.0106 0.0249 0.0000 0.04 3.55
0.0380 0.0163 0.0000 0.05 5.4
0.0832 0.0450 0.0000 0.13 12.8]
0.0391 0.0268 0.0000 0.07 6.66
0.0094 0.0268 0.0000 0.04 3.6:
0.0074 0.0208 0.0000 0.03 2.8:
0.0071 0.0117 0.0000 0.02 1 . 8<
Concord Scientific Corporation
GRID
UTM
AREA
GRID
(Ha)
10000
204
10000
205
10000
206
10000
207
10000
208
10000
209
10000
210
10000
211
10000
212
10000
213
10000
214
10000
215
10000
216
10000
217
10000
218
10000
219
10000
220
10000
221
10000
222
10000
223
B1-15
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSION
GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y)
GRID
UTM
AREA
GRID
(Ha)
10000
224
10000
225
10000
226
10000
227
10000
228
10000
229
10000
230
10000
231
10000
232
10000
233
10000
234
10000
235
10000
236
10000
237
10000
238
10000
239
10000
240
10000
241
10000
242
10000
243
0.
0105
0.
0178
0.
0000
0.
03
2.
84
0.
0203
0.
0605
0.
0000
0.
08
8.
07
0.
3178
0.
2793
0.
0000
0.
60
59.
71
0.
2799
0.
2623
0.
0000
0.
54
54.
22
0.
,0030
0.
,0068
0.
,0000
0,
,01
0.
,98
0.
0756
0.
0571
0.
,0000
0.
, 13
13,
,26
0.
,0060
0.
,0163
0,
,0000
0.
,02
2.
,22
0.
,0200
0,
.0246
0.
,0000
0,
,04
4.
,46
0.
.0262
0,
.0280
0.
,0000
0.
.05
5.
,42
0.
.0239
0.
.0223
0.
.0000
0.
.05
4.
.62
0.
.0285
0,
.0212
0,
.0000
0,
.05
4,
.97
0.
.0180
0,
.0200
0,
.0000
0.
.04
3,
.80
0,
.0134
0,
.0204
0,
.0000
0,
.03
3,
.38
0
.0044
0
.0068
0
.0000
0
.01
1,
.12
0
.0180
0,
.0302
0
.0000
0,
.05
4,
.82
0
. 1471
0
. 1512
0,
.0000
0
.30
29
.83
0
.3586
0
.2347
0
.0000
0
.59
59
.33
0
.0488
0
.0295
0
.0000
0
.08
7
.83
0
.0120
0
.0272
0
.0000
0
.04
3
.92
0
.2551
0
.3001
0
.0000
0
.56
55
. .2
Concord Scientit'ic Corporation
GRID
UTM
AREA
GRID
(Ha)
10000
244
10000
245
10000
246
10000
247
10000
248
10000
249
10000
250
10000
251
10000
252
10000
253
10000
254
10000
255
10000
256
250000
257
10000
258
10000
259
10000
260
10000
261
10000
262
10000
263
B1-16
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSION
GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y)
0.4389 0.4925 0.0000 0.93 93.13
0.0120 0.0166 0.0000 0.03 2.86
0.0352 0.0363 0.0000 0.07 7.15
0.0166 0.0310 0.0000 0.05 4.76
0.0182 0.0295 0.0000 0.05 4.77
0.0182 0.0295 0.0000 0.05 4.77
0.0283 0.0578 0.0000 0.09 8.62
0.0306 0.0903 0.0000 0.12 12.10
0.5445 0.5650 0.0000 1.11 110.95
0.^7001 0.5992 0.0000 1.40 139.93
0.0083 0.0181 0.0000 0.03 2.64
0.0046 0.0125 0.0000 0.02 1.71
0.0069 0.0147 0.0000 0.02 2.17
0.5142 0.6436 0.0000 1.16 4.63
0.0134 0.0076 0.0000 0.02 2.09
0.0244 0.0408 0.0000 0.07 6.52
0.0214 0.0325 0.0000 0.05 5.39
0.0304 0.0272 0.0000 0.06 5.76
0.0651 0.0208 0.0000 0.09 8.59
0.1138 0.1330 0.0000 0.25 24.68
Concord Scientit'ic Corp)ration
B1-17
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
GRID UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSION
AREA GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
(Ha) MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y)
10000 264
10000 265
10000 266
10000 267
10000 268
10000 269
10000 270
10000 271
10000 272
10000 273
10000 275
10000 276
10000 277
10000 278
10000 279
10000 280
10000 281
10000 282
10000 283
10000 284
0.
.0615
0,
.0593
0,
.0000
0.
. 12
12,
.08
0.
.0069
0,
.0163
0,
.0000
0.
.02
2,
.32
0,
.0509
0,
.0034
0,
.0000
0.
.05
5 ,
.43
0,
. 1950
0,
.1054
0,
.0000
0,
.30
30.
.04
0,
.0435
0,
.0571
0.
.0000
0,
.10
10,
.06
0.
.0403
0.
.0567
0,
.0000
0.
. 10
9,
.70
0,
.2418
0.
.1780
0.
.0000
0.
.42
41,
.98
0,
.2925
0.
