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 

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®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 concentrationsoccurring 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 otherphotochemical 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; effectsobserved 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 -0 3 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 2 O 2 , 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 NO 2 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

ixTABLE 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

xLIST 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 NO 3 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. NO x concentrations 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

xiiLIST 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, NO 2i 0 3 , 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 H 2 O 2 (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 10 2 ) 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 ofthe y eag

h t C na ti"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

xvLIST 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 atfour 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 

### CH 4 methane CO carbon monoxide
CO 2 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 HO Z 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/m 3 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 NH 3 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 NO x nitrogen oxides NO2 nitrogen dioxide NO3 nitrogen trioxide 

### N Z O nitrous oxide NR natural rubber NYCC New York City Driving Schedule


### 0 2 

### oxygen 

### 0 3 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/m 3 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

xxiiChapter 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

### Cha pter 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.

xxvChapter 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

xxviChapter 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.

xxviiChapter 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.

xxxSCIENCE 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 of Air 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 peroxyacylnitrates, 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 (0 3 ), 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 ofvarious 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 respiratorytract, 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 underprovisions 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 thiscountry 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 S0 2 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 S0 2 , 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 S0 2 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).

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

photochemical oxidants are discussed.in this document, although some of thetopics included are relevant to abatement strategies. Technology for controlling the emissions of'nitrogen oxides and of volatile organic compounds isdiscussed 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 scientificbasis 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 arenot 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) assessmentof 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 betterunderstanding 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 weremultiple-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 andsublethal effects. In the areas of mutagenesis, teratogenesis, and reproductive effects, however, results of studies conducted at much higher than ambientlevels 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 notrestricted 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

2-6effect. 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 NO x (NMOC/NO x ); and the reactivity of the organic precursors. This chapter describes the physical and chemical properties ofozone 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 (0 3 ) 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 O o C, 760 mm Hg At 25 0 C, 760 mm Hg Solubility at O o C (Indicated volume of ozone at O o C, 760 mm Hg)
Hen's Law constant, 

### 37CandpH=7
Conversion factors At O o C, 760 mm Hg

At 25 0 C, 760 mm Hg Colorless gas; blue violet liquid
0 3
48.0
-192.7 f 0.2 0 C
-111.9 t 0.3 0 C

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 -4 ppm
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 3 C(0)0 2 NO 2 , 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 2 O 2 . 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 researchersrely 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 0 C 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 (H 2 O 2 ) 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 2 O 2 is an inorganic aci-d that has a dissociation constant of 2.4 x 10 -12 and a pK of 11.62 (at 25 0 C) (Weast, 1977). Hydrogen peroxide has a standard potential of +1.776 

## in the redox pair, H 2 O 2 /H 2 O. The physical properties of H 2 O 2 are given in Table 3-4. Additional properties should be noted here that are of interest relative to whether effects of H 2 0 2 in biological receptors are of siynificance. First, H 2 O 2 , 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, @25 0 C Colorless liquid
Chemical formula CH3C00 2 NO 2 Molecular weight 121 Boiling point, o C 106t 2 a 103.9 b 

### Triple point, o C -50 t 0.5 a -48 t 0.5 b Vapor pressure, @room temperature —15 mm Hg Vapor pressure curve ln p=-4587/T + 18.76 a,c ln p = -4585/T + 18.79 b Hydrolysis t l 2 = 3.5 min at pH = 7.8 d ; In alkaline solution 

# ~ roducts include nitrite ion and molecular oxygen e
In acidic solution f @22 0 C, pH 5.6 4 x 10 -4 sec -1

@25 0 C, pH 5.6 6.8 x 10 -4 sec -lg Henry's Law constant, @10 0 C 5 t 1 M atm -lg
Conversion factors @O o C, 760 mm Hg 1 ppm = 5 398 /m 3 1 μg/m 3 = 1

~ 

### 52 x 10 -4 ,ppm
@25 0 C, 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 o K; 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 (1983 a ) 12.4 32.6c 13.6 15.8 13.4 Stephens (1964, 1969 b ) 10.0 23.6d 11.2 14.3 10.1

a 

### At 4 mbar; no diluent; resolution 1.2 cm -1

. b 

### At 7 mbar in N 2 diluent at 973 mb total pressure and 295 K; grating instrument.

c 

### Q branch resolved (1.5 torr).
d 

### Q branch not resolved (1.4 mbar).

TABLE 3-4. PHYSICAL PROPERTIES OF HYDROGEN PEROXIDE

Physical state, @25 0 C
Chemical formula Molecular weight Melting point, oC Boiling point, oC, @760 mm Hg Density, @25 0 C, 760 mm Hg Vapor pressure, @16.3 0 C
Conversion factors @OoC, 760 mm Hg

@25 0 C, 760 mm Hg

Source: Weast (1977). Colorless liquid
H 2 O 2
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 2 O2 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 2 O 2

.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 2 O0 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 0 C.



### 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 alongwith 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 (CH 4 ) 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 othergas-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 researchershave 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 hydrogenand 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 H 2n+2 . The first compound in the series is methane, CH4, which, because of its low reactivity, does notcontribute 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 H 2n . 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 10 H 16 . 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 H 8

. 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 carbonand 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 ofphotochemical reactivity,with benzene having a low photochemical reactivity and 1,3,5-trimethylbenzene, for example, showing high reactivity. 3.2.2.2 Aldehydes 

### . 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,

3-10since 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). 3.2.2.3 

### Other 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. 3.2.2.4 

### Volatility 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 atmas 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 ofhydrocarbons and of other volatile organic compounds is relevant to mechanistic studies in atmospheric chemistry, to modeling, and to otheroxidant-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 conditionsapproximating 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.

3-12The 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 correspondingto 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 (N 2 0). 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 NO 2 , 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( 3 P)] produced from the photolysis of NO2 with 0 2 (in the .• presence of a third, energy-absorbing molecule); and (3) the subsequent regeneration of NO 2 by the reaction of NO with 0 3 . Coupled with these basic reactions are reactions between NO and free radicais iri the atmosphere (hydroperoxy, alkylperoxy and acylperoxy) that oxidize NO to NO 2 , 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 (N 2 0 4 ).



## 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 formationof 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 (0 3 ) 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 (photoly

### sis) 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 0 3 produced from the reaction between atomic and molecular oxygen. In these cyclic reactions, no net increase in 0 3 occurs, with the result that an equilibrium is set up among 0 31 NO 21 and N0. Any reactions that produce NO 2 without destroying 0 3 will 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 2 without destroying 0 31 thusupsetting the equilibrium. The reactions of these radicals with NO also constitute a cyclic process. Since terminating reactions occur between NO 2 and these radicals, as well, which remove both NO 2 and the radicals from the photochemica1 reaction system, the cycles described above would gradually end, even in the presence of sunlight, unless fresh NO x emissions 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:NO x ratio. In the following sections, the processes just described in a simplistic andsummary 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 2 by near-ultraviolet solar radiation is a critical process:

NO 2 + 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) + 0 2 + M-> 0 3 + 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 + 0 2 (3-3)

As a result of the above three reactions, an equilibrium or steady-state condition is established among N0, NO 2 and 0 3 , and the concentration of 03 in the atmosphere is governed by the expression,
[NO 2 ] [0 3 

## 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 downwindsuburban areas to which the resulting NO2 is transported and then photodissociated, leading to 03 formation. The characteristic behavior of irradiated NMOC-NO x 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). 3.3.1.2 Formation 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 NO 2 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).

