United States Environmental Criteria and EPA/600/8-84/020bF 

Environmental Protection Assessment Office August 1986 

Agency Research Triangle Park, NC 27711 

rch and Development 

i ~ : .. 

Air Qua l itIA 

• • 

\ 

.. 

,1 

® 

EPA/600/8-84/020bF 

Augus# 1986 

Air Quality Criteria 

for ozone ancl other 

Photochemical oxiclants 

Volume II of V 

Environmental Criteria and Assessment Office 

Office of Health and Environmental Assessment 

Office of Research and Development 

U.S. Environmental Protection Agency 

Research Triangle Park, N.C. 27711 
DISCLAIMER 

Thi s document has been revi ewed i n accordance wi tli U. S.' Envi'ronme`nt`al 

Protection Agency policy and approved for publicat'ion,.` Mention of -trade'' 

names or commercial products does not constitute endorsement or recommeridation 

for use. 

ii 
ABSTRACT. 

Scientific information is presented and evaluated-relative to., :the„ health 

and welfare effects associated with exposure to ozone and other' photochemical. 

oxidants. Although it is not intended as a complete. and detailed`l,iterature , 

review, the document covers pertinent literature through early 1986. 

Data on health and welfare effects are emphasized, but additional information is provided for understanding the nature of the oxidant pollution problem and for evaluating the reliability of effects data as well as their 

relevance to potential exposures to ozone and other oxidants at concentrations 

occurring in ambient air. Information is provided on the following exposurerelated topics: nature, source, measurement, and concentrations of precursors 

to ozone and other photochemical oxidants; the formation of ozone and other 

photochemical oxidants and their transport once.formed; the properties, chemistry, and measurement of ozone and other photochemical oxidants; and the 

concentrations of ozone and other photochemical oxidants that are typically 

found in ambient air. 

The specific areas addressed by chapters on health and welfare effects 

are the toxicological appra,isal of effects of ozone and other oxidants; effects 

observed in controlled human exposures; effects observed in field and epidemiological studies; effects on vegetation seen in field and controlled exposures; 

effects on natural and agroecosystems; and effects on nonbiological materials 

observed in field and chamber studies. 

iii 
AIR QUALITY CRITERIA FOR OZONE 

AND OTHER PHOTOCHEMICAL OXIDANTS 

Page 

VOLUME I 

Chapter 1. Summary and Conclusions .............................. 1-1 

VOLUME II 

Chapter 2. Introduction ........................................ 2-1 

Chapter 3. Properties, Chemistry, and Transport of Ozone and 

Other Photochemical Oxidants and Their Precursors .... 3-1 

Chapter 4. Sampling and Measurement of Ozone and Other 

Photochemical Oxidants and Their Precursors .......... 4-1 

Chapter 5. Concentrations of Ozone and Other Photochemical 

Oxidants in Ambient Air .............................. 5-1 

VOLUME III 

Chapter 6. Effects of Ozone and Other Photochemical Oxidants 

on Vegetation ........................................ 6-1 

Chapter 7. Effects of Ozone on Natural Ecosystems and Their 

Components .......... .... ......... . ... ......... 7-1 

Chapter 8. Effects of Ozone and Other Photochemical Oxidants 

on Nonbiological Materials ........................... 8-1 

VOLUME IV 

Chapter 9. Toxicological Effects of Ozone and Other 

Photochemical Oxidants ............................... 9-1 

VOLUME V 

Chapter 10. Controlled Human Studies of the Effects of Ozone 

and Other Photochemical Oxidants ..................... 10-1 

Chapter 11. Field and Epidemiological Studies of the Effects 

of Ozone and Other Photochemical Oxidants ............ 11-1 

Chapter 12. Evaluation of Health Effects Data for Ozone and 

Other Photochemical Oxidants ......................... 12-1 

iv 
TABLE OF CONTENTS 

Page 

LIST OF TABLES ........................................................ t xi 

LIST OF FIGURES . . . ...... . ..................................... xiv 

LIST OF ABBREVIATIONS AND SYMBOLS ..................................... xvii 

AUTHORS, CONTRIBUTORS, AND REVIEWERS .................................. xxii 

2. INTRODUCTION ...................................................... 2-1 

2.1 PURPOSE AND LEGISLATIVE BASIS OF THIS DOCUMENT .............. 2-1 

2.2 THE OXIDANT PROBLEM ....... ..:............................... 2-2 

2.3 SCOPE AND ORGANIZATION OF THIS DOCUMENT ..................... 2-4 

2.4 REFERENCES .................................................. 2-7 

3. PROPERTIES, CHEMISTRY, AND TRANSPORT OF OZONE AND OTHER 

PHOTOCHEMICAL OXIDANTS AND THEIR PRECURSORS ...... :............... 3-1 

3.1 INTRODUCTION ..... .. .... ......... . . . ... ..... .... ............. 3-1 

3.2 DESCRIPTIONS AND PROPERTIES OF OXIDANTS AND THEIR 

PRECURSORS ................................................... 3-2 

3.2.1 Ozone and Other Photochemical Oxidants ............... 3-2 

3.2.1.1 Ozone ....................................... 3-2 

3.2.1.2 Peroxyacetyl Nitrate ........................ 3-3 

3.2.1.3 Hydrogen Peroxide ............ , ............. 3-4 

3.2.1.4 Formic Acid ................................. 3-7 

3.2.2 Organic Precursors .................................... 3-7 

3.2.2.1 Hydrocarbons ................................ 3-8 

3.2.2.2 Aldehydes . ... ............................. 3-10 

3.2.2.3 Other Organic Compounds ...................... 3-11 

3.2.2.4 Volatility and Reactivity ................... 3-11 

3.2.3 Nitrogen Oxides ...................................... 3-14 

3.3 ATMOSPHERIC CHEMICAL PROCESSES: FORMATION AND TRANSFORMATION 

OF OZONE AND OTHER PHOTOCHEMICAL OXIDANTS ................... 3-15 

3.3.1 Inorganic . . ..... ... .... 3-16 

3.3.1.1 

Reactions . . ..... 

Formation of Ozone: TheNO-NO • 2 -03 Cycle .... 3-16 

3.3.1.2 Formation of Radical Intermediates .......... 3-17 

3.3.1.3 Termination Reactions ...................... 3-21 

3.3.1.4 Reactions Invoiving Nitrous Acid ............ 3-22 

3.3.1.5 Reactions Involving Nitric Acid and 

Dinitrogen Pentoxide ........................ 3-23 

3.3..2 Organic Reactions .................................... 3-24 

3.3.2.1 Reactions with Hydroxyl Radicals ............ 3-25 

3.3.2.2 Reactions with Ozone ........................ 3-34 

3.3.2.3 Reactions with Nitrate Radicals ............. 3-41 

3.3.3 Atmospheric Lifetimes of Organic Compounds ........... 3-44 

3.3.4 Atmospheric Reactions of Peroxyacetyl Nitrate ........ 3-45 

3.3.5 Role of Ozone in Aerosol Formation ................... 3-47 

3.3.5.1 Formation of Sulfate Aerosol ................ 3-47 

3.3.5.2 Formation of Nitrate Aerosol ................ 3-48 

3.3.5.3 Formation of Organic Aerosols ............... 3-49 

v 
TABLE.OF CONTENTS 

(continued) 

Pa e 

3.3.6 Role of Ozone and Other Photochemical Oxidants in the 

Acidification of Rain ................................ 3-50 

3.3.6.1 Reactions of Ozone in Aqueous Droplets ...... 3-50 

3.3.6.2 Reactions of Hydrogen Peroxide in Aqueous 

Droplets .................................... 3-51 

3.3.6.3 Reactions of Formic Acid in Aqueo,us 

. Droplets .................................... 3-54 

3.4 METEOROLOGICAL AND CLIMATOLOGICAL PROCESSES ................. 3-54 

3.4.1 Atmospheric Mixing ................................... 3-54 

3.4.2 Wind Speed and Mixing ................................ 3-61 

3.4.3 Effects of Sunlight and Temperature .................. 3-66 

3.4.4 Transport of Ozone and Other Oxidants and Their 

Precursors .... ................ ................... 3-68 

3.4.5 Surface Scavenging in Relation to Transport .......... 3-74 

3.4.6 Stratospheric-Tropospheric Ozone Exchange ............ 3-75 

3.4.7 Stratospheric Ozone at Ground Level .................. 3-80 

3.4.8 Background Ozone from Photochemical Reactions ........ 3-84 

3.5 PRECURSORS TO OZONE AND OTHER PHOTOCHEMICAL OXIDANTS ........ 3-88 

3.5.1 Sources and Emissions ................................ 3-88 

3.5.1.1 Manmade Sources and Emissions ............... 3-88 

3.5.1.2 Natural Sources and Emissions ............... 3-98 

3.5.2 Representative Concentrations of Ozone Precursors 

in Ambient Air ............................ .......... 3-104 

3.5.2.1 Concentrations of Nonmethane Organic 

Compounds in Ambient Air ......... 3-104 

3.5.2:2 Concentrations of Nitrogen Oxides in 

Ambient Air ................................. 3-108 

3.6 SOURCE-RECEPTOR (OXIDANT-PRECURSOR) MODELS .............."..... 3-110 

3.6.1 Definitions, Descriptions, and Use ................... 3-112 

3.6.1.1 Statistical Models .............:............ 3-112 

3.6.1.2 Trajectory Models ........................... 3-113 

3.6.1.3 Fixed-Grid Models ........................... 3-119 

3.6.1.4 Box Models .................................. 3-121 

3.6.2 Validation and Sensitiv"ity Analyses for Dynamic 

Models............................................... 3-121 

3.7 SUMMARY ..................................................... 3-124 

3.7.1 Descriptions and Properties of Ozone and Other 

Photochemical Oxidants ............................... 3-124 

3.7.2 Nature of Precursors to Ozone and Other Photochemical 

Oxidants............................................. 3-125 

3.7.3 Atmospheric Reactions of Ozone and Other Oxidants 

Including Their Role in Aerosol Formation ............ 3-127 

3.7.3.1 Formation and Transformation of Ozone 

and Other Photochemical Oxidants ............ 3-127 

3.7.3.2 Atmospheric Chemical Processes Involving 

Ozone....................................... 3-128 

vi 
TABLE OF CONTENTS 

(continued) 

Page 

3.7.3.3 Atmospheric Reactions of PAN, H 2O2 , and 

HCOOH ...... ................................ 3-129 

3.7.4 Meteorological and Climatological Processes .......... 3-130 

3.7.4.1 Atmospheric Mixing ...................... .. 3-131 

3.7.4.2 Wind Speed and Direction .................... 3-132 

3.7.4.3 Effects of Sunlight and Temperature ......... 3-133 

3.7.4.4 Transport of Ozone and Other Oxidants 

and Their Precursors ........................ 3-133 

3.7.4.5 Stratospheric-Tropospheric Ozone Exchange ... 3-135 

3.7.4.6 Stratospheric Ozone at Ground Level ......... 3-135 

3.7.4.7 Background Ozone from Photochemical 

Reactions ............................... .. 3-136 

3.7.5 Sources, Emissions, and Concentrations of Precursors 

to Ozone and Other Photochemical Oxidants ............ 3-137 

3.7.5.1 Sources and Emissions of Precursors ......... 3-137 

3.7.5.2 Representative Concentrations in Ambient 

Air ............................. ........... ........... 3-138 

3.7.6 Source-Receptor (Oxidant-Precursor) Models ~ ........... 3-140 

3.7.6.1 Trajectory Models ........................... 3-140 

3.7.6.2 Fixed-Grid Models .. ....................... 3-141 

3.7.6.3 Box Models .................................. 3-142 

3.7.6.4 Validation and Sensitivity Analyses for 

Dynamic Models .............................. 3-142 

3.8 REFERENCES .................................................. 3-142 

4. SAMPLING AND MEASUREMENT OF OZONE AND OTHER PHOTOCHEMICAL 

OXIDANTS AND THEIR PRECURSORS ..................................... 4-1 

4.1 INTRODUCTION ................ . .. ... . 

.......

.

......... 

4-1 

4.2 QUALITY ASSURANCE AND OTHER SAMPLING FACTORS IN 

MONITORING FOR OZONE .............................. ......... 4-2 

4.2.1 Quality Assurance in Ambient Air Monitoring for 

Ozone................................................. 4-2 

4.2.2 Sampling Factors in Ambient Air Monitoring for 

Ozone................................................. 4-3 

4.2.2.1 Sampling Strategies and Air Monitoring 

Needs ........................................ 4-4 

4.2.2.2 Air Monitoring Site Selection ................ 4-4 

4.2.3 Measurement Methods for Total Oxidants and Ozone....... 4-6 

4.2.3.1 Total Oxidants ............................... 4-6 

4.2.3.2 Ozone ... .... .............................. 4-7 

4.2.4 Generation and Calibration Methods for Ozone .......... 4-13 

4.2.4.1 Generation ................................... 4-13 

4.2.4.2 Calibration .................................. 4-14 

4.2.5 Relationship between Methods for Total Oxidants 

and Ozone ............................................. 4-21 

4.2.5.1 Predicted Relationship ....................... 4-22 

vii 
5. 

TABLE OF CONTENTS 

(continued) 

Pa e 

4.2.5.2 Empirical Relationship Determined from 

Simultaneous Measurements .................... 4-24 

4.2.6 Methods for Sampling and Analysis of Peroxyacetyl 

Nitrate and Its Homologues ............................ 4-33 

4.2.6.1 Introduction ................................. 4-33 

4.2.6.2 Analytical Methods for PAN ................... 4-34 

4.2.6.3 Generation and Calibration of PAN ............ 4-39 

4.2.6.4 Methods of Analysis of Higher Homologues ..... 4-42 

4.2.7 Methods for Sampling and Analysis of Hydrogen 

Peroxide .............................................. 4-43 

4.2.7.1 Introduction ................................. 4-43 

4.2.7.2 Sampling ..................................... 4-44 

4.2.7.3 Measurement . .... .......... ....... ...... 4-45 

4.2.7.4 Generation and Calibration Methods ........... 4-48 

4.3 SAMPLING, MEASUREMENT, AND CALIBRATION METHODS FOR 

PRECURSORS TO OZONE AND OTHER PHOTOCHEMICAL OXIDANTS ....:.... 4-49 

4.3.1 Nonmethane Organic Compounds .......................... 4-50 

4.3.1.1 Nonmethane Hydrocarbons ...................... 4-50 

4.3.1.2 Aldehydes .................................... 4-61 

4.3.1.3 Other Oxygenated Organic Species ............. 4-66 

4.3.2 Nitrogen Oxides ....................................... 4-66 

4.3.2.1 Measurement Methods for NO2 and NO ........... 4-67 

4.3.2.2 Sampling Requirements ........................ 4-70 

4.3.2.3 Calibration .................................. 4-70 

4.4 SUMMARY ...................................................... 4-71 

4.4.1 Sampling and Measurement of Ozone and Other 

Photochemical Oxidants ................................ 4-71 

4.4.1.1 Quality Assurance and Sampling ............... 4-72 

4.4.1.2 Measurement Methods for Total Oxidants 

and Ozone .................................... 4-72 

4.4.1.3 Calibration Methods .......................... 4-74 

4.4.1.4 Relationships of Total Oxidants and Ozone 

Measurements ................................. 4-77 

4.4.1.5 Methods for Sampling and Analysis of Peroxyacetyl Nitrate and Its Homologues .......... 4-78 

4.4.1.6 Methods for Sampling and Analysis of 

Hydrogen Peroxide ............................ 4-80 

4.4.2 Measurement of Precursors to Ozone and Other 

Photochemical Oxidants ................................ 4-82 

4.4.2.1 Nonmethane Organic Compounds ................. 4-82 

4.4.2.2 Nitrogen Oxides .............................. 4-84 

4.5 REFERENCES ................................................... 4-86 

CONCENTRATIONS OF OZONE AND OTHER PHOTOCHEMICAL OXIDANTS 

INAMBIENT AIR .................................................... 5-1 

5.1 INTRODUCTION ................................................. 5-1 

5.2 TRENDS IN NATIONWIDE OZONE CONCENTRATIONS .................... 5-3 

viii 
TABLE OF CONTENTS 

(continued) 

Page 

5.3 OVERVIEW OF OZONE CONCENTRATIONS IN URBAN AREAS .............. 5-7 

5.4 OVERVIEW OF OZONE CONCENTRATIONS IN NONURBAN AREAS ..:........ 5-16 

5.4.1 National Air Pollution Background Network (NAPBN) ..... 5-16 

5.4.2 Sulfate Regional Experiment Sites (SURE) .............. 5-25 

5.5 VARIATIONS IN OZONE CONCENTRATIONS: DATA FROM SELECTED 

URBAN AND NONURBAN SITES ..................................... 5-?5 

5.5.1 Temporal Variations in Ozone Concentrations ........... 5-28 

5.5.1.1 Diurnal Variations in Ozone Concentrations ... 5L28 

5.5.1.2 Seasonal Variations in Ozone Concentrations .. 5-42 

5.5.1.3 Weekday-Weekend Variations in Ozone 

Concentrations ............................... 5-48 

5.5.2 Spatiai Variations in Ozone Concentrations ............ 5-49 

5.5.2.1 Urban Versus Nonurban Variations ............. 5-49 

5.5.2.2 Intracity Variations ......................... 5-50 

5.5.2.3 Indoor-Outdoor Concentration Ratios .......... 5-55 

5.5.2.4 Altitudinal and Latitudinal Variations ....... 5-60 

5.5.2.5 Vertical Gradients at Ground Level ........... 5-69 

5.6 CONCENTRATIONS OF PEROXYACETYL NITRATE AND PEROXYPROPIONYL 

NITRATE IN AMBIENT AIR ....................................... 5- 71 

5.6.1 Introduction .......................................... 5-71 

5.6.2 Historical Data ....................................... 5-72 

5.6.3 Ambient Air Concentrations of PAN and Its 

Homologues in Urban Areas ............................. 5-74 

5.6.4 Ambient Air Concentrations of PAN and Its 

Homologues in Nonurban Areas .................. ...... 5-, 80 

5.6.5 Temporal Variations in Ambient Air 

Concentrations of Peroxyacetyl Nitrate ................ 5-83 

5.6.5.1 Diurnal Patterns ............................. 5-83 

5.6.5.2 Seasonal Patterns ................. ........ 5-89 

5.6.6 Spatial Variations in Ambient Air Concentrations 

of Peroxyacetyl Nitrate ............................... 5-89 

5.6.6.1 Urban-Rural Gradients and Transport of PAN ... 5-89 

5.6.6.2 Intracity Variations ......................... 5-93 

5.6.6.3 Indoor-Outdoor Ratios of PAN Concentrations .. 5-93 

5.7 CONCENTRATIONS OF OTHER PHOTOCHEMICAL OXIDANTS IN 

AMBIENT AIR .................................................. 5=93 

5.8 SUMMARY ...................................................... 5-95 

5.8.1 Ozone Concentrations in Urban Areas ................... 5-98 

5.8.2 Trends in Nationwide Ozone Concentrations ............: 5-101 

5.8.3 Ozone Concentrations in Nonurban Areas ................ 5- 103 

5.8.4 Diurnal and Seasonal Patterns in Ozone 

Concentrations ........................:............... 5-105 

5.8.5 Spatial Patterns in Ozone Concentrations .............. 5T107 

5.8.5.1 Urban-Nonurban Differences in Ozone 

Concentrations ............................... 5-107 

5.8.5.2 Geographic, Vertical, and Altitudinal 

Variations in Ozone Concentrations ........... 5-109 

ix 
TABLE OF CONTENTS 

(continued) 

Pa e 

5.8.5.3 Other Spatial Variations in Ozone 

Concentrations ............................... 5-112 

5.8.6 Concentrations and Patterns of Other Photochemical 

Oxidants ..................................:........... 5-112 

5.8.6.1 Concentrations ............................... 5-112 

5.8.6.2 Patterns ..................................... 5-114 

5.8.7 Relationship Between Ozone and Other Photochemical 

Oxidants.............................................. 5-114 

5.9 REFERENCES ........................................................ 5-116 

I 

x 
LIST OF TABLES 

Table Pa e 

3-1 Physical properties of ozone .................................. 3-3 

3-2 Physical properties of peroxyacetyl nitrate ................... 3-5 

3-3 Infrared absorptivities of peroxyacetyl nitrate ............... 3-6 

3-4 Physical prope,rties of hydrogen peroxide ........................ 3-6 

3-5 Physical and chemical properties of nitric oxide and 

nitrogen dioxide ..................... 

I ............. ......... 3-15 

3-6 Calculated lifetimes of selected organic compounds resulting 

from atmospheric loss by reaction with 0 3 and OH and NO3 

radicals .................................... .... ........ . 3-44 

3-7 Published episodes of transport of stratospheric ozone to 

ground level .................................................. 3-82 

3-8 Emissions of VOC by decade, 1940 through 1980 ................. 3-90 

3-9 Emissions of NO by decade, 1940 through 1980 ................. 3-93 

3-10 Yearly quantities of motor vehicle fuels sold in the 

United States for highway use, 1980 through 1983 .............. 3-97 

3-11 Summary of NO emissions from mobile sources ................... 3-98 

3-12 Area-wide bio,enic emission.fluxes ................ :........... 3-101 

3-13 Global estimates of nitrogen transformation ................... 3-103 

3-14 Nonmethane hydrocarbon concentrations measured between 6:00 

and 9:00 a.m. in various United States cities ................. 3-106 

3-15 Nonmethane hydrocarbon concentrations measured in nonurban 

atmospheres................................................... 3-107 

3-16 Average 6:00 to 9:00 a.m. NOxconcentrations and HC/NO x 

ratios in urban areas ......................................... 3-109 

4-1 Performance specifications for automated methods of ozone 

analysis...................................................... 4-9 

4-2 List of designated reference and equivalent methods of 

ozoneanalysis ................................................ 4-10 

4-3 Factors for intercomparison of data caiibrated by UV , photometry versus KI colorimetry .......... . .................. 4-16 

4-4 Response of NBKI reagent and Mast meter to various 

oxidants...................................................... 4-23 

4-5 Comparison of corrected instrument readings to colorimetric 

oxidant readings during atmospheric sampiing .................. 4-27 

4-6 Summary of parameters used in determination of PAN by 

GC-ECD ........................................................ 4-35 

4-7 Infrared absorptivities of peroxyacetyl nitrate 

(Base 10) ............................. ....................... 4-38 

4-8 Measurement methods for hydrogen peroxide ..................... 4-46 

4-9 Percentage difference from known concentrations of 

nonmethane hydrocarbons obtained by sixteen users ............. 4-52 

4-10 Problems associated with gathering NMOC data with 

automated analyzers and recommendations for reducing 

these effects ................................................. 4-53 

4-11 Summary of advantages and disadvantages of primary 

collection media for NMOC analysis ............................ 4-58 

xi 
LIST OF TABLES 

(continued) 

Table Page 

4-12 GC/continuous NMOC analyzer comparisons, least-squares 

regressions ........ ..... ..... ...................... ..... 4-62 

4-13 Summary of ozone monitoring techniques ......................... 4-73 

4-14 Ozone calibration techniques ................................... 4-76 

5-1 Second-highest 1-hr ozone concentrations reported for 

Standard Metropolitan Statistical Areas having 

populations > 0.5 million, 1981 through 1983 ................... 5-13 

5-2 Annual ozone summary statistics for three si.tes of the 

National Air Pollution Background Network ...................... 5-18 

5-3 Concentrations of ozone during 6-day period of high 

values at NAPBN site in Mark Twain National Forest, 

Missouri, 1979 . ... ....................... .................. 5-20 

5-4 Percentile distributions of ozone concentrations at 

sites of National Air Pollution Background Network, 

aggregated by quarter across severa1 years ..................... 5-22 

5-5 Summary of ozone concentrations measured at Sulfate 

Regional Experiment (SURE) nonurban stations, August 

through December 1977 .......................................... 5-27 

5-6 Number of consecutive-day exposures or respites when,the 

daily 1-hr maximum ozone concentration was > 0.06 ppm, 

in four cities (April through September, 1979 through 1981) .... 5-38 

5-7 Number of consecutive-day exposures or respites when the 

daily 1-hr maximum ozone concentration was > 0.12 ppm, in 

four cities (April through September, 1979 through 1981) ....... 5-39 

5-8 Number of consecutive-day exposures or respites when the 

daily 1-hr maximum ozone concentration was > 0.18 ppm, in 

four cities (April through September, 1979 through 1981) ....... 5-40 

5-9 Number of consecutive-day exposures or respites when the 

daily 1-hr maximum ozone concentration was > 0.24 ppm, in 

four cities (April through September, 1979 through 1981) ....... 5-41 

5-10 Ozone concentrations at sites in and around New Haven, 

Connecticut, 1976 ............................................... 5-52 

5-11 Quarterly maximum 1-hour ozone values at sites in and 

around New Haven, Connecticut, 1976 ............................ 5-52 

5-12 Peak ozone concentrations at eight sites in New York 

City and adjacent Nassau County, 1980 .......................... 5-53 

5-13 Summary of reported indoor-outdoor ozone ratios ................ 5-57 

5-14 Comparison of ozone concentrations at three different 

elevations, High Point Mountain, NJ, and at Bayonne, NJ, 

July 1975 ....... .......... ............................. ..... 5-68 

5-15 Means and standard errors of ozone concentrations measured 

over 4 years at two sampling heights at three stations in 

the rural, upper-midwestern United States ..................... 5-70 

5-16 Summary of concentrations of peroxyacetyl nitrate in 

ambient air in urban areas of the United States ................ 5-75 

xii 
LIST OF TABLES 

(continued) 

Table Page 

5-17 Relationship of ozone and peroxyacetyl nitrate at urban 

and suburban sites in the United States ............ ........ 5-79 

5-18 Ambient air measurements of peroxypropionyl nitrate (PPN) 

concentrations by electron capture gas chromatography 

at urban sites in the United States ............................ 5-81 

5-19 Concentrations of peroxyacetyl and peroxypropionyl nitrates 

in Los Angeles, Oakland, and Phoenix, 1979 ..................... 5-82 

5-20 Concentrations in ambient air of peroxyacetyl and 

peroxypropionyl nitrates and ozone at nonurban remote sites 

in the United States ........................................... 5-84 

5-21 PAN and ozone concentrations in ambient air, New 

Brunswick, N.J., for September 25, 1978, to May 10, 1980 ..,.... 5-91 

5-22 Intracity variations in peak ozone and PAN concentrations 

in Houston, October 26 and 27, 1977 .......................... 5-94 

5-23 Concentrations of hydrogen peroxide in ambient air at 

urban and nonurban sites ....................................... 5-97 

5-24 Second-highest ozone concentrations among daily maximum 

1-hr values in 1983 in Standard Metropolitan Statistical 

Areas with populations > 1 million, given by census divisions 

andregions .................................................... 5-99 

xiii 
LIST OF FIGURES 

Fi ure Pac,~e 

3-1 Experimental time-concentration profiles for propene, N0, 

NO2i 03 , HCHO, and PAN,for an irradiated NO 

x

-propene-air 

mixture ... .... ......... ...... .. ................. ....... 3-18 

3-2 Reaction scheme for OH radical-initiated oxidation of propene 

in the presence of•NO .............. .. ........ ...... ....... 3-28 

3-3 Rate of aqueous-phase oxidation of S(IV) by 0 3 (30 ppb) and 

H2 O2 (1 ppb), as a function of solution pH ...................... 3-52 

3-4 Isopleths (m x 10 2 ) of inean summer morning mixing heights, AGL .. 3-59 

3-5 Isopleths (m x 102 ) of inean summer afternoon mixing 

heights, AGL .......... . .. ............... _.... . . 

~ 

3-59 

3-6 Percentage of summer 2315 GMT (6:15 p.m. EST, 3:15 p.m. PST) 

soundings withzan elevated inversion base between 1 and 

500 m above ground level ........................................ 3-60 

3-7 Mean resultant surface wind pattern for the United States 

for July.c Direction and length of arrows indicate monthly 

resultant wind...... .. . .. . .. ....."............ 3-62 

3-8 Percentage of summer 1115 GMT ~ (6:15 a.m..EST, 3:15 a.m. PST) 

soundings with an inversion base at the .surface and wind 

speeds at the surface <2.5 m%sec . .............. ........... 3-64 

3-9 Isopleths (m/sec) of inean summer wind speedaveraged 

through the morning mixing layer ....... ..:. ........... ... 3-65 

3-10 Isopleths (m/sec) of inean summer wind speed averaged 

through the.afternoon mixing layer . ... `...... .....

.

... 3-65 

- -3-11 Schematic.cross section, looking downwindalong the jet 

stream, of a tr.opopause folding event as modeled by 

Danielsen ................... ...... . .... ..... ........ 3-76 

3-12 Measured vertical cross sections of (A) 0 3 i (B) dewpoint, and 

(C) the 500 mb chart and the flight track for October 5, 1978 ... 3-78 

3-13 Hypothesized models of the process that mixes tropopause 

folding events into.the troposphere :... . . ...... . ....... 3-79 

3-14 National trend in estimated em issions of volatile organic' 

compounds, 1970 through 1983 ... .............. ...........:.. 3-91 

3-15 Comparative trends in highway vehicle emissions of nitrogen 

oxides (NO ) and.volatile organic compounds (VOC) versus 

vehicle mifes traveled, 1970-1983 ......... ........ .......... 3-92 

3-16 National-trend,in estimated emissions of nitrogen,oxides, 

1970 through 1983 ......................... ... ................... 3-94 

3-17 Schematics of the_three types.of dynamic models ................. 3-114 

' 3-18 Example of EKMA diagram for high-oxidant urban area ............. 3-116 

~ 4-1 Comparison of ozone and total oxidant concentrations in the 

Pasadena area, August 1955 .............. ... ... .... . 4-26 

r 4-2 Comparison of ozone and total oxidant concentrations inthe 

Los Angeles area, August 1955 ................... ............... 4-26 

4=3 Measurements of ozone and total oxidants in Los Angeles, 

September 4 through September 30, 1971 ..... .... ............. 4-29 

4-4 Measurements of ozone and total oxidants in St. Louis, 

October 14 through December 21, 1971 ......................... . 4-30 

xiv 
LIST 0F FIGUR.ES 

(continued) 

Figure Page, 

4-5 Meas,urements of ozone and total oxidants, Houston Ship 

Channel,,August 11, 1973 . . .:.. ............... ................ 4-32 

5-1 National trend in composite average of the second-highest 

value among daily maximum 1=hour concentrations ~ at 

selected groups of sites, 1975 through 1983 .................. . 5-4 

5-2 Comparison of the 1979-1980; 1981=1982, and 1983 

composite average of the second-highest daily'maximum 

1-hour ozone,concentrations across,EPA regions .................. 5-8 

5-3 Average daylight (6:00 a.m. to 8:00 p.m.) concentrations 

of ozone in the second and third quarters (April thr.ough 

September), 1981.... .......... .. .........'............ 5-9 

5-4 Average daylight.(6:00 a.m. to 8:00 p.m.) concentrations 

of ozone ,•"in the first and fourth quarters '(January through 

March and October through December), , 1981 .................. 5-10 

5-5 Distribution's of the three highest 1-hour"ozone 

concentrations at'valid sites (906 station-years) 

aggregated for 3 years (1979, 1980,'and 1981) and 

, the highest ozone concentrations at NAPBN sites aggre- 

"" Locations oftheyeagh t Cnati"orial

o

forests (NF) stations 

5-12 

56 -- g o ~~ ~~~~ 

constituting the.Nationa7 Air Po1lution Background Network 

(NAPBN) .......... .:.:. ........... ............. ....... 5-17 

5-7 Trajectory analysi`s plots for the NAPBN site at Mark Twain 

National Forest,.MQ, July 21, 1979. . ... . ...... ....... .: 

~ 

5-21 

5-8 (A) Second-highe'st value among maximuml-hrozone` 

conceritrations at five NAPBN monitoring stations, 1979 

through 1983. (B) Composite averages-of the second-highest , 

` value among daily maximum 1-hr ozone concentrations at 

five NAPBN stations, 1979 through 1983 ........................ ......... 5-24 

5-9 Location of;Sulfate Regional Experiment (SURE) monitoring 

stations .:. ....... .... ....... :.. ......:.... .......... 5-26 

> 5710 Diurnal pattei^n of 1-hr ozone concentrations on July 13, 

1979, Philadelphia, PA ............ ......'.... . ..........:.. 5-29 

5-11 Diurnal patterns of 1-hour ozone concentrations, ~ September 20 

and 21, 1980, Detroit, MI .... ................................. 5-30 

5-12 Diurnal and 1-month composi'te diurnal variations in ozone . 

concentrations; Washington, DC,"July 1981 .. .........:.... ... 5-31 

5-13 Diurnal and 1-month composite diurnal variations in ozone 

concentrations', St: Louis County, MO ; September 1981 ............ 5-31 

5-14 Diurnal and 17month composite diurnal variations , ~in-ozone 

concentrations, Alto'n, IL, October'1981 (fourth quarter) ........ 5-32 

5-15 Composite diurnal,patterns by quarter'of ozone 

concentrations',' A1ton, IL, 1981 :... ..... ........ ........ 5-33 

5-16 'Three-day sequence of hourly ozone concentrations at 

Montague, MA, SURE"station``showing locally generated 

midday peaks and transported late.peaks ............... ......... 5-36 

xv 
LIST OF FIGURES 

(continued) 

Fi c.ure Pa.ge 

5-17 Quarterly composite diurnal patterns of ozone concentrations 

at selected sites representing potential for exposure of 

major crops, 1981 ................................ .. ......... 5-43 

5-18 Composite diurnal ozone pattern at a rural NCLAN site in 

Argonne, IL, August 6 through September 30, 1980 ................ 5-45 

5-19 Daily 7-hour and 24-hour average ozone concentrations at 

a rural NCLAN site in Argonne, IL, 1980 ........... .......... . 5-45 

5-20 Seasonal variations in ozone concentrations as indicated 

by monthly averages and the 1-hour maximum in each month 

at selected sites, 1981 .. ... ................................ 5-46 

5-21 New York State air monitoring sites for Northeast Corridor 

Monitoring Program (NECRMP) ........ ... .................... 

. 

5-54 

5-22 Altitudinal sequence of monitoring sites in the San Bernardino 

Mountains ......... .... ........ ............. ............ 5-64 

5-23 Relationshipbetween elevation and diurnal patterns of total 

oxidant concentrations, temperature, and vapor pressure at 

four sites (A-D) in the San Bernardino Mountains, CA, JulyAugust 1969 ..................................................... 5-66 

5-24 Comparison of monthly daylight average and maximum PAN 

concentrations at Riverside, CA, for 1967-1968 and 1980 ......... 5-78 

5-25 Variation of inean 1-hour oxidant and PAN concentrations, 

by hour of day, in downtown Los Angeles, 1965 ................... 5-86 

5-26 Variation of inean 1-hour oxidant and PAN concentrations, by 

hour of day, Air Pollution Research Center, Riverside CA, 

September 1966 .................................................. 5-87 

5-27 Diurnal profiles of ozone and PAN at Claremont, CA, 

October 12 and 13, 1978, 2 days of a multi-day smog episode ..... 5-88 

5-28 Monthly variation of oxidant (Mast meter, continuous 24-hr) 

concentrations and PAN (GC-ECD, sequential, 6:00 a.m. to 

4:00-5:00 p.m.) concentrations, Air Pollution Research 

Center, Riverside, CA, June 1966-June 1967 ...................... 5-90 

5-29 Average daily profile by month (July 7-October 10) for 

PAN and ozone in New Brunswick, NJ, 1979 ........................ 5-92 

5-30 Diurnal profile of HCOOH, along with other oxidants and 

smog constituents, on October 12 and 13, 1978, at 

Claremont, CA .......... ........................... ............ 5-96 

5-31 National trend in composite average of the second highest 

value among daily maximum 1-hour ozone concentrations at 

selected groups of sites, 1975 through 1983 ................... 5-102 

5-32 Distributions of the three highest 1-hour ozone 

concentrations at valid sites (906 station-years) aggregated 

for 3 years (1979, 1980, and 1981) and the highest ozone 

concentrations at NAPBN sites aggregated for those years 

(24 station-years) .............................................. 5-104 

xvi 
LIST OF ABBREVIATIONS AND SYMBOLS 

~ approximately 

~ wavelength 

AFR air:fuel ratio 

APHA American Public Health Association 

aq aqueous 

AGL abov.e ground level 

ASL above sea level 

atm atmosphere 

avg average 

b.p. boiling point 

bz benzene 

C carbon 

°C degrees Celsius 

CA chromotropic acid 

CAMP Continuous Air Monitoring Program 

CARB California Air Resources Board 

cc cubic centimeter 

CH4 methane 

CO carbon monoxide 

CO2 carbon dioxide 

cm centimeter 

concn concentration 

DBH tree diameter at breast height 

DNPH 2,4-dinitrophenylhydrazine 

DOT Department of Transportation 

E6 normal electrode potential 

ECD electron-capture detector 

EKMA Empirical Kinetic Modeling Approach 

EPA U.S. Environmental Protection Agency 

EST eastern standard time 

FID flame ionization detector 

FRM Federal Reference Method 

ft foot 

xvii 
LIST OF ABBREVIATIONS AND SYMBOLS 

(continued) 

FTIR Fourier-transform infrared 

g gram(s), 

g/mi grams per mile 

GC gas chromatography 

GMT Greenwich mean time 

GPT gas-phase titration 

hr hour(s) 

hv photon 

HC hydrocarbons 

HCN hydrogen cyanide 

HCOOH formic acid 

HFET Highway Fuel Economy Driving Schedule 

Hg mercury 

H2O2 hydrogen peroxide 

HOZ hydroperoxy 

HONO nitrous acid 

HONO2 nitric acid 

HPLC high-pressure liquid chromatography; also, 

high-performance liquid chromatography 

HPPA 3-(p-hydroxyphenyl)propionic acid 

HRP horseradish peroxidase 

NZO water 

H2SO4 sulfuric acid 

in inch(es) 

IR infrared 

k constant 

KI potassium iodide 

km kilometer 

L liter.(s) 

LAAPCD Los Angeles Air Pollution Control District 

LCV leuco crystal violet 

ln natural logarithm (base e) 

LST local standard time 

xviii 
LIST OF ABBREVIATIONS AND SYMBOLS 

(continued) 

M molar 

m meter(s) 

mb millibar(s) 

MBTH 3-methyl-2-benzothiazolinone hydrazone 

mg milligram(s) 

mg/m3 milligrams per cubic meter 

MGE modified graphite electrode 

min minute(s) 

ml milliliter(s) 

mm millimeter(s) 

mM millimolar 

MMC mean meridional circulation 

M.P. melting point 

mph miles per hour 

MS mass spectrometry 

MSL mean sea level 

MTBE methyl tertiary butyl ether 

NA not available 

NAAQS National Ambient Air Quality Standard 

NADB National Aerometric Data Bank 

NAMS National Aerometric Monitoring Stations 

NAPBN National Air Pollution Background Network 

NAS National Academy of Sciences 

NBS National Bureau of Standards 

NECRMP Northeast Corridor Regional Modeling Project 

NEDS National Emissions Data System 

NEROS Northeast Regional Oxidant Study 

NH3 ammonia 

NH4NO3 ammonium nitrate 

NF National Forest 

mm nanometer(s) 

NMHC nonmethane hydrocarbons 

NMOC nonmethane organic compounds 

xix 
LIST OF ABBREVIATIONS AND SYMBOLS 

(continued) 

N0 nitric oxide 

NOx nitrogen oxides 

NO2 nitrogen dioxide 

NO3 nitrogen trioxide 

NZO nitrous oxide 

NR natural rubber 

NYCC New York City Driving Schedule 

02 oxygen 

03 ozone 

PAN peroxyacetyl nitrate 

PBzN peroxybenzoyl nitrate 

PNA peroxynitric acid 

PPN peroxypropionyl nitrate 

ppb parts per billion 

ppm parts per million 

ppt parts per trillion 

PSD Prevention of Significant Deterioration 

psig pounds per square inch gauge 

PST Pacific standard time 

PUFA polyunsaturated fatty acids 

RAPS Regional Air Pollution Study 

RTI Research Triangle Institute 

S.D. standard deviation 

SAROAD Storage and Retrieval of Aerometric Data 

SBR styrene-butadiene rubber 

SCAB South Coast Air Basin 

sec second(s) 

SLAMS State and Local Air Monitoring Stations 

SMSA Standard Metropolitan Statistical Area 

SRM Standard Reference Material 

SSET small-scale eddy transport 

STA seasonal tropopause adjustmeni; 

STP standard temperature and pressure 

xx 
LIST OF ABBREVIATIONS AND SYMBOLS 

(continued) 

SURE Sulfate Regional Experiment Sites 

TEL tetraethyl lead 

Tenax GC adsorbent used in NMOC analysis 

TF tropopause-folding events 

Tg/yr teragrams per year 

THC total hydrocarbon 

TML tetramethyl lead 

TNMHC total nonmethane hydrocarbons 

TWC three-way catalyst 

pg/m3 microgram(s) per cubic meter 

pM micromolar 

U uranium 

UHAC uranium hydroxamic acid chelates 

U.S. United States 

UV ultraviolet 

V vanadium 

v/v volume-volume 

VHAC vanadium hydroxamic acid chelates 

VOC volatile.organic compounds 

vol % volume percent 

w/w weight-weight 

WCOT wall-coated open tubular (column) 

XAD-2 absorbent used in NMOC analysis 

XO xylenol orange 

yr year(s) 

xXi 
AUTHORS, CONTRIBUTORS, AND REVIEWERS 

Chapter 3: 

Principal Authors 

*Mr. Thomas B. McMullen 

Environmental Criteria and Assessment Office 

MD-52 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

*Mr. Elmer Robinson 

Director, Mauna Loa Observatory 

National Oceanic and Atmospheric Administration (NOAA/CMCC) 

Hilo, HI 96720 

*Ms. Beverly E. Tilton 

Environmental Criteria and Assessment Office 

MD-52 

Environmental Protection Agency 

Research Triangle Park, NC 27711 

Dr. Halvor Westberg 

Director, Laboratory for Atmospheric Research, and 

Professor, Civil and Environmental Engineering 

Washington State University 

Pullman, WA 99164-2730 

*Dr. Arthur M. Winer 

Assistant Director 

Statewide Air Pollution Research Center 

University of California 

Riverside, CA 92521 

Contributinq Authors 

*Dr. A. Paul Altshuller 

Atmospheric Sciences Research Laboratory 

MD-59 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Dr. Basil Dimitriades 

Atmospheric Sciences Research Laboratory 

MD-59 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

xxii 
Chapter 3 Contributing Authors.(cont'd.) 

*Dr. Marcia C. Dodge 

Atmospheric Sciences Research Laboratory 

MD-84 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Mr. James M. Kawecki 

TRC Environmental Consultants, Inc. 

2001 Wisconsin Avenue, N.W. 

Suite 261 

Washington, DC 20007 

*Dr. Harold G. Richter 

Office of Air Quality Planning and Standards 

Monitoring and.Data Analysis Division 

MD-14 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

The following people reviewed Chapter 3 at the request of.EPA: 

Dr. Joseph J. Bufalini 

Atmospheric Sciences Research Laboratory 

MD-84 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Dr. Donald Fox 

Associate Professor 

Department of Environmental Science and Engineering 

School of Public Health, 201H 

University of North Carolina 

Chapel Hill, NC 27514 

Mr. Bruce Gay 

Atmospheric Sciences Research Laboratory 

MD-84 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Mr. Eric Ginsburg 

Office of Air Quality Planning and Standards 

Control Programs Development Division 

MD-15 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

xxiii 
Chapter 3 Reviewers (cont'd.) 

Mr. Gerald Gipson ` 

Office of Air Quality Planning and Standards 

Monitoring and Data Analysis Division 

MD-14 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Mr. Robert Hall 

Industrial Environmental Research Laboratory 

MD-65 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Dr. Michael R. Kuhlman 

Battelle, Columbus Laboratories 

505 King Avenue 

Columbus, OH 43201 

Mr. William A. Lonneman 

Atmospheric Sciences Research Laboratory 

MD-84 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Mr. Chuck Mann 

Office of Air Quality Planning and Standards 

Monitoring and Data Analysis Division 

MD-14 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Mr. E. L. Marti nez 

Office of Air Quality Planning and Standards 

Monitoring and Data Analysis Division 

MD-14 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Dr. Edwi n L. Meyer 

Office of Air Quality Planning and Standards 

Monitoring and Data Analysis Division 

MD-14 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Mr. Johnnie Pearson 

Office of Air Quality Planning and Standards 

Monitoring and Data Analysis Division 

MD-14 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

xxiv 
Chapter 3 Reviewers (cont'd.) 

Mr. Kenneth L. Schere 

Atmospheric Sciences Research Laboratory 

MD-80 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Mr. Stanley Sleva 

Office of Air Quality Planning and Standards 

Monitoring and Data Analysis Divsion 

MD-14 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Dr. Chester W. Spicer 

Battelle, Columbus Laboratories 

505 King Avenue 

Columbus, OH 43201 

Mr. Bruce Tichenor 

Industrial Environmental Research Laboratory 

MD-54 

U.S. Environmental Protection Agency 

Researcti Triangle Park, NC 27711 

*Authors also reviewed portions of this chapter. 

xxv 
Chapter 4 : Measurement of Ozone and Other Photochemical Oxidants and Their 

Precursors 

Principal Authors 

*Dr. Jimmie A. Hodgeson 

U.S. Environmental Protection Agency 

Environmental Monitoring Systems Laboratory 

26 W. St. Clair 

Cincinnati, OH 45268) 

*Mr. Michael W. - Holdren 

Battelle, Columbus Laboratories 

505 King Avenue 

Columbus, OH 43201 

Dr. M. Rene Surgi 

Department of Chemistry 

Louisiana State University 

Baton Rouge, LA 70803 

Contributing Authors 

Dr. Sandor Freedman 

Piedmont Technical Services 

Hillsborough, NC 27278 

*Ms. Beverly E. Tilton 

Environmental Criteria and Assessment Office 

MD-52 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

The followinq people reviewed Chapter 4 at the request of EPA : 

Dr. A. Paul Altshuller 

Atmospheric Sciences Research Laboratory 

MD-59 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Dr. Joseph J. Bufalini 

Atmospheric Sciences Resear•ch Laboratory 

MD-84 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

xxvi 
Chapter 4 Reviewers (cont'd) : 

Dr. Donald Fox 

Associate Professor 

Department of Environmental Science and Engineering 

School of Public Health, 201H 

University of North Carolina 

Chapel Hill, NC 27514 

Mr. Bruce Gay 

Atmospheric Sciences Research Laboratory 

MD-84 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Mr. William A. Lonneman 

Atmospheric Sciences Research Laboratory 

MD-84 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Mr. Kenneth Rehme 

Environmentai Monitoring Systems Laboratory 

MD-77 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Dr. Harold G. Richter 

Office of Air Quality Planning and Standards 

Monitoring and Data Analysis Division 

MD-14 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Mr. Stanley Sleva 

Office of Air Quality Planning and Standards 

Monitoring and Data Analysis Division 

MD-14 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Dr. Chester W. Spicer 

Battelle, Columbus Laboratories 

505 King Avenue 

Columbus, OH 43201 

Dr. Arthur M. Winer 

Assistant Director 

Statewide Air Pollution Research Center 

University of California 

Riverside, CA 92521 

*Authors also reviewed portions of this chapter. 

xxvii 
Chapter 5 : Concentrations of Ozone and Other Photochemical Oxidants in Ambient 

Air 

Principal Authors 

*Mr. Thomas B. McMullen 

Environmental Criteria and Assessment Office 

MD-52 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

*Mr. Elmer Robinson 

Director, Mauna Loa Observatory 

National Oceanic and Atmospheric Administration (NOAA/CMCC) 

Hilo, HI 96720 

*Ms. Beverly E. Tilton 

Environmental Criteria and Assessment Office 

MD-52 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Contributing Author 

Dr. Sandor Freedman 

Piedmont Technical Services 

Hillsborough, NC 27278 

The following people reviewed Chapter 5 at the request of EP A: 

Mr. Gerald Akland 

Environmental Monitoring Systems Laboratory 

MD-56 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Dr. Joseph J. Bufalini 

Environmental Sciences Research Laboratory 

MD-84 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Mr. Gary Evans 

Environmental Monitoring Systems Laboratory 

MD-56 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

xxviii 
Chapter 5 Reviewers (cont'd) : 

Dr. Donald Fox 

Associate Professor 

Department of Environmental Science and Engineering 

Schooi of Public Health, 201H 

University of North Carolina 

Chapei Hill, NC 27514 

Mr. Gerald Gipson 

Office of Air Quality Planning and Standards 

Monitoring and Data Analysis Division 

MD-14 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Dr. Jimmie A. Hodgeson 

Professor, Department of Chemistry 

407 Choppin Hail 

Louisiana State Un7versity 

Baton Rouge, LA 70803 

Mr. Michael W. Holdren 

Battelle, Columbus Laboratories 

505 King Avenue 

Columbus, OH 43201 

Dr. Michael R. Kuhlman 

Battelle, Columbus Laboratories 

505 King Avenue 

Columbus, OH 43201 

Mr. William A. Lonneman 

Atmospheric Sciences Research Laboratory 

MD-84 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Mr. Thomas McCurdy 

Office of Air Quality Planning and Standards 

Strategies and Air Standards Division 

MD-12 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Dr. Edwin L. Meyer 

Office of Air Quality Planning and Standards 

Monitoring and Data Analysis Division 

MD-14 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

xxix 
Chapter 5 Reviewers (cont'd) : 

Dr. Harold G. Richter 

Office of Air Quality Planning and Standards 

Monitoring and Data Analysis Division 

MD-14 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Dr. Chester W. Spicer 

Battelle, Columbus Laboratories 

505 King Avenue 

Columbus, OH 43201 

Dr. Arthur M. Winer 

Assistant Director 

Statewide Air Pollution Research Center 

University of California 

Riverside, CA 92521 

*Authors also reviewed portions of this chapter. 

xxx 
SCIENCE ADVISORY BOARD 

CLEAN AIR SCIENTIFIC ADVISORY COMMITTEE 

The substance of this document was reviewed by the Clean Air Scientific 

Advisory'Committee of the Science Advisory Board.in public sessions. 

SUBCOMMITTEE ON OZONE 

Chairman 

Dr. Morton Lippmann 

Professor 

Department of Environmental Medicine 

New York University Medical Center „ 

Tuxedo, New York 10987 

Members 

Dr. Mary 0. Amdur 

Senior Research Scientist 

Energy Laboratory 

Massachusetts Institute of Technology 

Cambridge, Massachusetts 02139 

Dr. Eileen G. Brennan 

Professor 

Department of Plant Pathology 

Martin Hall, Room 213, Lipman Drive 

Cook College-NJAES 

Rutgers University 

New Brunswick, New Jersey 08903 

Dr. Edward D. Crandall 

Professor of Medicine 

School of Medicine 

Cornell University 

New York, New York 10021 

Dr. James D. Crapo 

Associate Professor of Medicine 

Chief, Division of Allergy, Critical 

Care and Respiratory Medicine 

Duke University Medical Center 

Durham, North Carolina 27710 

Dr. Robert Frank 

Professor of Environmental Health 

Sciences 

Johns Hopkins School of Hyg.iene 

and Public Health 

615 N. Wolfe Street 

Baltimore, Maryland 21205 

Professor A. Myrick Freeman II 

Department of Economics 

Bowdoin College 

Brunswick, Maine 04011 

Dr. Ronald J. Hall 

Senior Research Scientist and Leader 

Aquatic and Terrestrial Ecosystems 

Section 

Ontario Ministry of the Environment 

Dorset Research Center 

Dorset, Ontario 

Canada POAlEO 

Dr. Jay S. Jacobson 

Plant Physiologist 

Boyce Thompson Institute 

Tower Road 

Ithaca, New York 14853 

xxxi 
Or. Warren B. Johnson 

Director, Atmospheric Science Center 

SRI International ` 

333 Ravenswood Avenue 

Menlo Park, California 94025 

Dr. Jane Q. Koeni g 

Research Associate Professor 

Department of Environmental Health 

University of Washington 

Seattle, Washington 98195 

Dr. Paul Kotin 

Adjunct Professor of Pathology 

University of Colorado Medical School 

4505 S. Yosemite, #339 

Denver, Colorado 80237 

Dr. Timothy Larson 

Associate Professor 

Environmental Engineering and 

Science Program 

Department of Civil Engineering 

University of Washington 

Seattle, Washington 98195 

Professor M. Granger Morgan 

Head, Department of Engineering 

and Public Policy 

Carnegie-Mellon University 

Pittsburgh, Pennsylvania 15253 

Dr. D. Warner North 

Principal 

Decision Focus Inc., Los Altos 

Office Center, Suite 200 

4984 E1 Camino Real 

Los Altos, California 94022 

Dr. Robert D. Rowe 

Vice President, Environmental and 

Resource Economics 

Energy and Resources Consultants, Inc. 

207 Canyon Boulevard 

Boulder, Colorado 80302 

Dr. George Taylor 

Environmental Sciences Division 

P.O. Box X 

Oak Ridge National Laboratory 

Oak Ridge, Tennessee 37831 

Dr. Michael Treshow 

Professor 

Department of Biology 

University of Utah 

Salt Lake City, Utah 84112 

Dr. Mark J. Utell 

Co-Director, Pulmonary Disease Unit 

Associate Professor of Medicine and 

Toxicology in Radiation Biology 

and Biophysics 

University of Rochester Medical 

Center 

Rochester, New York 14642 

Dr. James H. Ware 

Associ.ate Professor 

Harvard School of Public Health 

Department of Biostatistics 

677 Huntington Avenue 

Boston, Massachusetts 02115 

Dr. Jerry Wesolowski 

Air and Industrial Hygiene Laboratory 

California Department of Health 

2151 Berkeley Way 

Berkeley, California 94704 

Dr. James L. Whittenberger 

Director, University of California 

Southern Occupational Health Center 

Professor and Chair, Department of 

Community and Environmental Medicine 

California College of Medicine 

University of California - Irvine 

19772 MacArthur Boulevard 

Irvine, California 92717 

Dr. George T. Wolff 

Senior Staff Research Scientist 

General Motors Research Labs 

Environmental Science Department 

Warren, Michigan 48090 

xxxii 
PROJECT TEAM FOR DEVELOPMENT 

OF 

Air Quality Criteria for Ozone and Other Photochemical Oxidants 

Ms. Beverly E. Tilton, Project Manager 

and Coordinator for Chapters 1 through 5, Volumes I and II 

Environmental Criteria and Assessment Office (MD-52) 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Mr. Norman E. Childs 

Environmental Criteria and Assessment Office (MD-52) 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Dr. J.H.B. Garner 

Coordinator for Chapters 7 and 8, Volume III 

Environmental Criteria and Assessment Office (MD-52) 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Mr: Thomas B. McMullen 

Environmental Criteria and Assessment Office (MD-52) 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Mr. James A. Raub 

Coordinator for Chapters 9 through 12, Volumes IV and V 

Environmental Criteria and Assessment Office (MD-52) 

U.S. Environmental Protection Agency 

Research Triangle Park, NC 27711 

Dr. David T. Tingey 

Coordinator for Chapter 6, Volume III 

Environmental Research Laboratory 

U.S. Environmental Protection Agency 

200 S.W. 35th Street 

Corvallis, OR 97330 

xxxi i i 

2. INTRODUCTION 

2.1 PURPOSE AND LEGISLATIVE BASIS OF THIS DOCUMENT 

The Clean Air Act specifies that the Administrator of the United States 

Environmental Protection Agency (EPA) issue, and revise on a periodic basis, 

air quality criteria for certain air pollutants. Air quality.criteria may be 

defined as qualitative and quantitative information that describes the effects 

of a pollutant on public health and welfare in terms of the respective exposures 

that elicited them. According to section 108 of the Clean Air Act, as amended 

in 1977, criteria shall 

...accurately reflect the latest scientific knowledge useful in indicating 

the kind and extent of all identifiable effects on public health or 

welfare which may be expected from the presence of such pollutant in the 

ambient air, in varying quantities. 

(U.S.Code, 1982) 

Air quality criteria provide the Agency with a scientific basis for deciding 

whether regulations controlling given pollutants are necessary and for deriving 

such ambient air quality standards as may be needed. 

Among the air pollutants designated by the Administrator as criteria 

pollutants are those known as photochemical oxidants. This document is a 

revision ofAir Quality Criteria for Ozone and Other Photochemical Oxidants 

(U.S. Environmental Protection Agency, 1978a). Its purpose is to review and 

evaluate the scientific literature on ozone and related oxidants and to document 

their effects on public health and welfare. 

The term "photochemical oxidants" has historically been defined as those 

atmospheric pollutants that are products of photochemical reactions and that 

are capable of oxidizing neutral iodide ions (U.S. Environmental Protection 

Agency, 1978a). Research has established that photochemical oxidants in 

ambient air consist mainly of ozone, peroxyacetyl nitrate, and nitrogen dioxide, 

and of considerably lesser amounts of other peroxyacyl nitrates, hydrogen 

peroxide, alkyl hydroperoxides, nitric and nitrous acids, and formic acid. 

Other oxidants suspected to occur in ambient air but only in trace amounts 

include peracids and ozonides. 

2-1 
Although it is by definition a photochemical oxidant, nitrogen dioxide is 

not included among the oxidants discussed in this document. The formation of 

nitrogen dioxide clearly precedes the formation of ozone and other related 

oxidants in the ambient air. Thus, nitrogen dioxide is the dominant oxidant 

early in the day, while ozone and other related oxidants predoininate from late 

morning or midday through much of the afternoon. Nitrogen dioxide is known to 

exert deleterious effects on human health and welfare. This fact, coupled 

with temporal and spatial variations in concentrations that differ from those 

of ozone and related oxidants, underlies the listing of nitrogen dioxide under 

section 108(a)(1) of the Clean Air Act as a criteria pollutant separate from 

ozone and other photochemical oxidants. The second criteria document prepared 

by EPA on the oxides of nitrogen was completed in 1982 (O,S. Environmental 

Protection Agency, 1982a). That document discussed nitric and nitrous oxides, 

nitrogen dioxide, nitric and nitrous acids, and nitrosamines. As used in this 

document, the term "photochemical oxidants" refers to ozone, the peroxyacyl 

nitrates, hydrogen peroxide, and formic acid. The oxides of nitrogen are discussed, but only in the context of their role as precursors to ozone and 

related oxidants. 

2.2 THE OXIDANT PROBLEM 

Ozone (03 ), a reactive allotrope of oxygen (0 2 ), occurs as a natural 

component of the atmosphere. It is found in its highest concentrations in the 

stratosphere, where it is formed through cyclic reactions resulting from the 

photolysis of oxygen into atomic oxygen and the subsequent reaction of atomic 

oxygen with other oxygen molecules. 

Incursions of stratospherically produced ozone into the lower troposphere 

occur through meteorological and atmospheric exchange phenomena, resulting in 

a global background of ozone. To this global background of ozone of stratospheric 

origin are added ozone formed in the free troposphere and the contributions of 

ozone produced in the ambient air from photochemical reactions involving 

manmade emissions and natural products (e.g., natural emissions of volatile 

organic compounds). Manmade emissions of nitrogen oxides and volatile organic 

compounds are the chief contributors to the ozone burden found in the ambient 

air of urban areas. The presence of ozone in ambient air is the net result of 

various formation, stratospheric-tropospheric exchange, transport, and destruction processes. 

2-2 
Other photochemical oxidants occur in ambient air, but the nature and 

source of their global backgrounds are not well established. The additional 

photochemical oxidants of concern in this document, namely hydrogen peroxide, 

peroxyacetyl nitrate and its homologues, and formic acid, have, except for 

the latter, been detected in remote environments thought to be free of manmade 

influences; and all have been detected in the ambient air of urban areas. 

The toxicity of ozone is well known. It is a strong oxidizing agent that 

is highly reactive with a wide spectrum of chemical moieties. Since ozone is 

a gas, studies of its health-related toxicity have centered largely on its 

capacity for affecting pulmonary function and the morphology of the respiratory 

tract, which is now well documented. In addition, its effects on Oxtrapulmonary tissues and systems are also of concern and are the subject of some of 

the research discussed in this document. Studies of toxic effects of ozone on 

vegetation are also well documented and have focused on foliar injury and 

reduction in growth and yield. The toxicities of the peroxyacyl nitrates, of 

hydrogen peroxide, and of formic acid are less well documented than the toxicity 

of ozone, having been the focus of considerably less research because the 

levels at which these oxidants occur in the ambient air, even in urban areas, 

appear to warrant much less concern. 

Ozone, but not the other oxidants mentioned above, is regulated under 

provisions and procedures spelled out in the Clean Air Act. Its concentrations in ambient air are controlled through the promulgation and attainment of 

primary and secondary national ambient air quality standards (NAAQS). As 

described in the Clean Air Act, criteria pollutants are those atmospheric 

pollutants that are ubiquitous and are emitted into the air from numerous and 

diverse sources. While widespread, ozone and the other photochemical oxidants 

found in ambient air are not emitted into the air as primary pollutants. 

Rather, they are formed as secondary pollutants in the atmosphere from ubiquitous primary organic and inorganic precursors that are emitted by a multiplicity of sources. Consequently, photochemical oxidant pollution in this 

country is the result of a combination of many factors, such as local meteorological conditions and the concentrations, composition, and patterns of occurrence of the primary pollutants that give rise to the oxidants. 

2-3 
2.3 SCOPE AND ORGANIZATION OF THIS DOCUMENT 

The atmosphere does not easily lend itself to the partitioning required 

for documentation. Nevertheless, certain boundaries are logical for purposes 

of discussion as well as for purposes of regulatory decisions. Ozone and its 

organic precursors are known° to give rise to secondary orgarlic aerosols. 

Likewise, ozone and hydrogen peroxide both appear to participate in the atmospheric oxidation of nitrogen dioxide (NO 2 ) and sulfur dioxide (S0 2 ) to those 

inorganic aerosols leading to visibility degradation in the atmosphere and to 

acidic deposition. The contributi"ons of ozone and hydrogen peroxide to the 

oxidation of NO 2 and S02 cannot be quantified at present, but are known to be 

minor compared to the oxidation of these compounds by hydroxyl radicals. In 

addition, ozone, the principal photochemical oxidant in ambient air, has no 

direct effects on visibility since, unlike NO 2 and S02 , it does not absorb 

energy in the visible region of the spectrum. Thus, this document includes 

brief discussions, in Chapter 3, of the atmospheric chemistry of ozone and 

hydrogen peroxide relative to the formation of inorganic nitrogen and sulfur 

aerosols but does not include information on the actual effects associated 

with visibility degradation or acidic deposition. Since NO 2 and S02 are the 

immediate, direct precursors to the aerosol species involved, visibility 

degradation and acidic deposition are discussed in the respect:ive air quality 

criteria documents on oxides of nitrogen and on particulate matter and sulfur 

oxides (U.S. Environmental Protection Agency, 1982a,b). 

This document has been divided into five volumes for ease of review, 

printing, and distribution. Volume I consists of the summary and conclusions 

for the entire document. Volume II contains the introduction to the document 

(Chapter 2) and all chapters dealing with the formation, transport, and fate 

of photochemical oxidants (Chapter 3); the measurement of oxidants and their 

precursors (Chapter 4); and the concentrations of oxidants in ambient air 

(Chapter 5). Volume III contains the documentation of the effects of photochemical oxidants on vegetation, ecosystems, and nonbiological materials 

(Chapters 6, 7, and 8, respectively). Volume IV reviews the extensive body of 

data available on the toxicological effects of ozone and other oxidants in 

experimental animals and on in vitro effects on human cells and body fluids 

(Chapter 9). In Volume V, effects observed in human contro'Iled exposures 

(Chapter 10) and in field and epidemiological studies (Chapter 11) are presented. In addition, that volume contains an evaluation of the health effects 

data of probable consequence for regulatory purposes (Chapter 12). 

2-4 
Neither control techniques nor control strategies for the abatement of 

photochemical oxidants are discussed.in this document, although some of the 

topics included are relevant to abatement strategies. Technology for controlling the emissions of'nitrogen oxides and of volatile organic compounds is 

discussed in documents issued by the Office of Air Quality Planning and Standards (OAQPS) of the U.S. Environmental Protection Agency (e.g., U.S. Environmental Protection Agency, 1978b, 1983). Likewise, issues germane to the scientific 

basis for control strategies, but not pertinent to the development of criteria, are addressed in numerous documents issued by OAQPS. 

In addition, certain issues of direct relevance to standard-setting are 

not explicitly addressed in this document, but are addressed instead in documentation prepared by OAQPS as part of its regulatory analyses. Such analyses 

include: (1) discussion of what constitutes an "adverse effect," that is, the 

effect or effects the standard is intended to protect against; (2) assessment 

of risk; and (3) discussion of factors to be considered in providing an adequate margin of safety. While scientific .data contribute significantly to 

decisions regarding these three issues, their resolution cannot be achieved 

solely on the basis of experimentally acquired information. Final decisions 

on items (1) and (3) are made by the Administrator. 

A fourth issue directly pertinent to standard-setting is identification 

of the population at risk, which is basically a selection by the Agency of the 

population to be protected by the promulgation of a given standard. This 

issue is addressed only partially in this document. For example, information 

is presented in Chapter 12 on factors, such as pre-existing disease, that 

biologically may predispose individuals and subpopulations to adverse effects 

from exposures to ozone. The identification of a population at risk, however, 

requires information above and beyond data on bi'ological predisposition, such 

as information on levels of exposure, activity patterns, and personal habits. 

Such information is included in a staff paper developed by OAQPS. Thus, the 

identification of the population at risk reiative to standard-setting is the 

purview of OAQPS and is not addressed in this document. For information on the 

standard-setting process, see Padgett and Richmond (1983) and McKee et al. 

(1985). 

This document consists of the review and evaluation of reievant literature 

on ozone and other photochemical oxidants through early 1986. The material 

selected for review and comment in the text generally comes from the more recent 

2-5- 
literature, with emphasis on studies conducted at or near pollutant concentrations found in ambient air. Older literature that was cited in the previous 

criteria document for ozone and other photochemical oxidants (d.S. Environmental 

Protection Agency, 1978) has often been summarized and present.ed briefly. An 

attempt has been made, however, to discuss at greater length in the text older 

studies (1) judged significant because of their usefulness in deriving the 

1979 standards; (2) open to reinterpretation because of newer data; or 

(3) potentially useful in deriving subsequent standards. The newer information on oxidants now available may in some instances make possible a better 

understanding of the earlier studies, such that a more detailed and comprehensive picture of health effects is emerging on several issues. An attempt has 

been made to discuss key literature in the text and present it in tables as 

well. Reports of lesser importance to the purposes of this document may 

appear in tables only. 

Generally, .only published material that has undergone scientific peer 

review is included. In the interest of admitting new and important information, 

however, some material not published in the open literature but meeting other 

standards of scientific reporting may be ihcluded. Emphasis has been placed 

on studies in which exposure concentrations were <1 ppm. On this basis, 

studies in which the lowest concentration employed exceeded this level have 

been included only if they contain unique data, such as documentation of a 

previously unreported effect or of inechanisms of effects; or if they were 

multiple-concentration studies designed to provide information on concentrationresponse r.elationships. Application of.a concentration cutoff .of 1 ppm to 

health effects studies eliminates discussion of studies on mortality and 

sublethal effects. In the areas of mutagenesis, teratogenesis, and reproductive effects, however, results of studies conducted at much higher than ambient 

levels have been included because of the potential importance of these longterm effects to public health and welfare. 

In selecting studies for consideration, each paper or other publication 

was reviewed in detail. Technical considerations for inclusion of a specific 

study on health or welfare effects, for example, included, but were not 

restricted to, an analysis of the exposure method; specificity or appropriateness of the analytical method used to monitor the oxidant concentration; 

information on oxidant monitoring practices such as location, calibration, and 

sampling time; and the appropriateness of the technique used to measure the 

2-6 
effect. In addition, for health effects studies technical considerations 

included the characteristics of the subjects studied and the techniques used 

for obtaining or selecting the study cohorts. Interpretation of the results 

included consideration of the following factors: the end results of the 

statistical analysis; the degree to which the results are plausibie in the 

context of other extant data; the appropriateness of the-hypothesis developed; 

and the agreement between the hypothesis and the results reported. Unless 

otherwise stated, all, results cited in the text for health and vegetation 

studies are statistically significant at p<0.05. 

The general policy of EPA is to express concentrations of air pollutants 

in metric units, e.g., in micrograms per cubic meter (pg/m 3 ); as weil as in 

the more widely used units, parts per million (ppm) or parts per billion 

(ppb), which are neither metric nor English units. That policy has been 

followed in those chapters in which most of the data have been obtained from 

laboratory studies done at room temperature (e.g., Chapters 9 and 10). Data 

reported in ppm for studies conducted outdoors, such as field and open - top 

chamber vegetation studies, ambient air monitoring, and research on atmospheric 

chemistry, have not been converted. Conversion of reported ppm and ppb units 

is highly questionable in these cases because it assumes standard or uniform 

temperatures and pressures. For data in the health chapters, the conversion 

units used are 1 ppm ozone = 1962 pg/m 3 , and 1 ppm PAN = 4945 pg/m 3 ; at 

1 atmosphere pressure and 25°C. 

2.4 REFERENCES 

McKee, D.; Johnson, P.; Richmond, H.; Jones, M.; McCurdy, T.; Walton, T. 

(1985) Work plan for the ozone National Ambient Air Quaiity Standards. 

Presentation to the Clean Air Scientific Advisory Committee, March; 

Research Triangle Park, NC: U.S: Environmental Protection Agency, Office 

of Air Quality Planning and Standards. 

Padgett, J.; Richmond, H. (1983) The process of establishing and revising 

national ambient air quality standards. J. Air Pollut. Control Assoc. 

33:14. 

U.S. Code. (1982) Clean Air Act, §108, air quality criteria and control 

techniques. U.S.C. 42: §7408. 

2-7 
U.S. Environmental Protection Agency. (1978a) Air quality cri -teria for ozone 

and other photochemical oxidants. Research Triangle Park, NC: U.S. Environmental Protection Agency, Environmental Criteria and Assessment Office; 

EPA report no. EPA-600/8-78-004. Available from: NTIS, Springfield, VA; 

PB80-124753. 

U.S. Environmental Protection Agency. (1978b) Control techniques for volatile 

organic emissions from stationary sources. Research Triangle Park, NC: 

U.S. Environmental Protection Agency, Office of Air Quality, Planning and 

Standards; EPA report no. EPA-450/2-78-022. Available from: NTIS, 

Springfield, VA; PB-284804/2. 

U.S. Environmental Protection Agency. (1982a) Air quality criteria for oxides 

of nitrogen. Research Triangle Park, NC: U.S. Environmental Protection 

Agency, Environmental Criteria and Assessment Office; EPA report no. 

EPA-600/8-82-026. Available from: NTIS, Springfield, VA; PB83-163337. 

U.S. Environmental Protection Agency. (1982b) Air quality criteria for particulate matter and sulfur oxides. Research Triangle Park, NC: U.S. Environmental Protection Agency, Environmental Criteria and Assessment Office; 

EPA report no. EPA-600/8-82-029. .Available from: NTIS, Springfield, VA; 

PB84-156777. 

U.S. Environmental Protection Agency. (1983) Control technology for nitrogen 

oxides emissions from stationary sources. Revised second edition. 

Research Triangle Park, NC: U.S. Environmental Protection Agency, Office 

of Air Quality, Planning and Standards; EPA report no. EPA-450/3-83-002. 

Available from: NTIS, Springfield, VA; PB84-118330/REB. 

: 
3. PROPERTIES, CHEMISTRY, AND TRANSPORT OF OZONE AND,OTHER 

PHOTOCHEMICAL OXIDANTS AND THEIR PRECURSORS 

3.1 INTRODUCTION 

Ozone and other oxidants found in ambient air, such as the peroxyacyl 

nitrates and hydrogen peroxide, are formed as the result of atmospheric 

physical and chemical processes involving two classes of.precursor 

pollutants, volatile nonmethane organic compounds (NMOC) and nitrogen 

oxides (NO x). The formation of ozone and other oxidants from these 

precursors is a complex, nonlinear function of many factors, including the 

intensity and spectral distribution of sunlight; atmospheric mixing and 

related meteorological conditions; the concentrations of the precursors in 

ambient air and, within reasonable concentration ranges, the ratio between 

NMOC and NOx(NMOC/NOx); and the reactivity of the organic precursors. 

This chapter describes the physical and chemical properties of 

ozone and other photochemical oxidants (Section 3.2). It also characterizes the nature of the precursors in terms of their physical and chemical 

properties, their sources and emissions into the atmosphere, and their 

concentrations in ambient air (Section 3.5). In addition, a brief description 

is provided (Section 3.3) of the complex atmospheric chemical processes by 

which ozone and other photochemical oxidants are formed from their precursors. 

A brief discussion is also included of the relationship of ozone and other 

oxidants to atmospheric phenomena that result from the formation of secondary 

organic and inorganic aerosols. 

In addition to the information on the chemistry of oxidants and their 

precursors, the chapter includes a discussion of ineteorological processes 

(Section 3.4) that contribute to the formation of ozone and other oxidants. 

and that govern their transport and dispersion once formed. Finally, an 

overview is given (Section 3.6) of models of source-receptor relationships 

between precursor emissions and ozone formation in the atmosphere, which 

either implicitly or explicitly include the relevant emissions, atmospheric 

chemistry, and meteorological processes. 

3-1 
3.2 DESCRIPTIONS AND PROPERTIES OF OXIDANTS AND THEIR PRECURSORS 

3.2.1 Ozone and Other Photochemical Oxidants 

3.2.1.1 Ozone. Ozone (03 ) is a triangularly shaped molecule consisting 

of three oxygen atoms arranged in four basic resonance structures: 

:0: :0: :0: :0: 

:0: :0: :0: :0: :0: :0: :0: :0: 

+ — — — -- + 

(I) (II) (III) (IV) 

The first and fourth structures, which predominate, are characterized by 

the presence of a terminal oxygen atom having only six electrons. The 

resonance forms depicted above have no unshared electrons. As the result 

of the presence of only six electrons on one of the oxygen atoms in ozone, 

the chemical reactions of ozone are electrophilic; that is, ozone removes 

electrons from or shares electrons with other molecules or ions. By 

definition, then, ozone is an oxidant since the term "oxidant" 

characterizes an ion, atom, or molecule that is capable of removing one or 

more electrons from another ion, atom, or molecule, a process called 

"oxidation." A"reducing agent" adds one or more electrons to another 

ion, atom or molecule, a process called "reduction." Oxidation and 

reduction reactions occur in pairs and the coupled reactions are known as 

' ° redox reactions." In redox reactions, the oxidizing agent is reduced and 

the reducing agent is oxidized. The two components of such redox 

reactions are known as "redox pairs." The significance of redox reactions 

involving ozone is discussed in Chapters 6 and 9. The capability of' 

a chemical species for oxidizing or reducing is termed "redox potential" 

(positive or negative standard potential) and is expressed in volts. 

Uzone is, in fact, a strong oxidant having a standard potential of +2.07 

volts in aqueous systems (Weast, 1977). 

The physical properties of ozone are given in Table 3-1 (U. S. 

Department of Health, Education, and Welfare, 1970, modified). 

3-2 
TABLE 3--1. PHYSICAL PROPERTIES OF OZONE 

Physical state 

Chemical formula 

Molecular weight 

Melting point 

Boiling point 

Specific gravity relative to air 

Vapor density 

At OoC, 760 mm Hg 

At 250C, 760 mm Hg 

Solubility at OoC 

(Indicated volume of ozone at 

OoC, 760 mm Hg) 

Hen's Law constant, 

37CandpH=7 

Conversion factors 

At OoC, 760 mm Hg 

At 250C, 760 mm Hg 

Colorless gas; blue violet liquid 

03 

48.0 

-192.7 f 0.20C 

-111.9 t 0.30C 

1.658 

2.14 g/liter 

1.96 g/liter 

0.494 m1/100 ml water 

8666 atm/mole fraction a 

1 ppm = ?141 μg/m 3 4 

1 N,g/m = 4.670 x 10- 

1 ppm = 962 μg/m 3 

1 μg/m~ = 5.097 x 10 -4ppm 

aCalculated by formula of Roth and Sullivan (1981). 

Source: U. S. Department of Health, Education, and Welfare (1970), modified. 

3.2.1.2 Peroxyacetyl Nitrate. Peroxyacetyl nitrate (PAN) has been 

observed as a constituent of photochemical smog in many localities, though 

its concentrations and its ratio to ozone differ as a function of time at 

a given location as well as from place to place (Chapter 5). Peroxyacetyl 

nitrate, which has the formula CH 3C(0)02 NO2 , exists in a temperature 

dependent equilibrium with its decomposition products, NO 2 and acetylperoxy radicals. It can persist for substantial periods of time in the 

atmosphere, depending upon temperature and the NO 2/NO ratio (Cox and 

Roffey, 1977). 

3-3 
The chief property of interest regarding PAN is its oxidizing 

ability. A second property of PAN of interest is its thermal instability. 

In the laboratory, this thermal instability necessitates that precautions 

be taken in synthesizing, handling and storing PAN, since improper 

handling and storage have resulted in explosions (Stephens et al., 

1969). The ready thermal decomposition of PAN results in a notable 

temperature dependence in the rate of PAN decomposition in ambient air. 

Partly because of the thermal instability of PAN, its properties have 

not been as well characterized as those of 03 or H 2O2 . Recent work on the 

physical properties of PAN, however, has confirmed data reported earlier, 

and results of the earlier and more recent work are shown in Tables 3-2 

and 3-3 (Stephens, 1969; U. S. Dept. of Health, Education, and Welfare, 1970; 

Kacmarek et al., 1978; Bruckmann and Willner, 1983; Holdren et al., 1984). 

The infrared (IR) spectrum of PAN is important since most researchers 

rely on °it for establishing concentrations of PAN for calibration. Bruckmann and Willner (1983) reported the IR spectrum of pure PAN and the Raman 

spectrum of liquid PAN at -40 0C in an argon matrix. Their work confirmed 

effects that correlate with the ultraviolet (UV) spectrum published 

earlier by Stephens (1969); that is, PAN was shown to be stable at X>300 

nm but was efficiently photolyzed at %<300 nm (Bruckmann and Willner, 

1983). Actinic radiation falling upon the surface of the earth has 

wavelengths >295 nm, and it is light at wavelengths between —295 and —430 

nm which is involved in photochemical air.pollution formation. 

3.2.1.3 Hydrogen Peroxide. Hydrogen peroxide (H2O2 ) is an oxidant that 

occurs in ambient air as a component of photochemical smog. It is 

believed to be formed through the recombination of two hydroperoxy 

radicals (H0 2 ) in the presence of a third, energy-absorbing molecule 

(section 3.3.1.3). In aqueous media, H 2O2 is an inorganic aci-d that has a 

dissociation constant of 2.4 x 10 -12 and a pK of 11.62 (at 250C) (Weast, 

1977). Hydrogen peroxide has a standard potential of +1.776 in the redox 

pair, H2O2/H2O. The physical properties of H 2O2 are given in Table 3-4. 

Additional properties should be noted here that are of interest 

relative to whether effects of H 202 in biological receptors are of 

siynificance. First, H2O2 , though classed as a reasonably s -trong oxidant 

on the basis of its standard potential for the redox system H2O2/H2O, has 

3-4 
TABLE 3-2. PHYSICAL PROPERTIES OF PEROXYACETYL NITRATE 

Physical state, @250C Colorless liquid 

Chemical formula CH3C002NO2 

Molecular weight 121 

Boiling point, oC 106t 2a 

103.9b 

Triple point, oC -50 t 0.5a 

-48 t 0.5b 

Vapor pressure, 

@room temperature —15 mm Hg 

Vapor pressure curve ln p=-4587/T + 18.76 a,c 

ln p = -4585/T + 18.79b 

Hydrolysis t l 2 = 3.5 min at pH = 7.8d ; 

In alkaline solution ~ roducts include nitrite ion 

and molecular oxygene 

In acidic solution f 

@220C, pH 5.6 4 x 10-4sec -1 

@250C, pH 5.6 6.8 x 10-4sec -lg 

Henry's Law constant, 

@100C 5 t 1 M atm-lg 

Conversion factors 

@OoC, 760 mm Hg 1 ppm = 5398 /m3 

1 μg/m 3= 1 ~ 52 x 10-4,ppm 

@250C, 760 mm Hg 1 ppm =1945 μg/m3; 4 1 μg/m = 2.022 x 10 -ppm 

Sources: 

a Bruckmann and Wilner (1983). 

b Kacmarek et al. (1978). 

c Temperature, T, in oK; pressure, p, in torr. 

d Mudd (1966). 

e Stephens (1967); Nicksic et al. (1967). 

f Lee et al. (1983). 

gHoldren et al. (1984). 

Source of remainder of data: 'U. S. Department of Health, Education and 

Welfare (1970). 

3-5 
TABLE 3-3. INFRARED ABSORPTIVITIES - YF P~ ROXYA~ETYL NITRATE 

(RELATED TO 295 0 K and 973 mb) (ppm m x 10 )(base 10) 

Frequency, cm-1 

Reference 1842 1741 1302 1162.5 791.5 

Bruckmann and Willner (1983a ) 12.4 32.6c 13.6 15.8 13.4 

Stephens (1964, 1969b ) 10.0 23.6d 11.2 14.3 10.1 

aAt 4 mbar; no diluent; resolution 1.2 cm -1 . 

bAt 7 mbar in N2 diluent at 973 mb total pressure and 295 K; grating 

instrument. 

cQ branch resolved (1.5 torr). 

dQ branch not resolved (1.4 mbar). 

TABLE 3-4. PHYSICAL PROPERTIES OF HYDROGEN PEROXIDE 

Physical state, @25 0C 

Chemical formula 

Molecular weight 

Melting point, oC 

Boiling point, oC, @760 mm Hg 

Density, @25 0C, 760 mm Hg 

Vapor pressure, @16.3 0 C 

Conversion factors 

@OoC, 760 mm Hg 

@250C, 760 mm Hg 

Source: Weast (1977). 

Colorless liquid 

H2 O2 

34.01 

-0.41 

150.2 

1.4422 

—1 mm Hg 

1 ppm = 1520 μg/m3; 4 1 μg/m = 6.594 x 10 -ppm 

1 ppm = 1390 μg/m3; 

4 1 μy/m = 7.195 x 10-ppm 

3-6 
been reported to be a positive interference in measurements of total 

oxidants made by the Mast meter but to give a very slow response (slow 

color development) in the NBKI inethod for total oxidants (Chapter 4). 

This difference should be borne in mind when effects attributed to 

oxidants, as opposed to ozone, are evaluated. Second, H 2O2 occurs 

normally as a substrate•in biological systems and is involved in several 

redox pairs of biological importance (see, for example, West et al., 

1966). It should also be noted that enzymes are present, at least in 

mammalian systems,`that catalyze the breakdown of H 2O2 . 

3.2.1.4 Formic Acid. Formic acid is a stable product formed in photochemical air pollution from, for example,,the reaction of H02 radicals 

. . 

with yCHO and from the reactions'of the Criegee biradical CH 2O0 with water 

vapor (Atkinson and Lloyd, 1984). It has been detected in polluted 

ambient atmospheres by longpath infrared spectroscopy on the basis of its 

characteristic Q-branch absorption at 1105 cm -1 (Hanst et al., 1975; 

Tuazon et a1., 1978a, 1980, 1981a). 

Formic acid has the structure of both an acid and an aldehyde and 

hence it differs in chemical behavior from other carboxylic acids in which 

the carboxyl group is linked to a hydrocarbon residue rather than to a 

lone hydrogen atom. In concentrated form, HCOOH is a pungent-smelling, 

highly corrosive liyuid with a boiling point of 100.5 0C. 

3.2.2 Organic Precursors 

This section briefly describes and defines those hydrocarbons and 

other volatile organic compounds commonly found in the ambient air of the 

United States and provides relevant information about their chemical and 

physical properties. 

The term "hydrocarbon" has been used since the initial investigations 

of tropospheric photochemistry to represent those compounds of carbon and 

hydrogen that exist as gases in the ambient air and that participate along 

with oxides of nitrogen in reactions that form ozone and other photochemical oxidants. As knowledge of atmospheric chemistry has increased, some 

carbon compounds containing elements such as oxygen and the halogens have 

also been shown to be important in photochemical air pollution. Thus, the 

term "volatile organic compounds" (VOC) has come to be used to describe 

3-7 

~ 
stable organic compounds that exist as gases under normal atmospheric 

conditions, most of which can participate in the formation o -F photochemical oxidants. Recognition that methane (CH4 ) is virtual1y unreactive 

in the photochemical formation of ozone and other oxidants has given rise 

to the more accurate term, "nonmethane organic compounds" (NMOC), for 

describing those gas-phase organic compounds in ambient air that serve as 

precursors to ozone and other photochemical oxidants. While these three 

terms may sometimes appear to be used interchangeably in this chapter, the 

terminology used reflects that reported in the specific literature cited 

in this chapter, though in some instances differentiations may have been 

made for purposes of discussion. 

As discussed in Chapter 4, methods for measuring total gas-phase 

hydrocarbons are not specific for hydrocarbons but may also detect other 

gas-phase organic compounds, though they will not measure thein quantitatively. Where methods are used that permit speciation of the compounds 

measured, organic compounds other than hydrocarbons can be and usually are 

excluded from the summation of individual species used to arrive at a 

total nonmethane hydrocarbon (TNMHC) concentration. Where researchers 

have used methods that do not permit speciation, an indefinite and variable fraction of the reported TNMHC concentration may, in fact, be the 

result of the presence of nonhydrocarbon organics and such concentration 

data are more properly reported as total nonmethane organic compounds 

(NMOC). 

The discussion that follows is aimed at presenting basic facts on 

nomenclature and those characteristics of photochemically reactive 

volatile organic compounds that are relevant to the information given 

subsequently in this and later chapters. 

3.2.2.1 Hydrocarbons. Hydrocarbons are compounds consisting of hydrogen 

and carbon only. For a given homologous series the volatilities of hydrocarbons are related generally to the number of carbon atoms in each 

molecule, as well as to temperature. Hydrocarbons with a carbon number of 

one to four are gaseous at ordinary temperatures, while those with a 

carbon number of five or more are liquid or solid in pure state. Liquid 

mixtures of hydrocarbons such as gasoline may include some compounds in 

pure form that are gases, as well as those that are liquids. Likewise, 

ME 
gas-phase mixtures in ambient air will usually include compounds that are 

liquid or solid in their pure form. Hydrocarbons with a carbon number of 

about eight or less are abundant in ambient air, but those with a carbon 

number greater than about 12 have generally not been reported in the gas 

phase at significant concentrations, probably because of the inability of 

analytical techniques to detect these high molecular weight organics. 

A saturated hydrocarbon has each of its carbon atoms bonded to four 

other atoms; whereas an unsaturated hydrocarbon has two or more carbon 

atoms bonded to fewer than four other atoms. 

A1kanes . Alkanes, also known as paraffins, are saturated hydrocarbons having the general formula C n H2n+2 . The first compound in the 

series is methane, CH4, which, because of its low reactivity, does not 

contribute significantly to photochemical air pollution in urban atmospheres. Alkanes as a class are the least reactive of the photochemically 

important hydrocarbons (U. S. Environmental Protection Agency, 1978a,b). 

Alkanes may be straight- or branched-chain compounds, and comprise the 

open-chain (acyclic) hydrocarbons known as aliphatic hydrocarbons. 