.1327
0,
.0000
0.
.43
42.
.52
0.
. 1710
0.
.0283
0.
.0000
0,
.20
19,
.93
0.
."9 0 87
0.
,9169
0.
,0000
1.
,83
182.
.56
0.
. 1779
0,
.1746
0.
.0000
0,
.35
35.
.25
0.
.0069
0,
.0197
0.
.0000
0,
,03
2,
,66
0.
,0145
0.
.0129
0,
.0000
0,
.03
2.
,74
0,
.0150
0.
.0272
0.
.0000
0.
.04
4.
,22
0.
.0161
0.
.0261
0.
.0000
0.
,04
4.
,22
0,
.0203
0,
.0370
0.
.0000
0.
.06
5.
.73
0
.0440
0,
.0249
0,
.0000
0.
.07
6,
.89
0
.0822
0,
.0336
0,
.0000
0,
.12
11,
.58
0
.1872
0,
.1822
0,
.0000
0,
.37
36,
.94
0
.0030
0
.0042
0
.0000
0,
.01
0,
.72
Concord Scientific Corporation
GRID
UTM
AREA
GRID
(Ha)
10000
285
2500
286
2500
287
2500
288
2500
289
2500
290
2500
291
2500
292
2500
29 3
2500
294
2500
295
2500
296
2500
297
2500
298
2500
299
10000
300
10000
301
10000
302
10000
303
10000
304
B1-18
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSIO^
GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y)l
I
0.0078 0.0253 0.0000 0.03 3.32
0.2231 0.1901 0.0000 0.41 165.26
0.3872 0.3394 0.0000 0.73 290.63
0.8167 0.8500 0.0000 1.67 666.69
0.8414 0.9294 0.0000 1.77 708.30
0.5658 0.6077 0.0000 1.17 469.43
0.2304 0.0042 0.0000' 0.23 93.8ll
0.0617 0.0763 0.0000 0.14 55.22
0.1683 0.0998 0.0000 0.27 107. 2ll
072746 0.2627 0.0000 0.54 214.91
0.3724 0.2835 0.0000 0.66 262.35
0.0069 0.0204 0.0000 0.03 10.93
0.0377 0.0204 0.0000 0.06 23.26
0.0884 0.0204 0.0000 0.11 43.51
0.6246 0.3371 0.0000 0.96 384.70
0.0071 0.0159 0.0000 0.02 2.30
0.2659 0.3481 0.0000 0.61 61.40
0.0283 0.0140 0.0000 0.04 4.23
0.0569 0.0476 0.0000 0.10 10.45
0.2097 0.1187 0.0000 0.33 32.84
Concord Scientific Corporation
B1-19
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
GRID UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSION
AREA GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
(Ha) MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y)
10000 305
10000 306
10000 307
10000 308
10000 309
2500 310
2500 311
2500 312
2500 315
2500 315
10000 317
10000 318
10000 319
10000 320
10000 321
10000 322
10000 323
10000 324
10000 325
10000 326
0.
.5997
0,
,5242
0.
,0000
1,
. 12
112,
.39
0.
.8984
0,
,6145
0,
,0000
1,
.51
151,
,29
0.
.0709
0.
,0147
0,
.0000
0,
.09
8.
,56
0,
.0592
0,
,0412
0.
.0000
0,
,10
10.
,03
0.
.0378
0.
,0608
0,
.0000
0
.10
9,
,86
0,
,0005
0,
,0026
0.
.0000
0,
.00
1,
,24
0.
.4065
0.
,3345
0.
,0000
0.
,74
296.
,38
0,
.6076
0.
,4464
0.
,0000
1.
,05
421,
,59
0.
.2925
0.
,1073
0,
,0000
0.
,40
159,
,94
0,
,'l996
0.
,2151
0,
,0000
0,
,41
165,
,85
0,
.0069
0.
.0147
0.
.0000
0,
.02
2.
, 17
0.
.0081
0.
,0159
0,
.0000
0.
.02
2.
,39
0,
.0194
0.
.0257
0,
,0000
0,
.05
4,
,51
0,
.0247
0.
.0359
0,
.0000
0,
,06
6.
,06
0.
,4314
0.
,4713
0.
,0000
0,
.90
90.
,27
0,
.7606
0.
,6939
0,
,0000
1,
,45
145,
,45
0
.0071
0,
.0117
0,
,0000
0,
.02
1,
.89
0
.0378
0,
.0193
0,
.0000
0
.06
5.
.70
0,
.2087
0.
,0272
0,
.0000
0,
.24
23.
,59
0,
.3077
0,
.0026
0.
.0000
0,
.31
31.
,03
Concord Scientitic Corporation
B1-20
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSION
GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y)
GRID
UTM
AREA
GRID
(Ha)
10000
327
2500
329
2500
331
2500
332
10000
333
10000
334
10000
335
10000
336
10000
337
10000
338
10000
339
10000
340
10000
341
10000
342
10000
343
2500
344
2500
345
2500
346
2500
347
2500
349
0.
1758
0
0000
0
0000
0
18
0.
5591
0.
3583
0.
0000
0.
92
0
1981
0
2714
0.