3-170.6



### 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) + H 2 0 + 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)
H 2 + 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 + 0 2 

### + H0 2

### + M (3-10)



### These can then react with NO to form hydroxyl radicals:

H0 2 

### + 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 H0 2 radicals and hence OH 

### radical formation:

HCO + 0 2 

### + 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:

RCH 2 O + 0 2 + RCHO + H0 2 (3-14)



### These H0 2 

### radicals will also oxidize NO to NO 2 

### via reaction (3-11). Based upon environmental chamber data, computer modeling studies andmeasured 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 10 8 to 10 9 radicals cm

-3 , respectively.

3-20

### 3.3.1.2.2 Nitrate Radicals 

### . Ozone can react with NO 2 to pr.oduce the

nitrate radical and an oxygen molecule:

0 3 + NO 2 + NO 3 + 0 2 ; (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:

NO 3 + hv + NO 2 + 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 , NO 3 and dinitrogen pentoxide (N20 5 )- 3.3.1.3 Termination 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 theatmosphere. 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 + HNO 3 (3-18)

M H0 2 + NO 2 ; H0 2 NO 2 (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 NO x
. 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 H0 2 :

H0 2 + H02 + H 2 O 2 + 0 2 (3-20)

could be an important route to the oxidation of S0 2 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 , H0 2 NO 2 and H 2 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:

H0 2 + NO 2 + HONO + 0 2 (3-21)

3-22This 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 2 0 results in s-till another 

### potential source of HONO:

NO + NO 2 + H 2 0 ~ 

### 2 HONO (3-22)

2 HONO -> NO + NO 2 + H 2 0 (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 NO 2 with water may proceed in the gas phase or on surfaces:



### 2 NO 2 + H 2 0 -> 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 ; NO 3 , and 

### N 2 0 5 : 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 -1 1 cm 3 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

### N 2 0 5 + H 2 0 -> 2 HNO 3 (3-26)

Thus, for NO 2 and NO 3 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 + NO 3 ~ N 7 0 5 

### equilibrium constant of 3 x 10 -11 cm 3 molecule -1 , an HNO 3 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 cm 3 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 NO 2 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. ~



### 3.3.2 

### Organic 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' + H 2 0 (3-27)

M R' + 0 2 + R0 2 ' (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)

R0 2 ' + NO 2 

## E R0 2 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' + 0 2 -> R'OR + H0 2 ' (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. 3.3.2.1.2 Alkenes 

### . 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 0 2 addition, with the oxidation of NO to NO 2 ,by the resulting peroxy radical to form an alkoxy radical (Atkinson et al., 1985):

OH' + CH 3 CH=CH 2 

### + CH 3 CHCH 2 OH 

### (3-37)

00'

. I CH 3 CHCH 2 OH + 0 2 

### + CH 3 CHCH 2 OH 

### (3-38)

00' 0 ' CH 3 CHCH 2 OH 

### + NO + CH 3 CHCH 2 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:

CH 3 CHCH 2 OH + CH 3 CH0 + 'CH 2 OH 

### (3-40)

'CH 2 OH + 0 2 -). HCHO + H0 2 (3-41) 

# t NO

OH + NO 2
The overall reaction resulting from reactions (3-37) to (3-41) is:

OH + CH 3 CH=CH 2 

### + 20 2 + 2 NO -* CH 3 CH0 + HCHO + 2 NO 2 + OH (3-42)

1 hv, 0 2 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 CH 3 CHCH 2 OH

CH 3 CH0 + CH 2 OH

0 2 CH 3 CHOH + HCHO

0 2 OH
CH 3 CHCH 2 O' CH 3 CH=CH 2

(^-65%

CH 3 CHCH 2 OH (-35%)



# 7 H
CH 3 CHCH 2

. 0 2 0 2 NO NO 2 NO 31 , NO 2

H0 2+ HCHO CH 3 CH0 + H0 2

Figure 3-2. Reaction scheme for OH radical-initiated oxidation of propene in the presence of NO. Source: Atkinson (1985).

3-28The 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 thoseunsubstituted compounds emitted from vegetation. Examples include iso

### prene 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 andArnts, 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):


### I H3 

## r --> HOCH 2 -C-CH=CH 2 (3-43a)



## j H 3 OH + CH 2 =C-CH=CH 2



## j H 3 CH 2 =C-CH-CH 2 3-29 OH 

## j H 3 CH 2 

# ~ -CH=CH 2 (3-43b) OH
j H 3 CH 2 =C-CHCH 2 OH (3-43c)

( 3-43d )

### The hydroxyalkyl radicals formed in reaction,s (3-43a)-(3-43d) react .

rapidly with 0 2 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 0 2 (Lloyd et al., 1983):

CH 3 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 

## H 3 CH 2



### 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:



## CH 2 CH 2 O0' 

# 0
O + 0

2 

## -~ ~ ( 3-47 )



## 3-31

### H 2 00' CHZO'



# O + NO + O + NO 2 ( 3-48a')



### CH 2 ONO 2 

# 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 C 6 H 5 CH 2 O' radical then reacts with 0 2 to yield benzaldehyde and an H0 2 radical:



### CH 2 O' HO



# O + 0 2 ~ O + H0 2 (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 NO x -photooxidation chemistry of aromatic compounds is obtained.



### 3-323.3.2.1.4 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:

CH 3 CH0 + hv + CH 3 +.HCO (3-50)

Attack by OH radicals forms acetyl radicals which can successively add 0 2 and NO 2 to form PAN:

CH 3 CH0 + OH - ~ CH 3 C0 + H 2 0 (3-51)


# 11 NO2 i . CH 3 C0 + 0 2 + CH 3 C00' CH 3 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, S0 2 and CH 3 S0 3 H, together with CH 3 SN0 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 3 S and CH3SOH radicals.' Subsequent reactions of these CH 3 SOH and CH 3 S radicals then lead to the observed products: S02, HCHO, and CH 3 S0 3 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 H 2 SO 4 , 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 cm

3 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 cm 3 molecule -1 sec -1 can consumption by 0 3 be considered atmospherically;important. These ozone reactions of interest are summarized below. 3.3.2.2.1 Alkenes 

### . 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 cm 3 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 0 3 , rather than with OH radicals. It is now reasonably well established that the initial step in the 0 3 -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 ~ / R 3 ~ 

### / C=0 + C ~ 

## R 2 R 4 ~ 

### R1

~ 

### , / 0-0' ~ R 

### 3

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 0 3 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.