Alkenes . Alkenes, also known as olefins, have at least one 

unsaturated bond. The number of hydrogen atoms in the general formula of 

alkenes is decreased by two with respect to the alkanes for each double 

bond between carbon atoms; the general formula for alkenes with one double 

bond, for example, is C n H2n . The first compound in the alkene class is 

ethene, also known as ethylene; the second is propene, also known as 

propylene. For alkenes containing more than three carbon atoms the 

position of the double bond is specified by a numerical prefix (e.g., 1- 

butene). Compounds with carbon numbers three or higher can have two 

double bonds between the carbon atoms and are called dienes. The complete 

name of a diene is formed by including a prefix with numbers that indicate 

the location of the double bonds. Like alkanes, alkenes are aliphatic 

hydrocarbons and may exist as straight or branched chains. As a class, 

alkenes are among the most reactive hydrocarbons in photochemical systems 

(see Section 3.2.2.4). 

Terpenes . TerNenes are a naturally occurring subgroup of alkenes, 

many of which have the formula C 10H16 . Among the terpenes identified in 

ambient air a- and p-pinene have been most frequently studied. Both a - 

3-9 
and p-pinene contain six-membered rings, as do several other terpenes; but 

at least one commonly occurring member of this group, myrcene, is an 

acyclic or open-chain compound. Isoprene, also a naturally occurring 

alkene, is a hemiterpene having the formula C 5 H8 . 

Alkynes . Alkynes are open-chain hydrocarbons that contain one or 

more triple bonds. Acetylene, C2H21 is the simplest member of the class, 

which as a whole is often referred to as the acetylenes. The general 

formula for the acetylenes is CnH2n-2, and for each additional triple bond 

in the molecule four hydrogen atoms must be removed from the general 

formula. Acetylene is commonly present in ambient air, is thought to be 

emitted largely from mobile sources, and has often been taken to be an 

indicator of auto exhaust emissions, since it .is relatively unreactive in 

ambient air and persists in the atmosphere longer than most other exhaust 

components. 

Aromatics . Aromatic hydrocarbons include various compounds having 

atoms arranged in six-membered carbon rings with only one additional atom 

(of hydrogen or carbon) attached to each atom in the ring. 3enzene is the 

simplest compound in the series, having no side chains but only six carbon 

atoms and six hydrogen atoms, linked by three conjugated double bonds. 

Compounds containing the aromatic ring and elements other than carbon 

and hydrogen are included with aromatic hydrocarbons in the general classification "aromatics." The double bonds in aromatics are not nearly as 

chemically active as those in alkenes because of an effect called 

"resonance stabilization." As a class, aromatics exhibit a wide range of 

photochemical reactivity,with benzene having a low photochemical reactivity and 1,3,5-trimethylbenzene, for example, showing high reactivity. 

3.2.2.2Aldehydes . Aldehydes probably constitute the single most 

abundant group of volatile organic compounds other than hydrocarbons in 

ambient air. They are photochemically . important compounds because they 

photolyze to form free radicals that will react with oxygen an ambient air 

to form alkylperoxy or hydroperoxy radicals (National Research Council, 

1977a) (see below). 

Aldehydes are characterized by the presence of the form,yl functional 

group (CHO). A carbonyl group having a carbon-oxygen double bond, C=O, is 

part of the formyl group. The carbonyl group is not unique to aldehydes, 

3-10 
since it is found also in ketones and carboxylic acids; but it forms the 

basis for one of the analytical methods used for measuring aldehydes in 

ambient air (Chapter 4). 

3.2.2.3Other Organic Compounds . Other organic compounds found in 

ambient air are known to be photochemically reactive in the formation of 

ozone and other photochemical oxidants. These other organic compounds do 

not occur in ambient air collectively, much less singly, at concentrations 

that approach the total concentrations of nonmethane hydrocarbons. Some 

of them are suspected of having potentially adverse health effects, 

however, and are therefore under scrutiny by the U. S. Environmental 

Protection Agency. These compounds are mentioned here only because they 

are photochemically reactive, can serve as precursors to oxidants, and 

because they contribute a small but indeterminate fraction of the total 

NMOC concentrations reported when continuous hydrocarbon analyzers 

(Chapter 4) are used'to determine ambient levels of volatile organic 

compounds. 

Many of the volatile organics in ambient air that are not hydrocarbons are organic halides, in which one or more hydrogen atoms of a 

hydrocarbon have been replaced by a halogen atom such as chlorine, 

fluorine, or iodine. An enormous number of relatively simple organic 

halides are possible, since a halogen atom can be attached to an organic 

compound in many different positions. 

3.2.2.4Volatility and Reactivity . The physical and chemical propert.ies. 

of nonmethane organic compounds that are most pertinent to their role as 

precursors to ozone and other oxidants are those properties that govern 

their emission into the atmosphere (volatility) and their lifetime in the 

atmosphere, the latter being determined by photochemical reactions 

(reactivity) and other removal processes (e.g., gas-to-particle conversion 

and dry deposition). 

To be significant in atmospheric reactions, an organic.compound must 

have a sufficiently high volatility. Based upon a review of the available 

literature, Singh et al. (1984) have chosen a vapor pressure of 10 -8 atm 

as the criteria for deciding whether an organic compound should be described as volatile. Those compounds with vapor pressures less than 10 -8 

atm were considered by Singh et al. (1984) to occur predominately in 

3-11 
the condensed phase and therefore not to participate in atmospheric 

reactions. Clearly, a rigid cutoff for vapor pressures will not 

necessarily be applicable to all organic compounds. 

The photochemical reactivity of subclasses and individual species of 

hydrocarbons and of other volatile organic compounds is relevant to mechanistic studies in atmospheric chemistry, to modeling, and to other 

oxidant-control-related research; but it is not pertinent to the derivation of criteria. A major discussion of these properties therefore lies 

outside the scope of this document, but a brief discussion of the concept 

of hydrocarbon reactivity and its application is presented here. 

Differences in reactivities among volatile organic compounds have 

been the focus of considerable attention and research for nearly three 

decades. In early research on photochemical air pollution, inany different 

definitions or criteria were used to evaluate the reactivity of organic 

compounds. Examples of such criteria include: rate of NO-to-NO 2 conversion, maximum ozone concentration formed, initial rate of disappearance of 

the oryanic compound, eye irritation, damage to vegetation, and aerosol 

formation. 

Historically, reactivity classifications have been based on environmental chamber measurements of these criteria, as observed in the photooxidation of hydrocarbon-oxides of nitrogen mixtures under conditions 

approximating those of polluted ambient atmospheres. Many of the reactivity data that had been accumulated through 1969 for each of these 

manifestations (except plant damage) were critically reviewed by 

Altshuller and Bufalini in 1971. They noted general agreement in reactivity trends from studies employing different reactivity criteria, but 

they also cited a number of significant discrepancies in the specific 

assignments of reactivity to individual compounds and even -to whole 

classes of compounds. 

In recent years assessments of the reactivity of volati'le organic 

compounds have focused almost exclusively on the ability of an organic 

compound to produce ozone and other photochemical oxidants. This focus 

arises from interest in regulating most stringently the emissions of 

organic compounds having the highest potential for forming ozone and other 

photochemical oxidants. 

3-12 
The 1978 criteria document for ozone and other photochemical oxidants 

summarized reactivity data acquired from the mid-1960s to the mid-1970s 

(U. S. Environmental Protection Agency, 1978a). Reference to tables of 

reactivity schemes given in the 1978 document shows the relatively higher 

reactivities of internally double-bonded alkenes, of aliphatic aldehydes 

and other carbonyl compounds (such as branched alkylketones and 

unsaturated ketones), of dienes, of 1-alkenes, of partially halogenated 

alkenes, and of alkylbenzenes (primary and secondary monoalkylbenzenes, 

and di-, tri- and tetraalkylbenzenes). Other compounds also have 

relatively high reactivity but are not expected to be as abundant in 

ambient air as the compounds cited above. 

For more information the reader is referred to the 1978 criteria 

document (U. S. Environmental Protection Agency, 1978a) and the references 

therein [e.g., Dimitriades (1974) and Pitts et al. (1977)] for further 

information on reactivities of specific compounds. The reader is also 

referred to a more recent and comprehensive assessment of the present 

literature on the reactivity and volatility of 118 organic chemicals by 

Singh et al. (1984) in which a three-tiered classification scheme was 

developed based on the potential for involvement of a given chemical in 

photochemical air pollution formation. 

Since a key reaction of volatile organic compounds in ambient air, 

regardless of their class, is their oxidation via attack by hydroxyl 

radicals (Atkinson et al., 1979, 1982c; Atkinson, 1985), the basis of 

several proposed reactivity classifications is the rate of reaction 

between an organic compound and the OH radical. This reaction is the 

first step in a chain reaction that is propagated by various organic 

peroxy radicals. As discussed in section 3.3.2.1, reaction with the OH 

radical is thought to be the predominant loss process for most organics in 

the troposphere (Atkinson et al., 1979; Atkinson, 1985). On this basis, a 

five-class reactivity scale was proposed by Pitts and coworkers (Darnall 

et al., 1976; Pitts et al., 1977) based on the rate of reaction of more 

than 100 VOC with OH radicals. In this scale, each class spanned an order 

of magnitude in reactivity relative to methane, with Class I corresponding 

to an atmospheric half-life of greater than 10 days and Class V a halflife of less than 2.5 hours. The scale has the advantage that any 

3-13 
compound whose OH rate constant has been measured can be placed in a 

precise position in the scale. It has a number of limitations (Pitts et 

al., 1977), however, since it makes the implicit assumption that OH ' 

radical reaction is the sole loss process for an organic, and that the 

subsequent atmospheric chemistry is identical for all organic compounds. 

Both these limitations and the advantages of the OH radical reactivity 

scale are discussed in detail elsewhere (Pitts et al., 1977, 1985). 

3.2.3 Nitrogen Oxides 

The physical and chemical properties of the nitrogen oxides that 

serve as precursors in the formation of ozone and other photochemical 

oxidants have been documented in a recent air quality criteria document 

(U. S. Environmental Protection Agency, 1982a). The most pei°tinent 

properties are briefly summarized here. The role of nitrogen oxides in • 

the formation of oxidants in the troposphere is discussed in Section 3.3 

and in the document cited above. 

The three most abundant oxides of nitrogen in ambient air are nitric 

oxide (NO), nitrogen dioxide (NO 2 ), and nitrous oxide (N20). The latter, 

a product of soil microbiology, is not known to participate in photochemical reactions in the troposphere. The two important oxides of 

nitrogen relative to photochemical processes in the troposphere are NO and 

NO2 , which are abundant in ambient air and participate in cyclic reactions 

leading to the production of ozone and other oxidants, as described later 

in this chapter. 

The basic reactions of importance are (1) the photolysis of NO 2 

< 

430 nm); (2) subsequent formation of ozone from the reaction of atomic 

oxygen [0( 3P)] produced from the photolysis of NO2 with 0 2 (in the .• 

presence of a third, energy-absorbing molecule); and (3) the subsequent 

regeneration of NO2 by the reaction of NO with 03 . Coupled with these 

basic reactions are reactions between NO and free radicais iri the 

atmosphere (hydroperoxy, alkylperoxy and acylperoxy) that oxidize NO to 

NO2 , disturbing the NO-NO 2 equilibrium that would otherwise exist, and 

leading, then, to the buildup of 0 3 . These reactions, and further 

information on the source of the free radicals, are given below. Basic 

physical and chemical properties of NO and NO 2 are given in Table 3-5. 

3-14 
TABLE.3-5. PHYSICAL AND CHEMICAL PROPERTIES 

OF NITRIC OXIDE AND NITROGEN DIOXIDE 

Property NO NO2 

Odor None Pungent 

Taste - - 

Color None Reddish-brown 

Absorption <230 Broad range, 

x; nma both >400 

and <400 

Other Uneven number of Corrosive, strong oxidant. 

properties valence electrons Photolyzes at x<430 nm. 

of note: Low partiai pressure in ambient air. 

Uneven number of valence electrons. 

Forms dimers (N204 ). 

a Visible light x>400 nm; ultraviolet x<400 nm. Solar UV radiation in 

-the troposphere extends from about x290 nm to about x400 nm. 

Source: Derived from National Research Council (1977b) and U. S. Environmental Protection Agency (1982a). 

3.3 ATMOSPHERIC CHEMICAL PROCESSES:, FORMATION AND TRANSFORMATION OF 

OZONE AND OTHER PHOTOCHEMICAL OXIDANTS 

The photochemistry of the polluted atmosphere is exceedingly 

complex. Even if one considers only a single hydrocarbon pollutant, with 

typical concentrations of nitrogen oxides, carbon monoxide, water vapor, 

and other trace components of air, several hundred chemical reactions are 

inVolved in a realistic assessment of the chemical evolution of such a 

system. The.actual urban atmosphere contains not just one but hundreds of 

different hydrocarbons, each with its own reactivity and oxidation 

products. 

(National Research Council, 1977a) 

Despite the complexities of the chemistry of polluted atmospheres, it 

is sufficient to understand certain basic processes involved in the formation 

of photochemical oxidants from precursor compounds in the presence of sunlight. The concentrations of ozone and other oxidants found in urban areas 

and in downwind and rural receptor regions are the net result of at least 

three general processes: (1) the initial emission, dispersion, and transport 

of precursors; (2) the photochemical reactions that occur in the atmosphere 

3-15 
as the dispersion and transport take place; and (3) the scavenging processes 

along the trajectory that reduce the concentrations both of precursors and 

the resulting oxidants. 

Ozone (03 ) is formed in ambient air through the addition of an atom of 

oxygen (0) to a molecule of oxygen (0 2 ). The breakdown by sunlight (photolysis) of nitrogen dioxide (NO 2 ) into nitric oxide (NO) and atomic oxygen provide 

the atoms of oxygen involved. The NO formed in this reaction then reacts with 

the 03produced from the reaction between atomic and molecular oxygen. In 

these cyclic reactions, no net increase in 0 3occurs, with the result that an 

equilibrium is set up among 0 31 NO21 and N0. Any reactions that produce NO 2 

without destroying 03will upset this equilibrium, however, and will result in 

a net increase in 0 3 . In ambient air, the oxidation of photochemically reactive 

hydrocarbons and other nonmethane organic compounds (NMOC) provides a source of 

reactive species (radicals) that convert NO to NO 2without destroying 0 31 thus 

upsetting the equilibrium. The reactions of these radicals with NO also constitute a cyclic process. Since terminating reactions occur between NO 2and these 

radicals, as well, which remove both NO 2and the radicals from the photochemica1 

reaction system, the cycles described above would gradually end, even in the 

presence of sunlight, unless fresh NO xemissions were injected into the atmosphere. The complexity of these cyclic, coupled reactions is such that ozone 

concentrations in ambient air are a nonlinear function of the NMOC and NO x 

concentrations and, within realistic ranges of precursor concentrations, of the 

NMOC:NOxratio. 

In the following sections, the processes just described in a simplistic and 

summary manner are presented in more detail. For a complete and thorough discussion of the many complex reactions thought to take place in polluted atmospheres, the primary literature should be consulted (e.g., Demerjian et al., 1974; 

Finlayson and Pitts, 1976; Logan et al., 1981; Whitten, 1983; Atkinson and Lloyd, 

1984; and Atkinson, 1985). 

3.3.1 Inorganic Reactions 

3.3.1.1 Formation of Ozone: The NO-N0,-0, Cycle. Many aspects of the 

inorganic reaction systems in the atmosphere are now well understood. The 

photodissociation of NO 2by near-ultraviolet solar radiation is a critical 

process: 

NO2+ hv (295<-X<430 nm) 3 NO + 0( 3 P) (3-1) 

3-16 
The subsequent reaction of the resulting 0( 3 P) atom with molecular oxygen 

produces an ozone molecule: 

0( 3 P) + 02 + M-> 03 + M (3-2) 

where M is a"third body" molecule (e.g., N 2 ) which can carry away excess 

energy of reaction. In the absence of any competing reactions, the rapid 

reaction of NO with 0 3 completes this reaction cycle, regenerating an NO2 

molecule: 

NO + 0 3 + NO 2 + 02 (3-3) 

As a result of the above three reactions, an equilibrium or steady-state 

condition is established among N0, NO 2 and 03 , and the concentration of 03 

in the atmosphere is governed by the expression, 

[NO2 ] 

[03 1 = K NO (3-4) 

where K= k3-1/k3-3 which depends on the sunlight intensity. Typically, K 

in the lower troposphere is less than or equal to 0.025 ppm. 

Because reaction (3-3) is rapid, ozone concentrations in urban 

atmospheres cannot rise until most of the NO has been converted to NO 2 . 

This accounts in part for the fact that 0 3 levels may be lower on average 

in city centers where high NO emissions occur, but higher in downwind 

suburban areas to which the resulting NO2 is transported and then photodissociated, leading to 03 formation. The characteristic behavior of 

irradiated NMOC-NOx systems in producing 03 and other photochemical 

oxidants is shown in Figure 3-1, which depicts data obtained from an 

environmental chamber irradiation of a propene-NO-NO 2 mixture (Pitts et 

al., 1979). 

3.3.1.2Formation of Radical Intermediates . Reactions (3-1) through 

(3-3), however, cannot by themselves explain the buildup of ozone, since 

for each molecule of NO oxidized to NO2 in reaction (3-3) a molecule of 

ozone is also destroyed. An alternate pathway of conversion of NO to NO 2 

that does not destroy 03 is needed to explain the high ozone levels 

observed in urban environments. Such an aiternate pathway is available 

through the oxidation of reactive organic compounds. In the atmosphere, 

these compounds can be oxidized by ozone (0 3 ) and/or hydroxyl radicals (OH). 

3-17 
0.6 

E 0.s 

a 

1. 

~ 0.4 

Q 

~ 0.3 

z 

w L) 

z 0.2 

0 

C.) 

0.1 

C7 

0 60 120 180 240 300 360 420 

ELAPSED TIME, minutes 

Figure 3-1. Experimental time-concentration profiles for propene, NO, NO2, 

Og, HCHO, and PAN for an irradiated NOX propene-air mixture. 

Source: Pitts et al. (1979). 

3-18 
3.3.1.2.1 Hydroxyl and Hydroperoxyl Radicals. There are at least three 

significant formation routes leading to the production of hydroxyl 

radicals in the atmosphere. A pathway for OH radical formation, which 

becomes important in the afternoon as ozone concentrations rise, is the 

photolysis of 0 3 : 

0 3 + hv (%<319 nm) -> 0( 1 D) + 0 2 ( 1 og) (3-5) 

The electronically excited 0( 1 D) atoms may be quenched to ground state 

0( 3 P) atoms or may react with water vapor to yield OH radicals with an 

approximately 20 percent efficiency at 298 K and 50 percent relative 

humidity: 

0( 1 D) + H20 + 2,OH (3-6) ` 

Nitrous acid, which has been shown to accumulate to concentrations of 

—1 to 8 parts-per-billion (ppb) during the night in the Los Angeles basin 

(Platt et al., 1980a; Harris et al., 1982; Pitts et a1., 1984c) will 

photolyze at sunrise, producing a"pulse" of OH radicals (Harris et al., 

1982) : 

HONO + hv (x<400 nm) -* OH + NO (3-7) 

This photolytic reaction represents a major sink for HONO during daylight 

hours. (Other aspects of the formation and atmospheric chemistry of this 

important species are discussed in Section 3.3.1.4.) . 

A third significant source of OH radicals is the photolysis of HCHO: 

H + HCO (3-8a) 

HCHO + hv (%<370 nm) 

H2 + CO (3-8b) 

Formald-ehyde is both a primary (e.g., from motor vehicle exhausts) and 

secondary pollutant that may occur in significant concentrations in the 

morning hours as well as in the afternoon (Tuazon et al., 1978a, 1981a; 

Grosjean, 1982). 

3-19 
The H atoms formed in reaction (3-8a) or from reactions such as 

(3-9): 

OH + CO -> CO 2 + H (3-9) 

can react with oxygen to produce hydroperoxyl radicals: 

M 

H + 02 + H0 2 + M (3-10) 

These can then react with NO to form hydroxyl radicals: 

H02 + NO + OH + NO 2 (3-11) 

Reaction (3-11) then completes a chain reaction involving reactions (3-9), 

(3-10) and (3-11) and is a major pathway for the oxidation of nitric oxide 

in ambient air. 

The formyl radical in reaction (3-8a) may also serve as a precursor 

to H02 radicals and hence OH radical formation: 

HCO + 02 + H0 2 + CO (3-12) 

Other sources of forrrRyl radicals include the photolysis of higher 

aldehydes: 

RCHO + hv (%<350 nm) -* R+ HCO (3-13) 

In addition to reactions (3-11) and (3-12), H02 radicals are produced 

by H-atom abstraction from alkoxy radicals as discussed below: 

RCH2O + 02 + RCHO + H0 2 (3-14) 

These H02 radicals will also oxidize NO to NO2 via reaction (3-11). 

Based upon environmental chamber data, computer modeling studies and 

measured ambient concentrations of unreactive organics such as fluorochlorocarbons, concentrations of OH and H0 2 radicals in polluted 

atmospheres are believed to be in the ranges 5 x 10 5 to 5 x 10 6 and 108 to 

109 radicals cm-3 , respectively. 

3-20 
3.3.1.2.2Nitrate Radicals . Ozone can react with NO 2 to pr.oduce the 

nitrate radical and an oxygen molecule: 

03 + NO2 + NO3 + 02 ; (3-15) 

however, because of its large photolytic cross section (Graham and 

Johnston, 1978; Magnotta and Johnston, 1980) the NO3 radical photolyzes 

rapidly in sunlight: 

NO3 + hv + NO2 + 0( 3 P) (3-16a) 

+ NO + 0 2 (3-16b) 

Direct spectroscopic measurements (Platt et al., 1980b, 1984; Noxon, et al., 

1980; Pitts et al., 1984c) have confirmed that NO 3 radical concentrations only 

rise above part-per-trillion (ppt) levels after sunset. The atmospheric 

reactions of this important radical intermediate are discussed in Sections 

3.3.1.5 and 3.3.2.3, including the fact that at night the NO 3 radical will 

participate in a rapid equilibrium between NO 2 , NO3 and dinitrogen 

pentoxide (N20 5 )- 

3.3.1.3Termination Reactions . Although the photochemical reactions 

described above require sunlight, the presence of sunlight does not mean 

that the reactions continue indefinitely. Terminating reactions gradually 

remove NO and NO2 from the reaction mixtures such that the cycles would 

slowly come to an end unless fresh NO x emissions were injected into the 

atmosphere. Specifically, the inorganic chemistry system includes termination reactions for OH and H0 2 radicals with NO and NO 2 to form nitrogen 

acids such as nitrous acid (HONO), nitric acid (HNO 3 ) and peroxynitric 

acid (H02NO 2 ): 

M 

OH + NO + HONO (3-17) 

M 

OH + NO 2 + HNO3 (3-18) 

M 

H0 2 + NO 2 ; H02 NO2 (3-19) 

3-21 
Pernitric acid, however, thermally back-dissociates rapidly (ti 1/2 —10 sec 

at 298 K) and HONO photolyzes, so that under typical atmospheric 

conditions OH + NO 2 is the major sink of NOx . 

Nitrous acid and nitric acid have now both been reliably measured in 

ambient air by longpath spectroscopic techniques and, in the case of 

nitric acid by other techniques as well. Peroxynitric acid has not yet 

been observed in the atmosphere, although it has been predicted to be 

present at f ractional ppb concentrations and has been extensively studied 

in laboratory systems (Graham et a1.,.1977; Hanst and Gay, 1977; Howard, 

1977; Niki et al., 1977; Graham-et a1.,,1978). 

In the presence of N0, radical-radical reactions are generally not of 

major importance in the atmosphere, because concentrations of the radicals 

are low. In the absence of N0, however (for example at night), the 

reactions of peroxy radicals with H0 2 and with other peroxy and acylperoxy 

radicals, and the self-combination of H0 2 radicals, can become 

important. For example, the reaction of H02 : 

H02 + H02 + H2O2 + 02 (3-20) 

could be an important route to the oxidation of S02 to sulfate in solution 

(see section 3.3.4). 

Clearly, termination of the chain reactions can lead to the formation 

of other oxidants as well as relatively stable organic nitrates in the 

atmosphere. In addition to HONO, HNO 3 , H02 NO2 and H2 O

21 these oxidants 

include other peroxyacyl nitrates, organic hydroperoxides and organic 

peracids which have been observed either in polluted atmospheres or in 

irradiated laboratory mixtures (National Research Council, 1977a). These 

compounds almost always occur in low concentrations in ambient air, but 

they may play a significant or even criti.cal role in atmospheric chemistry 

(Pitts et al., 1983) as shown in the following sections. 

3.3.1.4 Reactions Involving Nitrous Acid. An alternative pathway for 

reaction (3-19) is the formation of. HONO: 

H02 + NO2 + HONO + 02 (3-21) 

3-22 
This reaction pathway has been shown, however (Graham et al., 1977, 1978; 

Howard, 1977), to be negligible compared to reaction (3-19). 

The equilibrium between N0, NO 2 , and H 20 results in s-till another 

potential source of HONO: 

NO + NO 2 + H20 ~ 2 HONO (3-22) 

2 HONO -> NO + NO 2 + H20 (3-23) 

Reactions (3-22) and (3 - 23) may proceed both homogeneously and hetero- ` 

geneously, but they appear to be too slow to be of atmospheric importance 

at part- per-million concentrations of NO x . 

Similarly, the reaction of NO2 with water may proceed in the gas 

phase or on surfaces: 

2 NO2 + H20 -> HONO + HNO 3 (3-24) 

Recent work (Sakamaki et al., 1983; Pitts et al., 1984a) has shown that 

HONO is produced in environmental chambers from the reaction of NO 2 with 

water vapor, almost certainly via heterogeneous processes (although HNO 3 

was not observed in these studies). This process may be a minor source 

of HONO in the atmosphere and in the exhaust plumes from combustion.sources 

(Pitts et al., 1984b). 

3.3.1.5 Reactions Involvina Nitric Acid and Dinitroaen Pentoxide. 

Another equilibrium reaction of importance is that involving NO 2 ; NO3 , and 

N205 : 

M 

NO2 + NO3 I> N205 (3-25a;b) 

The equilibrium constant for this system has been measured by several 

groups (Graham and Johnston, 1978; Kircher et al., 1984; Tuazon et al., 

1984; Perner et al., 1985) and appears to be 3(±0.5) x 10-11 cm3 molecule- 1 

at 298 K, which corresponds to a time to reach equilibrium of about one 

minute. The equilibrium constant remains somewhat uncertain, however. 

Dinitrogen pentoxide is a potentially important precursor to HNO 3 

(and hence acid deposition) through its reaction with water either in the 

gas phase (Tuazon et al., 1983) or on surfaces (Heikes and Thompson, 1983): 

3-23 
N205 + H20 -> 2 HNO 3 (3-26) 

Thus, for NO 2 and NO3 radical concentrations representative of receptor 

sites downwind from major urban areas such as Los Angeles (Pitts et al., 

1983, 1984c; Platt et al., 1984), and using an NO 2 + NO3 ~ N70 5 

equilibrium constant of 3 x 10 -11 cm3 molecule -1 , an HNO3 formation rate 

of several ppb hr-1 is obtained at ti50 percent relative humidity [assuming 

the upper limit rate constant of 1.3 x 10 -21 cm3 molecule -1 sec-1 for 

reaction (3-26)]. 

This estimated nighttime formation rate of HNO 3 via rear.tion (3-26) 

can be compared to a calculated daytime formation rate of til ppb hr -1 from 

reaction (3-18) for ti20 ppb of NO2 and 1 x 10 6 molecule cm-3 of OH radicals. 

Reaction (3-26) could potentially be an important loss process for NO x and a 

significant nighttime pathway for HNO 3 formation in urban atmospheres. ~ 

3.3.2Organic Reactions 

It is now well recognized that all organic compounds emitted into the 

atmosphere may be degraded by one or more of the following four pathways: 

reaction with hydroxyl radicals, reaction with ozone, reaction with 

nitrate radicals, or photolysis. Indeed, knowledge of the rates and 

mechanisms of these processes has advanced to the point tha -t the process that 

will predominate for a given compound can be predicted with a reasonable 

degree of certainty. 

This progress notwithstanding, there remain substantial differences 

in the degree to which the detailed atmospheric chemistry is understood 

for the principal classes of hydrocarbons found in polluted atmospheres: 

alkanes (paraffins), alkenes (olefins), and aromatics. Thus, the photooxidation reactions of the smaller alkanes and the simple alkenes, such 

as ethene, propene, and trans -2-butene, are fairly well understood. There 

is much less certainty, however, about the detailed reactions undergone by 

the higher alkanes, the higher alkenes, and the aromatics subsequent to 

their initial reactions with OH radicals, ozone, or NO 3 radicals. 

The following sections briefly summarize the basic features of the 

four reaction pathways identified above for organic compounds emitted into 

the atmosphere. 

3-24 
3.3.2.1 Reactions with Hydroxyl Radicals. The following sections treat 

separately the mechanisms of reaction of OH radicals with the major 

classes of organic compounds including alkanes, alkenes, aromatics and 

oxygenated compounds, as well as nitrogen- and sulfur-containing 

compounds. The treatment here of relevant reactions is necessarily an 

overview. For a comprehensive and current detailed description of the 

kinetics and mechanisms of the atmospheric reactions of OH radicals with 

organic compounds, the reader is referred to a review by Atkinson (1985). 

3.3.2.1.1 Alkanes. It is now well estabiished that the only significant 

atmospheric chemical loss process for the alkanes is reaction with OH 

radicals. These reactions proceed by hydrogen abstraction (Atkinson, 

1985) to produce alkyl radicals (R), which then add 0 2 to form alkyl 

peroxy radicals (R0 2 ): 

OH + RH + R' + H2 0 (3-27) 

M 

R' + 0 2 + R02 ' (3-28) 

In polluted atmospheres R0 2 radicals rapidly oxidize NO to NO 21 

forining alkoxy radicals (RO); or add NO 2 to form alkyl peroxynitrates: 

RU 2 ' + NO ->. RO' + NU 2 (3-29) 

R02 ' + NO 2 E R02 NO 2 (3-30a,b) 

The latter, however, are not expected to be present in ambient air at 

significant concentrations because of their short (51 sec at 298 K) 

lifetimes with respect to thermal decomposition (reaction 3-30b). 

Alkoxy radicals may also undergo hydrogen abstraction by molecular 

oxygen to form aldehydes or ketones (Baldwin et al., 1977): 

R'RCHO' + 02 -> R'OR + H02 ' (3-31) 

or they may decompose to form oxygenates (Baldwin et al., 1977; Batt, 

1979): 

R' RCHO' + RCHO + R " (3-32) 

R'CHO + R' (3-33) 

3-25 
In both of these reaction sequences, however, H0 2 radicals are formed, and 

hence OH radicals are regenerated. The carbonyl compounds thus formed may 

subsequently react with OH radicals or may photodecompose (see Section 

3.3.2.1.4). 

A further important reaction pathway for acyl radicals is the 

addition of 02, followed by reaction with NO2 to form peroxyacyl nitrates: 

0 0 

RC' + 0 2 -> RC00' (3-34) 

0 0 

RC00' + NO 2 E R COONO 2 (3-35) 

The simplest member of this class of compounds, peroxyacetyl nitrate, has 

been measured in polluted atmospheres throughout the world (see Chapter 5). 

The reactions described above suggest the importance of simple (<C 4 ) 

alkoxy and alkylperoxy radicals in atmospheric chemistry. In addition to 

these reactions for the simple alkanes, two other processes involving 

RO' and R0 2 ' reactions occur for larger (>C4) alkanes, namely alkoxy 

radical isomerization (Carter et al., 1976; Baldwin et al., 1977; Hendry 

et al., 1978; Batt, 1979; Batt and Robinson, 1979; Carter et al., 1979) 

and alkyl nitrate formation from the reaction of R0 2 ' with NO (Atkinson et 

al., 1982a, 1983, 1984f): 

M 

R0 2 ' + NO + RONO 2 (3-36) 

Discussion of these processes is beyond the scope of this chapter and the 

reader is referred to the original literature and to appropriate reviews 

(e.g., Atkinson, 1985) for summaries of these aspects of the atmospheric 

chemistry of longer-chain alkanes. 

3.3.2.1.2Alkenes . In polluted atmospheres, unsaturated hydrocarbons 

react primarily with OH radicals and with 0 3 (Herron and Huie, 1977, 1978; 

Dodye and Arnts, 1979; Akimoto et al., 1980; Kan et al., 1981; Niki et 

al., 1981; Atkinson et al., 1982c; Whitten, 1983; Atkinson, 1985). For 

most alkenes studied to date, the reaction with OH radicals proceeds 

3-26 
almost entirely by addition to the double bond. In the case of propene, 

for exampl.e, addition of the OH radical to the double bond is expected to 

be followed by 02 addition, with the oxidation of NO to NO 2 ,by the 

resulting peroxy radical to form an alkoxy radical (Atkinson et al., 1985): 

OH' + CH3CH=CH2 + CH3 CHCH2OH (3-37) 

00' 

. I 

CH3 CHCH2OH + 02 + CH3CHCH2OH (3-38) 

00' 0 ' 

CH3 CHCH2OH + NO + CH3CHCH2 OH + NO 2 (3-39) 

Decomposition of the alkoxy radical and subsequent reactions lead to 

the formation of acetaldehyde and formaldehyde, both of which can be 

detected in polluted atmospheres: 

CH3 CHCH2OH + CH3 CH0 + 'CH2 OH (3-40) 

'CH2OH + 02 -). HCHO + H0 2 (3-41) 

t NO 

OH + NO 2 

The overall reaction resulting from reactions (3-37) to (3-41) is: 

OH + CH3CH=CH2 + 202 + 2 NO -* CH3CH0 + HCHO + 2 NO 2 + OH (3-42) 

1 hv, 02 

0 3 

Thus, the reaction of OH radicals with alkenes increases the rate of NOto-NO2 conversion and hence increases the yield of ozone. The specific 

reaction sequence for the OH radical-initiated oxidation of propene 

(Atkinson, 1985) is shown in Figure 3-2. 

3-27 
- 0 

I 

CH3 CHCH2OH 

CH3CH0+ CH2OH 

02 

CH3CHOH +HCHO 

0 2 

OH 

CH3CHCH 2O' 

CH3CH=CH2 

(^-65% 

CH3CHCH2OH 

(-35%) 

7

H 

CH3CHCH 2

. 

0 2 0 2 

NO NO2 NO 31,NO2 

H02+HCHO CH3CH0 + H02 

Figure 3-2. Reaction scheme for OH radical-initiated oxidation of propene 

in the presence of NO. 

Source: Atkinson (1985). 

3-28 
The reaction schemes presented for propene illustrate the major ro2e 

that oryanic compounds (not only alkenes but also alkanes and aromatics) 

play in produciny photochemical air pollution, namely acceleration of the 

conversion of NO to NO 2 and the resulting formation of ozone. 

Alkenes Emitted Frorn Vegetation . A special class of alkenes 

receiving considerable attention over the past decade are those 

unsubstituted compounds emitted from vegetation. Examples include isoprene and monoterpenes such as a- and S-pinene, d-limonene, and myrcene. 

Much research and discussion have been devoted to assessments of the 

potential for such compounds, which are emitted from vegetation in large 

quantities, to contribute to photochemical air pollution (Coffey, 1977; 

Westberg, 1977; Arnts and Gay, 1979; Tingey and Burns, 1980; Bufalini and 

Arnts, 1981; Dimitriades, 1981; Altshuller, 1983), but a detailed treatment of this topic is beyond the scope of this document. Presented here, 

and in other appropriate parts of section 3.3, are relevant aspects of 

current knowledge of the atmospheric chemistry of organic compounds known 

to be emitted from vegetation. 

Although reliable rate constants for the reaction of these naturally 

emitted organic compounds with OH radicals are now available (Winer et 

al., 1976; Kleindienst et al., 1982; Atkinson, 1985), with the exception 

of isoprene their detailed atmospheric chemistry is still not well 

characterized. Summarized briefly here are the OH radical-initiated 

photooxidation reactions for isoprene and a-pinene. 

Isoprene . Based on the relative atmospheric concentrations of UH 

radicals and ozone, and their rate constants for reaction with isoprene, 

the dominant atmospheric reaction pathway for isoprene is expected to be 

OH radical addition to the olefinic double bonds (Lloyd et al., 1983): 

IH3 

r--> HOCH2-C-CH=CH 2 (3-43a) 

j H3 

OH + CH 2=C-CH=CH 2 

j H3 

CH 2=C-CH-CH 2 

3-29 OH 

j H 3 

CH2 ~-CH=CH2 (3-43b) 

OH 

jH3 

CH2=C-CHCH2 OH (3-43c) 

( 3-43d ) 
The hydroxyalkyl radicals formed in reaction,s (3-43a)-(3-43d) react . 

rapidly with 02 to form peroxy radicals, which can then rapidly oxidize N0. 

to NO2 (Atkinson and Lloyd, 1984). The resulting hydroxyalkoxy radicals 

decompose to form methyl vinyl ketone and methacrolein. The subsequent 

atmospheric reactions of these products are described in detail by Lloyd 

et al. (1983) and by Killus and Whitten (1984). 

a-Pinene. Based on the rate constant for its reaction with OH 

radicals (Atkinson et al., 1979; Kleindienst et al., 1982), a-pinene is 

expected to react exclusively by OH radical addition at the least 

substituted carbon atoms, with the resulting radical rapidly adding 02 

(Lloyd et al., 1983): 

CH3 CH 00' 

OH 

~ 0 2 H. 

OH + --> ( 3-44) 

A mechanism for the subsequent reaction pathways of this hydroxyperoxy 

radical has been reported (Lloyd et al., 1983), but it was of necessity 

largely parameterized because of the lack of data on the reaction products 

resulting from the photooxidation of a-pinene under atmospheric 

conditions. Even less information is available for other monoterpenes and 

any detailed consideration of their atmospheric chemistry would be highly : 

speculative. 

For further descriptions of the OH radical-initiated p,hotooxidations 

of isoprene and monoterpene, the reader should consult the-..primary 

literature (Lloyd et al., 1983; Killus and Whitten, 1984). 

3.3.2.1.3 Aromatics. The aromatic fraction,in gasolines has increased in 

recent years, partly as a result of the reduction of lead in gasoline. 

Given that there are also substantial emissions of aromatic compounds 

(e.g., benzene, toluene, xylenes, etc.) from a wide range of industrial 

processes, the importance of aromatics in the hydrocarbon distribution in 

ambient atmospheres has grown. 

Reactions with OH radicals constitute the sole atmospheric loss 

process for aromatic compounds. The available kinetic and mechanistic 

3-30 
data concerniny these reactions have been reviewed critically (Atkinson, 

1985). It is clear from these data that two reaction pathways are 

possible. The first of these is OH radical addition to the aromatic riny 

to form an initially energy-rich OH-aromatic adduct which can either 

decompose back to the reactants or be collisionally stabilized. Th -is is 

iTlustrated for toluene, an abu.ndant aromatic constituent in urban 

atmospheres: 

CH

(60 

CH 

OH 

H 

OH + ~ (plus other isomers) (3-45) 

For alkyl-substituted benzenes, the second reaction pathway involves . 

H-atom abstraction from the substituent group: 

0 

H3 CH2 

OH + 0 ~ H

2

0 + O (3-46) 

(A) 

- This latter is a minor (<10%) process at room temperature (Atkinson, 

1985). ~ The reaction pathways subsequent to the H-atom abstraction 

reaction pathway (3-46) are reasonably well understood {Atkinson and 

Lloyd, 1984). Thus, under atmospheric conditions the benzyl radical is 

expected to react via the following sequence of reactions: 

CH2 CH2O0' 

0O + 02-~ ~ ( 3-47 ) 

3-31 
H200' CHZO' 

O + NO + O + NO2 ( 3-48a') 

CH2ONO2 

M O + (3-48b) 

with reaction (3-48b) occurring approximately 10 percent of the time at 

atmospheric pressure and room temperature (Hoshino et al., 1978). 

The C6H5CH2O' radical then reacts with 0 2 to yield benzaldehyde and an H02 

radical: 

CH2O' HO 

O + 02 ~ O + H02 (3-49) 

Analogous reaction pathways are expected to be applicable to the other 

aromatic hydrocarbons, after H-atom abstraction from the substituent alkyl 

groups (Atkinson and Lloyd, 1984). 

The rates of reaction of OH radicals with aromatics are now well 

characterized (Atkinson, 1985), as is the relative importance of the 

alternative pathways (3-45) vs (3-46), at 'least for selected aromatics. 

In the case of toluene, for example, OH radical addition is expected to 

occur —80 percent of the time at the ortho position (Kenley et al., 1981). 

The fate, however, of the addition adducts formed from OH radicalaromatic reactions remains unclear (Atkinson, 1985), although several 

mechanisms have been proposed in the case of toluene (Atkinson et al., 

1980; Leone et al., 1985). Although substantial progress has been made in 

understanding reaction mechanisms for toluene and certain other aromatic 

hydrocarbons (Atkinson and Lloyd, 1984; Leone et al., 1985), much 

additional research is needed before a complete understanding of the 

complex NOx-photooxidation chemistry of aromatic compounds is obtained. 

3-32 
3.3.2.1.4 Aldehydes. Aldehydes are consumed in the atmosphere both by 

photolysis and attack by OH radicals. Photolysis of acetaldehyde, for 

example, leads to methyl and formyl radicals that then react as discussed 

above: 

CH3 CH0 + hv + CH 3 +.HCO (3-50) 

Attack by OH radicals forms acetyl radicals which can successively add 0 2 

and NO2 to form PAN: 

CH3 CH0 + OH -~ CH3 C0 + H2 0 (3-51) 

11 NO2 i . 

CH3 C0 + 02 + CH3 C00' CH3 COONO 2 (3-52) 

3.3.2.1.5 Nitrogen-Containing Compounds. Only limited information is 

available for the reactions of.OH radicals with nitrogen-containing 

compounds and their subsequent reactions under atmospheric conditions. 

The OH radical reactions with the aliphatic amines are rapid, with room 

temperature rate constants being in the range (24) x 10 -11 cm3 molecule -1 

sec -1 (Atkinson et al., 1985). For the methyl-substituted amines, the 

trend of the room temperature rate constants suggests that these reactions 

proceed via abstraction from the C-H bonds and, where possible, the' N-H 

bonds. Product studies of several irradiated amine-air systems have been 

reported in which plausible reaction,pathways following OH radical attack 

have been proposed (Pitts et al., 1978; Tuazon et al., 1978b; Lindley et 

al., 1979). 

To date, kinetic data are available only for OH radical reactions 

with hydrazine and methylhydrazine (Atkinson, 1985), and only limited 

product data are available for these reactions (Tuazon et al., 1981b, 

1982). Reactions of nitrites with OH radicals are expected to proceed via 

H-atom abstraction from the C-H bonds but, since no product data are 

presently available, no reliable assessment of the initial reaction 

pathway can be made. 

No product or direct mechanistic data are available for organic 

nitrates. However, the reactions of OH radicals with at least the smaller 

3-33 
alkyl nitrates (for which isomerization of the alkoxy radicals cannot 

occur) will probably ultimately yield NO 2 together with the corresponding 

aldehydes (Atkinson et al., 1982d). These reactions may be of importance 

in long-range transport and acid deposition, since alkyl nitrates are 

formed in significant yields from the atmospheric photooxidation of 

certain alkanes (Atkinson and Lloyd, 1984). 

3.3.2.1.6 Sulfur-Containing Compounds. The literature concerning reactions of sulfur-containing compounds with OH radicals has been reviewed 

recently by Atkinson (1985). Reactions of thiols with OH radicals must 

proceed via either H-atom abstraction from the weak S-H bonds or, more 

likely, by the formation of an OH-thiol adduct. The reaction of OH 

radicals with the sulfides, RSR, can proceed via either H-atom abstraction 

from the C-H bonds or OH radical addition to the sulfur atom. Product 

data for the reaction of OH radicals with dimethyl sulfide under 

atmospheric conditions have been obtained from numerous studies (Grosjean 

and Lewis, 1982; Hatakeyama et al., 1982; Hatakeyama and Akimoto, 1983; 

Niki et al., 1983; Grosjean, 1984); the major stable products are HCHO, 

S02 and CH3S03H, together with CH3SN0 as an intermediate product. 

Only for diinethyl disulfide have kinetic (Cox and Sheppard, 1980; 

Wine et al., 1981) and product (Hatakeyama and Akimoto, 1983) data been 

reported. On the basis of these data, it appears that the initial 

reaction proceeds via OH radical addition to form an adduct, followed by 

rapid decomposition of this adduct to CH 3S and CH3SOH radicals.' 

Subsequent reactions of these CH 3SOH and CH3S radicals then lead to the 

observed products: S02, HCHO, and CH 3S03 H. 

3.3.2.2 Reactions with Ozone. The atmospheric reactions of ozone are 

complex and result in products and processes that have significant 

environmental implications, including effects on biological systems, 

visibility, and materials. Ozone, for example, is highly reactive towards 

certain classes of organic compounds (e.g., alkenes) and certain of those 

reactions lead to the formation of secondary organic aerosols. Ozone may 

also play a role in the oxidation of SO Z to H2SO4 , both indirectly in the 

gas phase (via formation of OH radicals and Criegee biradicals) and 

directly in aqueous droplets. 

In the following sections, the atmospheric reactions of ozone with 

organic compounds are summarized in some detail, including the mechanisms 

3-34 
of certain of .these reactions. Emphasis is placed, whenever possible, on 

those reactions that lead to products or processes suspected or known to 

have effects on biological or other important receptors. 

In discussing the reactions of ozone with organic compounds in the 

troposphere, it is important to recognize that organics undergo competing 

reactions with OH radicals during daytime hours (Atkinson and Lloyd, 1984; 

Atkinson, 1985) and, in certain cases, they can photolyze or react with 

NO3 radicals at night (Japar and Niki, 1975; Carter et al., 1981a; 

Atkinson et al., 1984a,b,c,d; Winer et al., 1984). A11 organics except 

the perhaloalkanes exhibit room temperature OH radical rate constants`of , 

>5 x 10-15 cm3 molecule -1 sec -1 (Atkinson, 1985). Since the ratio of 03 

to OH radical concentrations in the unpolluted troposphere during daylight 

hours is believed to be of the order of 10 6 (Singh et al., 1978; Crutzen, 

1982), only for those organics whose 0 3 reaction rate constants are 

greater than —10 -21 cm3 molecule-1 sec-1 can consumption by 03 be 

considered atmospherically;important. These ozone reactions of interest 

are summarized below. 

3.3.2.2.1Alkenes . Ozone reacts rapidly with the acyclic mono-, di- and 

trialkenes and cyclic mono-, di-, and tri-alkenes. The rate constants for 

these reactions range from —10 -18 to —10-14 cm3 molecule -1 sec -1 (Atkinson 

and Carter, 1984), corresponding to atmospheric lifetimes ranging from a 

few minutes (for the more reactive cyclic alkenes such as the 

monoterpenes) to several days. In polluted atmospheres, especially in the 

afternoons during photochemical oxidant episodes, the principal 

consumption of the more reactive alkenes will therefore occur via reaction 

with 03 , rather than with OH radicals. 

It is now reasonably well established that the initial step in the 

03-alkene reaction involves the formation of a"molozonide" that rapidly 

decomposes (Harding and Goddard, 1978; Herron et al., 1982) to a carbonyl 

compound and a biradical (which is also initially energy rich): 

3-35 
0 + R1~C=C / R3 3 z ` 

R2 

--I R4 

o 

R 

* R1\ 0/ !0 / 

3 

+ R/ C--C ~ 

R 2 4 

a b 

(3-53) 

Rl-

~-' 

00-0~ / R3 ~ 

/C=0 + C ~ 

R2 R4 

~ 

R1~, / 0-0' ~ R3 

R / C• + O=C\ R 2 4 

where [P denotes an energy-rich species. 

8ased on an analysis of reported product and mechanistic studies of 

the reactions of 03 with ethene (Herron and Huie, 1977; Su et a1.,, 1980; 

Kan et al., 1981; Niki et al., 1981) and propene (Herron and Huie, 1978; 

Dodge and Arnts, 1979), and the much less extensive studies of the higher 

alkenes (Martinez et al., 1981), Atkinson and Lloyd (1984) have suggested 

that these initially energy-rich biradicals react under atmospheric 

conditions as shown below: 

M. . 

CH2O0 

[CH2O0]* 

/0]* 

/0' 

CH2~0~ [CH2`--, 0

. f 

/ CO2 + H2 

~ [HCOOH] *--! CO + H2 0 

~ H + HCO •2 

(40%) 

(129'0 ) 

(42%) (3-54) 

20 2 

2 HO ?+ CO 2(6%) 

and 

3-36 
m • e 

-> CH3CH00 (40%) 

i 

[CH 3CH00]~ 

0 ~ CH3 + CO + OH (19%) 

[CH3 COH]* ~ (3-55) 

CH3 + CO 2 + H (24°0) 

CH 0 * 0 

~C ~ [HCOCH 3]~ 

H HCO + CH30 ( 5%) 

CH4 + CO 2 (12%) 

. e • s 

where CH2O0 and CH 3CH00 denote thermalized biradicals. These thermalized 

biradicals have been shown (Calvert et al., 1978; Herron et al., 1982; 

Atkinson and Lloyd, 1984) to undergo bimolecular reactions with aldehydes, 

S0 2 1 CO 3 and H2O, and it is believed that they will also react with NO.and 

NO2 (Calvert et al., 1978;' Herron et al., 1982; Atkinson and Lloyd, 1984): 

- RCH00 + NO + RCHO + NO 2 (3-56) 

RCH00 + NO 2 -> RCHO + NO 3 (3-57) 

H20 

RCH00 + S0 2—, RCHO + H2 SO4 (3-58) 

RCN00 + H20 -> RCOOH + H 20 (3-59) 

RCH00 + CO products (3-60) 

0 —0 

RCH00 + R'CHO ~ RCH ~Oj CHR' (3-61) 

Under atmospheric conditions, the reactions with N0, NO 2 , or H20 are 

expected to be the dominant loss processes of these thermaiized 

biradicals, with the precise major reaction pathway depending on the 

relative concentrations of N0, NO 2 or H20 (Atkinson and Lloyd, 1984). 

3-37 
Hence, 03-alkene reactions in the atmosphere can lead ultimately to the 

formation of aldehydes and acids, as well as to the conversion of S0 2 to 

H2SO4 , although the latter is probably a minor process in the overall 

oxidation of S02 during long-range transport (Finlayson-Pitts and Pitts, 

1982). 

Isoprene and monoterpenes are alkenes emitted from vegetation. The 

role of these compounds in photochemical air pollution has been a subject 

of discussion, and sometimes controversy, for more than a decade. The 

case of ozone reactions with compounds such as isoprene and the monoterpenes 

is particularly interesting since these reactions can represent a sink for 

ozone as well as for the hydrocarbons themselves. This adds complexity 

to an overall assessment of the role of hydrocarbons emitted from 

vegetation since, depending upon the specific atmospheric conditions, 

they may be both sources and sinks for ozone (Dimitriades, 1981; Altshuller, 

1983). 

Again, as in the case of their OH radical reactions, the detailed 

reaction sequences following reaction of 0 3 with iosprene and the 

monoterpenes are not well understood although substantial kinetic data are 

available (Atkinson and Carter, 1984); thus only a brief suminary of 

available information is presented here. 

Isoprene. The reaction of 0 3 with isoprene (Kamens et al., 1982; 

Lloyd et al., 1983; Killus and Whitten, 1984) leads to molozonides that 

are presumed to decompose subsequently into stable products and r•adical 

intermediates analogous to those produced in other ozone-alkene reactions 

described earlier in this section. The major products, methlvinylketone 

and methacrolein, undergo further reactions with OH radicals and ozone 

(Lloyd et al., 1983; Killus and Whitten, 1984). 

The reaction of ozone with isoprene can also lead to aerosol 

formation (Kamens et al., 1982) and this is discussed in Section 3.3.4. 

Monoterpenes. The detailed pathway for the reaction of 0 3 with the 

monoterpenes under atmospheric conditions is unknown. Lloyd et al. (1983) 

have proposed a reaction sequence for a-pinene involving addition of 0 3 to 

the double bond to form a molozonide, with subsequent ring opening. They 

note, however, that this reaction sequence is entirely speculative, and 

3-38 
many additional kinetic and mechanistic data will be required to elucidate 

the detailed reactions of 03 with the monoterpenes. . 

3.3.2.2.2 Alkanes and Alkynes. Given reported rate constants of 10 -23 to 

10-26 cm3 molecule -1 sec -1 (Atkinson and Carter, 1984), there appears to 

be no convincing evidence in the literature for an elementary reaction 

between 03 and the alkanes. Similarly, although there is presently 

substantial uncertainty concerning the rate constants for the reactions of 

ozone with the simple alkynes (e.g., acetylene, propyne and 1-butyne), 

most of the available room temperature data for these <C 4 alkynes indicate 

03 rate constants in the range of —10 -20 to 10-19 cm3 molecule -1 sec -1 

(Atkinson and Carter, 1984). Thus, those reactions will also be 

unimportant in the atmosphere since the corresponding OH radical rate 

constants are, for example, —8 x 10 -13 for acetylene and —7 x 10 -1 2 cm3 

a molecule -1 sec-1 for propyne and 1-butyne (Atkinson, 1985). 

3.3.2.2.3 Aromatics. As in the case of the alkanes, the aromatic 

hydrocarbons react only very slowly with 0 3 (Atkinson and Carter, 1984) 

and these reactions are not expected to be.important in the atmosphere. 

Although the cresols are significantly more reactive than the aromatic 

hydrocarbons (Atkinson and Carter, 1984), under atmospheric conditions 

their reactions with 03 are minor compared to their reactions with OH 

radicals (Atkinson et al., 1978) or NO3 radicals (Carter et al., 1981a). 

3.3.2.2.4 Oxygen-Containing Organics. For those oxygen-containing 

compounds that contain no unsaturated carbon-carbon bonds (e.g., 

formaldehyde, acetaldehyde, glyoxal, and methylglyoxal) the reactions with 

ozone are very slow, and, by analogy, this is expected to be the case for 

all ethers, alcohols, aldehydes, and ketones containing no unsaturated 

carbon-carbon bonds. For the carbonyls and ethers (other than ketene) 

that contain unsaturated carbon-carbon bonds, however, much faster 

reactions are observed (Atkinson and Carter, 1984). 

Few data are available, however, concerning the mechanisms of the 

reactions of 0 3 with such oxygen-containing organics, the only published 

information being that of Kamens et al. (1982). From a study of the 

reactions of 03 with methacrolein and methylvinylketone, methylglyoxal was 

observed as a product, along with other minor products (Kamens et al., 

1982), as anticipated from the reaction schemes: 

3-39 
U~0~0 ' 

CH2=CHCOCH3+ 03 + CH2-CHCOCH3 (3-62). 

HCHO + [CH3COCH00]* CH3 COCH0 + [CH2O0]* (3-63) 

and 

~ 

0~0~0 

/CH3 t 1 /CH 3 CH2=C.` + 03 -~ CH2 -~C 

CHO CHO 

(3-64) 

CHO 

HCHO + CH3C00 CH3COCH0 + [CH2O0] 

Kamens et al. (1982) discuss the possible subsequent reactions. 

3..3.2.2.5Nitrogen-Containing Organics . Studies of the kinetics of the 

reactions of 03 with a variety of nitrites, nitriles, nitramines, nitrosoamines, amines, and hydrazines (Atkinson and Carter, 1984) indicate that 

only for the hydrazines are reactions with 03 sufficiently rapid to be of 

atmospheric importance (Carter et al., 1981b; Tuazon et a1., 1982). 

3.3.2.2.6Sulfur-Containing Organics . Based upon the kinetic data 

available for dimethyl sulfide, thiirane (C 2 H4S) and thiophene, it'appears 

at the present time that the rates of reaction of 03 with sulfurcontaining organics can be considered to be unimportant under• atmospheric 

conditions (Atkinson and Carter, 1984). 

3.3.2.2.7Organometallics . Rate constants have been reported only for` 

tetramethyl- and tetraethyl-lead (Harrison and Laxen, 1978), and no 

mechanistic or product data are available for these reactions. 

3.3.2.2.8 Radical Species . Because of the low. concentrati©n of 0 3 and 

of both alkyl and most alkoxy radicals in the atmosphere, and because 

these radicals react at significant rates with 02 (which is present at a 

concentration >10 5 higher than 03 in ambient atmospheres), the reactions 

of ozone with such radical species can be considered to be of negligible 

importance in the atmosphere. Of course, the reaction of ozone with the 

hydroxyl radical must be considered, as discussed earlier. 

3-40 
3.3.2.3 Reactions with Nitrate Radicals. The pioneering work of Niki and 

coworkers (Morris and Niki, 1974; Japar and Niki, 1975) showed that 

gaseous NO 3 radicals react with alkenes, with the rate constants 

increasing markedly with the degree of substitution on the double b,ond. 

Carter et al. (1981a) showed that the hydroxy-substituted aromatics 

(phenols and the cresols) also react rapidly with NO3 radicals. 

More recently, Atkinson and coworkers . have investigated the kinetics 

of the reactions of NO3 radicals with a wide range of organics at room 

temperature (Atkinson et al., 1984a-e) and f rom these and the earlier 

studies, information concerning the mechanisms of these reactions has been 

forthcoming. 

In the remainder of this section, current understanding of the 

mechanisms of reaction of NO 3 radicals with the various classes of 

organics is briefly summarized. 

3.3.2.3.1 Alkanes. The relevant kinetic data (Atkinson et al., 1984e) indicate that these reactions proceed via H-atom abstraction from the C-H 

bonds, almost certainly,predominantly from secondary or tertiary C-H bonds: 

NO3 + RH + HNO3 + R' 

Hence, these reactions lead directly to HNO 3 formation. The measured room 

temperature rate constants for the alkenes range between 3.6 x 10-17 cm 3 

molecule-1 sec-1 for n-butane to 2.2 x 10 -16 cm3 molecule -1 sec -1 for 2,3- 

dimethylbutane. 

, 

3.3.2.3.2 Alkenes. The reactions of NO3 radicals with the alkenes have 

been shown from.both kinetic (Japar and Niki, 1975; Atkinson et al., 

1984a) and product.(Bandow et al., 1980) studies to proceed via initial S . 

addition of the NO 3 radical to the olefinic double bond: 

ONO2 

~ . . 

2NO3 + 2CH3CH=CH2 

+ 

 CH3CHCH 2 + CH3CHCH2ONO 2 , 

with addition at the terminal carbon expected to dominate (Atkinson and 

Lloyd, 1984). Possible reaction sequences beyond this initial reaction 

~ 

have been discussed (Bandow et al., 1980; Atkinson and Lloyd, 1984) but 

are highly uncertain at the present time (Atkinson and Lloyd, 1984). 

3-41 
Thermally unstable nitro-peroxynitrates such as CH3CH(ONO2 )CH2O0NO2 , and 

stable dinitrates such as CH3CH(ONO2 )CH2ONO2 have been reported as 

products in NO3-NO 2-propene-air systems (Bandow et al., 1980). 

Monoterpenes . While no mechanistic information is available 

concerning the reactions of NO3 radicals with the monoterpenes, NO3 

radical reaction rate constants have recently been reported for a 

substantial number of these compounds (Atkinson et al., 1984c, 1985). 

Based in part on these kinetic data, Winer et al. (1984) have proposed 

that reaction with NO3 radicals at night may be an important reaction 

pathway for certain naturally occurring organics such as the monoterpenes 

and dimethyl sulfide. Conversely, these kinetic data also show that 

reactions with the more reactive alkenes, including isoprene and certain 

of the monoterpenes, as well as dimethyl sulfide and the hydroxysubstituted aromatics, can be important loss processes for NO 3 radicals at 

night (Winer et al., 1984). The importance of NO 3 radical reactions in 

determining the atmospheric lifetimes of the monoterpenes is discussed in 

section 3.3.3. 

3.3.2.3.3Aldehydes . Based upon the product data of Morris and Niki 

(1974), i.e., the observed formation of HNO 3 from the reaction of NO3 

radicals with CH3CH0, it is expected that these reactions proceed via Hatom abstraction from the relatively weak H-CO bonds: 

NO3 + RCHO -> RCO + HNO 3 (3-67) 

Thus the reaction of NO 3 radicals with acetaldehyde could be a nighttime 

source of peroxyacetyl nitrate (PAN): 

NO3 + CH3CH0 -> HNO 3+ CH3C0 (3-68) 

CH3C0 + 02 -> CH3 CO3 (3-69) 

1 CH3 CO3 ' + NO 2 + CH3 COONO 2 (3-70) 

Reaction of NO3 radicals with the higher aldehydes will lead, by analogous 

reaction schemes, to the higher peroxyacyl nitrates,RCO 3 NO2 . Reaction 

3-42 
with formaldehyde, however, will lead to H02 radical formation, since HCO 

reacts rapidly with 02 (Atkinson and Lloyd, 1984): 

HCO + 02 + H02 + CO (3-71) 

3.3.2.3.4 Aromatics. As discussed by Atkinson et al. (1984d), the 

reactions of NO 3 radicals with the monocyciic aromatic hydrocarbons and 

the hydroxy-substituted aromatics appear, based upon kinetic evidence, to 

proceed via H-atom abstraction from the C-H or 0-H bonds on the 

substituent groups. This conclusion is based upon the observation that 

for the xylenes and the cresols the meta-isomer reacts more slowly (by a 

factor of —2) than the ortho- and para-isomers. This is in contrast to 

the addition reactions of 0( 3 P) atoms and OH radicals, in which the metaisomer is the most reactive (Atkinson, 1985). Furthermore, o-nitrophenol 

has been tentatively identified as a product of N 20 5-NO 2-phenol-air 

reaction mixtures, presumably formed by the reaction sequence: 

OH 0' 

NO3 + ~ ~ HNO3 + ~ (3-72) 

foilowed by (Niki et al., 1979): 

' OH OH 

NO2 

O + NO 2 + ~ (3-73) 

NO2 

Thus, these reactions can also be a direct source of nitric acid as well 

as forming low-volatility organic nitro compounds. 

In summary, it is now clear that reaction with NO 3 radicals at night 

is a major atmospheric reaction pathway for many organic pollutants. It 

must therefore be considered, along with the reactions of OH radicals and 

03 and photolysis, as one of the dominant loss processes for organics in 

the atmosphere. a 

3-43 
3.3.3 Atmospheric Lifetimes of Organic Compounds 

Table 3-6 compares the atmospheric lifetimes (i.e., the time to reach 

1/e of the initial concentration) for selected organic compounds, arising` 

f rom manmade and natural sources, as the result of reaction with 0 3 over a 

24-hour period, with OH radicals during the day, and with NO 3 radicals at 

night. It can be seen that under the atinospheric conditions assumed reactions with 03 are important for the higher alkenes, including the monoterpenes during daylight when NO3 radical concentrations are low, and for the 

hydrazines. 'For the other organics for which kinetic data are available, 

including alkanes and aromatics, reactions with 03 are generally of negligible or minor importance in determining their atmospheric lifetimes. 

TABLE 3-6. CALCULATED LIFETIMES OF SELECTED ORGANIC COMPOUNDS RESULTING 

FROM ATMOSPHERIC LOSS BY REACTION WITH 0 3 AND OH AND NO3 RADICALS 

Organic lifetimes a,b 

Organic compound p3

, 

OH, NO31 '. 

24 hr daytime, nighttime 

Alkenes from manmade sources 

Ethene 2.7 days 16 hr 79 days 

Propene 11 hr 5.6 hr 1.1 days 

trans-2-Butene 35 min 2.0 hr 33 min 

2-Methyl-2-butene 17 min 1.6 hr 1.3 min 

2,3-Dimethyl-2-butene 6 min 1.3 hr 0.2 min 

Naturall_v emitted alkenes 

Isoprene 10 hr 1.4 hr 22 min 

a-Pinene 1.4 hr 2.3 hr 2 min 

p-Pinene 5.5 hr 1.8 hr 5 min 

n3-Carene 1.0 hr 1.7 hr 1.2 min 

d-Limonene 11 min 1.0 -hr 0.9 min 

aTime to reach 1/e of the initial concentration. 

bAssuming 100 ppb of 03 (24 hr average), 2 x 10 6molecule cm -3(0.08 ppt) 

of OH radicals during daylight hours, and 100 ppt of NO3 radicals during 

nighttime hours, and at room temperature. 

344 
3.3.4 Atmospheric Reactions of Peroxyacetyl Nitrate 

With the recognition in recent years that PAN is a ubiquitous 

nitr,ogenous species in•the troposphere (Singh and Hanst, 1981; Aikin et 

.al., 1983; Penkett, 1983; Singh and Salas, 1983; Spicer et al., .1983) and 

in the lower stratosphere (Aikin et. al., 1983), there has been renewed 

focus on the atmospheric role of this oryanic compound. 

Smog chamber studies have shown that, once formed, PAN can be 

relatively stable under atmospheric thermal conditions (Pitts et al ; ., 

1979; Akimoto et a1.., 1980). Since PAN is in equilibrium, however, with 

acetylperoxy radicals and NO2: 

0 

it 

CH3 COONO2E C H3 C00 + NO2 ( 3-74) 

any process that removes either acetyl peroxy radicals or NO2 will lead to 

the decomposition.of PAN. One such process is the reaction of NO with 

CH3C(0)02 radicals. Because PAN has been shown to persist through the 

night in urban atmospheres (Tuazon et al., 1980, 1981a) the reaction of 

PAN with NO during the-morning traffic peak can lead to the formation of 

OH radicals via the following,mechanism (Cox and itoffey, 1977;. Carter et 

al., 1981c): 

0 

CH LO , + NO _> CH CO' + NO 3 3 2 

0 

CH3 C!-0' -> CH3 ' + CO2 

M 

CH3 ' + 02 -> CH3 00 

CH3 00' + NO + CH3 0' + NO2 

CH3 0' + 02 ~ H02+ HCHO 

(3-76) 

(3-77) 

(3-78) 

(3-79) 

H02 + NO -> OH + NO2 (3-80) 

3-45 
M 

OH + NO -> HONO (3-81) 

M 

OH + NO 2 + HNO3 (3-82) 

Thus, the reaction with NO of PAN carried over from previous air 

pollution episodes will lead to enhanced smog formation on subsequent 

days. This enhancement in reactivity results both from the fact that 

these reactions form radicals that initiate the transformations occurring 

in photochemical smog and from the fact that these reactions convert NO to 

NO2 , which allows earlier formation of 0 3 and higher levels to be 

attained. It should be noted that this enhancement will result even if 

all of the PAN reacts with NO emitted at nighttime, since the NO 

conversion does not require sunlight and since at least some of the 

radicals formed will be "stored" as nitrous acid, to be released when 

photolysis begins at sunrise (Harris et al., 1982). 

These results could have important implications regarding multiday 

photochemical pollution episodes in which significant buildup of PAN is 

observed. Under such conditions, the carry-over of PAN may be a 

significant factor in promoting ozone formation on subsequent days and 

may, in part, contribute to the progressively higher 0 3 levels often 

observed during such episodes (Tuazon et al., 1981a). 

A second important role of PAN is its ability to contribute to the 

long-range transport of NO x . In the absence of significant levels of NO 

(i.e., in the cleaner troposphere) and in regions of lower temperature and 

in the upper troposphere, when the thermal decomposition of PAN becomes 

unimportant, the atmospheric lifetime of PAN will be determined by its 

reaction with OH radicals. This reaction is sufficiently slow (Wallington 

et al., 1984) that PAN will probably be long-lived and hence serve as a 

reservoir for odd nitrogen in a manner analogous to HNO 3 (Aikin et al., 

1983). Also analogous to the case for HNO3 , dry deposition o -F PAN may be 

a significant loss process in cleaner atmospheres. 

3-46 
3..3.5 Rol e of Ozone i n Aerosol Formati on 

In addition to having direct effects on human health and on vegetat1on, ecosystems, and nonbiological materials, ozone can contribute 

i:ndirectly to visibility degradation and to acidic deposition through its 

participation in the formation of both organic and inorganic aerosols 

(National Research Council, 1977a; U. S. Environmental Protection Agency, 

1982b). 

3.3.5.1 Formation of Sulfate Aerosol. It is well established that the 

source of the vast majority of manmade sulfate aerosol in the atmosphere 

is the oxidation of sulfur dioxide (S0 2 ), ultimately to sulfuric acid 

(H2SO4 ). The correlations between elevated levels of ozone and of sulfate 

aerosol in ambient air have been noted by several investigators in field 

studies concerned with visibility reduction by aerosols. Wilson (1978) 

and Gillani et al. (1981) have pointed out that atmospheric mixing 

intensity and the background 0 3 concentration are the two most important 

factors in determining S02 oxidation at relative humidities lower than 75 

percent. It is also clear, however, that the rate of reaction of 0 3 with 

S02 is far too slow to account for observed formation rates of sulfate 

aerosol (U. S. Environmental Protection Agency, 1982b). 

Of the many possible gas-phase reactions of S02, only a few appear to 

have any significance in the production of sulfate aerosol and the 

reaction of OH radicals with S02 appears to be the dominant pathway for 

the oxidation of S0 2 (Calvert and Stockwell, 1983, 1984; Calvert and Mohnen, 

1983). A recent analysis by Stockwell and Calvert (1983) shows that the 

formation of HOS02 radicals from the reaction of OH radicals with S0 2 , 

followed by reaction with 02 , is the reaction mechanism for the formation 

of S03 : 

OH + S02 (+ M) + HOS02 (+ M); (3-83) 

HOS0 2 + 02 -> H0 2 + S03 . (3-84) 

From the reaction of H02 radicals with N0, OH radicals are regenerated and 

the cycle begins again as discussed earlier in Section 3.3.1.2.1. 

3-47 
The importance of the reaction of OH radicals with S02 in the 

atmosphere is supported by observations of power plant plumes, in which n.o 

aerosol is formed at night when the OH radical concentration in ambient 

air is negligible; and none is formed during the day before the plume is 

well mixed with ambient air (the ambient air contains much higher 

concentrations of OH radicals and 03 than the plume) (Davis et al., 1979; 

Blumenthal et al., 1981). 

Though it does not react directly with S02 at an appreciable rate, by 

virtue of its role in OH radical production 03 plays an important indirect 

role in the transformation of S02 to sulfate aerosol via the homogeneous 

oxidation of S02 in both clean and polluted atmospheric systems. Ozone 

plays a further role in the oxidation of sulfur in aqueous droplets as 

discussed later in this section. 

3.3.5.2 Formation of Nitrate Aerosol.. Despite limited relevant data, 

the possible contribution of nitrate aerosol to visibility reduction 

should not be neglected and the role of 0 3 in the formation of this 

aerosol species is briefly considered here. 

The principal manmade nitrogen emissions in this case are NO and 

NO2 . Nitric oxide is relatively insoluble in aqueous systems (Section 

3.2) and does not react with water in any significant manner. Thus, NO 

must be converted to a more highly oxidized form, for example NO 2 , in 

order to participate in the formation of particulate nitrate. 

The oxidation of NO to NO 2 can occur through thermal oxidation at 

high concentrations of NO such as those in and very near the stacks of 

power plants (U. S. Environmental Protection Agency, 1982a). This 

generates only a small portion of the NO2 formed in the atmosphere, 

however. As previously'discussed (Section 3.3.1), the most important 

reactions leading to formation of NO 2 in ambient air are the reaction of 

NO with 03 and the oxidation of NO.to NO2 by hydroperoxyl radicals and 

other peroxy radicals (reactions 3-3 and 3-11, respectively). Thus, if 

NO2 is a precursor of nitrate aerosol, 0 3 plays a significant direct role 

in its formation by oxidizing N0, and an indirect role by leading to 

formation of UH radicals (Section 3.3.1). 

As discussed earlier, NO2 can be converted in the gas phase to nitric 

acid (HNO3) vapor by reaction with OH radicals during the day or by 

3-48 
reaction with 0 3 to form NO 3 radicals, which at night are in equilibrium with 

N20 5 . As shown in Section 3.3.1.5, homogeneous or heterogeneous hydrolysis, or 

both, of N 20 5 is an important nighttime pathway to nitric acid formation. Once 

it has been produced in the gas phase, HNO 3 is sufficiently volatile to remain 

in the atmosphere as a vapor. The available laboratory and ambient air data 

indicate, however, that HNO 3 vapor reacts with ammonia in a reversible reaction 

to form NH4NO3 (Doyle et al., 1979; Stelson et al., 1979; Appel et al., 1980), 

which, because of its low vapor pressure, will form nitrate aerosol particles: 

HNO3 + NH 3 --> NH4NO3 (3-85) 

If acidic sulfate is present, however, it will react with NH 4NO 3 to form HNO 3 

again. Consequently, reaction (3-85) is not a major sink for nitric acid in 

areas with high sulfate loading, such as the eastern United States. Evidence 

also indicates that HNO 3 vapor will:react with NaCI aerosol in the following 

way : 

HNO3 + NaCI -). NaNO 3 + HCl (3-86) 

This second reaction (equation 3-86) may account for the fact that much of the 

observed particulate nitrate in Los Angeles is.found in the coarse mode (Farber 

et al., 1982). Obviously, the importance of this mechanism for nitrate aerosol 

formation is determined by the availability of sea salt particles. 

3.3.5.3 Formation of Organic Aerosols. Sulfate and nitrate aerosols are 

present at significant levels in the atmosphere in the form of just a few 

compounds. In contrast, secondary organic aerosols are composed of a large 

number of species, but there is no clear consensus concerning which ones 

contribute most to the mass concentration. For all the species that are found 

in the secondary organic aerosol,.however, the fundamental formation mechanism 

is the same. The vapor-phase precursor undergoes some reaction that results 

in formation of a product having an equilibrium vapor pressure sufficiently 

low that condensation, nucleation, or both are possible at the gaseous concentration achieved. From the available data, it seems clear that the.more highly 

oxygenated, larger-carbon-number species generally are those precursors likely 

to form secondary aerosols in the atmosphere. 

3-49 
For an earlier but thorough review of the formation of secondary 

aerosol, the reader is referred to the 1977 monograph on ozone and other 

photochemical oxidants prepared by the National Research Council 

(1977a). This monograph reviews the reactions of manmade volatile organic 

compounds that produce aerosol. Biogenic as well as manmade volatile 

organic compounds, however, can participate in aerosol formation 

(Altshuller and Bufalini, 1971; Arnts and Gay, 1979). Direct experimental 

evidence of aerosol formation, along with product analysis, is only available, though, for a limited number of natural compounds, including isoarene (Kamens et al., 1982) and a-pinene and p-pinene (Schwartz, 1974; 

National Research Council, 1977a; Hull, 1981), mainly because the 

analysis and characterization of these kinds of products at ambient concentrations is extremely difficult. Hull (1981) has conducted experiments 

with a- and p-pinene at high concentrations in a small tube reactor. 

Analysis of the products showed, on a weight basis, that almost all of the 

reacted a-pinene carbon was found in the condensed materials extracted 

from the walls. Although the products identified from these experiments were either in the condensed phase or on the walls, Hull suggested 

that at the a-pinene levels found in ambient air these products have a 

high enough vapor pressure to exist both in the gas phase and in aerosols 

(Hull, 1981). In a recent review of the role of biogenic volatile organic 

compounds, Altshuller (1983) has discussed at length the contribution of 

these compounds to ambient air aerosols. 

3.3.6kole of Ozone and Other Photochemical Oxidants in the Acidification 

of Rai n 

Two recent criteria documents prepared by the U. S. Environinental 

Protection Agency (1982a; 1982b) contain thorough discussions of the 

contributions of ozone and of hydrogen peroxide, also an oxidant, to the 

oxidation of S0 2 and the roles of S0 2 as a precursor to acidic deposition., 

3.3.6.1Reactions of Ozone in Aqueous Droplets . While the thermal oxidation 

of S02 by ozone in the gas phase appears to be too slow to be important in 

acid deposition phenomena, the role of ozone in oxidizing S0 2 dissolved in 

water droplets (e.g., cloud, fog or rain) may be of considerable 

significance. At 25 0C ozone has a Henry's law constant of 10 -2 mol L-1 

atm-1 (Kirk-Othrner, 1981). Given ambient concentrations ranging from 30 

3-50 
to —300 ppb, 0 3would be expected to have concentrations in aqueous 

droplets in the atmosphere of approximately 3-30 x 10 -10 mol L-1 . The 

rate of reaction between 03and S02, when both are dissolved in aqueous 

droplets, has been shown in laboratory studies to be relatively fast 

(Penkett et al., '1979; Kunen et al., 1983; Brock and Durham, 1984; Hoffman 

and Jacob, 1984; Martin, 1984; Schwartz, 1984), but the rate of this 

reaction is pH dependent and decreases as the acidity of the solution 

increases. 

Figure 3-3 shows data reported by Schwartz (1984) for the rate of the 

aqueous-phase oxidation of S(IV) by 30 ppb of 0 3(and also by 1 ppb of 

H2O2) as a function of solution pH. The aqueous-phase oxidation rate, R, 

per part-per-billion S0 2partial pressure decreases with decreasing pH by 

roughly a factor of 20 per pH unit. This pH dependence reflects the 

solubility of S(IV), as well as a slight pH dependence of the second-order 

rate constant for the oxidation of S(IV) by 0 3(Erickson et al., 1977; 

Larson et al., 1978; Penkett et al., 1979). Schwartz (1984) concluded, 

from consideration of these data and uptake times for S0 2 , that the 

oxidation of S02by 03cannot produce solution pH values below —4.5. 

Schwartz (1984) has also, however, interpreted the field data of Hegg and 

Hobbs (1981) for sulfate production rates at the inf1ow and outflow 

regions of lenticular clouds as being consistent with the aqueous phase 

oxidations of S(IV) by 03. 

An additional aspect of the role of 0 3 in the chemistry of aqueous 

droplets concerns its photolysis to yield OH radicals in solution (Graedel 

and Weschler, 1981; Chameides and Davis, 1982): 

(0 3 ) aq+ hv 0( 1 D) aq + 02 aq (3-87) 

0( 1 D) aq+ (H2 0) aq 

 

2(OH) aq (3-88) 

and its reactions with aqueous OH - ions and H2O2to yield aqueous H0 2 

radicals (Chameides and Davis, 1982). The OH radicals formed by this 

aqueous process can result in the oxidation of S(IV). 

3.3.6.2 Reactions of Hydrogen Peroxide in Aqueous Droplets. Although 

hydrogen peroxide formed in the gas phase from the reactions of 

3-51 
1000 

100 

r 

M 

E 

10 M E ~ 

~ 

.`c 

0 

1 • ~ 

~ 

c. 

0.1 

0.01 

, 10-s 

10-4 

~ 

0 10-c 

~ 

.a 

a 

c. 

~ 

L 

s 10-6 

~ 

N 

O 

m 

a 

1-1 

CC 10-7 

10'b 

10-O 

2 3 4 5 6 

pH 

Figure 3-3. Rate of aqueous-phase oxidation 

of S(IV) by 03 (30 ppb) and H2O2 0 ppb), as a 

function of solution pH. Gas-mqueous equilibria 

are assumed for all reagents. R/pS02 represents aqueous reaction rate per ppb of gasphase S02; p/L represents rate of reaction 

referred to gas-phase S02 partial pressure per 

cm3—m-3 liquid water volume fraction. 

Source: Schwartz (1984). 

3-52 
hydroperoxyl radicals plays a role in HO x chemistry in the troposphere, 

and especially in the stratosphere (Crutzen and Fishman, 1977; Cox and 

Burrows, 1979), the major importance of hydrogen peroxlde arises f rom its 

high solubility in water. The latter ensures that a large fraction of 

gaseous H2O2 will be taken up in aqueous droplets. Over the past decade, 

evidence has accumulated that H 2O2 _dissolved in cloud, fog, and rainwater 

may play an important, and, in acidic droplets (i.e., pH <5), even a 

dominant role in the oxidation of S02.to H2SO4 (Hoffman and Edwards, 1975; 

Penkett et al., 1979; Dasgupta 1980a,b; Graedel and Weschler, 1981; Martin 

and Damschen, 1981; Charneides and Davis, 1982; Calvert and Stock'wel1, 

1983, 1984; Brock and Du'rham, 1984; Hoffman and Jacob, 1984; Schwartz, 

1984). Discussion of several proposed mechanisms for previous rate 

studies of the oxidation of S(IV) by H2O2 are beyond the scope of this 

document, but have recently been reviewed by several authors (see for 

example, Calvert and Stockwell, 1983, 1984). Hydrogen peroxide may also 

play a role in the oxidation of NO 2 dissolved in aqueous droplets, 

although relevant data are limited (Halfpenny and Robinson, 1952a,b; Anbar 

and Taube, 1954; Gertler et al., 1984) and additional research is 

required. In addition to the direct oxidation of S02 and NO2 dissolved in 

aqueous droplets, the photolysis of H 2O2 to produce aqueous OH radicals: 

(H2O2 ) aq + hv + 2(OH) aq (3-89) 

can lead to oxidation rates of S(IV) that can be competitive with calcu= 

lated oxidation rates of S(IV) by (H2O 2 ) aq and (03 ) aq (Chameides and 

Davis, 1982). 

It should be emphasized, however, that substantial uncertainties 

remain concerning the quantitative role of H 2O2 in the acidification of 

aqueous particles and droplets .(Richards et.al ., 1983). This is further 

complicated by the lack of reliable measurements of gas-phase H2O2 

concentrations in the atmosphere (see Chapter 4). Moreover, it has also 

been suggested recently that H 2O2 may be formed insituin aqueous 

droplets as the result of absorption of OH and H0 2 radicals and other 

precursors into solution from the gas phase (Graedel and Weschler, 1981; 

Chameides and Davis, 1982; Heikes et al., 1982). 

3-53 
3.3.6.3 Reactions of Formic Acid in Aqueous Droplets. As a gas-phase species, 

formic acid (HCOOH) cannot strictly be defined as 'a photochemical oxidant. It 

can be scavenged rapidly into water droplets, however, and can potentially 

function therefore as an oxidant in cloud water and rain water. It can also 

be di fferenti ated from other aci ds i n that i t i s formed readi 1y from the 

reactions of the Criegee intermediates discussed earlier and of hydroperoxyl 

radicals with formaldehyde (Calvert and Stockwell, 1983). The formation of 

other acids may be orders of magnitude slower as the result of both the 

apparently lower rates of reaction of H0 2 radicals with the higher aldehydes 

and the much lower atmospheric concentrations of the higher aldehydes (Grosjean, 

1982). Thus, formic acid is an example of a compound that is a non-oxidant or 

weak oxidant in the gas phase but that is transformed upon incorporation in 

aqueous solutions into an effective oxidizer of S(IV). 

Formic acid (as well as acetic acid) has been identified among the acidic 

components of rain (Galloway et al., 1982). Although much uncertainty remains 

concerning their quantitative roles, HCOOH and the higher organic acids potentially 

play a minor but still significant role in the acidification of rain. 

3.4 METEOROLOGICAL AND CLIMATOLOGICAL PROCESSES 

As discussed in Section 3.3, ozone and oxidants are formed by the action 

of sunlight on the precursors, NO 2 and hydrocarbons. The accumulation of the 

products to form an appreciable concentration is also dependent, however, on 

the prevailing meteorology in the vicinity of the precursor emissions. To 

understand the details of the effects of ineteorology on air quality requires a 

thorough knowledge of ineteorology and climatology, but an appreciation of the 

general factors important in the formation of elevated concentrations of 

oxidants is relatively easy to acquire. Following is a brief presentation of 

some features of atmospheric mixing and transport that will provide a basic 

understanding of the meteorological factors that affect the concentrations of 

ozone and other oxidants in urban and rural areas. 

3.4.1 Atmospheric Mixing 

The concentration of an air pollutant depends significantly on the degree 

of mixing that occurs between the time a pollutant or its precursors are 

emitted and the arrival of the pollutant at the receptor. Since, to a first 

3-54 
approximation, the diurnal cycle of weekday urban emission patterns for ozone 

and oxidant precursor pollutants is generaily uniform, it is reasonable to 

ascribe a significant proportion of the large day-to-day changes in pollutant 

concentrations to changes in meteorological mixing processes. The rate at 

which atmospheric mixing processes occur and the extent of the final dilution 

of the pollutants depends on the amount of turbulent mixing and on wind speed 

and wind direction. Moreover, the transport of pollutants and precursors from 

a source region to a distant receptor is also dependent on wind speed and wind 

direction. 

The degree of turbulent mixing can be characterized by atmospheric stability. In an atmospheric layer with relatively low turbulence, pollutants do 

not spread as rapidly as they do in an'unstable layer. Also, because a stable 

layer has a relatively low rate of mixing, pollutants in a lower layer will 

not mix through it to higher altitudes. The stable layer can act as a trap 

for air pollutants lying beneath it. Hence, an elevated inversion is often 

referred to as a"trapping" inversion. Also, if pollutants are emitted into a 

stable layer aloft, such as from a stack, the lack of turbulence will keep the 

effluents from reaching the ground while the inversion persists. 

In air pollution considerations, a stable atmospheric layer or situation 

is usually equated with a temperature inversion, which is a layer of the 

atmosphere in which the temperature increases with increasing altitude, because 

inversions are common and also represent the most stable atmospheric conditions. 

The lowest part of an inversion layer is called the base and is defined as the 

altitude at which the temperature begins to increase. The top of the inversion 

is the point at which the temperature begins to decrease with increasing 

altitude. The distance between the base and top of the inversion layer is the 

"depth" or "thickness" of the inversion. Inversion layers may begin at the 

ground surface (i.e., the altitude of the base is zero), or the entire inversion layer may be above the surface. The former is known as a"surface inversion" and the latter as an "elevated inversion." The two types are usually 

caused by different sets of weather conditions, but it is not unusual for both 

types of inversions to be present at a given location at the same time. In 

the United States, surface inversions are characteristic of nighttime and 

early morning hours except when heavy cloud cover or windy and stormy conditions 

prevail. 

Surface and elevated inversion layers are both important in determining 

pollutant concentration patterns since, as noted above, mixing and dilution 

3-55 
processes proceed at a relatively slow rate in such layers. Thus, if pollutants are emitted into an inversion layer, relatively high concentrations can 

persist for a considerable period of time or over a considerable distance of 

wind travel from the source. For example, a surface inversion in the morning 

could cause automotive exhaust pollutants released at the surface during the 

morning rush hours to persist with minimum dilution near the ground surface 

for an extended period of time, probably for 1 or 2 hours after sunrise, until 

solar radiation heats the ground and causes the inversion to disappear'or 

"break" (Hosler, 1961; Slade, 1968). High concentrations may occur at the 

ground even when an elevated inversion is present and the layers below the 

inversion are unstable and are undergoing good mixing. Such a persistent 

elevated inversion layer is a major meteorological factor that contributes to 

high pollutant concentrations and photochemical smog situations along the 

southern California coast (Holzworth, 1964; Hosler, 1961; Robinson, 1952). 

The vertical mixing profile through the lower layers of the atmosphere 

follows a typical and predictable cycle on a generally clear day. In such a 

situation a surface inversion would be expected to form during the early 

morning and to persist until surface heating becomes significant, probably 2 

or 3 hours after sunrise. Pollutants initially trapped in the surface inversion 

may cause relatively high, local concentrations, but these concentrations will 

decrease rapidly when the surface inversion is broken by surface heating, 

usually about midmorning. The surface inversion will begin to form again 

during the early evening hours and pollutants from near-surface sources such 

as automobiles or home fireplace chimneys will experience progressively less 

dilution as the surface inversion develops. 