0000
0
47
0
1150
0
2528
0.
0000
0
37
0
0085
0
0181
0
0000
0
03
0
0154
0
0351
0
0000
0
05
0
0154
0
0200
0
0000
0
04
0
0235
0
0253
0
0000
0
05
0
0359
0
0767
0
0000
0
11
0
'0 242
0
0540
0
0000
0
08
0
2560
0
2445
0
0000
0
50
0
5696
0
5507
0
0000
1
12
0
.0104
0
0193
0
0000
0
03
0
0672
0
1236
0
0000
0
19
0
2882
0
1584
0
0000
0
45
0
0426
0
0000
0
0000
0
04
0
.1233
0
.3258
0
0000
0
45
0
.5814
0
.6784
0
.0000
1
.26
0
.2244
0
.0000
0
.0000
0
.22
0
.3458
0
.6784
0
.0000
1
.02
Concord Scientitic Corporation
B1-21
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSION
GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y)
GRID
UTM
AREA
GRID
(Ha)
2500
350
10000
351
10000
352
10000
353
10000
354
10000
355
10000
356
10000
357
10000
358
10000
359
10000
360
10000
361
10000
362
2500
363
2500
364
2500
365
2500
366
2500
367
2500
368
2500
369
0.
4890
0.
8330
0.
0000
1.
32
528.
79
0.
0071
0.
0155
0.
0000
0.
02
2.
26
0.
0074
0.
0163
0.
0000
0.
02
2.
36
0.
0085
0.
0147
0.
0000
0.
02
2.
33
0.
0085
0.
0117
0.
,0000
0.
,02
2.
02
0.
0223
0.
0472
0.
,0000
0.
07
6.
,96
0.
1608
0.
0790
0.
,0000
0.
,24
23.
,97
0,
.0951
0.
.1153
0,
.0000
0.
,21
21,
,04
0.
.0164
0,
.0299
0,
.0000
0.
.05
4,
,62
0.
."0 59 6
0.
.0797
0.
.0000
0.
.14
13.
.94
0.
.0988
0.
.1236
0.
.0000
0,
.22
22.
.23
0.
.3625
0,
.2525
0.
.0000
0.
.62
61.
.50
1,
.6124
0.
,9003
0,
,0000
2,
.51
251,
.26
0
.0909
0,
.1742
0,
.0000
0.
,27
106,
.07
0
.4652
0
.5223
0
.0000
0,
,99
395
.00
0
.7256
0
.4959
0
.0000
1,
.22
488
.58
1
.3017
0
.6992
0
.0000
2
.00
800
.35
0
.3833
0
.2056
0
.0000
0
.59
235
.57
0
.3904
0
.3556
0
.0000
0
.75
298
.40
0
.4622
0
.5484
0
.0000
1
.01
404
.24
Concord Scientilic Corporation
B1-22
TABLE Bl-2 (cont'd]
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
GRID UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSION
AREA GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
(Ha) MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y)
2500 370
2500 371
2500 372
2500 373
10000 374
10000 375
10000 ■ 376
10000 377
10000 378
10000 379
10000 380
10000 381
10000 382
10000 383
10000 384
10000 385
2500 386
2500 387
2500 388
2500 389
1.
3732
1.
2211
0
0000
2.
59
1037
75
1.
9998
1.
2332
0
0000
3.
23
1293
22
1.
5791
1
2189
0
0000
2.
80
1119
18
1.
4849
0
9910
0
0000
2.
48
990
35
0
0074
0
0151
0
0000
0
02
2
25
0
1076
0
0922
0
0000
0
20
19
98
0
0071
0
0121
0
0000
0
02
1
92
0
0136
0
0242
0
0000
0
04
3
78
0
0076
0
0110
0
0000
0
02
1
86
0
0456
0
0246
0
0000
0
07
7
01
0
0090
0
0163
0
0000
0
03
2
52
0
0106
0
0200
0
0000
0
03
3
06
0
0270
0
0499
0
0000
0
08
7
68
0
0092
0
0163
0
.0000
0
03
2
55
0
0000
0
0000
0
.0000
0
00
0
00
0
1961
0
0707
0
0000
0
27
26
.68
0
.2275
0
.4350
0
.0000
0
66
264
.99
0
.8417
0
.6096
0
.0000
1
45
580
.53
1
.0972
0
.7162
0
.0000
1
.81
725
.36
0
.9374
0
.6115
0
.0000
1
.55
619
.55
Concord Scientific Cortx)ration
B1-23
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
GRID UTM TOTAL PAH INVENTORY FOR EACH ' TOTAL EMISSION
AREA GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
(Ha) MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y)
2500 390
2500 391
2500 392
2500 393
2500 394
2500 395
2500 396
2500 397
2500 398
2500 399
2500 400
10000 401
10000 402
10000 403
10000 404
10000 405
10000 406
10000 407
10000 408
10000 409
0
.4505
0
.8035
0
.0000
1
.25
501
.62
0
.3980
0
.6497
0,
.0000
1
.05
419
.08
0
.2739
0
.3288
0,
.0000
0
.60
241
.09
0,
.2063
0,
.2631
0,
.0000
0
.47
187
.72
0,
.3775
0,
.3526
0,
.0000
0
.73
292
.05
0,
.2995
0,
.5386
0,
.0000
0,
.84
335,
.24
0,
.7200
0,
.6765
0,
.0000
1,
.40
558,
.62
0.