. CH 2 O0

[CH 2 O0] *



### /0]* /0'

CH 2

~ 0

~ [ CH 

### 2`--, 0 . f / CO 2 + H 2

~ 

### [HCOOH] * --! CO + H 2 0



# ~ H + HCO

•2 (40%)



### (129'0 ) (42%) (3-54)

20 2

2 HO ? + CO 2 (6%)

and

3-36m • e 

### -> CH 3 CH00 (40%) i



### [CH 3 CH00] 

## ~ 0 ~ CH 3 + CO + OH (19%) [CH 3 COH] * 

## ~ (3-55)

CH 3 + CO 2 + H (24°0)



### CH 0 * 0 

## ~C ~ [HCOCH 3 

## ]~ H HCO + CH 3 0 ( 5%)



## CH 4 + CO 2 (12%)

. e • s 

### where CH 2 O0 and CH 3 CH00 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 H 2 O, and it is believed that they will also react with NO.and NO 2 (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)

H 2 0 RCH00 + S0 2 —, RCHO + H 2 SO 4 (3-58)

RCN00 + H 2 0 -> RCOOH + H 2 0 (3-59)

RCH00 + CO products (3-60)

0 —0 

### RCH00 + R'CHO ~ 

### RCH

~ O 

### j CHR' (3-61)

Under atmospheric conditions, the reactions with N0, NO 2 , or H 2 0 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 H 2 0 (Atkinson and Lloyd, 1984).



### 3-37

### Hence, 0 3 -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 H 2 SO 4 , although the latter is probably a minor process in the overall oxidation of S0 2 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 cm

3 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 cm 3 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 cm 3

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 NO 3 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 0 3 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:

3-39U~0~ 

### 0 ' 

### CH 2 =CHCOCH 3 + 0 3 + CH 2 -CHCOCH 3 (3-62).

HCHO + [CH 3 COCH00] * CH 3 COCH0 + [CH 2 O0] * (3-63) and

~
0~ 0

~ 

### 0 / CH 3 t 1 / CH 3 CH 2 =C. ` + 0 3

-~ CH 2

-~

C 

### CHO CHO (3-64) CHO HCHO + CH 3 C00 CH 3 COCH0 + [CH 2 O0]

Kamens et al. (1982) discuss the possible subsequent reactions. 

### 3..3.2.2.5 Nitrogen-Containing Organics 

### . Studies of the kinetics of the reactions of 0 3 with a variety of nitrites, nitriles, nitramines, nitrosoamines, amines, and hydrazines (Atkinson and Carter, 1984) indicate that only for the hydrazines are reactions with 0 3 sufficiently rapid to be of atmospheric importance (Carter et al., 1981b; Tuazon et a1., 1982). 3.3.2.2.6

Sulfur-Containing Organics 

### . Based upon the kinetic data available for dimethyl sulfide, thiirane (C 2 H 4 S) and thiophene, it'appearsat 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). 3.3.2.2.7 Organometallics 

### . 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 0 3 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. 

### 3-403.3.2.3 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 3radicals 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:



### NO 3 + RH + HNO 3 + 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 cm 3 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 3radical to the olefinic double bond:

ONO 2



### ~ . . 2NO 3 + 2CH3CH=CH2 + 

### CH 3 CHCH 2 + CH 3 CHCH 2 ONO 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 CH 3 CH(ONO 2 )CH 2 O0NO 2 , and

stable dinitrates such as CH 3 CH(ONO 2 )CH 2 ONO 2 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 NO 3 radicals with the monoterpenes, NO 3 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 NO 3 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 certainof 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. 3.3.2.3.3 Aldehydes 

### . Based upon the product data of Morris and Niki (1974), i.e., the observed formation of HNO 3 from the reaction of NO 3radicals with CH3CH0, it is expected that these reactions proceed via Hatom abstraction from the relatively weak H-CO bonds:

NO 3 + RCHO -> RCO + HNO 3 (3-67)

Thus the reaction of NO 3 radicals with acetaldehyde could be a nighttime source of peroxyacetyl nitrate (PAN):

NO 3 + CH 3 CH0 -> HNO 3 + CH 3 C0 (3-68)

CH 3 C0 + 0 2 -> CH 3 CO3 (3-69)

1 CH 3 CO 3 ' + NO 2 + CH 3 COONO 2 (3-70)

Reaction of NO 3 radicals with the higher aldehydes will lead, by analogous reaction schemes, to the higher peroxyacyl nitrates,RCO 3 NO 2 . Reaction

3-42with formaldehyde, however, will lead to H02 radical formation, since HCO

reacts rapidly with 0 2(Atkinson and Lloyd, 1984):

HCO + 0 2 + H0 2 + 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 2 0 5 -NO 2 -phenol-air reaction mixtures, presumably formed by the reaction sequence:

OH 0'

NO 3 

## + ~ ~ HNO 3 + ~ (3-72)

foilowed by (Niki et al., 1979):

' OH OH NO 2 

# O + NO 2 + 

# ~ (3-73)



### NO 2

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 atnight. It can be seen that under the atinospheric conditions assumed reactions with 0 3 are important for the higher alkenes, including the monoterpenes during daylight when NO 3 radical concentrations are low, and for the hydrazines. 'For the other organics for which kinetic data are available, including alkanes and aromatics, reactions with 0 3 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 NO 3 RADICALS
Organic lifetimes a,b Organic compound p 3

, OH, NO 31 '. 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 

## n 3 -Carene 1.0 hr 1.7 hr 1.2 min d-Limonene 11 min 1.0 -hr 0.9 min

a 

### Time to reach 1/e of the initial concentration. b 

### Assuming 100 ppb of 0 3 (24 hr average), 2 x 10 6 molecule 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 CH 3 COONO 2 E

C H 3 C0 0 + NO 2 ( 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 CH 3 C(0)0 2 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 CH 3 C ! -0' -> CH 3 ' + CO 2



### M CH 3 ' + 0

2 -> CH 3 00

CH 3 00' + NO + CH 3 0' + NO 2



### CH 3 0' + 0 2 ~ H0 2 

### + HCHO (3-76)

(3-77)

(3-78)

(3-79)

H0 2 + NO -> OH + NO 2 

### (3-80)

3-45

### M OH + NO -> HONO (3-81)

M OH + NO 2 + HNO 3 (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 NO 2 , 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 HNO 3 , 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 (H 2 SO 4 ). 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 S0 2 oxidation at relative humidities lower than 75 percent. It is also clear, however, that the rate of reaction of 0 3 with S0 2 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 S0 2 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 HOS0 2 radicals from the reaction of OH radicals with S0 2 , followed by reaction with 0 2 , is the reaction mechanism for the formation of S0 3 :

OH + S0 2 (+ M) + HOS0 2 (+ M); (3-83)

HOS0 2 + 0 2 -> H0 2 + S0 3

. (3-84)

From the reaction of H0 2 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 S0 2 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 0 3 than the plume) (Davis et al., 1979; Blumenthal et al., 1981). Though it does not react directly with S0 2 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 S0 2 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 NO 2 . 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 NO 2 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 NO 2 by hydroperoxyl radicals and other peroxy radicals (reactions 3-3 and 3-11, respectively). Thus, if NO 2 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 2 0 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 NH 4 NO 3 (Doyle et al., 1979; Stelson et al., 1979; Appel et al., 1980), which, because of its low vapor pressure, will form nitrate aerosol particles:

HNO 3 + NH 3 --> NH 4 NO 3 (3-85)

If acidic sulfate is present, however, it will react with NH 4 NO 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 :

HNO 3 + 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 sufficientlylow 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 experimentalevidence 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 theanalysis 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 extractedfrom 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.