Elevated inversions, when the base is above the ground, are also common 

occurrences (Hosler, 1961; Holzworth, 1964). Since these conditions, however, 

are identified with specific synoptic conditions, they are much less frequent 

than the nighttime radiation inversion. Because it may persist throughout the 

day and thus restrict vertical mixing, an elevated inversion is nevertheless a 

very significant air pollution feature. Smog-plagued southern California is 

adversely affected by persistent elevated inversions (Robinson, 1952). Wheri 

compared to a source near the surface and the effects of a radiation (surface) 

inversion, the pollutant dispersion pattern is quite different for an elevated 

source plume trapped in a layer near the base of an elevated inversion. This 

pl ume wi 11 not be i n contact wi th the ground surface i n the early morni ng 

3-56 
hours because there is no mixing downward through the surface radiation inversion. Thus, the elevated plume will not affect surface pollutant concentrations until the mixing processes become strong enough to reach the altitude of 

the plume. At that time, the plume may be mixed downward quite rapidly in a 

process cailed "fumigation.° During fumigation, surface.ozone concentrations . 

will increase if the morning ozone concentration is higher aloft than at.the 

ground, and if insufficient scavenging by NO occurs at ground-level. After 

this initial mixing, surface concentrations wil.l decrease as the us,ual daytime 

mixing processes continue to develop. If the daytime mixing becomes strong 

enough to break the upper inversion, the pollutants may be mixed through an 

increasingly deep layer of the atmosphere. When surface heating decreases in 

the late afternoon and early evening, both the surface and elevated inversions 

will form again. The surface inversion will again prevent pollutants from 

elevated sources from reaching the ground and surface scavenging processes 

will gradually reduce the concentrations of pollutants trapped during the 

formation of the surface inversion. 

Geography can have a significant impact on dispersion of pollutants 

(e.g., along the coast of an ocean or one of the Great,.Lakes). Near the coast 

or shore, the temperatures of land and water masses can be different, as can 

the temperature.of the air above such.land and water masses. When the water 

is warmer than the land, there is a,tendency toward reduction in the.frequency 

of surface inversion conditions inland.over a relatively narrow coastal strip 

(Hosler, 1961). This in turn tends to increase pollutant dispersion in such 

areas. Such conditions may occur frequently on the Gulf Coast and near the 

Great Lakes in the winter. The opposite conditi.on also occurs if . the water is 

cooler than the land, .as in summer or fall. Cool air near the water surface 

will tend to increase the stability of the boundary layer in the coastal zone, 

and thus decrease the mixing processes that act on pollutant emissions,. .These 

conditions occur frequently.along the New England coast (Hosler, 1961). 

Similarly, pollutants from, the Chicago area have been observed repeatedly to 

be influenced by a stable boundary,layer over Lake Michigan (Lyons and Olsson, 

1972). This has been observed especially in summer and fall when the lake 

surface is most likely to be cooler than the.air, that is carried over it from 

the adjacent iand. 

Since the diurnal mixing,conditions are such an important part of the 

meteorological parameters for understanding pollutant mixing and diffusion, it 

3-57 
is useful to have some knowledge of the mixing cycles that prevail over the 

United States. Figures 3-4 and 3-5 show the average summer morning and afternoon mixing heights (AGL, above ground level) as calculated on the basis of 

upper air temperature data and an estimated midmorning urban temperature. 

Since Holzworth (1972) attempted to include the influence of an urban heat 

island in this estimated temperature, the morning results in Figure 3-4 are 

probably most applicable to larger urban areas. Rural or nonurban areas would 

be expected to have lower mixing heights. 

Summer conditions are useful to consider because of the prevalence of 

high photochemical oxidant concentrations during this season. As shown in 

Figure 3-4, morning mixing heights are estimated to be greater than 300 meters 

except for the central part of the Great Basin, where a 200-meter isopleth 

includes parts of Oregon, Idaho, Utah, Arizona, and most of Nevada. By midafternoon (Figure 3-5), the estimated mixing height at the time of maximum 

temperature has increased markedly, and only a few coastal areas have an 

average afternoon maximum mixing height of less than 1000 meters. In contrast 

to the morning data, the central Great Basin area becomes the area of greatest 

mixing in the afternoon. This would be expected since this is a hot, arid, 

desert region, and the driving force generating the surface mixing layer is 

the solar heating of the ground surface. 

The magnitude of the afternoon mixing height is generally an indication 

of the potential for recurring urban air pollution problems. If a trapping, 

elevated inversion does not rise high enough in the afternoon to release the 

generated pollutants that are trapped, an accumulating episode is likely. 

From the average summer afternoon data shown in Figure 3-5, where the lowest 

average mixing height is 600 meters and almost all of the area has a value 

greater than 1500 meters, it would appear that urban air pollution should not 

be severe. On the average this is probably correct; however, there are several 

departures from the average, which result in relatively low mixing heights and 

adverse dispersion over many areas of the United States on a recurring basis. 

Figure 3-6 (Holzworth and Fisher, 1979) shows the frequency of occurrence 

of elevated inversions in summer having a base between 1 and 500 meters (1600 

feet) at the time of the afternoon upper air temperature measurement, 6:10 

p.m. EST or 3:15 p.m. PST. The California coastal conditions, in which low 

inversions occur with a frequency of nearly 90 percent, are clearly evident. 

The northeastern coastal area from New Jersey north to Maine, where cool ocean 

3-58 
Figure 3-4. Isopleths (rn x 102

) of inean summer morning mixing heights, AGL. 

Source: Holzworth (1972). 

s 

ia 

12 

io 

ltl 1s 

Figure 3-5. Isopleths (m x 10 2

) of inean summer afternoon mixing heights, AGL. 

Source: Holzworth (1972). 

3-59 
30 i7 0• 

e1 00 10 1 728 

•2 

ei •2 20 60 13 2• •0 •2 3• 

® 7 

22 

0• 21 

1• 0• •2 3 4 

83 •4 •2 

•1 

3 16 0 0 „ 2• 2 2 

1• • 2 ,. 

•1 •11 •5 

• 86 3 • 15 03 e4 •83 5 0 6• ~ 

2• •3 •1 •5 

10 

• 6• 

2 7 

e3 • 3 • 

3 2 

2 2 

Figure 3-6. Percentage of summer 2315 GMT (6:15 p.m. EST, 3:15 p.m. PST) 

soundings with an elevated inversion base between 1 and 500 m above ground level. 

Source: Adapted from Holzworth and Fisher (1979). 

water prevails, also has a relatively high percentage, above 20 percent, 

compared to most of the rest of the country. Stations bordering one of the 

Great Lakes--Green Bay, Sault St. Marie, and Buffalo--reflect a stabilizing 

lake effect with percentages above 5 percent. Except along the Pacific Coast, 

these ocean and lake coastal situations are probably limited to relatively 

narrow coastal zones (Hosler, 1961). Examples are evident in Figure 3-6, in 

which it may be noted that inversion frequencies of 21 to 28 percent occur in 

coastal New England compared to only 2 percent at Albany in upstate New York. 

A similar situation is evident in a comparison of the 3 percent inversion 

frequency at Washington, D.C., with the 16 percent frequency on the Delaware 

coast. A non-coastal region having summer afternoon low-level elevated inversions more than 5 percent of the time is the Southeast, where an area from 

Louisiana and Arkansas to the Atlantic coast shows frequency values between 5 

and 10 percent. Other seasons differ in details, but the general patterns are 

similar. 

3-60 
This means that, for most of the United States, low-level stable layers 

that persist through the afternoon hours are rare events, occurring`on less 

than 1 day in 20. Thus, air pollution situations in areas such as Kansas or 

Iowa, although they can be the result of transport, may commonly be related to 

the periods when the expected morning surface inversion persists later in the 

morning than usual and when winds are not strong enough to carry pollutants 

rapidly away from the local area. Along both the Pacific Coast and the 

Northeast Coast, low-level afternoon inversions are frequent enough to be a 

significant contributor to local and regional air pollution episodes. 

3.4.2Wind Speed and Mixing 

Another major meteorological factor in the urban pollutant dispersion 

problem is low-level or surface-layer wind. As would be expected, strong 

winds across a source area wi.11 dilute pollutant concentrations even though 

the're is a strong, low-level inversion base. San Francisco is one example of 

such a location where strong winds frequently provide good ventilation in 

spite of a low inversion. Conversely, light and variable or calm wind conditions over an area can iead to excessive pollutant accumulations even though 

the afternoon mixing depth is quite large. Thus, it is necessary to include 

wind direction and wind speed frequencies in any evaluation of air pollution 

potential for a given area. It must also be recognized that both elevated 

inversion conditions and surface wind patterns are governed to a major degree 

by the synoptic, or large-scale, weather patterns. Both wind and inversion 

factors tend to favor pollutant buildup when a deep, slow-moving high-pressure 

system dominates the weather across an area (Korshover, 1967; Korshover, 

1975). 

Figure 3-7 shows the wind ciimatology across the United States in the 

month of July by depicting the monthly resultant vector wind at major weather 

stations (U.S: Department of Commerce, 1968). Note that the flow across the 

West Coast is generaily directed inland, from west to east. This codtributes 

to a typical situation for major California cities: significant urban pollutant plumes are found east of the urban core source areas while the immediate 

coastline or beach areas are relatively poliutant-free. In the Northeast 

States, the average wind flow is from southwest to northeast more or less 

parallel to the coastline. As a result, pollutant plume.s from the major urban 

areas along this coast are frequently additive along the trajectory of the 

3-61 
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3-62 
wind. Polluted air moving toward the coast from major inland urban sources 

may also be a factor in this Northeast region. Along the Gulf Coast, the 

average winds form southerly, onshore flow: Under some weather situations, 

however, there i s often an offshore fl ow i n one area (e. g. , Texas) and an 

onshore flow in an adjacent area. Thus, because of this recirculation, the 

onshore Gulf air masses are not always pollutant-free (Price, 1976; Wolff et 

al., 1981). Before the situation was examined carefully, the recirculating 

pollutants were sometimes confused with natural background concentrations. 

Wind climatology provides an average wind flow pattern, but it does not 

provide a complete assessment of the influences of the wind on air polluti.on 

dispersion. Wind speed and, in particular, the frequency of weak winds are an 

important aspect to be considered. Figure 3-8, adapted from Holzworth and 

Fisher (1979), shows the frequency with which early morning (6:15 a.m. EST or 

3:15 a.m. PST) surface inversions occur with calm or weak surface winds; that 

is, wind speeds equal to or less than 2.5 m/sec or 6 mi/hr. There is considerable variation between stations because terrain and geography (e.g., coastal 

locations) influence both the wind flow and inversion frequency. It is clear, 

however, that over large areas of the United States, especially in heavi1y 

industrialized inland areas east of the Mississippi River, calm amd stab1e 

summer mornings are a frequent occurrence: 50 percent or more in many areas. 

This means that there will be frequent incidents of morning pollutant accumulation; but afternoon heating, as shown by Figure 3-5, will usually mix the 

pollutant accumulations through a deep mixing layer and' disperse them. Figure 

3-9 (Holzworth and Fisher, 1979) shows the average wind speed through the 

depth of the summer morning mixing layer. Note that the area east of the Rocky 

Mountains, except for the Appalachians, can on the average, expect winds of 4 

m/sec (about 10 mi/hr) or higher through the morning mixing layer. This 

probably would provide acceptable midmorning dilution of accumulated pollutants. In summer afternoons, as shown in Figure 3-10, the average wind speed 

within the mixing layer increases in all areas and may even double over some 

of the western mountain states. It should be noted, however, that since winds 

normally increase with altitude above the ground, much of the increase in the 

average afternoon mixing layer wind is probably the result of the considerable 

increase in the depth of the mixed layer, as shown by the differences between 

 Figures 3-4 and 3-5. 

3-63 
4 16 19 • 

28 

13• 50 

• 63 • 20 22 

• 31 

• 77 • 30 479 42 • 35 •16 249 

034 • 27 

440 5 

43 • 6 

160 120 •24 7 • 48 

50 

~ 13 •47 

• 

• 39 57 28 • 20 21 • 18 63 

13 • 10 o 42 

•14 45• 300 

~ 15 

• 9 

98 07 

510 420 •34 

200 • 36 68 

•24 • 490 

30 916 

0 

•7 058 • 

10 12 

__ 41 

Figure 3-8. Percentage of summer 1115 GMT (6:15 a.m. EST, 3:15 a.m. PST) soundings 

with an inversion base at the surface and wind speeds at the surface <2.5 m/sec. 

Source: Adapted from Holzworth and Fisher (1979). 

3-64 
Figure 3-9. Isopleths (m/sec) of inean summer wind,speed averaged 

through the morning mix:ng layer. 

Source: Holzworth and Fisher (1979). 

s s 

Figure 3-1 0. Isopleths (m/sec) of inean summer wind speed averaged 

through the afternoon rnixing layer. 

Source: Holzworth and Fisher (1979). 

3-65 
In summary, atmospheric mixing parameters of stability and wind in the 

pollutant transport layers can exert controlling effects on 03 and oxidant 

concentrations. The effects include the amount of dilution occurring in the 

source area, as well as along the trajectory followed by an urban or sourcearea plume. Regions having steady prevailing winds, such that a given air 

parcel can pass over a number of significant source areas, can develop significant levels of pollutants even in the absence of weather patterns that lead to 

the stagnation type of air pollution episodes. The Northeast states are 

highly susceptible to pollutant plume transport effects, although some notable 

stagnation episodes have also affected this area (e.g., Lynn et al., 1964). 

Along the Pacific Coast, especially along the coast of California, coastal 

winds and a persistent low inversion layer contribute to major pollutant 

buildups in urban source areas and downwind along the urban plume trajectory 

(Robinson, 1952; Neiburger et al., 1961). In the southern Appalachians, the 

weather favors longer-term air pollution episodes (Korshover, 1967; Korshover, 

1975). Generally, low pollution potential results from the conditions that 

occur in the Great Plains area and south to the Texas-Louisiana Gulf Coast; 

and between the Mississippi River and the crest of the Rocky Mountains. 

It should be clear even from this brief discussion that there are fundamental differences in regional meteorological conditions that cause the air 

pollution potential applicable generally to California to be more severe than 

in other parts of the United States. When adverse meteorology is coupled with 

high population and source concentrations, it is quite evident why California 

areas have severe photochemical air pollution problems. This very significant 

difference in the magnitude of the photochemical air pollution problem in 

California regions compared to non-California locations can serve as a basis 

for separating air pollution statistics into two sets for evaluation, namely, 

California and non-California groups. 

3.4.3 Effects of Sunlight and Temperature 

The significance of sunlight in photochemistry is related to its intensity 

and its spectral distribution, both of which have direct effects upon the 

specific chemical reaction steps that initiate and sustain oxidant formation. 

Sunlight intensity varies with season and geographical latitude but the latter 

effect is strong only during the winter months. During the summer, the maximum 

light intensity is fairly constant throughout the contiguous United States and 

only the duration of the solar day varies to a small degree with latitude. 

3-66 
The effects of light intensity on individual photolytic reaction steps 

and on the overall process of oxidant formation have been studied in the 

laboratory (Peterson, 1976; Demerjian et al., 1980). All of the early studies, 

however, employed constant light intensities, in contrast to the diurnally 

varying intensities that occur in the ambient atmosphere. More recently, the 

diurnal variation of light intensity has been recognized and studied .as a 

factor in photochemicai oxidant formation (Jeffries et al., 1975; Jeffries 

et al., 1976). Such studies showed that the effect of this factor varies with 

initial reactant concentrations. Most important was the observation that 

similar NMOC/NOxsystems showed different oxidant potential depending on 

whether studies of these were conducted using constant or diurnal light. This 

has led to incorporation of the effects of diurnai or variable light into 

photochemical models (Tilden and Seinfeld, 1982). 

While the effect of sunlight intensity is direct and has been amply 

demonstrated (Leighton, 1961; Winer et al., 1979), the effect of wavelength 

distribution on the overall oxidant formation process is subtle. Experimental 

studies have shown the photolysis of aldehydes to be strongly dependent on 

radiation wavelength in the near UV region (Leighton, 1961); and there is some 

indication (Bass et al., 1980) that the photolysis rates for aldehydes may be 

temperature-dependent. Since aldehydes are major products in the atmospheric 

photooxidation of NMOC/NOxmixtures, it is inferred that the radiation wavelength should have an effect on the overall photooxidation process. This 

inference was directly verified, at least for the propylene/NO xand n-butane/ 

NOxchemical systems, in smog chamber studies (Jaffee et al., 1974; Winer 

et al., 1979). In the ambient atmosphere, some variation in the wavelength 

distribution of sunlight does occur as a result of variations in time of day, 

stratospheric 0 3 , ambient aerosol (Stair, 1961), and cloud cover. 

It has been observed that days on which significant ozone-oxidant concentrations occur are usually days with warm, above-normal temperatures (Bach, 

1975). The possibility that photochemical reactions show some temperature 

dependence has been raised by smog chamber studies (e.g., Carter et al., 1979; 

Countess et al., 1981), but is usually thought to be the production of chamber 

artifacts concomitant with increases in temperature. In ambient air, the 

observed correlation between temperature and oxidant concentrations can be 

explained, at least in part, as a synoptic meteorological correlation rather 

than as a temperature-photochemical rate constant effect, in that periods of 

3-67 
clear skies and warm temperatures are periods of high air pollution potential, 

as discussed above. Because of the close correlation between above-normal 

temperatures and high photochemical air pollution potential, a maximum daily 

temperature forecasting procedure is often useful as a substitute for a more 

elaborate and specific program for forecasting possible photochemical air 

pollution. The correlation between temperature and, thus, synoptic weather 

conditions and photochemical air pollution intensity has been observed in a 

number of areas. Evaluation of photochemical air pollution in Los Angeles as 

early as 1948 showed a correlation with temperature. Recent studies of 03 

patterns in St. Louis, Missouri, have also shown a correspondence between 

°daily maximum 0 3 concentration and temperature (Shreffler and Evans, 1982). 

Wolff and Lioy (1978) have shown high correlations between ozone concentration 

and both the temperature for the current day and the temperature for the 

previous day. 

3.4.4 Transport of Ozone and Other Oxidants and Their Precursors 

The 1978 air quality criteria document for ozone and photochemical oxidants 

made a convincing case for the fact that ozone and other photochemical oxidants 

are transported from urban source areas, other than those in California, to 

downwind regions in concentrations of 0.1 ppm or greater (U.S. Environmental 

Protection Agency, 1978a, and references therein). The 1972 study by Research 

Triangle Institute at McHenry, Maryland, was an early examination of rural 

oxidant in the eastern United States (Ripperton et al., 1977). Bach (1975) 

examined the meteorology of these observed conditions and showed details of 

the influence of transient high pressure systems. The transport of large 

urban plumes in the northeastern states, especially from New York City into 

Connecticut, was the subject of an EPA field study in 1975 (Westberg et al., 

1976; Wolff et al., 1977d; Westberg et al., 1978a). Transport of ozone on a 

regional basis in the northeast was also described by Cleveland et al. (1976a, 

1976b), while Wolff et al. (1977a) presented details of several east coast 

urban oxidant plumes. The Northeast Regional Oxidant Study (NEROS) carried 

out by EPA in 1979 and 1980 in the corridor from Washington, D.C., to Boston, 

was designed specifically to support urban plume model development (Clarke et 

al., 1982). Plume models have been based on other urban plume investigations, 

as well (e.g., Wolff et al., 1977c), and reactive plume modeling for other 

urban areas has progressed in recent years (e.g., U.S. Environmental Protection 

9101 
Agency, 1981a). These several early research programs and analyses served to 

answer a number of perplexing problems that had been identified previously in 

studies of ozone'and other photochemical oxidants in nonurban areas (e.g., 

Ripperton et al., 1977; Bach, 1975). These questions included reasons for the 

occurrence of high ozone or total oxidant concentrations in areas remote from 

identifiable sources. 

Studies of the transport of ozone and other photochemical oxidants (0 3-0x) 

are classified into three regimes, depending upon transport distance (U.S. 

Environmental Protection Agency, 1978a). In the first, urban-scale transport, 

the occurrence of transport of photochemical pollutants can be detected in 

most large urban areas if there is sufficient 0 3 -0xmonitoring information. 

It has been'identified as a significant, characteristic feature of the 03 -0x 

problem in the Los Angeles basin (Tiao et al., 1975), as well as in San Fran- 

< 

cisco, New York, Houston, Phoenix, and St. Louis (Altshuller, 1975; Coffey and 

Stasiuk, 1975; Shreffler and Evans, 1982; Wolff et al., 1977a). As noted 

above, the recognition and assessment of the probable magnitude of urban 

oxidant problems in locations other than in California has been a major research 

topic since the mid-1970s (e.g., Bach, 1975; Cleveland et al., 1976a, 1976b; 

Westberg et al., 1976; Wolff et al., 1977a,b,c; and others). 

Urban-scale transport patterns result f'rom one or more of a combination 

of factors. First is the simple advection of the photochemically reacting air 

mass and •the devel opment of maximum 0 3-0xafter 1 or 2 hours of downwi nd 

travel. Maximum concentrations may be displaced up to 20 or so miles from the 

center of the major source area. It has been noted (U.S. Environmental Protection Agency, 1978a) that pollutant concentrations in air parcels in the central 

core area of major source areas may not be the most conducive for 0 3=0xformation because of the tendency toward occurrence there of more effective scavenging, 

especially scavenging related to NO and its reactions. 

The distance of.the peak 0 3 -0xconcentrations from the urban core area is 

dependent on the local wind pattern and is, in general, inversely related to. 

the peak 0 3-0xconcentration. Stronger winds will carry the air parcels 

farther during the reaction period, increasingly diluting pollutant concentrations along the trajectory. Weak winds and very restricted mixing heights 

will tend to cause higher 03 -0xconcentrations closer to the central source 

area. The diurnal wind cycle wiii also be an important factor, since in some 

situations calm condi.tions may prevail until late in the morning but in others 

3-69 
a steady wind may be present throughout the emission and reaction process. 

Wolff and Lioy (1978) were able to model urban ozone concentrations on the 

basis of ineteorological observations, especially temperature. 

The second, or mesoscale, kind of transport of 0 3-0x is in many respects 

an extension of the urban-scale transport and is characterized by urban plume 

development. A report by Bell (1960) described November 1959 0 3-0x incidents 

in northern coastal San Diego County, California. It showed conclusively that 

these were caused by the 03-0x and precursors formed and emitted, respectively, 

the previous day in the Los Angeles basin. The transport in these situations 

was over the coastal Pacific Ocean, and the 0 3-0x arrived at the San Diego 

receptor site as a contaminated sea breeze after overnight travel (Bell, 

 1960). The studies of Cleveland et al. (1976a, 1976b) are early documentation 

of a similar scale of transport in the New York-Connecticut area.. The results 

of extensive aircraft measurements and modeling assessments of studies in the 

Washington, D.C., to Boston corridor have also been reported by Wolff and his 

colleagues (Wolff et al., 1977a,c; Wolff and Lioy, 1978). 

In the 1978 03-0x criteria document, more than 30 references were cited 

relating to urban plume observations and investigations. Since that document 

was published, the results of the 1975 New England oxidant study have been 

published in detail (e.g., U.S. Environmental Protection Agency, 1977; Westberg 

et al., 1978a), and results of a more comprehensive 2-year field program 

carried out along the Washington, D.C.-Boston corridor in 1979 and 1980 have 

appeared in the literature (Clark and Clarke, 1982; Clarke et al., 1982; 

Vaughan et atl., 1982). A major field program supported by local industries 

has been conducted in Houston, Texas, although the 03-0x downwind plume phases 

of that study were not as extensive as in NEROS. Chicago, Detroit, and adjacent 

shoreline areas of Lake Michigan have also been the subject of a number of 

ground-level and airborne studies over distances of 70 to 300 kilometers 

(Lyons and Olsson, 1972; Sexton and Westberg, 1980; Westberg et al., 1981; 

Kelly et al., 1986). As described above, 0 3-0x plumes from major urban areas 

can extend about 100 to 200 miles, with widths of tens of miles (Sexton, 

1982), frequently up to half the length of the plume. Other field studies 

conclusively demonstrating mesoscale transport over New England have been 

reported (Spicer et al., 1979; Clarke et al., 1982; Cleveland et a1., 1976a,b; 

Rubino et al., 1976; Westberg et al., 1976; Westberg et al., 1978a; Wolff et 

al., 1977a). Although urban plumes are frequently thought of as a problem 

3-70 
related only to large source areas such as New York and other major metropolitan 

areas, measurements in plumes from smaller urban areas have shown that these 

sources cannot be ignored (Sticksel et al., 1979; Sexton, 1982; Spicer et al., 

1982; Wolff et al., 1977c). 

In the third kind of pollutant transport, synoptic-scale, the transport 

of 03 -0xand precursors is characterized by the general and widespread elevated 

concentrations of pollutants that can occur on an air-mass scale under certain 

favorable weather patterns. These weather situations are generally slow-moving, 

well-developed high-pressure, or anticyclonic systems. This type of deep 

high-pressure area was considered by Korshover (1967, 1975) as a prerequisite 

for stagnating air pollution episodes. 

A major criterion of these synoptic systems is the reinforcement of the 

surface high-pressure area by a warm high-pressure circulation in the upper 

air. The surface weather is then frequently characterized by weak windS, 

stable surface layers, and high pollution potential over regional or air-masssized areas. This is the generalized meteorological model pattern that is 

involved in synoptic-scale pollutant transport (Korshover, 1967; Korshover, 

1975). As with all generalized models, there have been specific oxidant 

episodes that departed from this model. In particular, weak winds are not 

always a prerequisite; and relatively strong winds have been observed to, be 

associated with some oxidant transport episodes (Mueller and Hidy, 1983; Wolff 

and Lioy, 1980), when such winds did not also produce rapid plume dilutiqn as 

is usually expected. 

While surface highs that are reinforced aloft by a warm high can lead to 

air pollution episodes, there are other high-pressure systems that usually 

cause few or no widespread air pollution problems. These are the strong 

surface highs found behind rapidly moving storm systems, in which the surface 

high underlies a cold low-pressure system aloft. These systems characteristically have good vertical mixing (instability), brisk winds, and low air 

pollution potential. 

Bach (1975) assessed in detail the relationship between elevated ozone 

levels and synoptic systems. Other investigators have since described specific 

instances of large-scale ozone transport and associated meteorological conditions (e.g., Wolff and Lioy, 1980; Wolff et al., 1980; Wolff et al., 1982). 

For example, Wolff et al. (1982) have described synoptic meteorological systems 

and the occurrence of haze and eievated ozone levels in an area extending from 

3-71 
the Midwest to the Gulf Coast. In a separate study, Wolff and Lioy (1980) 

examined the spatial and temporal distributions of ozone during three photochemical smog episodes in July 1977. A stagnating high-pressure system formed 

over the Gulf of Mexico and two high-pressure systems originating in Canada 

were described as the respective meteorological systems associated with the 

three episodes. In all three cases, elevated ozone concentrations (-120 to 

130 ppb) were found to extend, in a virtual "ozone river", from the Gulf Coast 

to New England (with 328 ppb ozone measured in Connecticut). The high-pressure 

system originating in the Gulf area affected the entire southeastern United 

States and extended from western Texas, northeastward through Illinois, and 

east to the Atlantic Ocean. Elevated ozone levels affected anywhere from a 

few hundred square miles to a thousand square miles during the 1-week period. 

The importance of synoptic-scale or air-mass pollutant situations has 

been recognized for many years, probably much longer than the importance of 

major plumes has been apparent. The Donora, Pennsylvania, smog episode in 

1948 (Schrenk et al., 1949), while not a photochemical smog situation, involved 

the occurrence over a wide area of a regional air mass having relatively high 

pollution levels simultaneous with the occurrence of a stagnating warm highpressure area over the Ohio Valley and the northern Appalachian area. Donora 

was an especially adversely affected pocket within this larger system; in that 

case, however, ozone was probably not one of the important pollutants. 

The synoptic-scale high-pressure air pollution system is not characterized by well-defined urban plumes. Rather, a warm, slow-moving or stagnant 

anti-cyclone provides a synoptic-scale weather system that, because of weak 

winds and limited vertical mixing, favors the accumulation of relatively high 

concentrations of air pollutants. On a climatological basis, these systems 

are most common in the summer and fall months over most of the United States, 

as shown by the work of Korshover (1967; 1975). In many cases, an anti-cyclone 

will stagnate or recurve and intensify over the Midwest or East as circulation 

patterns in the upper air change and become more supportive of the surface 

t anti-cyclonic pattern (Schrenk et al., 1949; Lynn et al., 1964). The typical 

paths followed by the air masses forming these slower-moving anticyclones in 

the summer and fall months, as described by Bach (1975) and.Wight et al. 

(1978), is southeastward from Canada into the upper Mississippi Valley, eastward 

across the Ohio Valley and then across the East Coast, either toward'the 

northeast into New England, east into the central Atlantic States, or southeast 

3-72 
into the southeastern states. The actual stagnation of a pressure system for 

one or more days, as occurred in the Qonora case, ,is, of course, the most 

severe example of a slow-moving system. 

Along the West Coast, air pollution problems are also the result of 

persistent high-pressure system influences. In this case, however, the high 

is the persistent subtropical anti-cyclone of the eastern Pacific rather than 

the series of transitory anti-cyclonic systems characteristic of the area east 

of the Rocky Mountains. The persistent or semipermanent subtropical anticyclone in the Pacific is linked to the large-scale general circulation of the 

atmosphere rather than to moving wave systems (Neiburger et al., 1961). The 

effect is much the same, except that the area of limited mixing and more 

adverse air pollutant effects is found on the eastern edge of the subtropical 

anti-cyclone rather than on the trailing western edge as in the transitory 

systems. 

It is worthwhile to point out that the typical pattern of ozone concentrations in the slow-moving midwest anticyclone, in contrast to the conditions in 

California, shows the lowest values in the eastern portion of the system, 

behind the advancing cold front. The central area of the anticyclonic system, 

where dispersion is usually at a minimum, shows a broad expanse of eleuated 

ozone concentrations; but the highest ozone values in the typical system are 

usually found in the western parts of the system, the so-called "back.side," 

where dispersion conditions, although better than in the central portion of 

the system, combine with increased residence time and longer exposur.e to 

emission sources to cause the maximum ozone concentrations in a given highpressure system. Investigators describing this ozone pattern over the midwestern 

and eastern parts of the United States include Bach (1975), Vukovich et al. 

(1977), Wolff et al. (1977b), Wight et al. (1978), and Westberg et al. (1978a, 

1981). 

The identification and understanding of photochemical 0 3 -0xand precursor 

transport by weather systems has provided a significant advance in comprehending 

photochemical air pollution and the potential extent of its effects. Considerable 

progress has been made in the development of long-range photochemical modeling 

techniques so that the likely impact of synoptic systems can be anticipated. 

Such tools are very much in the research stage because the local impact of 

0 3 -0xresults from a complex interaction of distant and local precursor,sources, 

3-73 
urban plumes, mixing processes, atmospheric chemical reactions, and general 

meteorology. 

3.4.5 Surface Scavenging in Relation to Transport 

Major scavenging processes for 0 3 in the atmospheric boundary layer are 

adsorption and subsequent destruction at the ground surface (i.e., dry deposition), and reactions with boundary layer pollutants, especially NO and 

alkenes. In dry deposition, eddy diffusion moves air parcels downward 'through 

the turbulent boundary layer to the laminar sub-layer where individual molecules, 

such as ozone, can move by Brownian motion to the underlying surface. There 

reactive molecules, such as ozone, can be removed by reactions at the surface. 

Chemical reactions between ozone and NO and reactive hydrocarbons are described 

in other sections of this document. When they occur in the near-surface 

layers, these reactions can play an important role in the boundary layer 

scavenging process. Dry deposition and boundary layer chemical reactions 

result in a vertical concentration gradient, with the lowest concentrations 

occurring at the surface of the ground. 

Because of this surface-scavenging process, ozone will persist in an 

atmospheric parcel in the absence of ozone-forming reactions only if the 

parcel is dispersed such that contacts with the ground surface or surface 

pollutant sources are minimized. It is likely that only in those air parcels 

moving above the surface layer will ozone escape the surface reactions and 

persist long enough to undergo long-distance transport or persist overnight. 

It should be noted here that ozone transported aloft, judging from the limited 

data available, may be 20 to 70 percent additive in urban areas, as determined 

by simulations using the EKMA chemica1 kinetics model and several different 

rates of vertical mixing (U.S. Environmental Protection Agency, 1977b). Thus, 

if 0.1 ppm ozone were to be transported aloft and was 40 percent additive at 

ground level, the contribution of transport to the peak ozone concentration 

downwind would be 0.04 ppm. 

Aircraft observations have documented frequently the occurrence of 

relatively high ozone concentrations above lower-concentration surface layers 

(e.g., Westberg et al., 1976). This is a clear indication that ozone is 

essentially preserved in layers above the surface and can be transported over 

relatively long distances even when continual replenishment through precursor 

3-74 
reactions is not a factor, such as at night. 8oundary-layer scavenging processes are also responsible for the fact that ozone concentrations in urban 

areas rapidly fall to zero during the night. 

3.4.6 Stratospheric-Tropospheric Ozone Exchanqe 

The fact that 0 3 is formed in the stratosphere, mixed downward, and 

incorporated into the troposphere, where it forms a more or less uniformly 

mixed background concentration, has been known in various degrees of detail 

for many years (Junge, 1963). 

It is widely accepted that the long-term average tropospheric background 

concentration at the surface ranges from about 30 ppb to 50 ppb (Fabian and 

Pruchniewicz, 1977; Oltmans, 1981); and that it results primarily, though not 

exclusively, from the transfer of stratospheric ozone into the upper troposphere, 

followed by subsequent dispersion throughout the troposphere (e.g., Singh et 

al., 1980; Kelly et al., 1982). Ozone residence time in the troposphere has 

been estimated at 1 to 2 months by Junge (1963); and at 1 to 2 months for 

spring and summer but at 2.5 to 3.5 months for fall and winter, respectively, 

by Singh et al. (1980). 

The mechanisms by which stratospheri.c air is mixed into the troposphere 

have been examined by a number of authors. Danielsen conducted extensive 

analyses of major synoptic weather events that injected stratospheric air into 

the troposphere (Danielsen, 1968; Danielsen and Mohnen, 1977; Danielsen, 

1980). Reiter has been especially active in describing the atmospheric 

mechanisms by which stratospheric air injection takes place and in relating 

these processes to the global circulation of the atmosphere (Reiter, 1963; 

Reiter and Mahlman, 1965; Reiter, 1975). As a result of such research, exchange 

between the stratosphere and troposphere in the middle latitudes has been 

determined to occur to a major extent in events called "tropopause folds." In 

a tropopause fold (TF), the jet stream in the upper troposphere plays a major 

role in directing stratospheric air and high ozone concentrations into the 

troposphere. Figure 3-11 is a schematic presentation of the intrusion process 

as described by Danielsen (1968). The subsidence occurs along the poleward 

side of the polar jet stream in the area where the jet is associated with a 

cold front at ground level. The result is downward transport in the cold air 

behind the cold front. 

3-75 
OZOCIE 

Q ~ J O 

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O OS HERE 

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Ne 

(JET 

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~~ ~•"'~' 

TROPOSPHERE x aIRECT CELL 

INDIRECT CELL 

Figure 3-11. Schematic cross section, looking downwind along the jet stream, 

of a tropopause folding event as modeled by ®anielsen (1968). 

Source: Johnson and Viezee (1981). 

3-76 
Since 1978, a considerable amount of research on TF and ozone injection 

has been done, especially by SRI-International (Johnson and Viezee, 1981; 

Ludwig et al., 1977; Singh et al., 1980; and Viezee et al., 1979). Figure 3-12 

from Johnson and Viezee (1981) shows one example of the probing by SRI of a TF 

event ih the midwestern United States. Concentrations of ozone in excess of 

90 ppb were found as low as 13,000 feet (ca. 2.5 miles or.3.9 kilometers), as 

shown in the upper part of Figure 3-12. These authors found that ozone intrusion 

was lower during this fall study (October 5, 1978) than in a number of other 

spring TF events. The dew point measurements in the second part of the figure 

confirm the stratospheric injection. The weather pattern accompanying this TF 

is shown at the bottom of Figure 3-12 by a 500 millibar (about 6 kilometer) 

chart; the surface cold front is also indicated. Note that the intrusion was 

detected well behind the cold front and appears to have assumed a layered 

formation in the altitude range of 8,000 to 12,000 feet (2.4 to 3.6 kilometers). 

From their analysis of ineasurement fiights in a number of TF situations, 

Johnson and Viezee (1981) concluded that the ozone-rich intrusions studied 

sloped downward toward the south. In terms of dimensions, the average crosswind 

width (north to south) at an altitude of 5.5 kilometers (ca. 18,000 feet or 

3.4 miies) for six spring intrusions averaged 226 kilometers (364 miles), and 

for four fall TF systems, 129 kilometers (208 miles). Ozone concentrations at 

5.5 kilometers (ca. 18,000 feet or 3.4 miles) averaged 108 ppb in the spring 

systems and 83 ppb in the fall systems. Previously it had been assumed that 

only a few fairly intense systems would produce a TF event and trans-tropopause 

mixing. From their data, however, Johnson and Viezee (1981) drew the very 

important conclusion that all low-pressure trough systems, such as that 

illustrated in Figure 3-12, may induce a TF event and cause the trans=tropopause 

movement of ozone-rich air into the troposphere. 

On the basis of their field studies and the earlier modeis and work of 

Danielsen (1968), Johnson and Viezee (1981) proposed a set of model mechanisms 

or types of TF injection, which are illustrated in Figure 3-13 and described 

in the following general manner: 

1. Type 1. The intrusion is broken up and dispersed by mixing and 

diffusion in the middle or free troposphere. 

2. Type 2. The intrusion persists down to the planetary boundary 

layer or the top of the mixed layer, where the lower 

part of the intrusion may be incorporated into the 

mixed layer and may subsequently reach the ground. 

3-77 
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DISTANCE FROM CNG, nautical mi and (km) 

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CONTOURS (DYNAMIC METERS) 

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(C) 500-mb CHART AND FLIdHT TRACK 

Figure 3-12. Measured vertical cross-sections of (A) ozone, (B) 

dew-point, and (C) the 500-mb chart and the flight track for 

October 5, 1978. CMI = Champaign, IL; Cf11G = Cunningham, KS; 

MEM = Memphis, TN. 

Source: Johnson and Viezee (1981). 

0 0 0 0 0 0 0 

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(8) CROSS-SECTION OF DEWPOINT TEMPERATURE, °C 

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3-78 
TERAL 

IXING 

.. 

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TYPE 1 TYPE 2 

TYPE 3 TYPE 4 

~ 

Figure 3-13. Hypothesized models of the process that mixes tropopause 

folding events into the troposphere. 

Source: Johnson and Viezee (1981). 

3-79 
3. Type 3 . The i'ntrusion occurs close behind the cold front, where 

the air parcels are caught by the downdrafts behind the 

cold front; and is brought to the ground by direct 

circulations associated with the front. 

4. Type 4. The ozone-rich parcels are incorporated into convective 

cells and brought to the ground in association with 

rain-showers and thunderstorm downdrafts; similar to 

Type 3. 

Johnson and Viezee (1981) summarized the possible impacts of these four 

types of TF events by noting that Types 1 and 2 should produce "relatively 

moderate effects" at the ground in comparison to those to be expected from 

Types 3 and 4. The latter two could cause "substantial" effects in terms of 

high surface ozone concentrations. The action described by Types 3 and 4 is 

supported by meteorological theory (Bjerknes, 1951) and by observations of 

surface ozone such as those made by'Danielsen and Mohnen (1977), Lamb (1977), 

and Davis and Jensen (1976). 

3.4.7Stratospheric Ozone at Ground Level 

After a detailed review of background tropospheric ozone, Viezee and 

Singh (1982) concluded, as also concluded by other investigators (e.g., Kelly 

et ai., 1982), that the stratosphere is a major , but not the sole source of 

background ozone in the unpolluted troposphere. This stratospheric ozone is 

brought to the surface mixed layer by vertical mixing p.rocesses that have been 

known for many years (Junge, 1963). In the northern hemisphere, at midlatitudes 

between 30°N and 504, annual average background surface ozone concentrations 

generally range between 30 and 50 ppb, but in the tropics, lower concentrations, 

15 to 20 ppb, prevail (e.g., Fabian and Pruchniewicz, 1977; Oltmans, 1981). 

The stratospheric ozone reservoir has a strong seasonal variation, with a 

maximum in the spring and a minimum in fall and winter months, especially at 

middle latitudes. This seasonal cycle is reflected at ground-level background 

observation stations, where the average spring background ozone at some stations 

may be as high as 80 ppb and tfie average fall values range between 20 and 40 ppb 

(e.g., Singh et al., 1977; Mohnen, 1977; U.S. Environmental Protection Agency, 

1978). In the troposphere, concentrations generally increase gradually to the 

tropopause, but the seasonal pattern is the same (Viezee and Singh, 1982). 

Using data acquired in their studies of TF events, researchers at SRIInternational examined the frequency with which stratospheric intrusions occur. 

EMO 
According to Singh et al. (1980), an intrusion can be expected to be present 

over the United States on about 90 percent of the days. This frequency 

diminishes somewhat in summer, however (Singh et al., 1980), which is the 

season when most ground-level smog episodes occur. 

Viezee and Singh (1982) concluded that relatively high ozone concentrations can occur for short periods of time, minutes to a few hours, over local 

areas as a result of stratospheric intrusions. They were able to document 

from published literature ten situations of probable intrusion of stratospheric 

ozone.. These instances are shown in Table 3-7, reproduced from Viezee et al. 

(1983). The concentrations reported in Table 3-7 were measured.atground-level 

stations. Note that all of the short-term situations in which peak concentrations exceeded 80 ppb occurred in fall, winter, or.spring months and not in 

the photochemically active summer season. Of the three summer instances that 

were reported, two at Whiteface Mountain, New York, and one at Pierre, South 

Dakota, the highest reported concentration was 56 ppb for a 1-hr average. The 

lower incidence in summer of reported ground-level impact by stratospheric 

ozone may be attributable in part to the reduced frequency of intrusions in 

summer, as reported by Singh et al. (1980). In addition, however, the potential 

for ground-level impact by stratospheric ozone in summer is lessened because 

of the stability provided by the upper-level, warm anticyclone present in the 

weather systems characteristic of summertime photochemical smog episodes. 

There have been a number of attempts to quantify the proportion of the 

surface ozone attributable to stratospheric sources. The most common method 

used is based on the assumption that 7Be is a unique tracer for air parcels of 

stratospheric origin. Both ozone and 7Be are measured and the proportion of 

the surface ozone that might be of stratospheric origin is calculated by using 

a derived ratio between ozone and 7 Be. Some of the investigators using this 

technique include Ferman and Monson (1978), Wolff et al. (1979), Husain et al. 

(1977), Dutkiewicz and Husain (1979), Singh et al. (1980), and Johnson and 

Viezee (1981), among others. 

Caiculated correlations between surface ozone and 7Be show that their 

relationship is variable. The results of Kelly et al. (1982), from studies of 

7

Be, 0 3 , and air mass classification in South Dakota, showed that continental 

tropical (cT) air masses frequently seemed to have higher 7 Be and 0 3 values on 

the western side of the traveling cT anticyclone. A similar relationship was 

not prominent in the polar air masses studied, however, and maritime tropical 

3-81 
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air masses did not reach the South Dakota site. Kelly et al. (1982) found a 

linear correlation, r, equal to 0.65, for 1978 summer measurements of 7Be and 

ozone in South Dakota. Ferman and Monson (1978) reported r= 0.60 at McKee, 

Kentucky, for 27 daily samples taken during August and September 1976. Johnson 

and Viezee (1981) reported an r value of about 0.50 for 7 Be and ozone data 

from Dodge City, Kansas, during April and May 1978. A much lower correlation, 

r= 0.15, was reported by Husain et al. (1977) for July and August 1975 measurements at Whiteface Mountain, New York. These data imply that the variabiiity 

of the surface ozone concentration that can be explained by 7 Be varies from 

about 40 percent to less than 5 percent. 

Singh et al. (1980) and Viezee and Singh (1982) have pointed out a number 

of problems with this technique in their detailed analyses of the application 

of 7Be measurements to the quantification of the amount of stratospheric ozone 

in surface air. Some of the problems encountered when applying 7 Be/03 ratios 

over short sampling periods (in contrast, for example, to seasonal averages) 

include the following: 

1. Because zBe is an aerosol and O a is a gas, they have fundamentally 

different atmospheric scavenging mechanisms and thus respond differently to tropospheric meteorological conditions. 

2. Although the stratosphere is its dominant source, on the average, 

7 Be is synthesized in both the troposphere and the stratosphere. 

The 7 Be source cannot be assumed to be equivalent to the 0 3 source 

without meteorological verification. 

3. 7 Be sampling data are primarily averaged over 24 hours, which does 

not give sufficient time resolution for disaggregating short-period, 

rapidly moving stratospheric intrusion events from longer-term 

processes. 

4. The lower stratosphere is a highly variable and pooriy characterized 

region for which no 7 Be%03 ratio has been firmly established. 

As a result of these and other factors, Singh et al. (1980) concluded that 

"the experimental technique involving a 7Be/03 ratio to estimate the daily 

stratospheric component of ground-level 0 3 is unverified and considered to be 

inadequate for air quality applications" (Singh et al., 1980, p. 1009). The 

investigators at SRI-International have suggested, however, that 7 Be may be 

used, under the appropriate meteorological conditions, as a quaiitative tracer 

for air masses of stratospheric origin (Johnson and Viezee, 1981; Viezee et 

al., 1979)., 

` 3-83 
Other methods have been used to estimate the quantitative contribution of 

stratospheric ozone to ozone concentrations measured at ground-level. These 

include: (1) aircraft observations of TF events (Johnson and Viezee, 1981), 

coupled with calculations of the downward flux of ozone in the troposphere 

(Viezee et al., 1983); (2) use of inean 90Sr surface measurements, from radioactive fallout studies, to calculate a mean stratospheric ozone contribution 

(Reiter, 1978); and (3) examination of data, especially multiyear data, on 

surface ozone concentrations at remote- sites (Viezee and Singh, 1982; Viezee 

et al., 1983). Using surface measurement data and their own data from aircraft 

measurements and calculated.ozone fluxes, Viezee et al. (1983) concluded that 

direct ground-level impacts by stratospheric ozone may be, infrequent, occurring 

=<1 percent of the time; that such ground-level events are short-lived and 

episodic; and that they are most likely to be associated with ozone concentrations <0.1 ppm. Viezee et al. (1983) recommended further study, however, on 

the possible contribution at ground-level of stratospheric ozone. 

On a qualitative basis, as mentioned earlier, there is no doubt that 

stratospheric ozone is present in, the atmospheric surface layers, and the 

meteorological mechanisms that bring this about have been described by,a 

humber of investigators, including Wolff et al. (1979 . ), Johnson and Viezee 

:(1981), and others. Most investigators cite the basic,meteorological analyses 

of Danielsen (1968) as a basis for. their exchange model. 

The downward transfer, of air parcels.and ozone from the stratosphere into 

the ti^oposphere has been described above. There is, of course, a compensating 

transfer of tropospheric air upward .i.nto the lower stratosphere. Reiter 

(1975) has examined various mechanisms that contribute to this transfer. Air 

parcels moving out of the troposphere will carry with them the background 

concentrati ons of ozone that they had . i n the troposphere, and, as the ai r 

parcels mix in the stratosphere, these ozone molecules will become part of the 

stratospheric background ozone. Since the ozone concentrations are very much 

lower in the troposphere compared,to the stratosphere, however, this exchange 

of tropospheric and stratospheric air.parcels will not result in a net upward 

transport of ozone .and is not considered to be ; a factor.inair pollution 

situations. 

3.4.8 'Background Ozone from Photochemical Reactions 

The apportionment of the background ozone concentrations measured at 

remote and other nonurban locations to stratospheric versus tropospheric 

3-84 1 
processes and sources is the subject of continuing discussion and research. 

In addition, the apportionment of ozone produced photochemically in the 

troposphere to natural versus manmade sources of precursors, and, in turn, 

their respective contributions to background levels of ozone, aiso remain a 

focus of discussion and research. The preceding discussion has emphasized the 

thesis that the major source of tropospheric background ozone is the 

stratosphere. It must be recognized, however, that some investigators argue 

for a much larger role for the formation of significant amounts of ozone 

within the troposphere (e.g., Fishman et al:, 1980; Fishman and Seiier,.1983; 

Fishman and Carney, 1984; Fishman et al. , 1985; Fishman and Crutzen, ,1978; 

Chameides and Walker, 1976). The question of the potential relative impacts 

of the stratospheric source of ozone and tropospheric photochemical sources 

has been evaluated critically by Singh et al. (1978, 1980). 

Singh et al. (1980) concluded that background ozone in the troposphere is 

"principally of stratospheric origin," and supported this position with,a 

number of arguments, including the following. First, the NO X concentrations 

in the free troposphere appear to be very low, probably 0.01 to 0..1 ppb; at 

NOx concentrations <0.03 ppb a tropospheric ozone reservoir cannot be generated, using photochemical simulation models. Second, Singh et al. (1980) 

pointed out that the tropospheric concentration pattern of ozone, with its 

spring maximum and fall minimum, is out_of phase and inconsistent with a 

photochemical source, which would be at a maximum during the summertime peak 

of solar radiation flux. Also, at low latitudes, where the solar fiux is 

relatively constant, the ozone seasonal cycle is.quite pronounced and .i.s 

consistent with the seasonal cycle of stratospheric-tropospheric exchange 

processes. In .addition, Singh et al. (1980) noted that the vertical gradient 

of ozone in the troposphere is contrary to a dominant photochemical source for 

background tropospheric ozone. 

Other evidence, however, especially more recent modeli.ngstudies, indicates 

that a substantial part of the 0 3 measured during the warmer months of the 

year over the United States and Western Europe is of photochemical origin 

(Altshuller, 1986). Attempts to obtain agreement with observed 0 3 concentrations 

with a general circulation model including only a stratospheric source of 03 

have been unsuccessful, especially over continents (Levy et al., 1985). From 

their tropospheric modeling studies, Fishman and coworkers have predicted the 

photochemical formation of near-surface, summer ozone at concentrations of:. 

3-85 
(1) 40 ppb, using surface NOxemissions corresponding to a concentration of 

N0.25 ppb (Fishman and Seiler, 1983; Fishman and Carney, 1984); (2) 15 ppb, 

assuming zero NO xemissions (Fishman and Seiler, 1983); and (3) 80 ppb, using 

NOxemissions corresponding to a concentration of about 5 ppb and ignoring 

contributions from individual urban or industrial plumes (Fishman et al., 

1985). Using another model, Dignon and Hameed (1985) have predicted the 

photochemical production of ozone at midlatitudes of the Northern Hemisphere 

at annual average concentrations of: (1) <60 ppb, using a 1980 NO xemissions 

inventory; or (2) —25 ppb, assuming zero manmade NO xemissions. On the basis 

of his review of these and other studies, Altshuller (1986) has estimated that 

average stratospheric contributions to near-surface ozone concentrations range 

from 10 to 15 ppb. 

As these data and the preceding discussion in Section 3.4.7 indicate, the 

portion of background ozone concentrations attributable to stratospheric 

sources versus tropospheric photochemistry remains uncertain, especially in 

the absence of a quantitative technique for determining whether ozone is of 

stratospheric origin. 

Investigations on the contributions of photochemistry to background ozone 

have focused on (1) the role of transport from urban into nonurban areas, 

which was discussed in Section 3.4 (see also Chapter,5); and (2) the respective 

roles of biogenic VOC and of NO x , from all sources, in the photochemical 

formation of ozone in nonurban areas. 

Altshuller (1983) has evaluated the specific question of whether naturally 

emitted VOC (i.e., from biogenic VOC sources) could be a significant source of 

background ozone. Since biogenic emissions are released into the atmospheric 

boundary layer, this potential source of ozone is expected to affect only 

ground-level or boundary-layer concentration patterns, in contrast to the 

whole troposphere as described in the preceding discussion. While a matter of 

controversy in recent years, the data on the role of biogenic VOC in boundarylayer photochemistry now appear more conclusive. Inventories of biogenic VOC 

emissions indicate that they are of the same order of magnitude as manmade VOC 

emissions, as described in Section 3.5.1 (below). On the other hand, concentration data indicate that biogenic VOC occur, even in nonurban areas, at low 

levels in ambient air relative to VOC from manmade sources (Section 3.5.2 

below). In an extensive review of the literature, Altshuller (1983) noted 

that the concentrations of biogenic hydrocarbons are very low, constituting 

:. 
even at rural sites probably 10 percent or less of the total nonmethane hydrocarbons. He concluded that the contribution of biogenic hydrocarbons to ozone 

formation, as a result, does not appear to be significant under most atmospheric 

conditions. 

The role of NOxin the photochemical formation of ozone in nonurban areas 

has been the subject of a number of modeling and experimentai studies, as well 

as a recent review (Altshuiler, 1986). As summarized in Section 3.5 (below), 

and as reviewed by Altshuller (1986), NO xconcentrations in nonurban areas of 

the United States , are appreciably lower than those in urban areas, sometimes 

by an order of magnitude. Likewise, concentrations of NO xin the western 

United States are usually lower by about a factor of ten than those in the 

northeastern states (Section 3.5.2.2.2). At nonurban locations inside of 

populated areas, however, the concentrations of NO xare much higher than those 

measured in clean air within the continental and maritime boundary layer and 

in the free troposphere (Altshuller, 1986). 

Giobal background concentrations of NOxwere previously thought to be 

lower than more recent measurements show them to be. For example, the analysis 

cited earlier in this section (Singh et al., 1980), on the contributions of 

photochemistry versus stratospheric intrusions, assumed NO xconcentrations in 

the free troposphere of 0.01 to 0.1 ppb and an NO xreservoir of <0.03 ppb; and 

the global models described in Kelly et al. (1984) assume an NO xbackground of 

0.1 to 0.2 ppb. In contrast, the mean NOxconcentrations tabulated by Altshuller 

(1986) for remote surface sites, while <1 ppb and often <0.5 ppb, are stiil in 

excess of the ranges described above. At nonurban sites east of the Rocky 

Mountains in the United States, mean NO xconcentrations range from 1 to 10 ppb 

(Altshuller, 1986). 

Martinez and Singh (1979) analyzed the role of NO xin nonurban ozone 

formation using theoretical approaches and aerometric data from the Sulfate 

Regional Experiment. They concluded that the impact of NO xon nonurban ozone 

is a function of geographical location since ozone production was NO x-limited 

at some sites they examined but not at others. Kelly et al. (1984) examined 

the role of photochemistry in nonurban ozone at three very different rural 

sites (in South Dakota, Louisiana, and Virginia, respectively) by ambient air 

analyses, captive air irradiations, and photochemical modeling. Ambient air 

,analyses indicated the formation of about 6 ppb ozone per ppb of NO x . Irradiation and modeling results indicated similar, but slightly higher, ozone formation per ppb of initial NO x . They concluded that photochemistry in these 

3-87 
rural areas was not NO x-limited but depended, as well, on hydrocarbon concentrations. In a study that included modeling and a comparison of the modeled 

predictions with observed ozone concentrations, Liu et,al. (1984) arrived at a 

similar conclusion; i.e., that nonurban ozone formation in the areas they 

studied was not exclusively controlled by the equilibrium between NO xand 

ozone (Equation 3-4, Section 3.3.1) but by the hydrocarbon reactions, as well. 

Consistent with the results of the Kelly et al. (1984) study, Parrish et al. 

(1986) developed empirical relationships between ozone and NOX concentrations 

(June to September) at a high-altitude location (Niwot Ridge, Colorado) and 

found the photochemical formation of about 17 ppb ozone for each ppb of NO x . 

The studies and analyses cited in this and the preceding section indicate 

that research is still needed to permit the apportionment of background ozone 

to stratospheric and respective tropospheric sources. From his review of data 

pertinent to the role of NO xin boundary-layer photochemistry, Altshuller 

(1986) concluded that photochemically generated ozone should equal or exceed 

ozone transported down from the stratosphere to relatively remote locations at 

lower elevations. At more polluted rural locations, photochemically generated 

ozone from manmade emissions are predi:cted to constitute most of the ozone 

measured during the warmer months of the year. Thus, the evidence supporting 

the stratosphere as being the major (but not exclusive) source for background 

tropospheric ozone is relatively strong but still not conclusive. Research in 

these topic areas is continuing, and the number of uncertainties and necessary 

assumptions needed for the respective conclusions cited above should be reduced 

as new data become available. 

3.5 PRECURSORS TO OZONE AND OTHER PHOTOCHEMICAL OXIDANTS 

3.5.1 Sources and Emissions 

3.5.1.1 Manmade Sources and Emissions. This section presents information on 

the manmade sources arid emissions of precursors to ozone and other photochemical 

oxidants. Estimates of annual emissions are presented by source category for 

volatile organic compounds and nitrogen oxides. In addition, information is 

presented on emission rates and the composition of emissions from principal 

stationary sources and from mobile sources. The annual estimates are based on 

emission inventories prepared according to procedures developed and published 

by EPA (U.S. Environmental Protection Agency, 1980a,b; 1981a,b,c,d,e,f). 

. .~ 
3.5.1.1.1 Trends in emissions of volatile organic compounds. Emission data 

on volatile organic compounds (VOC) include data on hydrocarbons as,well•as 

other organic compounds found in ambient air., Because of their negligible 

reactivity toward the photochemical production of ozone, methane, ethane, 

methylene chloride, and several hal.ogenated methanes and ethane,s are also 

excluded from emission inventories (U.S. Environmental Protection Agency, 

1980b). Estimates of total emissions of volatile ordanic compounds,(VOCs) 

provide a gross measure of compounds.avai.lable for photochemical production of 

ozone and other 'photochemical oxidants. Emissions of VOCs are reported here 

as the collective mass of reactive VOC. (See the footnote to-Table 3-8 for 

the calculation of emissions-based NMOC/NO X ratios.) 

Retrospective estimates of total manmade.VOC emissions in , the United 

States, based on records of economi c acti ,vity (e. g. , fuel usage, i ndustri al 

production) have been prepared by.decade, beginning with 1940 (U.S..Environmental Protection Agency, 1986). From a. level of 18.5 Tg/yr in 1940, VOC 

emissions increased about 14 percent each decade through 1970 ;then began to 

decline (Table 3-8): 

Figure 3-14 shows national trends in yearly emission,s ,o.f manmade VOC by 

general source category for the period 1970 through 1983 (U.S. Environmental 

Protection Agency, 1984a). . Emissions in 1983 of manmade VOCs .in the.United 

States have been estimated. at 19.9 Tg/yr (U. S. Environmen,tal Protection 

Agency, 1984a); total manmade VOC emissions nationwide were 26 percent.lower 

i n 1983 than i n 1970. The observed decrease i s attri butabl e 1 arge.ly , to a 

decrease of 30 percent in estimated VOC emissions from highway vehicles during 

this period. This decrease occurred in spite of a 42 percent increase in 

vehicle miles traveled. Trends in manmade VOC emissions versus vehicle miles 

traveled, urban only and total, for.1970 through 1983, are shown , in Figure 

3-15 (U.S. Environmental Protection Agency, 1984a; Motor Vehicle Manufacturers 

Association, 1984). The main sources nationwide of manmade VOC are industrial 

processes, which emit a wide variety of VOCs such as chemical solvents; and 

transportation, which includes the emission of VOCs,in gasoline vapor as well 

as in gasoline combustion products. 

3.5.1.1.2 Trends in emissions of nitrogen oxides. Emissions of nitrogen 

oxi des (NOx ) are reported here as the sum. of .NO and NO 2 , al l expressed as 

equivalent NO 2 . Retrospective estimates of total manmade NOxemissions in the 

.. 
TABLE 3-8. EMISSIONS OF VOC BY DECADE, 1940 THROUGH 1980 

(Tg/yr) a 

Source category 1940 1950 1960 1970 1980 

Transportation 5.2 7.9 11.1 12.3 8.2 

Stationary fuel combustion 4.7 3.1 1.9 1.0 2.1 

Industrial processes 3.2 5.2 6.1 8.7 8.9 

Solid waste 0.9 1.0 1.4 1.8 0.6 

Miscellaneous 4.5 3.6 3.1 3.3 2.9 

Total 18.5 20.8 23.6 27.1 22.7 

Source: U.S. Environmental Protection Agency (1986). 

aTo calculate emissions-based NMOC/NO molar ratios, as used in controlrelated data, a multistep procedure is required (Novak, 1986) as described 

by the equation: 

4 

Total NMOC (Tg/yr) x SUM[Pctix no. i /mwi ]/[Total NOx (Tg/yr)/461 

i=1 

where Pct = percent of a compound class present in the NMOC emissions; 

no. = average carbon number for a compound class; 

mw = average molecular weight for a compound class; 

i represents the four compound classes of interest; and 

NOxemissions are expressed as NO 2 (mw = 46). 

Values for terms in the equation vary in the published literature. Example 

values, where no. and mw are derived from emissions inventories (Novak, 1986) 

and pct from ambient air concentration data (Richter, 1986), are: paraffins 

=[63.5 x 4.56/95]; olefins =[15.5 x 3.57/43]; aromatics =[15.5 x 7.56/971; 

and aldehydes = [2.0 x 1.5/37]. 

United States have been prepared by decade, beginning with 1940 (U.S. Environmental Protection Agency, 1986). From a level of 6.8 Tg/yr in 1940, NO x 

°emissions increased almost 40 percent each decade through 1970 (Table 3-9). 

The increase from 1970 to 1980 was 13 percent. 

Total NOxemissions in the United States in 1983`(19.4 Tg/yr) were some 

17 percent above the 1970 level, but appear to have'declined slightly from 

about 1980 through 1983 (Figure 3-16) (U.S. Environmental Protection Agency, 

1984a). .The net increase over the period 1970 through 1983 may be attributed 

primarily to two causes: (1) increased fuel combustion in stationary sources 

such as power plants; and (2) increased fuel combustion in highway motor 

vehicles, as the result of the increase (42 percent) in vehicle miles driven 

over the 14 years in question. Emissions associated with industrial processes 

3-90 
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YEAR 

Figure 3-14. National trend in estimated emissions of volatile. organic compounds, 1970 through 1983. 

Source: U.S. Environmental Protection Agency (1984a). 

33 

3-91 
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Figure 3-15. Comparative trends in highway vehicle emissions of nitrogen oxides (IVOX) and 

volatile organic compounds (VOC) versus vehicle miles traveled, 1970-1983. 

Source: Motor Vehicle Manufacturers Association (1984); U.S. Environmental Protection 

Agency (1984a). 

3-92 
TABLE 3-9. EMISSIONS OF NO BY DECADE, 1940 THROUGH 1980 

(T~/yr) 

Source category 1940 1950 1960 1970 1980 

Transportation 2.2 3.5 4.9 7.6 9.2 

Stationary fuel combustion 3.4 4.7 6.7 9.1 10.2 

Industrial processes 0.2 0.3 0.5 0.7 0.7 

Sol id waste 0.1 0.2 0.3 0.4 0.1 

Miscellaneous 0.9 0.6 0.4 0.3 0.2 

Total 6.8 9.3 12.8 18.1 20.4 

Source: U.S. Environmental Protection Agency (1986). 

remained relatively constant, but solid waste and miscellaneous emissions 

decreased slightly. 

The national trends shown do not reflect the considerable local and 

regional differences that exist in the relative amounts of NO xemitted in the 

major source categories. For example, motor vehicle emissions in Los Angeles 

County, California, increased sixfold from 1940 to 1970 (Los Angeles County, 

1971), compared to a threefold national increase. Figure 3-15 compares the 

relative trends in mobile source NO xemissions versus vehicle miles traveled, 

urban only and total, from 1970 through 1983 (U.S. Environmental Protection 

Agency, 1984a; Motor Vehicle Manufacturers Association, 1984). 

3.5.1.1.3 Sources and emissions of volatile organic compounds. The source 

category contributing the largest percentage of VOC emissions in 1983, 37.7 

percent, is Industrial Processes (U.S. Environmental Protection Agency, 1984a). 

The category consists almost entirely of point sources. The composition of 

these emissions varies widely, depending on the process or product and the use 

of emission reduction equipment and operating practices. 

The second largest VOC source category, Transportation, accounting for 

36.2 percent of the annual total in 1983, is discussed beiow. 

The third largest VOC source category, Miscellaneous, accounts for 12.6 

percent of the annual total, over half of which consists of the subcategory, 

Miscellaneous Organic Solvents (U.S. Environmental Protection Agency, 1984a). 

These emissions generally qualify as area-source emissions. Some of these 

solvents are widely used in domestic products such as furniture polish, shoe 

polish, shaving soap, perfumes, cosmetics, shampoo, hair spray, hand lotion, 

3-93 
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to 

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1970 1971 1972 1973 1974 1975 1976 1977 1978 1979 1980 1981 1982 1983 

YEAR 

Figure 3-16. National trend in estimated emissions of nitrogen oxides, 1970 through 1983. 

Source: U.S. Environmental Protection Agency (1984a). 

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3-94 
rubbing alcohol, and nail poiish remover. The predominant compounds emitted 

are isopropyl alcohol and ethyl alcohol (Bucon et al., 1978). 

Emissions of volatile organic compounds from the production and marketing 

of gasolines and motor oils are classed in the Industrial Processes`category 

of VOC emissions. Emissions of VOC from these petroleum products after their 

sale to vehic7e owners are included in the Transportation category. 

A significant portion of VOC.emissions from gasoline-fueled mobile sources 

arises from evaporation, but the most conspicuous mobile source emissions are 

the combustion products. In a study by Black et al. (1980), evaporative 

emissions were found to constitute one-third to one-half of total hydrocarbon 

emissions from all of the vehicles tested '(n = 60). Based upon actual 

surveillance tests of in-use vehicles, Fisher (1980) found emission rates of 

2.53 g/mi in 1968 and 0.15 g/mi in 1980 for composite evaporative and crankcase 

emissions from gasoline-fueled cars. 

Exhaust emissions from gasoline-fueled vehicles typically 'contain fuel 

components and low-molecular-weight hydrocarbons that are not present in the 

fuel. Typicaily, exhaust from a catalyst-equipped automobile contains about 

62 percent alkanes, 17 percent aromatics, 18 percent alkenes, and 3 percent 

acetylene. This may be compared with the corresponding typical values for 

automobiles without converters: 40, 24, 26, and 11 percent, respectively. 

Methane levels generally range from about 10 to 30 percent (Black and Bradow, 

1975; Black, 1977). Exhaust gases from gasoline-fueled vehicles also contain 

organic compounds such as aldehydes, ketones, ethers, esters, acids, and 

phenols, amounting to as much as one-tenth of the total VOC content. 

Factors other than gasoline composition influence the composition of 

exhaust from internal combustion engines. These include driving patterns, the 

specific configuration of emission control devices, ambient temperature and 

humidity, and, of course, individual automobile parameters such as tuning, 

make, and model year. Fuel additives can also influence emissions. For 

example, in one study, tetraethyl lead increased hydrocarbon emissions by 

about 5 percent but did not change the type of emissions (Leihkanen and Beckman, 

1971). 

Evaporative emissions from diesel vehicles are negligible because of low 

fuel volatility (Linnell and Scott, 1962; McKee et al., 1962). In studies of 

exhaust emissions from diesel automobiles (Black and High, 1979; Gibbs et al., 

1983), 15 to 40 percent of the hydrocarbons emitted were found to be associated 

3-95 
with particles by the time the exhaust stream exited the tailpipe. Gibbs et 

al. (1.983) reported THC emissions from 19 in-use diesel automobiles, representing 1977 to 1979 model years, that were tested periodically over a 28-month 

period. Emissions of THC at the end of the period ranged from 0.17 to 0.88 

g/mi for individual vehicles and averaged 0.65 g/mi. It should be noted here 

that the population of diesel-powered passenger cars is not growing as rapidly 

as expected. Sales of diesel-powered cars peaked at 6 percent for 1981 models 

and dropped to less than 3 percent by 1983 (Automotive News, 1982a,b,; 1983a,b; 

Plegue, 1983). 

A brief summary of both exhaust and evaporative emissions characteristic 

of a variety of engines, fuel types, and control devices has been presented by 

1`ilton and Bruce (1981). Their summary indicates, for example, that exhaust 

emissions of aldehydes and ethylene increase in cars fueled with ethanolgasoline blends (gasohol) as opposed to gasoline. Exhaust emissions of total 

hydrocarbons (THC) decrease but evaporative emissions increase, for a net 

increase in emissions from gasohol-fueled cars. In diesels, evaporative THC 

emissions are virtually nonexistent, and net exhaust emissions of THC are 

lower. The percentage of carbonyl compounds (aldehydes, ketones) in the 

exhaust of diesels is higher than from gasoline combustion, but the net 

photochemical reactivity per gram of emissions is lower nevertheless (Tilton 

and Bruce, 1981). 

In vi ew of changes i n emi ssi ons wi th fuel type, i t i s of i nterest to 

examine fuel usage in the United States in recent years. Yearly sales of 

vehicle fuels fluctuate in consequence of several factors, including retail 

fuel price, general economic conditions, and the age and fuel efficiency of 

the composite vehicle population. In addition to the year-to-year influence 

of these factors on the sales of the principal vehicle fuels, gasoline and 

diesel fuel, a change in fuel composition is emerging with the introduction in 

the United States of gasolines containing a percentage of ethanol. As shown 

in Table 3-10, sales of gasoline containing up to 10 percent ethanol (gasohol) 

were first reported in 1981 and had increased almost sixfold 2 years later 

(NPN: National Petroleum News, 1981-1985 summary issues). 

3.5.1.1.4 Sources and emissions of nitrogen oxides. Fuel combustion is the 

dominant source of NOxemissions nationally. Stationary sources contributed 

50 percent and mobile sources contributed 45.4 percent in 1983 (U.S. Environmental Protection Agency, 1984a). In contrast to their contributions to VOC 

3-96 
TABLE 3-10. YEARLY QUANTITIES OF 

THE UNITED STATES FOR HIGHWAY 

MOTOR VEHICLE FUELS SOLD 

USE, 1980 THROUGH 1983 

IN 

Gallons x 109 

Fuel 1980 1981 1982: 1983 

Gasoline 106.45 104.14 100.70 101.55 

Gasohol — 0.72 2.16 4.29 

Diesel 13.60 14.30 14.67 15.87 

Source: NPN: National Petroleum News (1981-1985 summary issues). 

emissions, industrial process sources contributed only about 3 percent to the 

national total of manmade NOxeinissions in 1983 (U.S: Environmental Protection 

Agency, 1984a). 

Ratios of NO/NOxin emissions vary depending upon the source. Nitric 

oxi de (NO) i s the domi nant oxi de of ni trogen emi tted by most sources ; NO 2 

generally comprises less than 10 percent of the total NO xemissions. More than 

30 to 50 percent, however, of the total NOxemissions from certain diesel 

(Braddock and Bradow, 1976; Springer and Stahman, 1977 a,b) and jet turbine 

engines (Souza and Daley, 1978) can be NO 2under specific load conditions. 

Likewise, tail gas from nitric acid plants, if uncontrolled, may contain about 

50 percent NO 2(Gerstle and Peterson, 1966). Variations in NO/NO xratios by 

source type can be significant in local situations, as, for example, in the 

immediate vicinity of a high=volume roadway carrying a significant number of 

diesel-powered vehicles. 

Emissions of NO xfrom the principal categories of stationary combustion 

sources, electric utility boilers, industrial boilers, and industrial process 

heaters, vary according to fuei type and equipment design. Among electric 

utility boilers, coal-fired units produce the highest NOx emissions (994 ppm), 

as reported by Bartok et al. (1971). Combustion modifications can reduce NOx 

emissions from such coal-fired units by 50 percent (Crawford et al.; 1974, 

1978). Coal-fired industrial boiiers also emit greater amounts of NOx (475 

ppm) than oil-fired units emit (Cato et al., 1974, 1976). Of the industrial 

process heaters studied by Hunter et al. (1978,.1979), cement kilms emitted 

the highest levels of NOx (90 to 2250 ppm). 

Seasonal variations in'NOxemissions from stationary sources are expected 

since power plants using fossil fuels produce an estimated 15 percent more NO x 

3-97 
in the summer than in the spring (U.S. Department of Energy, 1978). Greater 

degrees of variation and different seasonal patterns have been reported for 

stationary sources in different regions of the country (California Board of 

Sanitation, 1966). 

Emissions of NOX from mobile sources, gasoline- and diesel-fueled vehicles, 

are affected by a number of variables such as speed, load, and air:fuel ratio 

(AFR), as reported by Hilliard and Wheeler (1979). Seasonal variations in NO X 

emissions from mobile sources will occur in relation to temperature (about a 

35 percent decrease in emissions per vehicle mile with an ambient temperature 

increase from 20 to 90 ° F) (Ashby et al., 1974), and number of vehicle miles 

traveled (about 18 percent higher in summer than in winter, nationwide) (Federal 

Highway Administration, 1978). Emissions of NO X also vary with vehicle miles 

traveled in urban versus rural areas and among states in different regions of 

the country (Federal Highway Administration, 1978). Diurnal variations in NO X 

emissions associated with motor vehicle traffic are especially important 

because of their potential impact on ambient air quality. Table 3-11 summarizes data on NOX emissions from mobile sources. 

TABLE 3-11. SUMMARY OF NO xEMISSIONS FROM MOBILE SOURCES 

Vehicle type Emissions, g/mi Comments Reference 

Gasoline-fueled 0.41 to 1.84 In-use; equipped Smith and 

passenger cars with three-way Black, 1980 

catalysts; four 

test cycles used 

Diesel passenger 0.84 to 3.15 In-use, 1977 to 1979 Gibbs et al., 

cars model years; tested 1983 

at ca. 35,000 mi 

Heavy-duty trucks 

Gasoline (leaded) 12.29 to 15.58 One test cycle Dietzman et al., 

Diesel 17.47 to 42.40 One test cycle; 1980, 1981 

several diesel fuels 

3.5.1.2 Natural Sources and Emissions 

3.5.1.2.1 Natural sources and emissions of volatile or anic compounds. This 

section presents a brief overview of the nature and quantity of hydrocarbon 

emissions from biogenic sources. For detailed information, the reader is 

00 
referred to severai excellent review articles (Altshuller, 1983; Bufalini and 

Arnts, 1981; Dimitriades, 1981). 

To date, isoprene and the monoterpenes are the only biogenic .hydrocarbons 

identified as emissions from vegetation (e.g., Sanadze and Dolidze, 1962; 

Rasmussen, 1964, 1970; Evans et al., 1982). Other volatile organic emissions 

from vegetation have been documented, but consist of oxygenated organic 

compounds, e.g., camphor and 1,8-cineole (see, e.g., Graedel, 1978). Isoprene 

and the monoterpenes are of interest to atmospheric scientists largely because 

they are volatiie enough to be released under normal environmental conditions, 

because they are abundant relative to other biogenic VOCs, and because they 

have been shown to be potential ozone precursors. The commonly identified 

monoterpenes are a-pinene, p-pinene, camphene, 0 3 -carene, limonene, myrcene, 

and p-phellandrene. As a general rule, coniferous trees emit primarily 

monoterpenes, and deciduous trees emit isoprene. 

Biogenic emission rates. Biogenic emission rates have been determined 

almost exclusively by techniques that involve enclosing the entire plant or a 

portion of it, such as a branch of a tree, in a bag or chamber constructed of 

light, transparent material. Isoprene and monoterpene emission rates given 

here [as pg (g dry biomass) -1 hr -~ ] were measured by the enclosure method. 

Because biogenic emission rates are species- and temperature-dependent, the 

species tested and the enclosure temperature are also given. 

Isoprene emission rates reported by Evans et al. (1982) for various 

species of trees range from 3 pg g -1 hr-1 for spruce (Picea sitchensis, 28 ° C) 

to 233 pg g-1 hr-1 for wiilow (Salix babylonica, 30°C). Reported emission 

rates for oak range from 9 pg g -1hr-1 (Quercus virginiana, 30 ° C; Zimmerman, 

1979) to 49 pg g-1 hr-1 (Quercus agrifolia, 30 ° C; Winer, 1983). The reader is 

referred to the studies by Zimmerman (1979), Evans et al. (1982), and Winer 

(1983) for additional.rate data and experimental details. 

Monoterpene emission rates are probably best - documented for pine. Rates 

obtained by Winer (1983) ranged from 0.6 pg g -1 hr -1 (Pinus radiata and Pinus 

halepensis, 30° C) to 2 pg g -1 hr-1 (Pinus canariensus, 30 0 C). For Pinus taeda, 

Arnts et al. (1978) reported a rate of 4 pg g -1 hr -1 (at 30 ° C) and Knoppel et 

al. (1982) reported a rate of 1 pg g -1 hr-1 (at 30 ° C). Among the higher monoter - 

pene emission rates are those for Pinus clausa, 11 pg g -1 hr 1 at 30°C 

(Zimmerman, 1979) and for pouglas fir (Pseudotsuga taxifolia), 15 pg g-1 hr 1 

at 30° C (Knoppel et al., 1982). See aiso the studies by Rasmussen (1972) and 

by Evans et al. (1982) for additional monoterpene emission rate data. 

3-99 
Besides temperature, biogenic emission rates are affected by other environmental factors. Rasmussen (1972) reported that emission rates varied with 

species, plant maturity, resin gland integrity, and leaf temperature. Dement 

et al. (1975) found that the emission rate of monoterpenes from Salvia mellifera 

(California Black Sage) is dependent on the vapor pressures of the terpenes, 

the humidity, and the amount of oil present on the surface of the leaf; but 

found that the emission rate is not directly dependent on the photosynthetic 

activity or on the stomatal openings of the plant. 

Tingey and his coworkers, in extensive studies on factors effecting 

isoprene emission rates in live-oak seedlings (Tingey et al., 1981), found 

light intensity to be a chief determinant of emissions, which decreased to 

near zero levels in the dark. In contrast to isoprene, monoterpene emission 

rates did not appear to be influenced by light intensity but were affected by 

temperature (Tingey et al., 1980); a log-linear increase in emission rates of 

monoterpenes with temperature was observed in studies of slash pine. 

The validity of emission rate data obtained by the bag enclosure technique 

has been widely discussed because of uncertainties associated with (1) isolating 

the vegetation in an artificial environment; (2) possible damage to isolated 

vegetation; (3) representativeness of emission rates measured from just one 

branch; and (4) relationship of emission rates to ambient air concentrations 

of biogenic hydrocarbons. 

Attempts to validate the bag enclosure method have focused on comparing 

enclosure emission estimates with those obtained by alternate procedures, such 

as micrometeorological gradient procedures (e.g., Lamb et al.', 1983; Knoerr 

and Mowry, 1981). In a study of Pennsylvania hardwood forest, the gradient 

profile procedure gave a flux of 8,000 pg m-2hr-1 , while the enclosure technique yielded 7,300 pg m-2hr-1(Lamb et al., 1983). Good agreement has been 

reported, also, for a-pinene emission fluxes measured by a micrometeorological 

procedure and by the enclosure method (Knoerr and Mowry, 1981). 

Although the micrometeorological approach yields mass fluxes similar to 

gthe enclosure method, it, too, has certain limitations. For example, the 

measurement of small vertical gradients above a forest canopy and the application of surface layer theory to non-ideal sites can lead to erroneous results. 

Such difficulties can largely be avoided by simulating the forest emissions 

with an inert tracer release (such as SF 6 ) and measuring ambient concentrations 

of the tracer and biogenic gases along downwind sample lines (e.g., Arnts and 

3-100 
Meeks, 1981). Isoprene fluxes obtained using the tracer procedure in a central 

Washington oak grove compared well with flux estimates determined simultaneously with the enciosur.e technique (Allwine et al., 1983). 

Biogenic emission inventories. Development of a biogenic emission inventory requires knowledge of (1) emission rates for individual species; (2) the 

vegetation coverage of an area, by species; (3) leaf biomass per tree (derived 

through allometric equations); and (4) a biomass factor for the forested area 

(derived from (2) and (3). 

Table 3-12 contains a listing of area-wide biogenic emission fluxes that 

have been reported for the United States and portions thereof. Accuracy of 

the estimates reported inr Table 3-12 depends upon the size of the area for 

which the inventory has been prepared. Many of the problems and uncertainties 

encountered in preparing inventories have been discussed in detail in the 

literature (Altshuller, 1983; Zimmerman, 1979, 1980; Wells, 1981; Box, 1981; 

Dimitriades, 1981). 

TABLE 3-12. AREA-WIDE BIOGENIC EMISSION FLUXES 

Emission 

flux, 

Location pg m'2 hr -1 Comment Reference 

South Coast Air Basin, <780 

California 

Lake Tahoe, Caiifornia 1950 Entire basin 

Lake Tahoe, California 2438 Forested area 

of basin 

San Francisco Bay Area, 1388 

California 

San Francisco Bay Area, 2265 Daytime 

California 

San Francisco Bay Area, 777 Nighttime 

California 

Tampa/St. Petersburg, 2540 

Florida 

Southeastern Virginia 8890 Forested area only 

Pennsylvania 1660 

Houston, Texas 1170 

United States 1712 

United States 1099 

United States 884 

Winer (1983) 

JSA, Inc. (1978) 

JSA, Inc. (1978) 

Sandberg et al. 

(1978) 

Hunsaker (1981) 

Hunsaker (1981) 

Zimmerman (1979) 

Salop et al. (1983) 

Flyckt et al. (1980) 

Zimmerman (1980) 

Marchesani et al. 

(1970) 

Zimmerman (1977) 

Zimmerman (1978) 

3-101 
Considering all the variables that affect biogenic emissions and their 

inventories, it is somewhat surprising that the area-wide emission fluxes 

listed in Tabie 3-13 show no more variation than they do. With the exception 

of the Southeastern Virginia area, which is a forested region with high biomass 

coverage, most of the values in Table 3-12 differ by less than a factor of 

three. 

3.5.1.2.2 Natural sources and emissions of nitro en oxides. Natural emissions 

of nitrogen oxides (NO x) originate from the oxidation of nitrogen gas by 

electrical discharge in the atmosphere, from the.ammonification of organic 

nitrogen during biological decomposition, and from the oxidation of organic 

nitrogen during forest fires. Nitrogen fixation and electrical discharge are 

normal processes of the nitrogen cycle that convert inert nitrogen gas to 

biologically useful nitrate or ammonia. For a discussion of the nitrogen 

cycle relative to NOxemissions, the reader is referred to Air Quality Criteria 

for Nitrogen Oxides (U. S. Environmental Protection Agency, 1982a). Table 3-13 

provides estimates of some of the nitrogenous compounds emitted from respective 

transformation processes. 

Although 40 to 108 tg NOX N per year has been estimated to be released 

from terrestrial sources to the atmosphere, the bulk is reabsorbed and only 8 

to 25 Tg NOx-N escapes to the troposphere (Robinson and Robbins, 1975; S6derlund 

and Svensson, 1976). Hill (1971) reported that NO and NO 2are absorbed, to 

some extent, from the atmosphere by plants. Using data obtained from the 

experiments of Marakov (1969), and those of Kim (1973), Soderlund and Svensson 

(1976) estimated that soil contributes to the atmosphere between 1 and 14 Tg N 

in the form of NO and NO 2 . They also hypothesized that a net flow occurs of 

NOxfrom terrestrial to aquatic systems, but concluded that losses from aquatic 

ecosystems to the atmosphere are insignificant. The principal source of 

gaseous NOxin terrestrial systems is believed to be the decomposition of 

nitrates (Soderlund and Svensson, 1976). 

Although more is known about the natural biological and chemical processes 

in the environment that produce emissions of nitrogen oxides -than those that 

produce VOC, the problems with actually quantifying such emissions exceed the 

problems associated with quantifying natural emissions of VOC. As indicated 

in Chapter 4, the limits of detection of inethods for measuring nitrogen oxides 

are such that low-level measurements are often not reliable. Techniques for 

estimating NOxemissions from such sources as lightning and biological processes 

3-102 
TABLE 3-13. GLOBAL ESTIMATES OF NITROGEN TRANSFORMATION 

(Tg N/yr) 

Range of estimates References a 

Biologicai fi+ation 

(N2 -> NH4 ) 54 to 270 1-5 

Electrochemical fixation 

lightning (N 2 NO ) 10 to 40 2, 6, 7, 8, 10 

atmospheric 02 602) 14 to 20 2, 4 

Biological denitrification 

(NO3 N2 ) 96 to 190 2, 3 

(NO3 N2 0) 20 to 340 2-4 

combined 83 to 270 5, 8, 9 

Industrial denitrification 

(Organic -N a NOx ) 14 to 19 2-4 

(Other a NOx ) 30 to 36 2, 3, 5 

Atmospheric denitrification 

(NH3 -> NOx) 3 to 30 2, 3 

Natural NO emissions 

from lan~ and sea 40 to 210 3, 4 

NH3 emissions to atmosphere 

from land and sea 110 to 850 2-4 

a1. Delwiche (1970). 

2. Burns and Hardy (1975). 

3. Soderlund and Svensson (1976). 

4. Robinson and Robbins (1975). 

5. Liu et al. (1977). 

6. Crutzen and Ehhalt (1977). 

7. Noxon (1976). 

8. Sze and Rice (1976). 

9. Council for Agricultural Science and Technology (1976). 

10. Chameides et al. (1977). 

3-103 
are virtually nonexistent. In addition, scaling of emissions from such sources 

as bacterial nitrification and denitrification for use in preparing area-wide 

emission inventories is not possible. Thus, the emissions reported in this 

section should be taken as very gross approximations that serve to identify 

natural sources of NOx and to present the estimated relative magnitude of 

emissions from such sources. 

3.5.2 Representative Concentrations of Ozone Precursors in Ambient Air 

As discussed earlier in this chapter (Section 3.3), nonmethane organic 

compounds (NMOC) and the oxides of'nitrogen (NOx ) in the presence of sunlight 

react to form ozone and other photochemical oxidants. The reaction sequence is 

complex, so that dependable precursor-oxidant relationships are difficult to 

establish. Factors such as absolute NMOC and NO xconcentrations, relative 

NMOC and NOxconcentrations (NMOC/NOxratios), NMOC reactivity, and NOxcomposition are known to affect the photochemical reactions that produce ozone and 

other oxidants in ambient atmospheres. Concentration-based NMOC/NO xratios 

are used in some precursor-oxidant models. Ratios of NMOC/NO xthat are 

concentration-based are calculated directly, using measured concentrations of 

NOx expressed as NO2(ppm or ppb) and measured concentrations of hydrocarbons 

expressed as carbon (ppm or•ppb C). The latter is easily obtained from gaschromatographic measurements, since the chromatograph yields a known response 

per carbon atom. This section provides summaries of NMOC and NO xconcentrations recorded at various urban and nonurban locations in the United States. 

In addition, HC/NOx(or NMOC/NOx) ratios are given for some urban areas. 

3.5.2.1 Concentrations of Nonmethane Organic Compounds in Ambient Air. The 

NMOC data in this section are segregated into (1) nonmethane hydrocarbons 

(NMHC) and (2) oxygenated hydrocarbons. The concentrations reported here for 

NMOC were obtained by gas chromatographic methods for the identification and 

quantification of individual NMOC species (see Chapter 4). A fairly.substantial 

data base exists for characterizing urban nonmethane hydrocarbon concentrations. 

Measurements of nonurban hydrocarbon levels, as well as both nonurban and 

urban oxygenated hydrocarbons, are much more limited. Among oxygenated hydrocarbons, aldehydes have received the most attention. Insufficient information 

exists for establishing ambient air concentrations of other classes of oxygenated 

hydrocarbons such as alcohols, ketones, acids, and ethers. 