.7612
0,
.6765
0.
.0000
1,
.44
575.
.09
0.
.7701
0.
.4619
0,
.0000
1,
,23
492,
.78
0.
.4907
0.
,4751
0.
,0000
0.
.97
386.
, 32
0.
.2739
0.
.3288
0.
,0000
0.
.60
241.
,09
0,
.0071
0.
.0144
0.
.0000
0.
.02
2.
, 15
0,
.0044
0,
.0083
0.
,0000
0,
,01
1.
,27
0.
,0193
0.
.0412
0.
.0000
0.
,06
6.
,05
0.
.0299
0.
,0098
0.
,0000
0,
.04
3.
98
0.
,0589
0,
.0355
0.
,0000
0.
.09
9.
45
0.
.0058
0.
.0102
0.
.0000
0.
.02
1.
60
0.
.0044
0.
.0117
0.
.0000
0,
.02
1.
61
0.
.0062
0.
,0147
0,
.0000
0.
.02
2.
10
0,
.0044
0.
.0076
0.
.0000
0.
.01
1.
19
Concord Scientific Corporation
B1-24
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
GRID UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSIO^
AREA GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
(Ha) MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y)
10000 410
10000 411
10000 412
10000 413
2500 414
2500 415
2500 416
2500 417
2500 418
2500 419
10000 420
10000 421
10000 422
2500 423
2500 424
2500 425
2500 426
2500 427
2500 428
250000 429
0.
1411
0.
2321
0
0000
0
37
0.
0000
0.
0000
0
0000
0
00
0.
0205
0
0665
0
0000
0
09
0
3392
0.
1134
0
0000
0
45
0
0803
0
0283
0
0000
0
11
0
0803
0.
1107
0
0000
0
19
0
4350
0
3995
0
0000
0
83
0
2739
0
3288
0
0000
0
60
0
6481
0
4751
0
0000
1
12
0
4405
0
4018
0
0000
0
84
1
0963
0
5809
0
0000
1
68
0
2960
0
2128
0
0000
0
51
0
.0523
0
0495
0
0000
0
10
0
0495
0
0900
0
0000
0
14
0
.2709
0
2502
0
0000
0
52
0
.1786
0
0000
0
0000
0
18
0
.2308
0
2676
0
.0000
0
50
0
.1425
0
.1765
0
.0000
0
.32
0
.1522
0
. 1825
0
.0000
0
.33
0
.3573
0
.4898
0
.0000
0
.85
Concord Sc lentitic CorjToration
Bl-25
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
GRID
UTM
TOTAL PAH INVENTORY
FOR EACH
TOTAL
EMISSION
AREA
GRID
GRID AND SOURCE (MT/YEAR)
PAH
DENSITY
(Ha)
MOBILE
RES. WOOD
FOR. FIRE
(MT/YEAR)
(g/ha/y)
250000
430
0.5639
0.5737
0.0000
1.14
4.55
250000
431
0.3642
0.6006
0.0000
0.96
3.86
250000
432
2.6367
1.8773
0.0000
4.51
18.06
250000
433
2.8053
2.2322
0.0000
5.04
20.15
250000
434
1.9859
1.5016
0.0000
3.49
13.95
10000
435
0. 1084
0.0370
0.0000
0.15
14.54
10000
436
0.0154
0.0094
0.0000
0.02
2.49
10000
437
0.0716
0.0563
0.0000
0. 13
12.79
10000
438
0.0081
0.0174
0.0000
0.03
2.54
10000
439
o.'oioi
0.0234
0.0000
0.03
3.36
10000
440
0.0012
0.0023
0.0000
0.00
0.34
10000
441
0.0131
0.0295
0.0000
0.04
4.26
10000
442
0.0060
0.0185
0.0000
0.02
2.45
10000
443
0.1511
0.1104
0.0000
0.26
26.15
10000
444
0.0058
0.0140
0.0000
0.02
1.97
10000
445
0.0012
0.0019
0.0000
0.00
0.30
10000
446
0.0058
0.0144
0.0000
0.02
2.01
10000
447
0.0041
0.0102
0.0000
0.01
1.44
10000
448
0.0028
0.0072
0.0000
0.01
0.99
10000
449
0.0016
0.0042
0.0000
0.01
0.58
Concord Scientific Corporation
B1-26
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
GRID
UTM
TOTAL PAH INVENTORY
FOR EACH
TOTAL
EMISSION
AREA
GRID
GRID AND SOURCE (MT/YEAR)
PAH
DENSITY
(Ha)
MOBILE
RES. WOOD
FOR. FIRE
(MT/YEAR)
(g/ha/y)
10000
450
0.0104
0.0189
0.0000
0.03
2.93
10000
451
0.1010
0.0178
0.0000
0. 12
11.88
10000
452
0. 1277
0.0238
0.0000
0.15
15.15
10000
453
0.2998
0. 1935
0.0000
0.49
49.33
10000
454
0.0431
0.0389
0.0000
0.08
8.20
10000
455
0.0060
0.0144
0.0000
0.02
2.04
10000
456
0.0055
0.0129
0.0000
0.02
1.84
10000
457
0.0055
0.0136
0.0000
0.02
1.91
10000
458
0.0051
0.0110
0.0000
0.02
1.60
10000
459
0.0025
0.0064
0.0000
0.01
0.90
10000
460
0.0014
0.0045
0.0000