3.3.6 kole 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., 3.3.6.1 Reactions of Ozone in Aqueous Droplets 

### . While the thermal oxidation of S0 2 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 0 C 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 3 would 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 0 3 and 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 2 partial 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 S0 2 by 0 3 cannot 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 + 0 2 aq (3-87)

0( 1 D) aq + (H 2 0) aq 2(OH) aq (3-88)

and its reactions with aqueous OH - ions and H 2 O 2 to 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

3-511000

100

r M 

### E 10 M 

### E

~

~ .`c 0 1 • ~

~ c.

0.1

0.01 , 1 0-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 equilibriaare 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
cm 3 —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 H 2 O 2 will be taken up in aqueous droplets. Over the past decade, evidence has accumulated that H 2 O 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 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 H 2 O 2 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 S0 2 and NO 2 dissolved in aqueous droplets, the photolysis of H 2 O2 to produce aqueous OH radicals:

(H 2 O 2 ) aq + hv + 2(OH) 

# a q (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 (0 3 ) aq (Chameides and Davis, 1982). It should be emphasized, however, that substantial uncertainties remain concerning the quantitative role of H 2 O 2 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 2 O 2 may be formed in 

### situ in 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).

3-533.3.6.3 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

3-54approximation, 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 theground 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

3-55processes 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

3-56hours 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

3-57is 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 isoplethincludes 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 10 2 ) 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 16 3

„ 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 Delawarecoast. 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. 

### 3-60This 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.

3.4.2 

### Wind 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 inspite 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 codtributesto 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

3-610 N t

~ 

### E o Z

~ W J

~ 

## U
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~

~ ^
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W L_

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# _ c D

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# ~ ` C



## ~ 

## 3

a~ c ` r
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## 3 a

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y 

## ~ ~

~ 'h+ 7 ~

r N a~ 3 

# g o , L 

# ~ cc 1 

# ~ M O ms L

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### cD c ir. 2 N 

# Z
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# _

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3-62wind. 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; thatis, 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. Thisprobably 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.

3-634 16 19 • 28 13• 50 • 63 • 20 22 • 31
• 77 • 30 479 42 • 35 249 •16 034 • 27 440 5 43 • 6 160 120 •24 7 • 48 50

~ 13 •47 • • 39 28 57 • 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).

3-65In 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 thesource area, as well as along the trajectory followed by an urban or sourcearea plume. Regions having steady prevailing winds, such that a given airparcel 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.

3-66The 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/NO x systems 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/NO x mixtures, 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 x and n-butane/ NO x chemical 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

3-67clear 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 0 3 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

9101Agency, 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 -0 x ) 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 -0 x monitoring information. It has been'identified as a significant, characteristic feature of the 0 3 -0 x 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 -0 x after 1 or 2 hours of downwi nd travel. Maximum concentrations may be displaced up to 20 or so miles from thecenter 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 =0 x formation 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 -0 x concentrations from the urban core area is dependent on the local wind pattern and is, in general, inversely related to. the peak 0 3 -0 x concentration. Stronger winds will carry the air parcelsfarther during the reaction period, increasingly diluting pollutant concentrations along the trajectory. Weak winds and very restricted mixing heights will tend to cause higher 0 3 -0 x concentrations 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

3-69a 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 -0 x 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 -0 x incidents in northern coastal San Diego County, California. It showed conclusively that these were caused by the 0 3 -0 x 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 -0 x 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 0 3 -0 x 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 a t l., 1982). A major field program supported by local industries has been conducted in Houston, Texas, although the 0 3 -0 x 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 -0 x 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

3-70related 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 0 3 -0 x and 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 surfacehigh 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 specificinstances 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

3-71the 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 highpollution 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

3-72into 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 eastof 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 toemission 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 -0 x and 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 -0 x results from a complex interaction of distant and local precursor,sources,

3-73urban 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 areadsorption 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

3-74reactions 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.

3-75OZOCIE

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### 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).

3-76Since 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.

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### 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).

3.4.7 

### Stratospheric 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.



## EMOAccording 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.at ground-levelstations. 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 7 Be is a unique tracer for air parcels of stratospheric origin. Both ozone and 7 Be 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 7 Be 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

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linear correlation, r, equal to 0.65, for 1978 summer measurements of 7 Be 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 7 Be measurements to the quantification of the amount of stratospheric ozone in surface air. Some of the problems encountered when applying 7 Be/0 3 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 fundamentallydifferent 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%0 3 ratio has been firmly established.

As a result of these and other factors, Singh et al. (1980) concluded that "the experimental technique involving a 7 Be/0 3 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).,

` 3-83Other 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 90 Sr 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 andepisodic; 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.in air 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 1processes 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 NO x 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.ng studies, 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 0 3 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 NO x emissions corresponding to a concentration of

N0.25 ppb (Fishman and Seiler, 1983; Fishman and Carney, 1984); (2) 15 ppb, assuming zero NO x emissions (Fishman and Seiler, 1983); and (3) 80 ppb, using NO x emissions 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 x emissions inventory; or (2) —25 ppb, assuming zero manmade NO x emissions. 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 ofcontroversy 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 VOCemissions, 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 NO x in 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 x concentrations 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 x in 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 x are 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 NO x were 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 x concentrations in the free troposphere of 0.01 to 0.1 ppb and an NO x reservoir of <0.03 ppb; and the global models described in Kelly et al. (1984) assume an NO x background of 0.1 to 0.2 ppb. In contrast, the mean NO x concentrations 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 x concentrations range from 1 to 10 ppb (Altshuller, 1986). Martinez and Singh (1979) analyzed the role of NO x in nonurban ozone formation using theoretical approaches and aerometric data from the Sulfate Regional Experiment. They concluded that the impact of NO x on 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

3-87rural 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 x and 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 x in 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 alproduction) 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 

### (NO x ) are reported here as the sum. of 

### .NO and 

### NO 2 , al l expressed as equivalent NO 2 . Retrospective estimates of total manmade 

### NO x emissions 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).

a To 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[Pct i x no. i /mw

i ]/[Total NO x (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 NO x emissions 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 NO x emissions 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

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Figure 3-14. National trend in estimated emissions of volatile. organic compounds, 1970 through 1983. Source: U.S. Environmental Protection Agency (1984a). 33



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Source: Motor Vehicle Manufacturers Association (1984); U.S. Environmental Protection Agency (1984a).