3-104 
3.5.2.1.1 Urban nonmethane hydrocarbon concentrations. Most of the data on 

ambient air concentrations of nonmethane hydrocarbons (NMHC) have been obtained 

during the 6:00 to 9:00 a.m. time period. Since urban hydrocarbon emissions 

peak during that period of the day and atmospheric dispersion is limited, 

these concentrations generally reflect maximum diurnal levels. Table 3-14 

lists the mean and range of NMHC concentrations recorded in a number of urban 

areas throughout the United States. For most urban areas included in the 

table, a mean NMHC value between 400 and 900 ppb C was observed, though mean 

concentrations in some cities (e.g., Houston, Las Vegas, and Los Angeles) are 

in excess of 1000 ppb C. The data in Table 3-14 are not meant to serve as a 

comparison of NMHC levels in various cities but rather are shown to indicate 

the mean and range of concentrations that have been reported. Comparisons are 

invalid because of major differences in sample numbers, site classifications, 

and seasonal sampling periods. In many cases, the range of values reporte'd1n 

Table 3-14 might not reflect the true maximum and minimum concentrations that 

occur in a particular urban area. Most of the hydrocarbon sampling programs 

were of short duration (M1 month) and in some cases were not operated on'a 

daily basis. For example, the relatively high mean values reported in Las 

Vegas are undoubtedly the result of the fact that ambient air samples were 

only analyzed for hydrocarbons on days when conditions were appropriate for 

oxidant 'formation. It is probably safe to assume, however, that NMHC levels 

during the 6:00 to 9:00 a.m. time period in major urban areas will usually 

exceed 50 ppb C but seldom surpass 10,000 ppb C. 

Species in the C2 -C10molecular-weight range dominate the hydrocarbon 

composition of urban atmospheres, with the alkanes generally constituting 50 

to 60 percent of the hydrocarbon burden, aromatics 20 to 30 percent, 'and 

alkenes and acetylene making up the remaining 5 to 15 percent (Sexton and 

Westberg, 1984). The alkane fraction is usually dominated by species in the 

C4-C6molecular-weight range. Predominant aromatics include benzene, toluene, 

ethylbenzene, and the 'xylenes. The most abundant alkenes are ethylene and 

propene. The studies cited in Table 3-14 provide information on the individual 

species of NMHC found in urban atmospheres. 

3.5.2.1.2 Nonurban nonmethane hydrocarbon concentrations . Nonurban nonmethane 

hydrocarbon concentrations are generally one to two orders of magnitude lower 

than those measured in urban areas (Ferman, 1981; Sexton and Westberg, 1984). 

Concentrations of individual species seldom exceed 10 ppb C. Total hydrocarbon 

3-105 
TABLE 3-14. NONMETHANE HYDROCARBON CONCENTRATIONS 

MEASURED BETWEEN 6:00 and 9:00 a.m. IN VARIOUS UNITED STATES CITIES 

Mean 

NMHC 

City concn., 

(Date) ppb C Range Reference 

Atlanta (1981) 491 113 to 1677 Westberg and Lamb (1983) 

Baltimore (1980) 659 51 to 2798 Sexton and Westberg (1984) 

`Boston (1980) 569 83 to 4750 Sexton and Westberg (1984) 

Cincinnati (1981) 840 260 to 1870 Holdren et al. (1982) 

Detroit (1981) • 330 60 to 1720 Kelly et al. (1986) 

Houston (1976) 1414 356 to 16,350 Sexton and Westberg (1984) 

Houston (1978) 1679 400 to 4500 Lonneman (1979) 

Las Vegas (1980) 2506 689 to 4515 Naylor et al. (1981) 

Las Vegas (1983) 2762 1835 to 4590 Naylor et al. (1984) 

Los Angeles (1968) 3388 N.A. a Lonneman (1977) 

Los Angeles (1982) 2920 390 to 6430 Grosjean and Fung (1984) 

Milwaukee (1981) 324 24 to 3116 Sexton and Westberg (1984) 

Newark (1980) 732 89 to 6946 Sexton and Westberg (1984) 

New York (1969) 830 N.A. Lonneman (1977) 

Philadelphia (1979) 669 305 to 1710 Sexton and Westberg (1984) 

St. Louis (1973) 817 N.A. Lonneman (1977) 

Tulsa 426 103 to 3684 Eaton et al. (1979) 

Washington, DC (1980) 671 210 to 2953 Sexton and Westberg (1984) 

aData are not available. 

concentrations range up to —150 ppb C, but usually fall in the range of about 

5 to 100 ppb C. Alkanes comprise the bulk of species present, with C 2-05 

compounds most abundant. Ethylene and propene are occasionally reported at 

concentrations of 1 ppb C or less, and toluene is usually present at —1 ppb C. 

Table 3-15 provides a summary of the range of hydrocarbon concentrations 

measured at various nonurban locations in the United States. Samples were 

carefully selected at most of the sites in order to guarantee their nonurban 

character. At the coastal and near-coastal sites, only those samples collected 

upwind of manmade sources (onshore advection) were included. The nonmethane 

3-106 
TABLE 3-15. NONMETHANE HYDROCARBON CONCENTRATIONS 

MEASURED IN NONURBAN ATMOSPHERES 

Location 

Species 

analyzed 

Concentration 

range, ppb C Reference 

Belfast, ME C 2 - C 5 10 to 22 Sexton and Westberg (1984) 

Benicia, CA C2 - C5 7 to 14 Sexton and Westberg (1984) 

Miami, FL C2 - C5 2 to 23 Sexton and Westberg (1984) 

Glascow, IL C2 - C lo 60 to 150 Chatfield and Rasmussen (1977) 

Janesville, WI C2 - -C1p 9 to 24 Sexton and Westberg (1984) 

Houston, TX C2 - C10 2 to 24 Sexton and Westberg (1984) 

Robinson, IL C2 - C lo 13 to 113 Sexton and Westberg (1984) 

Smoky Mtns., TN C2 - C 10 38 to 149 Cronn (1982) 

Northern Idaho Terpenes 0.1 to 18 Holdren et al. (1979) 

Virginia lsoprene 4 to 150 Ferman (1981) 

At1anta (urban) Isoprene 0 to 8 Westberg and Lamb (1983) 

Whiteface Mtn., 

NY Terpenes 6 to 84 Whitby and Coffey (1977) 

Elkton, MO Isoprene 0 to 28 Rasmussen et al. (1976.) 

Eastern TX a-pinene 0.1 to 8 Seila (1981) 

North Carolina a-pinene 0.6 to 13 Seila (1981) 

Colorado Terpenes 0 to 8 Roberts et al. (1983) 

hydrocarbon concentrations reported at coastal sites (Belfast, Benicia, Miami, 

and Houston) are definitely lower than those measured at most of the inland 

sites. It should be pointed out, however, that the numbers of samples measured 

for each of the nonurban locations listed in Table 3-15 is small. This, 

coupled with the fact that only a limited range of hydrocarbons were monitored 

~ 

in some cases, makes intersite comparisons tenuous at best. 

Ambient air concentrations of naturally emitted hydrocarbons (e.g., 

isoprene, a-pinene, P-pinene, ©-carene, and limonene) are generally reported 

only in nonurban hydrocarbon sampling programs. Because they are present at 

very low concentrations, natural hydrocarbons are extremely difficult to 

identify unequivocally when they mix with manmade emissions in an urban area. 

The one exception is isoprene, which has been reported in both urban and 

3-107 
nonurban sampling programs. Monoterpene (C10H16) concentrations in ambient 

air seldom exceed 20 ppb C. Average concentrations of a-pinene, the most 

commoniy reported monoterpene, are usually below 10 ppb C. During the summer 

months, isoprene concentrations as high as 150 ppb C have been measured (Ferman, 

1981), but maximum concentrations in the 30 to 40 ppb C range are more common. 

Ambient concentrations of the naturally emitted hydrocarbons are site-dependent, 

with the highest concentrations observed in or immediately adjacent to forested 

areas. Concentrations vary with season, as well, because natural hydrocarbon 

emission fluxes are directly related to the amount of biomass present and 

increase with temperature. In a recent review article, Altshuller (1983) has 

provided a more detailed discussion of natural hydrocarbons and their effect 

on air quality. 

3.5.2.2 Concentrations of Nitrogen Oxides in Ambient Air. Ambient air levels 

of nitrogen oxides have been monitored throughout the United States for a 

number of years. Since nitrogen dioxide (NO 2 ) is the only oxide of nitrogen 

for which an NAAQS has been promulgated, it has received the greatest attention. 

The emphasis here is on NO xmeasurements that can be related to the diurnal 

photochemical processes that produce ozone. 

3.5.2.2.1 Urban NO concentrations. Concentrations of NOx , like hydrocarbon 

concentrations, tend to peak in urban areas during the early morning period 

when atmospheric dispersion is limited and automobile traffic is dense. Most 

of the NOxis emitted as nitric oxide (NO), but the NO is converted rapidly to 

NO2by ozone and peroxy radicals produced in atmospheric photochemical reactions. Since ozone levels and photochemical activity vary diurnally and from 

day to day, the relative concentrations of NO and NO 2can fluctuate significantly. Generally, urban NO concentrations peak during the 6:00 to 9:00 a.m. 

period, followed by a rapid decrease caused by the photochemica1 conversion of 

NO and NO 2and increased atmosphere mixing. Nitric oxide levels remain low 

during the daytime and then usuaily build up again through the nighttime 

hours. Nitrogen dioxide concentrations typically increase during the midmorning hours and then abate as the afternoon progresses. Levels of NO 2 

increase again following the late afternoon rush-hour period, often continuing 

to increase during the nighttime. 

The average NO xconcentration in urban areas of the United States is 

about 70 ppb, with NO and NO 2contributing about equally (Logan, 1983). 

Monitoring data for 1975 through 1980 showed that peak 1-hr NO 2concentrations 

M ~ : 
equalled or exceeded 400 ppb in Los Angeles 'and severai other Ca7ifornia 

locations, as well as at sites in Kentucky (Ashland) and Michigan (Port Huron). 

Cities with one peak hourly concentration exceeding 270 ppb during those years 

include Phoenix; St. Louis; New York City; Springfield (Illinois); Cincinnati; 

Saginaw and Southfield (Michigan); and more than a dozen sites in California. 

Reported hourly concentrations in excess of 140 ppb were quite common nationwide 

during the years between 1975 and 1980 (U.S. Environmental Protection Agency, 

1982a). 

Urban NOxconcentrations during the 6:00 to 9:00 a.m. period are of 

primary importance in terms of oxidant production. Average NO xlevels recorded 

in several urban areas during this morning period are listed in Table 3-16, 

which shows mean 6:00 to 9:00 a.m. NO xconcentrations in the range of about 50. 

to 150 ppb. Concurrent 6:00 to 9:00 a.m. hydrocarbon samples were also 

obtained in the studies reported in Table 3-16, and the hydrocarbon-NO xratios 

in each of these urban areas are included. 

TABLE 3-16, AVERAGE 6:00 to 9:00 a.m. NO CONCENTRATIONS AND 

HC/NOX RATIOS IN URBAN ARgAS 

Average NO , Average 

City ppb x HC/NOx References 

Atlanta 57 9 Westberg and Lamb (1983) 

Baltimore 85 10 Richter (1983) 

Boston 63 10 Richter (1983) 

Houston 125 13 Westberg et al. (1978b) 

Detroit 67 5 Kelly et al. (1986) 

Linden, NJ 59 16 Richter (1983) 

Los Angeles 147 10 U.S. Environmental 

Protection Agency (1978a) 

Milwaukee 66 5 Westberg and Lamb (1983) 

St. Louis 77 8 U. S. Environmental Protection 

Agency (1978a) 

Tulsa 46 13 Eaton et al. (1979) 

Washington, DC 94 14 Richter (1983) 

3-109 
Hydrocarbon concentrations (ppb C) exceeded the NOxlevels by a factor of 

5 to 16 during the 6:00 to 9:00 a.m. period. Smog chamber experiments indicate 

that significant quantities of ozone can be produced when HC/NO xratios are in 

this range. Indeed, ozone production has been observed in the vicinity of 

most of the cities referenced in Table 3-16. 

3.5.2.2.2 Nonurban NO concentrations. Concentrations of NO xin "clean" 

remote environments are usually below 0.5 ppb (Logan, 1983). For example, 

median concentrations measured on Niwot Ridge in Colorado are about 0.3 ppb in 

the summer and 0.24 ppb in winter. In exceptionally clean air, NO xconcentrations as low as 0.015 have been recorded (Bollinger et al., 1982). Slightly 

higher NOxconcentrations have been reported at other remote locations in the 

western United States and Canada. Kelly et al. (1982) deduced a mean NOx 

concentration of about 1 ppb from measurements in South Dakota. At the South 

Dakota site, nitric oxide generally contributed, less than 20 percent of the 

total NOx . Measurements of NO xduring the 1970s at rural locations in Montana 

(Decker et al., 1978) and Saskatchewan (McElroy and Kerr, 1977) yielded average 

concentrations similar to those recorded in South Dakota.' 

At a rural site in Louisiana, Kelly et al. (1984) found mean•concentrations 

of N1 ppb NO and —4 ppb NO 2 . The same investigators observed mean concentrations 

of —0. 7 ppb NO and —1. 6 ppb NO 2at a rural si te i n Vi rgi ni a(Kel ly et al ., 

1984). 

In the northeastern United States, nonurban NO xconcentration.s appear to 

exceed those in the west by about a factor of ten. A median NO xconcentration 

of 6.6 ppb was derived from data collected at nine rural sites utilized in the 

Sulfate Regional Experiment (SURE) program (Mueller and Hidy, 1983). Median 

concentrations at the individual stations, which extended eastward from the 

Ohio River Valley to the Atlantic Coast, varied from 2 to 11 ppb. Measurements 

at nonurban sites in Pennsylvania and Louisiana, during the summer of 1975 

showed mean hourly NO xconcentrations of 4.7 and 4.1 ppb, respectively (Decker 

et al., 1978). Nitric oxide composed approximately 40 percent of the total 

NOxat these latter two nonurban sites. 

3.6 SOURCE-RECEPTOR (OXIDANT-PRECURSOR) MODELS 

In order to apply knowledge of the atmospheric chemistry of ozone and 

other photochemical oxidants and their precursors during their dispersion and 

3-110 
transport, models describing these phenomena have been developed in a variety of 

forms over the past 15 years. Most of these models relate the rates of precursor 

emissions from mobile and stationary sources, or precursor atmospheric concentrations, to.the resulting ambient concentrations of secondary pollutants that 

impact receptors at downwind sites. Eor this reason they have been described 

as source-receptor or oxidant-precursor models. 

A wide variety of source-receptor models exists, ranging in complexity 

from empirical relationships based on air monitoring or smog chamber data to 

complex computer-based grid or trajectory airshed models that may contain 

detailed emission inventories, sophisticated dispersion and transport submodels, and lengthy chemical reaction mechanisms. Moving-box models represent 

an approach of intermediate complexity. 

A11 presently available source-receptor models require a degree of simplifying assumptions to deal with practical limitations imposed by existing 

computer capabilities, time and cost constraints, or lack of knowledge concerning inputs such as boundary conditions, emissions, wind fields, or detailed 

reaction mechanisms. The reliability and applicability of any particular 

model therefore depend upon its specific limitations, data requirements, and 

degree of validation against experimental data from ambient air measurements 

or environmental chamber runs. 

A detailed discussion of the range of available source-receptor models, 

and their validation and applications, is beyond the scope of this document. 

Instead, brief conceptual descriptions are provided of the major classes of 

source-receptor models. It is important to recognize that such models are 

continually undergoing evolution, revision, and refinement, particularly as 

knowledge of atmospheric chemistry grows and, for example, as more sophisticated approaches to dealing with boundary conditions become available. Even 

under the best circumstances, however, the present generation of sourcereceptor models should be viewed as being most useful for investigating the 

relative effects on air quality of particular emission sources or emission 

control strategies, rather than for predicting absolute concentrations of 

secondary pollutants resulting from specific precursor emissions. 

3-111 
3.6.1Definitions, Descriptions, and Use 

Current air-quality, or source-receptor, models can be classified as 

either statistical or computational/dynamic. Statistical models are generally 

based on an analysis o.f historical air quality data. An example of a statistical 

model is the linear rollback concept. 

Computational, or dynamic, models attempt to describe mathematically the 

atmospheric chemical and physical processes i.nfluencing air pollution formation 

and impacts. Examples of computational models include trajectory and grid 

airshed models. The basis of these models is the solution of the atmospheric 

diffusion equation (Bird et al., 1960; Liu and Seinfeld, 1975). 

Two phenomenologically different approaches have been employed in dynamic 

models with respect to the coordinate systems chosen. A coordinate system 

fixed with respect to the earth is termed Euierian, while in Lagrangian models 

the reference frame moves with the air parcel whose behavior is being simulated. 

In the following section these and other models are briefly described. 

3.6.1.1Statistical Models . Two widely used models of this kind are linear 

rollback and the Appendix J approach, both of which were employed by EPA prior 

to the advent of more sophisticated dynamic modeling approaches. The concept 

of linear rollback is based on the assumption that'ambient concentrations of 

air pollutants are directly proportional to emissions; and that a given reduction in emissions will result in a proportional decrease in the maximum ambient 

concentrations of that po1lutant. In principle, linear rollback models should 

be applied only to inert primary pollutants and their original use was for 

unreactive pollutants such as carbon monoxide (Larsen, 1969). Such models 

have been applied, however, in modified form to secondary pollutants such as 

ozone (Barth, 1970). 

A prominent example of a statistical model was the Appendix J relationship 

developed by EPA to relate maximum 1-hour average ozone concentrations in 

several United States cities to 6:00-to-9:00 a.m. average nonmethane hydrocarbon 

concentrations in those cities (F.R., 1971). This relationship was used to 

calculate the amount of NMHC control needed to achieve the Federal standard 

for photochemical oxidants. 

Two important limitations of past statistical methods were their failure 

to take into account the transport of primary and secondary pollutants from 

source areas to downwind receptor sites, and lack of recognition of the role 

of oxides of nitrogen in the formation of ozone and other photochemical oxidants. 

3-112 
These and other weaknesses led EPA to abandon the use of statistical models in 

state implementation plans (F.R., 1981). 

3.6.1.2Trajectory Models . Figure 3-17(a) contains a schematic representation 

of the trajectory model approach, in which a hypothetical air parcel moves 

through the area of interest along a path calculated from wind trajectories. 

Thus, a moving-coordinate system (Lagrangian) describes pollutant transport 

under the influence of local meteorological conditions. Emissions are injected 

into the air parcel and undergo vertical mixing and chemical transformations. 

The data requirements fior trajectory modeis include: (1) initial concentrations of all relevant pollutants and species; (2) rates of emissions of 

NMOC and NOX precursors into the parcel along its trajectory; (3) meteorological 

characteristics such as wind speed and direction; and (4) solar ultraviolet 

radiation. Various trajectory models exhibit a range of sophistication and 

complexity with regard to such elements as chemical mechanisms (Atkinson 

et a1., 1982c), emission inventories (Braverman and Layland, 1982), treatment 

of vertical mixing (Whitten and Hogo, 1978; Drivas, 1977; Meyers'et al., 1979; 

Lloyd et al., 1979; Lurmann et al., 1979), and trajectory determination (U.S. 

Environmental Protection Agency, 1980c; Whitten and Hogo, 1978). Basic limitations of trajectory models include the amount and density of data required 

for precise calculations of emissions input, chemical transformations, and 

diiution; neglect of horizontal wind shear; neglect of cell volume changes 

resulting from convergence and divergence of the wind field; and uncertainties 

in boundary conditions, including conditions aloft (Liu and Seinfeld, 1975). 

Conversely, moving-cell models provide a dynamic description of atmospheric 

source-receptor relationships that is simpler and less expensive to derive 

than that obtained from fixed-cell models. 

The simplest form of trajectory model is the empirical kinetic modeling 

approach (EKMA). This modeling approach was developed from earlier efforts 

(Dimitriades, 1972) to use smog chamber data to develop graphical relationships 

between morning NMOC and NOxlevels and afternoon ozone maximum concentrations. 

Dodge (1977a,b) presented an approach in which smog charober data (Dimitriades, 

1970; 1972) were used to test and validate a photochemical kinetics model 

(Durbin et al., 1975). Dimitriades (1977) used the resulting 03 isopleths to 

define a method for obtaining the relative degree of precursor emissions 

control needed to achieve a given percentage reduction in ozone. In the EKMA 

approach, which has been extensively utilized (U.S. Environmental Protection 

3-113 
\ ~Z 

4CE/TIME TRACK THROUGH SOURCE\ } 

GRID; DEFRIVED FROM WIND DATA ~F 

SUNLIGHT GIVEN AS 

FUNCTION OF TIME 

i 

~ TIME-DEPENDENT MIXING AND 

~ REACTION COMPUTED FOR AIR 

 PARCEL UP TO MIXING HEIGHT, h 

h 

U 

POLLUTANT INFLUXES AT ANY 

ELEVATION (INCLUDING THE GROUND); ~x 

IMPOSED BY EMISSION'SOURCE FUNCTIONS 

(a) TRAJECTORY MODEL 

1*3 TRANSPORT TOP OF MODELING 

_ REGION 

CHEMISTRY•ELEVATED EMISSIONS 

 TRANSPORT TOP OF TRANSP_ORT 

_ MI X ED 

E TRANSPORT 

CHEMISTRY-ELEVATED EMISSIONS 

LAYER 

w TRANSPORT TRANSPORT 

~ 

4if 

TRANSPORT 

$ TRANSPORT

CHEMISTRY-ELEVATEDEMISSIONS

TRANSPORT 

* TRANSPORT 

~ 

~ 

CHEMISTRV-ELEVATED EMISSIONS 

~ TRANSPORT TRANSPORT 

. SURFACE iREMOVAL 

1 TO tO km -~. 

_IGROUNDSURFACE 

GRID CHEMISTRY AND TRANSPORT 

L GRID SPECIFICATION (b) GRID MODEL ~ 

RISING MIXED 

HEIGHT ENTRAINMENT OF 

t POLLUTANTS ALOFT 

WIND DIRECTION ~ II 

I 

I 

~ S..CHEMICAL ` ~ -' • ~ 

ADVECTIVE TRANSFORMATIONS 

Z(U) INFLOW 

r

SOURCE 

—

EMISSIONS L ~ 

~ A VECTIVE 

Qx Ox OUTFLOW 

(c) BOX MODEL 

Figure 3-17. Schematics of the three types of dynamic models. 

3-114 
Agency, 1978a,b; 1980c), the Ozone Isopleth Plotting Package (OZIPP) (Whitten 

and Hogo, 1978) is used to generate ozone isopleths at various levels of 

sophistication corresponding to "standard" EKMA, "city-specific" EKMA, or the 

"simplified trajectory" model (F.R., 1979). These models are designated as 

Levels IV, III, and II, respectively, with Level IV being the least sophisticated and Level II the most sophisticated. Substantial documentation and 

guidance concerning the use of OZIPP and a more flexibie modified version of 

the Program (OZIPM) are available (U.S. Environmental Protection Agency, 

1981a, 1984b,c). 

The sensitivity of this method to variables such as the input hydrocarbon 

composition and the choice of chemical kinetics mechanisms has been reported 

(Carter et al ., 1982; Jeffri es et al ., 1981; Shafer and Sei nfel d, 1985) and 

further refinements in the EKMA approach to accomodate these factors have been 

made. For example, site-specific versions of EKMA allow the user to select 

particular dilution rates, emissions, and solar intensity applicable to the 

city or airshed of interest. Another version allows the user to employ an 

alternative mechanism for making the EKMA calculations (U.S. Environmental 

Protection Agency, 1984b). 

Because it is often recommended by EPA for use in determining needed 

precursor reductions and is in widespread use, EKMA is discussed here in 

detail. An example of an EKMA diagram is presented in Figure 3-18, which 

shows ozone isopleths for sites downwind of an urban source area in which 

morning precursor emissions are high. The isopleths in this diagram depict 

downwind, peak 1-hour ozone concentrations as an explicit function of initial 

(i.e., morning) concentrations of nonmethane hydrocarbons (NMHC) and nitrogen 

oxides (NOx ); and as indirect functions of (1) NMHC and NO xemissions occurring 

later in the day; (2) specified meteorological conditions; (3) reactivity of 

the precursor mix; and (4) concentrations of ozone and precursors transported 

from upwind areas (U.S. Environmental Protection Agency, 1977). The relationships between ozone and its precursors that are depicted in Figure 3-18 are 

based on empirical data and the application of a chemical kinetics model 

(Dodge, 1977a; Whitten and Hogo, 1977) that has been adjusted by comparing 

model predictions against smog-chamber data obtained by irradiating automobile 

exhaust (Dimitriades, 1972). Alternatively, EKMA diagrams can be constructed 

using more recent chemical mechanisms that have been tested against smog 

chamber data (Gipson, 1984; Whitten and Gery, 1986). 

3-115 
E 

a 

a 

Z O 

H 

Q 

o: 

H 

z 

W 

U 

Z 

O 

C~ 

y 

W ~ 

C 

X O 

Z 

W C9 I 

O 

cr. F- :_.. 

~ 

28 

24 

20 

16 

12 

08 

D4 

u u.Z 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8 2.0 

NONMETHANE HYDROCARBON CONCENTRATION, ppm 

Figure 3:18 Example of EKMA diagram for high-oxidant urban area. 

3-116 
For the EKMA-diagram given here as Figure 3-18, several general, interrelated features are of interest. First, the isopleth iines in the lower 

right quadrant are more or less parallel to the abscissa (the NMHC concentration). Second, the isopleth lines in the upper left quadrant are slanted 

with respect to the ordinate (the NOX concentration). Because of the shape of 

the isopleth lines, varying the NMOC or NOxconcentration will have different 

effects on ozone at different NMHC/NO xratios. 

These two features of the diagram are related to the underlying photochemistry (see, e.g., Whitten, 1983)`. For extremely low NO xconcentrations 

(i.e., high NMHC/NOxratios), where the lines parallel the abscissa, the 

formation of 0 3is (1) insensitive to changes in NMHC concentrations, and (2) 

NOx-limited; that is, changes in NO xconcentrations cause co-directi,onal 

changes in peak 0 3concentrations. As described early in Section 3.3 and in 

Section 3.3.1, the atmospheric oxidation of hydrocarbons produces an abundance 

of peroxy radicals (R0 2 •), more than enough to oxidize NO to NO 2rapidly 'and 

completely. In addition, the NO xpresent, before being removed from the 

cyclic reactions via termination reaction.s with various radicals, completes a 

number of N0 QR 2 ~ -. NO2--- -d NO (+ 03 ) cycles, thus producing a number 

of 0 3molecules. Small-to-moderate changes in NMHC concentrations will tfierefore have little impact, since there will still be an abundance of R0 2 • radicals 

in the atmosphere. On the other hand, changing the already low concentration 

of NOxdoes not have an appreciable impact on the R02 radicals, but it do-es 

change co-directionally the number of 0 3molecules produced, since the photolysis 

of NO2and the oxidation of NO to NO xare essential for 0 3production. 

For moderately higher concentrations of NOx(lower NMHC/NO xratios), 

there is no longer a large abundance of R0 2 • relative to the NO present; and 

changes in NMHC, therefore, have a co-directional effect on R0 2 • and, hence,on 

03 . Thus, for moderate NMHC/NO xratios, the effects on 0 3formation of varying 

the concentrations of NMHC or NOxare simiiar in direction. 

Finally, for much higher NO xconcentrations (i.e., very low NMHC/NO x 

ratios), the dominant effects are the relative depletion of radicals, through 

the reactions of NO 2with these radicals (Equations 3-18 and 3-35, Section 

3.3); with a resulting decrease in the rate of reaction of peroxy radicals 

with NO (Equation 3-29, Section 3.3). The consequence of these effects is 

that the conversion of NO to NO2is very slow. At such low NMHC/NO xratios, 

therefore, increasing the NMHC concentration enhances 0 3formation but 

increasing NO xconcentrations inhibits it. 

3-117 
As examination of Figure 3-18 reveals, for an NMHC concentration of 0.6 

ppm C, for example, increasing NO xleads to increased 03until NMHC/NOxratios 

of about 5:1 to 6:1 are reached; further NOxincreases, leading to lower 

NMHC/NOxratios, inhibit 0 3formation. Thus, in this example, there is a 

"critical" ratio (approximately 5.6:1) at which the NO xeffect on 0 3changes 

direction. Besides this "critical" ratio, an "equal controi" NMHC/NOxratio 

also exists, above which the reduction of NO xis more beneficial in terms of 

03 reduction than an equal percentage reduction in NMHC. This ratio, for the 

isopleths shown in Figure 3-18, is roughly 8:1 to 9:1 for low levels of control, 

and as high as 20:1 for the levels of control needed to reduce 0 3to 0.12 ppm. 

Thus, for this particular case (Figure 3-18), the chemical mechanism modeling 

evidence suggests that (1) NOxcontrol will increase the peak downwind 0 3 

concentration at NMHC/NOx ratios of 5.6:1 or lower; (2) both NO xcontrol and 

NMHC control will be beneficial at somewhat higher ratios, with control of 

NMHC being more effective; and (3) for ratios above 20:1, NO xcontrol is 

relatively more effective in reducing 03 . 

The calculation of precursor controls necessary to reduce 0 3to 0.12 ppm, 

from the isopleths given in Figure 3-18, shows that NO xcontrol, although at 

first beneficial, is ultimately detrimental because it makes the reduction of 

03to 0.12 ppm more difficult: This can be demonstrated through use of the 

EKMA isopleths in Figure 3-18 as follows. For a high-oxidant atmosphere with, 

for example, a peak 0 3concentration of 0.30 ppm and an NMHC/NO xratio of 

12:1, a 74 percent control of NMHC (1.84 ppm C to 0.48 ppm C) will be needed 

in order to reduce downwind, 1-hour peak 03to 0.12 ppm through the unilateral 

control of NMHC (line AB in Figure 3-18). If, however, NO xis first controlled 

by, for example, 29 percent (from 0.152 ppm to 0.108 ppm, line AA'), this will 

cause a 15 percent reduction in 0 3(line AA') but it will also increase the 

NMHC control requirement (to reduce 0 3to 0.12 ppm) from 74 percent (line AB) 

to 81 percent (line A'B'). Since it is not ordinarily feasible to reduce 0 3 

to 0.12 ppm in a high-oxidant area through the unilateral control of NO x(it 

would take, in this case, an almost 85 percent control of NO x , line AC), it 

follows that 29 percent control (or any control) of NOx , would ultimately be 

detrimental for the situation represented by these isopleths. There may be 

situations, however, in which the control of NOxwill not increase the NMHC 

control required to achieve a given percentage reduction in ozone. It must 

be emphasized that the preceding discussion of the implications of controlling 

VOC and NOxis not applicable to all situations. These diagrams may differ 

3-118 
when different chemical mechanisms or other model input data are used. Furthermore, conclusions drawn from Figure 3-18 may differ if a different starting 

point on the diagram is used. 

As noted early in Section 3.3, the effects on 0 3formation of controlling 

NOX emissions are a matter of continuing discussion and research. The inhibition 

of 0 3formation by NOxunder some circumstances, as demonstrated by the EKMA 

diagram in Figure 3-18, has been shown in smog chamber studies (e.g., Glasson 

and Tuesday, 1970; Dimitriades, 1972). In a more recent study, in which he 

irradiated ambient air in Teflon bags, Kelly (1985) showed that the addition 

of NOxto the Detroit ambient air samples suppressed 0 3formation. Kelly also 

compared EKMA predictions with the results of his smog chamber study and 

concluded that EKMA correctly describes the respective effects of NO xand NMHC 

(or NMOC, as reported in the study) on ozone maxima. 

The studies cited above, as well as other documentation (e.g., Dimitriades, 

1977;.Liu and Grisinger, 1981; Chock et al., 1981; Kelly, 1985; and references 

therein), should be consulted for more complete discussions of the respective 

effects ,of NMHC and NO x.controls on resulting peak ozone concentrations. 

3.6.1.3 Fixed-Grid Models. Fixed-grid models, also called regional airshed 

models, are based on two- or three-dimensional arrays of grid cells fixed with 

respect to the earth (Eulerian) and are the most sophisticated source-receptor 

models presently available. As indicated in Figure 3-15(b), these models take 

into consideration the transfer of pollutants between adjacent cells, in 

addition to accounting for emissions, atmospheric chemistry, vertical mixing, 

and meteorological parameters. Such models are computationally complex and 

require the most extensive set of input data, but they also provide the most 

realistic treatment of the various processes involved in photochemical air 

pollution formation. 

Examples of grid models that have been applied to emission control 

strategies and their assessment include the Systems Applications, Inc., airshed 

model (Reynolds et al., 1979) and the Livermore Regional Air Quality (LIRAQ) 

model (MacCracken et al., 1978). These models are most often used to simulate 

days for which detailed aerometric data are available. The availability of 

such data depends upon the number and spacing of ground monitoring stations, 

the amount of data on wind and turbulence aloft, and the frequency of vertical 

soundings. Wind, temperature, inversion, and diffusivity data are interpolated 

over the field of interest as part of the process of preparing inputs for grid 

3-119 
models (Roth et al., 1976). In addition to atmospheric data, various emissions 

sources must be considered in each model. At the most sophisticated level, 

emission models and inventories are constructed to provide estimates, by 

category, of vehicular, aircraft, power plant, refinery, and distributed 

source emission rates, including their temporal and seasonal variations (Roth 

et al., 1976; Braverman and Layland, 1982; F.R., 1979). 

Shortcomings in grid models stem from the theoretical and computational 

complexities that are necessary in this type of simulation. Inaccuracies in 

grid-model predictions arise from: (1) theoretical deficiencies in the mathematical representation of atmospheric processes; (2) numerical inaccuracy in 

the solution of the atmospheric diffusion equation (Liu and Seinfeld, 1975); 

and (3) inadequate input data resulting from incomplete data bases. An 

incomplete understanding of advection and turbulent diffusion, necessitates 

the use of estimates or parameterizations to provide appropriate values 

(Seinfeld and Wilson, 1977). Atmospheric chemical kinetics descriptions are 

continually updated as new information-is obtained, but uncertainties associated with these mechanisms may be propagated during solution. In addition, 

sparse and often unrepresentative data are utilized to derive continuous 

fields (wind fields, turbulence, and mixing depths) over the region (Seinfeld 

and Wilson, 1977), a problem that is common to all dynamic models. 'In general, 

wind and turbulence data are rare1y collected aloft; surface data are much 

more abundant but still vary widely in terms of number, frequency, and quality 

of ineasurements (Roth et al., 1976). This implies that critical values, 

especially aloft, must often be estimated to provide initial, boundary, and 

operating conditions (Seinfeld and Wilson, 1977). Finally, uncertainty in 

grid-model solutions also arises from emission inventories that are poorly 

resolved, either spatially, temporally, or with respect to hydrocarbon 

reactivity specifications (Braverman and Layland,`1982). 

The structure and complexity of grid models also account for their utility. 

The increased temporal resolution afforded by grid models can provide minuteby-minute concentration estimates in each cell, and there can be as much 

spatial resolution as the data will allow. Since it accounts for specific 

atmospheric processes within the system, the model allows explicit insertion 

of new information (e.g., on meteorological or chemical processes) into the 

structure of the model. In addition, the impact of individual precursor 

sources may be analyzed with this type of model (Association of Bay Area 

Governments, 1979a,b). 

3-120 
3.6.1.4Box Models . Box models (Hanna, 1973; Demerjian and Schere, 1979; 

Derwent and Hov, 1980) are the simplest of dynamic models. They treat the , 

atmosphere as a single cell, bounded by the mixing layer,, with an.area on the 

order of 100 square miles [see Fi.gure 3-15(c)]. The chemist.ry within the box 

is affected by: (1) instantaneously mixed regional emissions, (2), dilution 

from lifting of the inversion, (3) ventilation and transport resulting from a 

characteristic wind,field, and (4) entrai,nment o.f species from aloft. Because 

the only consideration of spatial resolution occurs when the modeling boundaries 

are chosen, data requirements are minimal. Results can only be interpreted 

temporally, however, for a mass average of a species, and results can be 

strongly affected by uncertainties in boundary conditions.. 

3.6.2Validation and Sensitivity Analyses for Dynamic Models 

Dynamic models are mathematical. representations of atmospheric processes. 

They are based on many assumptions, however, and.can only be considered approximations of real processes. Therefore, it is, important toAnvestigate the 

extent to which model predictions disagree with actual measurements, Deviations 

occur for two basic reasons: (1) a complete1y valAd mathematical description 

of natural systems does not presently exist; and (2).input data and data for 

comparison with predictions are often unresolved and imprecise (Seinfeld and 

Wilson, 1977). It is therefore difficult to. determine,numerically the overall 

accuracy of model calculations.. Rather, attempts are made to validate model 

predictions by comparing them with real observations and operating parameters 

are often varied to determine the sensitivity of the model (Ge1inas and Vajk, 

1979). In addition, the,extent of agreement between the results.from two 

simulations can be tested. ~''In thi.s way, completely different models may be 

compared, or an internai component,.such,as the chemical kinetics mechanism, 

may be substituted and a model re-run to„aascertain the effects of such substitutions. 

It shoul d be noted that the val i di ty of al l,dynami c.model,s depends, i n 

part, on the quality of the chemical kinetics mechanisms used to define the 

03-HC-NOx relationship. These mechanisms have the advantage of being causeand-effect descriptions derived from actual experimental data. The data are 

subject, however, to the effects of smog chamber artifacts (Carter et al., 

1982), which may or may not occur in the atmosphere. Also, there remain substantial uncertainties in.the detailed chemistry of certain classes of organics 

3-121 
such as the aromatics (Section 3.3). In addition, only recently have data 

become available so that mechanisms can be tested against data bases in which 

the hydrocarbon composition has been systematically varied or in which dynamic 

dilution and injection of new reactants has occurred (Jeffries et al., 1981). 

Therefore, dynamic models using existing chemical mechanisms may not accurately 

describe all of the conditions that apply in the atmosphere (Jeffries et al., 

1981). 

Evaluations of complex dynamic models have been of two forms: numerical 

sensitivity analyses and simulation performance studies using ambient air 

data. Sensitivity analyses have considered the effects of varying meteorological factors, initial and boundary .air quality data and emissions inputs, 

model structure and computational factors, and reactions within the chemical 

kinetics mechanism (U.S. Environmental Protection Agency, 1981g; Liu and 

Sei nfel d, 1975; Gel i nas and Vaj k, 1979; Ti 1 den and Sei nfel d, 1982; Dunker, 

1980, 1981; Falls et al., 1979). 

A comparative study of the Photochemical Box Model (PBM) of Demerjian and 

Schere (1979), the Lagrangian Photochemical Model (LPM) developed by Lurmann 

et al. (1979), and the Urban Airshed Model (UAM) (Ames et al., 1978) was 

performed in which the models were compared in "off-the-shelf" use (Schere and 

Shreffler, 1982a,b). That is, no effort was made to adjust the model predictions, although great care was taken in preparation of data and in model 

execution. Based upon application of these models to St. Louis air quality 

data, a fourth model, LIRAQ, was shown to be unsatisfactory and was not evaluated further (Schere and Shreffler, 1982a,b). 

The remaining three models yielded adequate source-receptor information, 

provided knowledgeable interpretation of the output was applied. For instance, 

the PBM averaged 23 percent overprediction of ozone concentration over all 

test days, but in the 5 stagnation days for which the maximum ozone observed 

occurred within the PBM domain the average overprediction was only 8 percent. 

The LPM showed the largest variance in the ozone concentration residuals, 

possibly because the input data were not precise enough to fulfill the temporal 

and spatial demands of the model. As in previous studies (4lhitten and Hogo, 

1981; Reynolds et al., 1982; Cole et al., 1982), the UAM predicted ozone 

maximum concentrations with little bias (about 4 percent overprediction), but 

had difficulty placing the "ozone cloud" at the correct time and place. 

Again, this suggests uncertainty in specifying the wind field data. The user 

; 

3-122 
of any of these models must have a strong scientific background and must 

exercise extreme care in implementing air quality simulations (Schere and 

Shreffler, 1982a,b). 

Because of their wide use, Levels II, III, and IV of EKMA have been the 

subject of many evaluation efforts. These include parameter sensitivity 

studies, comparison between the various levels of the approach, and studies 

comparing EKMA with other dynamic models (Meyer et al., 1981). 

Jeffries and coworkers (1981) evaluated the performance of EKMA Levels II 

and III using 10 days of data from the 1976 St. Louis Regional Air Pollution 

Study. To evaluate the effects of chemistry and meteorology inputs, four 

different chemical kinetics mechanisms and three methods of calculating mixingheight profiles were employed. The choice of inechanism, trajectory, and 

mixing- height profile proved to have a large effect on the prediction of 

absolute ozone levels. No one mechanism or mixing-height profile was superior, 

however, at producing the "best fit" over all days. When the EKMA procedure 

was used in a relative sense to estimate needed reductions in NMHC, results 

obtained were not consistent. Improvements needed in the simple trajectory 

model (Level III) were identified as: (1) improved chemical kinetics descriptions, (2) smoother and more defined trajectories, (3) better treatment of 

point sources, and (4) improved mixing-height profiles. 

In another sensitivity study, standard and city-specific versions of 

EKMA were used to simulate 100 pre-selected test days (Maxwell and Martinez, 

1982). A statistical analysis was performed to determine how accurately these 

models, using three different chemical mechanisms, could predict absolute 

ozone 1 evel s. As above (Jeffri es et al ., 1981) , none of the model s was a 

consistently good predictor of ozone and ozone levels were usually overpredicted 

by more than 20 percent. 

Finally, all three levels of EKMA were compared on a limited number of 

test days with respect to: (1) level of EKMA, (2) chemical kinetics'mechanisms, 

and (3) isopleth diagram entry parameters (Hayes and Hogo, 1982). Again it 

was found that substituting chemical mechanisms produced significant differences 

in the shape of the isopleth curves. Also, and partly because of . this, EKMA 

predictions were found to be quite sensitive to low NMHC/NO xprecursor ratios. 

Precursor-ozone relationships derived for various levels of EKMA were not 

particularly different, nor did they appear to be sensitive to the choice of 

trajectory (for Level II analysis). 

3-123 
Studies comparing EKMA performance with that of more complex dynamic 

models have also been carried out (Whitten and Hogo, 1981; U.S. Environmental 

Protection Agency, 1981a). Input to the EKMA model was often generated by the 

comparison model so that specific features could be compared. In all studies,. 

the most sensitive difference, in terms of absolute prediction of ozone levels, 

was found to be the choice of chemical mechanism. It was shown, however, that 

application of the EKMA produced precursor-ozone descriptions similar to those 

from the more complex models when the same chemical mechanism was used in each 

of the models. Finally, it should be noted that the non-linear relationships 

between 03 and its precursors means that good model performance in replicating 

base •case conditions does not ensure accurate emission control calculations. 

To address this problem, emission control estimates obtained with EKMA have 

been compared with trends and emission reduction estimates obtained with grid 

models (Meyer et al., 1981). Further information on the comparative performance 

of EKMA is found in Dimitriades and Dodge (1983). 

Selection of a modeling approach for determining ozone concentrations 

that is acceptable and appropriate for given circumstances necessitates making 

many interrelated decisions. A11 models considered should be able, of course, 

to simulate the physical and chemical processes known or suspected to be 

important. Potential users must then weigh the advantages of greater credibility and capability against the disadvantages of greater cost, time, and 

personnel requirements (Association of Bay Area Governments, 1979a). In 

addition to the technical aspects of potential modeling approaches, specific 

selection constraints also include: (1) extent of data requirements; (2) costs 

of data collection, model implementation, and operation; (3) computer constraints; (4) personnel requirements; and (5) schedule constraints. These 

criteria are not independent of one another, and time spent defining a selection 

plan can result in substantial benefits throughout the model.ing exercise. 

3.7 SUMMARY 

3.7.1 Descriptions and Properties of Ozone and Other Photochemical Oxidants 

Ozone (03 ) and other photochemical oxidants occurring at low concentrations in ambient air, such as peroxyacetyl nitrate (PAN), hydrogen peroxide 

(H2O2 ), and formic acid (HCOOH), are characterized chiefly by their ability to 

remove electrons from, or to share electrons with, other molecules or ions (i.e., 

3-124 
oxidation). The capability of a chemical species for oxidizing or reducing 

other chemical species is termed "redox potential" (positive or negative standard potential) and is expressed in volts. A reactive allotrope of oxygen 

that is only about one-tenth as soluble as oxygen in water, ozone has a standard potential of +2.07 volts in aqueous systems for the redox pair, 0 3IH20 

(Weast, 1977). Hydrogen peroxide, which is highly soluble in water and other 

polar solvents, has a standard potential of +1.776 in the redox pair, H 2O 2/H 20 

(Weast, 1977). No standard potential for peroxyacetyl nitrate in neutral or 

buffered aqueous systems, such as those that occur in biological systems, 

appears in the literature. In acidic solution (pH 5 to 6), PAN hydrolyzes 

fairly rapidly (Lee et al., 1983; Holdren et al., 1984); in alkaline solution 

it decomposes with the production of nitrite ion and molecular oxygen (Stephens, 

1967; Nicksic et al., 1967). An important property of PAN, especially in the 

laboratory, is its thermal instability. Its explosiveness dictates its synthesis 

for experimental and calibration purposes by experienced personnel only. 

Formic acid is formed as a stable product in photochemical air pollution. 

It has the structure of both an acid and an aldehyde and in concentrated form 

is a pungent-smelling, highly corrosive liquid. 

The toxic effects of oxidants are attributable to their oxidizing ability. Their oxidizing properties also form the basis of several measurement 

techniques for 0 3 and PAN. The calibration of ozone and PAN measurements, 

however, is achieved via their spectra in the ultraviolet and infrared regions, 

respectively. A11 three pollutants of most concern in this document (0 3 , PAN, 

and H2O2 ) must be generated in situ for the calibration of ineasurement techniques. For ozone and H 2O2 , generation of calibration gases is reasonably 

straightforward. 

3.7.2 Nature of Precursors to Ozone and Other Photochemical Oxidants 

Photochemical oxidants are products of atmospheric reactions involving 

volatile organic compounds (VOC) and oxides of nitrogen (NO x), as well as 

hydroxyl (OH) and other radicals, oxygen, and sunlight (see, e.g., Demerjian 

et al., 1974; National Research Council, 1977a; U.S. Environmental Protection 

Agency, 1978a; Atkinson, 1985). The oxidants are largely secondary pollutants 

formed in the atmosphere from their precursors by processes that are a complex, 

nonlinear function of precursor emissions and meteorological factors. 

3-125 
The properties of organic compounds that are most relevant to their role 

as precursors to ozone and other oxidants are their volatility, which governs 

their emissions into the atmosphere; and their chemical reactivity, which 

determines their lifetime in the atmosphere. Although vapor-phase hydrocarbons (compounds of carbon and hydrogen only) are the predominant organic 

compounds in ambient air that serve as precursors to photochemical oxidants, 

other volatile organic compounds are also photochemically reactive in those 

atmospheric processes that give rise to oxidants. In particular, halogenated 

organics (e.g., haloalkenes) that participate in photochemical reactions are 

present in ambient air, although at lower concentrations than the hydrocarbons. 

They are oxidized through the same initial step involved in the oxidation of 

the hydrocarbons; that is, attack by hydroxyl radicals. Alkenes, haloalkenes, 

and aliphatic aldehydes are, as classes, among the most reactive organic 

compounds found in ambient air (e.g., Altshuller and Bufalini, 1971; Darnall 

et al., 1976; Pitts et al., 1977; U.S. Environmental Protection Agency, 1978a, 

and references therein). Alkenes and haloalkenes are unique among VOC in 

ambient air in that they are susceptible both to attack by OH radicals (OH) 

and by ozone (Niki et al., 1983). Methane, halomethanes, and certain haloethenes are of negligible reactivity i,n ambient air and have been classed as 

unreactive by the U.S. Environmental Protection Agency (1980a,b). Since methane 

is considered only negligibly reactive in ambient air, the volatile organic 

compounds of importance as oxidant precursors are usually referred to as 

nonmethane hydrocarbons (NMHC) or, more properly, as nonmethane organic 

compounds (NMOC). 

The oxides of nitrogen that are important as precursors to ozone and 

other photochemical oxidants are nitrogen dioxide (NO 2 ) and nitric oxide (NO). 

Nitrogen dioxide is itself an oxidant that produces deleterious effects, which 

are the subject of a separate criteria document (U.S. Environmental Protection 

Agency, 1982a). Nitrogen dioxide is an important precursor (1) because its 

photolysis in ambient air leads to the formation of oxygen atoms that combine 

with molecular oxygen to form ozone; and (2) because it reacts with acetylperoxy radicals to form peroxyacetyl nitrate, a phytotoxicant and a lachrymator. Although ubiquitous, nitrous oxide (N 20) is unimportant in the production 

of oxidants in ambient air because it is virtually inert in the troposphere. 

3-126 
3.7.3 ' Atmospheric Reactions of Ozone and Other Oxidants Including Their Role 

in Aerosol Formation 

The chemistry of the polluted atmosphere is exceedingly complex, but an 

understanding of the basic phenomena is not difficult to acquire. Three 

processes occur: the emission of precursors to ozone from predominantly 

manmade sources; photochemical reactions that take.place during ~ the dispersion and transport of these precursors; and scavenging processes that reduce 

the concentrations of both 0 3 and precursors along the trajectory. 

The specific atmospheric reactions of ozone and of other photochemical 

oxidants such as peroxyacetyl nitrate and hydrogen peroxide are becoming 

increasingly well-characterized. The reactions of these species result in 

products and processes that may have significant environmental and health- and 

welfare-related implications, including effects on biol'ogical systems, nonbiological materials, and such phenomena as visibility degradation and acidification of cloud and rain water. 

3.7.3.1 Formation and Transformation of Ozone and Other Photochemical Oxidants. In the troposphere, ozone is formed through the dissociation of NO 2 by 

sunl,ight to yield an oxygen atom, which then reacts with molecular oxygen (0 2 ) 

to produce an 0 3 molecule. If it is present, NO can react rapidly with 0 3 to 

form NO 2 and an 0 2 molecule. In the absence of competing reactions, a steadystate or equilibrium concentration of 0 3 is soon established between 0 3 , NO2 , 

and NO (National Research Council, 1977a). The injection of organic compounds 

into the atmosphere upsets the equilibrium and allows the ozone to accumulate 

at higher than steady - state concentrations. The length of the induction 

period before the accumulation of 0 3 begins depends heavily on the initial 

NO/NO2 and NMOC/NO x ratios (National Research Council, 1977a). 

The major role played by organic compounds in smog reactions is attributable to the hydroxyl radical (OH), since it reacts with essentially all organic 

compounds (e.g., Atkinson, 1985; Herron and Huie, 1977, 1978; Dodge and Arnts, 

1979; Niki et al., 1981). Aldehydes, which are constituents of automobile 

exhaust as weli as decomposition products of most atmospheric photochemical 

reactions involving hydrocarbons, and nitrous acid (HONO), are important 

sources of OH radicals, as is 0 3 itself. Other free radicals, such as hydroand alkylperoxy radicals and the nitrate (NO 3 ) radical play important roles in 

photochemical air pollution. 

3-127 
The presence of organic compounds, oxides of nitrogen, and sunlight does 

not mean that the photochemical reactions will continue indefinitely. Termination reactions gradually remove NO2 from the reaction ~ •mixtures, such that 

the photochemical cycles slowly come to an end unless fresh NO and NO 2 emissions are injected into the atmosphere. Compounds containing Ritrogen, such 

as PAN, nitric acid (HNO 3 ), and peroxynitric acid (HN04), as well as organic. 

and inorganic nitrates, are formed in these termination reactions. 

Recent studies on the photooxidation of organic compounds under simulated 

atmospheric conditions have been reasonably successful. The rate constants 

for the reaction of OH radicals with a large number of organic compounds have 

been measured (e.g., Atkinson et al., 1979; Atkinson et al., 1985). The 

mechanisms of the reactions of paraffinic compounds are fairly well understood, as are those of olefinic compounds, at least for the smaller compounds. 

Photooxidation reactions of the aromatic compounds, however, are poorly understood. 

In the presence of NOx , natural hydrocarbons (i.e., those organic compounds emitted from vegetation) can also undergo photooxidation reactions to 

yield 03 , although most naturally emitted hydrocarbons are olefins and are 

scavengers as well as producers of 0 3 (e.g., Lloyd et al., 1983; Atkinson 

et al., 1979; Kamens et al., 1982; Killus and Whitten, 1984; Atkinson and 

Carter, 1984). 

3.7.3.2Atmospheric Chemical Processes Involving Ozone. Ozone can react with 

organic compounds in the boundary layer of the troposphere (Atkinson and 

Carter, 1984). It is important to recognize, however, that organics undergo 

competing reactions with OH radicals in the daytime (Atkinson et al., 1979; 

Atkinson, 1985) and, in certain cases, with NO 3 radicals during the night 

(Japar and Niki, 1975; Carter et al., 1981a; Atkinson et.al ., 1984a,b,c,d,e; 

Winer et al., 1984), as well as photolysis, in the case of aldehydes and other 

oxygenated organics. Only for organics whose ozone reaction rate constants 

are greater than —10-21cm3molecule-1sec-1can consumption by ozone be 

considered to be atmospherically important (Atkinson and Carter, 1984). 

Ozone reacts rapidly with the acyclic mono-, di-, and tri-alkenes and 

with cyclic alkenes. The rate constants for these reactions range from —10

-18 

to —10-14cm3molecule-1sec-1 (Atkinson and Carter, 1984), corresponding to 

atmospheric lifetimes ranging from a few minutes for the more reactive cyclic 

alkenes, such as the monoterpenes, to several days. In polluted atmospheres, 

3-128 
a significant portion of the con.sumption of the more reactive alkenes will 

occur via reaction with ozone rather than with OH radicals, especialiy'in the 

afternoons during photochemical oxidant episodes. Reactions between ozone and 

alkenes can result in aerosol formation (National Research Council, 1977a; 

Schuetzle and Rasmus'sen, 1978), with alkenes of higher carbon numbers the 

chief contributors. 

Because of their respective rate - constants, neither alkanes (Atkinson and 

Carter, 1984) nor alkynes (Atkinson and Aschmann, 1984) are'expected to react 

with ozone in the atmosphere, since competing reactions with OH radicals have 

higher rate constants (Atkinson et al., 197YI; Atkinson, 1985). 

The aromatics react w'ith ozone, but quite slowly (Atkinson aind Carter, 

1984), such that their reactions with ozone are expected to be unimportant in 

the atmosphere. Cresols are more reactive toward ozone than the aromatic 

hydrocarbons (Atkinson and Carter, 1984), but their reactions with OH radicals 

(Atkinson, 1985) or NO 3 radicals (Carter et al., 1981a; Atkinson et al., 

1984d) predominate. 

For oxygen-containing organic compounds, especially those without carboncarbon double bonds, reactions with.ozone are slow. For carbonyls and ethers 

(other than ketene) that contain unsaturated carbon-carbon bonds, however, 

much faster reactions are observed (Atkinson and Carter, 1984). 

Certain reactions of ozone other than its reactions with organic compounds are important in the atmosphere. Ozone reacts, rapidly with NO to form 

NO2 , and subsequently with NO2 to produce 'the nitrate (NO 3 ) radical and an 

oxygen molecule. Photolysis of ozone can be a significant pathway for the 

formation of OH 'radicals, particularly in poll'uted "atmospheres when ozone 

concentrations are at their peak. 

Ozone may play a role in the oxidation of S0 2 to H2SO4, both indirectly 

in the gas phase (via formation of OH radicals and Criegee biradicals) and 

directly in aqueous droplets.- 

3.7.3.3Atmospheric Reactions of PAN, , H„0„ and HCOOH . Because PAN is in 

equilibrium with acetyl peroxy radicals - and NO2 , any process that leads to the 

removal of either of these species will lead to the decomposition of PAN. One 

such process i s the reacti on of NO wi th acetyl peroxy radi cal s. Thi s can 

lead, however, to the formation of OH radicals. Thus, PAN remaining overnight 

from an episode on the previous day can react with NO emitted from morning 

traffic to produce OH radicals (Cox and Roffey -, 1977; Carter 'et al., 1981c) 

3-129 
that will enhance smog formation on that day (e.g., Tuazon et al., 1981a). In 

the absence of significant NO concentrations, and in regions of moderate to 

lower temperatures, PAN will persist in the atmosphere (wallington et al., 

1984; Aikin et al., 1983) and contribute to the long-range transport of NO x . 

Although hydrogen peroxide formed in the gas phase from the reactions of 

hydroperoxyl radicals plays a role in HO x chemistry in the troposphere, and 

especially in the stratosphere (Crutzen and Fishman, 1977; Cox and Burrows, 

1979), its major importance arises from its high solubility in water. The 

latter ensures that a large fraction of gaseous H 2O2 will be taken up in 

aqueous droplets. Over the past decade, evidence has accumulated that H 2O 2 

dissolved in cloud, fog, and rainwater may play an important, and, in acidic 

droplets (i.e., pH <5), even a dominant role in the oxidation of S0 2 to H 2SO4 

(e.g., Hoffman and Edwards, 1975; Martin and Damschen, 1981; Chameides and 

Davis, 1982; Calvert and Stockwell, 1983, 1984; Schwartz, 1984). Hydrogen 

peroxide may also play a role in the oxidation of NO 2 dissolved in aqueous 

droplets, although relevant data are limited and additional research is required 

(see, e.g., Gertler et al., 1984). Substantial uncertainties remain concerning 

the quantitative role of H 2O 2 in acidification of aqueous particles and droplets 

(Richards et al., 1983). 

Because it can be scavenged rapidly into water droplets, formic acid can 

E 

potentially function as an oxidant in cloud water and rain water. Thus, HCOOH 

is an example of a compound that is a non-oxidant or weak oxidant in the gas 

phase but that is transformed upon incorporation in aqueous solutions into an 

effective oxidizer of S(IV). Although much uncertainty remains concerning the 

quantitative role of HCOOH and the higher organic acids, they potentially play 

a minor but still significant role in the acidification of rain. 

3.7.4 Meteorological and Climatological Processes 

Meteorological and climatological processes are important in determining 

the extent to which precursors to ozone and other photochemical oxidants can 

accumulate, and thereby the concentrations of ozone and other oxidants that 

can result. The meteorological factors most important in the formation and 

transport of ozone and other photochemical oxidants in the lower troposphere 

are: (1) degree of atmospheric stability; (2) wind speed and direction; 

(3) intensity and wavelength of sunlight; and (4) synoptic weather conditions. 

These factors are in turn dependent upon or interrelated with geographic, 

seasonal, and other climatological factors. 

3-130 
Incursions of ozone from the stratosphere are an additional source of the 

ozone found in the lower troposphere. The physical and meteorological mechanisms 

by which ozone is brought into the troposphere from the stratosphere are 

important in determining the resulting ground-level concentrations, ground-level 

locations irnpacted, and the seasonality of incursions of stratospheric ozone. 

3.7.4.1Atmospheric Mixing . The concentration of a pollutant in ambient air 

depends significantly on the degree of atmospheric mixing that occurs from the 

time the pollutant or its precursors are emitted and the arrival of the pollutant at the receptor. The rate at which atmospheric mixing proceeds and the 

extent of the final dilution depends on the amount of turbulent mixing that 

occurs and on wind speed and direction. Atmospheric stability is one of the 

chief determinants of turbulent mixing since poilutants do not spread rapidly 

within stable layers nor do they mix upward through stable layers to higher 

altitudes. 

Temperature inversions, in which the temperature increases with increasing 

altitude, represent the most stable atmospheric conditions. Surface inversions (base at ground level) and elevated inversions (the entire layer is 

above the surface) are both common (Hosler, 1961; Holzworth, 1964) and both 

can occur simultaneously at the same location. Surface inversions show a 

diurnal pattern, forming at night in the absence of solar radiation but breaking up by about mid-morning as the result of surface heating by the sun (Hosler, 

1961; Slade, 1968). Elevated inversions can persist throughout the day and 

pollutants can be trapped between the ground surface and the base of the 

inversion. The persistence of elevated inversions is a major meteorological 

factor contributing to high pollutant concentrations and photochemical smog 

conditions along the Caiifornia coast (Hosler, 1961; Holzworth, 1964; Robinson, 

1952). In coastal areas generally, such as the New Engiand coast (Hosler, 

1961) and along the Great Lakes (Lyons and Olsson, 1972), increased atmospheric 

stability (and diminished mixing) occurs in summer and fall as the result of 

the temperature differential between the water and the land mass. 

The depth of the layer in which turbulent mixing can occur (i.e., the 

"mixing height") shows geographical dependence. Summer morning mixing heights 

are usually >300 m in the United States except for the Great Basin (part of 

Oregon, Idaho, Utah, Arizona, and most of Nevada), where the mixing height is 

—200 m(Holzworth, 1972). By mid-morning, mixing heights increase markedly 

such that only a few coastal areas have mixing heights <1000 m. 

3-131 
Summer afternoon mixing heights are generally an indication of the potential for recurring photochemical oxidant problems. Photochemical smog problems 

in the United States are somewhat unexpected since the lowest afternoon mixing 

height is —600 m(Holzworth, 1972). Elevated inversions having bases <500 m 

(i . e. , 1 ow-1 evel i nversi ons) occur i n the Uni ted States , however, wi th the 

following frequencies: 90 percent on the California coast; >20 percent on the 

Atlantic coast (New Jersey to Maine); >5 percent along the Great Lakes; and 5 

to 10 percent from Louisiana to Arkansas and eastward to about Atlanta, Georgia. 

For most areas of the United States, though, the persistence through the 

afternoon of low-level stable layers is a rare event, occurring on <1 day in 

20 (Holzworth and Fisher, 1979). 

3.7.4.2 Wind Speed and Direction. For areas in which mixing heights are not 

restrictive, wind speed and, in some cases, wind direction are major determinants of pollution potential. Since strong winds dilute precursors to ozone 

and other photochemical oxidants, a location may have good ventilation despite 

the occurrence of persistent inversions (e.g., San Francisco). Conversely, 

light winds can result in high oxidant levels even if the mixing layer is 

deep. 

The frequency of weak winds, then, is important in oxidant formation. In 

industrialized, inland areas east of the Mississippi River, surface inversions 

in the morning coupled with wind speeds <2.5 m/sec (<6 mi/hr) occur with a 

frequency >50 percent (Holzworth and Fisher, 1979). These surface inversions 

break up by afternoon, however, permitting dispersion. 

The effects of wind speed and direction include the amount of dilution 

occurring in the source areas, as well as along the trajectory followed by an 

urban or source-area plume. Regions having steady prevailing,winds, such that 

a given air parcel can pass over a number of significant source areas, can 

develop significant levels of pollutants even in the absence of weather patterns 

that lead to the stagnation type of air pollution episodes. The Northeast 

states are highly susceptible to pollutant plume transport effects, although 

some notable stagnation episodes have also affected this area (e.g., Lynn 

et al., 1964). Along the Pacific Coast, especially along the coast of 

California, coastal winds and a persistent low inversion layer contribute to 

major pollutant buildups in urban source areas and downwind along the urban 

plume trajectory (Robinson, 1952; Neiburger et al., 1961). 

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3.7.4.3 Effects of Sunlight and Temperature. The effects of sunlight on 

photochemical oxidant formation, aside from the role of solar radiation in 

meteorological processes, are related to its intensity and its spectral distribution. Intensity varies diurnally, seasonally, and with latitude, but the 

effect of latitude is strong only in the winter. Experimental studies have 

verified the effects on oxidant formation of light intensity (Peterson, 1976; 

Demerjian et al., 1980) and its diurnal variations (Jeffries et al., 1975; 

1976), as well as on the overall photooxidation process (Jaffee et al., 1974; 

Winer et al., 1979). 

A correlation between high oxidant concentrations and warm, above-ndrmal 

temperatures has been demonstrated generally (Bach, 1975; Wolff and Lioy, 

1978) and-for specific locations, e.g., St. Louis (Shreffler and Evans, 1982). 

Coincident meteorology appears to be the cause of the observed correlatio.n. 

Certain synoptic weather conditions are favorable both for the occurrence of 

higher temperatures and for the formation of ozone and other oxidants, so that 

temperature is often used to forecast the potential for high oxidant concentrations (e.g., Wolff and Lioy, 1978; Shreffler and Evans, 1982). Data from 

smog chamber studies show an effect of temperature on ozone formation (e.g., 

Carter et ai., 1979b; Countess et al., 1981), but the effect is thought to 

result from the volatilization and reaction of chamber wall contaminants as 

the temperature is increased. 

3.7.4.4 Transport of Ozone and Other Oxidants and Their Precursors. The 

levels of ozone and other oxidants that wili occur at a given receptor site 

downwind of a precursor source area depend upon many interrelated factors, 

which include but are not restricted to: (1) the concentrations of respective 

precursors leaving the source area; (2) induction time; (3) turbulent mixing; 

(4) wind speed and wind direction; (5) scavenging during transport; (6) injection of new emissions from source areas in the trajectory of the air mass; 

and (7) local and synoptic weather conditions. 

Ozone and other photochemical oxidants can be transported hundreds of 

miles from the place of origin of their precursors, as documented by the 

numerous studies on transport phenomena that were described in the 1978 criteria document for ozone and other photochemical oxidants (U.S. Environmental 

Protection Agency, 1978a). In that document, transport phenomena were classified into three categories, depending upon transport distance: urban-scale, 

3-133 
mesoscale, and synoptic-scale. In urban-scale transport, maximum concentrations of 03 are produced about 20 miles or so (and about 2 to 3 hours) downwind from the major pollutant source areas. In mesoscale transport, 0 3 has 

been observed up to 200 miles downwind from the sources of its precursors. 

Synoptic-scale transport is associated with large-scale, high-pressure air 

masses that may extend over and persist for many hundreds of miles. 

Urban-scale transport has been identified as a significant, characteristic feature of the oxidant problem in the Los Angeles Basin (Tiao et al., 

1975), as well as in San Franciso, New York, Houston, Phoenix, and St. Louis 

(e.g., Altshuller, 1975; Coffey and Stasiuk, 1975; Shreffler and Evans, 1982; 

Wolff et al., 1977a). Simple advection of a photochemically reactive air 

mass, local wind patterns, and diurnal wind cycles appear to be the main 

factors involved in urban-scale transport. 

Mesoscale transport is in many respects an extension of .urban-scale 

transport and is characterized by the development of urban plumes. Bell 

documented cases in 1959 in which precursors from the Los Angeles Basin and 

the resultant oxidant plume were transported over the coastal Pacific Ocean, 

producing elevated oxidant concentrations in San Diego County the next day 

(Bell, 1960). Similar scales of transport have been reported by Cleveland 

et al. (1976a,b) for the New York-Connecticut area; by Wolff and coworkers and 

others (Wolff et al., 1977a,c; Wolff and Lioy, 1978; Clark and Clarke, 1982; 

Clarke et al., 1982; Vaughan et al., 1982) for the Washington, DC-Boston 

corridor; and by Westberg and coworkers for the Chicago-Great Lakes area 

(Sexton and Westberg, 1980; Westberg et al., 1981). These and other studies 

have demonstrated that ozone-oxidant plumes from major urban areas can extend 

downwind about 100 to 200 miles and can have widths of tens of miles (Sexton, 

1982), frequently up to half the length of the plume. 

Synoptic-scale transport is characterized by the general and widespread 

occurrence of elevated oxidants and precursors on a regional or air-mass scale 

as the result of certain favorable weather conditions, notably, slow-moving, 

well-developed high-pressure, or anti-cyclonic, systems characterized by weak 

winds and limited vertical mixing (Korshover, 1967; 1975). The size of the 

region that can be affected has been described by Wolff and coworkers, who 

reported the occurrence of haze and elevated ozone levels in an area extending 

from the Midwest to the Gulf Coast (Wolff et al., 1982) and the occurrence of 

elevated ozone concentrations extending in a virtual "ozone river" from the 

3-134 
Gulf Coast to New England that affected anywhere from a few hundred square 

miles to a thousand square miles during a 1-week period in July 1977 (Wolff 

and Lioy, 1980). 

3.7.4.5 Stratospheric-Tropospheric Ozone Exchange. The fact that ozone is 

formed in the stratosphere, mixed downward, and incorporated into the troposphere, where it forms a more or less uniformly mixed background concentration, has been known in various degrees of detail for many years (Junpe, 

1963). It is widely accepted that the long-term average tropospheric background concentration of about 30 ppb to 50 ppb results primarily, though not 

exclusively, from the transfer of stratospheric ozone into the upper troposphere, followed by subsequent dispersion throughout the troposphere (e.g., 

Kelly et al., 1982). 

The exchange of ozone between the stratosphere and the troposphere in the 

middle latitudes occurs to a major extent in events called "tropopause folds" 

(TF) (Reiter, 1963; Reiter and Mahlman, 1965; Danielsen, 1968; Reiter, 1975; 

Danielsen and Mohnen, 1977; Danielsen, 1980), in which the polar jet stream 

plays a major role. From recent studies, Johnson and Viezee (1981) proposed 

four types or mechanisms of TF injection and concluded that two of these,, both 

of which are consistent with-theory, could cause substantial effects in terms 

of high ozone concentrations at ground level. They concluded, in addition, 

that all low-pressure trough systems may induce TF events and cause the trahstropopause movement of ozone-rich air into the troposphere (Johnson and Viezee, 

1981). 

3.7.4.6 Stratospheric Ozone at Ground Level. From a detailed review, of 

studies on background tropospheric ozone, Viezee and Singh (1982) concluded 

that the stratosphere is a major but not the sole source of background ozone 

in the unpolluted troposphere, a conclusion reached by other investigators as 

well (e.g., Kelly et al., 1982). The stratospheric ozone reservoir shows a 

strong seasonal cycle that is reflected at ground-level. At some stations 

that monitor background ozone levels, average spring background levels may be 

as high as 80 ppb, with average fa11 levels ranging from 20 to 40 ppb (e.g. ~ 

Singh et a1., 1977; Mohnen, 1977; U.S. Environmental Protection Agency, 1978a). 

Viezee and Singh (1982) and Viezee et al. (1983) concluded that relatively 

high ozone concentrations can occur for short periods of time (minutes tp a 

few hours) over local areas as a resuit of stratospheric intrusions. 

3-135 
A number of investigators have attempted to quantify the amount of the 

surface ozone that can be attributed to stratospheric sources. The method 

most commonly used is based on the assumption that beryllium-7 ( 7 Be) is a 

unique tracer for air parcels of stratospheric origin. Calculated correlations 

between surface ozone and 7Be show, however, that their relationship is highly 

variable (e.g., Kelly et al., 1982; Ferman and Monson, 1978; Johnson and 

Viezee, 1981; Husain et al., 1977). Singh et al. (1980) and Viezee and Singh 

(1982) have pointed out problems with using this technique to quantify the 

contribution of stratospheric ozone to surface ozone. Singh et al. (1980) 

concluded that "the experimental techni,que involving a 7Be/03 ratio to estimate the daily stratospheric component of ground-level 0 3 is unverified and 

considered to be inadequate for air quality applications" (p. 1009). This 

group of investigators have suggested, however, that 7Be may be used, under 

appropriate meteorological conditions, as a qualitative tracer for air masses 

of stratospheric origin (Johnson and Viezee, 1981; Viezee et al., 1979). 

Other methods used to attempt to quantify the stratospheric component of 

surface ozone include aircraft observations o.f TF events coupled with calculations of downward ozone flux, and examination of surface ozone data records. 

From such data, Viezee et al. (1983) concluded that direct ground-level contributions from stratospheric ozone are infrequent (<1 percent of the time), 

short-lived, and associated with ozone concentrations <0.1 ppm. 

Notwithstanding difficulties with quantifying its contribution to surface 

ozone, however, stratospheric ozone is clearly present in atmospheric surface 

layers, and the meteorological mechanisms responsible have been described by a 

number of investigators (e.g., Danielsen, 1968; Wolff et al., 1979; Johnson 

and Viezee, 1981). 

3.7.4.7 Background Ozone from Photochemical Reactions. Whereas stratospheric 

ozone is thought by many investigators to be the dominant contributor to 

background levels of ozone, as discussed above, other inves -tigators have 

concluded that as much as two-thirds of the annual average background concentrations may result from photochemical reactions. For example, Altshuller 

(1986), in a recent review article, has concluded that photochemically generated 

ozone should equal or exceed the stratospheric contribution at lower-elevation 

remote locations; and that photochemically generated ozone from manmade emissions 

probably constitutes most of the ozone measured at more polluted rural locations 

during the warmest months of the year. His conclusions were based, in part, 

3-136 
on an analysis of global circulation (e.g., Levy et al., 1985) and photochemical 

modeling .studies (e.g., Fishman and Seiler, 1983; Fishman and Carney, 1984; 

Fishman et al., 1985; Dignon and Hameed, 1985). In these modeling studies, 

the photochemical contribution to background ozone levels was estimated to 

range from —15 ppb (long-term) to —80 ppb (summertime), depending on the level 

of NOxemissions assumed. 

Studies on the role of NOxin nonurban ozone photochemistry have shown 

that ozone formation at many of the locations is not NOX limited, but depends 

on VOC reactions, as well (e.g., Martinez and Singh, 1979; Kelly et al., 1984; 

Liu et al., 1984). Background NO xconcentrations at most remote, clean l.ocations range from <0.05 ppb upward. Mean concentrations of NO xat nonurban 

locations in the United States east of the Rocky Mountains range from —1 ppb 

to 10 ppb (Altshuller, 1986; see also Sections 3.5 and 3.7.5). These background 

concentrations of NO xare higher than previously thought (see, e.g., Singh et 

al:, 1980; Kelly et al., 1984, regarding global modeis and assumed reservoirs 

of NOx ). 

The contributions of biogenic VOC to background ozone, although a matter 

of controversy in recent years, appear not to be significant under most atmospheric conditions, since ambient air concentrations of biogenic VOC are quite 

low, even at rural sites (Altshuller, 1983). 

Thus, photochemistry and stratospheric intrusions are both regarded as 

contributing to background ozone concentrations, but the apportionment of 

background to respective sources remains a matter of investigation. 

3.7.5 Sources, Emissions, and Concentrations of Precursors to Ozone and Other 

Photochemical Oxidants 

As noted earlier, photochemical production of ozone depends both on the 

presence of precursors, volatile organic compounds (VOCs) and nitrogen oxides 

(NOx ), emitted by manmade and by natural sources; and on suitable conditions 

of sunlight, temperature, and other meteorological factors. Because of the 

intervening requirement for meteorological conditions conducive to the photochemical generation of ozone, emission inventories are not as direct predictors of ambient concentrations of secondary pollutants such as ozone and other 

oxidants as they are for primary pollutants. 

3.7.5.1 Sources and Emissions of Precursors. Emissions of manmade VOCs 

(excluding several relatively unreactive compounds such as methane) in the 

3-137 
United States have been estimated at 19.9 Tg/yr for 1983 (U:S. Environmental 

Protection Agency, 1984a). Retrospective estimates show that manmade VOC 

emissions rose from about 18.5 Tg/yr in 1940 to about 27.1 Tg/yr in 1970 (U.S. 

Environmental Protection Agency, 1986). An examination of trends in manmade 

VOC emissions for 1970 through 1983 shows that the annual emission rate for 

manmade VOCs decreased some 26 percent during this period. The main sources 

nationwide are industrial processes, which emit a wide variety of VOCs, such 

as chemical solvents; and transportation, which includes the emission of VOCs 

in gasoline vapor as well as in gasoline combustion products. Estimates of 

biogenic emissions of organic compounds in the United States are highly 

inferential but data suggest that the yearly rate is the same order of magnitude as manmade emissions. Most of the biogenic emissions actually occur 

during the growing season, however, and the kinds of compounds emitted are 

different from those arising from manmade sources.' 

Emissions of manmade NO xin the United States were estimated at 19.4 Tg/yr 

for 1983. Retrospective estimates show that manmade NO xemissions rose from 

about 6.8 Tg/yr in 1940 to about 18.1 Tg/yr in 1970 (U.S. Environmental Protection Agency, 1986). Annual emissions of manmade NOxwere some 12 percent 

higher in 1983 than in 1970, but the rate leveled off in the late 1970s and 

exhibited a small decline from about 1980 through 1982 (U.S. Environmental 

Protection Agency, 1984a). The increase over the period 1970 through 1983 had 

two main causes: (1) increased fuel combustion in stationary sources such as 

power plants; and (2) increased fuel combustion in highway motor vehicles, as 

the result of the increase in vehicle miles driven. Total vehicle miles 

driven increased by 42 percent over the 14 years in quest7on. 

Estimated biogenic NOx emissions are based on uncertain extrapolations 

from very limited studies, but appear to be about an order of magnitude less 

than manmade NO xemissions. 

3.7.5.2 Representative Concentrations in Ambient Air. 

3.7.5.2.1 Hydrocarbons in urban areas. Most of the available ambient air 

data on the concentrations of nonmethane hydrocarbons (NMHC) in urban areas 

have been obtained during the 6:00 to 9:00 a.m. period. Since hydrocarbon 

emissions are at their peak during that period of the day, and since atmospheric 

dispersion is limited that early in the morning, NMHC concentrations measured 

then generally reflect maximum diurnal levels. Representative data for urban 

areas show mean NMHC concentrations between 0.4 and 0.9 ppm.' 

3-138 
The hydrocarbon composition of urban atmospheres is dominated by species 

in the C 2to C10molecular-weight range. The paraffinic hydrocarbons (alkanes) 

are most prominent, followed by aromatics and alkenes. Based on speciation 

data obtained in a number of urban areas, alkanes generally constitute 50 to 

60 percent of the hydrocarbon burden in ambient air, aromatics 20 to 30 percent, 

with alkenes and acetylene making up the remaining 5 to 15 percent (Sexton and 

Westberg, 1984). 

3.7.5.2.2Hydrocarbons in nonurban areas . Rural nonmethane hydrocarbon 

concentrations are usually one to two orders of magnitude lower than those 

measured in urban areas (Ferman, 1981; Sexton and Westberg, 1984). In samples 

from sites carefully selected to guarantee their rural character, total NMHC 

concentrations ranged from 0.006 to 0.150 ppm C(e.g., Cronn, 1982; Seila, 

1981; Holdren et al., 1979). Concentrations of individual species seldom 

exceeded 0.010 ppm C. The bulk of species present in rural areas are alkanes; 

ethane, propane, n-butane, iso-pentane, and n-pentane are most abundant. 

Ethylene and propene are sometimes present at <0.001 ppm C, and toluene is 

usually present at —0.001 ppm C. Monoterpene concentrations are usually 

<0.020 ppm C. Quring the summer months, isoprene concentrations as high as 

0.150 ppm C have been measured (Ferman, 1981). The maximum concentrations of 

isoprene usually encountered, however, are in the range of 0.030 to 0.040 

ppm C. 

3.7.5.2.3Nitrogen oxides in urban areas . Concentrations of NO x , like hy,drocarbon concentrations, tend to peak in urban areas during the early morning, 

when atmospheric dispersion is limited and automobile traffic is dense. Most 

NOxis emitted as nitric oxide (NO), but the NO is rapidly converted to NO 2 , 

initially by thermal oxidation and subsequently by ozone and peroxy radicals 

produced in atmospheric photochemical reactions. The relative concentrations 

of NO versus NO 2 fluctuate day-to-day, depending on diurnal and day-to-day 

fluctuations in ozone levels and photochemical activity. 

Urban NOx concentrations' during the 6:00 to 9:00 a.m. period in 10 cities 

ranged from 0.05 to 0.15 ppm in studies done in the last 5 to 7 years (e.g., 

Westberg and Lamb, 1983; Richter, 1983; Eaton et al., 1979), although concentrations two to three times higher occur in cities such as Los Angeles. . 

Concurrent NMHC measurements for these 10 cities showed that NMHCINOxratios 

ranged from 5 to 16. 
3.7.5.2.4 Nitrogen oxides in nonurban areas.. Concentrations of NO 

x 

 in clean 

remote environments are_ usually <0.5 ppb (Logan, 1983). In exceptionally 

clean air, NOx concentrations as low as 0.015 ppb have been recorded (Bollinger 

et al., 1982). Concentrations of NOxat nonurban sites•.an the northeastern 

United States appear to be higher than NO X concentrations in the west by a 

factor of ten (Mueller and Hidy, 1983). From the limited amount of data 

available, NOx concentrations in unpopulated nonurban areas in the west average 

<1 ppb; but in nonurban northeastern areas average NOX can exceed 10 ppb. 

3.7.6 Source-Receptor (Oxidant-Precursor) Model.s. 

In order to apply knowledge of the atmospheric chemistry of precursors, 

and of ozone and other photochemical oxidants, during their dispersion and 

transport, models describing these phenomena have been developed in a variety 

of forms over the past 15 •years. Most of these model s rel ate the rates of 

precursor emissions from mobile andt stationary sources, or precursor atmospheric concentrations, to the resulting ambient concentrations, of secondary 

pollutants that impact receptors at•downwind sites.. For this reason they have 

been described as source-receptor models. • 

Current air quality, or source-receptor, models can be classified as 

either statistical or computational-dynamic. Statistical models are 

generally based on a statistical analysis of historical air quality data, 

and a're not explicitly concerned with atmospheric chemistry or meteorology. 

An example of empirical models is the linear rollback concept. 

Computational, or dynamic, models attempt to describe mathematically the 

atmospheric chemical and physical processes influencing air pollution formation and impacts. Examples of computational models include trajectory and 

fixed-grid airshed models. Two phenomenologically different approaches have 

been employed in dynamic models with respect to the coordinate systems chosen. 

A coordinate system fixed with respect to the earth is termed Eulerian, while 

in Lagrangian models the reference frame moves with the air parcel whose 

behavior is being simulated. 

3.7.6.1 Trajectory Models. In trajectory models, a moving-coordinate system 

describes pollutant transport as influenced by local meteorological conditions. 

Trajectory models provide dynamic descriptions of atmospheric source-receptor 

relationships that are simpler,and less expensive to derive than those obtained 

from fixed-cell models. 

3-140 
The simplest form`of trajectory model is the empirical kinetic modeling 

approach (EKMA). This approach ,was developed from earlier efforts 

(Dimitriades, 1972) to use smog chamber data to develop graphical relationships 

between° morning NMOC and NO xlevels and afternoon maximum concentrations of 

ozone. In applying EKMA, the Ozone Isopleth Plotting Package (OZIPP) (Whitten 

and Hogo, 1978) is used to generate ozone isopleths at various levels of 

sophistication corresponding to "standard" EKMA, "city-specific" EKMA, or the 

simplified trajectory model (F.R., 1979). The EKMA isopleths generated are used 

to determine the relative degree of control of precursor emissions needed to 

achieve a given percentage reduction in ozone. 

The use of EKMA in ozone abatement programs is relatively widespread. It 

is therefore worth noting the general control implications of EKMA isopleths. 

For areas with high levels of morning precursor emissions and meteorologyconducive to oxidant formation, such as Los Angeles, for example, EKMA isopleths 

predict that (1) at high NMOC/NO xconcentration ratios, reductions in NO xwill 

decrease ozone formation; (2) at moderate NMOC/NO xratios, reductions in NMOC 

and NOxwill decrease ozone formation; and (3) at very low NMOC/NO xratios, 

increases in NO xwill inhibit ozone formation. These predictions cannot be 

assumed to apply to all urban areas, or even to all high-oxidant urban areas, 

since the shape of the EKMA isopleths is a function of numerous factors, many 

of which are location-specific. For discussions of the specific assumptions 

employed in EKMA and the underlying chemistry and meteorology, the primary 

literature should be consulted (e.g., Dimitriades, 1970, 1972, 1977a,b; Dodge, 

1977a,b; Whitten and Hogo, 1977; U.S. Environmental Protection Agency, 1977, 

1978a; Whitten, 1983). Likewise, the primary literature should be consulted 

for additional data and discussions on the respective effects on ozone formation of controlling NMHC and NO x(e.g., Liu and Grisinger, 1981; Chock et al., 

1981; Kelly, 1985; Kelly et al., 1986; Glasson and Tuesday, 1970; Dimitriades, 

1970, 1972, 1977a,b). 

3.7.6.2Fixed-Grid Models . Fixed-grid models, also called regional airshed 

models, are based on two- or three-dimensional arrays of grid cells and are 

the most sophisticated source-receptor models presently available. Such 

models are computationally complex and require the most extensive set of input 

data; but they also provide the most realistic treatment of the various processes 

involved in photochemical air pollution formation. 

3-141 
3.7.6.3 Box Models. Box models (Hanna, 1973;-Demerjian and Schere, 1979; 

Derwent and Hov, 1980) are the simplest of dyriamic models. They treat the 

atmosphere as a single cell, bounded by the mixing layer, having an area on 

the order of 100 square miles. 

3.7.6.4 Validation and Sensitivity Anal ses for Dynamic Models. A11 presently available source-receptor models require a degree of simplifying assumptions to deal with practical limitations imposed by existing computer capabilities, time and cost constraints, or lack of knowledge concerning inputs such 

as boundary conditions, emissions, or detailed reaction mechanisms. The 

reliability and applicability of any particular model therefore depends upon 

its specific limitations, data requirements, and degree of validation against 

experimental data from ambient air measurements or environmental chamber runs. 

Reliability and applicability also depend on the quality of the chemical 

kinetics mechanisms used to define the 0 3-HC-NOx relationship. 

Attempts are made to validate model predictions by comparing them with 

real observations; and operating parameters are often varied to determine the 

sensitivity of the model to respective parameter changes (Gelinas and Vajk, 

1979). In addition, the extent of agreement between the results from two, 

simulations can be tested. In this way, completely different models may be 

compared, or an internal component, such as the chemical kinetics mechanisms, 

may be substituted and the model run again to ascertain the effect of such 

substitutions. 

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MISTT. Atmos. Environ. 12: 537-548. 

3-178 
Wine, P. H.; Kreutter, N. M.; Gump, C. A.; Ravishankara, A. R. (1981) Kinetics 

of OH reactions with the atmospheric suifur compounds H 2 S, CH3SH, CH 3 SCH3 

and CH3 SSCH3 . J. Phys. Chem. 85: 2660-2665. 

Winer, A. M. (1983) Investigation of the role of natural hydrocarbons in photochemical smog formation in California: final report. Contract no. AO-056-32. 

Sacramento; CA: California Air Resources Board., 

Winer, A. M.; Lloyd, A. C.; Darnall, K. R.; Pitts, J. N., Jr. (1976) Relative 

rate constants for the reaction of the hydroxyl radicai with selected 

ketone, chloroethenes and monoterpene hydrocarbons. J. Phys. Chem. 80: 

1635-1639. 

Winer, A. M.; Breuer, G. M.; Carter, W.- P. L.; Darnall, K. R.; Pitts, J. N., 

Jr. (1979) Effects of ultraviolet spectral distribution on the photochemistry of simulated polluted atmospheres. Atmos. Environ. 13: 989-998. 

Winer, A. M.; Atkinson, R.; Pitts, J. N., Jr. (1984) Gaseous nitrate radical: 

possible nighttime atmospheric sink for biogenic organic compounds. 

Science (Washington, DC) 224: 156-159. 

Wolff, G. T.; Lioy, P. J. (1978) An empirical model for forecasting - maximum 

daily ozone levels in the northeastern U.S. J. Air Pollut. Control Assoc. 

28: 1034-1038. 

Wolff, G. T.; Lioy, P. J. (1980) Development of an ozone river associated with 

synoptic-scale episodes in the eastern United States. Environ. Sci. Technol. 

14: 1257-1260. 