0.01
0.59
10000
461
0.0081
0.0159
0.0000
0.02
2.39
10000
462
0.0104
0.0163
0.0000
0.03
2.66
10000
463
0.0191
0.0348
0.0000
0.05
5.39
10000
464
0.1651
0.0884
0.0000
0.25
25.35
10000
465
0.3210
0.2457
0.0000
0.57
56.67
10000
466
0.3626
0.2264
0.0000
0.59
58.90
10000
467
0.0700
0.0155
0.0000
0.09
8.55
10000
468
0.0244
0.0465
0.0000
0.07
7.09
10000
469
0.0735
0.0590
0.0000
0.13
13.24
Concord Scientific Corporation
B1-27
TABLE Bl-2 (cont'd)
Regional Annual PAH Elmissions and Elmission Density
from Three Major Sources to the Atmosphere in Ontario
GRID
UTM
TOTAL PAH INVENTORY
FOR EACH
TOTAL
EMISSION
AREA
GRID
GRID AND SOURCE (MT/YEAR)
PAH
DENSITY
(Ha)
MOBILE
RES. WOOD
FOR. FIRE
(MT/YEAR)
(g/ha/y)
10000
470
0.0253
0.0442
0.0000
0.07
6.96
10000
471
0.0173
0.0382
0.0000
0.06
5.55
10000
472
0.3779
0.3640
0.0000
0.74
74.18
10000
473
0.4046
0.3621
0.0000
0.77
76.66
10000
474
0.0012
0.0023
0.0000
0.00
0.34
10000
475
0.0076
0.0151
0.0000
0.02
2.27
10000
476
0.0834
0.0658
0.0000
0.15
14.92
10000
477
0.0101
0.0204
0.0000
0.03
3.05
10000
478
0.0818
0.0820
0.0000
0.16
16.38
10000
479
0.0171
0.0472
0.0000
0.06
6.43
10000
480
0.0115
0.0242
0.0000
0.04
3.57
10000
481
0.0028
0.0068
0.0000
0.01
0.96
10000
482
0.0824
0.0144
0.0000
0. 10
9.68
10000
483
0.1198
0.0767
0.0000
0.20
19.65
10000
484
0. 1264
0.0472
0.0000
0.17
17.36
10000
485
0.0661
0.0646
0.0000
0.13
13.07
10000
486
0.2136
0. 1739
0.0000
0.39
38.75
10000
487
0.1349
0.0790
0.0000
0.21
21.39
10000
488
0.0622
0.0336
0.0000
0.10
9.58
10000
489
0.1073
0.0865
0.0000
0.19
19.39
Concord Scientiiic Corporation
B1-28
TABLE Bl-2 (cont'd!
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
GRID
UTM
TOTAL PAH INVENTORY
FOR EACH
TOTAL
EMISSIOr
AREA
GRID
GRID AND SOURCE (MT/YEAR)
PAH
DENSITY
(Ha)
MOBILE
RES. WOOD
FOR. FIRE
(MT/YEAR)
(g/ha/y)
250000
490
0.6588
0.7831
0.0000
1.44
5.7'.
250000
491
0.2428
0.3148
0.0000
0.56
2.2:
250000
492
0.5915
0.8213
0.0000
1.41
5.65
250000
493
2.2969
1.2257
0.0000
3.52
14. OS
250000
494
0.4371
0.5401
0.0222
1.00
4.0C
250000
495
0.5304
0.6758
0.0000
1.21
4.82
250000
496
0. 1762
0.2657
0.0934
0.54
2. 14
250000
497
0.1566
0.2294
0.0000
0.39
1.54
250000
498
0.2487
0.3277
0.0000
0.58
2.31
250000
499
0.'8504
0.6856
0.0000
1.54
6. 14
250000
500
0.1214
0.1130
0.0000
0.23
0.94
250000
501
0.0062
0.0094
0.0000
0.02
0.06
250000
502
0.0219
0.0404
0.0000
0.06
0.25
250000
503
0.5202
0.3693
0.0474
0.94
3.75
250000
504
0.0730
0. 1107
0.0113
0.20
0.78
250000
505
0.0654
0. 1247
0.0113
0.20
0.81
250000
506
0.1422
0.1232
0.0241
0.29
1.16
250000
507
0.0283
0.0540
0.0000
0.08
0.33
250000
508
0.6407
0.6145
0.0039
1.26
5.04
250000
509
0.5575
0.6297
0.0000
1.19
4.75
Concord Scientitic CorfX)rjition
B1-29
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Elmission Density
from Three Major Sources to the Atmosphere in Ontario
GRID
UTM
TOTAL PAH INVENTORY
FOR EACH
TOTAL
EMISSION
AREA
GRID
GRID AND SOURCE (MT/YEAR)
PAH
DENSITY
(Ha)
MOBILE
RES. WOOD
FOR. FIRE
(MT/YEAR)
(g/ha/y)
250000
510
0.6447
0.4959
0.0000
1. 14
4. 56
250000
511
0.5756
0.5303
0.0000
1. 11
4.42
250000
512
0.0513
0.0336
0.0000
0.08
0.34
250000
513
0.0253
0.0253
0.0000
0.05
0.20
250000
514
0.5019
0.1542
0.0105
0.67
2.67
250000
515
0.0843
0.1380
0.0105
0.23
0.93
250000
516
0.3350
0.2324
0.0109
0.58
2.31
250000
517
0.0177
0.0314
0.0004
0.05
0.20
250000
518