### 3-92TABLE 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 x emitted 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 x emissions 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,

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### 3-94rubbing 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

3-95with 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 exhaustemissions 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 NO x emissions nationally. Stationary sources contributed50 percent and mobile sources contributed 45.4 percent in 1983 (U.S. Environmental Protection Agency, 1984a). In contrast to their contributions to VOC

3-96TABLE 3-10. YEARLY QUANTITIES OF THE UNITED STATES FOR HIGHWAY MOTOR VEHICLE FUELS SOLD USE, 1980 THROUGH 1983 IN



### Gallons 

# x 10 9 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 NO x einissions in 1983 (U.S: Environmental Protection Agency, 1984a). Ratios of NO/NO x in 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 x emissions. More than 30 to 50 percent, however, of the total NO x emissions from certain diesel (Braddock and Bradow, 1976; Springer and Stahman, 1977 a,b) and jet turbine engines (Souza and Daley, 1978) can be NO 2 under 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 x ratios 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 x from 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'NO x emissions from stationary sources are expected since power plants using fossil fuels produce an estimated 15 percent more NO x

3-97in 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 NO X 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 importantbecause of their potential impact on ambient air quality. Table 3-11 summarizes data on NO X emissions from mobile sources.

TABLE 3-11. SUMMARY OF NO x EMISSIONS 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

00referred 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 -1 hr -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.

3-99Besides 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 -2 hr -1 , while the enclosure technique yielded 7,300 pg m -2 hr -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
g the enclosure method, it, too, has certain limitations. For example, themeasurement 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

3-100Meeks, 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)

3-101Considering 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 NO x emissions, 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 NO x -N escapes to the troposphere (Robinson and Robbins, 1975; S6derlund and Svensson, 1976). Hill (1971) reported that NO and NO 2 are 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 NO x from terrestrial to aquatic systems, but concluded that losses from aquatic ecosystems to the atmosphere are insignificant. The principal source of gaseous NO x in 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 NO x emissions 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 (N 2 -> NH 4 ) 54 to 270 1-5
Electrochemical fixation lightning (N 2 NO ) 10 to 40 2, 6, 7, 8, 10 atmospheric 0 2 60 2) 14 to 20 2, 4 Biological denitrification (NO 3 N 2 ) 96 to 190 2, 3 (NO 3 N 2 0) 20 to 340 2-4 combined 83 to 270 5, 8, 9
Industrial denitrification (Organic -N a NO x ) 14 to 19 2-4 (Other a NO x ) 30 to 36 2, 3, 5 Atmospheric denitrification (NH 3 -> NO x ) 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

a 

### 1. 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).

3-103are 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 NO x 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 (NO x ) 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 x concentrations, relative NMOC and NO x concentrations (NMOC/NO x ratios), NMOC reactivity, and NO x composition are known to affect the photochemical reactions that produce ozone and other oxidants in ambient atmospheres. Concentration-based NMOC/NO x ratios are used in some precursor-oxidant models. Ratios of NMOC/NO x that are concentration-based are calculated directly, using measured concentrations of 

### NO x expressed as NO 2 (ppm or ppb) and measured concentrations of hydrocarbonsexpressed 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 x concentrations recorded at various urban and nonurban locations in the United States. In addition, HC/NO x (or NMOC/NO x ) 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 andurban 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.

3-1043.5.2.1.1 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 C 2 -C 10 molecular-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 C 4 -C 6 molecular-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)

a 

### Data 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 -0 5 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 C 2 - C 5 7 to 14 Sexton and Westberg (1984) Miami, FL C 2 - C 5 2 to 23 Sexton and Westberg (1984) Glascow, IL C 2 - C lo 60 to 150 Chatfield and Rasmussen (1977) Janesville, WI C2

- -C1p 9 to 24 Sexton and Westberg (1984) Houston, TX C 2 - C 10 2 to 24 Sexton and Westberg (1984) Robinson, IL C 2 - C lo 13 to 113 Sexton and Westberg (1984) 

### Smoky Mtns., TN C 2 - 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

3-107nonurban 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 x measurements that can be related to the diurnal photochemical processes that produce ozone. 3.5.2.2.1 Urban NO concentrations. Concentrations of NO x , 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 NO x is emitted as nitric oxide (NO), but the NO is converted rapidly to NO 2 by 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 2 can 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 2 and increased atmosphere mixing. Nitric oxide levels remain low during the daytime and then usuaily build up again through the nighttimehours. 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 x concentration in urban areas of the United States is about 70 ppb, with NO and NO 2 contributing about equally (Logan, 1983). Monitoring data for 1975 through 1980 showed that peak 1-hr NO 2 concentrations



### 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 NO x concentrations during the 6:00 to 9:00 a.m. period are of primary importance in terms of oxidant production. Average NO x levels 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 x concentrations 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 x ratios in each of these urban areas are included.

TABLE 3-16, AVERAGE 6:00 to 9:00 a.m. NO CONCENTRATIONS AND

HC/NO X RATIOS IN URBAN ARgAS
Average NO , Average City ppb x HC/NO x 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)

3-109Hydrocarbon concentrations (ppb C) exceeded the NO x levels 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 x ratios 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 x in "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 x concentrations as low as 0.015 have been recorded (Bollinger et al., 1982). Slightly higher NO x concentrations have been reported at other remote locations in the western United States and Canada. Kelly et al. (1982) deduced a mean NO x 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 NO x . Measurements of NO x during 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 2 at a rural si te i n Vi rgi ni a(Kel ly et al ., 1984). In the northeastern United States, nonurban NO x concentration.s appear to exceed those in the west by about a factor of ten. A median NO x concentration 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 x concentrations of 4.7 and 4.1 ppb, respectively (Decker et al., 1978). Nitric oxide composed approximately 40 percent of the total NO x at 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

3-110transport, 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 precursoremissions 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 containdetailed 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 existingcomputer 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 asknowledge of atmospheric chemistry grows and, for example, as more sophisticated approaches to dealing with boundary conditions become available. Evenunder 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.

3-1113.6.1 

### Definitions, 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. 3.6.1.1 

### Statistical 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 ofair 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.

3-112These and other weaknesses led EPA to abandon the use of statistical models in

state implementation plans (F.R., 1981). 3.6.1.2 

### Trajectory 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 NO

X 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 NO x levels 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 

# i REMOVAL
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 ~ I I I I

~ S..CHEMICAL ` ~ -' • ~ AD VECTI VE TRANSFORMATIONS Z(U) INFLOW

r

S OURC E

— EMISSIONS 

# ~ L

~ A VECTIVE Qx Ox OUTFLOW
(c) BOX MODEL Figure 3-17. Schematics of the three types of dynamic models.



### 3-114Agency, 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 asLevels 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 (NO x ); and as indirect functions of (1) NMHC and NO x emissions occurring later in the day; (2) specified meteorological conditions; (3) reactivity of the precursor 

### mix; and (4) concentrations of ozone and precursors transportedfrom 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).