Wolff, G. T.; Lioy, P. J.; Meyers, R. E.; Cederwall, R. T.; Wight, G. D.; Pasceri, 

R: E.; Taylor, R. S. (1977a) Anatomy of two ozone transport episodes in 

the Washington, D.C., to Boston, Mass., corridor. Environ. Sci.'Technol. 

11: 506-510. 

Wolff, G. T.; Lioy, P. J.; Wight, G. D.; Meyers, R. E.; Cederwall, R. T. (1977b) 

An investigation of long-range transport of ozone across the midwestern 

and eastern United States. Atmos. Environ. 11: 797-802. 

Wolff, G. T.; Lioy, P. J.; Wight, G. D.; Pasceri, R. E. (1977c) Aerial investigation of the ozone plume phenomenon. J. Air Pollut. Control Assoc. 

27: 460-463. 

Wolff, G. T.; Lioy, P. J.; Wight, G. D.; Pasceri, R. E. (1977d) Aerial investigation of photochemical oxidants over the northeast. In: Bufalini, J. J.; 

Lonneman, W. A., eds. Proceedings of symposium on 1975 northeast oxidant 

transport study. Research Triangle Park, NC: U.S. Environmental Protection 

Agency. Report no: EPA-600/3-77-017. 

Wolff, G. T.; Ferman, M. A.; Monson, P. R. (1979) The distribution of 

beryllium-7 within high-pressure systems in the eastern United States. 

Geophys. Res. Lett. 6: 637-639. (See also "Correction," Geophys. Res. 

Lett. 6: 816, 1979.) 

3-179 
Wolff, G. T.; Lioy, P. J.; Wight, G. D. (1980) Transport of ozone associated 

with an air mass. J. Environ. Sci. Health. A15: 183-199. 

Wolff, G. T.; Kelly, N. A.; Ferman, M. A. (1981) On the sources of summertime 

haze in the eastern United Sta'tes. Science (Washington, DC) 211:703-706. 

Wolff, G. T.; Kelly, N. A.; Ferman, M. A. (1982) Source regions of summertime 

ozone and haze episodes in the eastern United States. Water, Air, Soil 

Pollut. 18: 65-81. 

Zika, R.; Saltzman, E.; Chameides, W. L.; Davis, D. D. (1982) H 2 O2 levels in 

rainwater collected in South Florida and the Bahama Islands. JGR J. Geophys. 

Res: 87: 5015-5017. 

Zimmerman, P. R. (1977) Procedures for conducting hydrocarbon emission inventories of biogenic sources and some results of recent investigations. 

In: Emission inventory/factor workshop, v. 2. Research Triangle Park, 

NC: U.S. Environmental Protection Agency, Office of Air Quality Planning 

and Standards; EPA report. Available from: NTIS,, Springfield, VA; 

PB-289123. 

Zimmerman, P. R. (1978) Testing of Hydrocarbon Emissions from Vegetation and 

Development of a Methodology for Estimating Emission Rates from Foliage. 

Final Report, Draft. Prepared by Washington State University for U.S. 

Environmental Protection Agency under°Contract No. DU-77-0063. U.S. 

Environmental Protection Agency, Research Triangle Park, NC. 

Zimmerman, P. R. (1979) Determination of emission rates of hydrocarbons from 

indigenous species of vegetation in the Tampa/St. Petersburg, Florida, 

area: final report; appendix C. Research Triangle Park, NC: U.S. Environmental Protection Agency; EPA report no. EPA-904/9-77-028. Available from: 

NTIS, Springfield, VA; PB-297057. 

Zimmerman, P. R. (1980) Natural sources of ozone in Houston: natural organics. 

In: Proceedings of specialty conference on ozone/oxidants--interactions 

with the total environment; October 1979; Houston, TX. Pittsburgh, PA: 

Air Pollution Control Association; pp. 299-310. 

3-180 
4. SAMPLING AND MEASUREMENT OF OZONE AND OTHER 

PHOTOCHEMICAL OXIDANTS AND THEIR PRECURSORS 

4.1 INTRODUCTION 

Detailed information is presented in this chapter on methods for sampling 

and measuring ozone, "total oxidants," hydrogen peroxide, and peroxyacetyl 

nitrate and its higher homologues. Because of their utility in atmospheric 

research and in the application of oxidant abatement programs, methods for 

sampling and measuring the organic and inorganic precursors to oxidants are 

described as well. The information presented here should prove helpful to 

state and local air pollution agencies and to researchers investigating health 

and welfare effects. The chief reason for presenting such information, however, 

is to provide relevant information: (1) for assessing the accuracy of aerometric 

data on these pollutants; and (2) for determining the impact of respective 

measurement and calibration methods on existing data on the health and welfare 

effects of ozone, total oxidants, and individual other oxidants. Pr.imary 

emphasis is placed in this chapter on techniques considered satisfactory for 

routine monitoring, on the effects of changes in calibration procedures for 

ozone measurements, on the relationship between ozone and "total oxidant" 

measurements, and on the accuracy and reliability of inethods for measuring 

oxidants not routinely monitored in ambient air. 

Since the publication of the 1978 criteria document on ozone and.other 

photochemical oxidants (U.S. Environmental Protection'Agency, 1978a), a new 

procedure for calibrating ozone measurements has been promulgated by EPA as 

the Federal Reference Method for calibration. In addition, EPA has continued 

efforts to institute and codify a formal nationwide program of quality assurance 

for the routi ne moni tori ng of pol 1 utants i n ambi ent ai r. Some exampl es of 

these procedures are documented in this chapter as they apply to actual operation of the analytical instrumentation. Detailed descriptions of analytical 

procedures, quality assurance procedures, and reporting requirements are 

contained in Quality Assurance Handbook for Air Pollution Measurement Systems 

(U.S. Environmental Protection Agency, 1977b). Pertinent rules and regulations 

are contained in the Federal Register (1979a,b,c,d,e). 

4-1 
Brief summaries are provided below of requirements pertaining to quality 

assurance and sampling for ozone monitoring. 

4.2 QUALITY ASSURANCE AND OTHER SAMPLING FACTORS IN MONITORING FOR OZONE 

4.2.1 Quality Assurance in Ambient Air Monitoring for Ozone 

Quality assurance as defined by EPA rules and regulations consists of two 

distinct functions. One is the assessment of the quality of monitoring data 

by estimating their precision and accuracy. The other is the control and 

possible improvement of the quality of the ambient air data by implementation 

of quality control po1icies, procedures, and corrective actions. 

Each quality control program, developed by the individual States and 

approved by the EPA Regional Administrator, must include operational procedures 

for each of the following activities: 

1. Selection of inethods, analyzers, or samplers (prescribed reference and equivalent methods for ambient air monitoring are 

described elsewhere in this chapter); 

2. Installation of equipment; 

3. Calibration (test concentrations for ozone must be obtained by 

means of the ultraviolet (UV) photometric calibration procedure 

described elsewhere in this chapter or by means of a certified 

ozone transfer standard); 

4. Zero/span checks and adjustments of automated analyzers; 

5. Control checks and their frequency; 

6. Control limits for zero, span, and other control checks, and 

respective corrective actions when such limits are surpassed; 

7. Calibration and zero/span checks for multiple range analyzers; 

8. Preventive and remedial maintenance; 

9. Quality control procedures for air pollution episode monitoring; 

10. Recording and validation of data; 

11. Documentation of quality control information. 

A one-point precision check must be carried out at least every 2 weeks on 

each automated analyzer used for ozone, using a precision test gas of known 

concentration. Each calendar quarter, at least 25 percent of the analyzers 

4-2 
used by the State and Local Air Monitoring Stations (SLAMS) for monitoring 

ozone must be formally audited by an independent operator by challenging with 

at least one audit gas of known concentration in each of the four concentration 

ranges. Similar requirements are set forth for monitoring networks designed 

to assess Prevention of Significant Deterioration (PSD) requirements. 

In addition to requirements and recommendations associated with the 

selection, instal.lation, and maintenance of monitoring equipment, the above-cited 

Federai Register publications discuss certain design criteria for monitoring 

networks (SLAMS and the National Aerometric Monitoring Stations, NAMS). Included 

are requirements on siting of monitors in order to obtain ozone.concentrations 

that are representative of regions of varying dimensions. For example, a 

"middle scale" monitor would represent conditions close to sources of NO xsuch 

that local ozone scavenging effects might be of significance. A"neighborhood 

scale" monitor, on the other hand, would be located somewhere in a reasonably 

homogeneous urban subregion having dimensions of a few kilometers. Other 

"scales" applicable to siting of ozone monitors include urban scale, which 

would be used to estimate concentrations characteristic of an area having 

dimensions between several and 50 kilometers or to measure high concentrations 

downwind of an area with high precursor emissions; and regional scale, used to 

typify concentrations over portions of a major metropolitan complex up to 

dimensions of hundreds of kilometers. For ozone SLAMS stations, applicable 

scales are middle, neighborhood, urban, and regional. Requirements for NAMS 

stations for ozone are neighborhood and urban scale. Two ozone NAMS stations 

are expected to be sufficient for each urban area: one for specific transport 

conditions leading to high ozone; and the other for monitoring peak concentrations relative to population exposure. 

4.2.2 Sampling Factors in Ambient Air Monitoring for Ozone 

Sampling factors may have a crucial effect on the quality and utility of 

measurements both in ambient air and in controlled laboratory situations. 

Sampling techniques and strategies must preserve the integrity of a representative fraction of ambient air and must be consistent with the specific purpose 

of the measurement. In this section, the significance of some sampling factors 

will be discussed briefly. For more detailed discussions of this subject, the 

reader is referred to Ott (1977) and to reports prepared for EPA by the National 

Research Council (U.S. Environmental Protection Agency, 1977a; National Research 

Council, 1977). 

4-3 
4.2.2.1 Sampling Strategies and Air Monitoring Needs. Air monitoring data 

relevant to assessing ambient 0 3 or oxidant levels are collected for a variety 

of specific needs, including: 

1. Data to be used in trend analysis as indicators of the state of 

attainment of ambient air quality standards. 

2. Data to be used i n devel opment of 0 3 control strategi es and 

evaluation of their effectiveness. 

3. Data to be used in the development and validation of air quality 

simulation models capable of application to the 03 problem. 

4. Data to be used in investigation of causes of the ozone problem 

both in general and in specific localities. 

5. Data to be used in special research studies such.as the effects 

of ambient air pollution on human health and welfare. 

Each specific purpose or need requires special considerations with regard 

to air sampling strategy. For example; 5 or more years of 0 3 data might be 

required for the adequate assessment of trends that resulted from the application of a particular control strategy rather than trends that resulted from 

chance local meteorological conditions. In contrast, the validation of an air 

quality simulation model might require only a few carefully chosen days of 

very detailed measurements of 03 , hydrocarbons, and NOx , as well as detailed 

meteorological data and time-varying emissions along the trajectory of the air 

parcel in question. 

4.2.2.2 Air Monitoring Site Selection. Ozone in the lower troposphere is a 

product of photochemical reactions that involve sunlight, hydrocarbons, and 

oxides of nitrogen. In typical urban atmospheres, ozone precursors react to 

produce ozone at such a rate that the 0 3 reaches its daily peak level in the 

middle of the day at locations downwind from the source- i ntensi ve center-city 

area. Thus, if peak 0 3 concentrations are to be measured, monitoring stations 

should, in general, be located downwind from city centers. This downwind 

distance may be on the order of 15 to 30 kilometers (9 to 19 miles), depending 

on predominant wind patterns in the area (U.S. Environmental Protection Agency, 

1977a). This distance may be highly area-specific, however. For example, 

ozone maxima in the Los Angeles plume.have been observed as far downwind as 50 

to 70 km. 

4-4 
Once a station is located, additional sampling .considerations,arise 

because of the chemical reactivity and instability of the 03 molecule. Ozone 

reacts extremely rapidly with NO and with some hydrocarbon compounds, including 

most of those emitted by vegetation. A1so,-0 3 decomposes readily on contact 

with the surface of many materials. Consideration of these effects led to the 

development of specific criteria for locating an 03 monitoring station (U.S. 

Environmental Protection Agency, 1977a; National Research Council, 1977). 

Briefly, the inlet of the sampling probe of the ozone analyzer shouid be 

positioned 3 to 15 meters (10 to 49 feet) above ground, at least 4 meters 

(13 feet) from large trees., and 120 meters (349 feet) from heavy automobile 

traffic.. Sampling probes, should be designed so as to minimize 0 3 destruction 

by surface reaction or by reaction with N0. 

Another consideration that has significance for the selection of sites 

for air monitoring stations is the fact that ambient ,monitoring data, as 

routihely obtained, have limited validity as absolute measures of air,quality. 

This limitation arises from the fact that, at ground level, the ambient atmosphere is inhomogeneous as a result of a continuous influx of fresh emissions, 

incomplete mixing, and ~ destruction of 0 3 by fresh and unreacted.emissions, and 

destruction on surfaces. In view of such inhomogeneity, monitoring data from 

a fixed network provide measures of air quality at a discrete number of elocations but may not detect temporal and spatial variations in ozone concentrations of a localized nature. fihis probiem can be allevi.ated by use of a 

greater density.of monitoring stations or by , use of a.validated air quality 

model. Such models are capable of helping quantify the emission, dispersion, 

and chemical reaction processes. Their outputs can provide data on,the distribution of air quality concentrations between widely.space,d ambient monitors. 

The emphasis An this section has been on a brief discussion of sampling 

strategies. The word sampling is also widely considered to mean those techniques that are required to obtain a parcel of air that is.representat'ive of 

the polluted atmosphere, and to maintain its integrity until a measurement of 

concentration has been carried.out. . Considerations relating to this;meaning 

of sampling are discussed as appropriate in the following sections on measurement techniques. 

4-5 
4.2.3 Measurement Methods for Total Oxidants and Ozone 

4.2.3.1 Total Oxidants. Although ozone was first unambiguously identified in 

polluted atmospheres by infrared spectroscopy (Stephens et al., 1956b; Scott 

et al., 1957), the earliest procedures for routinely monitoring 0 3 and other 

oxidizing species in the atmosphere were based on iodometry. Iodometric 

techniques are inherently non-specific in that a variety of oxidizing species 

in addition to 03 may be positive interferences, whereas reducing agents are 

negative interferences. Thus, the name "total oxidants" was coined because 

the technique responded to 0 3 and other oxidants such as peroxides, peroxyacetyl 

nitrate (PAN), and nitrogen dioxide (NO2 ). Total oxidants are then actually 

defined by the particular iodometric procedure used, since the response to the 

various oxidizing species present will depend on the details of the procedure. 

This will be more evident when interferences are discussed below. The use of 

the word "total" is in itself a misnomer. The measurement does not reflect a 

sum of the oxidizing species present because the various oxidants present in 

the atmosphere react to produce iodine at different stoichiometries and different rates. In spite of these difficulties, the measurement of total oxidants 

was a useful method for characterization of the atmosphere because of its 

correlation with the principal oxidant, 0 3 ; and, consequently, there is a 

large oxidant data base available. For these reasons, the two principal 

methods used for monitoring total oxidants are discussed briefly below. 

The bulk of the total oxidants data base was obtained by the use of two 

types of continuous monitoring instruments. In both types, an air sample is 

continuously scrubbed by an aqueous reagent containing potassium iodide (KI). 

Ozone and other oxidants produce iodine (tri-iodide ion) according to the 

reaction: 

03 + 3I + H20 -> I 3 + 02 + 20H (4-1) 

In colorimetric oxidant instruments, the iodine is measured photometrically by 

ultraviolet absorption (Littman and Benoliel, 1953). In the other common type 

analyzer the iodine produced is measured by electrochemical means (Brewer and 

Milford, 1960; Mast and Saunders, 1962). Many other chemical techniques for 

oxidants have been proposed and in some cases applied, but for these reference 

is made to the original literature (Hodgeson, 1972; Katz, 1976). 

4-6 
The interferences for both colorimetric and amperometric 0 3 analyzers are 

other oxidizing and reducing species in the atmosphere. The major oxidant in 

ambient air by far is 03 (Chapter 5); and the other oxidants present, except 

NO2 , are considered part of the total oxidants measured rather than interferences. The only significant reducing interference known is S0 2 . 

The magnitude of the NO 2 interference is variable (Tokiwa et al., 1972; 

Intersociety Committee, 1970). For the Brewer amperometric cell, the interference from NO2 is only 6 percent of an equivalent concentration of 0 3 (Tokiwa 

et al., 1972). For colorimetric oxidant analyzers, NO 2 interference equivalents 

vary from 20 to 32 percent depending on 0 3 concentration (Tokiwa et al., 

1972). A"corrected" oxidant measurement is obtained by simultaneous measurement of NO2 and correction of the corresponding total oxidant measurement. 

The interference from S02 is quantitative for both colorimetric and electrochemical oxidant measurements, with one mole of S0 2 consuming one mole of 

tri-iodide ion. If the S0 2 concentration is less than that of totai oxidant 

and S02 is simultaneously measured, the total oxidant may also be corrected 

for S02 . This was the procedure previously applied in the older aerometric 

data for California. For many areas of the East Coast and Midwest, such a 

correction was not possible and preferential S0 2 scrubbers were used (Saltzman 

and Wartburg, 1965; Mueller et al., 1973). These scrubbers could be effective 

i n the hands of ski 11 ed operators but thei r use was not wi thout probl ems. 

Among these problems were partial oxidation of NO to NO 2 and of H2S to S0 2 , 

and partial removal of 03 when the scrubber was wet or contaminated (Hodgeson, 

1972). 

4.2.3.2 Ozone 

4.2.3.2.1 Gas-phase chemiluminescence. Many of the 0 3 oxidation reactions 

are sufficiently energetic that they produce electronically excited products, 

intermediates, or reactants, which in turn may chemiluminesce (Zocher and 

Kautsky, 1923; Bowman and Alexander, 1966). Although well known for many 

years, such reactions were not applied to chemical analysis until the 1960s. 

In 1965, Nederbragt reported a detector that employed chemiluminescence from 

the reaction of 0 3 with ethylene for measurement of 0 3 in the vicinity of 

large accelerators (Nederbragt et al., 1965; Warren and Babcock, 1970). 

Applications to atmospheric analysis were a natural consequence (Stevens and 

Hodgeson, 1973). The reference method for 0 3 originally promulgated by EPA 

for compliance monitoring was the 03-ethylene chemiluminescence method (F.R., 

4-7 
1971). Appendix D of 40 CFR, Part 50, describes the principle of the method, 

including a method of calibration (F.R., 1971; C.F.R., 1971). Since then, the 

measurement principle has remained the same but calibration procedures have 

undergone extensive revision as discussed below (Section 4.2.4). It is also 

noteworthy that the reference method is specific for 0 3 , whereas. the data used 

for establishing the standard were based on measurement of total oxidants. 

This issue is addressed in Section 4.2.5. 

A flow of sample air (1 to 5 liter/min) containing 03 and a small flow of 

pure ethylene are mixed at atmospheric pressure in a small reaction chamber 

closely coupled to the photocathode of a photomultiplier tube. The reaction 

between 03 and ethylene produces a small fraction of electronically excited 

formaldehyde. Chemiluminescence from this excited state results in a broad 

emission band centered at 430 nm (Finlayson et al., 1974). The emission 

intensity that is monitored is a linear function of 0 3 concentration from 

0.001 to greater than 1 ppm. The relation between intensity and concentration, 

i.e., instrument calibration, must be determined for each instrument with 

standard concentrations of 0 3 in air. The minimum detection limit and the 

response time are functions of detector design. Detection limits of 0.005 ppm 

and response times of less than 30 seconds are readily attained, however, with 

modest design features. For example, cooling the photomultiplier improves the 

sensitivity but is not normally required. There are no known interferences 

among the common atmospheric pollutants. There have been reports of a positive 

interference when 03 is measured in the presence of water vapor, i.e., a 

signai enhancement of 3 to 12 percent in high humidity'as opposed to measurement of the same concentration of 03 in dry air (California Air Resources 

Board, 1976). Where this may be a real problem, it can be minimized by performing calibrations with humidified air. Finally, in order to obtain acceptable measurement precision and constant span, analyzers must contain means for 

maintaining constant air and ethylene flow rates. 

Ambient air monitoring reference and equivalent methods have been published by EPA (F.R., 1975a). This regulation prescribes methods of testing and 

performance specifications that commercial analyzers must meet in order to be 

designated as a reference method or as an equivalent method. An analyzer may 

be designated as a reference method if it is based on the same principle as 

the reference chemiluminescence method and meets performance specifications. 

An automated equivalent method must meet the prescribed.performance specifications and show a consistent relationship with a reference method. These 

, • 
specifications for 0 3 analyzers are listed in Table 4-1. Commercial analyzers 

that have been designated as reference or equivalent methods are listed in 

Table 4-2. Information concerning the applications supporting the designation 

of analyzers as reference or equivalent methods may be obtained by writing the 

U.S. Environmental Protection Agency, Environmental Monitoring Systems 

Laboratory, Research Triangie Park, NC 27711. 

4.2.3.2.2 Gas-solid chemiluminescence. The first chemiluminescence technique 

for 0 3 was developed by Regener for stratospheric measurements (Regener, 1960) 

and later for measurements in the troposphere (Regener, 1964). The reaction 

of 03 with Rhodamine-B adsorbed on activated silica gel produces chemiluminescence in the red region of the spectrum characteristic of the fluorescence 

spectrum of Rhodamine-B. The intensity is a linear function of 0 3 concentration, 

the minimum detection limit can be lower than 0.001 ppm, and no atmospheric 

TABLE 4-1. PERFORMANCE SPECIFICATIONS FOR AUTOMATED METHODS 

OF OZONE ANALYSIS 

Performance parameter Units Specification 

Range ppm 0 to 0.5 

Noise ppm 0.005 

Lower detectable limit ppm 0.01 

Interference equivalent 

Each interference ppm ±0.02 

Total interference ppm 0.06 

Zero drift, 12 and 24 hour ppm ±0.02 

Span drift, 24 hour 

20% of upper range limit percent ±20.0 

80% of upper range limit percent ±5.0 

Lag time minutes 20 

Rise time minutes . 15 

Fall time minutes 15 

Precision 

20% of upper range limit ppm 0.01 

80% of upper range limit ppm 0.01 

Source: F.R. (1975a); C.F.R. (1975). 

4-9 
TABLE 4-2. LIST OF DESIGNATED REFERENCE AND EQUIVALENT METHODS 

OF OZONE ANALYSIS 

Designati on 

Fed. Re ister notice (E=equiv. 

Identification and source Vo . Page Date Ref. R=ref.) 

Beckman Model 950A Ozone 

Analyzer 

Beckman Instruments 

2500 Harbor Boulevard 

Fullerton, CA 92634 

Bendix Model 8002 Ozone 

Analyzer 

The Bendix Corporation 

Post Office Drawer 831 

Lewisburg, WV 24901 

Columbia Scientific Industries 

Model 2000 Ozone Meter 

11950 Jollyville Road 

Austin, TX 78759 

Dasibi Model 1008-AH Ozone 

Analyzer 

Dasibi Model 1003-AH 

1003-PC or 1003-RS Ozone 

Analyzers 

Dasibi Environmental Corp. 

616 E. Colorado Street 

Glendale, CA 91205 

MEC Model 1100-1 Ozone Meter, 

MEC Model 1100-2 Ozone Meter, 

or MEC Model 1100-3 Ozone Meter 

Columbia Scientific Industries 

11950 Jollyville Road 

P.O. Box 9908 

Austin, TX 78766 

Meloy Model OA 325-2111 Ozone 

Analyzer 

Meloy Model OA 350-2R Ozone 

Analyzer 

Columbia Scientific Industries 

11950 Jollyville Road 

Austin, TX 78759 

42 28571 6/03/77 (F.R., 1977a) R 

41 5145 2/04/76 (F.R., 1976a) 1;7 

45 18474 3/21/80 (F.R., 1980a) 

44 10429 2/20/79 (F.R.,1979a) R 

48 10126 3/10/83 (F.R., 1983) E 

42 28571 6/03/77 (F.R., 1977a) E 

41 46647 10/22/76 (F.R., 1976b) R 

42 30235 6/13/77 (F.R., 1977b) 

40 54856 11/26/75 (F.R., 1975b) 3 

40 54856 11/26/75 (F.R., 1975b) f V 

4-10 
TABLE 4-2 (continued). LIST 

OF 

OF DESIGNATED REFERENCE AND 

OZONE`ANALYSIS 

EQUIVALENT METHODS 

Desi gnation 

Fed. Re ister not ice (E=equiv. 

Identification and source Vol. Page Date Ref. R=ref.) 

Monitor Labs Model 8810 46 52224 10/26/81 (F.R., 1981) E 

Photometric Ozone Analyzer 

Monitor Labs, Incorporated 

10180 Scripps Ranch Boulevard 

San Diego, CA 92131 

Monitor Labs Model 8410E 4]. 53684 12/08/76 (F.R., 1976d) R 

Ozone Analyzer 

Monitor Labs, Incorporated 

10180 Scripps Ranch 

San Diego, CA 92121 

PCI Ozone Corporation Model 47 13572 03/31/82 (F.R., 1982) E 

LC-12 Ozone Analyzer 

PCI Ozone Corporation 

One Fairfield Crescent 

West Caldwell, NJ 07006 

Philips PW9771 0 Analyzer 42 38931 08/01/77 (F.R., 1977c) E 

Philips Electrgnic Instruments, 42 57156 11/01/77 (F.R., 1977d) 

Incorporated 

85 McKee Drive 

Mahwah, NJ 07430 

Thermo Electron Model 49 

UV Photometric Ambient 03 45 

Analyzer 

Thermo Electron Corporation 

Environmental Instruments Division 

108 South Street 

Hopkinton, MA 01748 

57168 08/27/80 (F.R., 1980b) E 

interferences have been observed (Hodgeson et al., 1970). The technique is, 

in fact, more sensitive than the gas-phase Nederbragt method and does not 

require critical control of flow rate. It had the disadvantage in the original 

analyzer built, however, that frequent and periodic internal calibration 

cycles were required to compensate for changes and decaying sensitivity of the 

surface of the detector (Regener, 1964; Hodgeson et al., 1970). 

Improvement was made in the stability of the surface response in a modification added by Bersis and Vassiliou (1966), in which gallic acid is also 

4-11 
adsorbed on the surface in excess. The 03 apparently reacts with and consumes 

the gallic acid rather than Rhodamine-B. An energy transfer step to Rhodamine-B 

subsequent to the initial reaction results in the same chemiluminescence from 

the dye compound, which is now no longer consumed. A commercial analyzer, 

Phillips Model PW9771, is based on this principle and has been designated as 

an equivalent method under EPA regulations. 

4.2.3.2.3 Ultraviolet photometry. Ozone has a moderately strong absorption 

band in the ultraviolet (UV), with a maximum very near the mercury 254 nm 

emission line. This band is essentially a continuum near 250 nm. The molar 

absorption coefficient at the mercury line has been measured by several investigators with good agreement and has an accepted value of 134 ± 2 M 1cm-1 

(base 10) at 0 ° C and 1 atm (Hampson et al., 1973). Ultraviolet absorption has 

long been used as a method of ineasuring gas-phase 03 in fundamental chemical 

and physical studies. Some of the first atmospheric 0 3 measurements were, in 

fact, made by UV photometry; e.g., the Kruger Photometer. These early instruments and the problems with their use are described more completely in the 

first criteria document for photochemical oxidants (National Air Pollution 

Control Administration, 1970). The major problem with the older photometric 

instruments was the larrge imprecision involved in measuring the very small absorbance values obtained. 

This problem of adequate sensitivity with moderate pathlengths has been 

overcome by modern digital techniques for measuring small absorbancies. The 

first instrument of this new generation of photometers was marketed by Dasibi 

of Glendale, California, in the early 1970s. The details of this instrument 

have been described by Bowman and Horak (1972). Other commercial instruments 

have since been marketed and, along with the Dasibi, have been designated as 

equivalent methods by EPA (Table 4-2). All of these instruments operate 

effectively as double-beam digital photometers. A transmission signal is 

averaged over a finite period of time with 03 present and is compared to a 

similar transmission signal obtained through an otherwise identical reference 

air stream from which the 0 3 has been preferentially scrubbed; e.g., using a 

manganese dioxide scrubber. The electronic comparison of the two signals can 

be converted directly into a digital display of.0 3 concentration. 

The UV photometric technique has the advantages, like gas-solid chemiluminescence, that a reagent gas flow is not required and that sample air flow 

control is not critical. In addition, the measurement is in principle an 

4-12 
absolute one, in that the concentration can be computed directly from the 

measured absorbance since the absorption coefficient and the pathlength are 

known. Thi-s capabil,ity is used extensively for the purpose of 0 3 calibration 

as discussed in Section 5.5.5. Commercial UV photometers for 0 3 can serve a 

dual function as a standard for 0 3 calibration and as a means for measuring 

ambient ozone concentrations. In practice, UV photometric analyzers that are 

used for monitoring 03 concentrations in the atmosphere are calibrated with 

standard 03 samples in order to compensate for possible 0 3 losses in the 

sampling and inlet systems. A UV photometric analyzer has the potential 

disadvantage that any molecular species that absorbs at 254 nm (e.g., S0 , 2 , 

benzene, mercury vapor) and that may also be removed along with 0 3 during the 

reference cycle can interfere. Documentation of such interference during 

atmospheric monitoring is. lacking at present. 

4.2.4 Generation and Calibration Methods for Ozone 

Unlike the other criteria pollutants, 0 3 is a thermally unstable species 

that must be generated in situduring the calibration of analyzers. used for 

atmospheric monitoring. This creates special requirements not encountered 

with other pollutants and thus this section deals with means for generating 

dynamic air streams containing stable 03 concentrations and chemical and 

physical means for absolute measurement of these concentrations. 

4.2.4.1Generation . Ozonized sampies of air can be produced by a number of 

means, including photolysis (Brewer and`Milford, 1960), electrical discharge 

(Toyama and Kobayashi, 1966), and radiochemical methods (Steinberg and Dietz, 

1969). Electrical discharges are useful for producing high concentrations of 

03 in air for other appl i cati ons; e.g., 0 3 chemistry. Radiochemical methods 

would be ideal except for their cost and required safety features. By far the 

most common method, however, for generating low concentrations.of 0 3 in air involves the photolysis of molecular oxygen. 

02 + hy(x<200 nm) -> 20 (4- 2) 

0+ 02 + M(M = N2 or 02) _> 03 + M (4- 3) 

One of the most common photolytic generators uses a mercury vapor 6- or 8-inch 

PenRay photolysis lamp positioned parallel to a quartz tube through which air 

flows at a controlled rate. The 03 concentration is simply varied by means 

4-13 
of an adjustable and calibrated mechanical sleeve placed over the lamp envelope 

(Hodgeson et al., 1972) or by varying the voltage or current supplied to the 

lamp. 

4.2.4.2Calibration 

4.2.4.2.1KI procedures: original EPA reference method . The output of 

photolytic 03 generators can provide air samples containing stable 0 8 concentrations over a considerable period of time with careful control of flow rate, 

lamp voltage, temperature, and pressure. It is necessary to calibrate these 

generators periodically with an absolute reference method. Prior to 1975, 

there were as many as seven different calibration methods for 0 3 employed to 

varying extents in this country (National Research Council, 1977). In an 

attempt to standardize the methodology, EPA published a reference calibration 

procedure with the reference method in 1971 when the oxidant (as 0 3 ) standards 

were promulgated (F.R., 1971). This method was the 1 percent neutral buffered 

potassium iodide (NBKI) procedure, a technique that had been used by EPA and 

other agencies for some time. 

During the early 1970s, it became evident that there were serious deficiencies with the NBKI reference method. Several problems with the NBKI procedure, summarized by a joint EPA-NBS workshop in 1974 (Clements, 1975), included the gradual continued release of iodine after sampling, variable results 

obtained with different types of impingers, reagent impurities, and a positive 

bias when compared to other 0 3 measurement methods. An interagency collaborative study was undertaken to intercompare iodometric methods used by the Los 

Angeles Air Pollution Control District (LAAPCD), the California Air Resources 

Board (CARB), and EPA, using UV photometric 0 3 measurements as the rieference. 

The results of this study (DeMore et al., 1976) demonstrated the positive bias 

of the NBKI methods. Concurrent with and after these earlier reports, a large 

number of individual studies ensued. The history of these studies has been 

reviewed in the previous criteria document (U.S. Environmental Protection 

Agency, 1978a) and by Burton et al. (1976). The major conclusions from these 

studies are presented below. 

1. Results obtained by NBKI procedures are higher than those obtained 

by UV photometry or gas-phase titration by 5 to 25 percent, 

depending on details of the procedure. 

4-14 
2. When 03 i s measured i n the presence of humi di fi ed ai r, NBKI 

results tend to be even higher by another 5 to 10 percent (e.g., 

California Air Resources Board, 1976). The reason for this 

apparent moisture effect is not known. 

3. In general, NBKI techniques are subject to large imprecision 

'because of procedural variation. 

The EPA then evaluated four alternative calibration procedures (Rehme et al., 

1981) and selected UV photometry as the reference procedure because of its 

superior accuracy and precision and its simplicity of use. 

Although NBKI methods are no longer used in this country for the purporse 

of calibration, there is a considerable data base available on health and welfare effects, as well as atmospheric chemistry and monitoring, that is based 

on these methods as standards. Therefore, it is important to consider how 

these data may be evaluated and compared to newer effects and aerometric data 

based on the new UV calibration standard. Since a systematic bias is known to 

exist between calibrations by KI methods and UV photometric methods, it should 

be possible, in principle, to apply correction factors to convert from a KI 

reference to a UV photometric reference. There are several problems inherent 

in attempting such corrections, however. A fairly wide range of variations has 

been reported in the literature on the comparison of KI and UV photometric 

measurements. As discussed previously, the presence of moisture in the calibration air increases the magnitude -of the bias. Fortunately; both the CARB 

and the LAAPCD procedures called for the consistent use of humidified air, 

whereas the EPA reference method prescribed the use of dry air. In addition, 

the elapsed time between sample collection and color measurement will also 

affect the magni tude of the bi as because of the sl ow 1 i berati on of i odi ne 

after sampling (Clements, 1975; Beard et al., 1977; Hodgeson, 1976). Other 

unknown experimental factors may also influence the bias, e.g., impinger 

design (Clements, 1975; Beard et al., 1977). 

An assessment has been made of the previous KI versus UV intercompari .sons, 

and recommendations are given in Table 4-3 for correction factors to apply to 

calibration data for conversion from UV to a KI reference or vice versa. It 

should be emphasized that these factors could validly be applied to correct 

for a calibration bias only and can not be applied for comparison of data 

where other effects are present, e.g., the comparison of oxidants versus 0 3 

4-15 
TABLE 4-3. FACTORS FOR INTERCOMPARISON OF DATA CALIBRATED BY 

UV PHOTOMETRY VERSUS KI COLORIMETRY 

Calibration method Ratio, KI/UV 

EPA, 1% NBKI 1.12 ± 0.05 

CARB, 2% NBKI 1.20 ± 0.05 

LAAPCD, 1% UKI 0.96a 

aCorrection for this method not recommended; only one intercomparison has been 

reported. 

data where the effects of oxidizing or reducing interferences must be considered. In this assessment, consideration was given only to those studies in 

which the KI procedure was compared directly to UV photometry. The recommended 

value for data based on the CARB method assumes the use of humidified air. 

The value recommended for the EPA method assumes that dry air was used and 

that color measurement was made immediately after sample collection. 

The uncertainties assigned reflect the fact that a range of values has 

been reported for the ratios in previous studies. Finally, whenever any 

attempt is made to convert from one data base to another, these uncertainties 

must be added by conventional error propagation techniques to the uncertainty 

inherent with the original measurement. 

4.2.4.2.2Ultraviolet photometry method . A major reason UV photometry was 

designated as the calibration procedure was the excellent precision of the 

photometric measurement. In the collaborative study by Rehme et al. (1981), 

measurement with ten individual UV photometers gave only a 3.4 percent variability when compared to a reference measurement system. Other significant 

factors in the selection of UV photometry were the inherent simplicity of UV 

photometric measurements and the ready availability of commercial instruments 

 that can also serve well as transfer standards between laboratory photometers 

and field 03 analyzers. (See National Research Council, 1977, and McElroy, 

1979, for a discussion of transfer standards.). 

It was also presumed that UV photometry gives more accurate results, 

since the accuracy is determined primarily by the 0 3 absorption coefficient, 

which is well known (Hampson et al., 1973; DeMore and Patapoff, 1976). Although 

4-16 
there is little doubt that the accuracy of 0 3 measurements has been significantly improved by conversion to the UV basis, some question still exists 

regarding the absolute relation between 0 3 measurements by UV photometry and 

03 measurements by GPT measurements based on either an NBS standard reference 

material (SRM) nitric oxide (NO) gas cylinder or an NO 2 SRM permeation tube. 

These intercomparisons have been made by several investigators over the past 

10 years and have been summarized by Burton et al. (1976) and Paur and McElroy 

(1979). The agreement between GPT and UV measurements was generally close to 

unity, although in some cases 03 measurements by GPT have shown a small positive 

bias with respect to UV measurements (e.g., Rehme et ai., 1981; DeMore et al'., 

1976). DeMore and Patapoff (1976) reported a ratio of unity between simultaneous measurements of 03 by GPT and UV with a 5 percent uncertainty on the 

ratio of these measurements. In a recent detailed study conducted at the 

National Bureau of Standards (NBS) (Fried and Hodgeson, 1982), 0 3 measurements 

made with an NBS standard photometer (Bass et al., 1977) were compared to GPT 

measurements of 0 3 that were standardized against both NO cylinders (NBS SRM) 

and NO 2 permeation tubes (NBS SRM). Since the measurement of flow rates is a 

critical GPT variable and has been considered as a major source of error in 

GPT measurements (DeMore and Patapoff, 1976), NBS facilities were used for 

making absolute flow measurements by both gravimetric and volumetric means. 

The results of this study were that values of 0 3 measured by GPT based on NO 2 

or NO SRMs agreed to within less than 1 percent, but that,values of 0 3 measured 

by UV were lower by 3 percent. For a consideration of possible error sources, 

reference is made to the original article (Fried and Hodgeson, 1982). In 

summary, the UV photometric 0 3 standard agrees quite closely with the NO and 

NO2 measurement standards by GPT, as it should in principle. The resolution 

of any small biases that remain seems an appropriate matter for consideration 

by EPA and NBS. 

The measurement principle for the absolute measurement of 0 3 by UV photometry is the same as that used by instruments for monitoring atmospheric 0 3 as 

described in Section 4.2.3.2.3 (Bowman and Horak, 1972; DeMore et al., 1976; 

Bass et al., 1977). Ozone is measured in a dynamic flow system by measuring 

the transmission, I/Io, of ozonized clean air in an absorption cell of pathlength Q. When the concentration is to be expressed in units of ppm, measurement of temperature and pressure is also required. The 03 concentration 

may then be calculated directly from the Beer-Lambert equation: 
_ 106 T 760 Io (4-4) 

[03 ~ ppm aQ X ~ X ~ lnT_ 

where a= 03 absorption coefficient at 254 nm, 1 atm, and 0°C, 

= 308 ± 4 atm lcm-i (log base e), 

and 

T = temperature, ° K; 

P = pressure, torr. 

Laboratory photometers used for primary 0 3 calibrations have pathlengths of 1 

to 5 meters and sophisticated digital electronic means for measuring small 

absorbancies (Bass et al., 1977; Bowman and Horak, 1972). 

The conversion of a commercial 03 photometric monitor to a photometer for 

use as a transfer standard for calibration has been described by Paur and 

McElroy (1979). Definitions are in order here. A primary standard UV photometer 

is one that meets the requirements and specifications given in the 1979 revision 

of the 03 measurement and calibration procedures (F.R., 1979e). A transfer 

standard as used by EPA is a device or a method that can be calibrated against 

a primary photometer and transferred to another location for calibration of 0 3 

analyzers. Commercial 03 photometers have served well in this regard, but 

other devices have been used as well; e.g., calibrated generators and GPT 

apparatus. Guidelines on transfer standards for 0 3 have been published by EPA 

(McElroy, 1979), and reference has already.been made to the NAS discussion on 

transfer standards (National Research Council, 1977). 

The use of UV photometry is unique in air pollution measurements in that 

it is based on a physical measurement principle rather than a chemical standard. It is then worthwhile to trace how the measurement chain works from a 

primary standard to field measurements. The primary standard is referenced to 

the accepted 0 3 absorption coefficient. Transfer standards are then calibrated 

with primary photometers maintained at EPA, NBS, and elsewhere. The use of 

commercial photometers in this regard has been described by several investigators (DeMore et al., 1976; Hodgeson et al., 1977). These and other kinds of 

transfer standards are then used to calibrate 03 analyzers used for field 

measurements. 

4-18 
4.2.4.2.3 Other procedures . Although UV photometry has been specified as the 

reference calibration procedure, other procedures are available that can g'ive 

equivalent results. These include a modified KI method, which was allowed as an 

interim alternative method for the calibration of 03 monitors when the UV 

method was designated in 1979. Other KI methods that have been used successfully in Europe are also briefly discussed here. Finally, the GPT method is 

reviewed since it has been used extensively in this country and was discussed 

above with regard to the cross-check of inethod accuracies. 

A major problem with NBKI techniques is the slow release of iodine and 

continued color development after sampling. Flamm (1977) evaluated the.rate 

of this iodine production and found that it was the same as the rate with 

which hydrogen peroxide (H 2O2 ) releases iodine from the same solution. Based 

on this observation and a consideration of other possible species that might 

be responsible, Flamm concluded that certain buffer anions, including phosphate, 

catalyze the formation of H 2O 2 and yield stoichiometries for iodine production 

greater than 1. Measurements made with a 1 percent KI reagent containing 0.1 

M boric acid (BAKI), pH=5, did not exhibit this phenomenon, nor did the original 

EPA method when phosphate was omitted from the reagent (Hodgeson, 1976). 

The BAKI method was evaluated as one of four alternative technique$ ih 

the collaborative study conducted by EPA (Rehme et al., 1981). No significant 

bias was observed between BAKI and the reference technique based on UV'photometry. An analysis, however, of BAKI measurements by ten volunteers revealed 

a large system-dependent variability, and thus the BAKI technique was not. 

recommended as an independent calibration method. It is noteworthy that the 

system variability attributabie to calibration was reduced somewhat if each 

operator assumed a molar absorption coefficient for iodine (as I 2 ) of 2.56 x 

104 M 1cm-1 at 352 nm rather than independently measuring the absorption 

coefficient with standard I 2 solutions as these procedures usually prescribe. 

Measurement systems based on the BAKI procedure may still be certified as 

transfer standards provided the guidelines for certification given in the EPA 

technical assistance document for such standards are foliowed (McElroy, 1979). 

Methods based on iodometry have been used in Europe for some time for the 

calibration of 0 3 analyzers. Bergshoeff (1970) described a method for use in 

the Netherlands, in which thiosulfate is added to the KI reagent (KIT method) 

along with 0.1 M phosphate buffer. The iodine released is immediately reduced 

by the thiosulfate and the amount of iodine consumed is determined by backtitration of the thiosulfate. This method has'the advantage that problems 

4-19 
associated with iodine instability in solution are eliminated. In the Federal 

Republic of Germany, the standard is based on a 2 percent KI reagent with 2 

percent KBr (KIBr Method) and a low concentration (0.02 to 0.03 M) of phosphate 

buffer (Van de Wiel et al., 1979). These techniques have been compared to UV 

and GPT measurement procedures by Van de Wiel. et al. (1979). Measurements 

made with the KIBr method were in essential agreement with measurements by UV 

or GPT, while measurements by the KIT method were too high by 15 to 25 percent, 

depending on the relative humidity of the samples. Modifications have since 

been made in the KIT method by the addition of KBr and reduction of the phosphate concentration. 

The gas-phase titration (GPT) method employs the moderately rapid bimolecular reaction between 0 3 and NO to produce NO2 (Rehme et al., 1974): 

NO + 03 -> NO2 + 02 (4-5) 

This approach was, in fact, one of the early methods used to measure the 

absorption coefficient of 0 3 (Clyne and Coxon, 1968) and yielded excellent 

agreement with other absolute techniques (DeMore and Patapoff, 1976). When NO 

i s present i n excess, no si de reacti ons occur and the stoi chi ometry i s as 

given above. This method has the distinct advantage that it gives an absolute 

relation among three common pollutants. A measurement of the quantity of NO 

or 03 consumed or NO 2 produced provides a simultaneous measurement of the 

other two species and the GPT procedure has been used in all three modes. 

This calibration technique is often used in the calibration of chemiluminescence NOX (NO + NO2 ) analyzers. In order to obtain accurate concentration 

measurements in the procedure as normally employed, accurate flow measurements 

are required; and this is the principal complexity and difficulty with this 

procedure (DeMore and Patapoff, 1976). Stedman et al. (1976) has employed an 

appropriate NO detector to make flow ratio measurements and thus avoid the 

requirement for absolute flow measurements. Because of unexplained biases 

between GPT measurement systems and the UV reference in the EPA collaborative 

study, the GPT method was not recommended as an independent calibration technique (Rehme et al., 1981). It is still allowed, however, as a transfer 

standard in accordance with the EPA guidelines for these standards (McElroy, 

1979). 

4-20 
4.2.5 Relationship between Methods for Total Oxidants and Ozone 

When the ambient air quality, standards for cri.teria pollutants were 

originally established, a numerical standard was set for photochemical oxidants 

as defined by measurements based on iodometric.techniques. Much of the healthrelated and welfare-related evidence used as the basis for the standards was 

obtained using the total oxidant instrumentation discussed above. The reference 

method specified in 1971, however, was the chemiluminescence measurement 

of 03 . Instrumental methods for the specific measurement of atmospheric 0 3 

became commercially available in 1970. These had several practical advantages 

over total oxidant KI-based instruments. These advantages were greater sensitivity, precision, specificity- - no interferences from ambient S02 and NO2--and 

improved reliability in routine monitoring. Second, the data available showed 

that 03 was the major contributor to total oxidant measurements, that 0 3 was 

the major contributor to observed health and welfare effects, and that 0 3 

could probably serve, then, as the best surrogate for measurements of total 

oxidants and for controlling effects of oxidants in ambient air (see reviews 

in Burton et al., 1976; U.S. Environmental Protection Agency, 1978a). 

Notwithstanding the promulgation of standards for ozone rather than 

photochemical oxidants by EPA in 1979, an examination of the temporal.and 

quantitative relationships between total oxidant and 03 data remains of considerable interest, largely because earlier data and many newer data on health 

and welfare effects were obtained by means of total oxidant methods. Aside 

from the relative paucity of data on simultaneous measurements, there are two 

distinct problems in making such comparisons. The first is the difficulty'in 

estimating the contributions to the total oxidant measurements from other 

oxidizing species such as NO 2 and from reducing species such as S02 . 'The 

presence of such species could cause the total oxidant measurements to be 

either higher or lower than 03 concentrations. The second difficulty is in 

estimating the bias created between past and present data as a result of the 

change from the NBKI to the UV photometry calibration procedures. Fortunately, 

these two problems can be treated separately and the iatter problem vanishes 

for comparison of simultaneous 0 3 and oxidant data obtained using the same 

calibration procedure. 

In the sections below, the relationship that should exist between total 

oxidant and 03 is considered from an evaluation of the response of NBKI'measurements to other oxidizing and reducing species. The predicted relationship is 

4-21 
then compared to data obtained in simultaneous field measurements of total 

oxidant and 03 . 

4.2.5.1 Predicted Relationship. The predicted total oxidant measurements can 

be expressed as the sum of the contributions from oxidizing and reducing 

species that release or consume iodine in NBKI reagent: 

[Total Ox] = a[0 3 ] + 7- ib i[Ox] i+ c[NO2 ] (4-6) 

- d[S0 2 ] - 7- ie i[Red] i 

In this equation, [Ox] iand [Red] irepresent the concentrations of other 

oxidizing and reducing species in the atmosphere. The atmospheric concentrations of other reducing species, such as H 2S, are normally quite low compared 

to S02concentrations (Stevens et al., 1972b, and references therein) and 

these species will not be considered further here. If the concentrations 

above are true atmospheric concentrations, the constants a, b, c, and d represent the efficiencies with which the various species release or consume iodine. 

For example, the value of the constant a for an oxidant instrument calibrated 

by the CARB 2 percent NBKI method would be approximately 1.2 (Section 4.2.4.2). 

Since the instruments are calibrated with ozonized air, the factor a represents the bias of the calibration method used. If the 0 3concentration is 

overestimated because of calibration bias, then so are the contributions of 

the other species by the same factor; i.e., the constants b, c, and d are all 

higher than their true values by the same constant, a. Therefore, it should 

in principle be possible to correct total oxidant data for calibration bias by 

dividing both sides of the equation above by a. 

[Total Ox]' = [Total Ox]/a (4-7) 

= [03 ] + 7ib' i [Ox]i + c'[NO2 ] - d'[S0 2 ] 

Other atmospheric oxidants that have been identified and that may contribute to the total oxidant reading are hydrogen peroxide (H 20 2 ), small organic 

peroxides (e.g., methyl and ethyl hydroperoxide), peracetic acid, peroxyacyl 

nitrates (Cohen et al., 1967), and pernitric acid (Niki et al., 1976). An 

estimation of the contribution of these species to the total oxidant measurement is quite difficult because individual b i 's have not been measured and 

there are few data available on atmospheric concentrations of individual 

species. The magnitude of the efficiency term will depend not only on the 

4-22 
stoichiometry of the oxidation reaction, but also on the rate. A summary of 

the effects of various oxidants on NBKI reagent and the Mast oxidant meter is 

given in Table 4-4 (Cohen et al., 1967; Purcell and Cohen, 1967; Burton et 

a1., 1976). 

TABLE 4-4. RESPONSE OF NBKI REAGENT AND MAST METER TO VARIOUS OXIDANTS 

NBKI Mast 

Ozone Aa E 

Peracetic acid A NAb 

Nydrogen peroxide B D 

Acetyl peroxide B N 

Ethyl hydroperoxide B NA 

n-Butyl hydroperoxide B NA 

tert-Butyl hydroperox7de B NA 

Nitrogen oxides (NO x) D(10% as NO2 ) D(10%) 

Peroxyacetyl nitrate (PAN) D N 

Peroxypropionyl nitrate (PPN) D D 

aA = immediate color development; B= slow color development; D= positive 

interference; E= good response; and N= no response (or negligible). 

b NA = Data not available. 

Source: Cohen et al. (1967). 

By contrast, reaction efficiencies for NO 2 and S0 2 are relatively well 

known. Tokiwa et al. (1972) observed reaction efficiencies of 6 percent for 

the Mast oxidant meter, 22 percent for a 10 percent KI colorimetric analyzer, 

and variable (20 to 32 percent) for a 20 percent KI colorimetric analyzer. It 

is well documented that S0 2 is a quantitative negative interference with a 100 

percent efficiency for reducing the oxidant reading by an amount equi.valent to 

the S0 2 concentration present (Cherniack and Bryan, 1965; Saltzman and Wartburg, 

1965). 

4-23 
Returning to the analytical expression for total oxidant,, an oxidant 

value corrected for NO2 and S02interferences can be expressed "as below, assuming that no other significant reducing interferences are present: 

[Total Ox] corr. = [Total Ox] - c[NO2 ] + [S0 2 ] ( 4-8) 

= [03 ] + Fi b i[Ox] i 

Thus, a total oxidant measurement for which legitimate corrections or compensations for NO 2and S02have been made should always be higher than a 

simultaneous 03 measurement by an amount that is a function of the type and 

concentrations of other oxidants present. The only major qualifications to this 

prediction are that both types of ineasurements must be sampling the same air 

mass and be calibrated with respect to the same reference; that no other 

significant reducing interferences are present; and that 03 losses within 

sample inlet systems are insignificant. On the other hand, total oxidant data 

uncorrected for S02and NO 2interferences may be higher.or lower than corresponding 0 3 data, depending on the concentrations of these,pollutants. Because 

of the potential presence of these interferences, it is quite difficult or 

impractical to compare oxidant and 0 3measurements during evening and early 

morning hours, when 0 3concentrations are quite low. Therefore, in the comparison of total oxidant and 0 3 simultaneous field measurements below, emphasis 

is placed on comparison of peak hourly averages. 

4.2.5.2 Empirical Relationship Determined from Simultaneous Measurements. 

Several precautions should be taken in performing simultaneous measurements. 

Both kinds of instruments must be calibrated frequently with the same ozonized 

air stream that has been analyzed by a common refer.ence method. In a simultaneous comparison, daily calibrations should be made with an 03generator and 

the generator output should be analyzed weekly. Both instruments should 

sample the same air parcel. Routine maintenance should be frequent to ensure 

constant gas and reagent flow rates, clean sample inlet systems, etc. Finally, 

in any meaningful comparison of 0 3 and oxidant data, simultaneous measurements 

of NO2and S02should be made. If chromium trioxide scrubbers are used to 

remove S02 in the inlet to the oxidant instrument, these must be frequently 

tested to ensure that 03is not also removed during continued use, particularly under very humid conditions. These scrubbers may cause some additional 

bias by oxidation of NO to NO2 . 

4-24 
The eariiest comparative study reported was by Renzetti and Romanovsky 

(1956). This study compared a phenolphthalein total oxidant monitor, a KI 

continuous oxidant -monitor, a rubber-cracking apparatus, and an open-path 

ultraviolet spectrometer, which monitored the UV absorption at characteristic 

03 absorption wavelengths. The only meaningful measurements for consideration 

here are the KI oxidant and UV 03 measurements,.since these are similar to 

measurement methods used later. The UV 0 3 spectrometer differed in a number 

of respects from the 0 3 photometers in use today. Measurements were made 

across an open optical path of 325 ft of the transmission at three wavelengths, 

x1 = 265 nm, X2 = 313 nm, and X3 = 280 nm. Intensity ratios at the three 

wavelengths were used to minimize the effects of other UV absorbers and of 

particulate scattering. 'Some non-03 absorption may still have been present, 

and, if so, the measured values would be higher than the true 03 values. The 

published absorption coefficients at these three wavelengths were used to compute 0 3 concentrations (Vigroux, 1952). Measurements were made over a 4-month 

period. Figures 4-1 and 4-2 are illustrative of the data obtained for a 

monthly average and a single day, respectively. Peaks of tota1 oxidant and of 

03 occurred at the same time, but the 0 3 maximum was usually less than the 

total oxidant maximum. The UV-03 data were usually higher in the wings at low 

0 3 concentrations. Renzetti and Romanovsky attributed the higher total oxidant 

reading to the presence of "other. oxidants" and estimated concentrations of 

other oxidants of 0.1 to 0.4 ppm, depending on 03 concentration. Since the 

total oxidant instrument was calibrated by an NBKI method, this estimate is 

almost certainly too high and a large portion of the difference between oxidant 

and 0 3 may have been a result of the 20 to 25 percent positive calibration 

bias. Since interferences may be present in the UV measurement and simultaneous measurements of NO 2 and S02 were not available at the time, no attempt is 

made to make any more quantitative assessment of this study. 

A later study (Cherniack and Bryan, 1965) compared a 10 percent colorimetric KI oxidant instrument, a Mast oxidant meter (Brewer and Milford, 1960), a 

galvanic-celi oxidant instrument (Hersch and Deuringer, 1963), and a UV 0 3 

photometer (Bryan and Romanovsky, 1956). This latter instrument was similar 

in principle to present-day photometers.. The precautions noted above were 

taken. A11 the instruments were calibrated with respect to the 2 percent UKI 

calibration procedure used by the LAAPCD. Simultaneous S0 2 and NO2 measurements were made, but no corrections were made because the concentrations were 

reported to be quite small during the period of comparison. Atmospheric sampling 

4-25 
E 

.c 

~. 

a 

z 

O 

Q ~ cc H 

Z 

w 

U 

Z 

O 

V l 

12 1 2 3 4 5 6 7 8 9 10 11 12 1 2 3 4 5 6 7 8 9 10 11 12 

a.m. AUGUST 1955 p.m. P.S.T. 

Figure 4-1. Comparison of ozone and total oxidant concentrations in the Pasadena area, August 1955. 

Source: Renzetti and Romanovsky (1956). 

~ 

r 

a 

a. 

Z O 

Q 

~ N- Z w 

V Z O 

V 

12 1 2 3 4 5'6 7 8 9 10 11 12 1 2 3 4 5 6 7 8 9 10 11 12 

a.m. AUGUST 27, 1955 O.M. P.S.T. 

Figure 4-2. Comparison of ozone and total oxidant concentrations in the Los Angeles area, 

August 1955, 

Source: Renzetti and Romanovsky (1956). 

4-26 
was conducted over an unspecified period of time, and the data were referenced 

to the colorimetric oxidant measurements. The linear regression analysis of 

the data over the concentration range 0 to 0.6 ppm gave the relationships 

shown in Table 4-5 after correction for calibration factors. Thus, the data: 

show a much better absolute agreement and correlation between 0 3 measurements 

and colorimetric total oxidant than between electrochemical total oxidant and 

colorimetric total oxidant. In addition, these data do not indicate any 

significant contribution by "other oxidants" to the total oxidant measurement'. 

TABLE 4-5. COMPARISON OF CORRECTED INSTRUMENT READINGS TO COLORIMETRIC 

OXIDANT READINGS DURING ATMOSPHERIC SAMPLING 

Instrument m b r 

Mast meter 0.896 -0.013 0.868 

Galvanic cell 0.776 +0.004 0.867 

Ozone photometer 0.980 -0.005 0.982 

(y = Instrument reading 

= mx + b; x= colorimetric oxidant measurement; m= slope; b= intercept; 

r = correlation coefficient) 

Source: Cherniack and Bryan (1965). 

During the 1970s, several studies were conducted on the intercomparison 

of 0 3 and total oxidant instrumentation by the Research Triangle Institute 

(RTI) of North Carolina (Ballard et al., 1971a,b; Stevens et ai., 1972a,b). 

Measurements were made for 0 3 by chemiluminescence and for total oxidant by a 

colorimetric KI analyzer and a Mast meter. Calibrations were carried out 

Mast meter. Calibrations were carried out frequently with an 03 generator 

calibrated by the 1 percent NBKI method. Simultaneous NO 2 and S02 measurements 

were also made and the oxidant data reported were corrected for these interferences. Clark et al. (1974) intercompared a commercial UV photometer, three 

different commercial gas-phase chemiluminescence analyzers, and a gas-solid 

chemiluminescence analyzer by monitoring in a rural environment. The instru - 

ments were all calibrated by a common reference procedure and hourly-averaged 

field measurements were collected over a 1-month period in August 1972. Davis 

and Jensen (1976) reported intercomparisons of Mast meter total oxidant,measure - 

ments and chemiluminescence 0 3 measurements. The instruments were not calibrated 

by the same procedure, however, nor were any corrections attempted for S0 2 and 

4-27 
NO2 interferences. Okita and Inugami (1971) reported an intercomparison of KI 

total oxidant measurements with chemiluminescence 0 3 measurements in the urban 

atmosphere of Musashino, Japan. An intercomparison of total oxidant by KI and 

03 by chemiluminescence in irradiated auto exhaust was reported by Carroll et 

al. (1972). In another extensive field study conducted at an air monitoring 

station near the Houston ship channel, Severs and coworkers (Severs, 1975; 

Neal et al., 1976) examined the relationship between ozone and total oxidants 

for this area by making simultaneous measurements with a gas-phase chemiluminescence 03 monitor and a Beckman colorimetric total oxidant analyzer. Primary 

calibrations of the instruments were performed periodically using the EPA 1 

percent NBKI method. No corrections were attempted for NO 2 or S02 interferences, but during the latter part of this study a chromium trioxide scrubber 

was placed in the inlet of the total oxidant analyzer. 

All of these 1970 studies were reviewed in the previous criteria document 

(U.S. Environmental Protection Agency, 1978a). Only the major conclusions are 

repeated here. In general, the averaged data showed fairly good qualitative 

and quantitative agreement between the diurnal variations of total oxidants 

and 03 . The usual trend was a slightly higher value for the total oxidants 

measurement at the maximum, a not unexpected result in view of the discussion 

above. Comparisons of monthly-averaged data taken from studies in Los Angeles 

and St. Louis are shown in Figures 4-3 and 4-4 (Stevens et al. , 1972a,b). 

The total oxidant data shown in Figure 4-4 are uncorrected and show distinct 

morning and evening peaks resulting from NO 2 interference (see Chapter 3 for 

diurnal patterns of NO 2 ). Examination of data taken from individual days 

shows considerably more variation among the methods, with total oxidant measurements both higher and lower than 03 measurements. Intercomparisons of only UV 

photometric and chemiluminescence 03 analyzers have not shown these large 

variations (Clark et al., 1974; Wendt, 1975). In all probability, these 

variations result from the large imprecision and interferences in total oxidant 

measurement. 

Two of the studies described above reported consistently lower total 

oxidant measurements. In one of these (Davis and Jensen, 1976), the reference 

KI method was used for calibrating the chemiluminescence analyzer•while a 

factory calibration was used for the Mast meter. As pointed out above, other 

studies have found low oxidant readings for the Mast meter as compared to 

colorimetric analyzers (Cherniack and Bryan, 1965; Tokiwa et al., 1972; Stevens 

et al., 1972a,b). The use of the factory calibration would cause the 

4-28 
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Mast readings to be even lower because of a calibration bias (Cherniack and 

Bryan, 1965; Tokiwa et al., 1972). The results reported by Severs and coworkers 

are more difficult to evaluate. Chemiluminescence 0 3 values generally higher, 

and sometimes considerably higher, than total oxidant measurements were reported, although the measurements were referenced to the same calibration 

procedure. Correlations were reported both with and without a chromium trioxide 

scrubber in the oxidant inlet. These results are inconsistent with the known 

responses of the instruments and the results of other investigators. The data 

reported for one day of high 03 , but abnormally low oxidants, are shown in 

Figure 4-5. It is highly improbable that the problem is with the 

chemiluminescence 0 3 measurement, since this response is typical of a normal 

0 3 diurnal variation and no other species are known to interfere. It is more 

probable that some other species of pollutant in the highly industrialized area 

of the Houston ship channel repressed the response of the total oxidant analyzer, 

which thus does not respond to 0 3 , much less to any other oxidant. 

The most recent comparison in the literature involved simultaneous 0 3 and 

total oxidant measurements in the Los Angeles basin by the California Air 

Resources Board (1978) in the years 1974, 1976, and 1978. The maximum hourly 

data pairs were correlated (Chock et al., 1982) and yielded the foilowing 

regression equation for 1978 data, in which a large number (927) of data pairs 

were available: 

Oxidant (ppm) = 0.870 0 3 + 0.005 (4-9) 

(correlation coefficient = 0.92) 

Thus, when the 0 3 levels were relatively high, they were actually slightly 

higher than total oxidant. The total oxidant data were uncorrected for NO 2 

and S0 2 interferences. 

In summary, specific 0 3 measurements agree fairly well with total oxidant 

corrected for NO2 and S02 interferences and 0 3 is the dominant contributor to 

total oxidant. Indeed, it is difficult to discern the presence of other 

oxidants in most total oxidant data. There can, however, be major temporal 

discrepancies between 0 3 and oxidant data, which are primarily a result of 

oxidizing and reducing interferences with KI measurements. As a result of 

these interferences, on any given day the total oxidant data may be higher 

4-31 
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4-32 
than or lower than simultaneous 0 3 data. The quantitative relationship between oxidant and 03 data, such as that used by Chock et al. (1982), is probably quite location-dependent. From a methodologic standpoint, the measurement 

of 0 3 is a more reliable indicator than total oxidant measurements of oxidant 

air quality, and such difficulties and controversy as may be involved in the 

intercomparison of 03 and oxidant measurements are eliminated if the air 

quality standard is defined in terms of 0 3 . 

4.2.6 Methods for Sampling and Anal sis of Perox acet 1 Nitrate and Its 

omoloc ~ues 

4.2.6.1 Introduc tion. Since the discovery of "compound X" (Stephens et al., 

1956a,b), later unambiguously identified as peroxyacetyl nitrate (PAN) (Stephens 

et al., 1961), much effort has been directed toward its atmospheric measurement. 

Peroxyacetyl nitrate and its homologues are products of photochemical reactions 

involving hydrocarbons and oxides of nitrogen (NO x) in the atmosphere (Stephens 

et al.,. 1961). The significance of atmospheric PAN is twofold. It is a 

potent lachrymator and phytotoxicant in the ppb concentration range (Stephens 

et al., 1961; Heuss and Glasson, 1968). Because of the reversible thermal 

equilibrium (Hendry and Kenley, 1977), 

CH30(02 )NO2E CH3C0(02 )• + NO 2 , (4-10) 

which is sensitive to the NO 2/NO ratio, PAN may serve as an important reservoir 

for peroxy radicals and NO 2 (Singh and Salas, 1983a,b) and may play a significant role in both the atmospheric nitrogen cycle and in tropospheric ozone 

formation (Spicer et al., 1983). 

Oniy two analytical techniques have been used to obtain significant data 

on ambient PAN concentrations. These are gas chromatography with electron 

capture detection (GC-ECD) and long-path Fourier-transform infrared (FTIR) 

spectrometry. Atmospheric data on PAN concentrations have been obtained 

predominantly by GC-ECD because of its relative simplicity and superior sensitivity. These analytical techniques are described in Section 4.2.6.2 along 

with attendant methods of sampling. Peroxyacetyl nitrate is somewhat analogous 

to 03 in that it is a thermodynamically unstable oxidant and PAN standards 

must be generated and analyzed by some absolute technique for the purpose of 

calibrating the GC-ECD. Generation and calibration techniques are discussed 

separately in Section 4.2.6.3. Finally, the analysis of PAN homologues is 

discussed briefly in Section 4.2.6.4. 

4-33 
4.2.6.2Analytical Methods for PAN . By far the most widely used technique 

for the quantitative determination of ppb concentrations of PAN is GC=ECD 

(Darley et al., 1963; Stephens, 1969). With.Carbowax or SE30 as a stationary 

phase, PAN can be separated from components such as ai r, water, and other 

atmospheric compounds, as well as ethyl nitrate, methyl nitrate, and other 

contaminants that are present in PAN synthetic mixtures. Electron-capture 

detection (using a nickel-63 source and a pulsed-current detector or a tritium 

source and a direct-current detector) provides sensitivity for PAN in the ppb 

and sub-ppb ranges. A typical column for the separation of PAN would be 3 to 

5 feet in length and 1/8 inch in diameter (i.d.), and would be run isothermally 

at 250 to 60°C. Under these conditions, a peak assignable to PAN appears 

after 2 to 3 minutes. Table 4-6 shows parameters used by several investigators 

to determine trace levels of PAN by GC-ECD. 

Sample injection into the GC is accomplished by means of a gas-sampling 

valve with a gas-sampling loop of a few milliliters volume (Stephens and 

Price, 1973). Sample injection may be performed manually or automatically. 

Typically, manual air samples are collected in 50 to 200 ml ungreased glass 

syringes, and purged through the gas-sampling valve. Samples collected from 

the atmosphere should be analyzed as soon as possible because PAN undergoes 

thermal decomposition in the gas phase and at the surface of containers. 

Automatic sample collection and injection may be accomplished by using a small 

pump to pull ambient air continuously through the sampling loop of an automatic 

sampling valve, which periodically injects the sample onto the column (Stephens 

and Price, 1973). Recently, Singh and Salas (1983b) have used cryogenic trapping of PAN, with liquid argon, from relatively large air samples for the 

purpose of ineasuring PAN concentrations in the sub-ppb range. 

Most of the atmospheric PAN measurements have been made in polluted urban 

environments, where maximum concentrations of 5 to 50 ppb and average concentrations of a few ppb may occur (Stephens, 1969; Lonneman et al., 1976). For 

the purpose of such measurements, chromatographic detection limits of 0.1 to 

1 ppb are sufficient. The recent work of Singh and Salas (1983a,b) on the 

measurement of PAN in the free (unpolluted) troposphere is illustrative of 

current capabilities for measuring low concentrations. A 50- to 200-m1 volume 

of air was collected by preconcentration into an unpacked 0.15 cm o.d. stainless 

tube of 1.24 ml volume, at liquid argon temperature, prior to analysis (see 

Table 4-6). For measurements in humid environments, the air sample was passed 

4-34 
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through a Nafion drier (Foulger and Simmonds, 1979) to reduce the humidity 

prior to preconcentration. A minimum detection limit of 0.01 ppb was obtained. 

Free tropospheric concentrations in the 0.01 to 0.1 ppb range were always 

observed and indicated that PAN is a natural constituent of the atmosphere and 

may constitute a significant fraction of the reactive nitrogen. 

There are conflicting reports in the literature on the effects of variable 

relative humidity on PAN measurements by GC-ECD. In 1973, Stephens and Price 

stated that in preparing PAN calibration samples "the diluent (gas) should be 

of normal humidity so that the chromatogram will be a realistic one." Subsequently, Holdren and Rasmussen (1976) observed a reduced response to PAN 

calibration samples when the relative humidity of the sample was 30 percent or 

lower and a tenfold decrease in PAN response when the relative humidity approached 0 percent. This' effect was attributed to. an interaction between 

' sample and GC column. Nieboer and van Ham (1976) reported that "the elution 

gas stream was previously humidified ... because it appeared that the height 

of the PAN peak depends on the relative humidity of ambient air if dry elution 

gas was used." In contrast to these studies, Lonneman (1977) observed no 

effect on peak height in PAN calibration samples in which the relative humidity 

varied from 10 to 50 percent. 

In 1978, Watanabe and Stephens reported on a reexamination of the moisture 

anomaly and investigated the effects of humidity on PAN storage flasks, columns, 

and detectors. A consistent PAN loss to the walls of dry acid-washed glass 

storage flasks was observed, but the PAN could be recovered by the addition of 

moisture to the flasks. A tritium direct-current detector showed no humidity 

effect except for a small (5 to 10 percent) decrease in peak height in a few 

cases at very low humidities (2 to 3 percent). With a different GC instrument 

employing a Ni-63 detector, erratic responses were observed at low humidities, 

with responses reduced 30 to 95 percent from that obtained at 53 percent 

relative humidity. No conclusion was drawn on whether this difference reflected a humi di ty effect on the detector, the col umn, or the sampl i ng val ue. 

Finally, the moisture anomaly did not appear to depend on column history or 

loading even after a bake-out treatment. , 

In the most recent study (Grosjean et al., 1984), the humidity effect on 

PAN calibration samples prepared by the dynamic and static irradiation of 

CH3CH0-C1 2-NO2 mixtures was examined. A small decrease (<3 percent) was 

observed in PAN peak height when the dry air stream was passed either over an 

impinger containing water or directed through the water impinger. These 

4-36 
results are in contradiction to all of the above, in which the response to 

humidified PAN samples is either greater than or the same as dry PAN samples. 

It is noteworthy that the chromatographic systems employed by different 

investigators often employ different materials :(e.g., glass, Teflon, stainless 

steel), column loadings, and detectors. The resolution of all the differences 

noted above in regard to a suspected humidity effect. might require considerable 

effort. For the present, if a moisture effect is suspected in a PAN analysis, 

the bulk of this evidence suggests that humidification of PAN calibration 

samples (to a range approximating the humidity of the samples being analyzed) 

would be advisable. 

Conventional long-path infrared spectroscopy and Fourier-transform infrared 

spectros`copy (FTIR) have been used to detect and measure atmospheric.PAN., 

Sensitivity is enhanced by the use of FTIR over conventional l.ong-path infrared 

spectroscopy. Accurate knowledge of the absorptivities of many IR bands , 

ass.ignable to PAN makes possible the quantitative analysis of PAN without the 

use of calibration standards. The most frequently used IR bands have been 

assigned and the absorptivities shown in Table 4-7 have been reported. Only 

the key bands are shown, but all 27 fundamentals are infrared active and 

Bruckmann and Willner (1983) have assigned most of them. The assignment by 

Adamson and Guenthard (1980) of the bands at 1435, 1300, and 990 cm -1 to an 

impurity, CH 30NO2 , is apparently incorrect. Bruckmann and Willner (1983) observed these same bands in a 99 percent pure PAN sampTe. - 

The initial di.scovery of PAN in'simulated photochemical smog was . accomplished by long-path infrared absorption spectrometry (Stephens et a1., ; 1956b). 

Some recent simultaneous measurements of PAN and other atmospheric pollutants 

such as 0 3 , HNO3 , HCOOH, and HCHO have been made by long-path FTIR spectrometry during smog episodes in the Los Angeles Basin. Tuazon et al. (1978) have 

described an FTIR system operable at pathlengths up to 2 km for ambient measurements of PAN and their trace constituents. This system employed an eight-mirror 

multiple reflection cell with a 22.5-m base path. The spectral windows available at pathlengths of 1 km were 760-1300, 2000-2230, and 2390-3000 cm l . 

Thus, PAN could be detected by the bands at 793 and 1162 cm-1 . The 793 cm 1 

band is characteristic of peroxynitrates, while the 1162 cm -1 band is reportedly 

caused by PAN only (Stephens, 1969; Hanst.et al.., 1982). Tuazon et al. (1981a,b) 

reported ambient measurements with this system during a smog episode in 

Claremont, CA, in 1978. zMaximum PAN concentrations ranged from 6 to 37 ppb 

over a 5-day episode; the report presented diurnal patterns for PAN and several 

019QQ/A 4-37 
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other pollutants for the 2 most severe days. The detection limit given for 

PAN at a 900-m pathlength was 3 ppb. 

Hanst et al. (1982) modified the FTIR system used by Tuazon et al. (1978) 

by changing it from an eight-mirror to a three-mirror cell configuration and 

by considerably reducing the cell volume. Measurements were made over a 

1260-m optical path folded along a 23-m base path at 0.25 cm 1 resolution. 

Measurements were reported for PAN and a variety of other pollutants for a 

2-day smog episode at California'State University, Los Angeles, in 1980. The 

maximum PAN concentration observed was 15 ppb for this period of only moderate 

smog intensity. An upper limit of 1 ppb of peroxybenzoyl nitrate (PbzN) was 

placed based on observations in the vicinity of the PBzN band at 990 cm -1 . 

The reports by Tuazon et al. (1978) and Hanst et al. (1982) both refer to 

earlier FTIR ambient air studies. 

Sampling may constitute one of the major problems in the analysis of 

trace reactive species, such as PAN, by long-path FTIR spectrometry. The 

folded-path White cells have a large internal volume (15 m 3 for Tuazon et al., 

1978; 3 m3 for Hanst et al., 1982). The large internal surface area may serve 

to promote the decomposition or irreversible adsorption of reactive trace 

species. To minimize these effects, both Tuazon et al. and Hanst et al. 

employed high-speed blowers to pul1 ambient air through the ce11s at high 

velocities. For interior cell linings, Hanst et al. employed 0.5 mm polyvinyl 

chloride sheeting and Tuazon et al. used Plexiglas and FFP Teflon. 

Pitts et al. (1973) proposed a chemiluminescence technique for continuous 

monitoring of ambient concentrations of PAN. The reactions of both PAN and 0 3 

with triethylamine in the gas phase produce chemiluminescence. The spectra 

reported overlap somewhat with aILmax of 520 nm for the 0 3 reaction and "max 

of 650 nm for the PAN reaction. Pitts proposed a technique that included the 

measurement of the emission intensity in the two regions by the use of optical 

cut-off filters. Thus, the PAN concentration can be determined, provided the 

absolute 03 concentration is simultaneously measured. Concentrations of 6 ppb 

PAN were detected and a lower limit of detection of 1 ppb was estimated. No 

interfering emissions were observed from methyl nitrate, ethyl nitrate, ethyl 

nitrite, or NO2 . No'further work has been reported on the development of this 

technique, and there have been no atmospheric applications. 

4.2.6.3Generation and Calibration of PAN . Because of the thermal instability 

of dilute PAN samples and the explosive nature of liquid PAN, calibration 

samples are not commercially available; each laboratory involved in making 

4-39 
such measurements must prepare its own standards. The PAN samples are prepared 

by various means at concentrations in the ppm range and these must be analyzed 

by some absolute technique. The analyzed samples must then be diluted to 

obtain gas-phase samples in the 1ow ppb range for direct calibration of GC 

instruments. Thus, this section includes descriptions of various means of PAN 

generation, methods of analysis, and the procedures for sample handling and 

storage where applicable. 

The earlier methods used for the preparation of PAN have been summarized 

by Stephens (1969). These included (1) the photolysis of mixtures of nitrogen 

oxides with organic compounds in air or oxygen (Stephens et al., 1956b; 

Stephens et al., 1961); (2) the photolysis of alkyl nitrite vapor in oxygen 

(Stephens et al., 1965); (3) the dark reaction of aldehyde vapor with nitrogen 

pentoxide (Tuesday, 1961); and (4) the nitration of peracetic acid. Of these 

methods, the photolysis of alkyl nitrites was favored and used extensively by 

Stephens and other investigators. As described by Stephens et al. (1965), the 

liquefied crude mixture obtained at the outlet of the photolysis chamber is 

purified by preparation-scale GC. [CAUTION: Both the liquid crude mixture 

and the purified PAN samples are violently explosive and should be handled 

behind explosion shields using plastic fu11-face protection, gloves, and a 

leather coat at a11 times. These PAN samples should be handled in the frozen 

or gaseous state whenever possible.] The pure PAN is usually diluted to about 

1000 ppm in cylinders pressurized with nitrogen to approximately 100 psig. 

When refrigerated at <15°C, PAN losses are 1ess than 5 percent per month 

(Stephens et al., 1965). Lonneman et al. (1976) used the photolysis products 

without purification for the calibration of GC instruments in the field and 

discussed the use of Tedlar bags for the preparation and transport of calibration samples. 

Gay et al. (1976) have used the photolysis of C1 2 : aldehyde: NO 2 mixtures 

in air or oxygen for the preparation of PAN and a number of its homologues at 

high yields: 

C1 2+ hv = 2 C1 (4-11) 

0 0 

u u 

C1 + RC-H R-C' + HC1 (4-12) 

4-40 
0 0 

RC' + 02 + M RC-02 ' + M (4-13) 

0 0 

RC-02 ' + NO2 =RC-02-NO2 (4-14) 

This procedure has been utilized in a portable PAN generator that can be used 

for the calibration of GC-ECD instruments in the field (Grosjean, 1983; 

Grosjean et al., 1984). The output of this generator is a dynamic flow of PAN 

in air at a concentration of about 2 to 450 ppb. Dilute concentrations of 

reactant gases for the photolysis chamber are obtained by passing a controlled 

flow of air over C1 2 , NO2 , and acetaldehyde permeation tubes. 

The other technique for PAN preparation in current use involves the 

nitration of peracetic acid. In the 1969 review (Stephens, 1969), this approach 

was considered not useful for synthesis. Several investigators, however, have 

recently reported on a condensed-phase synthesis of PAN with peracetic acid 

that produces high yields of a pure product free of other alkyl nitrates 

(Hendry and Kenley, 1977; Kravetz et al., 1980; Nielsen et,al., 1982; Holdren 

and Spicer, 1984). Most of these procedures call for the addition of peracetic 

acid (40 percent in acetic acid) to a hydrocarbon solvent (pentane, heptane, 

octane) maintained at -80° C in a dry-ice acetone bath, followed by acidification 

with sulfuric acid. Nitric acid is formed in situwith stirring by the slow 

addition of sodium nitrate. After the nitration is complete, the hydrocarbon 

fraction, containing PAN concentrations of 2 to 4 mg/ml (Nielsen et al., 

1982), can be stored at -20°C for periods longer than a year (Holdren and 

Spicer, 1984). After analysis, the PAN-hydrocarbon solutions can be used 

directly for calibration by the evaporation of ineasured microliter volumes of 

solution into Tedlar bags containing known volumes of clean air. 

The most direct method for absolute analysis is by infrared absorption 

using absorptivities given in Table 4-7. This is the technique used by Stephens 

(1969; analysis of PAN in N2 cylinders), Lonneman et al. (1976; analysis of 

gas-phase products from photolysis of ethyl nitrite); and Holdren and Spicer 

(1984; analysis of PAN in octane solutions). Whereas long, folded-path cells 

and FTIR spectrometry are required for the analysis of atmospheric PAN, conventional IR instruments and 10-cm gas cells can analyze gas standards with 

concentrations greater than 35 ppm (Stephens, 1969) and Holdren and Spicer 

4-41 
(1984) used 50-pm liquid microcells for the analysis of PAN in octane solutions. 

Another candidate technique for absolute PAN analysis was gas-phase coulometry 

using a tandem el ectro n- capture detector (Lovelock et al., 1971). Singh and 

Salas (1983b) have shown, however, that this technique is unsuitable for absolute PAN analysis because a significant fraction of the PAN is destroyed prior 

to coulometric detection. 

The alkaline hydrolysis of PAN to acetate ion and nitrite ion in quantitative yield (Nicksic et al., 1967) provides a means independent of infrared 

for the quantitative analysis of PAN. Molecular oxygen is also produced in 

quantitative yield by the reaction (Stephens, 1967): 

0 0 

11 _ 11 - 

- CH3COONO2 + 20H ~ CH3C0 + 02 + NO2 + H20 (4-15) 

The colorimetric determination of nitrite ion with Saltzman reagent was first 

used to measure PAN quantitatively (Nicksic et al., 1967; Stephens, 1969; 

Kravetz et al., 1980). Nielsen et al. (1982) analyzed the hydrolyzed products 

of pure PAN samples by ion chromatography for nitrite and nitrate and found 

that 4 percent of the nitrite had been oxidized to nitrate. Some gas-phase 

PAN calibration samplEs (e.g., photolysis of C1 2 : acetaldehyde: NO2 ) contain 

impurities such as NO 2

. that will yield nitrite and nitrate in aqueous solution. 

Thus, Grosjean (1983) and Grosjean et al. (1984) performed ion chromatographic 

analysis of the acetate ion to determine the PAN output of a portable generator. 

An alternate calibration procedure has been proposed based on the thermal 

decomposition of PAN in the presence of excess N0, and measurement by chemiluminescence of the NO consumed (Lonneman et al., 1982). The acetylperoxy radical, 

CH3C(0)021 and its decomposition products rapidly oxidize NO to NO2 . In the 

presence of a small amount of benzaldehyde, which is used to scavenge the 

hydroxyl radical and control the stoichiometry, simulation models predict that 

5 molecules of NO will be removed per PAN molecule present. By the use of NO 

and PAN standard mixtures and the chemiluminescent measurement of the NO 

consumed, the experimental value was determined to be ONO/OPAN = 4.7 ± 0.2. 

This measurement could be performed in field stations where chemiluminescent 

NO analyzers are usually available. 

4.2.6.4 Methods of Analysis of Higher Homologues. The GC-ECD analyzer is 

likewise used for the higher homologues of PAN (Darley et al., 1963; Stephens, 

1969; Heuss and Glasson, 1968). The higher homologues elute with longer 

4-42 
retention times. The first observation of peroxypropionyl nitrate (PPN) in 

heavily polluted air was by Darley et al. (1963), who also measured peroxybutyryl (PBN) nitrate in synthetic mixtures by GC-ECD. The concentrations of 

the higher homo.logues in ambient air are usually below the detection limits of 

the GC-ECD technique. Heuss and Glasson (1968) measured peroxybenzoyl nitrate 

(PBzN) in irradiated auto exhaust samples by GC-ECD and reported that this 

homologue. was 100 times more potent than PAN as a lachrymator. The direct 

analysis of PBzN by.GC-ECD is reported to be complicated by interferences 

(Appel, 1973). Therefore, an analytical technique was developed in which the 

PBzN was quantitatively hydrolyzed to methyl benzoate (MeOBz), followed by GC 

analysis for MeOBz using a flame ionization detector (Appel, 1973). An upper 

limit of 0.07 ppb was placed on the concentration of PBzN in the San Francisco 

bay area. The analysis for the higher homologues of PAN in the atmosphere by 

FTIR spectrometry is not feasible because of inadequate sensitivity, although 

Hanst et al. (1982) placed an upper limit for PBzN in smoggy Los Angeles air 

of 1 ppb based on absorption in the 990 cm 1region. 

The higher homologues of PAN may be prepared for use in calibration in 

the same manner as PAN by the use of a compound contai`ni ng the parent al kyl 

group. Thus, PPN and PBzN have been prepared by the photolysis of alkyl 

nitrates in oxygen (Stephens, 1969) and parent aldehydes plus chlorine and NO 2 

(Gay et al., 1976). The study of Gay et al. (1976) confirmed that the first 

member of the series, peroxyformyl nitrate [HC(0)0 2NO2 1, is too unstable to be 

observed. There have been few reports of the absolute analyses for the higher 

homologues. Infrared absorption analysis of purified samples should be the 

preferred technique. Infrared absorptivities of homologues have been reported 

by Stephens (1969) and Gay et al. (1976). 

4.2,.7Methods for Sampling and Analysis of Hydrogen Peroxide 

4.2.7.1Introduction . Hydrogen peroxide (H 2O2 ) is mechanistically significant 

in photochemical smog as a chain terminator , and as an index of the hydroperoxyl 

radical (H0 2 ) concentration (Bufalini and Brubaker, 1969; Demerjian et al., 

1974). The major reaction leading to the formatiorr of H 2O2 is the recombination of the hydroperoxyl radical (Graedel et al., 1976): 

H02 + H02 + M-> H2O2 + 02 + M (4-16) 

4-43 
Recent studies have implicated atmospheric H2O2 in the aqueous-phase oxidation 

of S02 to SO42 and in the acidification of rain (Penkett et al., 1979; 

Dasgupta, 1980a,b; Martin and Damschen, 1981; Overton and Durham, 1982). 

One of the major problems in assessing the role of atmospheric H 2O 2 has 

been a lack of adequate measurement methodology. Earlier measurements for 

atmospheric H 2O2 were by the titanium colorimetric method (Gay and Bufalini, 

1972a,b; Bufalini et al., 1972) and by a chemiluminescence technique (Kok et 

al., 1978a,b). The reported H 2O 2 concentrations (0.01 to 0.18 ppm) are now 

believed to be far too high, primarily as a result of artifact H 2O 2 formation 

from reactions of absorbed 0 3 (Zika and Saltzman, 1982; Heikes et al., 1982; 

Heikes, 1984). Furthermore, maximum tropospheric 'H 2O 2 concentrations predicted by modeling calculations (Chameides and Tan, 1981; Logan et al., 1981) 

and observed in recent field studies (Das et al., 1983) are on the order of 1 

ppb. In a recent study, a chemiluminescence technique was employed with an 

argon sample purge to remove 03 interference, and a maximum H2O 2 concentration 

of 1.2 ppb was observed in a polluted urban environment (Das et al., 1983). 

Promising techniques that have been used or proposed for aqueous- and gas-phase 

H2O2 are discussed below, as well as methods for sampling, generation, and 

standardization'of 'H2O 2 samples. 

4.2.7.2 Sampling. Almost all of the methods used for the measurement of 

atmospheric H2O2 have used aqueous traps for sampling. Midget impingers (Gay 

and Bufalini, 1972a,b; Kok et al., 1978b; Das et al., 1983), continuous extractors (Kok et al., 1978a), and gas washing traps (Zika and Saltzman, 1982) have 

been used. Aqueous traps have been found to be highly efficient in removing 

trace concentrations of H 2O 2 from gas streams (Zika and Saltzman, 1982). 

Atmospheric 03 , however, which is also absorbed at concentrations much higher 

than H2O2 , reacts in the bul'k aqueous phase and at surfaces to produce H2O 2 

and thus is a serious interference (Zika and Saltzman, 1982; Heikes et~ al., 

1982; Heikes, 1984). Details of the aqueous chemistry of 0 3 can also be found 

in other sources (Hoigne and Bader, 1976; Kilpatrick et al., 1955; Taube and 

Bray, 1940). Because the rate of H 2O 2 production is relatively slow (Heikes, 

1984), the removal of absorbed 03 by purging immediately after sample collection may remove or substantially reduce this interference. This was the 

approach employed by Zika and Saltzman (1982) and by Das et al. (1983). 