0.0385
0.0949
0.0004
0.13
0.53
250000
519
0.'l545
0.1644
0.0000
0.32
1.28
250000
520
0.4505
0.4074
0.0000
0.86
3.43
10000
521
0.0479
0.0265
0.0000
0.07
7.43
10000
522
0.1167
0.1081
0.0000
0.22
22.48
10000
523
0.0930
0.0703
0.0000
0.16
16.33
10000
524
0.5365
0.7328
0.0000
1.27
126.93
10000
525
0.0051
0.0094
0.0000
0.01
1.45
10000
526
0.0187
0.0310
0.0000
0.05
4.96
10000
527
0.0274
0.0359
0.0000
0.06
6.33
10000
528
0.0143
0.0265
0.0000
0.04
4.07
10000
529
0.0778
0.1145
0.0000
0.19
19.23
Concord Scientific Corporation
B1-30
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
GRID UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSION
AREA GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
(Ha) MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y)
10000 530
10000 531
10000 532
10000 533
10000 534
10000 535
10000 538
10000 539
10000 540
10000 541
250000 542
250000 543
250000 544
250000 545
250000 546
250000 547
250000 548
250000 549
250000 550
250000 551
0.
3881
0
2219
0
0000
0
61
2.
1210
1
7945
0
0000
3
92
0
2944
0
3024
0
0000
0
60
0
2221
0
0389
0
0000
0
26
0
1278
0
0624
0
0000
0
19
0
0111
0
0197
0
0000
0
03
0
5566
0
4887
0
0000
1
05
0
3614
0
2033
0
0000
0
56
0
2021
0
0586
0
0000
0
26
0
'0364
0
0548
0
0000
0
09
0
9286
0
5499
0
0000
1
48
0
2137
0
2200
0
0000
0
43
0
0025
0
0064
0
0000
0
.01
0
0154
0
0265
0
0000
0
04
0
1938
0
0847
0
0000
0
28
0
0136
0
0147
0
0000
0
03
0
.0076
0
0110
0
0105
0
03
0
.0113
0
.0178
0
0000
0
.03
0
.2123
0
.1402
0
.0105
0
36
0
.0035
0
.0045
0
.0004
0
.01
Concord Scientific Cor['x)mtion
B1-31
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
GRID
UTM
TOTAL PAH INVENTORY
FOR EACH
TOTAL
EMISSION
AREA
GRID
GRID AND SOURCE (MT/YEAR)
PAH
DENSITY
(Ha)
MOBILE
RES. WOOD
FOR. FIRE
(MT/YEAR)
(g/ha/y)
250000
552
0.0002
0.0008
0.0004
0.00
0.01
250000
553
0.2596
0.2978
0.0004
0.56
2.23
250000
554
0.3965
0.2778
0.0625
0.74
2.95
250000
555
0.0970
0.0722
0.0000
0. 17
0.68
250000
556
0. 1487
0.0491
0.0000
0.20
0.79
250000
557
0.2283
0.2022
0.0069
0.44
1.75
250000
558
0.2434
0.1323
0.0130
0.39
1.55
250000
559
0.2227
0.2547
0.0158
0.49
1.97
10000
566
0.0012
0.0026
0.0000
0.00
0.38
10000
569
0.7354
0.7544
0.0000
1.49
148.98
10000
570
0.0062
0.0076
0.0000
0.01
1.38
10000
571
0.0378
0.0231
0.0000
0.06
6.08
10000
572
0.0742
0.0801
0.0000
0.15
15.43
10000
573
0.0327
0.0344
0.0000
0.07
6.71
10000
574
0.0251
0.0306
0.0000
0.06
5.57
250000
575
0.2241
0.0858
0.0000
0.31
1.24
250000
576
0.7284
0.5934
0.0465
1.37
5.47
250000
577
0.2584
0.1803
0.0465
0.49
1.94
250000
578
0.0870
0.0480
0.0000
0. 14
0.54
250000
579
0.0345
0.0049
0.0000
0.04
0.16
Concord Scientific Corporation
B1-32
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSION
GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y)
GRID
UTM
AREA
GRID
(Ha)
10000
580
10000
581
10000
582
10000
583
10000
584
10000
585
10000
586
1000000
587
250000
588
250000
591
250000
592
250000
593
250000
594
250000
595
250000
59-8
250000
599
250000
600
250000
601
250000
602
1000000
603
0
0025
0
0034
0
0000
0
01
0
59
0
0569
0
0506
0
0314
0
14
13
89
0
9098
0
7597
0
0314
1
70
170
09
0
0000
0
0000
0
0000
0
00
0
00
0
0465
0
0038
0
0000
0
05
5
03
0
0387
0
0038
0
0000
0
04
4
24
0
0025
0
0038
0
0000
0
01
0
63
0
0838
0
0094
0
0000
0
09
0
09
0
0000
0
0000
0
0000
0
00
0
00
0
'0189
0
0238
0
0158
0
06
0
23
0
3770
0
3938