3-115E



### 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 lowerright 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 NO x concentration will have different effects on ozone at different NMHC/NO x ratios.These two features of the diagram are related to the underlying photochemistry (see, e.g., Whitten, 1983)`. For extremely low NO x concentrations 

### (i.e., high NMHC/NO x ratios), where the lines parallel the abscissa, the formation of 0 3 is (1) insensitive to changes in NMHC concentrations, and (2) NO x -limited; that is, changes in NO x concentrations cause co-directi,onal changes in peak 0 3 concentrations. 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 2 rapidly 'and completely. In addition, the NO x present, before being removed from the cyclic reactions via termination reaction.s with various radicals, completes a number of N0 R Q 2
~ 

### -. NO 2 --- -d NO (+ 0 3 ) cycles, thus producing a number of 0 3 molecules. 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 NO x does not have an appreciable impact on the R02 radicals, but it do

- es change co-directionally the number of 0 3 molecules produced, since the photolysis of NO 2 and the oxidation of NO to NO x are essential for 0 3 production. For moderately higher concentrations of NO x (lower NMHC/NO x ratios), 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 0 3 . Thus, for moderate NMHC/NO x ratios, the effects on 0 3 formation of varying the concentrations of NMHC or NO x are simiiar in direction. Finally, for much higher NO x concentrations (i.e., very low NMHC/NO x ratios), the dominant effects are the relative depletion of radicals, through the reactions of NO 2 with 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 NO 2 is very slow. At such low NMHC/NO x ratios, therefore, increasing the NMHC concentration enhances 0 3 formation but increasing NO x concentrations inhibits it.
3-117As examination of Figure 3-18 reveals, for an NMHC concentration of 0.6

ppm C, for example, increasing NO x leads to increased 0 3 until NMHC/NO x ratios of about 5:1 to 6:1 are reached; further NO x increases, leading to lower NMHC/NO x ratios, inhibit 0 3 formation. Thus, in this example, there is a "critical" ratio (approximately 5.6:1) at which the NO x effect on 0 3 changes direction. Besides this "critical" ratio, an "equal controi" NMHC/NO x ratio also exists, above which the reduction of NO x is more beneficial in terms of 0 3 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 3 to 0.12 ppm. Thus, for this particular case (Figure 3-18), the chemical mechanism modeling evidence suggests that (1) NO x control will increase the peak downwind 0 3 concentration at NMHC/NO x ratios of 5.6:1 or lower; (2) both NO x control 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 x control is relatively more effective in reducing 0 3

. The calculation of precursor controls necessary to reduce 0 3 to 0.12 ppm, from the isopleths given in Figure 3-18, shows that NO x control, although at first beneficial, is ultimately detrimental because it makes the reduction of 0 3 to 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 3 concentration of 0.30 ppm and an NMHC/NO x ratio 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 0 3 to 0.12 ppm through the unilateral control of NMHC (line AB in Figure 3-18). If, however, NO x is 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 3 to 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 NO x , would ultimately be detrimental for the situation represented by these isopleths. There may be situations, however, in which the control of NO x will 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 NO x is not applicable to all situations. These diagrams may differ

3-118when 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 3 formation of controlling NO X emissions are a matter of continuing discussion and research. The inhibition of 0 3 formation by NO x under 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 NO x to the Detroit ambient air samples suppressed 0 3 formation. Kelly also compared EKMA predictions with the results of his smog chamber study and concluded that EKMA correctly describes the respective effects of NO x and 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

3-119models (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 ingrid-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 arecontinually 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).

3-1203.6.1.4 

### Box 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..

3.6.2 

### Validation 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„ a ascertain 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-NO x

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

3-121such 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 airdata. 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 andShreffler, 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 qualitydata, 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

;

3-122of 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, fourdifferent 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 trajectorymodel (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 x precursor 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).

3-123Studies 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 0 3

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 beimportant. 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) costsof 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 (0 3 ) and other photochemical oxidants occurring at low concentrations in ambient air, such as peroxyacetyl nitrate (PAN), hydrogen peroxide (H 2 O 2 ), 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.,

3-124oxidation). 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 oxygenthat 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 3 IH 2 0 (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 2 O 2 /H 2 0 (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 H 2 O 2 ) must be generated in situ for the calibration of ineasurement techniques. For ozone and H 2 O 2 , 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.

3-125The 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, whichdetermines 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 combinewith 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 2 0) is unimportant in the production of oxidants in ambient air because it is virtually inert in the troposphere.

3-1263.7.3 ' 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- andwelfare-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 2by 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 , NO 2 , 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/NO 2 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.



### 3-127The presence of organic compounds, oxides of nitrogen, and sunlight does

not mean that the photochemical reactions will continue indefinitely. Termination reactions gradually remove NO 2 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 (HN0 4 ), 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). Themechanisms 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 NO x , natural hydrocarbons (i.e., those organic compounds emitted from vegetation) can also undergo photooxidation reactions to yield 0 3 , 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). 3.7.3.2 

### Atmospheric 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 -21 cm 3 molecule -1 sec -1 can 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 -14 cm

3 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,



### 3-128a 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 NO 2 , and subsequently with NO 2to 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.- 3.7.3.3 

### Atmospheric Reactions of PAN,
, H„0„ and HCOOH . Because PAN is in equilibrium with acetyl peroxy radicals

- and NO 2 , 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)

3-129that 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 2 O 2 will be taken up in aqueous droplets. Over the past decade, evidence has accumulated that H 2 O 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 2 SO 4 (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 2 O 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.

3-130Incursions 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. 3.7.4.1 

### Atmospheric Mixing . The concentration of a pollutant in ambient air depends significantly on the degree of atmospheric mixing that occurs from thetime 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 increasingaltitude, 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 adiurnal 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.

3-131Summer 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 notrestrictive, 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).

3-1323.7.4.3 Effects of Sunlight and Temperature. The effects of sunlight on

photochemical oxidant formation, aside from the role of solar radiation inmeteorological 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 thattemperature 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 thenumerous studies on transport phenomena that were described in the 1978 criteria document for ozone and other photochemical oxidants (U.S. EnvironmentalProtection Agency, 1978a). In that document, transport phenomena were classified into three categories, depending upon transport distance: urban-scale,

3-133mesoscale, and synoptic-scale. In urban-scale transport, maximum concentrations of 0 3 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

3-134Gulf 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 isformed 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 notexclusively, 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.