Another problem identified with aqueous sampling is that other atmospheric 

species (in particular, S02) may interfere with the generation of H 2O 2 in 

4-44 
aqueous traps and also react with collected H 2O2 to reduce the apparent-H 2O2 

concentration measured (Heikes et al., 1982). 

4.2.7.3 Measurement. A number of inethods for measuring low levels of H 2O2 

have been reported, including the following: 

1. Titanium colorimetric methods. 

2. Chemiluminescence methods. 

3. Enzyme-catalyzed methods. 

4. Laser diode infrared method. 

5. Fourier-transform infrared method (Niki et al., 1980; 

Hanst et al., 1975, 1982). 

6. Electrochemical methods (Pisarevskii et al.,, 1980). 

7. H202-olefi.n reaction (Hauser and Kolar, 1968). 

8. ~ Mixed-ligand complexes (Csanyi, 1981; Meloan et a7., 1961). 

9. Iodometry. 

Of these techniques, only the chemiluminescence and enzyme-catalyzed methods 

are summarized below. A summary of inethods reported in the literature is 

given in Table 4-8. Although titanium colorimetric methods have been applied 

in atmospheric measurements, these techniques are now thought to have inadequate sensitivity for the actual atmospheric concentrations that are present. 

In addition, these techniques have questionable specificity. The same comments 

apply to.methods 6 through.8 and only the primary references are given above 

for those methods. At the present state-of-art, the inevitable presence of 

water vapor absorption limits the use of Fourier-transform infrared methods to 

concentrations above about 0.040 ppm (Hanst et al., 1982). The use of a 

tunable diode infrared laser.source should eliminate the problem associated 

with nearby water bands, and this method is currently under investigation for 

atmospheric measurements (unpublished work in progress, Schiff, 1985). Iodometric techniques are useful only for calibration of H202 standards and will be 

discussed in that section. 

4.2.7.3.1 Chemiluminescence. Hydrogen peroxide in the atmosphere may be 

detected at low concentrations by the chemiluminescence obtained from Cu(II)- 

catalyzed oxidation of luminol (5 - amino- 2,3- dihydro-1,4-phthalazinedione) by 

H2 02 (Armstrong and Humphreys,, 1965; Kok et al., 1978a,b). The reagent is a 

solution containing NaOH (pH = 12.8) and iuminol, 10 - 5 M Cu(II). The products 

of the reaction with H 2 O 2 are 3-amino-phthalic acid, a nitrogen molecule, and 

4-45 
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a photon of light at 450 nm. The detection limit for atmospheric samples was 

given as 0.001 ppm, and the linear dynamic range is 0.001 to 1 ppm. This 

technique as initially employed--5uffered the interferences from 0 3 and S02 

discussed above for aqueous traps. Das et ai. (1982) employed a static version 

of the method of Kok et al. (1978a) to measure H 2O2 concentrations in the 0.01 

to 1 ppb range. In addition, samples were purged with argon immediately after 

collection to eliminate, reportedly, the 0 3 interference. 

Recently, a modified chemiluminescence method has been reported that used 

hemin, a blood component, as a catalyst for the luminol-based H2O2 oxidation 

(Yoshizumi et al., 1984). This method was applied to the measurement of H2O 2 

in rainwater. There was no interference for S0 2 but a significant positive 0 3 

ynterference was reported. 

4.2.7.3.2Enzyme-Catalyzed Methods (Peroxidase) . This general method involves 

three components: a substrate that is oxidizable; the enzyme, horseradish 

peroxidase (HRP); and H 2O2 . The production or decay of the fluorescence intensity of the substrate or reaction product is measured as it is oxidized by 

H2O2 , catalyzed by HRP. Some of the more widely used chromogenic substrates 

have been scopoletin (6-methoxy-7-hydroxy-1,2-benzopyrone) (Andreae, 1955; 

Perschke and Broda, 1961); 3-(p-hydroxyphenyl)propionic acid (HPPA) (Zaitsu 

and Okhura, 1980); and leuco crystal violet (LCV) (Mottola et al., 1970). 

In the scopoletin method, the reagent solution is mixed with a second 

solution containing the H2O2 , the concentration of which must not be less than 

0.33 nor more than 0.84 times the concentration of scopoletin (Perschke and 

Broda, 1961). The disappearance of scopoletin fluorescence is monitored and 

the fluorescence intensity can be used to obtain the concentration of H 2O2 

from a calibratibn curve. 'The most significant advantages of this method of 

analysis are the specificity for H 2O 2 , the sensitivity, and the stoichiometry 

of the scopoletin:H 2O2 reaction (1:1 mole per mole as long as scopoletin is 

present in a 20-fold excess over HRP). The chief disadvantage of the scopoletin 

method i s that the concentrati on of H 2O2 must be wi thi n a narrow range i n 

order to obtain an accurately measurable decrease in fluorescence. This 

limits the usefulness of the technique in determining unknown H 2 O 2 concentrations over several orders of magnitude (Armstrong and Humphreys, 1965; Andreae, 

1955). Detection limits for this technique are quite low and are in the range 

of 1.5 x 10 -11M(Perschke and Broda, 1961) to 2 x 10-10M(Andreae, 1955). 

This technique has been applied to measurements of H 202 in rainwater by Zika 

and Saltzman (1982). 
With the leuco crystal violet (LCV) substrate, intensely colored crystal 

violet is formed from the reaction of H20 2 with LCV, catalyzed by HRP (Mottola 

et al., 1970). The absorbance is measured at 596 nm, where the absorption 

coefficient of crystal violet is 105M 1cm 1 , a very high value and an inherent 

advantage of this method. The concentration of H 2O 2 is a linear function of 

the concentration of crystal violet produced. The detection limit reported is 

about 10-8M H2O2 for an absorbance of 0.005 in a 5-cm cell. 

The HRP catalyzes the oxidation of a wide variety of hydrogen-donating 

substrates by H202 . Zaitsu and Okhura (1980) have tested a number of 4-hydroxyphenyl compounds and found that 3-(p-hydroxyphenyl) propionic acid (HPPA) provided the most sensitive and rapid means for determining H2O2 . When HPPA 

reagent solution is mixed with HRP solution and a test solution containing 

H2O2 , a product is formed that fluoresces at 404 nm following excitation at 

320 nm. The intensity of this fluorescence is monitored as a function of H 2O2 

concentration. The detection limit was reported to be 10 -10mole H2O2 with a 

linear range extending to 10-8mole H2O2 when a test solution of only 0.1 ml 

volume was used. Presumably the molar sensitivity could be improved by the 

use of larger sample volumes. No interference studies were reported. More 

recently, the acetic acid homologue of HPPA has been employed (Kunen et al., 

1983; Dasgupta and Hwang, 1985). The fluorescence intensity of the product 

dimer is directly proportional to H20 2 concentration. 

The enzymatic methods appear to be the most promising aqueous, colorimetric methods for H 2O2 and have considerably greater sensitivity than other 

colorimetric methods. Studies of potential atmospheric interferences, however, 

have not been reported for any of these three substrates. 

4.2.7.4 Generation and Calibration Methods 

4.2.7.4.1 Generation. Standard samples containing trace concentrations of 

H2O2 are required for testing and calibration of various measurement methodologies. As with 0 3 , such standards are not available and are usually prepared 

at the time of use. A number of techniques have been employed for generating 

aqueous standards, but convenient methods for the generation of gas-phase 

standards are noticeably lacking. 

Techniques for the generation of high concentrations of H 2 0 2 have been 

discussed by Shanley (1951). Commercial solutions of 30 percent aqueous H 2 O 2 

are readily available. Trace levels of H2O2 in water may be generated by the 

irradiation of water with 6000 y-radiation (Hochanadel, 1952; Armstrong and 

Humphreys, 1965)- and by enzymatic means (Andreae, 1955). By far the most 

4-48 
convenient method for generating aqueous standards containing micromolar 

concentrations of H 202 is simply the serial dilution of commercial-grade=30 

percent H2O2 (Fisher Analytical Reagent). Samples prepared in this manner 

must be standardized and the method usually employed is the iodometric technique discussed below. Stock standard solutions of H2O 2 as low in concentration as 10-4 M have been found to be stable for many weeks if kept in the dark 

(Armstrong and Humphreys, 1965). 

Techniques for the convenient generation of gas-phase standards are not 

available. With the increased interest in atmospheric H 2O 2 , there is an 

obvious need for an H 2O2 generator comparable to the photolytic 0 3 .gener,ator 

discussed in Section 4.2.4.1. A technique that has often been used for generating ppm concentration levels of H 2O2 in air has been described by Cohen and 

Purcell (1967). Microliter quantities of 30 percent H 2O2 solution are injected 

into a metered stream of air that flows into a Teflon bag.. The concentration 

of H2O2 in the bag is then determined by the iodometric method discussed 

below. 

4.2.7.4.2Calibration . By far the most common method for standardizing low 

concentrations of H 2O2 is based on iodometry (Allen et al., 1952; Hochanadel, 

1952; Cohen et al., 1967). Hydrogen peroxide liberates iodine from an iodide 

solution quite slowly, but in the presence of a molybdate catalyst the reaction 

is quite rapid. The iodine liberated may be determined by titration with 

standard thiosulfate at higher concentrations or by photometric.measurement of 

the tri-iodide ion_ at low concentrations.. The molar absorption coefficient of 

the tri-iodide ion at 350 nm. has been determined to be 2.44 x 104by measuring 

10-5to 10-4M H2O2 solutions prepared from 0.2 M stock H 2O2 solution standardized against primary grade arsenious oxide (Armstrong and Humphreys, 1965). 

The stoichiometry is apparently l mole of iodine released per mole of H 2 02 . 

Definitive studies of the stoichiometry, however, have not been performed to 

the same extent as those of the stoichiometry for the iodometric determination 

of 0 3 . 

4.3 SAMPLING, MEASUREMENT, AND CALIBRATION METHODS FOR PRECURSORS TO OZONE 

AND OTHER PHOTOCHEMICAL OXIDANTS 

During the last decade, a number of advances have been made in the methodology for determining nonmethane organic compounds and oxides of nitrogen. An 

overview of these advances is discussed in the following sections. In the case 

of ineasurement methods for nonmethane organic compounds (NMOCs), early ozone 

4-49 
control models did not require speciation of the complex mixture of organics 

in ambient air. As a result, non-speciation methods were employed for the 

purpose of providing a data base for early photochemical modeling studies. As 

the air quality models grew more sophisticated, however, the need arose for 

more specific information concerning the organic composition of the atmosphere. 

Consequently, methodology was developed to provide for detailed speciation of 

NMOCs. In addition to improving the data base for photochemical modeling, the 

NMOC speciation techniques have also been utilized to characterize various 

sources of pollution (e.g., mobile versus stationary). 

The development of inethodology for oxides of nitrogen has likewise advanced since the original EPA Federal Reference Method for measurement of 

nitrogen dioxide, the Jacobs-Hochheiser technique, was withdrawn in 1973. A 

number of inethods for nitric oxide and nitrogen dioxide have been proposed and 

evaluated since then and are described here. 

4.3.1 Nonmethane Organic Compounds 

Numerous sampling, anaiytical, and calibration methods have been employed 

to determine vapor-phase NMOCs in ambient air. Some of the analytical methods 

utilize detection techniques that are highly selective and sensitive to specific 

functional groups or atoms of a compound (e.g., formyl group of aldehydes, 

halogen), while others respond in a more universal manner; that is, to the 

number of carbon atoms present in the organic molecule. In this overview of 

the most pertinent measurement methods, NMOC have been arranged in three major 

classifications: nonmethane hydrocarbons, aldehydes, and other oxygenated 

compounds. Sampling, analytical, and calibration procedures are discussed for 

each classification. Reference is also made to those analytical methods 

utilized in more than one of the above classifications. 

4.3.1.1 Nonmethane Hydrocarbons. Nonmethane hydrocarbons (NMHC) constitute 

the major portion of NMOC in ambient air (Chapter 3). Traditionally, NMHC 

have been measured by methods that employ a flame ionization detector (FID) as 

the sensing element. This detector was originally developed for gas chromatography and employs a sensitive electrometer that measures a change in ion 

intensity resulting from the combustion of air containing organic compounds. 

Ion formation is essentially proportional to the number of carbon atoms present 

in the organic molecule (Sevcik, 1975). Thus, aliphatic, aromatic, alkenic, 

and acetylenic compounds all respond similarly to give relative responses of 

1.00 ± 0.10 for each carbon atom present in the molecule (e.g., 1 ppm hexane = 

4-50 
6 ppm C; .1 ppm benzene = 6 ppm C; 1 ppm propane = 3 ppm C). Carbon atoms 

bound to oxygen, nitrogen, or halogens give reduced relative responses (Dietz, 

1967). Consequently, the FID, which is primarily used as a hydrocarbon measuring method, should more correctly be viewed as an organic carbon analyzer. 

In the following sections, discussion focuses on the various methods 

utilizing this detector to measure total nonmethane organics. Methods in 

which no compound speciation is obtained are covered first. Methods for 

determining individual organic compounds are then discussed. 

4.3.1.1.1 Non-speciation methods. The EPA reference method for nonmethane 

organic compounds, which was promulgated in.1971, invoives the gas chromatographic separation of inethane (CH4) from the remaining organics in an,air 

sample (F.R., 1971). A second sample is injected directly to the detector 

without methane separation. Subtraction of the first value from the second 

produces a nonmethane organic concentration. 

A number of studies of commercial analyzers employing the Federal Reference Method have been reported (Reckner, 1974; McElroy and Thompson, 1975; 

Harrison et al., 1977; Sexton et al., 1982). In one of the first studies, 

analyses of known synthetic mixtures of hydrocarbons were conducted by 16 

users of the reference method (Reckner, 1974). The nonmethane concentration$ 

tested in this study were 0.23 and 2.90 ppm C. The results shown in Table 4-9 

indicate the percentage error from the two known concentrations. At the 0.23 

ppm level, the majority of the measurements were in error by amounts greater 

than 50 percent. At 2.'90 ppm, most of the measurements were in error by only 

20 percent or iess. 

In general, all of the above studies indicated an overall poor performance 

of the commercial instruments when either calibration or ambient mixtures 

containing NMOC concentrations less than 1 ppm C were used. The major problems 

associated with using these NMOC instruments have been reported in a recent 

technical assistance document (U.S. Environmental Protection Agency, 1981). 

The technical assistance document also suggests ways to reduce the effects of 

existing problems. Table 4-10 summarizes the major problem areas and lists 

recommendations for reducing these effects. 

Other approaches to the measurement of nonmethane organics have also been 

investigated. One such method, developed in 1973, utilizes the fact that CH 4 

requires more heat for combustion than other organics (Poli and Zinn, 1973). 

One portion of the air sample passes through a catalyst bed where all hydrocarbons except CH4are combusted. This sample stream then enters an FID where 

4-51 
TABLE 4-9. PERCENTAGE DIFFERENCE FROM KNOWN CONCENTRATIONS 

OF NONMETHANE HYDROCARBONS OBTAINED BY SIXTEEN USERS 

Known % difference from given concentration 

concentration, 

ppm >100 50 to 100 20 to 50 10 to 20 0 to 10 

0.23 6 4 3 2 1 

2.90 2 0 3 2 9 

Source: Reckner (1974). 

the CH4 concentration alone is recorded. The remaining portion of the sample 

passes directly to a second FID for a total organic carbon measurement. The 

simultaneous2 processing of both signals yields an NMOC value. Although it 

provides a continuous measurement of NMOC levels, this method is also subject 

to many of the same shortcomings attributed to the EPA reference method. 

Recently, a prototype instrument that measures NMOCs by optical absorption 

has been developed (Manos et al., 1982). The unit oxidizes NMOCs to carbon 

dioxide (CO2) and uses a non-dispersive infrared absorption technique to 

measure the organic burden indirectly. Ascarite serves to remove CO 2 initially 

present in air and a hopcalite catalyst selectively oxidizes organics other 

than methane to CO2 and H2O. Since carbon monoxide (CO) will also oxidize to 

CO2 during this process, a dual-channel system is utilized to correct for the 

contribution from ambient CO concentrations. This unit performed well during 

a brief laboratory evaluation using calibration standards; however, more 

extensive laboratory and field tests are needed before the unit can be considered suitable for NMOC measurement. 

Other methods under development and evaluation include oxidation-reduction 

schemes in which nonmethane organics are chromatographically separated from 

methane and non-organic species and then oxidized to CO 2 , reduced to CH 41 and 

detected by FID (U.S. Environmental Protection Agency, 1978b). When organic 

carbon concentrations are greater than 100 ppb, the reduction step can be 

eliminated and a non-dispersive infrared analyzer can be used to detect the CO 2 

formed during the oxidation step (Salo et al., 1975). 

A method for measuring NMOC directly has been reported by Cox et al. 

(1982). This approach involves the cryogenic preconcentration of nonmethane 

organic compounds and the measurement of the revolatilized NMOCs using flame 

4-52 
TABLE 4-10. PROBLEMS ASSOCIATED WITH GATHERING NMOC DATA 

WITH AUTOMATED ANALYZERS AND RECOMMENDATIONS FOR REDUCING THESE EFFECTS 

Problem Areas: 

l. Contaminants may be present in compressed-gas cylinders containing 

calibration gases. 

2. Compressed-gas cylinders of calibration gases sometimes contain the 

standard in a nitrogen or argon background. When no oxygen is blended 

with these gases, FID sensitivity is altered. 

3. The assay of calibration gases contained in compressed-gas cylinders (as 

received from the supplier) is sometimes incorrect. 1 

4. There are wide differences in the per-carbon response to different NMOC 

species. 

5. FID analyzers require hydrogen, which presents a potential- operational 

hazard. 

6. The NMOC concentration is obtained by subtraction of two relatively large 

and nearly equal numbers (TOC-CH 4=NMOC) and thus is subject to large; relative errors. 

7. NMOC analyzers'may exhibit excessive zero and span drift during unattended 

operation. 

8. The complex design of some NMOC analyzers creates unique problems that are 

generally not experienced in other pollutant analyzers. Meticulous set-up, 

calibration, and operation procedures (which are analyzer-specific) are 

difficult to understand and follow. 

Recommendations: 

1. Calibration gases should be checked to determine the concentration of . 

contaminants. , 

2. Calibration concentrations should be obtained by dynamic dilution of a 

pollutant standard with zero-grade air containing oxygen. The d7iution 

ratio should be sufficiently high (ti100:1) to ensure that the calibration 

sample contains 20.9% ± 0.37o oxygen. 

3. A11 calibration standards contained in compressed-gas cylinders should be 

traceable to Standard Reference Materials from the National Bureau of. 

Standards. 

4. The NMOC response should be calibrated to a propane standard. 

5. The operator should use documented procedures for hydrogen safety. 

6. All channels should be properly calibrated. 

7. The FIDs should be operated in accordance with instructions supplied by 

the manufacturer and this document. ' 

8. The training of qualified operators should be augmented with a Technical 

Assistance Document, which provides detailed calibration and operation 

procedures for NMOC analyzers. 

Source: U.S. Enuironmental Protection Agency (1981). 

4-53 
ionization detection. The procedure involves the following steps. A fixed 

volume of sample is drawn through a trap cooled to liquid argon temperature 

(liquid N2 cannot be used since it will also condense methane and oxygen). At 

this temperature all NMOCs are condensed into the trap. After the residual 

CH4 and oxygen are cleared from the trap by the helium carrier gas, the trap 

temperature is raised to revolatilize the NMOC. Oxygen removal reduces the 

variation in response to different organic compounds. 

Jayanty et al. (1982) used a similar system to study the responses of a 

F variety of aliphatic and aromatic compounds. Good linearity was observed with 

the cryogenic trapping procedure (500 ml of air) over a concentration range of 

50 to 5,000 ppb C. Humidity did not generally interfere with the analysis. 

Sample precisions of ±5 percent for single- and multiple-component gas standards 

and ±10 percent for ambient samples were consistently achieved. Responses for 

aromatic compounds, however, were less than expected. The researchers recommended additional testing and instrument refinement in order to resolve this 

problem. 

4.3.1.1.2Speciation methods . The primary separation technique utilized for 

NMOC speciation is gas chromatography (GC). Coupled with flame ionization 

detection, this analytical method permits the separation and identification of 

many of the organic species present in ambient air. 

Compound separation is accomplished by means of both packed and capillary 

GC columns. If high resolution is not required and large sample volumes are 

to.be injected, packed columns are employed. The traditional packed column 

may contain either (1) a solid polymeric adsorbent (gas-solid chromatography) 

or (2) an inert support, coated with a liquid (gas-liquid chromatography). 

Packed columns containing an adsorbent substrate are required to separate 

C2-C3 compounds. The second type of column can be a support-coated or wallcoated open tubular capillary column. The latter column has been widely used 

for environmental analysis because of its superior resolution and broader 

applicability. The wall-coated capillary column consists of a liquid stationary phase coated or bonded (cross-linked) to the specially treated glass or 

fused-silica tubing. Fused-silica tubing is most commonly used because of its . 

physical durability and flexibility. When a complex mixture is introduced 

into a GC column, the carrier gas (mobile phase) moves the sample through the 

packed or coated capillary column (stationary phase). The chromatographic 

process occurs as a result of repeated sorption-desorption of the sample 

4-54 
components (solute) as they move along the stationary phase. Separation 

results from the different affinities that the solute components have for the 

stationary phase. 

The GC-FID technique has been used by numerous researchers to obtain 

ambient NMOC data. Singh (1980) drew on the cumulative experience of these 

researchers to prepare a guidance document for state and local air pollution 

agencies interested in obtaining speciation data. In general, most researchers 

have employed two gas chromatographic units to carry out analyses of NMOC 

species in ambient air. The more volatile organic compounds (C 2 through C5 ) 

are generally measured on one unit using packed-column technology, while the 

other GC separates the less volatile organics using a capillary column. In 

typical chromatograms of urban air, all major peaks are identified and, on a 

mass basis, represent from 65 to 90 percent of the measurable nonmethane 

organic burden. Identification of GC peaks is based upon matching retention 

times of unknown compounds with those of standard mixtures. The use of dedicated computer systems facilitates this task, but close scrutiny of the data 

is still necessary to correct periodic misidentification of unknown compounds 

resulting from variations in retention time. Subsequent verification of the 

individual species is normally accomplished with gas chromatographic/mass 

spectrometric (GC/MS) techniques. Compound-specific detection systems,,such 

as electron capture, flame photometry, and spectroscopic techniques, have also 

been employed for peak identifications. A discussion of these systems, however, 

is beyond the scope of this report. Several documents covering these detection 

systems are available (Lamb et al., 1980; Riggin, 1983). 1 

Because the organic components of the ambient atmosphere are present at 

ppb levels or lower, some means of sample preconcentration is necessary to 

provide sufficient material for the GC-FID system. The two primary techniques 

utilized for this purpose are the use of solid adsorbents and cryogenic collection. The more commonly used sorbent mate.rials are divided into three categories: (1) organic polymeric adsorbents, (2) inorganic adsorbents, and (3) 

carbon adsorbents. Primary organic polymeric adsorbents used for NMOC analyses 

include the materials Tenax GC and XAD-2. These materials have a low retention 

of water vapor and, hence, large volumes of air can be collected. These 

materials do not, however, efficiently capture highly voiatile compounds such 

as C2 to C5 hydrocarbons, nor certain polar compounds such as methanol and 

acetone. Primary inorganic adsorbents are silica gel, alumina, and molecular 

sieves. These materials are more polar than the organic polymeric adsorbents 

4-55 
and are thus more efficient for the collection of the more volatile and polar 

compounds. Unfortunately, water is also efficiently collected, which in many 

instances leads to rapid deactivation of the adsorbent. , Carbon adsorbents are 

less polar than the inorganic adsorbents and, as a result, water adsorption by 

carbon adsorbents is a less significant problem. The carbon-based materials 

also tend to exhibit much stronger adsorption properties than organic polymeric 

adsorbents; thus, lighter-molecular-weight species are more easily retained. 

These same adsorption effects result, however, in irreversible adsorption of 

many compounds. Furthermore, the very high thermal desorption temperatures 

required (350 to 400 ° C) limit their use and also may lead to degradation of 

labile compounds. The commonly available classes of carbon adsorbents include: 

(1) various conventional activated carbons; (2) carbon molecular sieves (Spherocarb, Carbosphere, Carbosieve); and° (3) carbonaceous polymeric adsorbents 

(Ambersorb XE-340, XE-347, XE-348). I 

Although a number of researchers have employed solid adsorbents for the 

characterization of selected organic species in air, only a few attempts have 

been made to identify and quantitate the range of organic compounds from C 2 

and above. Westberg et al. (1980) evaluated several carbon and organic polymeric adsorbents and found that Tenax-GC exhibited good collection and recovery 

efficiencies for >C 6 organics; the remaining adsorbents tested (XAD-4, XE-340) 

were found unacceptable for the lighter organic fraction. The XAD-4 retained 

>C2 organic gases, but it was impossible to completely desorb these species 

without partially decomposing the XAD-4. Good collection and recovery efficiencies were provided by XE-340 only for organics of C4 and above. 

Ogle et al. (1982) used a combination of adsorbents in series and designed 

an automated GC-FID system for analyzing C 2 through C 10 hydrocarbons. Tenax 

GC was utilized for C 6 and above, while Carbosieve S trapped C 3 through C5 

organics. Silica gel followed these adsorbents and effectively removed water 

vapor while passing the C 2 hydrocarbons onto a molecular-sieve 5A adsorbent. 

The combined sorbents have been laboratory-tested with a 38-component hydrocarbon mixture. Good collection and recovery efficiencies were obtained. 

Preliminary field tests have also been successful, but`a very limited data 

base exists. Futhermore, the current chromatographic procedures utilize 

packed-column technology. The addition of capillary columns to this system 

would permit better resolution of the complex mixtures typically found in 

ambient air. 

4-56 
The preferred method for obtaining NMOC data is cryogenic preconcentration (Singh, 1980). Sample preconcentration is accomplished by directing air 

through a packed trap immersed in either liquid oxygen (b.p. -183 ° C) or liquid 

argon (b.p. -186° C). For the detection of about 1 ppb C of an individual 

compound, a 250-cc air sample is normally processed. The collection trap is 

generally filled with deactivated 60I80 mesh glass beads (Westberg et al., 

1974), although coated chromatographic supports.have also been used (Lonneman 

et al., 1974). Both of the above cryogens are suf.ficiently warm to allow air 

to pass completely through the trap, yet cold enough to collect trace organics 

efficiently. This collection procedure also condenses. water.vapor. An air 

volume of 250 cc at 50 percent relative humidity and 25 °C contains approximately 2.5 mg of water, which appears as ice: in the collection trap. The 

collected ice at times will plug the trap and stop the sample flow; furthermore, 

water transferred to the capillary column during the thermal desorption step 

occasionally causes plugging and other deleterious column effects. To circumvent water condensation problems, McClenny et al. (1984) have employed a Nafion 

tube drier to remove water vapor seiectively during the sample collection 

step. These researchers have constructed an automated cryogenic preconcentration sampling and analysis GC system.usi.ng this drier and are currently conducting field evaluaton studies on their system.. 

In addition to direct sampling via preconcentr.ati.on with sorbents and 

cryogenic techniques, collection of whole air samples is frequently used to 

obtain NMOC data. Rigid devices such as syringes, glass bulbs, or metal 

containers and non-rigid devices such as Tedlar and Teflon plastic bags are 

often utilized during sampling. The primary purpose of whole-air collection 

is to store an air sample temporarily:until subsequent laboratory analysis is 

performed. The major problem with this approach is assuring the integrity of 

the sample contents prior to analysis. Of concern is whether sample,components 

of interest are adsorbed or decomposed th.rough interaction with.the container 

walis.or reaction with co-collected gases such as ozone and nitrogen dioxide. 

Sample condensation may also occur at elevated.concentrations or when samples 

are stored under high pressures (i.e., in metal containers). Contamination 

from sampling containers is likewise a frequent occurrence (Lonneman et al., 

1981; Seila et al., 1976). Table 4-11 summarizes the advantages and disadvantages•of the primary collection media for NMOC analysis.. 

4.3.1.1.3 Calibration. Calibration procedures for NMOC instrumentation 

require the generation of dilute mixtures at concentrations expected to be 

4-57 
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•r- •r- •r- li3 Rf O O 7 T3 r 4- +-) OtLf L rl 

S —I tm N N u tn "0 d N O N O H N 

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4-) 

N 

'L3 

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~ f.

co

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~ ~~ 4J ~ O -~ 

t[3 S_ S- `~- 

•~- v 

co 

~ Gl to ~ 4F- +-~ 

~ •r T >

- °-' N >, N to 

+~ v ~ iV +J E r +-) tLi 

o fv d'r- S- N O N • 

i-) tO O N E r C1 r U 'F'> r i~. 

O Rf 

O SO4--) 

r.. .O ~~ tLf i. 

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•rt 

~ ~ s - A a~ o i + > '— O,,.r- O.. (U N . O . tG S-. 

to N d 4-> N N r N C N '"7 "i3 

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.G •~- ~ O ~ G1 r •

r"'

r" .~ ~ r 't 

/'~S 

C G! S-. 'C7 O S. u N N O = E O N C 

to L O C to N X 0.0 N Om O O. 00 • 

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+-) 4- 

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4-59 
found in ambient air. Methods for generating such mixtures are classified as 

static or dynamic systems. 

Static systems are generally preferred for quantitating NMOCs. The most 

commonly used static system is a compressed gas cylinder containing the appropriate concentration of the compound of interest. These cylinder gases may 

also be diluted with hydrocarbon-free air to provide multi-point calibrations. 

Calibration and hydrocarbon-free air cylinders are available commercially. 

Additionally, some standard gases such as propane and benzene are available 

from the National Bureau of Standards (NBS) as certified standard reference 

materials (SRM). Commercial mixtures are generally referenced against these 

NBS standards. In its recent technical assistance document for operating and 

calibrating continuous NMOC analyzers, the U.S. Environmental Protection 

Agency (1981) recommended propane-in-air standards for calibration. Some 

commercially available propane cylinders have been found to'contain other 

hydrocarbons (Cox et al., 1982), so that all calibration data should be referenced to NBS standards. Because of the uniform carbon response of a GC-FID 

system (±10 percent) to hydrocarbons (Dietz, 1967), a common response factor 

is assigned to both identified and unknown compounds obtained from the speciation systems. If these compounds are oxygenated species, an underestimation 

of the actual concentrations will be reported. Dynamic calibration systems 

are employed when better accuracy is needed for these oxygenated hydrocarbon 

species. Dynamic systems are normally employed to generate in situconcentrations of the individual compound of concern and include devices such as permeation and diffusion tubes and syringe delivery systems. 

4.3.1.1.4Comparison of non-speciation versus speciation methods . Speciation 

methods involving cryogenic preconcentration have been compared with non-speciation NMOC analyzers in the following studies. 

Jayanty et al. (1982) conducted a laboratory comparison between the prototype non-speciation method described earlier (Section 4.3.1.1.1) and their 

gas chromatographic separation method. Comparison of the two methods for 12 

ambient air samples collected in stainless steel canisters showed agreement 

within ±15 percent. Ambient air concentrations ranged from 100 to 1000 ppb C. 

Lonneman (1979) compared total NMOC and speciation methods during field 

studies in Houston in 1978. Samples were collected during 3-hour integrated 

time periods (6 to 9 a.m., 1 to 4 p.m.) in Tedlar bags for subsequent analysis. 

The correlation coefficients for 150 measurement pairs from five sites averaged 

0.74. For data pairs of 500 ppb C and less, an average correlation coefficient 

4-60 
(r) of 0.55 was calculated, with a low value of 0.12 at one site. Lonneman 

attributed the low correlations to maintenance and calibration problems in the 

continuous analyzers and concluded that the results from the. non-speciation 

method are "at best marginal for use in photochemical modei applic.ations." 

Holdren et al. (1982) made a similar comparison during a 2-month study at 

urban sites in Cincinnati and Cleveland, Ohio. They utilized a GC/cryogenic 

trapping technique and compared their results with data from state-operated 

NMOC analyzers (SAROAD data). Ambient air samples were collected in Teflon 

bags (6 to 9 a.m. integrated collection) and were transferred immediately to 

pretreated aluminum cylinders for shipment and analysis at the central laboratory. Concentrations of NMOC ranged from 200 to 2100 ppb C. Linear regression analyses of the non-speciated versus speciated data resulted in correlation coefficients that ranged from 0.75 to 0.92 for the four urban sites 

(total of 67 comparisons). Limiting the comparisons to concentrations of 500 

ppb C and lower resulted in an average correlation coefficient of 0.10. 

Richter (1983) compared continuous total'NMOC with GC speciation results 

obtained at seven fixed ground-level sites used in the Northeast Corridor 

Regional Modeling Project (NECRMP). The NMOC data were obtained in real time, 

while Teflon bags were used to collected integrated samples (6 to 9 a.m.) for 

the GC/cryogenic analyses. Over 60 comparisons were available from each site. 

Table 4-12 summarizes statistical information obtained from least-squares 

analysis of the data (Richter, 1983). As the table indicates, only data from 

the East Boston site exhibited a high correlation coefficient. This study 

represents the most extensive effort made yet to compare the two NMOC measuring 

methods. Emphasis was placed on correct instrument operation, calibration, 

etc.; and only verified data were compared. Yet the above results indicate that 

much more work is needed to resolve the differences between the two methods. 

4.3.1.2Aldehydes . Historically, the major problem in measuring concentrations 

of aldehydes in ambient air. has been to find an appropriate monitoring technique 

that is sensitive to low concentrations and specific for the various homologues. 

Early techniques for measuring formaldehyde, the most abundant aldehyde, were 

subject to many interferences and lacked sensitivity at low ppb concentrations. 

The more recently developed techniques can be utilized to accurately measure 

the various types and amounts of aldehydes at ppb levels. This section describes those methods currently used for measuring aldehydes in ambient air. 

4=61 
TABLE 4-12. GC/CONTINUOUS NMOC ANALYZER COMPARISONS, 

LEAST-SQUARES REGRESSIONS 

Location Slope Intercept, ppm C 

Standard 

error r2 

West End Library, 0.552 -0.552 0.672 0.169 

Washington, DC 

Read Street, 0.113 -0.283 0.713 0.0077 

Baltimore, MD 

Essex, MD 0.835 -0.101 0.599 0.354 

Linden, NJ 0.531 NAa 0.865 0.141 

Newark, NJ 0.987 -0.277 0.574 '0.467 

East Boston, MA 1.108 +0.095 0.327 0.887 

Watertown, MA 0.750 -0.568 0.574 0.475 

aData not available. 

Source: Richter (1983). 

These include the chromotropic acid (CA) method for formaldehyde, the 3-methyl2-benzothiazolone hydrazone (MBTH) technique for total aldehydes, Fourier-trans - 

form infrared (FTIR) spectroscopy, and the high-performance liquid chromatography (HPLC) method employing 2,4-dinitrophenylhydrazine (DNPH) derivatization. 

4.3.1.2.1Chromotropic acid method . The chromotropic acid method (CA) involves 

the collection of formaldehyde in a midget impinger containing an aqueous 

mixture of chromotropic and sulfuric acids, followed by the spectrophotometric 

measurement of absorbance of the resulting color (Altshuller and McPherson, 

1963; U.S. Dept, of Health, Education and Welfare, 1965). A modification 

described by Johnson et al. (1981) improved the accuracy and sensitivity of 

the CA method by reducing the concentration of sulfuric acid and by eliminating 

a heating cycle, relying solely on the heat of solution generated by sulfuric 

acid (Altshuller et al., 1961; Olansky and Deming, 1976). Trapping formaldehyde in a 1 percent bisulfite solution before adding the CA solution increased 

collection efficiency from 84 percent to 92 percent with no sulfite interferences. 

4-62 
The CA method for measuring formaldehyde has been widely studied (Salas 

and Singh, 1982; Grosjean'and Kok, 1981; National Research Council, 1981; 

Tuazon et al., 1980; Lioyd, 1979). Originally developed as a spot-test by 

Eegriwe (1937), it was adopted to quantitate formaldehyde spectrophotometrically (Bricker and Johnson,.1945; West and Sen, 1956) and was modified for 

ambient air measurements in the early 1960s (Aitshuller et al., 1961; Altshuller 

and McPherson, 1963; U.S. Dept. of Health, Education, and Welfare, 1965). 

While used widely today for both occupational and ambient air environments, 

its specificity for formaldehyde, which accounts for approximately half of 

total ambient air aldehydes (see Section 3.5), limits its usefulness for 

characterizing aldehyde concentrations in ambient air. 

The CA method has been reported to be sensitive to acrolein, acetaldehyde, 

phenol, nitrogen dioxide, and nitrates (National Research Council, 1981; Krug 

and Hirt, 1977; U.S. Dept. of Health, Education, and Welfare, 1973; Sleva, 

1965; Altshuller et al., 1961). Recent work, however, indicates that neither 

nitrates, nitrites, NO2 , nor acetaldehyde at elevated ambient air levels 

interfere with the CA analysis (Johnson et al., 1981; Grosjean and Kok, 1981). 

Relevant data on other interfering agents were not found. 

4.3.1.2.2 MBTH method. A spectrophotometric technique for total aldehydes 

was developed in the early 1960s by Sawicki et al. (1961). Known as the MBTH 

method, it involves the reaction of aldehyde with 3-methyl -2-benzothi azol one 

hydrazone to form an azine that is oxidized by a ferric chloric-sulfamic acid 

so.lution to form a blue cationic dye (Altshuller, 1983; U.S. Dept. of Health, 

Education, and Welfare, 1965; Hauser and Cummins, 1964; Altshuller and McPherson, 

1963; Altshulier and Leng, 1963; Altshuller et al., 1961). 

The MBTH method has a reported sensitivity of 15 ppb for, primarily, 

low-moiecular-weight aldehydes (National Research Council, 1981). The method 

is subject to interferences by NO2 and gives an inconsistent response to 

higher-molecular-weight aldehydes (Sawicki et al., 1961; Altshuller et al., 

1961). Nonetheless, the Intersociety Committee of the American Public Health 

Association recommends the MBTH colorimetric method for determining total 

aldehydes in air (Intersociety Committee, 1977a). v Miksch and Anthon (1982) 

devised a sampling and analysis scheme that permits a single MBTH sample to be 

used for both formaldehyde and total aliphatic aldehyde determinations. 

4.3.1.2.3 Fourier-transform infrared spectroscopy. Infrared absorption 

spectroscopy has been used to identify and measure aldehydes in ambient air 

(Hanst et al., 1975, 1982; Tuazon et al., 1978, 1980, 1981a). These studies 

 r 

4-63 
employ Fourier-transform infrared (FTIR) spectrometers interfaced to multiple 

reflection cells operating at total optical paths of up to 1 km. At such 

pathlengths, a detection limit of a few ppb is achieved for formaldehyde. The 

advantages of the long-pathlength FTIR method for ambient air aldehyde measurements (i.e., good specificity and direct in situ analysis) are offset by the 

relatively high cost and lack of portability of such systems. 

4.3.1.2.4 High-performance liquid chromatography (HPLC) 2,4-dinitrophenylhydrazine (DNPH) method. This method takes advantage of the reaction of carbonyl 

compounds with 2,4-dinitrophenylhydrazine tb form a 2,4-dinitrophenylhydrazone: 

RR'C=0 + NH 2NHC6H 3 (NO 2 )2 ~ H20 + RR'C=NNHC6H3 (NO2 ) 2 (4-17) 

Since DNPH is a weak nucleophile, the reaction is carried out in the presence 

of acid in order to increase protonation of the carbonyl. 

The HPLC-DNPH method is the preferred way of ineasuring aldehydes in 

ambient air. Atmospheric sampling is usually conducted with micro-impingers 

containing an organic solvent and aqueous, acidified DNPH reagent (Papa and 

Turner, 1972; Katz, 1976; Smith and Drummond, 1979; Fung and Grosjean, 1981). 

After sampling is completed, the hydrazone derivatives are extracted and the 

extract is washed with deionized water to remove the remaining acid and unreacted DNPH reagent. The organic layer is then evaporated to dryness, subsequently dissolved in a small volume of solvent, and analyzed by reversed-phase 

liquid chromatographic techniques employing an ultraviolet (UV) detection 

system. Analysis by a flame ionization detection (FID) system has proved less 

successful than UV because the derivatives are.not really amenable to GC-FID 

analysis. 

An improved procedure has been reported that is much simpler than the 

above aqueous impinger method (Lipari and Swarin, 1982; Kuntz et al., 1980; 

Tanner and Meng, 1984). This scheme utilizes a midget impinger containing an 

acetonitrile solution of DNPH and an acid catalyst. After sampling, an aliquot 

of the original collection solution is directly injected into the liquid 

chromatograph. This approach eliminates the extraction step and several 

sample-handling procedures associated with the DNPH-aqueous solution; and 

provides much better recovery efficiencies. Lipari and Swarin (1982) have 

reported detection limits of 20, 10, 5, and 4 ppb for formaldehyde, acetaldehyde, acrolein, and benzaldehyde, respectively, in 20-liter air samples. 

These researchers have also developed a newer technique employing a solid 

4-64 
adsorbent cartridge containing Florisil coated with 2,4-dinitrophenyldrazine 

(Lipari and Swarin, 1985). Collection efficiencies greater than 90 percent 

were obtained at low ppb levels. A detection limit of 1 ppb of HCHO for a 

100-liter air sample was.achieved. 

4.3.1.2.5 Calibration of aldehyde measurements. Since they ar.e reactive 

compounds, it is extremely difficult to make stabie calibration mixtures of 

aldehydes in pressurized gas cylinders. Although gas-phase aldehyde. standards 

are available commercially, the vendors do not guarantee any level of accuracy. 

Formaldehyde standards are generally prepared by one of several methods. 

The first method utilizes dilute commercial formalin (37 percent formaldehyde, 

w/w). Cal7bration is accomplished by the direct spiking into sampling impingers of the diluted mixture or by evaporation into known test volumes, followed 

by impinger collection. Formaldehyde can also be prepared by heating known 

amounts of paraformaldehyde, passing the effluent gases through a methanolliquid nitrogen slush trap to remove impurities, and collecting the remaining 

f ormal dehyde. Paraformaldehyde permeation tubes have also been used (Tanner 

and Meng, 1984). 

For the higher-molecular-weight aldehydes, liquid solutions can be evaporated or pure aldehyde vapor can be generated in dynamic gas-flow systems 

(permeation tubes, diffusion tubes, syringe delivery systems, etc.,). These 

test atmospheres are then passed through the appropriate aldehyde collection 

system and analyzed. A comparison of these data, with the direct spiking of 

liquid aldehydes into the particular collection system, provides : a measure of 

the overall collection efficiency. 

4.3.1.2.6 Comparison of ineasurement methods. Several side-by-side comparisons 

of the chromotropic acid method (CA) with other methods have been reported. 

Grosjean and Kok (1981) compared the CA method (Johnson et al., 1981) with 

HPLC-DNPH (Fung and Grosjean, 1981) and FTIR spectroscopy (Tuazon et al., 

1978). They found fairly close agreement between the CA and HPLC-DNPH methods, 

but noted consistently higher results with FTIR. Corse (1981) sampled ambient 

air with CA (U.S. Dept. of Health, Educati on,, and Welfare, 1965), MBTH (U.S. 

Dept. of Health, Education, and Welfare, 1965), and HPLC-DNPH methods (Kuntz 

et al., 1980). An examination of tabulated data from the Corse study shows a 

consistent and considerable difference between CA and HPLC measurements. For 

25 CA measurements, formaldehyde averaged 8.8 ppb; while HPLC measurements 

from the same sampling train averaged 14.2 ppb. Overall, formaldehyde.levels 

were approximately 60 percent higher with HPLC than with CA measurements. 

4-65 
Because blanks were not utilized for the HPLC analyses, however, the HPLC data 

are subject to uncertainty, since blank corrections can affect results substantially (Altshuller, 1983a). During laboratory studies, Kuntz et al. (1980) 

reported reasonable agreement (±7 percent) among the HPLC-DNPH, CA, FTIR, and 

GC-FID methods when low ppb levels of formaldehyde, acetaldehyde, propionaldehyde, hexanal, and benzaldehyde were generated. 

4.3.1.3 Other Oxygenated Organic Species. As mentioned earlier, unsubstituted hydrocarbons constitute the major fraction of vapor-phase organic compounds occurring in ambient air. Aldehydes as a class of volatile organics 

appear second in abundance. With the exception of formic acid (Hanst et al., 

1982; Tuazon et al., 1981a, 1980, 1978), other oxygenated species are seldom 

reported. The lack of data for oxygenated hydrocarbons is somewhat surprising 

since significant quantities of these species are emitted directly into the 

atmosphere by solvent-related industries (methanol, ethanol, acetone, etc.; 

see Chapter 3). Along with manmade emissions, natura7 sources of oxygenated 

hydrocarbons also contribute to this total. In addition to these direct 

emissions, it is also expected that photochemical reactions of hydrocarbons 

with oxides of nitrogen, ozone, and hydroxyl radicals will produce significant 

quantities of oxygenated products. 

Difficulties in sample collection and analysis may account for this lack 

of data. The adsorptive nature of the surfaces that contact these oxygenated 

species often complicates the process of compound quantitation. The approach 

used for analysis of oxygenated and other polar organic compounds is to decrease 

adsorption by deactivating the interior surfaces of analytical hardware. A 

novel method has been reported, however, in which the reactive compounds of 

interest are modified rather than the surfaces with which these compounds 

interact (Osman et al., 1979; Westberg et al., 1980). In these studies, the 

laboratory derivatization of vapor-phase alcohols and acids (silylation) was 

investigated to evaluate the potential of such a procedure for stabilizing 

these polar compounds prior to analysis. Results indicate that silylation 

procedures greatly reduce adsorption of alcohols and acids and that, qualitatively, the silylated derivatives can be detected via the GC-FID system. 

4.3.2 Nitrogen Oxides 

In highly polluted urban air, the two most abundant oxides of nitrogen 

compounds are nitric oxide (NO) and nitrogen dioxide (NO 2 ). Both compounds, 

together designated as NO x , participate in the photooxidation reactions in the 

4-66 
atmosphere that lead to the formation of ozone (Chapter 3). Other m7nor 

reactive oxides of nitrogen in ambient air include peroxyacyl nitrates, nitrogen 

trioxide, dinitrogen pentoxide, and peroxynitric acid. In rural areas, and in 

urban areas late in the day and at night, these species may constitute a 

significant fraction of the total oxides of nitrogen in ambient air (Singh and 

Hanst, 1.981; Killus and Whitten, 1985). 

Analytical methods for NO 2 and NO are briefly described in this section. 

Older methods, including the former Federal Reference Method (Jacobs-Hochheiser 

method), are described in the criteria document on nitrogen oxides prepared by 

the U.S. Environmental Protection Agency (1982). 

4.3.2.1Measurement Methods for N0, and N0 . In 1976, the continuous chemiluminescence method was promulgated as the new Federal Reference Method 

(F.R., 1976c). Th7s method measures atmospheric concentrations of NO 2 indirectly 

by first reducing or thermally decomposing the gas quantitat9veiy to NO (with 

a converter), reacting NO with 0 3 , and measuring the light intensity from the 

chemiluminescent reaction (Fontijn et al., 1970). Two types of converters have 

been employed for converting NO 2 to N0. In carbon or ferrous sulfate converters, 

NO2 is chemically reduced to NO with the concurrent oxidation of converter 

material. Thermal converters, such as stainless or molybdenum steel converters, 

spontaneously decompose NO 2 to NO at elevated temperatures. The type and 

severity of interfering species will depend upon the converter used and the 

temperature of operation. The NO in the air stream is measured separately and 

subtracted from the previous NO x (NO plus NO2 ) measurement to yield the NO 2 

concentration. Typical commercial chemiluminescence 7nstruments can detect 

levels as low as 2.5 pg/m3 (0.002 ppm) (Katz, 1976). 

While all oxides of nitrogen and organic nitrogen compounds are thermodynamically unstable with respect to the formation of N0, the rate of the conver- 

~ 

sion to NO is infinitesmally slow under normal conditions. In the presence of 

reducing agents or at elevated temperatures, however, the conversion may 

become quite rapid. A number of studies have shown for example, that some 

species found in ambient air can undergo reduction by converters, resulting in 

positive interferences. Winer et al. (1974), for example, found that peroxyacetyl nitrate (PAN) and various nitrogen compounds are reduced by the converter 

to NO and that nitroethane and nitric acid are partially reduced in the system 

when a carbon (reducing) converter or a molybdenum (thermal decomposition) 

converter is used. Joshi and Bufalini (1978) reported positive interferences 

from halocarbons when a heated carbon converter was used; they also suggested 

4-67 
that stainless steel converters may be subject to interferences from chlorinated 

hydrocarbons. Other evidence suggests that ammonia (NH 3 ) may be converted to 

NO in high-temperature thermal converters (U.S. Environmental Protection 

Agency, 1982). Positive interferences resulting from the presence of PAN and 

HNO3 on afternoons of high oxidant concentrations can exceed 30 percent of the 

NO2 concentrations (Spicer et al., 1979). Grosjean (1982) also reported that 

positive interferences from nitric acid and PAN during NO 2 analysis by chemiluminescence can cause a 50 to 60 percent NO 2 overestimation during smog 

conditions in Los Angeles. In less severe smog, the overestimation should not 

be thi s hi gh. 

Other acceptable methods for measuring ambient NO 2 levels, including two 

methods designated as equivalent methods, are the Lyshkow-modified GriessSaltzman method, the instrumental oolorimetric Griess-Saltzman method, the 

triethanolamine method, the sodium arsenite method, and the TGS-ANSA method 

[TGS-ANSA = triethanolamine, guaiacol (o-methoxyphenol), sodium metabisulfite, 

and 8-anilino-l-naphthalene sulfonic acid]. The sodium arsenite method and 

the TGS-ANSA method were designated as equivalent methods in 1977. While some 

of these methods measure the species of interest directly, others require 

oxidation, reduction, or thermal decomposition of the sample, or separation 

from interferences, before measurement. These methods have been described in 

the 1982 criteria document on nitrogen oxides (U.S. Environmental Protection 

Agency, 1982). 

In addition to the wet chemical methods for measuring NO2 , other techniques have been investigated. Maeda et al. (1980) have reported a new chemiluminescence method based on the reaction of NO 2 with luminol (5-amino-2,3- 

dihydro-1,4-phthalazine dione) ; with a detection limit of about 50 parts per 

trillion (ppt) and linearity over a range of 0.5 ppb to 100 ppm. Work is 

under way by Maeda and coworkers to remove the interferences of 0 3 and S02 . 

Wendel et al. (1983) have reported the development of a luminol-based instrument 

for the continuous measurement of NO 2 in ambient air. In the early instrument 

developed by Maeda et al. (1980), response time was slow and the luminol solution 

pool in the detector cell showed chemiluminescence long after NO 2 was removed. 

The Wendel et al. (1983) instrument has rapid response time and luminol contacts 

the air sample via a filter paper strip rather than a pool of liquid. The 

method has a detection limit of 30 parts per trillion and has been freed of 0 3 

interference through modifications to the inlet system and the addition of 

~ .: 
Na2S03 to the luminol solution. Other workers have endeavored to improve 

chemiluminescence analyzers through physical. modifications (e.g., Ridley and 

Howlett, 1974; Schiff et al., 1979). 

Lipari (1984) has developed a solid-sorbent method for measuring NO2 in 

ambient air. The NO 2 is quantitatively collected by means of a commericially 

available air sampling cartridge containing Florisil (magnesium silicate) 

coated with diphenylamine (DPA). The NO 2 reacts with DPA to form 2-nitro-4- 

nitro-, and N-nitroso-DPA derivatives. The products are eluted from the 

cartridge with methanol, acidified with a 1 N HC1 catalyst to ensure recovery 

of the DPA derivatives, and analyzed by HPLC-UV. No interferences from N0,' 

03 , HNO 3 , S02 and humidity were found, but PAN was found to produce a 50% 

positive interference. Storage stability tests indicated that cartridge 

blanks are stable for at least 3 months and that samples can be transported, 

stored, and analyzed for at least a period of 4 weeks after collection without 

significant sample loss or degradation. The reported limit of detection is 

0.1 ppb NO 2 for a 2000-liter air sample, which corresponds to an 8-hr sampling 

period at a flow rate of 4 liters/min. 

Molecular correlation spectrometry, in which an absorption band of a 

sample is compared with a corresponding band stored in the spectrometer, has 

been applied in analysis of NO 2 (Williams and Kolitz, 1968). Instruments 

processing the second derivative of sample transmissivity have also been used 

(Hagar and Anderson, 1970), as have infrared lasers and infrared spectrometers 

(Hanst, 1970; Hinkley and Kelley, 1971; and Kreuzer and Patel, 1971). Tucker 

et al. (1973, 1975) reported on instruments based on the principle of laserinduced fluorescence at optical frequencies. Fincher et al. (1977) described 

detection of 1 ppb NO 2 with a technique based on fluorescence by a pulsed 

xenon flashlamp. Long-pathlength differential optical absorption spectroscopy 

has also been employed to monitor NO 2 in the troposphere (Platt et al., 1980, 

1984). 

The preferred approach for measuring NO is also the continuous chemiluminescence method. O.ther methods for measuring NO directly include ferrous 

sulfate absorption and spectrophotometric measurement of the resulting ion 

(Norwitz, 1966), ultraviolet spectroscopy (Sweeny et al., 1964), infrared 

spectroscopy (Lord et al., 1975), and ultraviolet fluorescence (Okabe and 

Schwartz, 1975). Mass spectrometry and gas chromatography may also be employed. 

, R. 
4.3.2.2 Sampling Requirements. When sampling for NO X , long residence times 

in sampling lines should be avoided. In the ambient air, the rate of photolysis of NO2 (forming NO and 0, and thus 0 3 ) is almost equal to the rate of the 

reaction of the NO and 0 3 to form NO2 . In sampling lines, photolysis stops 

but NO continues to react with 0 3 , producing NO 2 . The magnitude of the dark 

reaction of NO with 0 3 depends, of course, on the concentrations of NO and 0 3 

in the sample being analyzed, as well as on the residence time of the sample 

in the line. This dark reaction has practical consequences only under certain 

conditions. In moderately polluted urban areas, steady-state concentrations 

of NO are almost certainly too low at the period,of maximal 03 to cause significant errors in obtaining NO x or 03 measurements. Conversely, when 03 is at a 

minimum, as in the early morning or possibly even in the late afternoon, NO 

(and NO2) may be at maximal levels, and no significant errors would be introduced. If the concentrati ons of NO and 0 3 are both 1 ow as i n some rural 

areas, or during those brief periods in polluted areas when NO and 0 3 diurnal 

patterns cross, then measurement errors could be introduced. Such errors are 

not likely to be significant, however. 

Techniques for limiting sampling errors sampling to given levels of 

tolerance are reviewed by Butcher and Ruff (1971). In general, only glass or 

Teflon materials should be used in sampling trains. Among absorbents, granu1es 

impregnated with triethanolamine are reported to be the best, converting only 

2 to 4 percent of the incoming NO 2 to NO (Intersociety Committee, 1977b; 

Huygen, 1970). The most frequently used oxidizer is chromic oxide on a firebrick granule support (Intersociety Committee, 1977b; Levaggi et al., 1974). 

4.3.2.3 Calibration. Calibration procedures for NO X measurements methods are 

critical for obtaining accurate analyses. Measurement methods for NO and NO 2 

are calibrated (1) by sampling a gas stream of known concentration prepared 

from standard reference materials (SRMs); and (2) in the case of chemiluminescence analyzers, by determining converter efficiency. The SRM for NO is a 

cylinder of compressed NO in N2 (50 and 100 ppm). The initial accuracy and 

the stability with time of this mixture were found to be quite good in a study 

conducted at the National Bureau of Standards (Hughes, 1975). An accuracy of 

1 percent of the labeled concentrations was obtained as determined from pressure 

or gravimetric measurements. Concentration changes of less than 1 percent 

were observed over a 7-month period. The SRM for NO 2 is the NO2 permeation 

tube (0'Keeffe and Ortman, 1966; Scaringelli et al., 1970). These'tubes are 

4-70 
caiibrated by determining the permeation rate by means of weight-loss measurements at a constant temperature; or, in some cases, by means of micromanometric 

measurements. Both the NO and NO 2 SRMs are commercially available from sources 

traceable to SRMs maintained at the National Bureau of Standards. 

Procedures for preparing calibrated gas streams call for an accurately 

measured dilution of the output of the NO and NO 2 SRMs. Likewise, the flow 

from the NO-in-N 2 cylinder must be accurately measured, as well as the diluent 

air flow. The flow of diluent air over the NO 2 permeation tube, maintained at 

constant temperature, must also be accurately measured. An alternative procedur.e commonly employed for the calibration of N0, NO 2 , and NOx measurements is 

the use of gas-phase titration (Rehme et al., 1974). A constant flow of 0 3 is 

added to a diluted gas stream of NO at a higher concentration and the reaction 

mixture is monitored with an NO monitor, e.g., chemiluminescence. Because of 

the straightforward bimolecular reaction, 

NO + 03 -> NO2 + 02 (4-18) 

the N0 consumed is equivalent to the 0 3 added and the NO 2 produced. Thus, if 

the concentration of one of the constituents is known, the concentrations of 

the other two are readily determined. The U.S. Environmental Protection 

Agency (1976a,b) recommends the combined use of permeation tubes and,gas-phase 

titration, using one technique to check the othe.r. 

To help ensure the quality of aerometric and NO x exposure data, the U.S. 

Environmental Protection Agency issued a technical assistance document for NO x 

measurement and calibration methods. The information summarized above is 

presented in detail in that document (U.S. Environmental Protection Agency, 

1975). 

4.4 SUMMARY 

4.4.1Sampling and Measurement of Ozone and Other Photochemical Oxidants 

The analysis of ozone and other, reiated atmospheric oxidants includes 

requirements for representative sampling, specific and sensitive measurement 

methodologies, methods for the generation of standard samples, absolute methods 

for the calibration of these standards, and procedures by which to provide 

quality assurance for the whole measurement process. In this summary, recommended procedures are given for all of these analytical steps. Because of the 

4- 71 
large existing data base that employed measurements for "total oxidants," 

non-specific iodometric techniques are discussed and compared to current 

specific 0 3 measurements. 

4.4.1.1Quality Assurance and Sampling . A quality assurance program is 

employed by the U.S. Environmental Protection Agency for assessing the accuracy 

and precision of monitoring data and for maintaining and improving the quality 

of ambient air data. Procedures and operational details have been prescribed 

in each of the following areas: selection of analytical methods and instrumentation (i.e., reference and equivalent methods); method specifications for 

gaseous standards; methods for primary and secondary (transfer standards) 

calibration; instrumental zero and span check requirements, including frequency 

of checks, multiple-point calibration procedures, and preventive and remedial 

maintenance requirements; procedures for air pollution episode monitoring; 

methods for recording and validating data; and information on documenting 

quality control (U.S. Environmental Protection Agency, 1977b). Design criteria 

for 03 monitoring stations, to help ensure the quality of aerometric data, 

have been established (U.S. Environmental Protection Agency, 1977a; National 

Research Council, 1977). 

4.4.1.2Measurement Methods for Total Oxidants and Ozone . Techniques for the 

continuous monitoring of total oxidants and 03 in ambient air are summarized 

in Table 4-13. The earliest methods used for routine1y monitoring oxidants in 

the atmosphere were based on iodometry. When atmospheric oxidants are absorbed 

in potassium iodide (KI) reagent, the iodine produced, 

03 + 3I- + H20 a 13 + 02 + 20H (4-18) 

is measured by ultraviolet absorption in colorimetric instruments and by 

amperometric means in electrochemical instruments. The term "total oxidants" 

is of historical significance only and should not be construed to mean that 

such measurements yield a sum of the concentrations of the oxidants in the 

atmosphere. The various oxidants in the atmosphere react to yield iodine at 

different rates and with different stoichiometries. Only ozone reacts immediately to give a quantitative yield of iodine. As discussed below, the total 

oxidants measurement correlates fairly well with the specific measurement of 

ozone, except during periods when significant nitrogen dioxide (NO 2 ) and 

sulfur dioxide (S0 2 ) interferences are present. The major problem with the 

total oxidants measurement was the effect •of these interferences. Total 

4-72 
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4-73 
oxidants instruments have now been replaced by specific ozone monitors in all 

aerometric networks and in most research laboratories. Biases among and 

between total oxidants and ozone methods are still important, however, for 

evaluating existing data on health and welfare effects levels where concentrations were measured by total oxidants methods. 

The reference method promulgated by EPA for compliance monitoring for 

ozone is the chemiluminescence technique based on the gas-phase ozone-ethylene 

reaction (F.R., 1971). The technique is specific for ozone, the response is a 

linear function of concentration, detection limits of 0.001 to 0.005 ppm are 

readily obtained, and response times are 30 seconds or less. Prescribed 

methods of testing and prescribed performance specifications that a commercial 

analyzer must meet in order to be designated as a reference method or an 

equivalent method have been published by EPA (F.R., 1975b). An analyzer may be 

designated as a reference method if it is based'on the same principle as the 

reference method and meets performance specifiCations. An acceptable equiva-, 

lent method must meet the prescribed performance specifications and also show 

a consistent relation with the reference method. 

. The designated equivalent methods are based on either the gas-solid 

chemiluminescence procedure or the ultraviolet absorption procedure (Table 

4-13). The first designated equivalent method was based on ultraviolet absorption by ozone of the mercury 254 nm emission line. The measurement is in 

principle an absolute one, in that the ozone concentration,can be computed 

from the measured transmission signal since the absorption coefficient and 

pathlength are accurately known. In the gas-solid chemiluminescence analyzer, 

the reaction between ozone and Rhodamine-B adsorbed on activated silica produces chemiluminescence, the intensity of which is directly proportional to 

ozone concentration. 

4.4.1.3 Calibration Methods. A11 the analyzers discussed above must be 

calibrated periodically with ozonized air streams, in which the ozone concentration has been determined by some absolute technique. This includes the 

ultraviolet (UV) absorption analyzer, which, when used for continuous ambient 

monitoring, may experience ozone losses in the inlet system because of contamination. 

A primary ozone calibration system consists of a clean air source, ozone 

generator, sampling manifold, and means for measuring absolute ozone concentration. The ozone generator most often used is a photolytic source employing a 

mercury lamp that irradiates a quartz tube through which clean air flows at a 

4-74 
controlled rate (Hodgeson et al., 1972). Once the output of the generator has 

been calibrated by a primary reference method; it may be used to calibrate 0 3 

transfer standards, which are portable generators, instruments, or other 

devices used to calibrate analyzers in the field. Reference calibration 

procedures that have been used for total oxidants and ozone-specific analyzers 

in the United States are summarized in Table 4-14. 

The original reference calibration procedure promulgated by EPA was the 

1 percent neutral buffered potassium iodide (NBKI) method (F.R., 1971). This 

technique was employed in most of the United States, with the exception of 

Califot^nia. The California Air Resources Board (CARB) (1976) and the Los 

Angeles Air Pollution Control District (LAAPCD) employed different versioris of 

iodometric techniques. Procedural details of the calibration methods are 

summarized in Table 4-14. A number of studies conducted between 1974 and 1978 

revealed several deficiencies with KI methods, the most notable of which were 

poor precision or interlaboratory comparability and a positive bias of NBKI 

measurements relative to simultaneous absolute UV absorption measurements. 

The positive bias observed is peculiar to the .use of phosphate'buffer in the 

NBKI techniques. The bias was not observed in the unbuffered LAAPCD method 

(which nevertheless suffered from poor precision), nor in the 1 percent EPA KI 

method without phosphate buffer (Hodgeson et al., 1977), nor in a KI procedure 

that used boric acid as buffer (Flamm, 1977). A summary of results of these 

prior studies was presented in the previous criteria document (U.S. Environmental Protection Agency, 1978a) and in a review by Burton et al. (1976). 

Correction factors for converting' NBKI calibration data to a UV photometry 

basis are given in Table 4-14 and discussed in Section 4.2.4.2.1. 

Subsequently,. EPA evaluated four alternative reference calibration procedures and selected UV photometry -on the basis of superior accuracy and precision 

and simplicity of use (Rehme et ai., 1981). In 1979 regulations were amended 

to specify UV photometry as the reference calibration procedure (F.R., 1979e). 

Laboratory photometers used as reference systems for absolute 0 3 measurements 

have been described by DeMore and Patapoff (1976) and Bass et al. (1977). 

These laboratory photometers contain long path cells (1 to 5 m) and 

employ sophisticated digital techniques for making effective double beam 

measurements of small absorbancies at low ozone concentrations. A primary 

standard UV photometer is one that meets the requirements and specifications 

given in the revised ozone calibration procedures (F.R., 1979e). Since these 

are currently available in only a few laboratories, EPA has allowed the use of 

4-75 
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4-76 
transfer standards, which are devices or methods that can be calibrated against 

a primary standard and transferred..to another.location for calibration of 0 3 

analyzei^s. Examples of transfer standards that have been used are commercial 

03 photometers, calibrated generators, and gas-phase titration (GPT) apparatus. 

Gui'delines on transfer standards have been published by EPA (McElroy, 1979). 

4.4.1.4~~ Relationships of Total Oxidants and Ozone Measurements . The temporal 

and quantitative relationship between simultaneous total oxidants and ozone 

measurements has been examined in this chapter because of the existence of a 

data base obtained by total oxidants measurements. Such a comparison is complicated by the relative scarcity of simultaneous data, the presence of both 

positive (NO 2 ) and negative (S0 2 ) interferences in total oxidants measurements 

of ambient air, and the change in the basis of calibration. In particular, 

the presence of NO 2 and S02 interferences prevent the establishment of a 

definite quantitative relationship between ozone and oxidants data. Nevertheless, some interesting conclusions can be drawn and are summarized below. 

Studies concluded in the early to mid-1970s were reviewed.,in the previous 

criteria document (U.S. Environmental Protection Agency, 1978a). Averaged data 

showed fairly good qualitative and quantitative agreement.between diurnal 

variations of total oxidants and ozone. For example, uncorrected monthly 

averaged data from Los Angeles and St. Louis showed distinct morning and 

evening peaks resulting from NO 2 interference (Stevens et al., 1972a,b). The 

most recent comparison in the 17terature.involved s;imultaneous ozone and total 

oxidant measurements, in the Los Angeles basin by the Californ7a Air Resources 

Board (1978) i n 1974, 1976, and 1978. The maximum hourly data pai rs were 

correlated (Chock et al., 1982) and.yielded the following regression equation 

for 1978, in which a large number (927) of,data pairs were available: 

Oxidant (ppm) = 0.870 0 3 + 0.005 

(Correlation coefficient = 0.92) (4-19) 

The oxidant data were uncorrected for NO 2 and S0 2 interferences, and on individual days maximum oxidant averages were both higher than and lower than ozone 

averages. 

In summary, specific ozone measurements agree fairly well with total oxidants corrected for NO2 and S02 interfei^ences, and in such corrected total 

oxidants measurements ozone is the dominant contributor. Indeed, it is difficul t to di scern the - presence of -other oxi dan ~ts i n corrected .total oxi dant 

r~Nj~ 
data. Without corrections there can be major temporal discrepancies between 

ozone and oxidants data, which are primarily a result of oxidizing,and reducing 

interferences with KI measurements. As a result of these interferences, on 

any given day the total oxidant values may be higher than or lower than simultaneous ozone data. The measurement of ozone is a more reliable i,ndicator 

than total oxidant measurements of oxidant air quality. 

4.4.1.5 Methods for Sampling and Analysis of Peroxyacetyl Nitrate.and Its 

Homologues. Only two analytical techniques have been used to obtain.signifi- 

•cant data on ambient peroxyacetyl nitrate (PAN) concentrations. These are gas 

chromatography with electron capture detection (GC-ECD) and long-path Fourier 

transform infrared (FTIR) spectrometry. Atmospheric data on PAN concentrations 

have been obtained predominantly by GC-ECD because of its relative simplicity 

and superior sensitivity. These techniques have been described in this chapter 

along with attendant methods of sampling, PAN generation, absolute analysis, 

and calibration. 

By far the most widely used technique for the quantitative determination 

of ppb concentrations of PAN and its homologues is GC-ECD (Darley et al., 

1963; Stephens, 1969). With carbowax or SE30 as a stationary phase, PAN, 

peroxypropionyl nitrate (PPN), peroxybenzoyl nitrate (PBzN), and other homologues (e.g., peroxybutyryl nitrate) are readily separated from components such 

as air, water, and other atmospheric compounds, as well as ethyl nitrate, 

methyl nitrate, and other contaminants that are present in synthetic mixtures. 

El ectron- capture detection provides sensitivities in the ppb and sub-ppb 

ranges. Typically, manual air samples are collected in 50- to 200-m1 ungreased 

glass syringes and purged through the gas-sampling valve. Continuous analyses 

are performed by pumping ambient air through a gas sampling loop of an automatic valve, which periodically injects, the sample onto the column. Samples 

collected from the atmosphere should be analyzed as soon as possible because 

PAN and its homologues undergo thermal decomposition in the gas phase and at 

the surface of containers. The recent work of Singh and Salas (1983a,b) on 

the measurement of PAN in the free (unpolluted) troposphere (see Chapter 5) is 

illustrative of current capabilities for measuring low concentrations. A 

minimum detection limit of 0.010 ppb was obtained. 

The literature contains conflicting reports on the effects of variable 

relative humidity on PAN measurements by GC-ECD. If a moisture,effect is 

4-78 
suspected in a PAN analysis, the bulk of this evidence suggests that humidification of PAN calibration samples (to a range approximating the humidity of 

the samples being analyzed) would be advisable. 

Conventional long-path infrared spectroscopy and Fourier-transform infrared spectroscopy (FTIR) have been used to detect and measure atmospheric 

PAN. Sensitivity is enhanced by the use of FTIR. The most frequently used IR 

bands have been assigned and the absorptivities reported in the literature 

(Stephens, 1964; Bruckmann and Willner, 1983; Holdren and Spicer, 1984) permit 

the quantitative analysis of PAN without calibration standards. The.absorptivity of the 990 cm 1band of PBzN, a higher homologue of PAN, has been reported 

V S'tephens (1969). Tuazon et al. (1978) describes an FTIR system operable at 

pathlengths up to 2 km for ambient measurements of PAN and other trace constituents. This system employed an eight-mirror multiple reflection cell with a 

22.5-m base path. Tuazon et al. (1981b) reported maximum PAN concentrations 

ranging from 6 to 37 ppb over a 5-day smog episode in Claremont, CA. Hanst et 

al. (1982) made measurements with a 1260-m folded optical path system during a 

2-day smog episode in Los Angeles in 1980. An upper limit of 1 ppb of PBzN 

was placed, and the maximum PAN concentration observed was 15 ppb. The.large 

internal surface area of the White cells may serve to promote the decomposition 

or irreversible adsorption of reactive trace species such as PAN: High volume 

sampling rates and inert internal surface materials are used to minimize these 

effects. 

Because of the thermal instability of dilute PAN samples and the explosive 

nature of liquefied PAN, calibration samples are not commercially available 

and must be prepared. Earlier methods used to synthesize PAN have been summarized by Stephens (1969). The photolysis of alkyl nitrites in oxygen was the 

most commonly used procedure and may still be used by some investigators. As 

described by Stephens et al. (1965), the liquefied crude mixture obtained at 

the outlet of the photolysis chamber is purified by preparative-scale GC. 

[CAUTION: Both the liquid crude mixture and the purified PAN samples. are 

violently explosive and should be handled behind explosion shields using 

plastic full-face protection, gloves, and a leather coat at all times.] The 

pure PAN is usually diluted to about 1000 ppm in cylinders pressurized with 

nitrogen to 100 psig and stored at reduced temperatures, <15 ° C. 

Gay et al. (1976) have used the photolysis of chlorine: aldehyde: nitrogen 

dioxide mixtures in air or oxygen for the preparation of PAN and a number of 

its homologues at high yields. This procedure has been utilized in a portable 

4-79 
PAN generator that can be used for the calibration of GC-ECD instruments in 

the field (Grosjean, 1983; Grosjean et al., 1984). 

Several investigators have recently reported on a condensed-phase synthesis 

of PAN by nitration of peracetic acid in a hydrocarbon solvent. High.yields 

are produced of a pure product free of other alkyl nitrates (Hendry,and..Kenley, 

1977; Kravetz et al., 1980; Nielsen et al., 1982; Holdren and Spicer, 1984). 

After the nitration is complete, the hydrocarbon fraction containing PAN 

concentrations of 2 to 4 mg/ml can be stored at -20 ° C for periods longer than 

a year (Holdren and Spicer, 1984). 

The most direct method for absolute analysis of these PAN sampl,es is by 

infrared absorption, using the IR absorptivities mentioned earlier. Conyentional IR instruments and 10-cm gas cells can analyze gas standards of concentrations >35 ppm. Liquid microcells can be used for the analysis of PAN in 

isooctane solutions. The alkaline hydrolysis of PAN to acetate ion and nitrite 

ion in quantitative ,yield (Nicksic et al., 1967) provides a means independent 

of infrared for the quantitative analysis of PAN. Following hydrolysis, 

nitrite ion may be quantitatively analyzed by the Saltzman colorimetric procedure (Stephens, 1969). The favored procedures now use ion chromatography to 

analyze for nitrite (Nielsen et al., 1982) or acetate (Grosjean, 1983; Grosjean 

et al., 1984) ions. Another calibration procedure has been proposed that is 

based on the thermal decomposition of PAN in the presence of excess and measured 

NO concentrations (Lonneman et al., 1982). The acetylperoxy radical, CH 3C(0)021 

and its decomposition products rapidly oxidize nitric oxide (NO) to NO 2 with a 

stoichiometry that has been experimentally determined. 

4.4.1.6 Methods for Sampling and Analysis of Hydrogen Peroxide. Hydrogen 

peroxide (H2O2) is significant in photochemical smog as a chain terminator; as 

an index of the hydroperoxyl radical (H0 2 ) concentration (Bufalini and Brubaker, 

1969; Demerjian et al., 1974); and as a reactant in the aqueous-phase oxidation 

of S02 to SO4-2 and in the acidification of rain (Penkett et al., 1979; Dasgupta, 

1980a,b; Martin and Damschen, 1981; Overton and Durham, 1982). 

One of the major problems, however, in assessing the role of atmospheric 

H2O2 has been a lack of adequate measurement methodology. Earlier measurements 

(Gay and Bufalini, 1972a,b; Bufalini et al., 1972; Kok et a1., 1978a,b) reporting 

H2O2 concentrations of 0.01 to 0.18 ppm are now believed to be far too high, 

and to be the result of artifact H 2O2 formation from reactions of absorbed 0 3 

(Zika and Saltzman, 1982; Heikes et al., 1982; Heikes, 1984). Maximum tropospheric H2O2 concentrations predicted by modeling calculations (Chameides and 

sm 
Tan;'`1981; Logan et al., 1981) and observed in recent fiEld studies (Das et 

al., 1983) are on the order of 1 ppb. 

Almost all of the methods used for the measurement of atmospheric H 202 

have used aqueous traps for sampling. Atmospheric 0 3 , however, which is also 

absorbed at concentrations much higher than H 2O2 , reacts in the bulk aqueous 

phase and at surfaces to produce H 2O2 and thus is a serious interference (Zika 

and Saltzman, 1982; Heikes et al., 1982; Heikes, 1984). The removal of°absorbed 

03 by purging immediately after sample collection may remove or substantially 

reduce this interference (Zika and Saltzman, 1982; Das et al., 1983). Another 

problem identified with aqueous sampling is that other atmospheric species (in 

particular, S02 ) may interfere with the generation of H2O2 in aqueous,traps 

and also react with collected H 2O2 to reduce the apparent concentration measured 

(Heikes et al., 1982). 

Of the techniques that have been used for the measurement of aqueous and 

gas-phase H 2O2 , only the chemiluminescence and enzyme-catalyzed methods are 

summarized below. The other techniques are now believed to have inadequate 

sensitivity and specificity for the atmospheric concentrations actually present. 

In addition, the use of a tunable diode infrared laser source should eliminate 

the problem associated with nearby water bands, and this method is currently 

under investigation for atmospheric measurements (unpublished work in progress, 

Schiff, 1985). 

Hydrogen peroxide in the atmosphere may be detected at 1ow concentrations 

by the chemiluminescence obtained from copper(II)-catalyzed oxidaton of luminol 

(5-amino-2,3-dihydro-1,4-phthalazinedione) by H20

2 (Armstrong and Humphreys, 

1955; Kok et al., 1978a,b). Th7s technique as initially employed suffered the 

interferences from 0 3 and S02 discussed above for aqueous traps. Das et al. 

(1982) employed a static version of the.method of Kok et al. (1978a) to measure 

H2O2 concentrations in the 0.01 to 1 ppb range. In addition, samples were 

purged with argon immediately after collection to eliminate, reportedly, the 

03 interference. Recently, a modified chemiluminescence method has been 

reported which used hemin, a blood component, as a catalyst for the luminolbased H2O2 oxidation (Yoshizumi et a1., 1984). 

The most promising chemical approach employs the catalytic activity of 

the enzyme horseradish peroxidase (HRP) on the oxidation of organic substrates 

by H2O2 . The production or decay of the fluorescence intensity of the substrate 

or reaction product is measured as it is oxidized by H2O2 , catalyzed by HRP. 

r~ : 1 
Some of the more widely used substrates have been scopoletin (6-met,hoxy-7- 

hydroxyl,2-benzopyrone) (Andreae, 1955; Perschke and Broda, 1961); 3-(p-hydroxyphenyl)propionic acid (HPPA) (Zaitsu and Okhura, 1980); and leuco crystal 

violet (LCV) (Mottola et al., 1970). 

The decrease in the fluorescence intensity of scopoletin is measured as a 

function of H 0 concentration. Detection limits have been reported to be 

quite low (10?H M). The chief disadvantage of this approach is that the 

concentration of H202 must be within a narrow range to obtain an accurately 

measureable decrease in fluorescence. Oxidation of LCV produces intensely 

colored crystal violet, which has a molar absorption coefficient of 10 5 M 1 

cm-1 at the analytical wavelength, 596 nm. The detection limit reported was 

10-8 M in 5 cm cells. Two quite similar hydrogen donor substrates have been 

used. Zaitsu and Okhura (1980) employed 3-(p-hydroxyphenyl) propionic acid 

and more recently the p-hydroxyphenyl acetic acid homologue is being used 

(Kunen et al., 1983; Dasgupta and Hwang, 1985). The measurement of the fluorescence intensity of the product dimer provides a quite sensitive means for the 

assay of H2O2 . 

As with 03 , H202 calibration standards are not commercially available and 

are usually prepared at the time of use. The most convenient method for preparing aqueous samples containing micromolar concentrations of H 2O 2 is simply 

the serial di1ution of commercia1 grade 30 percent H 2O2 (Fisher Analytical 

Reagent). Techniques for the convenient generation of gas-phase standards are 

not available. A technique often used for generating ppm concentrations of 

H2O2 in air involves the injection of microliter quantities of 30 percent H202 

solution into a metered stream of air that flows into a Teflon bag. Aqueous 

and gas-phase samples are then standardized by conventional iodometric procedures (Allen et al., 1952; Cohen et al., 1967). 

4.4.2 Measurement of Precursors to Ozone and Other Photochemical Oxidants 

4.4.2.1 Nonmethane Organic Compounds. Numerous analytical methods have been 

employed to determine nonmethane organic compounds (NMOC) in ambient air. 

Measurement methods for the organic species may be grouped according to three 

major classifications: nonmethane hydrocarbons, aldehydes, and other oxygenated 

compounds. 

Nonmethane hydrocarbons have been determined primarily by methods that 

employ a flame ionization detector (FID) as the sensing element. Early methods 

for the measurement of total nonmethane hydrocarbons did not provide for 

4-82 
speciation of the complex mixture of organics in ambient air. These methods, 

still in use for analysis of total nonmethane organic compounds, are essentially 

organic carbon analyzers, since the response of the FID detector is essent7ally 

proportional to the number of carbon atoms present in the organic molecule 

(Sevcik, 1975). Carbon atoms bound, however, to oxygen, nitrogen, or halogens 

give - reduced relative responses (Dietz, 1967). The FID detector has been 

utilized both as a stand-alone continuous detection system (non-speciation 

analyzers have indicated an overall poor performance of the commercial instruments when calibration or ambient mixtures containing nonmethane organic 

compounds (NMOC) concentrations less than 1 ppm C were analyzed (e.g., Reckner, 

1974; McElroy and Thompson, 1975; Sexton et al., 1982). The major problems 

associated with the non-speciation analyzers have been summarized in a recent 

technical assistance document published by the U.S. Environmental Protectipn 

Agency (1981). The document also presents ways to reduce some of the existing 

problems. 

Because of the above deficiencies, other approaches to the measurement of 

nonmethane hydrocarbons are currently under development. The use of gas 

chromatography coupled to an FID system circumvents many of the problems 

associated with continuous non-speciation analyzers. This method, however, 

requires sample preconcentration because the organic components are present et 

part-per-billion (ppb) levels or lower in ambient air. The two main preconcentration techniques in present use are cryogenic. collection and the use of 

solid adsorbents (McClenny et al., 1984; Jayanty et al., 1982; Westberg. 

et al., 1980; Ogle et al., 1982). The preferred preconcentration method for 

obtaining speciated data is cryogenic collection. Speciation methods involving 

cryogenic preconcentrationhave a7so been compared with continuous nonspeciation 

analyzers (e.g., Richter, 1983). Results indicate poor correlation between 

methods at ambient concentrations below 1 part-per-million carbon (ppm C). 

Aldehydes, which are both primary and secondary pollutants in ambient 

air, are detected by total NMOC and NMHC speciation methods but can not be 

quantitatively determined by those methods. Primary measurement techniques 

for aldehydes include the chromotropic acid (CA) method for formaldehyde 

(Altshuller and McPherson, 1963; Johnson et al., 1981), the 3-methyl-2-benzothiazolene (MBTH) technique for total aldehydes (e.g., Sawicki et al., 1961; 

Hauser and Cummins, 1964), Fourier-transform infrared (FTIR) spectroscopy 

4-83 
(e.g., Hanst et al., 1982; Tuazon et al., 1978, 1980, 1981a), and high-`performance liquid chromatography employing 2,4-dinitrophenyl-hydrazine derivatization (HPLC-DNPH) for aldehyde speciation (e.g., Lipari and Swarin, 1982; Kuntz 

et al., 1980). The CA and MBTH methods utilize wet chemical procedures and 

spectrophotometric detection. Interferences from other compounds have.been 

reported for both techniques. The FTIR method offers good specificity and 

direct in situ analysis of ambient air. These advantages are offset, however, 

by the relatively high cost and lack of portability of the instrumentation. 

On the other hand, the HPLC-DNPH method not only offers good specificity but 

can also be easily transported to field sites. A few intercomparison studies 

of the above methods have been conducted and considerable differences in 

measured concentrations were found. The data base is still quite limited at 

present, however, and further intercomparisons are needed. 

Literature reports describing the vapor-phase organic composition of 

ambient air indicate that the major fraction of material consists of unsubstituted hydrocarbons and aldehydes. With the exception of formic acid, other 

oxygenated species are seldom reported. The lack of oxygenated hydrocarbon 

data is somewhat surprising since significant quantities of these species are 

emi-tted into the atmosphere by solvent-related industries and since at least 

some oxygenated species appear to be emitted by vegetation. In addition to 

direct emissions, it is also expected that photochemical reactions of hydrocarbons with oxides of nitrogen, 03 , and hydroxyl radicals will produce significant quantities of oxygenated species. Difficulties during sampie collection 

and analysis may account for the apparent lack of data. Attempts have been 

made to decrease adsorption by deactivating the reactive surface or by modifying 

the compound of interest (Osman et al., 1979; Wostberg et al., 1980). Additional 

research efforts should focus on this area. 

4.4.2.2 Nitrogen Oxides. Aside from the essentially unreactive nitrous oxide 

(N20), only two oxides of nitrogen occur in ambient air at appreciable concen- 

: trations: nitric oxide (NO) and nitrogen dioxide (NO 2 ). Both compounds, 

together designated as NO x , participate in the cyclic reactions in the atmosphere 

that lead to the formation of ozone. Other minor reactive oxides of nitrogen 

in ambient air include peroxyacyl nitrates, nitrogen trioxide, dinitrogen 

pentoxide, and peroxynitric acid. 

The preferred means (Federal Reference Method) of ineasuring NO and NO 2 is 

the chemiluminescence method (F.R., 1976c). The measurement principle is the 

gas-phase chemiluminescent reaction of 0 3 and NO (Fontijn et al., 1970). 

4-84 
Whi1e;:N0 is determined directly in this fashion, NO 2 is detected indirectly by 

first` reducing or thermally decomposing the gas quantitatively to NO with a 

converter. The reaction of NO and 03 forms excited NO 2 molecules that release 

light energy that is proportional to the NO concentration. Although the NO 

chemiluminescence is interference-free, other nitrogen compounds do interfere 

wherr:directed through the NO 2 converter. The magnitude of these interferences 

is dependent upon the type of converter used (Winer et al., 1974; Joshi and 

Bufalini, 1978). The detection limit of commercial chemiluminescence instruments for NO 2 measurement is 2.5 pg/m 3 (0.002 ppm) (Katz, 1976). 

Development of an instrument based on the chemiluminescent reaction of 

NO2 with luminol (5-amino-2, 3-dihydro-1, 4-phthalazine dione) has been reported 

by Maeda et al. (1980). Wendel et ai. (1983) have reported modifications of 

this luminol-based method in which better response time and less interference 

from 0 3 have been achieved. 

Other acceptable methods for measuring ambient NO 2 levels, including two 

methods designated as equivalent methods, are the Lyshkow-modified GriessSaltzman method, the instrumental colorimetric uriess-Saltzman method, the 

triethanolamine method, the sodium arsenite method, and the TGS-ANSA method 

[TGS-ANSA = triethanolamine, guaiacol (o-methoxyphenol), sodium,metabisulfite, 

and 8-anilino-l-naphthalene sulfonic acid]. The sodium arsenite method,and 

the TGS-ANSA method were designated as equivalent methods in'1977. Eor descriptions of these methods, the reader is referred to the EPA criteria docWment 

for nitrogen oxides (U.S. Environmental Protection Agency, 1982). While some 

of these methods measure the species of interest directly, others requi,re 

oxidation, reduction, or thermal decomposition of the sample, or separation 

from interferences, before measurement. None of these other techniques, 

however, is widely used to monitor air quality. 

Careful adherence to specified calibration procedures is essential for 

obtaining accurate NO x measurements. The U.S. Environmental Protection Agency 

(1975) has issued a technical assistance document that describes in detail the 

two acceptable calibration methods for NO x : (1) the use of standard reference 

materials (SRMs) and (2) gas-phase titration (GPT) of NO with 0 3 . The SRM for 

NO i s a cyl i nder of compressed NO i n N2 ; the mi xture i s both accurate and 

stable (Hughes, 1975). The SRM for NO 2 is the NO2 permeation tube (0'Keeffe 

and Ortman, 1966; Scaringelli et al., 1970). The gas-phase titration, described 

by Rehme et al. (1974), is based upon the bimolecular reaction of NO with 0 3 

to form NO 2 plus 02 . The U.S. Environmental Protection Agency (1975) recommends 

4-85 
the combined use of GPT plus SRM procedures, using one technique to check the 

other. 

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13: 975-985. 