0
0000
0
77
3
08
0
0173
0
0238
0
0000
0
04
0
16
0
0854
0
0174
0
0000
0
10
0
41
0
0037
0
0053
0
0000
0
01
0
04
0
0062
0
0087
0
0158
0
03
0
12
0
0062
0
0087
0
0158
0
03
0
12
0
.0062
0
0087
0
0000
0
01
0
06
0
.0099
0
0151
0
2675
0
29
1
17
0
. 1446
0
1293
0
0132
0
29
1
15
0
.0537
0
0850
1
0569
1
20
1
20
Concord Scientitic Corporation
B1-33
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
GRID UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSION
AREA GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
(Ha) MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y)
1000000 604
1000000 605
250000 606
250000 607
250000 608
250000 609
250000 610 ■
250000 611
250000 612
250000 613
250000 614
250000 615
1000000 616
250000 617
250000 618
250000 619
250000 620
250000 623
250000 624
1000000 627
Concord Scientitic Corporation
0.
1006
0.
0627
0.
.0249
0.
. 19
0.
,19
0.
0567
0.
.0620
0.
.0624
0,
.18
0.
, 18
0.
,0062
0.
.0087
0.
.0158
0.
.03
0.
,12
0,
,0166
0,
.0227
0.
.0158
0.
,06
0.
,22
0.
.0065
0,
.0098
0,
,0158
0,
,03
0,
.13
0.
,0122
0.
,0204
0.
,0000
0,
.03
0,
.13
0.
,2852
0,
. 1965
0.
,0000
0.
.48
1.
.93
0,
.0154
0.
.0215
0.
,0000
0.
,04
0.
. 15
0,
.0025
0.
.0045
0,
,0000
0.
,01
0.
.03
0.
.'0039
0.
,0068
0.
.0000
0,
.01
0.
.04
0.
.0389
0.
.0102
0,
,0000
0.
.05
0.
.20
0,
.1426
0.
.2460
2,
.8877
3,
.28
13.
.11
0.
,0233
0.
,0253
0,
.0000
0,
.05
0.
.05
0,
.0000
0,
,0000
0
.0000
0,
,00
0.
.00
0.
,0000
0.
.0000
0,
.0000
0.
,00
0.
.00
0.
,4460
0,
.0488
0
.2154
0.
.71
2,
,84
0
.0861
0
.0000
0
.0000
0
.09
0,
,34
0
.0670
0
.0000
0
.0000
0
.07
0,
.27
0
.1297
0
.0737
0
.0133
0
.22
0
.87
0
.1482
0
.0333
0
.0976
0
.28
0
.28
B1-34
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
GRID UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSION
AREA GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
(Ha) MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y)
250000 628 0.2053 0.0125 0.1435 0.36 1.45
250000 629 0.0090 0.0042 0.0000 0.01 0.05
250000 630 0.0030 0.0042 0.0000 0.01 0.03
250000 631 0.0030 0.0042 0.0000 0.01 0.03
250000 632 0.0030 0.0042 0.0000 0.01 0.03
250000 633 0.0030 0.0042 0.0000 0.01 0,03
250000 634 0.0030 0.0042 0.0000 0.01 0.03
250000 635 0.0030 0.0042 0.0000 0.01 0.03
250000 636 0.0030 0.0042 0.0000 0.01 O.OI-
1000000 640 0."0012 0.0000 0.0000 0.00 O.OG
1000000 644 0.0106 0.0042 0.2766 0.29 0.2
1000000 647 0.0074 0.0102 0.0000 0.02 0.0
1000000 648 0.0041 0.0000 0.0000 0.00 0.00
1000000 649 0.0012 0.0011 0.0000 0.00 0.00
1000000 650 0.0088 0.0140 0.0000 0.02 0.02
1000000 653 0.0972 0.0941 0.0000 0.19 0.19
1000000 654 0.1877 0.1587 0.6042 0.95 0.95
1000000 655 0.1327 0.0756 0.1385 0.35 0.35
1000000 656 0.1353 0.0189 0.0000 0.15 0.15
10000 657 0.0088 0.0136 0.0000 0.02 2.24
Concord Scientilic Corp<.)ration
B1-35
TABLE Bl-2 (cont'd)
Regional Annual PAH Elmissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
GRID UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSION
AREA GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
(Ha) MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y)
10000 658
10000 659
10000 660
10000 661
10000 662
10000 663
10000 665
10000 666
10000 667
250000 683
1000000 684
1000000 685
1000000 686
250000 688
1000000 689
1000000 690
1000000 691
1000000 692
1000000 693
1000000 694
Concord Scientific Corporation
0.