3-135A 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 7 Be 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 7 Be/0 3 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 7 Be 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 ofsurface 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 haveconcluded 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,

3-136on 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 NO x emissions assumed. Studies on the role of NO x in 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 x concentrations at most remote, clean l.ocations range from <0.05 ppb upward. Mean concentrations of NO x at 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 x are higher than previously thought (see, e.g., Singh et al:, 1980; Kelly et al., 1984, regarding global modeis and assumed reservoirs of NO x ). The contributions of biogenic VOC to background ozone, although a matterof 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 (NO x ), emitted by manmade and by natural sources; and on suitable conditions of sunlight, temperature, and other meteorological factors. Because of theintervening 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

3-137United 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 highlyinferential 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 x in the United States were estimated at 19.4 Tg/yr for 1983. Retrospective estimates show that manmade NO x emissions rose fromabout 6.8 Tg/yr in 1940 to about 18.1 Tg/yr in 1970 (U.S. Environmental Protection Agency, 1986). Annual emissions of manmade NO x were 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 NO x emissions are based on uncertain extrapolations from very limited studies, but appear to be about an order of magnitude less than manmade NO x emissions. 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.'

3-138The hydrocarbon composition of urban atmospheres is dominated by species

in the C 2 to C 10

molecular-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). 3.7.5.2.2 

### Hydrocarbons 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. 3.7.5.2.3 

### Nitrogen 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 NO x is 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 NO x 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 NMHCINO x ratios 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, NO x concentrations as low as 0.015 ppb have been recorded (Bollinger et al., 1982). Concentrations of NO x at 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, NO x 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 ofprecursor 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 theatmospheric 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.

3-140The 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 x levels 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 x concentration ratios, reductions in NO x will decrease ozone formation; (2) at moderate NMOC/NO x ratios, reductions in NMOC and NO x will decrease ozone formation; and (3) at very low NMOC/NO x ratios, increases in NO x will 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 consultedfor 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). 3.7.6.2 

### Fixed-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-NO x 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.

3.8 REFERENCES

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United States. J. Air Pollut. Control Assoc. 25: 19-24. Altshuller, A. P. (1983) Review: natural volatile organic substances and their effect on air quality in the United States. Atmos. Environ. 17: 2131-2165. ,Aitshuller, A. P. (1986) Review paper: the role of nitrogen oxides in nonurban ozone formation in the planetary boundary layer over N America, W Europe and adjacent areas of ocean. Atmos. Environment. 20: 245-268.Altshuller, A. P.; Bufalini, J. J. (1971) Photochemical aspects of air polluti on: a revi ew. Envi ron. Sci . Technol . 5: 39-64.Aitshuller, A. P.; McPherson, S. P. (1963) Spectrophotometric analysis of aldehydes in the Los Angeles atmosphere. J. Air Pollut. Control Assoc. 13: 109-111. Ames, J.; Myers, T. C.; Reid, L. E.; Whitney, D. C.; Golding, S. H.; Hayes, S. R.; Reynolds, S. D. (1978) The user's manual for the SAI airshed model. San Raphael, CA: Systems Applications, Inc. No. EM78-89. Anbar, M.; Taube, H. (1954) Interaction of nitrous acid with hydrogen peroxide and with water. J. Am. Chem. Soc. 76: 6243-6247. Anderson, G. E.; Hayes, S. R.; Hillyer, M. J.; Killus, J. P.; Mundkur, P. V. (1977) .Air quality in the Denver metropolitan region: 1974-2000. Denver, C0: Region VIII, U.S. Environmental Protection Agency. EPA-908/1-77-002. Andrews, A. W.; Thibault, L. H.; Lijinsky, W. (1978) The relationship between mutagenicity and carcinogenicity of some nitrosamines. Mutat. Res. 51: 319-326. Appel, B. R.; Kothny, E. L.; Hoffer, E. M.; Wesolowski, J. J. (1980) Sulfate and nitrate data from the California Aerosol Characterization Exper.iment (ACHEX). In: Hidy, G.M.; Muelier, P. R.; Grosjean, D.; Appel, B. R.; Wesolowski, J. J., eds. The character and origins of smog aerosols: a digest of resuits from the California Aerosol Characterization Experiment(ACHEX). New York, NY: John Wiley and Sons; pp. 315-335. (Advances in environmental science and technology: v. 9). Arnts, R. R.; Gay, B. W., Jr. (1979) Photochemistry of some naturally 'emitted hydrocarbons. Research Triangle Park, NC: U.S. Environmental Protection Agency, Office of Research and Development; EPA report no. EPA-600/3-79-081. Available from: NTIS, Springfield, VA; PB80-131980. Arnts, R. R.; Meeks, S. A. (1981) Biogenic hydrocarbons: contribution to the ambient air in selected areas. Atmos. Environ. 15: 1643-1651. Arnts, R. R.; Seila, R. L.; Kuntz, R. L.; Mowry, F. L.; Knoerr, K. R.; Dudgeon, A. C. (1978) Measurement of a7pha-pinene fluxes from a loblolly pine forest. In: Proceedings of the 4th joint conference on sensing of environmental pollutants, American Chemical Society. Washington, DC: American Chemical Society:

3-143Arnts, R. R.; Gay, B. W.; Bufalini, J. J. (1981) Photochemical oxidant potential of the biogenic hydrocarbons. In: Bufalini, J. J.; Arnts, R. R., eds. Atmospheric biogenic hydrocarbons, v. II. Ann Arbor, MI: Ann Arbor Science Publishers; p. 117. Ashby, H. A.; Stahman, R. C.; Eccleston, B. H.; Hurn, R. W. (1974) Vehicle emissions--summer to winter. In: Society of Automotive Engineers meeting; October; Toronto, Canada. Warrendale, PA: Society of Automotive Engineers; SAE technical paper 741053. Association of Bay Area Governments. (1979a) Application of photochemical models, Vol. I: The use of photochemical models in urban ozone studies. Research Triangle Park, NC: U. S. Environmental Protection Agency; EPA report no. EPA-450/4-79-025. Association of Bay Area Governments. (1979b) Application of photochemical models, Vol. III: Recent sensitivity tests and other applications of the LIRAQ model. U. S. Environmental Protection Agency, Research Triangle Park, NC; EPA report no. EPA-450/4-79-027. Atkinson, R. (1985) Kinetics and mechanisms of the gas phase reactions of the hydroxyl radical with organic compounds under atmospheric conditions. Chem. Rev.: in press. Atkinson, R.; Aschmann, S. M. (1984) Rate constants for the reactions of 0 3 and OH radicals with a series of alkynes. Int. J. Chem. Kinet. 16: 259-268. Atkinson, R.; Carter, W. P. L. (1984) Kinetics and mechanisms of the reactions of ozone with organic compounds in the gas phase. Chem. Rev. 84: 437-470. Atkinson, R.; Lloyd, A. C. (1981) Smog chemistry and urban airshed modeling. In: Rodgers, M. A. J. ; Powers, E. L. , eds. Oxygen and oxy-radi cal s i n chemistry and biology. New York; Academic Press. pp. 559-592. Atkinson, R.; Lloyd, A. C. (1984) Evaluation of kinetic and mechanistic data for modeling of photochemical smog. J. Phys. Chem. Ref. Data 13: 315-444. Atkinson, R.; Darnall, K. R.; Pitts, J. N., Jr. (1978) Rate constants for the reactions of OH radicals and ozone with cresols at 300 ± 1 K. J. Phys. Chem. 82: 2759-2761. Atkinson, R.; Darnall, K. R.; Lloyd, A. C. ; Winer, A. M.; Pitts, J. N., Jr. (1979) Kinetics and mechanisms of the reactions of the hydroxyl radical with organic compounds in the gas phase. In: Pitts, J. N., Jr.; Hammond, G. S.; Gollnick, K.; Grosjean, D., eds. Advances in photochemistry, v. 11. New York, NY: John Wiley and Sons. pp. 375-488.
Atkinson, R.; Carter, W. P. L.; Darnall, K. R.; Winer, A. M.; Pitts, J. N., Jr. (1980) A smog chamber and modeling study of the gas phase NO -air photooxidation of toluene and the cresols. Int. J. Chem. Kinet. 12x 779-836.
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## A. M.; Pitts, J. N., Jr. (1981) Rate constants for the gas-phase reactions of 0 3 with a series of carbonyls at 296 K. Int. J. Chem. Kinet. 13: 1133-1142.