Spicer, C. W.; Holdren, M. W.; Keigley, G. W. (1983) The ubiquity of peroxyacetyl nitrate in the continental boundary layer. Atmos. Environ. 17: 

1055-1058. 

:0 t~I~I 
Stedman, D. H.; Kok, G.; Delumyea, R.; Alvord, H. H. (1976) Redundan-t. ~ calibration of nitric oxide, carbon monoxide, nitrogen dioxide, and ozone air 

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4-102 
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0 

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4-105 

5. CONCENTRATIONS OF OZONE AND - OTHER PHOTOCHEMICAL OXIDANTS IN AMBIENT AIR 

5.1 INTRODUCTION 

The data presented in this chapter on the concentrations of ozone and 

other photochemical oxidants in ambient air are intended to support and complement information presented in subsequent chapters on the effects of these 

compounds. Thus, this chapter describes potential exposures of human populations, crops and other vegetation, ecosystems, and nonbiological materials in 

general terms for the entire nation and in specific terms for selected areas 

of the country. Since the health and welfare effects of ozone have been much 

more thoroughly documented than those of other related oxidants, primary 

emphasis in this chapter is placed on the concentrations of ozone found in 

ambient air. Potential exposures are described by presenting data on peak and 

average concentrations nationwide and on seasonal and diurnal patterns in 

selected urban and nonurban areas. The recurrence on consecutive days of 

selected levels of ozone has been examined for a few urban sites to aid in 

understanding the significance of health effects documented in subsequent 

chapters. Likewise, data have been included that portray representative urban 

and rural concentrations by season and by time of day. Spatial variations in 

ozone concentrations are briefly addressed since the.effects on concentrations 

of latitude, altitude, intracity variations, and indoor-outdoor gradients are 

pertinent to the assessment of potential exposures of human populations, and, 

except for the indoor-outdoor gradients, of crops and other vegetation and 

ecosystems. 

Ozone is the only photochemical oxidant other than nitrogen dioxide that 

is routinely monitored and for which a comprehensive aerometric data base 

exists. Data for peroxyacetyl nitrate (PAN) and its homologues and for hydrogen peroxide (H2O 2 ) and formic acid (HCOOH) have all been obtained as part of 

special research investigations. Consequently, no data on nationwide patterns 

of occurrence are available for these non-ozone oxidants; nor are extensive 

data available on the correlations of levels and patterns of these oxidants 

with those of ozone. Data on these oxidants are considerably more abundant 

now, however, than in 1978, when the previous criteria document for ozone and 

other photochemical oxidants was published (U.S. Environmental Protection 

Agency, 1978). Sections 5.6 and 5.7 summarize the available data on these 

other oxidants. 

5-1 
The concentrations of ozone and related photochemical oxidants observed 

in ambient air are the net result, as shown in Chapter 3, of various combinations of a variety of atmospheric processes, including: 

1. Local photochemical production from oxides of nitrogen and 

reactive volatile organic compounds. 

2. Atmospheric mixing after sunrise, such that ozone-rich air from 

above the nocturnal inversion layer is mixed to the surface, 

resulting in a steady increase in ozone during the morning and 

early afternoon. This increase, from ozone preserved aloft, is 

independent of photochemistry, occurring even in the absence of 

precursors and photochemical processes. 

3. Transport of ozone produced photochemically but not locally. 

4. Intrusion into the troposphere, even to ground level, of ozonerich air from the stratospheric reservoir. 

5. Formation of ozone photochemically in the mid-troposphere, with 

subsequent intrusion into the boundary layer. 

6. Chemical scavenging in the atmosphere of ozone and other oxidants; e.g., the reaction of ozone with nitric oxide (NO) or the 

reaction of H2O 2 with sulfur dioxide (S02 ). 

7. Physical scavenging in the atmosphere of ozone and other oxidants; e.g., the temperature-dependent decomposition of PAN, 

the precipitation scavenging of H 2O 2 , and the photolytic dissociation of ozone. 

8. Combined physical and chemical scavenging processe5 at the 

surface of the earth; e.g., the deposition of ozone on reactive 

biological or nonbiological surfaces, such as vegetation, 

soils, .or certain polymers. 

These processes include, obviously, both manmade and natural processes 

and driving mechanisms. The occurrence of high ozone concentrations is most 

commonly associated with recognized meteorological conditions that involve 

intense sunlight and elevated temperatures, and the variety of processes 

involved contribute to strong diurnal cycles. Peak concentrations have been 

observed to occur, however, at almost any time of day. Ozone may be transported after its formation for distances up to 1000 km or more (e.g., Hov et 

al., 1978; Wolff and Lioy, 1980). As a result, high concentrations of ozone 

and related oxidants occur not only near large sources of precursors but a-lso 

in downwind nonurban areas, and usually later in the day at these downwind 

receptor sites. Ozone, and apparently PAN, as well, can be transported at 

5-2 
night above the surface pollutant and nocturnal inversion layer (Chapter 3), 

but during daylight hours can be transported considerable distances at or near 

ground level (Coffey and Stasiuk; 1975a,b). 

The emphasis of this chapter is on the documentation of concentrations 

rather than on explanations of possible causes of observed concentrations. 

Probable causes and explanations are mentioned, however, where they are pertinent to the discussion. 

Most of the data presented in this chapter to characterize both nationwide 

and site-specific ozone concentrations in ambient air were obtained after 

1978. Two factors influenced the use of post-1978 data. First, the current 

Federal Reference Method for ozone, chemiiuminescence, and the equivalent UV 

method were almost universally-employed by 1979. Second, EPA promulgated a UV 

calibration method for ozone in 1979. Thus, these data form a relatively 

homogeneous set for purposes of intercomparison. Because of the well-recognized difficulties in converting from older data sets to the current reference 

method, few pre-1979 aerometric data for ozone are presented in this chapter, 

and then only to demonstrate spatial variations in concentrations that are 

pertinent to exposure assessments. 

5.2 TRENDS IN NATIONWIDE OZONE CONCENTRATIONS 

Whether ozone concentrations in ambient air are static, rising, or declining 

over time must be determined from statistical tests using comparable aerometric 

data for a number of years. The national trend in ozone concentrations is 

shown in Figure 5-1 for the 9-year period, 1975 through 1983 (U.S. Environmental 

Protection Agency, 1984a). The concentrations depicted are the average 

second-highest 1-hour concentrations for selected stations for each year. 

In this context, the second-highest 1-hour value for each station is selected 

from all daily maximum 1-hour values (n < 365) recorded per year at that station. 

Subsequently, this statistic will be referred to as the second-highest daily 

maximum 1-hour value or simply the second-highest 1-hour value. The 176 

monitoring stations included in this analysis reported at least 50 percent of 

the possible hourly values for the ozone season in at ieast 7 of these 9 years. 

(The ozone season varies at the respective sites from 4 to 12 months, depending 

upon local conditions. The sampling period is specified in State Implementation 

Plans. ) 

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Figure 5-1. National trend in composite average of the second highest 

value among daily maximum 1-hour ozone concentrations at selected 

groups of sites, 1975 through 1983. - 

Source: U.S. Environmental Protection Agency (1984a). 

5-4 
Ambient air monitoring stations in the nation are generally operated by 

state and local agencies, but among the stations is a small group of National 

Air Monitoring Stations (NAMS) whose data are reported directly to EPA. The 

trend li.ne for the subset of the 62 NAMS ozone stations, also shown in Figure 

5-1, tracks fairly closely the trend for all 176 stations. To permit a comparison of aerometric data for California, a high-oxidant area, with the 

nationwide trends data, data for California and data for all states other than 

California are also plotted separately in Figure 5-1 and are compared below. 

For the entire 9-year period, 1975 through 1983, all subsets of monitoring 

stations show a decline in the composite second-highest daily maximum 1-hour 

ozone concentration. In 1979, a new, more accurate ozone calibration procedure 

was promulgated by the U.S. Environmental Protection Agency (see Chapter 4). 

Between 1979 and 1982, a small decline of 9 to 10 percent in nationwide ozone 

concentrations occurred. From 1982 to 1983, however, concentrations increased 

by about 10 percent in California, by about 12 percent nationwide, and by,about 

14 percent outside California. Recently published data for 1984 from a somewhat 

smaller number of stations (163 sites), not depicted in Figure 5-1, show a 

decrease in second-highest 1-hour concentrations nationwide, with ozone levels 

in 1984 approximating those recorded in 1981, which are shown in Figure 5-1 

(U. S. Environmental Protection Agency, 1986). 

Note the influence of.ozone data from Calfornia. The California data 

heavily influence the trends data because of the number of California monitoring sites represented in the data and because of the actual concentrations of 

ozone in California. 

Evaluation of national trends, as well as local or regional trends, in 

reported concentrations of ozone in ambient air over the past 5 to 10 years is 

complicated by a number of factors, including: (1) the change in calibration 

procedure recommended by EPA in 1978 and promulgated in 1979 (see Chapter 4); 

(2) the possible effects on aerometric data of quality assurance procedures 

instituted by EPA in 1979; (3) the influence of diverse regional meteorological 

conditions; and (4) changes in precursor emissions. 

How much of the observed decline in ozone concentrations from 1975 through 

1982 is attributable to the 1979 promulgation of the ultraviolet (UV) calibration method as the Federal Reference Method is uncertain. To determine that, 

the monitoring practices at each of the 176 sites would have to be examined in 

detail to find out the calibration methods used, the date when the UV method 
was adopted for this site, and whether states applied correction factors to 

prior data (see, e.g., Walker, 1985; Hoggan, 1986; Walker, 1986). Such an 

examination would be necessary because not all monitoring sites switched to 

the use of the UV method simultaneously. For example, the state of California 

(in EPA Region IX) had already been using the UV method before'it was 

promulgated in February 1979. In addition,' other states, in other regions, 

may have used the boric acid-potassium iodide (BAKI) method before, after, or 

both before and after promulgation of the UV method, since the'BAKI procedure 

was allowed by EPA as an interim method for 18 months followirig'the 1979 UV 

•promulgation (see Chapter 4). Likewi`se, ot'her states used gas-phase titration 

prior° to 1979 but eithei^ BAKI or UV procedures following the UV promulgation. 

In addition, random errors associated with individual operator practices 

occur. Thus, the relationships among these three methods', even if monitoring 

practices at individual sites were known, are,complex and would preclude the 

simple application of a single correction factor to'these trend data (see 

Chapter 4). 

Hunt and Curran (1982) have noted that , only EPA Region IX showed improvement in ozone air quality between 1979 and 1981 but, in contrast to other EPA 

regions, showed no improvement in that period versus the 1975 to 1979 period. 

Since California, which dominates Region IX, changed its calibration method in 

1975, the decrease in ozone concentrations seen in Ca1ifornia from 1979 through 

1982 (Figure 5-1) cannot be attributed to the introduction''of the UV , cal i bration method. The shape of the trend line for California is quite similar to 

that for the rest of the nation oVer the°4-year period (1980 through 1983) 

fol'lowing the promulgation of the UV method. From 1982 to 1983, an increase 

in ozone concentrations occurred nationwide, with the percentage increase in 

California roughly paralleling that for the rest of the nation. Increased 

precursor emissions and meteorological conditions conducive to oxidant formation appear to be the most likely causes of the increase in 1983 in the composite average of the second-highest daily maximum l-hour ozone concentrations. 

An examination of emissions` data, however; indicates that NO xemissions did 

not change significantly from 1982 to 1983 and that VOC emissions rose only 3 

percent in that period (U:S. Environmental Protection Agency, 1984b; also see 

Chapter 3). Therefore, precursor emissions are not thought to accou'nt for all 

of the 12 percent increase in 03firbm 1982 to 1983 nationwide (U.S. Environmental Protection Agency, 1984a). 

5-6 
The influence of ineteorological conditions on ozone concentrations has 

been explored for 10 cities using an.experimental index based on meteorological 

parameters conducive to ozone formation (U.S. Environmental Protection Agency, 

1984a). The index'suggests that the potential for ozone formation was greater 

in 1983 than in 1981 and 1982; however, the index also indicates that even 

greater potential for ozone formation existed in 1980 than in 1983 and that 

conditions in 1979 might be comparable to those in 1981 and 1982. Thus, this 

simple index may indicate the direction but not the magnitude of the effects 

of ineteorological factors on ozone formation. 

From a study of climate in the upper Midwest during the summer of 1983, 

Wendland et al. (1984) have reported that temperatures, which correlate reasonably well with insolation, were generally higher in the summer of 1983 than 

the 1950- through- 1980 norm for the 12-state region; and that cooling-degree 

days over about one-third of the region were 50 percent higher than normal. 

While'not quanti,tative or conclusive, these studies do suggest that meteorological conditions in 1983 contributed to the increase in the second-highest 

value among daily maximum 1-hour ozone concentrations. 

Trends in the composite average second-highest daily maximum 1-hour ozone 

values in the 10 EPA regions are shown in Figure 5-2 (U.S. Environmental 

Protection Agency, 1984a). , The use of data beginning with 1979 avoids some of 

the potential effects of changes in calibration and quality.assurance procedures 

mentioned earlier. Nine of the 10 regions show a 7 to 15 percent decrease in 

this statistic from the 1979-1980 period to the 1981-1982 period. The same 

nine regions showed increases of 6 to 16 percent from the 1981-1982 period to 

1983, demonstrating the pervasiveness of the trend. Only in.the Pacific 

Northwest, Region X, was there an opposite trend: +6 percent in 1981-1982 and 

-9 percent in 1983. 

5.3 OVERVIEW OF OZONE CONCENTRATIONS IN URBAN AREAS 

An overview of nationwide urban ozone concentrations for 1981 is:provided 

in Figures 5-3"and,5-4, which depict graphically the average daylight concentrations. Figure 5-3 shows data for spring and summer months, the months that.make 

up the smog season in most though not al1 areas of the nation; and Figure 5-4 

shows daylight concentrations during the fall and winter months. The daylight 

period of 6:00 a.m. to 8:00 p.m. includes the hours of greatest human activity 

5-7 
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Figure 5-2. Comparison of the 1979-1980, 1981-1982, and 1983 

composite average of the second highest daily max6mum 1-hour ozone 

concentrations across EPA Regions. 

allote: Regions are composed of the following states: 

I CT, MA, ME, NH, RI, VT 

11 NJ, NY, PR, VI 

111 DE, MD, PA, VA, WV 

IV AL, FL, GA, KY, MS, NC, SC, TN 

V IL, IN, MI, MN, OH, WI 

VI IA, KS, MO, NE 

VII AR, LA, NM, OK, TX 

VIII CO, MT, ND, SD, UT, WY 

IX AZ, CA, HI, NV 

X AK, ID, OR, WA 

Source: U.S. Environmental Protection Agency (1984a). 

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outdoors; the hours when exposure of vegetation and ecosystems would be expected 

to have the greatest consequences (stomata are open in daylight and photosynthesis is taking place; see Chapter 6); and the hours of greatest local formation of ozone and other oxidants via photochemistry in the atmosphere (see 

Chapter 3). The average concentrations during the spring and summer months 

(second and third quarters of the year) are clustered mainly in the 0.04 to 

0.06 ppm range. Averages for the winter and fall months (first and fourth 

quarters) are clustered mainly in the 0.02 to 0.04 ppm range. 

The stations used in Figures 5-3 and 5-4 reported at least 75 percent of 

the possible 1-hour values per quarter. Some ,stations, however, monitor ozone 

only during the months when the potential for photochemical ozone formation is 

significant in those localities: Also, certain areas of the United States are 

not monitored routinely for ozone because of the lack of emission sources or 

transport events and thus the low potential for significant ozone or oxidant 

concentrations. The Great Basin and the Great'P°lains, for example, are such 

areas. 

Figure 5-5 shows the nationwide frequency distributions of the first-, 

second-, and third-highest 1-hour 0 3 concentrations at gpredominantly urban 

stations aggregated for 1979, 1980, and 1981, as well as the highest l-hour 0 3 

concentration at sites of the Nat.ional Air Pollution Background Network (NAPBN) 

(see Section 5.4.1) aggregated .for the same 3 years. Only data collected by 

the Federal Reference Method (chemiluminescence) or the equivalent UV method 

(see:Chapter 4) have been used in this analysis. A"valid site" is one reporting at least 75 percent of the 87Wpossible 1-hour values in a year. There 

were 282 such sites in 1979, 266"in 1980, and 358 in 1981 (U.S. Environmental 

Protection Agency, 1980, 1981, 1982a). As ° shown by Figure 5-5, 50 percent of 

the second-highest 1-hour values from non-NAPBN sites in this 3-year period 

we.re0.12 ppm or less and about 10 percent were equal to or greater than 0.20 

ppm. At the NAPBN sites, the collective 3-year distr•ibution (1979 through 

1981) is such that about 60 percent of the values are less than 0.10 ppm and 

fewer'than 20 percent are higher than 0.12 ppm. 

Table 5-1 lists the second-highest 1-hour 0 3 values reported for 1981 

through 1983 for the 80'most populous Standard Metropolitan Statistical Areas 

(SMSAs), grouped by population (see Table 5-1 for a definition of the "second- 

. ., 

highest" values). CollectiveYy these SMSAs account for about 53 to 54 percent 

of the 1981 United States population of 229.3 million. The significant observation to be drawn from this table of second-highest values is that the lowest 

5-11 
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5-12 
TABLE 5-1. SECOND-HIGHEST 1-hr OZONE CONCENTRATIONSa' bREPORTED FOR 

STANDARD METROPOLITAN STATISTICAL AREAS HAVING POPULATIONS >0.5 MILLION, 

1981 THR011GH 1983 

Ozone concentration, ppm 

Standard Metropolitan Statistical Area 1981 1982 1983 

Population >2 million 

New York, NY - NJ 

Los Angeles - Long Beach, CA 

Chicago, IL 

Philadelphia, PA - NJ 

Detroit, MI 

San Francisco - Oakland, CA 

Washington, DC - MD - VA 

Dallas - Fort Worth, TX 

Houston, TX 

Boston, MA 

Nassau - Suffolk, NY 

St. Louis, MO - IL 

Pittsburgh, PA 

Baltimore, MD 

Minneapolis - St. Paul, MN -.WI 

Atlanta, GA 

Summary statistics: 

Minimum 1-.hour value 

Median 1-hour value 

Maximum 1=hour value 

Population 1 to < 2 million 

Newark, NJ 

Anaheim - Santa Ana - Garden Grove, CA 

Cleveland, OH 

San Diego, CA 

Miami, FL 

Denver - Boulder, CO 

Seattle - Everett, WA 

Tampa - St. Petersburg, FL 

Riverside - San Bernardino - Ontario, CA 

Phoenix, AZ 

Cincinnati, OH - KY - IN 

Milwaukee, WI 

Kansas City, MO - KS 

San Jose, CA 

Buffalo, NY 

Portland, OR - WA 

New Orleans, LA 

Indianapolis, IN 

Columbus, OH 

0.18 0.17 0.19 

0.35 0.32 0.37 

0.14 0.12 0.17 

0.17 0.18 0.10 

0.15 0.16 0.17 

0.14 0.14 0.17 

0.15 0.15 0.17 

0.15 0.17 0.16 

0.23 0.21 0.28 

0.13 0.16c 0.18 

0.14 0.13 0.17 

0.15 0.16 0.18 

0.16 0.14 0.14 

0.17 0.14 0.19 

0.10 0.10 0.13 

0.14 0.14 0.17 

0.10 0.10 0.10 

0.15' 0.15 0:17 

0.35 0.32 0.37 

0.14 0.17 0.25 

0.31 0.18c 0.28 

0.12 0.13 0.15 

0.24 0.21 0.20 

0.14 0.14 0.12 

0.13 0.14 0.14 

0.12 0.09 0.10 

0.12 0.12 0.14 

0.34 0.32 0..34 

0.16 0.12 0.16 

0.13 0.13 0.15 

0.17 0.13 0.18 

0.12 0.10 0.13. 

0.14 0.14 0.16 

0.12c 0.11 0.12 

0.15 0.12 0.12 

0.12 0.17 0.12 

0.13 0.12 0.14 

0.11 0.13 0.12 

5-13 
TABLE 5-1 (continued). SECOND-HIGHEST 1-hr OZONE CONCENTRATIONS a ' bREPORTED 

FOR STANDARD METROPOLITAN STATISTICAL AREAS HAVING POPULATIONS > 0.5 MILLION, 

1981 THROUGH 1983 

Ozone concentration, ppm 

Standard Metropolitan Statistical Area 1981 1982 1983 

San Juan, PR 0.07 0.04c NDd 

San Antonio, TX 0.12 0.14 0.12 

Fort Lauderdale - Hollywood, FL 0.11 0.09 0.10 

Sacramento, CA 0.17 0.16 0.15 

Summary statistics: 

Minimum 1-hour value 0.07 0.04 0.10 

Median 1-hour value 0.13 0.13 0.14 

Maximum 1-hour value 0.34 0.32 0.34 

Ponulation 0.5 to < 1 million 

Rochester, NY 0.12 0.11 0.12 

Salt Lake City - Ogden, UT 0.16 0.14 0.14 

Providence - Harwick - Pawtucket, RI - MA 0.15 0.15 0.15 

Memphis, TN - AR - MS 0.14 0.13 0.15 

Louisville, KY - IN 0.14 0.14 0.16 

Nashville - Davidson, TN 0.13 0.11 0.12 

Birmingham, AL 0.16 0.15 0.15 

Oklahoma City, OK 0.11 O.11c 0.11 

Dayton, OH 0.13 0.16 0.13 

Greensboro - Winston-Salem - High Point, NC 0.11 0.11 0.12 

Norfolk - Virginia Beach - Portsmouth, VA - NC 0.12 0.11 0.13 

Albany - Schenectady - Troy, NY 0.12 0.12 0.12 

Toledo, OH - MI 0.13 0.13 0.13 

Honolulu, HI 0.05 0.07 0.06 

Jacksonville, FL 0.11 0.12 0.14 

Hartford, CT 0.15 0.17 0.19 

Orlando, FL 0.10 O.IOc 0.11 

Tulsa, OK 0.15 0.13 0.13 

Akron, OH 0.24e 0.14 0.13 

Gary - Hammond - East Chicago, IN 0.14 0.13 0.17 

Syracuse, NY 0.11 0.12 0.08 

Northeast Pennsylvania 0.10 0.16 0.13 

Charlotte - Gastonia, NC 0.12 0.12 0.15 

Allentown - Bethlehem - Easton, PA - NJ 0.12 0.15 0.14 

Richmond, VA 0.12 0.12 0.14 

Grand Rapids, MI 0.11 0.11 0.14 

New Brunswick - Perth Amboy - Sayreville, NJ 0.13 0.16 0.19 

West Palm Beach - Boca Raton, FL 0.09 0.09 0.09 

Omaha, NE - IA 0.08 0.09 0.09 

Greenville - Spartanburg, SC 0.11 0.11 0.11 

Jersey City, NJ 0.14 0.14 0.16 

Austin, TX 0.12 0.11 0.12 

5-14 
TABLE 5-1 (continued). SECOND-HIGHEST 1-hr OZONE CONCENTRATIONS a ' bREPORTED 

FOR STANDARD METROPOLITAN STATISTICAL AREAS HAVING POPULATIONS >0.5 MILLION, 

1981 THROUGH 1983 

Ozone concentration, ppm 

Standard Metropolitan Statistical Area 1981 1982 1983 

Youngstown - Warren, OH 

Tucson, AZ 

Raleigh - Durham, NC 

Springfield - Chicopee - Holyoke, MA 

Oxnard - Simi Valley - Ventura, CA 

Wilmington, DE - NJ - MD 

Flint, MI 

Fresno, CA 

Long Branch - Asbury Park, NJ 

Summary statistics: 

Minimum 1-hour value 

Median 1-hour value 

Maximum 1-hour value 

0.13 0.11 0.11 

0.10 0.12 0.11 

0.12 0.09 0.13 

- CT 0.16 0.16 0.19 

0.20 0.22 0.21 

0.12c 0.16 0.18 

0.11 0.11 0.11 

0.14 0. lg 0.1R 

ND ND ND 

0.05 0.07 0.06 

0.12 0.12 0.13 

0.24 0.22 0.21 

aThe maximum 1-hour value for each day forms the data set from which the highest 

and "second-highest" concentrations are determined. Thus, the "second-highest" 

1-hour concentration occurs on a different day from the highest and therefore 

is not necessarily the absolute second-highest concentration for the monitoring 

period. 

bThese values permit a comparison of potentiai exposures and are not necessarily 

equivalent to the "design value" used for control strategy development. 

c Less than 50% of days in ozone season. 

dND = no data. 

eQuestionable data; third-highest value for 1981 was 0.125 ppm. 

Source: U.S. Environmental Protection Agency (1984a). 

5-15 
median concentrations in 1981, 1982, and 1983, for SMSAs having populations of 

0.5 to 1 million, are 0.12, 0.12, and 0.13 ppm, values that equal or exceed 

the current national ambient air quality standard for ozone. With the possible 

exception of the concentration reported for Akron, Ohio, in 1981, which appears 

to be questionable (see footnote, Table 5-1), the highest of all the secondhighest 1-hour concentrations in these 80 SMSAs occurred in California in all 

3 years. 

5.4 OVERVIEW OF OZONE CONCENTRATIONS IN NONURBAN AREAS 

As mentioned in the preceding section, very few ozone monitoring stations 

are located in nonurban areas. Consequently, the aerometric data base for 

nonurban areas is not comparable to that for urban areas. The nonurban data 

presented in this section were obtained from two special-purpose monitoring 

networks that were designed to measure ozone concentrations at sites specifically selected to represent a variety of pristine, rural, or suburban environments. These sites do not all represent areas totally unaffected by manmade 

ozone or its precursors, as shown by the fact that some data records contain a 

significant number of high values that are best explained as resulting from 

the transport of ozone or its precursors from upwind urban areas. The data 

given here are intended to show an overview of nonurban concentrations in 

areas with relatively infrequent urban influences. Additional data on specific rural areas are presented in Sections 5.5.1 and 5.5.2. 

5.4.1National Air Pollution Background Network (NAPBN ) 

The NAPBN consists of eight stations located in eight National Forests 

(NF) across the country (Figure 5-6). The first three stations began operation in 1976 (Green Mountain NF, Vermont; Kisatchie NF, Louisiana; and Custer 

NF, Montana); the second three in 1978 (Chequamegon NF, Wisconsin; Mark Twain 

NF, Missouri; and Croatan NF, North Carolina); and the last two in 1979 (Apache 

NF, Arizona; and Ochoco NF, Oregon). Yearly summaries of ozone concentrations 

through 1980 are shown in Table 5-2 for the three sites established first 

(Evans et al., 1983). The principal points of interest in these summary statistics are the range and the arithmetic mean of the ozone concentrations 

measured at these National Forest sites. The arithmetic mean concentrations 

for the three sites ranged from 0.027 ± 0.015 ppm at the Kisatchie NF site in 

5-16 
Figure 5-6. Locations of the eight national forest (NF) stations 

constituting the National Air Pollution Background Network 

(NAPBN). 

Source: Evans et al. (1983). 

5-17 
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5-18 
1979 to 0.040 ± 0.011 ppm at the Custer NF site in 1977. The arithmetic mean 

concentration across ail years and all three sites was 0.033 ppm. Fluctuations in the observed concentrations from year-to-year and site-to-site are 

demonstrated by the range of concentrations measured and by the size of the 

standard deviations as well. The lowest concentrations seen were below the 

limits of detection of the chemiluminescence monitor employed, but the highest 

concentrations observed at the Kisatchie NF and Green Mt. NF sites were both 

above the present ozone standard of 0.12 ppm. In 1979, only one excursion 

over 0.12 ppm 0 3 was recorded at an NAPBN site. That excursion was recorded 

at the Mark Twain NF site in Missouri (Evans et al., 1983; Lefohn, 1984). In 

1980, seven 1-hour excursions over 0.12 ppm were reported for Croatan NF, North 

Carolina (Lefohn, 1984). None of the NAPBN sites recorded an ozone concentration greater than 0.12 ppm in 1981 (Evans et al., 1983). 

These summary statistics show somewhat higher mean concentrations, lower 

maximum concentrations, and lower,standard deviations in data obtained at the 

Custer NF site than at the other two, which may indicate that meteorological 

conditions are less variable at that site or that the site is much less affected, if not altogether unaffected, by manmade ozone or its precursors. 

Duri ng a 6-day peri od i n 1979, the NAPBN s i te i n the Mark Twai n NF, 

Missouri, showed ozone concentrations well in excess of typicai values. 

Table 5-3 shows the peak 1-hour value for each of the 6 days. A 1-hour value 

of 0.125 ppm, the maximum observed at any NAPBN site in 1979, was measured at 

that site on July 21, 1979. Evans et al. (1983) calculated the trajectories 

of air masses reaching the site during the 6-day period of July 18 through 

July 23, 1979. They ascribed the unusually high values, including the peak 

value on the 21st, to pollutants picked up as the trajectory passed over urban 

areas in the Ohio River Valley and the Great Lakes region. Figure 5-7 shows 

the trajectories for the air parcels reaching the Mark Twain NF site at midnight 

(0000), 8 a.m. (0800), noon (1200), and 6 p.m. (1800) on July 21, 1979. On 

July 23, clouds and rain spread over the region and the airflow trajectories 

shifted to the east and south, reducing both the quantities of transported 

precursors and the potential for photochemical ozone generation. 

More recent and more comprehensive data from the NAPBN sites are presented 

in Table 5-4 and in Figures 5-8A and 5-8B. Table 5-4 presents the percentile 

distributions of ozone concentrations at all eight of the sites, aggregated by 

quarter across severai years. 

5-19 
TABLE 5-3. CONCENTRATIONS OF OZONE DURING 6-day PERIOD OF HIGH 

VALUES AT NAPBN SITE IN MARK.TWAIN NATIONAL FOREST, MISSOURI, 1979 

1-hr maximum 

Date 0 3 concentration, ppm 

July 18 0.080 

July 19 0.100 

July 20 0.115 

July 21 0.125 

July 22 0.120 

July 23 0.050 

Source: Evans et al. (1983) 

As the data in Table 5-4 show, the arithmetic mean concentrations of 

ozone, for the years of data averaged, are generally higher in the second 

quarter of the year (April, May, June) than at other times at the NAPBN sites. 

Although a few excursions of the 1-hour concentration above 0.12 ppm were 

recorded at some of these sites, as discussed above (Evans et al., 1983; 

Lefohn, 1984), the distribution given in Table 5-4 clearly indicates that 99 

percent of the 1-hour ozone concentrations measured at these sites are well 

below the present ozone standard. This is true even at those sites thought to 

be influenced by transport of ozone (Green Mt. NF) or demonstrated to be 

influenced by transport (Mark Twain NF, Evans et al., 1983). As shown in 

Tabie 5-4, the highest 99th percentile value was 0.093 ppm, reached at both 

the Green Mt. NF, VT, site (second quarter) and the Mark Twain NF, M0, site 

(third quarter). The maximum 1-hour ozone concentrations at these sites ranged 

from 0.050 ppm at Custer NF, MT (in the fourth quarter) to 0.155 ppm at Mark 

Twain NF, MO (in the third quarter). The second-highest'1-hour concentration 

among maximum daily 1-hour values ranged from 0.050 ppm at Custer NF, MT 

(fourth quarter) to 0.150 ppm at Mark Twain NF, MO (third quarter). 

Five of the NAPBN monitoring stations (Apache NF, AZ; Mark Twain NF, M0; 

Custer NF, MT; Croatan NF, NC; Ochoco NF, OR) reported sufficient data for 

1979 through 1983 to support an examination of their second-highest 1-hour 

5-20 
MINNEAPOLIS • 

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DES MOINES CHICAGO 

• i 

~ INDIANAPOLlS ~ 

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KANSAS CITY INCINNATI 

ST. LOUIS " 

MARK TWAIN N.F. 

EM ~HIS 

Figure 5-7. Trajectory analysis plots for the 

NAPBN site at Mark Twain National Forest, MO, 

July 21, 1979 (distance between bars represents 

12 hr). 

Source: Evans et al. (1983). 

5-21 
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5-23 
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c. o. 

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U 

w 

Z 

O N 

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o 

[4 

0. 

0. 

0. 

0. 

0. 

0. 

0. 

0. 

0. 

0.- - 

1979 1980 1981 1982 1983 

PERIOD, years 

Figure 5-8A. Second-highest value among maximum 1-hr ozone 

concentrations at five NAPBN monitoring stations, 1979 through 

1983. 

Source: Derived from SAROAD (1985 b-f). 

0.14 

c. 0.1: 

a 

Z 

O 

H 

cr. 0.1( 

f2 

w 

U 

p 0.01 

C) 

w 

Z 

O 

N 

O 0.01 

0.0, 

i`.1i`.1 7`JZfll l.ya, 1.704 1 zlaJ 

PERIOD, years 

Figure 5-8B. Composite averages of the second-highest value 

among daily maximum 1-hr ozone concentrations at five NAPBN 

stations, 1979 through 1983. 

Source: Derived from SAROAD (1985 b-f). 

5-24 
ozone values for evidence of trends. As seen in Figure 5-8B, however, the 

diversity of the NAPBN site characteristics, the consequent large variation in 

recorded values, and the sma.il sample size result in such a large confidence 

interval around the average for each year that no trend can be determined. 

The data presented in Figures 5-8A and 5-8B show the NAPBN second-highest 

daily maximum 1-hour ozone concentrations to be about one-half those reported 

for urban-suburban sites for the same period (Figure 5-1). 

5.4.2Sulfate Regional Experiment Sites (SURE) 

As part of a comprehensive air monitoring project sponsored by the Electric 

Power Research Institute (Martinez and Singh, 1979), ozone measurements were 

made by the chemiluminescence method in the last 6 months of 1977 at nine 

"nonurban" SURE sites in the eastern United States, shown in Figure 5-9. 'On 

the basis of diurnal NOxpatterns that indicated the influence of traffic 

emissions, five of the sites were classed as "suburban"; the other four were 

classed as "rural." The ozone data from these nine stations are summarized in 

Table 5-5. Martinez and Singh (1979) noted that the four rural stations 

occasionally recorded high values comparable with those in urban areas, but that 

the incidence was low. They concluded that infrequent transport of ozone or 

its precursors, or both, rather than local ozone generation, was the most probable cause of these high values. 

5.5 VARIATIONS IN OZONE CONCENTRATIONS: DATA FROM SELECTED URBAN AND NONURBAN 

SITES 

Variations of ozone concentrations by season and by time of day have 

been long known and are well documented. First studied in smog chambers, 

diurnal patterns have since been corroborated by field investigations, and 

exceptions to such general patterns have been examined and documented. Likewise, field investigations have substantiated general seasonal patterns and 

exceptions to them, and have also established a number of spatial variations 

in concentration, such as those that occur with latitude or with altitude. 

While it is difficult to discuss temporal and spatial variations separately, 

this section is subdivided along those lines for convenience. 

5-25 
MN MI 

ME 

W I V;T. 

NY NH 

IA MI SITE 1 MA 

SITE 2 ♦ 

CT 

SITE 7 ♦ 

1 L 

OH 

PA 

IN SITE 4 ♦ 

NJ 

MD DL~ 

MO vw SITE 3 SITE 5A SITE 9 

KY VA 

AR 

TN ♦ SITE 6 SITE 8 ♦ ~ 

NC 

MS AL GA SC 0 50 150 250 

LA km ~ 

Figure 5-9. Location of Sulfate Regional Experiment (SURE) 

monitoring stations. 

Source: Martinez and Singh (1979). 

5-26 
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5-27 
5.5.1 Temporal Variations in Ozone Concentrations 

5.5.1.1 Diurnal Variations in Ozone Concentrations. By definition, diurnal 

variations are those that occur during a 24-hour period. Diurnal patterns of 

ozone may be expected to vary with location, depending on the balance among 

the many factors affecting ozone formation, transport, and destruction, as described in Chapter 3 and noted in Section 5.1. Although they vary with locality, 

diurnal patterns for ozone typically show a rise in concentration from low or 

near-zero levels to an early afternoon peak. The 1978 criteria document 

ascribed the diurnal pattern of concentrations to three simultaneous processes: 

(1) downward transport of ozone from layers aloft; (2) destruction of ozone 

through contact with surfaces and through reaction with nitric oxide (NO) at 

ground level; and (3) in situ photochemical production of ozone (U.S. Environmental Protection Agency, 1978; Coffey et al., 1977; Mohnen, 1977; Reiter, 

1977a). Figure 5-10 shows the diurnal pattern of ozone concentrations on 

July 13, 1979, in Philadelphia, Pennsylvania. On this day a peak 1-hour average 

concentration of 0.20 ppm, the highest for the month, was reached at 2:00 p.m., 

presumably as the result of ineteorological factors, such as atmospheric mixing, 

and local photochemical processes. The severe depression of concentrations to 

~ 

below detection limits (less than 10 ppb) between 3:00 and 6:00 a.m. is usually 

explained as resulting from the scavenging of ozone by local nitric oxide 

emissions. In this regard, this station is typical of most urban locations. 

Diurnal profiles of ozone concentrations can vary from day to day at a 

specific site, however, because of changes in the various factors that influence 

concentrations. Such day-to-day variations are clearly demonstrated in Figure 

5-11 (SAROAD, 1985c), which shows diurnal variations in ozone concentrations on 

2 consecutive days at the same monitoring site in Detroit, Michigan. Differences in timing and magnitude occur that are especially noticeable between 

midnight and about 7:00 a.m. Transport is probably involved in these nighttime 

variations. The afternoon peak concentrations, the actual maxima for the 2 

days, differ in magnitude but not in timing. 

Composite diurnal data, that is, concentrations for each hour of the day 

averaged over multiple days or months, often differ markedly from the diurnal 

cycle shown by concentrations for a specific day. In Figures 5-12 through 

5-14 (SAROAD, 1985d), diurnal data for 2 consecutive days are compared with 

composite diurnal data (1-month averages of hour-by-hour measurements) at 

three different kinds of sites: center city (Washington, DC), rural but near 

A;K_3 
E 

a 

a 

z 

O 

B 

~ H 

z 

W 

v 

z 

0 

IL) 

W 

z 

O 

N 

O 

0.19 

0.18 

0.17 

0.16 

0.15 

0.14 

0.13 

0.12 

0.11 

0.10 

0.09 

0.08 

0.07 

0.06 

0.05 

0.04 

0.03 

0.02 

0.01 

0 

11 2 1 2 3 4 5 6 7 S 9 1n 11 i 1 2 3 4 5 R 7 R 9 1(1 1' 

NOON 

~--a.m. HOUR OF DAY p.m.-~ 

Figure 5-1 O. Diurnal pattern of 1-hr ozone 

concentrations on July 13, 1979, Philadelphia, PA. 

Source: SAROAD (1985b). 

5-29 
0.12 

o, 0.10 

a. 

Z 

0 0.08 FQ 

~ FW 0.06 

U 

z 

U 0.04 

w 

z 

O 

~ 0.02 

1"o 

• SEPTEMBER 20 

O SEPTEMBER 21 

240 

220 

200 

180 

160 

140 

120 

100 

80 

60 

40 

20 

U 

U 0 Ln 

N 

U~ 

M E 

NI 

~ 

~ 

Z 

O 

FQ 

~ 

Fz 

w 

U 

z 

O 

U 

w 

Z 

O 

N 

O 

19 1 9 3 49; 6 7 R 9 1(1 11 T 1 9 3 4-9; R 7 R Q 1n 1 

NOON 

~--~a.m. HOUR OF DAY 

Figure 5-11. Diurnal patterns of 1-hr ozone 

concentrations, September 20 and 21, 1980, 

Detroit, M I. 

Source: SAROAD (1985c). 

5-30 
0.12 

E 

a 0.10 z_ O 

F 0.08 

Q 

~ 

a 

W 0'06 

t) 

z 

~ 0.04 

W 

z 

r®rN 0.02 

O 

cENTER cITY-c0NIMERC12 

— -- JULY 8-9 

o ~ 1-rnonth 

e 

o ~` ~ e 

r ~ 1 

I 

a 

JL 1 

1_ 

0 

24 4 8 12 16 20 24 4 8 12 16 20 24 

1i--a.m.--NOON--p.rn.—o i

-a—a.m.—NOON—p.m.— I 

HOUR OF DAY 

Figure 5-12. Diurnal and 1-rnonth composite 

diurnal variations in ozone concentrations, . 

Washington, DC, July 1981. 

Source: SAROAD (1985d). 

0.12 

~, 0.10 

A. 

Z 

0 0.08 

Q 

cr 

z 0.06 

W 

0 

z 

v 0.04 

W 

z 

N 0.02 

O 

0 

RURAL-NEAR URBAN --- SEPT. 29-30 

1-month 

24 4 8 12 16 20 24 4 8 12 16` 20 24 

14—a.m: —NOON--p.m.-- i

a--a.m.—NOON--p.m.---H 

HOUR OF DAY 

Figure 5-13. Diurnal and 1-month composite 

diurnal variations in ozone concentrations, St. 

Louis County, MO, September 1981. 

Source: SAROAD (1985d). 

5-31 
0.12 

a 0.10 

a 

Z 

0 0.08 

Q 

~ 

W0.06 

V 

Z 

v 0.04 

W 

Z 

0 0.02 

0 

SUBURBAN-RESIDENTIAL 

— — — OCT. 11-12 

1-month 

~ 

♦ 

/ ♦ • e` • ♦ 

~ . 

_ ~ 1 ~~_ ~• 

24 4 8 12 16 20 24 4 8 12 16 20 24 

~

*_a.m..—NOON.—p.m.--P+*—a.m..—NOON— p.m.—so-1 

HOUR OF DAY 

Figure 5-14. Diurnal and 1-month composite 

diurnal variations in ozone concentrations, Alton, 

IL, October 1981 (fourth quarter). 

Source: SAROAD (1985d). 

an urban area (St. Louis County, MO), and suburban-residential but in a cropgrowing area (Alton, IL). Several obvious points of interest present themselves 

in these graphs: (1) at some sites, at least, peaks can occur at virtually 

any hour of the day or night but these peaks may not show up strongly in the 

longer-term average data; (2) some sites may be exposed to multiple peaks 

during a 24-hour period; and (3) disparities, some of them large, can exist 

between peaks (the diurnal data) and the 1-month mean (the composite diurnal 

data) of hourly ozone concentrations. These figures are given simply as examples of the differences that can occur between daily and monthly mean concentration patterns. 

The effects of averaging are readily apparent when diurnal or short-term 

composite diurnal ozone concentrations are compared with longer-term composite 

diurnal ozone concentrations. When compared with Figures 5-12 through 5-14 

(daily values and 1-month averages), Figure 5-15 (SAROAD, 1985d), based on 

3-month averages, demonstrates rather graphically the effects of lengthening 

the period of time over which values are averaged. This figure shows a 

composite diurnal pattern calculated on the basis of 3 months. While seasonal 

5-32 
~ ri 

z 

O 

f- Q 

FE 

z 

W 

0 

z O 

0 

W 

z 

N 

O 

0.10 

0.09 

0.08 

0.07 

0.66 

0.05 

0.04 

0.03 

0.02 

0.01 

0 

ALTON, IL 

~- ~ • •-• , 3rd Q - 

2nd Q~ 

... ..

....... 

 . 4th Q ~ • ~ ~--~ 

~:./. ... 

... .. _ 

24 2 4 6 8 10 12 14 16 18 20 22 24 

~~-a.m.• NOON•--- 

HOUR OF DAY 

Figure 5-15. Composite diurnal patterns by 

quarter of ozone concentrations, Alton, IL, 1981. 

Source: SAR®AD (1985d). 

differences are seen, and will be discussed later, the comparison of 3-month 

and 1-month composite diurnal concentrations at the Aiton, IL, site readily 

demonstrates the smoothing out of peak concentrations as the averaging period 

is lengthened. Thus, a fourth pertinent point of interest, related to the 

third point cited above, emerges from the data presented above: that is, 

increasing the averaging time obscures the magnitude and time of occurrence of 

peak ozone concentrations. This is an obvious and familiar resuit in the 

statisticai treatment of monitoring data, but one that is highly pertinent to 

the protection of human health and welfare from the effects of ozone. 

The significance of the relationships of peak and mean concentrations 

depends upon whether the health and welfare effects of exposure to ozone are 

solely concentration-dependent, heavily concentration-dependent, or both 

concentration- and time-dependent. Depending upon whether acute or chronic 

exposures elicit the health and welfare effects of concern, careful attention 

5-33 
may have to be paid to the relationship of short-term (e.g., 1-hour) versus 

longer averaging times. 

Quantitative analyses of the relationships among maximum or second-highest 

1-hour concentrations and daylight, diurnal, monthly, seasonal, and yearly 

average ozone concentrations lie outside the scope of this document. An 

example of one study, however, serves to describe the kinds of analyses that 

may be done to determine relationships among various exposure statistics or 

"averaging times." Lefohn and Benedict (1985) have examined the relationship 

between 7-hour and 12-hour (daylight) mean ozone concentrations averaged over 

a season (7 months). They also examined the relationships between each of 

these two exposure statistics and the 1-hour values during the same season 

>0.10 ppm. For their analyses they evaluated data records from all sites in 

the SAROAD network that met the following criteria: (1) existence of 4 years 

of monitoring data (1978 through 1981); (2) >4300 1-hour concentrations re= 

ported; and (3) the occurrence of exposures >10.2 ppm-hours, which was an 

index of sites exposed repeatedly to ozone concentrations >0.10 ppm over the 

specified time period. Sites so selected were then subjected to additional 

criteria that resulted in the identification of monitoring sites located in 

areas where agriculturally important cash crops were grown. 

On the basis of calculated integrated exposures (hours times 1-hour 

concentrations >0.10 ppm), the monitoring sites thus identified were subdivided 

by Lefohn and Benedict (1985) into two categories: (1) those with 80 to 200 

occurrences of 1-hour val ues <0.10 ppm (i . e. , 80 to 200 hours at >0.10 ppm 

ozone), and (2) those with 200 to 1612 hours at >0.10 ppm 03 . To describe the 

quantitative relationships among the exposure statistics at sites in the two 

categories, a Pearson linear correlation matrix was calculated. For sites 

with 200 to 1612 occurrences of 1-hour concentrations >0.10 ppm, the correlation was 0.88 between the number of occurrences and the 7-hour seasonal mean 

and 0.93 between the number of occurrences and the 12-hour seasonal mean. The 

correlation between the 7- and 12-hour seasonal means was 0.96. When correlations were calculated from data obtained at sites with 80 to 200 occurrences 

of 1-hour concentrations >0.10 ppm, the correlations between occurrences and 

the 7- and 12-hour seasonal means were 0.40 and 0.47, respectively. The 

correlation for the two seasonal means at these sites was 0.94 (Lefohn and 

Benedict, 1985). 

5-34 
No attempt is made in this section to document the respective contributions of local formation of ozone versus transport of ozone; however, the 

occurrence of multiple peak,ozone concentrations within a 24-hour period is 

usually construed as indicating the presence of ozone transported to the site 

from elsewhere. Figure 5-16 illustrates the diurnal variations that can be 

seen when transport occurs. Note the occurrence of dual peaks on each of 3 

successive days at this site, part of the Sulfate Regional Experiment (SURE) 

network. 

A familiar measure of ozone air quality is the number or percentage of 

days on which some specified concentration is equailed or exceeded. This 

measure, however, does not shed light on one of the more important questions 

regarding the effects of ozone on both people and plants; that is, the possible 

significance of high concentrations lasting 1 hour or longer and then recurring 

on 2 or more successive days. 

In human controlled exposures, attenuation of responses to ozone has been 

observed at about 0.20 to 0.50 ppm in exercising subjects receiving repeated, 

consecutive-day exposures (see Chapter 10). That attenuation is lost after 

exposures to those levels cease (see Chapter 10 for the time course of loss of 

attenuation). It becomes of interest, therefore, to examine how many days in 

a row the maximum 1-hour ozone concentration reaches or exceeds_specified 

levels in communities in high-ozone areas, as well as in other parts of the 

country. 

The recurrence of high ozone concentrations on consecutive days was examined in data (SAROAD, 1985b-d) for one site in each, of four cities: Pasadena 

and Pomona, CA; Washington, DC; and Dallas, TX. The numbers of multiple-day 

events were tallied by length of event (i.e., how many consecutive days) using 

data for the daylight hours (6:00 a.m. to 8:00 p.m.) in the second and third 

quarters of 1979 through 1981. These sites were chosen because they include 

areas known to experience high ozone concentrations (California), and because 

they represent different geographic regions of the country (west, southwest, 

and east). Similar data could be compiled for any city for which sufficient 

aerometric data exist. The choice of the 14-hour daylight period and of the 

second and third quarters is consistent with known diurnal and seasonal patterns 

of ozone concentrations and with typical human, crop, and ecosystem exposures. 

In this discussion of the recurrence of respective specified ozone levels, 

a day or series of days on which the daily 1-hour maximum reached or exceeded 

5-35 
a • - - - - - . N 

~ 

~ 

z~ ~ r% W 

a 

~ D 

D 

« Z ~ a 

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f 

Z `► 

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y z I z Q 

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cp N 

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p O O O O O O O 

wdd 'N011t/alN30N00 5-36 
the specified level is called an "exposure"; the intervening day or days when 

that level was not reached is calied a"respite." Four ozone concentrations 

were chosen: 0.06, 0.12, 0.18, and 0.24 ppm. The 1- to 7-day events (i.e., 

exposures or respites) are individually tabulated and events longer than 

7 days are grouped together. 

The summaries of "exposure" and "respite" events in Tables 5-6 through 

5-9 show expected regional differences in the successive recurrence of these 

respective 1-hour concentrations. At the Dallas station, for example, the 

3-year data tally shows 11 exposure events when the daily l-hour maximum ozone 

value equalled or exceeded 0.06 ppm for more than 7 days in a row (Table 5-6). 

At first glance, the Pasadena station, with 10 such exposures, appears similar, 

but those 10 exposure events spanned 443 days; in Dallas the 11 exposures involved only 168 days. At this lowest concentration (> 0.06 ppm), the Dallas 

station recorded more short-duration (< 7 days) exposures (45) involving more 

days (159) than the Pasadena station (14 exposures over 45 days), simply because 

the daily 1-hour maximum statistic in Pasadena remained above 0.06 ppm for such 

protracted periods. At concentrations > 0.12 ppm (Table 5-7), the lengthy 

exposures at the Pasadena site resolved into numerous shorter exposures; whereas 

in Dallas the exposures markedly dwindled in number and duration. At concentrations > 0.18 ppm (Table 5-8), short-term (< 7 days) exposures in Pasadena 

were yet more numerous; Dallas had only two 1-day exposures. At > 0.24 ppm 

(Table 5-9), the incidence rate of exposures at the Pasadena station finally 

decreased. 

Note that these tables (Tables 5-6 through 5-9) also present the single 

and multiple-day periods when concentrations were lower than specified levels 

("respites"); "event-days" include all "exposures" to and "respites" from 

specified concentrations.- , such that "total event-days" for a given site equal 

the total days monitored. 

These tables give a more extensive indication of the long-term severity 

of ozone air quality at some of the selected sites than is given by the use of 

a 1-hour exposure statistic alone. The human responses to single acute exposures to ozone are well documented (Chapter 10), but the full significance of 

the attenuation of human responses to ozone over the course of a multiple-day 

exposure, and any possible consequences of the repetition of such multiple-day 

exposures within a smog season or over a number of years remain uncertain. 

5-37 
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5-41 
5.5.1.2 Seasonal Variations in Ozone Concentrations. In addition to the 

diurnal cycles and between-day variations discussed in the preceding section, 

seasonal variations in ozone concentrations occur (for,the reasons discussed 

in Chapter 3 and Section 5.1) and usually assume characteristic patterns. 

In order to compile an assessment of potential ozone damage to the six 

leading commercial crops in the United States (corn,' soybeans, hay, wheat, 

cotton, and tobacco), Lefohn (1982) surveyed 304 ozone monitoring stations and 

identified 24 that (1) were located in counties producing significant quantities of one or more of these six crops in 1978; (2) reported at least 50 

percent of possible hourly data in 1978; (3).reported an hourly maximum of at 

least 0.1 ppm 0 3 ; and (4) ranked high in cumulative.ozone exposure for the 

period April to October 1978. Six of these sites represented counties with 

high soybean, wheat, or hay production. Quarterly composite diurnal patterns 

for six of these sites with reasonably complete (>75 percent) 1981 data are 

shown in Figure 5-17 (SAROAD, 1985d). The average levels are apparently 

comparable with the long-term averages at the NAPBN sites previously discussed 

(Section 5.4.1). In addition, the diurnal patterns for these sites clearly 

show the division of the afternoon ozorie concentrations into two seasonal 

patterns, the low "winter" levels in the first and fourth quarters and the 

higher "summer" levels in the second and third quarters of the year. 

Although averaging causes details to be obscured, the average diurnal 

patterns in Figure 5-17 show that the time of occurrence of peaks differs 

among sites, and, to an extent, between seasons. The seasonal differences'in 

time of day are especially noticeable for the.fourth quarter. Among'the sites 

shown in Figure 5-17, ozone concentrations appear to peak at 2:00 to 2:30 p.m. 

in Little Rock in the higher-concentration second.and third quarters. At 

Bakersfield in the second and third quarters, there is evidence, even from 

these composite data, of two peaks, the first at.about 1:00 p.m. and the second 

at 5:00 to 6:00 p.m. At the Clark County, Ohio, site, the peak concentrations 

in the second and third quarters center around about 5:00 p.m., but they do not 

return to "baseline" until after midnight. The patterns at the Bakersfield and 

Clark County sites appear to indicate transport into the areas. It is also 

possible that single peaks that are shifted to mid- to late afternoon, as at 

the Little Rock site, are the product of transport. Depending upon proximity 

to urban centers and wind speed and direction, rural areas are typically 

exposed to their peak concentrations later than urban areas, usually within 

daylight hours but not always. 

5-42 
0.10 

0.09 

0.08 

0.07 

0.06 

0.05 

0.04 

0.03 

0.02 

0.01 

0 

N. LITTLE ROCK, AR 

(A) 

1st Q 

- • - • 2nd Q 

- - -- 3rd Q ........ 4th Q 

s 

\ , 1 • d ~ ~ • ``` ` • 

24 2 4 6 8 10 12 14 16 18 20 22 24 

~--- a.m. p.m. --- 0 0.10 

0.09 SACRAMENTO, CA 

0.08 (C) 9 

E 

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24 2 4 6 8 10 12 14 16 18 20 22 24 24 2 4 6 8 10 12 14 16 18 20 22 24 

i 4.-a.m. +--- P.M. of-_a.m. --vI a--p.m.-_i

~ 

HOUR OF DAY, LST HOUR OF DAY; LST ' 

Figure 5-17. (A-F). Quarterly composite diurnal patterns of ozone 

concentrations at selected sites representing potential for exposure of major 

crops, 1981. 

Source: SAROAD (1985d). 

0.07 

0.06 

0.05 

0.04 

0.03 

0.02 

0.01 -•,..,.,,.

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5-43 
Composite diurnal variations in ozone concentrations at a rural site in 

Argonne, Illinois, over a 7-week period of the third quarter of 1980 are shown 

in Figure 5-18 (Kress and Miller, 1983). The actual day-to-day variations in 

ozone concentration over the entire third quarter of 1980 at that site are 

shown for comparison in Figure 5-19 (Kress and Miller, 1983). As part of the 

National Crop Loss Assessment Network, the site at Argonne monitors ozone 

concentrations over a 7-hour daytime period approximating the period of peak 

photosynthesis in crops. The data in Figure 5-18 yield a 24-hour average 

ozone concentration of 0.026 ppm (Kress and Miller, 1983). The 7-hour average 

for the same 7-week period at the Argonne site is 0.042 ppm (Kress and Miller, 

1983). The day-to-day variations in both the 7-hour and the 24-hour averages 

(Figure 5-19) generally appear to be greater than the average difference 

within a day for either the 7-hour or 24-hour periods (Figure 5-18). The 

fluctuations in 1-hour values either within a day or from day to day, however, 

would be larger than within-day or between-day variations in either the 7-hour 

or the 24-hour average. The 7-hour average will be higher than the 24-hour 

average because the former excludes the low nighttime concentrations. As 

Figure 5-19 and its data illustrate, the selection of the appropriate averaging 

time is critical for the accurate description of dose-response relationships 

and for the protection of human, vegetation, and other receptors from the 

effects of ozone. 

It is worth noting in this context that air pollution exposure statistics 

are of two basic functional types: descriptive and preventive. Descriptive 

exposure statistics simply define the conditions of concentration and exposure 

duration (averaging time) under which a specified effect has been observed or 

detected. Preventive exposure statistics are used prescriptively with the 

expectation of keeping a specified effect from happening. To be effective 

bases for an air quality standard, preventive exposure statistics must be 

(1) related to the true time-course of development of the specified effect(s), 

and (2) based on observed or predictable distributions, or both, of exposure 

conditions over the range of concentrations and durations producing the 

effect(s) of concern. 

In Figure 5-20 (A-H), seasonal variations in ozone concentrations in 1981 

are depicted using 1-month averages and the single 1-hour maximum concentration within the month for eight sites across the nation (SAROAD, 1985d). 

The data from most of these sites exhibit the expected pattern of high ozone 

5-44 
E 0.0s 

a 

c. 

z 

O 

F- Q 

~ 

~ 

W0.04 

0 

z 

O 

V 

LLI 

z 

O 

N 

O0.02 

0 

24 a.m. IVOON p.m. 24 

Figure 5-18. Composite diurnal ozone pattern at a rural 

NCLAN site in Argonne, IL, August 6 through Septerrtber 

30, 1980. 

Source: Kress and IVliller (1983). 

o.1s 

E 

a 

a 

z 

~ 0.10 

H 

Q 

cc 

~ 

z 

W 

V 

z 

V 0.05 

W Z 

O 

N 

O 

0 

JUL AUG SEP OCT 

MONTH OF YEAR 

Figure 5-19. Daily 7-hour and 24-hour average 

ozone concentrations at a rural NCLAN site in 

Argonne, IL, 1980. 

Source: Kress and IVliller (1983). 

5-45 
0.35 

0.30 

E 

a 

a 

Z 0.25 

O 

Q ~ 0.20 

z 

W 

U 

z00.15 

U 

W 

~ 0.10 

N 

O 

TUCSON, AZ 

(A) 

MAXIMUM 

AVERAGE 

0.35 

0.30 

0.25 

0.20 

0.15 

0.10 

LENNOX, CA 

(B) 

0.05 

0 

J F M A M J J A S O N D 

0.05 

0 

J F M A M J J A S O N D 

0.35 

0.30 

E 

a 

~ 0.25 

O 

F» 

~ 0.20 

~ 

W 

z 0.15 

O 

U 

2 0.10 

CO 

N 

O 

0.05 

0 

FOMONA, CA 

(C) 

0.35 

0.30 

0.25 

0.20 

0.15 

0.10 

0.05 

0 

DENVER, CO 

(D) 

J F M A M J J A S O N D J F M A M J J A S O N D 

MONTH OF YEAR MONTH OF YEAR 

Figure 5-20. (A-H). Seasonal variations in ozone concentrations as indicated 

by monthly averages and the 1-hour maximum in each month at selected sites, 

1981. 

5-46 
0.35 0.: 

WASHINGTON, DC TAMPA, FL 

0.30 (E) 0.: (F) 

E 

CL 

CL 

;i 0.25 

0 

C 0.20 r 

z 

Z 0.15 

0 

L) 

LU 

z

0 

 0.10 

N 

0 

0.05 

0 

MAXIMUM 

AVERAGE 

0.25 

0.20 

0.15 

0.10 

0.05 

0 

J F M A M J J A S 0 N D 

0.35 

J F M A M J J A S 0 N D 

0.35 

0.30 

E 

CL 

CL 0.25 

0 

0.20 

z 

z 0.15 

0 0 

Z

LU 0.10 

0 N 

0 

0.05 

0 

MORRIS CO., NJ 

(G) 

~

j

~ I i~ !E, ~ rs ~

da 

0.30 

0.25 

0.20 

0.15 

0.10 

0.05 

0 

DALLAS, TX 

(H) 

J F M A M J J A S 0 N D J F M A M J J A S 0 N D 

MONTH OF YEAR MONTH OF YEAR 

Figure 5-20. (A-H) (continued). Seasonal variations in ozone concentrations as 

indicated by monthly averages and the 1 -hour maximum in each month at 

selected sites, 1981. 

Source: SAROAD (1 985d). 

5-47 
levels in late spring or in summer and low levels in the winter. Data for 

Pomona (Figure 5-20C) and Denver (Figure 5-20D) show summer maxima. Tampa 

shows a late spring maximum but with concentrations in the fall (October) 

approaching those of spring (June) (Figure 5-20F). Dallas data also tend to 

be skewed toward higher spring concentrations; but note that November concentrations are also relatively high (Figure 5-20H). Averaging together data for 

several years would give a smoother "characteristic" pattern but also would 

obscure the fact that local, and even national, weather in a particular year 

plays at least as big a role in the formation of ozone as the regular seasonal 

changes in the elevation of the sun and the resulting variations in insolation. 

Because of seasonal changes in storm tracks from year to year, the general 

weather conditions in a given year may be more favorable for the formation of 

ozone and other oxidants than during the prior or following year. Thus, 

short-term concentration trends may not be indicative of real changes in air 

quality. 

5.5.1.3 Weekday-Weekend Variations in Ozone Concentrations. Atmospheric 

ozone concentrations represent the combined effects of emission sources and 

meteorological conditions. The various sections of this chapter have been 

based on the assumption that ozone precursor sources operate on a generally 

steady-state or at least on an average, repeatable diurnal cycle. For the 

most part, urban source patterns of oxidant precursors do appear to be reasonably constant; however, in most urban areas changes occur in traffic and , commercial emission patterns that are keyed to a weekday-weekend activity 

cycle. The effects of these changes have been observed in corresponding 

changes in ozone concentration patterns. 

Debate i n the 1960s over the rol e of ni tri c oxi de (NO) i n scavengi ng 

ozone led to the examination of whether weekday-weekend differences in ozone 

concentrations might occur in urban areas, on the assumption that NO emissions 

would be lower on weekends when some NOx-emitting sources virtually shut down. 

Altshuller (1975) reported the observation that no alerts.had ever occurred in 

the Los Angeles Basin on a Sunday. Likewise, Elkus and Wilson (1976), Levitt 

and Chock (1976), and Graedel et al. (1977) reported the existence of a weekdayweekend difference in ozone concentrations. While not reporting lower average 

ozone concentrations on weekends for the Los Angeles Basin, Schuck et al. 

(1966) noted a spatial shift of concentrations on summer weekends, away from 

the central commercial and urban areas to suburban and coastal areas. They 

5=48 
attributed the spatial shift in concentrations to changes in traffic patterns 

resulting from a shift from job-related travel to recreational travel. 

No newer reports of possible weekend-weekday variations are avaiiable in 

the literature, but any exhaustive exposure assessment should take into consideration the possibility that subtle differences in absolute concentrations or 

their spatial and temporal patterns could have some effect on total exposure. 

5.5.2 Spatial Variations in Ozone Concentrations 

Ozone is commonly thought of as a relatively homogeneous, regional pollutant, largely because of the nature of the sources of its precursors and the 

processes that contribute to its formation as a secondary pollutant. Numerous 

factors affect the occurrence of ozone in ambient air, however, so that differences in the true exposures sustained by respective receptors (e.g., human 

populations, crops, natural ecosystems) will vary within regions, airsheds, or 

other defined areas. The data presented in this section demonstrate a few of 

the known spatial variations in ozone concentrations that should be taken into 

account when assessing actual or even potential exposures. The data presented 

in this brief section are intended to be illustrative rather than exhaustive, 

since numerous meteorological, topographical, and physicochemical factors will 

influence the spatial and temporal distributions of ozone concentrations at 

specific sites. 

5.5.2.1 Urban versus Nonurban Variations. Data were presented in the 1978 

criteria document demonstrating that peak concentrations of ozone in rural 

areas are generally lower than those in urban areas, but that "dosages or 

average concentrations in rural areas are comparable to or even higher than 

those in urban areas" (U.S. Environmental Protection Agency, 1978). The 

diurnal concentration data presented in the preceding section indicate that 

peak ozone concentrations can occur later in the day in rural areas than in 

urban, with the distance downwind from urban centers generally determining how 

much later the peaks occur. The data presented in the preceding section for 

Montague, Massachusetts, in Figure 5-16 (Martinez and Singh, 1979) exemplify 

high late-afternoon secondary peak concentrations resulting from transport. 

The NAPBN and other nonurban data presented in Section 5.4 illustrate the 

urban-nonurban gradient in ozone concentrations that generally exists, and 

support the statement quoted above from the 1978 document. While corroboration 

of that statement would require the calculation of ineans or ppm-hours, the 

5-49 
data given in Section 5.4.2 tend to support that conclusion. For example, 

data presented in that section (Table 5-5) showed that rural sites of the SURE 

network were exposed, on the average, to daily maximum 1-hour ozone concentrations in excess of those to which the suburban sites were exposed (Martinez 

and Singh, 1979). While one rural site recorded 21 occurrences of 1-hour 

ozone concentrations >0.12 ppm during August through December 1977 (refei-red 

to below as A-D), the suburban sites recorded no such occurrences. A11 four 

of the rural sites in the SURE network recorded two or more occurrences (and 

one recorded 33 such occurrences) of daily maximum 1-hour concentrations >0.10 

ppm; only one of the five suburban sites recorded the occurrence (n = 10) of 

that concentration over the same period (A-D). A11 exposure statistics examined by Martinez and Singh (1979) for the SURE concentration data were higher 

at the rural sites, on the average, than at the suburban sites. The respective 

statistics, weighted by the number of observations in the data record, were: 

average of 24-hour mean concentrations (A-D), 0.028 ppm at rural versus 0.024 

ppm at suburban sites; mean of daily maximum 1-hour (A-D), 0.050 ppm at rural 

versus 0.044 ppm at suburban sites; and the 1-hour maximum for the whole 

period (A-D), 0.121 ppm at rural sites versus 0.094 ppm at suburban sites. 

In another example, data reported by Lefohn (1984) from sites classified 

as rural in the SAROAD data bank showed 9 occurrences of daily maximum 1-hour 

ozone concentrations >0.12 ppm at a rural Preble County, Ohio, site; and 19 

sucha occurrences at a rural site in Camden County, New Jersey. 

In addition to the occurrence of higher mean concentrations and occasionally higher peak conccentrations in nonurban areas than in urban, it is well 

documented that ozone persists longer in nonurban than in urban areas (Coffey 

et al. , 1977; Wolff et al ., 1977; Isaksen et al ., 1978). The absence of 

chemical scavengers appears to be the chief reason. 

5.5.2.2 Intracity Variations. Despite relative intraregional homogeneity, 

evidence exists for intracity variations in concentrations that are pertinent 

to potential exposures of human populations and to the assessment of actual 

exposures sustained in epidemiologic studies. Two illustrative pieces of data 

are presented in this section, one a case of relative homogeneity in a city 

with a population under 500,000 (New Haven, Connecticut) and one a case of 

relative inhomogeneity of concentrations in a city of greater than 9 million 

population (New York City) (U.S. Department of Commerce, 1982). 

5-50 
New Haven, Connecticut, was the site of an epidemiological study in 1976 

by Zagraniski et al. (1979). Symptoms recorded in diaries kept by study 

participants were correlated with ozone concentrations measured by the chemiluminescence method at a downtown New Haven site characterized as Center 

City-Residential. Table 5-10 shows several percentiles in the distribution of 

hourly values for that site plus two other monitoring stations in the county 

that were operating at the time, one in Derby, Connecticut, 9 miles west of 

New Haven, and one in Hamden, Connecticut, 6 miles north of New Haven. The 

Derby site also is characterized as Center City-Commercial, the Hamden site as 

Rural-Agricultural. The general similarity of values among the three sites 

appears to substantiate the New Haven data used in the epidemiological study 

since there was probably a reasonable temporal correlation between these close 

sites. Thus, wherever individuals might have traveled about the county, they 

probably were exposed to similar concentrations of ambient ozone. This conclusion is reinforced by the data in Table 5-11, showing the date and time of the 

maximum hourly concentrations by quarter at these three sites. The significant 

data are those for the second and third quarters when the potential for 0 3 

formation and for exposure was the greatest. Differences in peak concentrations 

varied from 0.006 ppm in the fourth quarter to 0.055 ppm in the third quarter 

among sites. 

The source of much of the ozone found in the New Haven, Connecticut, area 

is the greater New York City area (e.g., Wo1ff et al., 1975; Cieveland et al., 

1976a,b) and an urban plume transported over the distance from New York City 

to New Haven would tend to be relatively well-mixed and uniform, such that 

intracity variations in New Haven would probably be minimal. 

The highest or second-highest 1-hour maximum ozone concentration reported 

from a given station during a given year frequently gives an indication of the 

potential for repeated human exposure to high ozone levels. Nevertheless, a 

one-to-one correspondence between peak levels and either the number of days or 

the number of hours that a given level may be exceeded does not necessarily 

exist. Data obtained in the metropolitan New York area in 1980 illustrate 

this latter fact (Smith, 1981); and illustrate, as well, that intracity 

gradients can exist that should be taken into account in exposure assessment. 

The data given in Table 5-12 were obtained at the monitoring sites shown in 

Figure 5-21. The second-highest 1-hour ozone readings at the Eisenhower Park 

and Queens College stations have values only a few percentage points apart, 

5-51 
TABLE 5-10. OZONE CONCENTRATIONS AT SITES IN AND AROUND 

NEW HAVEN, CONNECTICUT, 1976 

(CHEMILUMINESCENCE METHOD, HOURLY VALUES IN ppm) 

% of values < stated concentration 

No. 

Site (SAROAD No.) measurements 50% 90% 95% 99% Max concn. 

New Haven, CT 4119 0.021 0.035 0.091 0.162 0.274 

(070700123F01) 

Derby, CT 5698 0.023 0.038 0.071 0.095 0.290 

(070190123F01) 

Hamden, CT 3853 0.030 0.045 0.075 0.098 0.240 

(070400001F01) 

Source: SAROAD (1985a). 

TABLE 5-11. QUARTERLY MAXIMUM 1-HOUR OZONE VALUES AT SITES 

IN AND AROUND NEW HAVEN, CONNECTICUT, 1976 

(CHEMILUMINESCENCE METHOD, HOURLY VALUES IN ppm) 

uarter of Year 

1 2 3 4 

New Haven , CT 

o. measurements 10 1964 2079 66 

Max 1-hr, ppm 0.045 0.274 0.235 0.066 

Hour of day 11:00 a.m. 2:00 p.m. 2:00 p.m. 10:00 P.M. 

Date 3/29 6/24 8/12 10/3 

Derby, CT 

No. measurements 11 2140 2187 1360 

Max 1-hr, ppm 0.015 0.280 0.290 0.060 

Hour of day 11: 00 p. m. 2: 00 p. m. 2: 00 p. m. 7: 00 p. m. 

Date 3/31 6/24 8/12 12120 

Mamden, CT 

No. measurements 56 2065 1446 286 

Max 1-hr, ppm 0.050 0.240 0.240 0.065 

Hour of day Noon 3:00 p.m. 1:00 P.M. 3:00 p.m. 

Date 3/29 6/24 7/20 10/7 

Source: SAROAD (1985a). 

5-52 
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5-53 
SITES 

1. Susan Wagner High School 

2. Mabel Dean High School 

3. Woolsey Post Office (Astoria) 

4. Mamaroneck 

4 

5. Public School 321 BRONX 

6. Sheepshead Bay High School o 

7. Queens College ~ 

8. Eisenhower Park (Nassau Co.) Z, . 

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NEW JERSEY 2 **~ C76 

QUEENS 8 o 

05 

STATEN BROOKLYN ISLAND 

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Figure 5-21. New York State air monitoring sites for 

Northeast Corridor Monitoring Program (NECRMP). 

Source: Smith (1981). 

5-54 
yet there were 51 hours of ozone concentrations exceeding 0.12 ppm and 15 days 

when ozone levels exceeded 0.12 for at least 1 hour at the Queens College 

Station; whereas corresponding values were recorded at Eisenhower Park for 

only 7 hours during 2 days. At both stations, data for about 94 percent of 

possible hours were recorded as valid.. 

The range of first-, second-, third-, and fourth-highest values, along 

with frequencies'of values >0.12 ppm, establishes an apparent-concentration 

gradient in the area from sites 6 and 5 to site 4. Exposures- of human populations living and working in metropolitan New York City could di'ffer appreciably 

if the residences were located and all activities were centered in lower. 

Brooklyn as opposed to the upper Bronx. Differences in peak concentrations at 

the respective sites varied by date (6/14 to 8/28) and by ievel on the same 

day (8/28), when ozone was 0.080 ppm at site 4 and 0.174 at site 7, a difference of 0.094 ppm (a factor of 2.18). 

Intracity gradients in ozone concentrations have also been reported by 

Kelly et al. (1986) for a 1981 study in Detroit, MI. Ozone concentrati,ons 

were measured for about 3 months at 16 sites in the metropolitan Detroit area 

and in nearby Ontario, Canada. Values at 15 sites were correlated with those 

at a site adjacent to the Detroit Science Center, about 3 km north of the 

central business district 'in Detroit. In general, the correlation decreased 

as distance from the Science Center site increased; and, in general, the 

actual concentrations increased with distance from that site toward the northnortheast. The highest ozone concentrations were recorded at sites about 10 

to 70 km north-northeast of the urban core. At greater distances or in other 

directions, ozone maxima decreased. 

5.5.2.3 Indoor-Outdoor Concentration Ratios. Most people in the United 

States spend a large proportion of their time indoors. Although knowiedge of 

actual exposures of populations to ozone is essential for optimal interpretation and use of the results of epidemiological studies, essentially all air 

pollution monitoring is done on outdoor air. The modeling of actual exposures, 

as opposed to potential exposures, therefore necessitates knowing general 

activity patterns and at least approximate indoor/outdoor ratios (I/0) of 

ozone. 

For slowly reacting compounds such as carbon monoxide, long-term average 

ratios of indoor to outdoor concentrations tend to be close to unity, in the 

absence of indoor sources. 'Over short time periods, however, the I/0 may be 

5-55 
significantly different because of non-equilibrium factors (Yocom, 1982). The 

situation for reactive pollutants such as ozone is much more complex, and 

reported I/0 values for ozone are highly variable. Unfortunately, the number 

of experiments and kinds of structures examined to date provide only limited 

data for use in modeling indoor exposures. Yocom (1982) has presented a 

chronological summary of studies in which either ozone or photochemical oxidant 

indoor-outdoor gradients in buildings and residences were measured. Studies 

summarized by Yocom have been conducted over the period 1971 through the 

present (one ongoing study) by five research organizations: University of 

California, the California lnstitute of Technology, GEOMET, Inc., Lawrence 

Berkeley Laboratory, and TRC Environmental Consultants. Structures examined 

have included hospitals, schools, office buildings, single-dwelling homes, 

"experimental" dwellings, apartments, and mobile homes. Private homes included 

those with and without gas stoves and fireplaces, and those inhabited by 

smokers versus nonsmokers. Areas of the country in which the buildings were 

located ranged from Southern California to Boston, including such cities in 

between as Denver, Chicago, Washington, Baltimore, Pittsburgh, as well as 

other unspecified locations (Yocom, 1982). The indoor-outdoor ratios reported 

from these studies are summarized in Table 5-13 and are discussed later. 

Among newer reports of indoor/outdoor gradi`ents in ozone concentrations, 

also summarized in Table 5-13, are the studies of Stock et al. (1983) and 

Contant et al. (1985), undertaken to provide exposure assessments for an 

epidemiological study of asthmatics in Houston, Texas, in 1981 (see Holguin et 

al., 1985, in Chapter 11). Stock et al. (1983) found I/0 ratios of nearly 

zero to 0.09 in one air-conditioned residence (maximum ozone concentration of 

5 ppb) and an I/0 ratio of 1.0 in a residence ventilated completely by outdoor 

air. (Indoor concentrations were monitored via a sampling manifold connected 

by Teflon lines to a chemiluminescence analyzer in a mobile van parked near 

the respective houses.) Contant et al. (1985), in a continuation and extension 

of the same study, used "personal monitors" (i.e., portable analyzers) to 

measure indoor and outdoor air in the immediate environs of participants in 

the asthma study. Mean, median, and maximal ozone concentrations, respectively, 

were 10.8, 5, and 147 ppb indoors and 51.8, 42, and 250 ppb outdoors. The 

respective I/0 ratios were 0.21, 0.12, and 0.59. 

Davies et al. (1984) found I/0 ratios of 0.7 ± 0.1 at an art gallery in 

rural Norwich, England, where the outside ozone concentrations ranged from 18 

5-56 
TABLE 5-13. SUMMARY OF REPORTED INDOOR-OUTDOOR OZONE RATIOS 

Indoor-outdoor 

Structure ratio (I/0) Reference 

Residence 0.60a Thompson et al. (1973) 

(with evaporative cooler) 

Office Sabersky et ai. (1973) 

(air-conditioned; 100% outside 0.80 ± 0.10 

air intake) 

(air-conditioned; 70% outside 0.65 ± 0.10 

air intake) 

Residence 0.70 Sabersky et.al . (1973) 

Residence 0.50 to 0.70 Moschandreas et al. (1978) 

Residence Moschandreas et al. (1981) 

(gas stoves) 0.19 

(all electric) 0120 

Office 0.29 Moschandreas et al. (1978) 

School room 0.19 Berk et al. (1980) 

(maximum concn.) 

Residence 0.10 to 0.25 Berk et al. (1981) 

Residences (1 each) Stock et al. (1983) 

(air-conditioned) 0.00 to 0.09 

(100% outside air; no 1.0 

air conditioning) 

Residences (12) 0.21 (mean concn.) Contant et al. (1985) 

(air-conditioned) 0.12 (median concn.) 

0.59 (maximum concn.) 

Art gallery 0.70 + 0.10 Davies et al. (1984) 

(three modes of ventilation (mean concn.) 

in each 24-hr period: 

recirculation, mixture of 

recirculated and outside 

air, and 100% outside air) 

aMeasured as total oxidants. 

5-57 
to 58 ppb in the 3-week study period. The interior of the modernistic gallery 

had few reactive surfaces (two plate glass walls and two walls and ceiling of 

painted aluminum, with minimal other walls and furnishings). In addition, 

ventilation cycled between recirculation air, a mixture of outside-and recirculation air, and 100 percent outside air over each 24-hour period. 

The results of all the studies cited above are shown in Table 5-13. 

These results are highly variable, to say the least. The variability is not 

surprising, however, considering the diversity of structures and locations 

included in the studies. In Table 5-13, the highest I/0, 1.0, was reported 

for a residence ventilated by outside air (Stock et al., 1983). The secondhighest I/0, 0.80, was determined by smog-season measurements made in a multi- 

-story, air-conditioned building on the Pasadena campus of the California 

Institute of Technology (Sabersky et al., 1973). Air exchange in this building 

was at a rate of 10 changes per hour with 100 percent outside air (i.e., no 

recirculation of inside air). For another Cal Tech building, in which there 

was a mixture of 70 percent outside air and 30 percent recirculated inside air, 

Sabersky et al. (1973) found an indoor-outdoor ozone ratio of 0.65. The 

lowest indoor-outdoor ozone concentration ratios shown in Table 5-13 are those 

reported by Stock et al. (1983) for one residence; the indoor concentration 

recorded is barely above the limit of detection of the analyzer. 

A relatively large number of factors can affect the difference in ozone 

concentrations between the inside of a structure and the outside air. In 

general, outside air infiltration or exchange rates, interior air circulation 

rates, and interior surface composition (e.g., rugs, draperies, furniture, 

walls) affect the balance between replenishment and decomposition of ozone 

within buildings (Thompson et al., 1973; Sabersky et al., 1973; Berk et al., 

1980; Moschandreas et al., 1978). The rate at which exterior air enters a 

building depends on local wind speed and direction, on how well-sealed the 

building is, on how frequently doors and windows are opened, and on the operating chariacteristics and cycles of heating/air conditioning/ventilating systems. 

A significant factor that increases infiltration is an increasing temperature 

differential between warm interior air and cold outside air (Moschandreas et 

al., 1978), although a high differential would be unusual for a photochemical 

smog period. Moschandreas et al. (1978) reported exterior-interior exchange 

y rates ranging from ten changes per hour in an office building to one change 

every 5 hours in a residence. At the higher exchange rates, inducted ozone 

5-58 
remains at a level indoors that is closer to the outdoor level. As the exchange 

rate decreases, surface decomposition processes can result in progressively 

lower equilibrium ozone concentrations. Other factors, such as relative 

humidity, also affect decomposition. The half-life of ozone inside residences 

has been estimated at 2 to 6 minutes (Moschandreas et al., 1978; Mueller 

et al., 1973; Sabersky et al., 1973), while its half-life in an office environment has been estimated at 11 minutes (Mueller et al., 1973). These results 

are indicative of the relatively rapid reaction rate that can be expected for 

ozone in a building or room environment. The problem of I/0 values in buildings 

was the subject of a model development program by Shai•r and Heitner (1974) in 

which they tried to account for ventilation and for losses by reactions and 

surface scavenging. Considering the research results shown in Table 5-13 and 

summarized by Yocom (1982), any estimates of indoor ozone exposures must be 

considered as having a large degree of probable variability. 

Automobiles and other vehicles constitute another indoor environment in 

which people may spend appreciable amounts of time. As with buildings, the 

mode of ventilation and cooling helps determine the inside concentrations. 

Peterson and Sabersky (1975) reported ozone measurements made inside and 

outside a car traveling a freeway in southern California. Concentrations 

inside were higher when the air-conditioning was ope.rating on "maximum" than 

when it was operating on "normal," largely because at higher air turnover 

rates ozone intake exceeded ozone decay. When windows'were opened, an equilibrium between inside and outside ozone concentrations was established in about 

3 minutes, with inside concentrations then remaining significantly lower than 

outside concentrations as the result of the exponential decay of ozone from 

contact with interior surfaces. Inside ozone concentrations were one-third or 

less of the outside concentrations (I/0 < 0.33), on the average, under all 

ventiiation and cooling conditions. Ozone concentrations measured outside the 

car were lower than those measured at nearby fixed sites, a result attributed 

to scavenging of ozone by nitric oxide. In the exposure study of Contant et 

al. (1985) in Houston, I/0 ratios from 49 measurements inside vehicles were 

0.44 for mean, 0.33 for median, and 0.56 for maximum concentrations measured. 

Driving routes and ventilation and cooling modes were not described. 

Concentrations of ozone inside vehicles obviously depend first on the 

concentration outside and then on the various effects of cooling and ventilation 

conditions and surface decay. Thus, the titration of 0 3 by NO on and near 

5-59 
roadways is an important factor. The titration reaction is rapid but depends 

on the amount of NO emitted, which in turn depends on the amount of traffic on 

a given road or freeway. Rhodes and Holland (1981) found average ozone levels 

on the nine-lane freeway between Los Angeles and San Diego, which carries 

200,000 cars each day, to be one-ninth to one-sixth the levels found 30 meters 

upwind. 

At present there are no long-term monitoring data on indoor air pollutant 

concentrations that are comparable to the concentration data available for 

outdoor locations. Thus, for estimates of the exposure of building or vehicle 

occupants to ozone and other photochemical oxidants, it is necessary to rely 

on extrapolations of very limited I/0 data such as those presented above and 

in Table 5-13. 

5.5.2.4Altitudinal and Latitudinal Variations . Concentrations of ozone vary 

with altitude and with latitude. These variations occur for a number of 

reasons, including the nature of the interchange mechanisms involved in stratospheric-tropospheric exchange, the decay of stratospheric ozone as it traverses 

the troposphere, and the known production of ozone in the apparently unpolluted 

free troposphere at certain altitude ranges above mean sea level (MSL). In 

addition, other meteorological, physical, and chemica1 factors contribute to 

the concentration gradients found with changes in altitude and latitude at the 

surface of the earth. 

Among specific factors known to contribute to variations in ozone concentrations with altitude and latitude, and, to a much lesser extent, longitude, 

are the following: 

1. Incursions of stratospheric ozone; 

2. Global circulation patterns, along with accompanying altitudinal and 

geographic differences in atmospheric pressure, direction of prevailing winds, wind speeds, temperature, humidity, and precipitation; 

3. Intensity and spectral distribution of sunlight; and 

4. Meteorological factors, such as mixing heights, frequency and persistence of inversions, upslope flows in mountainous areas, and cloud 

cover and precipitation patterns. 

Most of these factors were discussed in Chapter 3, as well as in the 1978 

criteria document (U.S. Environmental Protection Agency, 1978). A few are 

briefly noted here to present an overview of the macroscale differences in 

ozone concentrations that may be expected. Concentration gradients found with 

increases in altitude at the earth's surface are presented in more detail 

because of their relevance for specific kinds of exposure assessments. 

5-60 
Latitudinal differences in ozone concentrations have been demonstrated. 

These are in part attributable to the variations, temporal as well as geographic, in actinic irradiance at the earth's surface. Both the intensity and 

the spectral distribution of sunlight have direct effects on the photochemical 

reactions that initiate and sustain oxidant formation (Chapter 3 and U.S. 

Environmental Protection Agency, 1978). The effect of latitude on ozone 

formation is significant during the winter but not during the summer, since in 

summer, for the contiguous United States, the maximum light intensity is 

fairly constant and the duration of the solar day varies only slightly with 

latitude. Light intensity varies somewhat with longitude in the summer, with 

the highest intensities occurring in the western states. 

Incursions of stratospheric ozone also contribute to variations in ozone 

concentrations with latitude. These are expected to be strohgest in the 

mid-latitudes of the northern hemisphere because of the nature of the mechanisms 

by which stratospheric-tropospheric exchange occurs. 

The 1978 criteria document presented discussions of the effects of tropopause-folding events (TF) and of the seasonal tropopause adjustment (STA) and 

small-scale eddy transport (SSET) mechanisms on stratospheric-trospheric 

exchange: As described previously in Chapter 3 and in the 1978 document, TF 

events would be expected to produce sporadic increases in ground-level ozone 

concentrations, resulting from strong incursions of stratospheric ozone, in 

the southern and eastern United States (latitudes of < 37 ° N and longitudes of 

< 90° ) (U.S. Environmental Protection Agency, 1978). The STA-produced incursions also occur in these latitudes. Both STA and TF produce incursions in 

the winter or early spring. 

Logan et al. (1981) summarized earlier measurements of ozone in the lower 

troposphere. Hemispheric asymmetry is apparent in the data they presented, 

with higher concentrations occurring in the northern hemisphere than in the 

southern. Also apparent is a seasonal increase in iower-tropospheric ozone in 

the summer at mid-latitudes in the northern hemisphere. The authors noted, 

however, that ozone data for the lower troposphere are sparse, especially for 

the southern hemisphere. Compared with ozone data from two east-coast sites 

at mid-latitudes in the northern hemisphere, ozone data from one site in 

Australia showed (1) iittle seasonal va.riation at <3 km (MSL), (2) about 40 

percent lower concentrations at 2 km (MSL) in summer, and (3) similar concentrations in autumn and winter. Altitudinal profiles of ozone concentrations 

5-61 
for the tropical northern hemisphere (Panama, 9 ° N, and Hawaii, 19.5°N) show a 

considerably lower gradient with altitude than that seen at mid-latitudes 

(Oltmans, 1981; cited in Logan et al., 1981). Logan et al. (1981) noted the 

difficulty, given the sparseness of the data, of separating hemispheric effects 

from variations in concentrations attributable to other fac-tors (e.g., oceanic, 

coastal, inland meteorology). 