,0779
0.
,0722
0,
.0000
0.
,15
15,
,01
0.
,1410
0.
, 1213
0.
.0000
0.
,26
26,
,23
0.
0221
0.
.0193
0.
.0000
0.
,04
4.
, 14
0.
0977
0.
,0639
0,
.0000
0.
.16
16,
. 15
1.
0984
0.
,8545
0.
,0000
1,
,95
195.
,29
0.
0074
0.
,0113
0.
,0000
0,
.02
1,
,87
0.
,1527
0.
.1179
0.
,0000
0,
.27
27.
.06
0.
.0214
0.
,0000
0.
.0000
0,
.02
2,
. 14
0,
.0189
0.
.0000
0.
,0000
0.
,02
1.
.89
0.
.0474
0.
.0060
0.
.0000
0.
.05
0,
.21
0.
. 1193
0.
.0782
0.
.0319
0,
.23
0.
.23
0.
.1635
0.
,0922
1.
,0569
1.
.31
1.
.31
0.
,1027
0,
,0586
0,
.0000
0,
.16
0.
,16
0.
.0755
0.
,0548
0.
,0108
0,
,14
0.
,56
0.
.3250
0.
,2831
4.
.3276
4.
.94
4.
,94
0.
.3691
0,
.1489
0.
.2458
0.
.76
0.
.76
0
.1515
0
.0952
8,
,2075
8.
,45
8.
.45
0
.0062
0
.0117
1,
.0569
1,
.07
1,
.07
0
.0122
0
.0200
1
.0569
1,
.09
1.
,09
0
.0569
0
.0272
1
.0569
1
.14
1,
,14
B1-36
TABLE Bl-2 (cont'd)
Regional Annual PAH Emissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
GRID UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSIOli
AREA GRID GRID AND SOURCE (MT/YEAR) PAH DENSIT'^
(Ha) MOBILE RES. WOOD FOR. FIRE (MT/YEAR) (g/ha/y
1000000 695
1000000 696
1000000 • 697
1000000 698
1000000 699
1000000 700
1000000 701
1000000 702
1000000 703
1000000 706
1000000 707
1000000 708
1000000 709
1000000 710
1000000 711
1000000 717
1000000 723
1000000 724
1000000 725
1000000 726
Concord S( lentitit Q)r|X)rciti()n
0
1099
0
1051
14
2953
14
51
0
0490
0
0302
0
0000
0
08
0
0071
0
0181
0
0000
0
03
0
0399
0
0998
41
5101
41
65
0
0071
0
0189
0
0000
0
03
0
0062
0
0170
0
0000
0
02
0
0000
0
0000
0
0000
0
00
0
0000
0
0000
0
0000
0
00
0
0048
0
0060
0
0000
0
01
0
0094
0
0185
0
0000
0
03
0
0060
0
0140
0
0000
0
02
0
0060
0
0159
0
0000
0
02
0
0009
0
0023
0
0000
0
00
0
0048
0
0117
0
0000
0
02
0
0002
0
0011
0
0000
0
00
0
0023
0
0053
0
0000
0
01
0
.0000
0
.0000
0
0000
0
00
0
.0025
0
.0049
0
.0000
0
.01
0
.0035
0
.0072
0
.0000
0
.01
0
.0039
0
.0079
0
.0000
0
.01
B1-37
TABLE Bl-2 (cont'd)
Regional Annual PAH Elmissions and Emission Density
from Three Major Sources to the Atmosphere in Ontario
GRID UTM TOTAL PAH INVENTORY FOR EACH TOTAL EMISSION
AREA GRID GRID AND SOURCE (MT/YEAR) PAH DENSITY
(Ha) MOBILE RES. WOOD FnOR.FIRE (MT/YEAR) (g/ha/y)
1000000 731 0.0023 0.0038 0.0000 0.01 0.01
1000000 732 0.0014 0.0030 0.0000 0.00 0.00
1000000 733 0.0025 0.0049 0.0000 0.01 0.01
1000000 740 0.0014 0.0023 0.0000 0.00 0.00
1000000 743 0.0012 0.0015 0.0000 0.00 0.00
TOTAL 102.17 88.89 79.28 270.34
Concord Scientitic Corporation
B1-38
TABLE B1-3
Colour Coding for Illustrative Regional PAH Maps
Colour Code
PAH Emission
Density
(gyha/yr)
blue
green
yellow
red
violet
<10
10-50
>50-100
>100-500
>500
B1-39
FIGURE B1-1
PAH Emission Densities for Southwestern Ontario
Concord Scientific CorfX)ration
^J*^.
)\i >--•«
B1-40
FIGURE B1-2
PAH Emission Densities for Southeastern Ontario
Concord Scientific Corporation
■ 1
:•:•-'
.-•'.•■». *\
. *. *
: < ^ n
-4- ^■••1
. < •!: -1
.^^^— ^
B1-41
FIGURE B1-3
PAH Emission Densities for Central Ontario
Concord Scientific Corfwration
B1-42
FIGURE B1-4
PAH Emission Densities for Northern Ontario
Concord Scientitic CorfDomtion
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