In a more recent report, the same investigators presented other data 

showing hemispheric differences in ozone concentrations and the seasonal 

distributions of those concentrations (McElroy et al., 1985). Ozone concentrations in the lower troposphere measured at Cape Grim, Tasmania (ca. 41°S), 

showed strong similarities, including seasonal distributions, to those measured 

at three different sites, at different longitudes, in Canada (latitudes of ca. 

530N to 590N). Ozone concentrations from sites in West Germany, Switzerland, 

and the United States showed even greater similarity, both in level and seasonal 

distribution. 

Although of interest and concern when estimating global ozone budgets, 

variations in ozone concentrations with latitude have little practical significance for assessing exposure within the contiguous United States. The effects 

on ozone concentrations of latitude, as well as longitude, within the contiguous 

states are minor. 

A number of additional reports are available on the increase of tropospheric 

ozone concentrations with altitude (e.g., Kroening and Ney, 1962; Galbally and 

Roy, 1980). The data cited above showing latitudinal effects also show altitudinal effects, particularly at mid-latitudes in both the northern and southern 

hemispheres (Oltmans, 1981; cited in Logan et al., 1981). At mid-latitudes, 

the increase in background ozone concentrations rises rather sharply (from a 

spring-summer mixing ratio of about 40 to 60 ppb to one of about 100 ppb) 

between altitudes of about 8 and 12 km (ca. 5 to 7.4 mi, or ca. 26,400 to 

39,600 ft). Between the surface and about 2 km (ca. 1.2 mi, or ca. 6600 ft), 

another relatively sharp increase in ozone concentrations is observed for the 

mid-latitudes (from a spring-summer mixing ratio of about 35 ppb to one of 

about 50 to 55 ppb) (Oltmans, 1981; cited in Logan et al., 1981). Other data 

corroborate these findings. Seiler and Fishman (1981) reported ozone measurements taken on flights in remote tropospheric air during July and August 1974. 

Their data also show that ozone concentrations increase with increasing altitude 

and in general substantiate the accepted view that the lower atmosphere and 

the surface of the earth act as ozone sinks. 

5-62 
Increases in ozone concentrations at higher altitudes are of potential 

concern relative to exposures sustained in-flight by airline passengers and 

employees on high-altitude flights (see Chapter 11). Air-filtration systems 

are employed on airplanes, however, that reduce ozone concentrations. Increases 

in concentrations with altitude at lower altitudes (e.g., in the zero to 2 km 

range) have potentially greater importance, since these gradients may affect 

forest ecosystems in mountainous areas. (It must be borne in mind, however, 

that atmospheric pressure decreases with altitude and thus, for a given concentration, the mass of ozone per cubic meter of air also decreases with altitude.) 

Contributors to increased ozone concentrations at the higher altitudes of 

relevance for forest ecosystems include stratospheric ozone intrusions and ozone 

transported aloft and conserved overnight in nocturnal inversions layers. The 

latter, in particular, may be an important consideration for the accurate 

assessment of the ozone exposures sustained by mountain forest ecosystems, as 

discussed below (see Chapter 7 also). 

While a number of reports contain data on ozone concentrations at high 

altitudes (e.g., Coffey et al., 1977; Reiter, 1977b; Singh et al., 1977; 

Evans, 1985; Lefohn and Jones, 1986), fewer reports are available that present 

data for different elevations at the same locality. Studies by Berry (1964) 

and Miller and Ahrens (1969) are among earlier reports documenting ozone 

gradients with altitude in forested areas. Two more recent studies are discussed here because the investigators acquired and presented data on the 

diurna1 patterns of ozone at those sites, as well as on differences in ozone 

concentrations at the respective altitudes. In addition, the two studies were 

conducted in different parts of the country and presented data that demonstrate 

differences in the meteorological influences on the concentrations found in 

the respective studies. 

The mixed-conifer forest ecosystem of the San Bernardino Mountains in 

California is the most thoroughly studied forest ecosystem in the United 

States (see Chapter 7). In conjunction with vegetation and forest ecosystem 

studies conducted there in the 1960s and 1970s, a number of monitoring stations 

were established at different elevations on the south-facing slopes of the San 

Bernardino Mountains. The mountain range begins about 80 miles east of Los 

Angeles and its center is almost due east of Pasadena. The first four monitoring stations were established at Highland, City Creek, Mud Flat, and Rim 

Forest, located as shown in Figure 5-22. 

5-63 
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y2135 

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305 

0 

0 4 8 12 16 20 24 28 32 36 

DISTANCE, kilometers 

Figure 5-22. Altitudinal sequence of monitoring sites in the San 

Bernardino Mountains. 

Source: Adapted from Miller et al. (1972). 

5-64 
Total,oxidant, temperature, and reiative humidity (RH) were measured 

continuously during 16 days in July and August at all but the Highland station 

(Miller et al., 1972). The vapor pressure gradient was calculated from RH and 

temperature data. The time of the daily peak oxidant concentration was progressively later at stations of higher elevation, as shown in Figure 5-23. 

Temperatures and vapor-pressure gradients were also progressively lower at 

higher elevations at the time of the oxidant peak. The.average duration of 

oxidant concentrations exceeding 200 pg/m 3 (0.10 ppm) was 9,'13, 9, and 

8 hr/day going from lower to higher stations for June, July, and August 1969. 

The longer duration at City Creek (elevation 817 m; 2680 ft) r probably coincided 

with the zone where the inversion layer most often contacted the mountain slope. 

The oxidant concentrations rarely decreased below 98 pg/m 3 (0.05 ppm) at night 

on the slope of the mountain crest, whereas they usually decayed to near zero 

at the basin station (Highland, elevation 442 m; 1459 ft). 

The vertical and horizontai distributions of oxidant air pollution in the 

Los Angeles Basin have been described by several investigators and presented 

in the 1978 criteria document (Blumenthal et al., 1974; Edinger, 1973; Edinger 

et al., 1972; Miller et al., 1972; as cited in U.S. Environmental Protection 

Agency, 1978). In Los Angeles, a marine temperature inversion layer frequently 

forms above the heavily urbanized metropolitan area and often extends inland 

as far as 144 km (90 miles), depending on season and time of day. Surface 

heating of air under the inversion increases with distance eastward in the 

basin and often disrupts the inversion by roidmorning at its eastern edge. The 

marine temperature inversion layer encounters the mountain slopes, usually 

below 1200 m(4000 ft). In the morning, the temperature inversion often 

remains intact at this juncture, and air pollutants are confined beneath it. 

The heated mountain slopes act to vent oxidant air pollutants over the crest 

of the mountains and cause the injection of pollutants into the stable inversion 

layer horizontally away from the slope. Oxidant concentrations within the 

inversion are not uniform, but occur in multiple layers and strong vertical 

gradients. In some cases, the inversion may serve as a reservoir for ozone, 

which may arrive at downwind locations along the mountain slopes relatively 

undiluted because .of a lack of vertical mixing within the inversion layer and 

a lack of contact with ozone-destroying materiai generated at the ground. The 

important result .of the trapping of oxidant in these iayers is its prolonged 

contact with high ,terrain at night. 

5-65 
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total oxidant concentrations, temperature, and vapor pressure at 

four sites (A-D) in the San Bernardino Mountains, CA, July-August 

1969. 

Source: Miller et al. (1972). 

5-66 
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Figure 5-23 (Cont'd). Relationship between elevation and diurnal 

patterns of total oxidant concentrations, temperature, and vapor 

pressure at four sites (A-D) in the San Bernardino Mountains, CA, 

July-August 1969. 

Source: Miller et a6. (1972). 

5-67 
The studies in the Los Angeles Basin and the nearby San Bernardino Mountains underscore the importance of understanding the interactions between 

topography and local meteorology for specific areas. Consideration of those 

factors is essential for accurate exposure assessments at higher altitudes, 

especially in the absence of monitoring sites and adequate aerometric data. 

Wolff et al. (1986) conducted a study of the effects of altitude on ozone 

concentrations at three sites located at three separate elevations on High 

Point Mountain in northwestern New Jersey. Ozone concentrations and temperature 

were measured at ground level at the respective sites. In addition, temperature 

and relative humidity were measured above the surface by balloon-borne sensors, 

with wind direction and speed data obtained by means of a tracking theodolite. 

Data for several days (July 1975) indicate that in mid-day, when atmospheric mixing was good, vertical profiles were nearly constant, with concentrations increasing only slightly with elevation. Likewise, the daily ozone 

maxima were similar at different elevations. At night, however, ozone concentrations were nearly zero in the valley (the lowest-elevation site) and increased with elevation. Comparison of the ozone dosages at the three sites 

(number of hours > 0.08 ppm) showed that greater cumulative doses were sustained 

at the higher elevations (Table 5-14). Comparable data from an urban area 

(Bayonne, NJ) about 80 km southeast of High Point Mountain showed that the 

cumulative doses were higher at all three of the mountain sites than in the 

urban area (Table 5-14) (Wolff et al., 1986). 

TABLE 5-14. COMPARISON OF OZONE CONCENTRATIONS AT THREE DIFFERENT 

ELEVATIONS, HIGH POINT MOUNTAIN, NJ, AND 

AT BAYONNE, NJ, JULY 1975 

1-hr 24-hr 1-hr 24-hr 

max, mean, max, mean, No. hr 

Location/elevation ppb ppb ppb ppb >80 ppb 

High Point, 500 m 66 49 130 81 13 

`High Point, 300 m 61 38 110 61 9 

High Point, 140 m 59 26 120 52 9 

Bayonne, sea level 69 33 119 48 7 

Source: Wolff et al. (1986). 

.: 
The maximum depth of the nocturnal inversion at High Point Mouritain was 

about 900 m(ca. 2970 ft). Extrapolating from the gradient observed from the 

ground sites, Wolff et al. (1986) estimated that mountain peaks exceeding 900 

m(2970 ft) would sustain exposures to relatively constant diurnal concentrations that would approximate the maximum ozone concentration found at the 

valley site. The investigators concluded from their concentration and meteorological data that elevated, mountainous sites in the eastern United States may. 

-'`be expected to be exposed to higher ozone dosages than valley sites throughout 

= the year. 

5.5.2.5Vertical Gradients at Ground Level . Just as macroscale variations in 

ozone concentrations have been observed by measuring vertical and horizontal 

profiles at various altitudes, so too have microscale vertical variations at 

ground level been observed as a function of placement of sampling probes. 

Data drawn from a recent study on rural ozone concentrations illustrate the 

effects on concentration data of the placement of monitoring probes. These 

effects are pertinent for vegetation exposures, in particular, although the 

vertical gradient at ground level has obvious implications for human exposures 

as we11. 

Pratt et ai. (1983) studied concentrations of ozone and oxides of nitrogen 

in the upper-midwestern part of the United States. Concentration data were 

obtained over 4 years by means of monitors placed at two di:fferent sampling 

heights (ca. 3 and 6, or 3 and 9 meters) at three air quality monitoring 

sites: LaMoure County, North Dakota; Traverse County, Minnesota; and Wright 

County, Minnesota. All stations were rural sites. The mean ozone concentrations did not differ greatly among the sites, but in at least some instances 

the mean differences between sampling heights were as.large or larger than the 

differences among the scattered sites. Table 5-15 presents the mean ozone 

concentrations measured at two separate sampling heights (Pratt et al., 1983). 

Annual average concentrations were 1 to 3 ppb lower at the 3.05-meter height 

than at the 6.10 or 9.14-meter heights, reflecting the depletion of ozone near 

the surface. As might be expected, the gradient:was especialiy conspicuous at 

night because of continued surface scavenging and a decrease in the rate of 

transfer from layers aloft. The concentrations of ozone occurring at these 

sites were near background in all years measured. In areas with higher ozone 

concentrations, one would expect to see larger absolute gradients between 

5-69 
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5-70 

~ 
monitors at different heights. In fact, the careful measurement of concentration gradients over 'di,stances"of l to 10 meters above a surface is a recognized 

method for estimating the scavenging potential'of the surface. 

In this context, it should be noted that the dry deposition of ozone on 

various surfaces has been fairly extensively studied. Summaries on surface 

scavenging and other natural removai processes appeared in the previous criteria 

document for ozone and other photochemical oxidants (U.S. Environmental Protection Agency, 1978) and in the review by-the National Research Council (National 

Research Council, 1977). Newer review articles on dry deposition and surface 

scavenging have been published by Wesely (1983) and Galbally and Roy (1980), 

among others. 

5.6 CONCENTRATIONS OF PEROXYACETYL NITRATE AND PEROXYPROPIONYL NITRATE IN 

AMBIENT AIR 

5.6.1Introduction 

As noted in the introduction to this chapter (Section 5.1), pub1ished 

data on.the concentrations in ambient air of photochemical oxidants other than 

ozone are not comprehensive or abundant. Much more is known now, however, 

about their atmospheric•concentrations than was known when the 1978 criteria 

document for ozone and other photochemical oxidants was published. Review of 

the data that follow will show that peroxyacetyl.nitrate (PAN) and peroxypropionyl nitrate (PPN) are the most abundant of.the non-ozone oxidants in ambient 

air in the United States other than the inorganic nitrogenous oxidants such as 

nitrogen dioxide (NO 2 ); and possibly nitric acid (HNO3), in some areas. The 

inorganic nitrogenous oxidants founci in ambient.air are reviewed in air quality 

criteria documents on the nitrogen oxides,and are not treated in this document 

except for the review of the role of nitric oxide (NO) and NO2 in atmospheric 

photochemistry. 

Available, data show that PAN is toxic to vegetation and is potentiaily 

toxic to humans. No data indicating potentially adverse effects of PPN are 

available. It exists in trace quantities and appar.ently no research has ever 

been undertaken to :determine its potential toxicity. At least one study 

(Heuss and Glasson; 1968) has reported that a higher homologue of the series, 

peroxybenzoyl nitrate (PBzN), like PAN, is a lachrymator. No unambiguous 

identification of ~PBzN in the ambient air of the United States has been made, 

however. 

5-71 
Given the information available on PAN, the concentrations of PAN that 

are of most concern are those to which vegetation could potentially be exposed, 

especially during daylight hours in agricultural areas; followed in importance 

by those both indoors and outdoors, in both urban and nonurban areas, to which 

human populations could potentially be exposed. Most of the available data on 

concentrations of PAN and PPN in ambient air are from urban areas. The levels 

to be found in nonurban areas will be highly dependent upon the transport of 

PAN and PPN or their precursors from urban areas, since the concentrations of 

the NOxprecursors to these compounds are considerably lower in nonurban than 

in urban areas. 

6 The following sections present historical and recent data on the cdncentrations of PAN and, where available, on PPN, as well as the patterns those 

concentrations assume. 

5.6.2 Historical Data 

In the 1970 criteria document for photochemical oxidants (U.S. Department 

of Health, Education, and Welfare, 1970), concentrations for total oxidants 

and PAN were reported for Los Angeles and Riverside, California. In Los 

Angeles, composite diurnal data for PAN (obtained by GC-ECD) showed average 

peak 1-hr concentrations of about 40 ppb in September 1965 and about 60 ppb in 

October 1965. The September peaks occurred around noon and the October peaks 

occurred shortly after 1:00 p.m., with the difference in time possibly being a 

function of the temperature-dependence of PAN formation and persistence. In 

Riverside, two peak concentrations were observed in composite diurnal data, 

one of > 4 ppb around 10:00 a.m. and one of —10 ppb between 4:00 and 6:00 p.m. 

Seasonal data from Riverside for June 1966 to June 1967 showed that PAN concentrations were highest in September 1966 and in March and June 1967. 

Total oxidants (by Mast meter) in these Los Angeles sites reached a peak 

concentration (the same composite diurnal data as above) of close to 140 ppb 

in September 1965 at about the same hour of day as the PAN peak. In October, 

total oxidants peaked at nearly 200 ppb, again coinciding in time with peak 

PAN concentrations. In Riverside, the morning PAN peak preceded the oxidant 

peak (-110 ppb around noon) by almost 2 hours but the afternoon PAN peak 

trailed the afternoon oxidant peak (-160 ppb around 2:00 to 4:00 p.m.) by 

about 2 hours. The ozone/PAN ratio was variable from month to month but was 

generally lower from November through April than during the rest of the year, 

5-72 
i.e., PAN was generally a greater percentage of the totai oxidant during the 

November through April period. 

In the 1978 criteria document for ozone and other photochemical oxidants 

(U.S: Environmental Protection Agency, 1978), additional PAN data for Los 

Angeles as well as two other cities were presented. Lonneman et al. (1976) 

measured PAN and total oxidants in Los Angeles in 1968. In 118 samples collected for the period 10:00 a.m. to 4:00 p.m. during the study, the median PAN 

concentration was 13 ppb and the average PAN concentration was 18 ppb. The 

median total oxidant concentration (measured by UBKI) was 97 ppb and the 

average was 117 ppb. Thus, the median oxidant/PAN ratio was 7.5 and the average 

oxidant/PAN ratio was 6.5; the median PAN concentration was 13.4 percent of 

the median total oxidant concentration, and the average PAN concentration was 

15.4 percent of the average totai oxidant concentration. 

Lonneman et al. (1976) conducted similar studies in Hoboken, New Jer'sey; 

in 1970 and in St. Louis, Missouri, in 1973. Samples were measured during the 

period 10:00 a.m. to 4:00 p.m. over the course of the study. In Hoboken, PAN 

concentrations averaged 3.7 ppb. Ozone concentrations (measured by chemiluminescence) averaged 90.5 ppb. In St. Louis, PAN averaged 6.4 ppb, ozone (measured by chemiluminescence) averaged 50.1 ppb, and total oxidants (by UBKI) 

averaged 74.3 ppb. 

From these 1966 through 1973 urban data, it is clear that PAN conGentrar 

tions in urban areas are appreciably lower than those of ozone, even in , the 

winter when PAN is higher relative to ozone than in summer. 

In addition to urban data, the 1978 criteria document (U.S. Environmental 

Protection Agency, 1978) also included PAN concentration data from one nonurban 

agricultural area, Wilmington, Ohio (Lonneman et al., 1976). The maximum PAN 

concentration observed in 1500 samples taken during August 1974 was 4.,1 ppb. 

The daily maximum PAN concentration rarely exceeded 3.0 ppb even though the 

daily maximum ozone concentration frequently exceeded 80 ppb. 

Conclusions reached in the 1978 document were (1) that PAN concentrations 

are much lower in the ambient air of nonurban than of urban areas; and (2) 

that ozone/PAN or total oxidant/PAN ratios vary with location, such ratios 

being higher in nonurban areas than in urban; that is, PAN concentrations are 

a lower percentage of total oxidant concentrations in nonurban areas (U.S. 

Environmental Protection Agency, 1978). From data presented in the 1970 

document, it may be concluded (1) that oxidant/PAN ratios vary with season; 

5-73 
(2) that PAN concentrations are lower than ozone concentrations in urban 

areas; and (3) that PAN and ozone concentrations exhibit similar but not 

identical diurnal patterns. 

The historical data presented above have been given in some.deta,il because 

the information about,PAN conveyed by those data remains valid.. Examination 

of the more recent data presented in subsequent sections shows that newer data 

extend and corroborate, for the most part, the findings of the older literature. 

5.6.3 Ambient Air Concentrations of,PAN .and Its Homologues in Urban Areas 

Additional studies on concentrations of PAN and its homologues in both 

urban and nonurban areas are now available. Newer data are presented in this 

section in summary form, where possible; and the'individual studies or examples 

presented are merely a few of many, chosen to represent currerit knowledge 

regarding ambient air PAN concentrations and their patterns. 

The existing literature on the concentrations in ambient ai•r of the peroxyacyl nitrates, PAN and its higher homologues, has been compiled and examined 

in two recent review articles (Temple and Tay1or, 1983; Altshuller, 1983). In 

the first, Temple and Taylor reviewed the concentrations of PAN'in the ambient 

air in Europe, Japan, and North America in the context of the phjitotoxicity of 

PAN. Altshuller, in the second, reviewed the published concentrations of PAN 

and of PPN i n ambi ent ai r, al so wi thi n and outsi de the Uni ted `States. 'I n 

addition, Altshuller analyzed the relationships to'ozone of PAN and other 

photochemical reaction products. The reader is referred to these reviews for 

detailed information`and for references therein. 

Table 5-16 presents a summary of PAN concentrations obse_rved in the 

ambient air of urban areas of the.United States. Data in this table include 

the results of studies cited in Section 5:6.2 above, along with the results of 

newer studies. This table was derived from the reviews of Altshuller (1983) and 

Temple and Taylor (1983), as well as from a few additional sources (Jorgen et 

al ., 1978; Lewi s et al ., 1983) ., The data are summari zed i n the tabl e by 

region of the United States and by date, with the rrewer studies.reported first 

for each region. 

Because of variations in diurnal patterns of PAN by location and season, 

and because no national, uniform aerometric .data base for PAN exists, few of 

the data reported in Table 5-16 are really comparable. Thus, data in this 

table lend themselves to general conclusions but not to the analysis of trends 

5-74 
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over the past decade and a half or-necessarily to analysis of between-city or 

between-region similarities or differences. For example, concentrations of 

PAN in Los Angeles appear to have been a great deal higher in 1965 (Mayrsohn 

and Brooks, 1965) than in 1980 (Hanst et al., 1982) for nearly the same time 

of day, but the 1965 concentrations were measured in September and October and 

the 1980 concentrations were measured in June. Other data from California 

indicate that September and October are more likely to be part of the smog 

season there than June. A comparison by Temple and Taylor (1983) of PAN 

concentrations in Riverside in 1980 with those in 1967 and 1968 indicates 

little difference. Again, however, the sampling periods were not identical 

relative to averaging time or time of year. Tabulated data from Temple and 

Taylor (1983) have been plotted in Figure 5-24. 

In addition to compiling existing data on the concentrations of PAN in 

ambient air, Altshuller (1983) also related PAN concentrations to ozone concentrations where data for both exist. It must be borne in mind for this review, 

as well, that sampling periods (years, time of year, number of ineasurements) 

are not identical and in many cases are not even similar between studies (see 

Table 5-16). Neither are the averaging times over which samples were collected 

and calculated within respective studies identical or even necessarily similar. 

Nevertheless, the data of Altshuller constitute a comprehensive review and 

examination of the relationships among respective photochemical oxidants in 

urban areas of the United States. Concentrations of PAN as a percentage of 

ozone concentrations are given in Table 5-17, but Table 5-16 should be consulted 

for information on sampling periods and averaging times. 

The existence of peroxybenzoyl nitrate (PbzN) in ambient air of urban 

areas was postulated in the 1978 criteria document for ozone and other oxidants 

(U.S. Environmental Protection Agency, 1978), but PBzN has not been clearly 

identified in ambient air in the United States. Hanst et al. (1982) estimated 

that as much as 2 ppb PBzN would be clearly detectable in FTIR measurements 

but reported no clear absorption band for PBzN in their measurements during a 

smoggy period in Los Angeles. They estimated an upper limit of 1 ppb PBzN 

during their 1980 study (the maximum ozone concentration was 272 ppb and the 

maximum PAN concentration was 16 ppb du;ing that period). 

The only homologue of PAN that has been unambiguously identified in 

ambient air in the United States is peroxypropionyl nitrate (PPN). In his 

5-77 
0.07 

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Figure 5-24. Comparison of monthly daylight average and 

maximum PAN concentrations at Riverside, CA, for 1967- 

1968 and 1980. 

Source: Derived from Temple and Taylor (1983). 

-I = 
TABLE 5-17. RELATIONSHIP OF OZONE AND PEROXYACETYL NITRATE AT 

URBAN AND SUBURBAN SITES IN THE UNITED STATES 

PAN/0n , % 

Site/year Avg. At 0 3 peak Reference 

California 

Downtown Los Angeles, 1960 8 7 Renzetti and Bryan (1961) 

Downtown Los Angeles, 1965 NA 7 Mayrsohn and Brooks (1965) 

Downtown Los Angeles, 1968 13 NAa Lonneman et al. (1976) 

West Los Angeles, 1980 9 6 Hanst et al. (1982) 

Pasadena, 1973 10 8 Hanst et al. (1975) 

West Covina, 1977 20 12 Spicer (1977) 

Claremont, 1978 7 6 Tuazon et al. (1981a, 1981b) 

Claremont, 1979 4 4 Tuazon et al. (1981a) 

Riverside, 1967-1968 8 NA Taylor (1969) 

Riverside, 1975-1976 9 5 Pitts and Grosjean (1979) 

Riverside, 1976 5 4 Tuazon et al. (1978) 

Riverside, 1977 4 4 Tuazon et al. (1980) 

Riverside, 1977 4 NA Singh et al. (1979) 

Southwest 

Houston, TX, 1976 3 3 Westberg et al. (1978) 

Midwest 

St. Louis, M0, 1973 13 NA Lonneman et al. (1976) 

St. Louis, M0, 1973 5 5 Spicer (1977) 

Dayton, OH, 1974 2 1 Spicer et al. (1976) 

(Huber Hts, OH) 

East 

Hoboken, NJ, 1970 4 NA Lonneman et al. (1976) 

New Brunswick, NJ, 1978-1980 4 2 Brennan (1980) 

allot available. 

Source: Adapted from Altshuller (1983). 

5-79 
review of existing literature, Altshuller (1983) compiled data on the concentrations of PPN in ambient air in urban areas. In addition, fie calculated 

ratios of the concentrations of PPN and PAN., His data, expressed as 

percentages [(PPN/PAN) x 100], are presented as Table 5-18 (Altshuller, 1983). 

As Altshuller has pointed out, average PPN concentrations are 10 to 30 

percent of the average PAN concentrations shown in Table 5-18 except for those 

reported for San Jose (8 percent) and Oakland (42 percent). The maximum PPN 

concentrations reported are highly variable, however, ranging from 0.13 ppb in 

San Jose (August 1978) to 6.0 ppb in Riverside (month and year as well as 

sampling period of day unknown). Thus, the PPN/PAN ratio at maximum concentrations of PPN is highly variable, as well. Among more recent data, the maximum 

PPN concentration was 5.0 ppb in St. Louis in August 1973. Note, however, 

that the sampl i ng peri od i n St. Loui s was 10: 00 a. m. to 3: 30 p. m. Dependi ng 

upon temperature, concentrations of precursors, and other factors, a true PPN 

maximum may or may not have occurred by 3:30 p.m. 

Table 5-19 presents PAN and PPN concentrations reported by Singh et al. 

(1981) for three cities, Los Angeles, Oakland, and Phoenix; as well as PPN 

concentrations as percentages of PAN concentrations (PPN/PAN x 100). Comparison 

of the data from these three cities helps demonstrate the variability of PPN 

and PAN concentrations with location. 

5.6.4 Ambient Air Concentrations of PAN and Its Homologues in Nonurban Areas 

Data on the concentrations of PAN and PPN in agricultural and other nonurban areas of the United States are sparse. They include data from the study 

done by Lonneman et al. (1976) in Wilmington, Ohio, in August 1974, and cited 

earlier in Section 5.6.2. In that study, measurements made by GC-ECD from 

10:00 a.m. to 4:00 p.m., local,time, showed a maximum concentration for the 

study period of 4.1 ppb. The average daily maximum was 2.0 ppb. While the 

4:00 p.m. cutoff used by Lonneman et al. (1976) could possibly have resulted 

in missing some peak PAN concentrations, especially in transported air masses, 

the data are within the range reported by Westberg et al. (1978) and by Spicer 

and Sverdrup (1981) for 24-hour measurements at nonurban sites. At the Sheldon 

Wildlife Reserve, Texas, Westberg et al. (1978) found a 24-hour average PAN 

concentration of 0.64 ppb and a maximum concentration of.2.8 ppb for the study 

period (October). Spicer and Sverdrup (1981) found a 24-hour average PAN 

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concentration of 0.50 ppb and a maximum of 6.5 ppb for the study period (July 

and August) at Van Hiseville, New Jersey, in the New Jersey pine barrens; and 

Spicer et al. (1983) reported averages of 0.46 ppb at a nonurban site in 

Indiana (Huntington Lake) and 0.74 at "a nonurban site in east central Missouri 

(42 km west of St. Louis). 

In contrast with these values, 24-hour average concentrations of PAN at 

one nonurban and four remote sites (see Table 5-20) were reported by Singh et al. 

(1979) to range from 0.08 to 0.30 ppb, with maxima at those sites ranging from 

0.22 to 0.83 ppb. The Texas, Ohio, and New Jersey sites sustained higher PAN 

concentrations than the nonurban sites where Singh et al. (1979) measured PAN. 

The higher concentrations may reflect the influence at those sites of nearby 

metropolitan areas. 

Concentrations of PAN have recently been reported by Singh and Salas 

(1983) for a Pacific marine site, Point Arena, California, at which earlier 

measurements were also made and reported by Singh et al. (1979) (see Table 

5-19). Data collected in August 1982 showed concentrations of PAN ranging 

from 0.01 to 0.12 ppb during the 5-day study period. The average. concentration 

for the period was 0.032 ± 0.024 ppb. Winds were west-to-northwest 90 percent 

of the time and northerly the rest of the time. Modeled trajectories confirmed 

that air masses passing over the site were of a marine origin. The site is 

thought to be free of manmade pollutants. 

Data from two nonurban sites in Canada are of interest even though they 

are outside the United States. Cherniak and Corkum (1981; cited in Temple and 

Taylor, 1983) measured PAN at a nonurban site in Simcoe, Ontario, Canada, for 

6 months. Measurements made by GC-ECD showed monthly means of <2 ppb and a 

maximum concentration during the study of 5.6 ppb. At a remote site in the 

Kananaskis Valley of Alberta, Canada, monthly mean concentrations were <1 ppb 

for samples taken at half-hour intervals, 24 hr/day, for 110 days (Peake et al., 

1983). The site, located at the base of a mountain range and about 50 miles west 

of Calgary, is thought to be free of manmade pollutants, including transported 

pollutants. 

5.6.5 Temporal Variations in Ambient Air Concentrations of Peroxyacetyl Nitrate 

5.6.5.1 Diurnal Patterns. Concentration data obtained in the 1960s were 

briefly discussed in Section 5.6.2, where it was noted that the first criteria 

5-83 
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5-84 

~ 
document for photochemical oxidants (U.S. Department of Health, Education, and 

Welfare, 1970) reported concentrations and patterns for PAN that remain valid 

now. In that document, the general proximity in time of PAN and oxidant peaks 

was shown in data from Los Angeles and Riverside, California. Maximum PAN 

concentrations, although varying from location to location, generally occurred 

in midday; i.e., late morning to mid-afternoon. Figures 5-25 and 5-26, taken 

from the 1970 criteria document (U.S. Department of Health, Education, and 

Welfare, 1970), graphically present the diurnal patterns of PAN in Los Angeles 

in 1965 and in Riverside in 1966. The occurrence of the second PAN peak in 

Riverside, which appears to trail a second total oxidant peak by an hour or 

two, was ascribed to transport, as verified by the occurrence of maximum 

oxidant concentrations at three receptor sites east of West Los Angeles (downtown Los Angeles, Azusa, and Riverside), at times that corresponded, wind 

speed factored in, with respective distances from West Los Angeles. 

Examples drawn from recent data substantiate that the general diurnal 

pattern (as it appears in composite diurnal data averaged over a week, a 

month, or longer) remains the same as the pattern established by data obtained 

in the mid-1960s. 

Using FTIR spectroscopy, Tuazon and coworkers (Tuazon et al., 1978, 1980, 

1981b) measured concentrations of PAN at Claremont and Riverside, California, 

over a 5-year period. Concentrations of PAN ranged from about 5 to 40 ppb 

over the course of the study. The diurnal profiles for PAN and ozone at 

Claremont are shown for 2 days of a multiday smog episode in October 1978 in 

Figure 5-27 (Tuazon et al., 1981b). Note the qualitative relationship of the 

two pollutants, with the peak concentrations of the two occurring almost 

simultaneously. The relationship between PAN and ozone concentrations and 

behavior in the atmosphere is neither constant nor monotonic, however, as is 

borne out by the siight differences in time of occurrence of their peak concentrations but especially by the persistence of somewhat elevated PAN concentrations before return to "baseline" levels. It appears that PAN concentrations, 

in this instance at least, closely parallei the nitric acid (HNO 3 ) concentrations, persisting after ozone concentrations have subsided. The percentage of 

PAN relative to ozone differed slightly at the peak concentrations of the two 

on the 2 days. On October 12, the peak PAN concentration was close to 6 percent 

of the peak ozone concentration; on October 13, the peak PAN concentration was 

nearly 8 percent of the peak ozone concentration. 

5-85 
cvs 0.16 

z 

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Fz 

W 

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a 

a 

AVERAGES: 

19 WEEKDAYS, 

— • — ~ OCTOBER 

} 16 WEEKDAYS, 

SEPTEMBER 

OXIDANT 

i 

/ 

/ 

P 

/ 

/ 

/ 

i 

/ 

/ 

/ 

~ 

~ OXIDANT 

\ 

O♦ 

♦ 

♦ 

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8 9 10 11 12 1 2 3 

a.m. 1

4-----;p.m. ---~ 

HOUR OF DAY, PST . 

Figure 5-25. Variation of inean 1-66ur oxidant 

and PAN concentrations, by hour of day, in 

downtown Los Angeles, 1965. 

Source: U.S. Department of Health, Education, 

and Welfare (1970). 

:. 
0.18 

E 0.16 

a 

a 

z

0 

 0.14 

i= 

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E 0.010 Q 

Q i 1 1 1 I 1 1 1 0 

6 8 10 12 2 4 6 8 

~-- a.m.— ~--- p.m. ---~I 

I-IO U R 0 F DAY, PST 

Figure 5-26. Variation of inean 1-hour 

oxidant and PAN concentrations, by hour of 

day, Air Pollution Research Center, 

Riverside, CA, September, 1966. 

Source: U.S. Department of Health, 

Education and Welfare (1970). 

5-87 
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5-88 
5.6.5.2 Seasonal Patterns. Seasonal differences in PAN concentrations were 

alluded to in Section 5.6.3 and mean and maximum PAN concentrations were presented by month for 2 years (1967-1968 and 1980) for Riverside, California, in 

Figure 5-24. That seasonal differences exist in PAN concentrations was also 

documented in the 1970 criteria document for photochemical oxidants (U.S. 

Department of Health, Education, and Welfare, 1970). Total oxidants and PAN 

were monitored for 13 months in Riverside, California. The oxidant concentrations were obtained by continuous Mast meter measurements, 24 hr/day. Concentrations of PAN were measured in sequential samples analyzed by GC-ECD from 

6:00 a.m. to about 4:00 or 5:00 p.m. The data are not strictly comparable, 

since the shorter, daylight averaging time for PAN would be expected to result 

in somewhat higher mean concentrations of PAN than would be obtained across a 

24-hour averaging period. Nevertheless, the patterns given in Figure 5-28 are 

of interest and demonstrate that peak PAN concentrations can constitute a 

higher percentage of the peak ozone concentrations during winter months than 

during the rest of the year. This observation is still valid (Spicer et al., 

1983) and has been attributed by Lewis et al. (1983) and by Holdren et al. 

(1984) to greater PAN stabiiity in the winter months because of cooler temperatures (Cox and Roffey, 1977). The possibility, however, that the somewhat 

greater NOx emissions of the winter heating season also contribute to this 

phenomenon should not be overlooked. 

Data from one additional study complement the older data already presented 

on diurnal and seasonal patterns (Section 5.4.2). Lewis et al. (1983) measured 

PAN and ozone concentrations from September 25, 1978, to May 10, 1980, in New 

Brunswick, New Jersey. Average (10-hr and 24-hr) and maximum concentrations 

of both pollutants are given in Table 5-21 by month of the year (Lewis et al., 

1983). Note that the highest monthly mean concentrations, both the 10-hr and 

24-hr means, occurred during the smog season (August and September) but that 

the next highest occurred in October and February, respectively. Average 

diurnal profiles were obtained during this same study and are shown, by month, 

in Figure 5-29 (Lewis et al., 1983). 

5.6.6 Spatial Variations in Ambient Air Concentrations of Peroxyacetyl Nitrate 

5.6.6.1 Urban-Rural Gradients and Transport of PAN. As noted earlier, precursors to PAN, especially NO 2 , are lower in nonurban than in urban areas, such 

that little local formation is expected in nonurban areas. Available data on 

5-89 
E 

a 0.22 

Z 0.20 

O 

~ 0.18 

~ 0.16 

H 

W 0.14 

Z 0.12 

V 0.10 

Z 0.08 

p 0.06 

~ 0.04 

z 0.02 

ua 0.00 

MONTHLY MEANS OF DAILY MAXIMUM 

1-hour AVERAGE CONCENTRATIONS 

OXIDANT MONTHLY MEANS OF 1-hour AVERAGE 

CONCENTRATIONS 

E 

0.010 a 

0.008 z z QO 

0006 Z ~ 

0.004 W ~ 

0.002 w 

0.000 z 

~ JUN. JUL. AUG. SEP. OCT. NOV. DEC. JAN. FEB. MAR. APR. MAY JUN. V 

MONTH 

1966 1967 ---~--yI 

Figure 5-28. Monthly variation of oxidant (Mast meter, continuous 

24-hr) concentrations and PAN (GC-ECD, sequential, 6:00 a.m. to 

4:00-5:00 p.m.) concentrations, Air Pollution Research Center, 

Riverside, CA, June 1966 - June 1967. 

Source: U.S. Department of Health, Education, and Welfare (1970). 

5-90 
TABLE 5-21. PAN AND OZONE CONCENTRATIONS IN AMBIENT AIR, 

NEw BRUNSWICK, N.J., FOR SEPTEMBER 25, 1978, TO MAY 10, 1980 

PAN concentration, ppb 03concentration, ppb 

24-hr 10-hr Hourly 24-hr 10-hr Hourly 

Month average average maximum average average maximum 

January 0.12 0.19 1.3 11.5 15.5 34 

February 0.61 0.69 4 17.2 23.2 40 

March 0.36 0.41 1.3 23 29.1 58 

Apri1 `''0.45 0.57 2.5 284.5 37.3 80, 

May 0.23 0.28 1 31.4 40.9 78 

Junea 0.09 0.17 0.8 NA NA NA 

July 0.26 ` 0.44 3.5 37.5 57.6 130 

August 1.17' 1.63 10.6 37.4 55.9 145 

September .1.04 1.41 7.5 22.4 33.9 110 

Octob,er 0.93 1.08 5.8 15.8 22.6 68.' 

~ November 0.25. 0.31 3.5 11.6 15.8 40 

=~ December 0.57 0.62 2.5 9.7 12.8 35 k+. . . . . , ... . . . 

aThese res''ults are lower than expected`; however, there was no evidence of 

instrument malfunction. 

Source: Lewis et al. (1983). 

PAN concentrations indicate clearly that they are lower in nonurban areas than 

in urban (Section 5.5.3). It should be noted, however, that few data exist on 

concentrations in agricultural areas and that the possib.ility that PAN is 

transported is therefore important in,.assessing exposures.of vegetation to 

PAN. Lonneman et al. (1976) and Nieboer and Van Ham (1976), in studies cited 

in the 1978 criteria document (U.S. Environmental Protection Agency, 1978), 

reported the transport of PAN. The more recent study by Nielsen et al. (1981) 

has confirmed that PAN can be present at relatively high concentrations in 

photochemically polluted air after long-range transport. Variations in concentrations of PAN and other oxidants measured in Claremont, California 

(Grosjean, 1983), are consistent with transport patterns. The recent work of 

Singh and Salas (1983) has shown that significant nighttime PAN concentrations 

can occur aloft, at least in a relatively clean environment. It is possible 

that the transport of PAN occurs aloft, as with ozone, and that under favorable 

conditions PAN can be transported long distances. 

5-91 
0.080 

0.060 

E 0.040 

a 

a 

Z 0.020 

O 

Q 0 

~ fW 0.025 

c~ 

z 

O 

V0.015 

OZONE 8 

7 

9 

10 

8 

PAN 

9 

10 

7 

~fi~➢ 

0 12 24 

~s --a.m. ►}A p.m.----~

I 

TIME OF DAY, hour 

Figure 5-29. Average daily profile by month 

(July 7- October 10) for PAN and ozone in 

New Brunswick, NJ, 1979. Numbers refer to 

months of the year. 

Source: Lewis et al. (1983). 

5-g2 
5.6.6.2Intracity Variations . Few data on PAN concentrations at different 

sites in the same city are available. One study is available for Houston, 

Texas (Jorgen et al., 1978), in which PAN was measured on October 26 and 27, 

1977, at three separate sites, two in Houston and one north of the city. 

Comparison of the levels of ozone and PAN peaks among the three sites on those 

2 days reveals significant differences (Table 5-22). On October 26, the 

highest ozone concentrations (two peaks of 110 and 95 ppb at —3:00 p.m. and 

—7:00 p.m., respectively) were seen at site 3, in southeast Houston; while the 

highest PAN concentration on that day was a secondary peak of 15 ppb at —8: 100 

p.m. at site 2(mid-city). On October 27, the highest ozone concentration, 

180 ppb, occurred at site 1, about 4 miles north-northeast of Houston, at 

—1:00 p.m. That was accompanied at the same time by the highest PAN concentration for the day, 16 ppb. 

Examination of Table 5-22 shows the lack of a consistent, quantitative 

relationship between PAN and ozone as indicated by intracity differences of 

more than twofold in ozone concentrations (afternoon peaks, sites 1 and 2, 

October 27) and differences of about fivefold in PAN concentrations between 

sites (afternoon peaks, sites 1 and 2, October 27). 

5.6.6.3 Indoor-Outdoor Ratios of PAN Concentrations . No recent studies 

appear in the literature on indoor concentrations of PAN or indoor-outdoor 

ratios (I/0). In three school buildings in southern California, Thompson et 

al. (1973) found I/0 ratios (expressed here as percentages) of 89, 97, and 148 

percent, respectively, in the absence of air conditioning. With air conditioning, the I/0 ratios were 75, 108, and 117 percent, respectively. Total 

oxidants were.neariy constant all day, remaining about 30 percent (in air 

conditioning) of the average outside concentration. The higher I/0 gradients 

for PAN than for oxidants were attributed by the authors to the greater breakdown of ozone ("oxidants") through its reaction with surfaces. The cooler 

temperatures indoors are the probable cause of the greater persistence of PAN 

indoors. Like ozone, PAN also decays indoors, but over an extended period 

(Thompson et al., 1973). 

5.7 CONCENTRATIONS OF OTHER PHOTOCHEMICAL OXIDANTS IN AMBIENT AIR 

Concentrations of nitrogen dioxide (NO 2 ) and related nitrogenous oxidants 

are presented in a recent criteria document on oxides of nitrogen (U.S. Environmental Protection Agency, 1982b) and are not given here. In addition, the 

5-93 
TABLE 5-22. INTRACITY VARIATIONS IN PEAK OZONE AND PAN CONCENTRATIONS IN 

HOUSTON, OCTOBER 26 AND 27, 1977 

03 , PAN, 

Site Date ppb Time of peak ppb Time of peak 

1(N4 mi NNE 10/26 75 1:00 P.M. 2 9:00 a.m. - 5:00 

of Houston) p.m. (plateau) 

10/27 180 2: 00 p. m. 16 2: 00 p. m. 

Z(mid-city) 10/26 50 2:00-3:00 p.m. 2 2:00-6:00 p.m. 

(plateau) 

10/26 50 8: 00 p. m. 10 7: 00 p. m. 

10/27 70 1:00-3:00 p.m. 3 12:30-3:00 p.m. 

3(SE Houston) 10/26 110 3:00 p.m. 8 2:00 p.m. 

10/26 95 7: 00 p. m. 12 5: 30 p. m. 

10/27 90 1:00 p.m. 4 11:30 a.m. 

10/27 55 7:30 p.m. 4 6:30 p.m. 

Source: Jorgen et al. (1978). 

comprehensive review by Altshuller (1983) also documents available information 

on nitrogenous oxidants such as nitric acid (HNO3) and peroxynitric acid, as 

well as on formic acid (HCOOH) and hydrogen peroxide (H 2O 2 ). The reader is 

referred to these reviews for information on these oxidants. The two nonnitrogenous oxidants, formic acid and hydrogen peroxide, are appropriate 

concerns for this document, however, and the limited information on concentrations of these two pollutants is summarized below. 

Studies on the toxicologic effects of formic acid, though limited in 

number, have shown only negligible effects, even in animals exposed to levels 

as high as 20 ppm (aerosol vapor) (see Chapter 9). These levels are three 

orders of magnitude greater than the concentrations seen in polluted urban 

atmospheres. For example, maximum concentrations of HCOOH observed by Tuazon 

and coworkers, using FTIR (Tuazon et al., 1978; 1980; 1981b), in Claremont and 

Riverside, California, were in the range of 5 to 20 ppb in a study covering 

5 years. The ranges of concentrations of HCOOH measured by Tuazon et al. were 

consistent with those found in a long-path FTIR study by Hanst et al. (1982). 

The FTIR method offers a reliable assessment of the ambient air concentrations 

of HCOOH and reported concentrations are.believed to be accurate. 

5-94 
Data on HCOOH concentrati ons for 2 days i n October 1978 are shown i n 

Figure 5-30 (Tuazon et ai.,' 1981b). The diurnal pattern is similar to that of 

related oxidants and of some of the other smog products. ° 

The measurement of hydrogen peroxide (H2O2 ) in ambient air is fraught 

with difficulties that remain unresolved. Ozone itself is known to be an 

interference in virtually all of the past and current measurement methods for 

H2O2 (Chapter 4). In his review of non-ozone oxidants and other smog constituents and photochemical products, Altshuller (1983) examined data obtained in 

the South Coast Ai r Basi n of Caa i forni a i n the 1 ate 1970s and i n 1980 for 

possible consistency in the interference of ozone in H 2 O2 measurements. 

Laboratory experiments (Heikes et al., 1982) have indicated that 1 ppb H 2O 2 

would be generated per 100 ppb ozone. The analysis by Altshuller shows that 

this relationship does not hold in ambient air in the South Coast Air Basin 

once H2O 2 levels exceed about 5 to'10 ppb; and Altshuller (1983) concluded 

that variations in H 2O 2 measurements there remain unexplained. 

Because of ineasurement probiems, the true levels of H 2O2 in ambient air 

are unknown, especially in polluted areas, where multiple interferences may 

possibly occur. Attempts to detect hydrogen peroxide by means of FTIR spectroscopy have all been negative, even in polluted areas. The method can measure 

H 2O 2 in ambient air with specificity at H 2O 2 levels > 40 ppb, which is the 

limit of detection for an FTIR instrument with a 1-km pathlength (see Chapter 

4). 

Notwithstanding measurement difficuities, some ranges of H 2O 2 concentrations at urban and nonurban sites have been reported in the literature. These 

are given in Tabie 5-23, along with the general type of ineasurement method 

used to obtain the reported concentrations. It must be kept in mind, however, 

that the reported concentrations, though they represent state-of-the-art 

measurements, are thought not to be accurate. 

5.8 SUMMARY 

In the context of this document, the concentrations of ozone and other 

photochemical oxidants found in ambient air are important for: (1) assessing 

potential exposures of human and other receptors; (2) determining the range of 

ambient air concentrations of ozone and other photocheinical oxidants relative 

5-95 
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5-96 
TABLE 5-23. CONCENTRATIONS aOF HYDROGEN PEROXIDE IN AMBIENT AIR 

AT URBAN AND NONURBAN SITES 

(ppb) 

Concns., and 

Location Date comments Method Reference 

Hoboken, NJ 1970 < 40 Titanium (IV) Bufalini et 

(urban) sulfate/8-quino- al. (1972) 

linol 

Riverside, CA 1970 <180 (during Titanium (IV) Bufalini et 

(urban) smog episode sulfate/8-quino- al. (1972) 

with 650 ppb linol 

oxidants) 

Riverside and 

Claremont, CA 

July-Aug. 

1977 

100 max. (ozone 

also 100 and 

Luminol 

chemiluminescence 

Kok et al. 

(1978) 

(urban) increasing); 

10 to 50 on 

most days 

Minneapolis, MN NAb < 6 Wet chemical c Heikes et al. 

(urban) (1982) 

Boulder, CO NA < 0.5 Wet chemical c Heikes et al. 

(urban) (1982) 

Boulder, CO Feb. 1978 0.2 to 3 Luminol Kelly et al. 

(nonurban, chemiluminescence (1979) 

east of 

Boulder) 

Tucson; AZ NA < 7 Luminol Farmer and 

(nonurban, chemiluminesence Dawson (1982) 

54 km SE 

of Tucson) 

Tucson, AZ NA —1 Luminol Farmer and 

(remote, near chemiluminescence Dawson (1982) 

Tucson) 

allote: Since methods used to obtain these data are all subject to positive interference by ozone, data presented here are not reliable. They are included as a 

summary of reported concentrations. 

bNot available. 

cSee Chapter 4 for method used by Heikes et al. (1982). 

Source: Derived from data in Altshuller (1983). 

5-97 
to demonstrated "effects levels" (Chapters 6-12); (3) determining indooroutdoor gradients for exposure analyses; (4) assessing whether the concentrations of oxidants other than ozone, singly, collectively, or in comb'ination 

with ozone, are cause for concern; and (5) evaluating the adequacy of ozone as 

a control surrogate for other photochemical oxidants, if concentrations of the 

other oxidants are cause for concern given the effects and the "effects levels" 

of those oxidants. 

5.8.1 Ozone Concentrations in Urban Areas 

In Table 5-24, 1983 ozone concentrations for Standard Metropolitan Statistical Areas (SMSAs) having populations > 1 million are given by geographic 

area, demarcated according to United States Census divisions and regions (U.S. 

Department of Commerce, 1982). The second-highest concentrations among daily 

maximum 1-hour values measured in 1983 in the 38 SMSAs having populations of 

at least 1 million ranged fr.om0.10 ppm in the Ft. Lauderdale, Florida; Philadelphia, Pennsylvania; and Seattle, Washington, areas to 0.37 ppm in the Los 

Angeles-Long Beach, California, area. The second=highest value among daily 

maximum 1-hour ozone concentrations for 35 of the 38 SMSAs in Table 5-24 

equaled or exceeded 0.12 ppm. The data clearly show, as well, that the 

highest 1-hour ozone concentrations in the United States occurred in the 

northeast (New England and Middle Atlantic states), in the Gulf Coast area 

(West South Central states), and on the west coast (Pacific states). Secondhighest daily maximum 1-hour concentrations in 1983 in the SMSAs,within each 

of these three areas averaged 0.16, 0:17, and 0.21 ppm, respectively. It 

should be emphasized that these three areas of the United States are subject 

to those meteorological and climatological factors that are conducive to local 

oxidant formation, or transport, or both. It should also be emphasized that 9 

of the 16 SMSAs in the country with populations > 2 million are located in 

these areas. 

Emissions of manmade oxidant precursors are usually correlated with 

population, especially emissions from area source categories such as transportation and residential heating (Chapter 3): Accordingly, when grouped by 

population, the 80 largest SMSAs had the following median values for their 

collective second-highest daily maximum 1-hour ozone concentrations in 1983: 

populations > 2 million, 0.17 ppm 03; populations of 1 to 2 million, 0.14 ppm 

03 ; and populations of 0.5 to 1 million, 0.13 ppm 0 3 . As noted above, however, 

.: 
TABLE 5-24. SECOND-HIGHEST OZONE CONCENTRATIONS AMONG DAILY MAXIMUM 1-hr 

VALUES IN 1983 IN STANDARD METROPOLITAN STATISTICAL AREAS•WITH POPULATIONS 

> 1 MILLION, GIVEN BY CENSUS DIVISIONS AND REGIONS a 

Division 

and region SMSA 

SMSA 

population, 

millions 

Second-highest 

1 1983 03 

concn., ppm 

Northeast 

New England Boston, MA _>2 0.18 

Middle Atlantic Buffalo, NY 1 to <2 0.12 

Nassau-Suffolk, NY >2 0.17 

Newark, NJ 1 to <2 0.25 

New York, NY/NJ >2 0.19 

Philadelphia, PA/NJ >2 0.10 

Pittsburgh, PA >2 0.14 

South 

South Atlantic Atlanta, GA >2 0.17 

Baitimore, MD >2 0.19 

Ft. Lauderdale-Ho1lywood, FL 1 to <2 0.10 

Miami, FL 1 to <2 0.12 

Tampa-St. - Petersburg, FL 1 to <2. 0.14 

Washington, DC/MD/VA >2 0.17 

South 

West South 

Central Dallas-Ft. Worth, TX >2 0.16 

Houston, TX >2 0.28 

New Orleans,`LA 1 to <2 0.12 

San Antonio,.TX 1 to <2 0.12 

North Central 

East North 

Central Chicago, IL >2 0.17 

Detroit, MI >2 0.17 

Cleveland, OH 1 to <2 0.15 

Cincinnati, OH/KY/IN 1 to <2 0.15 

Milwaukee, WI 1 to <2 0.18 

Indianapolis, IN l to <2 0.14 

Columbus, OH` 1 to <2 0.12 

West North 

Central St. Louis, MO/IL >2 0.18 

Minneapolis-St. Pau1,-MN/WI >2 0.13 

Kansas City, MO/KS 1 to <2 0.13 

5-99 
TABLE 5-24 (cont'd). SECOND-HIGHEST OZONE CONCENTRATIONS AMONG DAILY MAXIMUM 

1-hr VALUES IN 1983 IN STANDARD METROPOLITAN STATISTICAL AREAS 

WITH > 1 MILLION, GIVEN BY CENSUS DIVISIONS AND REGIONS a 

Division 

and region SMSA 

SMSA 

population, 

millions 

Second-highest 

1983 03 

concn., ppm 

West 

Mountain Denver-Boulder, CO 1 to <2 0.14 

Phoenix, AZ - 1 to <2 0.16 

Pacific Los Angeles-Long Beach, CA >2 0.37 

San Francisco-Oakland, CA >2 0.17 

Anaheim-Santa AnaGarden Grove, CA 1 to <2 0.28 

San Diego, CA 1 to <2 0.20 

Seattle-Everett, WA 1 to <2 0.10 

Riverside-San Bernardino- 1 to <2 0.34 

Ontario, CA 

San Jose, CA 1 to <2 0.16 

Portland, OR/WA 1 to <2 0.12 

Sacramento, CA 1 to <2 0.15 

aStandard Metropolitan Statistical Areas and geographic divisions and regions 

as defined by Statistical Abstract of the United States (U.S. Department of 

Commerce, 1982). 

Source; U.S. Environmental Protection Agency (1984a). 

coincident meteorology favorable for oxidant formation undoubtedly contributes 

to the apparent correlation between population and ozone levels. 

Among all stations reporting valid ozone data (> 75 percent of possible 

hourly values per year) in 1979, 1980, and 1981 (collectively, 906 stationyears) in the United States, the median second-highest 1-hour ozone value was 

0.12 ppm, and 5 percent of the stations reported second-highest 1-hour values 

> 0.28 ppm. 

A pattern of concern in assessing responses to ozone in human populations 

and in vegetation is the occurrence of repeated or prolonged multiday periods 

when the ozone concentrations in ambient air are in the range of those known 

to elicit responses (see Chapters 10 and 12). In addition, the number of days 

of respite between such multiple-day periods of high ozone is of possible 

consequence. Data show that repeated, consecutive-day exposures to or respites 

5-100 
from specified concentrations are location-specific. At a site in Dallas, 

Texas, for example, daily maximum 1-hour concentrations were > 0.06 ppm for 

2 to 7 days in a row 37 times in a 3-year period (1979 through 1981). A concentration of >0.18 ppm was recorded at that site on only 2 single days, 

however, and no multiple-day recurrences of that concentration or greater were 

recorded over the 3-year period. At a site in Pasadena, California, daily 

maximum 1-hour concentrations >0.18 ppm recurred on 2 to 7 consecutive days 

33 times in that same 3-year period (1979 through 1981) and occurred, as well, 

on 21 separate days. These and other data demonstrate the occurrence in some 

urban areas of multiple-day potential exposures to relatively high concentrations of ozone. 

5.8.2.Trends in Nationwide Ozone Concentrations 

Trends in ozone concentrations nationwide are important for estimating 

potential exposures in the future of human populations and other receptors, as 

well as for examining the effectiveness of abatement programs. The determination of nationwide trends requires the application of statistical tests to 

comparable, representative, multiyear aerometric data. The derivation of 

recent trends in ozone concentrations and the interpretation of those trends 

is complicated by two potentially significant factors that have affected 

aerometric data since 1979: (1) the promulgation by EPA in 1979 of a new 

calibration procedure for ozone monitoring (see Chapter 4); and (2) the introduction by EPA of a quality-assurance program that has resulted.in improved 

data-quality audits. The effects of these factors on ozone concentration 

measurements are superimposed on the effects on concentrations of any changes 

in meteorology or in precursor emission rates that may have occurred over the 

same time span. 

The nationwide trends in ozone concentrations for a 9-year period, 1975 

through 1983, are shown in Figure 5-31 (U. S. Environmental Protection Agency, 

1984a). The data given are trends as gauged by the composite average of the 

second-highest value among daily maximum 1-hour ozone concentrations. Data 

from four subsets of monitoring stations, most of them urban stations, are 

given: (1) California stations only; (2) all stations except those in 

California; (3) all stations including those in California; and (4) all 

National Air Monitoring Stations (NAMS), which report data directly to EPA. 

Only stations reporting > 75 percent of possible hourly values in the respective 

years are represented in the data. 

5-101 
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A CALIFORNIA STATIONS (27) 

0 ALL STATIONS EXCEPT 

CALIFORNIA (149) 

T , l ~----~--- 

----~--- 

1975 1976 1977 1978 1979 1980 1981 1982 1983 

YEAR 

Figure 5-31. National trend in composite average of the second highest value 

among daily maximum 1-hour ozone concentrations at selected groups of 

sites, 1975 through 1983. 

Source: U.S. Environmental Protection Agency (1984a). 

5-102 
For the entire 9-year period, 1975 through 1983, all subsets of monitoring 

stations show a decline in the composite second-highest daily maximum 1-hour 

ozone concentration. Between 1979, when the new, more accurate calibration 

procedure was promulgated, and 1982, a small decline of 9 to 10 percent in 

nationwide ozone concentrations occurred. From.1982 to 1983, however, concentrations increased by about 10 percent in California, by about 14 percent 

outside California, and by about 12 percent nationwide (all states). Recently 

published data for 1984 from a somewhat smaller number of sites (163) (U.S. 

Environmental Protection Agency, 1986) show a decrease in nationwide ozone 

concentrations from 1983 to 1984, with 1984 levels approximating those recorded 

in 1981. The portion of the apparent decline in ozone nationwide from 1975 

through 1984 that is attributable to the calibration change of 1979 cannot be 

determined by simply applying a correction factor to all data, since not all 

monitoring stations began using the UV calibration procedure in the same year. 

Figure,5-32 shows the nationwide frequency distributions of the first-, 

second-, and third-highest 1-hour 0 3 concentrations at predominantly urban 

stations aggregated for 1979, 1980, and 1981, as well as the highest 1-hour 0 3 

concentration at site of the National Air Pollution Background Network (NAPBN) 

aggregated for the same 3 years. As shown by Figure 5-32, 50 percent of the 

second-highest 1-hour values from non-NAPBN sites in this 3-year period were 

0.12 ppm or iess and about 10 percent were equal to or greater than 0.20 ppm. 

At the NAPBN sites, the collective 3-year distribution (1979 through 1981) is 

such that about 6 percent of the vaiues are less than 0.10 ppm and fewer than 

20 percent are higher than 0.12 ppm. 

5.8.3. Ozone Concentrations in Nonurban Areas 

Few nonurban areas have been routinely monitored for ozone concentrations. 

Consequently, the aerometric data base for nonurban areas is considerably less 

substantial than for urban areas. Data are available, however, from two 

special-purpose networks, the National Air Pollution Background Network (NAPBN) 

and the Sulfate Regional Experiment network (SURE). Data on maximum 1-hour 

concentrations and arithmetic mean 1-hour concentrations reveal that maximum 

1-hour concentrations at nonurban sites classified as rural (SURE study, 

Martinez and Singh, 1979; NAPBN studies, Evans et al. , 1983) can sometimes 

exceed the concentrations observed at sites classified as suburban (SURE 

study, Martinez and Singh, 1979). For example, maximum 1-hour ozone concentrations measured in 1980 at Kisatchie National Forest (NF), Louisiana; Custer 

5-103 
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5-104 
NF, Montana; and Green Mt. NF, 1/ermont, were 0.105, 0.070, and 0.115 ppm, 

respectively. Arithmetic mean' 1-hour concentrations for 1980 were 0.028, 

0.037, and 0.032 ppm at the respective sites. For four nonurban (rural) siteS 

in the SURE study, maximum 1-hour ozone concentrations were 0.106, 0.107, 

0.117, and 0.153; and mean 1-hour concentrations ranged from 0.021 to 0.035 

ppm. At the five nonurban (suburban) sites of the SURE study, maximum concentrations were 0.077, 0.099, 0.099, 0.080, and 0.118 ppm, respectively. The 

mean 1-hour concentrations at those sites were 0.023, 0.030, 0.025, 0.020, and 

0.025 ppm, respectively. 

Ranges of concentrations and the maximum 1-hour concentrations at some of 

the NAPBN and SURE sites show the probable influence of ozone transported from 

urban areas. In one documented case, for example, a 1-hour peak ozone concentration of 0.125 ppm at an NAPBN site in Mark Twain National Forest, Missouri, 

was measured during passage of an air mass whose trajectory was calculated to 

have included Detroit, Cincinnati, and Louisville in the preceding hours 

(Evans et al., 1983). 

The second-highest concentration among all the daily maximum 1-hour 

concentrations measured at the NAPBN sites appear to be about one-half the 

corresponding concentrations measured at urban sites in the same years.. No 

trend in concentrations at these NAPBN sites is discernible in the data re.cord 

for 1979 through 1983. 

These data corroborate the conclusion given in the 1978 criteria docqment 

(U.S. Environmental Protection Agency, 1978) regarding urban-nonurban and 

urban-suburban gradients; i.e., nonurban areas may sometimes sustain higher 

peak ozone concentrations than those found in urban areas. 

5.8.4. Diurnal and Seasonal Patterns in Ozone Concentrations 

Since the photochemical reactions of precursors that result in ozone formation are driven by sunlight, as well as by emissions, the patterns of ozone 

occurrence in ambient air depend on daily and seasonal variations in sunlight 

intensity. The typical diurnal pattern of ozone in ambient air has a minimum 

ozone level around sunrise (near zero in most urban areas), increasing through 

the morning to a peak concentration in early afternoon, and decreasing.toward 

minimal levels again in the evening. The 1978 criteria document ascribed,the 

daily ozone pattern to three simuitaneous processes: (1) downward transport of 

ozone from layers aloft; (2) destruction of ozone through contact with surfaces 

and through reaction with nitric oxide (NO) at ground level; and (3) in s`itu — —~ 

5-105 
photochemical production of ozone (U.S. Environmental Protection Agency, 1978; 

Coffey et al., 1977; Mohnen, 1977; Reiter, 1977a). Obviously, meteorology is 

a controlling factor; if strong winds disperse the precursors or heavy clouds 

intercept the sunlight, high ozone levels will not develop. Another important 

variation on the basic diurnal pattern appears in some localities as a secondary 

peak in addition to the early afternoon peak. This secondary peak may occur 

any time from midafternoon to the middle of the night and is attributed to 

ozone transported from upwind areas where high ozone levels have occurred 

earlier in the day. Secondary peak concentrations can be higher than concentrations resulting from the photochemical reactions of locally emitted precursors 

(Martinez and Singh, 1979). At one nonurban site in Massachusetts (August 

1977), for example, primary peak concentrations of about 0.11, 0.14, and 0.14 

occurred at noon, from noon to about 4:00 p.m., and at noon, respectively, on 

3 successive days of high ozone levels (Martinez and Singh, 1979). Secondary 

peaks at the same site for those same 3 days were about 0.150, 0.157, and 

0.130 ppm at about 6:00 p.m., 8:00 p.m., and 8:00 p.m., respectively (Martinez 

and Singh, 1979). 

Because weather patterns, ambient temperatures, and the intensity and 

wavelengths of sunlight all play important roies in oxidant formation, strong 

seasonal as well as diurnal patterns exist. The highest ozone levels generally 

occur in the spring and summer (second and third quarters), when sunlight 

reaching the lower troposphere is most intense and stagnant meteorological 

conditions augment the potential for ozone formation and accumu1ation. Average 

summer afternoon levels can be from 150 to 250 percent of the average winter 

afternoon levels. Minor variations in the smog season occur with location, 

however. In addition, it is possible for the maximum and second-highest 

1-hour ozone concentration to occur outside the two.quarters of highest.average 

ozone concentrations. Exceptions to seasonal patterns are potentially important 

considerations with regard to the protection of crops from ozone damage, 

especially since respective crops have different growing seasons in terms of 

length, time of year, and areas of the country in which they are grown. 

Tn addition to the seasonal meteorological conditions that obtain in the 

lower troposphere, stratospheric-tropospheric exchange mechanisms exist that 

produce relatively frequent but sporadic, short-term incursions into the 

troposphere of stratospheric ozone (see Chapter 3). Such incursions show a 

seasonal pattern, usually occurring in late winter or early spring. 

5-106 
Percentile distributions, by season of the year, of concentration data.. 

from all eight NAPBN sites show that the arithmetic mean 1-hour concentration 

(averaged over a minimum of 3 years of data at each site, for 1977 through. 

1983) was higher in the second quarter of the year (April, May, June) at seven 

of the eight stations; and was only negligibly lower than the third-quarter 

value at the eighth station. The maximum l-hour concentrations at respective 

sites, aggregated over 3 to 6 years, depending on the data record for each 

site, ranged from 0.050 ppm at Custer NF, MT (in the fourth quarter) to. 

0.155 ppm at Mark Twain NF, MO (in the third quarter). The second-highest 

1-hour concentration among maximum daily 1-hour values ranged from 0.050 ppm 

at Custer NF, MT (in the fourth quarter), to 0.150 ppm at Mark Twain NF, MO 

(in the third quarter). The data also show that 99 percent of the 1-hour 

concentrations measured were well below 0.12 ppm, even in the second quarter 

of the year, when incursions of stratospheric ozone are expected to affect, 

at least to some degree, the concentrations measured at these stations. 

Excursions above 0.12 ppm were recorded in 1979 and 1980 at NAPBN sites; but 

none were recorded in 1981 (Evans et al., 1983; Lefohn, 1984). 

Because of the diurnal patterns of ozone, averaging across -longer- term 

periods such as a month, a season, or longer masks the occurrence of peak: 

concentrations (see, e.g., Lefohn and Benedict, 1985). This is an obvious.and 

familiar statistical phenomenon. It is pointed out, however, because it has 

direct reievance to the protection of public health and welfare. Averaging 

times must correspond to, ~ or be related in a consistent manner to, the pattern 

of exposure that elicits untoward responses. 

5.8.5 Spatial Patterns in Ozone Concentrations 1 . . 1 

In addition to temporal variations, both macro- and microscale spatial 

variations in ozone concentrations occur that have relevance ranging from 

important to inconsequential for exposure assessments. Differences in concentrations or patterns of occurrence, or both, are known to exist, for exampie, 

between urban and nonurban areas, between indoor and outdoor air, within large 

metropolises, and between lower and higher elevations. The more important 

variations are summarized below. 

5.8.5.1 Urban-Nonurban Differences in Ozone Concentrations. Ozone concentrations differ between urban and rural, between urban and remote, and even 

between rural and remote sites, as discussed in part in the preceding section 

5-107 
on temporal variations. The variations with area and type of site are variations both in the timing and the magnitude of the peak concentrations, and, in 

the case of transported ozone, are related to the temporal variations between 

urban and nonurban areas discussed above. Data from urban, suburban, rural, 

and remote sites (see, e.g., SAROAD, 1985a-f; Martinez and Singh, 1979; Lefohn, 

1984; Evans, 1985; respectively) corroborate the conclusion drawn in the 1978 

criteria document (U.S. Environmental Protection Agency, 1978) that ozone 

concentrations can sometimes be higher in some suburban or even rural areas 

downwind of urban plumes than in the urban areas themselves; and, furthermore, 

that higher concentrations can persist longer in rural and remote areas, 

iargely because of the absence of nitric oxide (NO) for chemical scavenging. 

In nonurban areas downwind of urban plumes, peak concentrations can 

occur, as the result of transport, at virtually any hour of the day or night, 

depending upon many factors, such as the strength of the emission source, 

induction time, scavenging, and wind speed (travel time) and other meteorological 

factors. The dependence of the timing of peak exposures upon these transportrelated factors is well-known and is illustrated here by two studies. Evans 

et al. (1983) calculated multiday trajectories for air parcels reaching a 

nonurban sites in the Mark Twain National Forest, Missouri, during an episode. 

Four separate trajectories, all of which passed over the Ohio River Valley and 

the Great Lakes region, impacted the forest site at different times in a 

24-hour period (in which the maximum 1-hour concentration measured was 0.125 

ppm). Subsequently, regional cloud cover and rains produced shifts in air 

flow and also reduced the potential for ozone formation, alleviating the 

impact at the site. Kelly et al. (1986) showed in the Detroit area that peak 

ozone concentrations occurred at distances of 10 to 70 km (ca. 6 to 43 mi) 

north-northeast of the urban center. Consequently, it would be possible for 

peak ozone concentrations to occur in the late afternoon or early evening in 

nonurban areas downwind of Detroit. Kelly et al. (1986) also found that 

concentrations diminished again beyond 70 km (ca. 43 mi) downwind of the urban 

center. Thus, as illustrated by these and similar data, beyond the distance 

traversed in the time required for maximum ozone formation in an urban plume, 

ozone concentrations will decrease (unless fresh emissions are injected into 

the plume) as the rate of ozone formation decreases, the plume disperses, 

surface deposition or other scavenging occurs, and meteorological conditions 

intervene. 

~ ~: 
It is not surprising, therefore, that in rural areas lying beyond the 

point of maximum ozone formation, for a given set of conditions, peak concentrations are lower and average diurnal profiles are flatter than in urban and 

near-urban areas (see, e.g., SAROAD, 1985b-f, for rural and remote sites). In 

remote areas beyond the influence of urban plumes, average peak concentrations 

will be still lower and average diurnal profiles still flatter (see e.g., 

Evans, 1985). Exceptions to these generalizations occur, of course, because 

of the complex interactions of topography, meteorology, and photochemistry. 

Such temporal and spatial differences between ozone concentrations in 

urban versus nonurban areas are important considerations for accurately assessing 

actual and potential exposures for human populations and for vegetation in 

nonurban areas, especially since the aerometric data for nonurban areas are 

far from abundant. 

5.$.5.2 Geographic, Vertical, and Aititudinal Variations in Ozone Concentrations . 

Although of interest and concern when estimating global ozone budgets, demonstrated variations in ozone concentrations with latitude and the lesser variations 

with longitude have little practical significance for assessing exposure 

within the contiguous llnited States. The effects on ozone concentrations of 

latitude and longitude within the contiguous states are minor, and are reflected 

in the aerometric data bases. Of more importance, ozone concentrations are 

known to increase with increasing height above the surface of the earth. 

Conversely, they may be viewed as decreasing with proximity to the surface of 

the earth, since the earth's surface acts as a sink for ozone (see, e.g. , 

Seiler and Fishman, 1981; Galbally and Roy, 1980; Oltmans, 1981, cited in 

Logan et al., 1981). The most pertinent vertical and altitudinal gradients in 

ozone concentrations are: (1) increases in concentration with height above 

the surface of the ground (regardless of altitude); (2) increases in concentrati on wi th al ti tude; and (3) vari ati ons i n concentrati ons wi th el evati on i n 

mountainous areas, attributable to transport and overnight conservation of 

ozone aloft, nocturnal inversions, trapping inversions, upslope flows, and 

other, often location-specific interactions between topography, meteorology, 

and photochemistry. 

5-109 
The importance of monitoring concentrations at the proper height,above 

the surface of the ground has been known for a long time, and EPA guidance on 

the placement of monitoring instruments. (see Chapter 4) is predicated on the 

existence of a vertical gradient as ozone is depleted by reaction with groundlevel emissions of NO or by deposition on reactive surfaces such as vegetation. 

Data illustrative of the near-surface gradient were reported by Pratt et al. 

(1983), who measured ozone concentrations at two separate heights (3 and 1 9 or 

6 and 9 meters) above the ground at three rural,,vegetated sites. Although 

the maximum mean difference between 3 and 9 meters was 3 ppb, this difference 

was similar to the mean difference between sites at the same height. Given 

the height of some vegetation canopies, especially forests, even such small 

differences over a spread of 6 meters should probably be taken into consideration when interpreting reported dose-response functions. 

The gradual increase in ozone conce'ntrations with altitude has been 

documented by a number of workers (see e.g., Viezee et al., 1979; Seiler and 

Fishman, 1981; Oltmans, 1981, as cited in Logan et al., 1981). There is a 

general increase in concentration with altitude, but as described by Seiler 

and Fishman (1981) and Oltmans (1981; cited in.Logan et al., 1981), for example, 

two relatively pronounced gradients.exist, one between the surface of the 

earth and 2 km (ca. 1 mi) and one even more pronounced between 8 and 12 km 

(ca. 5 and 7.5 mi). 

Increases in concentrations with altitude could potentially be of some 

consequence for passengers and airline personnel on high-altitude flights in 

the absence of adequate ventilation-filtration systems (see Chapter 11). 

Variations with height above the surface and with elevation, in mountainous 

areas, however, should be taken into account to ensure the accurate assessment 

of exposures and the accurate derivation of dose-response functions, especially 

for forests and other vegetation. 

Variations in ozone concentrations with elevation, not always consistent 

or predictable, have been reported by researchers investigating the effects of 

ozone on ~the mixed-conifer forest ecosystem of the San Bernardino Mountains of 

California. Measurements taken at four monitoring stations at four different 

elevations showed that peak ozone concentrations occurred progressively later 

in the day at progressively higher elevations (Miller et al., 1972). Ozone 

concentrations >0.10 ppm occurred for average durati_ons of 9, 13, 9, and 

5-110 
8 hr/day at the four respective stations; going from iower.to higher e7evations. 

The occurrence for 13 hr/day of concentrations >0.10, ppm at the station at 

817 m(2860 ft) was probably the result of contact of that.zone of the mountainside with the inversion layer (U.S. Environmental Protection Agency, 1978). 

Nighttime concentrations rarely decreased below 0.05 ppm_at the mountain crest; 

whereas at the lowest elevation, the basin station at 442 m(1459 ft), the 

nighttime. concentration usually decayed to near zero. Trapping inversions 

were major..contributor.s to the elevational gradients observed i,n this study, 

which was conducted in the 1970s.. Oxidant concentrations within the inversion 

were found not to be uniform but to occur in multiple layers. and strong vertical 

gradients. The important result of the trapping of oxidants in the inversion 

layers was the prolonged contact of high terrain with oxidants at night (U.S. 

Environmental Protection Agency, 1978). 

In`a more recent report, Wolff et al. (1986)-described measurements made 

in July 1975 at three separate elevations at Hig.h Point Mountain in northeastern 

New Jersey. The'daily ozone maxima were similar at different elevations. At 

night, however, ozone concentrations were nearly zero in the .valley but increased 

with elevation on the mountainside. Greater cumulative doses .(number of hours 

at >0.08 ppm) were sustained at the higher elevations, 300 and 500 m, respectively (ca. 990 and 1650 ft, respectively): Wolff et al. (1986) related this 

phenomenon to the depth of the nocturnal inversion layer and the contact with 

the mountainside of.ozone conserved aloft at night. 

These -concentration gradients with increased elevation are important for 

accurately describing concentrations at which injury or damage to vegetation, 

especially forests, may occur. Researchers investigating the effects of ozone 

on forest ecosystems have seldom measured nighttime ozone concentrations 

because the stomates of most , species are thought tot be closed at night, thus 

preventing the internal flux of ozone that produces injury.or damage (see 

Chapter 6). If stomates remain even partially open at night, however, the 

possible occurrence of nighttime peak concentrations of ozone, the occurrence 

of multiple peaks in a 24-hour,period, or the persistence of elevated concentrations that do not decay to near zero overnight should not be overlooked. 

Furthermore, the lack of N0 for nighttime scavenging in nonurban areas and the 

persistence of ozone overnight at,higher elevations will result in the presence 

of relatively higher concentrations in.such areas at sunrise when the stomates 

5-111 
open and photosynthesis begins. This possibility requires that exposure 

assessments, in the absence of sufficient aerometric data for forests and 

other vegetated areas, take such factors into consideration. 

5.8.5.3Other Spatial Variations in Ozone Concentrations . Other spatial variations are important for exposure assessments for human populations. Indooroutdoor gradients in ozone concentrations are known to occur even in buildings 

or vehicles ventilated by fresh air rather than air conditioning (e.g., Sabersky 

et al., 1973; Thompson et al., 1973; Peterson and Sabersky, 1975). Ozone reacts 

with surfaces inside buildings, so that decay may occur fairly rapidly, depending 

upon the nature of interior surfaces and furnishings (e.g., Davies et al., 1984; 

Contant et al., 1985). Ratios of indoor-to-outdoor (I/0) ozone concentrations 

are quite variable, however, since cooling and ventilation systems, air infiltration or exchange rates, interior air circulation rates, and the composition 

of interior surfaces all affect indoor ozone concentrations. Ratios (I/0, 

expressed as percentages) in the literature thus vary from 100 percent in a 

non-air-conditioned residence (Contant et al., 1985); to 80 ± 10 percent 

(Sabersky et al., 1973) in an air-conditioned office building (but with 100 percent outside air intake); to 10 to 25 percent in air-conditioned residences 

(Berk et al., 1981); and to as low as near zero in air-conditioned residences 

(Stock et al., 1983; Contant et al., 1985). 

On a larger scale, within-city variations in ozone concentrations can 

occur, even though ozone is a"regional" pollutant. Data show, for example, 

relatively homogeneous ozone concentrations in New Haven, Connecticut (SAROAD, 

1985a), a moderately large city that is downwind of a reasonably well-mixed 

urban plume (Wolff et al., 1975; Cleveland et al.; 1976a,b). In a large 

metropolis, however, appreciable gradients in ozone concentrations can exist 

from one side of the city to the other, as demonstrated for New York Ci.ty 

(Smith, 1981), and for Detroit (Kelly et al., 1986). Such gradients should be 

taken into consideration, where possible, in exposure assessments. 

5.8.6Concentrations and Patterns of Other Photochemical Oxidants 

5.8.6.1Concentrations . No aerometric data are routinely obtained by Federal, 

state, or local air pollution agencies for any photochemical oxidants other 

than nitrogen dioxide and ozone. The concentrations presented in this document 

for non-ozone oxidants were all obtained in special field investigations. The 

limitations in the number of locations and areas of the country represented in 

5-112 
the information presented simply reflect the relative paucity of data in the, 

published literature. 

The four non-ozone photochemical oxidants for which at least minimai 

concentration data are available are.formi.c acid, peroxyacetyl nitrate (PAN), 

peroxypropionyl nitrate (PPN), and hydrogen.peroxide. (H 2O2 ). Peroxybenzoyl 

nitrate has not been clearly identified in ambient air in the United States. 

The highest concentrations of PAN reported in the older literature, 1960 

through the present, were those found in the Los Angeles area: 70 ppb (1960), 

214 ppb (1965); and 68 ppb (1968) (Renzetti and Bryan, 1961; Mayrsohn and 

Brooks, 1965; Lonneman et al., 1976; respectively). 

The highest concentrations of PAN measured and reported in urban areas in 

the past 5 years were 42 ppb at Riverside, California, in 1980 (Temple and 

Taylor, 1983) and 47 ppb at Claremont, California, also in 1980 (Grosjean 

1981). These are clearly the maximum concentrations of PAN reported for 

California and for the entire country in this period. Other maximum PAN 

concentrations measured in the last decade in-the Los Angeles Basin have been 

in the range of 11 to 37 ppb. Average concentrations of PAN in the Los Angeles 

Basin in.the past 5 years have ranged from 4 to 13 ppb (Tuazon et al.-, 1981a; 

Grosjean, 1983; respectively). The only published study covering urban PAN 

concentrations outside California in the past 5 years is that of Lewis et al. 

(1983) for New Brunswick, New Jersey, in which the average PAN concentration 

was 0.5 ppb and the maximum was 11 ppb during September 1978 through May 1980. 

Studies outside California from the early 1970s through 1978 showed average 

PAN concentrations ranging from 0.4 ppb in Houston, Texas, in 1976 (Westberg 

et al., 1978) to 6.3 ppb in St. Louis, Missouri, in 1973 (Lonneman et al., 

1976). Maximum PAN concentrations outside California for the same period 

ranged from 10 ppb in Dayton, Ohio, in 1974 (Spicer et al., 1976) to 25 ppb in 

St. Louis (Lonneman et al., 1976). 

The highest PPN concentration reported in studies over the period 1963 

through the present was 6 ppb in Riverside, California (Darley et al., 1963). 

The next highest reported PPN concentration was 5 ppb at St. Louis, Missouri, 

in 1973 (Lonneman et al., 1976). Among more recent data, maximum PPN concentrations at respective sites ranged from 0.07 ppb in Pittsburgh, Pennsylvania, in 

1981 (Singh et al., 1982) to 3.1 ppb at Staten Island, New York (Singh et al., 

1982). California concentrations fell within this range. Average PPN concentrations at the respective sites for the more recent data ranged from 0.05 ppb at 

Denver and Pittsburgh to 0.7 ppb at Los Angeles in 1979 (Singh et al., 1981). 

5-113 
Altshuller (1983) has succinctly summarized the nonurban concentrations 

of PAN and PPN by pointing out that they overlap the lower end of the range of 

urban concentrations at sites outside California. At remote locations, PAN 

and PPN concentrations are lower than even the lowest of the urban concentrations by a factor of 3 to 4. 

The concentrati ons of H2O2 reported i n. the 1 i terature to date must be 

regarded as inaccurate since ozone is now thought to be an interference in all 

methods used to date except FTIR (Chapter 4). Measurements by FTIR, the most 

specific and accurate method now available, have not demonstrated unambiguously 

the presence of H 2O2 in ambient air, even in the high-oxidant atmosphere of the 

Los Angeles area. (The limit of detection for a 1-km-pathlength FTIR system 

is around 0.04 ppm.) 

Recent data indicate the presence in urban atmospheres of only trace 

amounts of formic acid: < 15 ppb, measured by FTIR (Tuazon et al., 1981b). 

Estimates in the earlier literature (1950s) of 600 to 700 ppb of formic acid 

in smoggy atmospheres were erroneous because of faulty measurement methodology 

(Hanst et al., 1982). 

5.8.6.2 Patterns. The patterns of formic acid (HCOOH), PAN, PPN, and H 2O2 

may be summarized fairly succinctly. Qualitatively, diurnal patterns are 

similar to those of ozone, with peak concentrations of each of these occurring 

in close proximity to the time of the ozone peak. The correspondence in time 

of day is not exact, but is close. As demonstrated by the wor.k of Tuazon 

et al. (1981b), ozone concentrations return to baseline levels somewhat faster 

than the concentrations of PAN, HCOOH, or H 2O2 (PPN was not measured). 

Seasonally, winter concentrations (third and fourth quarters) of PAN are 

lower than summer°concentrations (second and third quarters). The percentage 

of PAN concentrations (PAN/0 3 x 100) relative to ozone, however, is higher in 

winter than in summer. Data are not readily available on the seasonal patterns 

of the other non-ozone oxidants. 

Indoor-outdoor data on PAN are limited to one report (Thompson et al., 

1973), which confirms the pattern to be expected from the known chemistry of 

PAN; that i s, i t persi sts 1 onger i ndoors than ozone. Data are 1 acki ng on 

indoor-outdoor ratios for the other non-ozone oxidants. 

5.8.7 Relationship Between Ozone and Other Photochemical Oxidants 

The relationship between ozone concentrations and the concentrations of 

PAN, PPN, H2O2 , and HCOOH is important only if these non-ozone oxidants are 

5-114 
shown to produce potentially adverse health or welfare effects, ~ singly, in 

combination with each other, or in various combinations with ozone at concentrati ons correpondi ng to those found -i n ainbi ent ai r. _ If only ozone i s shown to 

produce adverse health or welfare effects in the concentration ranges of 

concern, then only ozone must be controlled. If any or all'of these other 

four oxidants are shown to produce potentially adverse health or welfare 

effects, at or near levels found in ambient air, then such oxidants will also 

have to be controiled. Since ozone and all four of the other oxidants arise 

from reactions among`the same organic and inorganic precursors, an obvious 

question is whether the control of ozone will also result in the control of 

the other four oxidants. 

Controlled-exposure studies on these non-ozone oxidants have employed 

concentrations much higher than those found in ambient air (see Chapters 9 and 

10). Because PAN m'ay have contributed, however, to the eye irritation symptoms 

reported in earlier epidemiological studies, and because PAN is the most. 

abundant of these non-ozone oxidants, the relationship between ozone and PAN 

concentrations in ambient air remains of interest. 

- The patterns of PAN and ozone concentrations are not quantitatively 

similar but do show'quaiitative similarities for most locations at which both 

pollutants have been measured in the same study. That a quantitative, monotonic 

relationship between ozone and PAN is'lacking, however, is shown by the range 

of PAN-to-ozone ratios, expressed as'percentages, between locations and.at the 

same location, as reported in the review of Altshuller (1983). 

Certain other information bears out the lack of a monotonic relationship 

between PAN and ozone. Not only are PAN-ozone relationships not consistent 

between different urban areas, but they are not consistent in urban versus 

nonurban areas, in summer versus winter, in indoor versus outdoor environments, 

or even, as the data show, in location, timing, o.r magnitude of diurnal peak 

concentrations within the same city. Data obtained in Nouston by Jorgen 

et al'. (1978), for example, show variations in peak concentrations of PAN and 

in relationships to ozone concentrations of those peaks among three separate 

monitoring sites.` Temple and Taylor (1983) have shown that PAN concentrations 

are a greater percentage of ozone 'concentrations in winter than-in the.remainder 

of the year in California. Lonneman et al. (1976) demonstrated that PAN, 

absolutely and as a percentage o:f ozone, is considerably lower "in nonurban 

than in urban areas. Thompson et al. (1973), in what is apparently the only 

5-115 
published report on indoor concentrations of PAN, showed that PAN persists 

longer than ozone indoors. (This is to be expected from its enhanced stability 

at cooler-than-ambient temperatures such as found in air-conditioned buildings.) 

Tuazon et al. (1981b) demonstrated that PAN persists in ambient air longer 

than ozone, its persistence paralleling that of nitric acid, at least in the 

locality studied (Claremont, CA). Reactivity data presented in the 1978 

criteria document for ozone and other photochemical oxidants indicated that 

all precursors that give rise to PAN also give rise to ozone. The data also 

showed, however, that not all precursors giving rise to ozone also give rise 

to PAN, and that not all that give rise to both are equally reactive toward 

both, with some precursors preferentially giving rise, on the basis of units 

of product per unit of reactant, to more of one product than the other (U.S. 

Environmental Protection Agency, 1978). 

In the review cited earlier, Altshuller (1983) examined the relationships 

between ozone and a variety of other smog components, including PAN, PPN, 

H2O 2 , HCOOH, aldehydes, aerosols, and nitric acid. He concluded that "the 

ambient air measurements indicate that ozone may serve directionally, but 

cannot be expected to serve quantitatively, as a surrogate for the other 

products" (Altshu11er, 1983). It must be emphasized that the issue Altshuller 

examined was whether ozone could serve as an abatement surrogate for all 

photochemical products, not just the subset of non-ozone oxidants of concern 

in this document. Nevertheless, a review of the data presented indicates that 

his conclusion is applicable to the non-ozone oxidants examined in this document. 

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