3-144Atkinson, R.; Aschmann, S. M.; Carter, W. P. L.; Winer, A. M.; Pitts, J. N.,

Jr. (1982a) Alkyl nitrate formation from the NO x -air photooxidations of C Z -C $ n-alkanes. J. Phys. Chem. 86: 4563-4569.
Atkinson, R.; Aschmann, S. M.; Fitz, D. R.; Winer, A. M.; Pitts, J. N., Jr. (1982b) Rate constants for the gas phase reactions of 0 3 with selected organics at 296 K. Int. J. Chem. Kinet. 14: 13-18. Atkinson, R.; Lloyd, A. C.; Winges, L. (1982c) An updated chemical mechanism for hydrocarbon/NO /SO photooxidation suitable for inclusion in atmospheric simulation models. x Atmos. Environ. 16: 1341-1355.
Atkinson, R. ; Aschmann, S. M. ; Carter, W. P. L. ; Winer, A. M. ; Pitts, J. N. , Jr. (1982d) Kinetics of the gas-phase reactions of OH radicals with alkyl nitrates at 299 ± 2 K. Int. J. Chem. Kinet. 14: 919-926. Atkinson, R.; Carter, W. P. L.; Winer, A. M. (1983) Effects of temperature and pressure on alkyl nitrate yields in the NO photooxidations of n-pentane,and n-heptane. J. Phys. Chem. 87: 20124018.
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Atkinson, R.; Carter, W. P. L.; Plum, C. N.; Winer, A. M.; Pitts, J. N., Jr. (1984d) Kinetics of the gas-phase reactions of NO 3 radicais with a series of aromatics at 296 ± 2 K. Int. J. Chem. Kinet. 16: 887-898.
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3-145Automotive News. (1982a) Sales of diesel-powered cars in the U.S. In: 1982

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### Black, F. M. (1977) The impact of emissions control technology on passenger

car hydrocarbon emisson rates and patterns. In: Dimitriades, B., ed. International conference on photochemical oxidant pollution and its control; September 1976;=Raleigh, NC. Proceedings: v. II. Research Triangle Park, NC: U.S. Environmental Protection Agency; pp. 1053-1068; EPA report no. EPA-600/3-77-00Ib. Available from: NTIS, Springfield, VA; PB-264233. Black, F. M.; Bradow, R. L. (1975) Patterns of hydrocarbon emissions from 1975 production cars. Presented at: Society of Automotive Engineers' national fuels and lubricants meeting; June; Houston, TX. Warrendale, PA: Society of Automotive Engineers; SAE paper no. 750681. Black, F. M.; High, L. E. (1979) Diesel hydrocarbon emissions, particulate and gas phase. Presented at: Society of Automotive Engineers' symposium on diesel particulate emissions measurement characterization; February; Detroit, MI. Warrendale, PA: Society of Automotive Engineers.Black, F. M.; High, L. E.; Lang, J. M. (1980) Composition of automobile evaporative and tailpipe hydrocarbon emissions. J. Air Pollut. Control Assoc. 30: 1216-1221. Blumenthal, D. L.; Richards, L. W.; Macias, E. S.; Bergstrom, R. W.; Wilson, W. E.; Bhardwaja, P. S. (1981) Effects of a coal-fired power plant and other sources on southwestern visibility (interim summary of EPA's project VISTTA). Atmos. Environ. 15: 1955-1969. Bollinger, M. J.; Parrish, D. D.; Hahn, C.; Albritton, D. L.; Fehsenfeld, F. C. (1982) NO measurements in clean continental air. In: Proceedings of2nd symposiug on composition of the nonurban troposphere; May; Williamsburg, VA. (see also, Bollinger, M., et al. (1984) NO measurements in clean continental air and analysis of the contributi&g meteorology. J. Geophys. Res. 89: 9623-9631). Box, E. 0. (1981) Foliar biomass: data base of the International BiologicalProgram and other sources. In: Bufalini, J. J.; Arnts, R. R., eds. Atmospheric biogenic hydrocarbons, v. 1. Ann Arbor, MI: Ann Arbor Science; pp. 107-119.Braddock, J. N.; Bradow, R. L. (1976) Emission patterns of diesel-powered passenger cars. Presented at meeting of Society of Automotive Engineers, Houston, TX, June 3-5, 1975. SAE Paper No. 750682.Braverman, T. N.; Layland, D. E. (1982) Application of urban emission inventories to photochemical modeling. Proceedings of Air Pollution Control Association Specialty Conference on Emission Inventories and Air Quality Management, Kansas Ci.ty, M0, April.

y 

### Breeding, R. J.; Lodge,:J'. P., Jr.; Pate, J. B.; Sheesley, D. C.; Klonis, H. B.; Fogle, B.; Anderson, J. A.; Englert, T. R.; Haagenson, P. L.; McBeth, R. B.; Morris, A. L.; Pogue, R.; Wartburg, A. F. (1973) Background trace gas concentrations in the central United States. JGR J. Geophys. Res. 78: 7057-7064.

3-147Brock, J. R.; Durham, J. L. (1984) Aqueous aerosol as a reactant i the atmosphere. In: Calvert, J. G., ed. S0 2 , NO and NO 2 oxidation mechanisms: atmospheric considerations. Boston, MA: Butterworth Publishers; pp. 209-250. (Teasley, J. I., ed. Acid precipitation series, v. 3).Bruckmann, P. W.; Willner, H. ('1983) Infrared spectroscopic study of peroxyacetyl nitrate (PAN) and its decomposition products. Environ. Sci. Technol. 17: 352-357. Bucon, H. W.; Macko, J. .F.; Taback, H. J. (1978) Volatile organic compound (VOC) species data manual. Prepared,by KVB Engineering, Inc., Tustin, CA. Research Triangle Park, NC: U.S. Environmental Protection Agency; EPA report no. EPA-450/3-78-119. Available from: NTIS, Springfield, VA; PB80-113822. Bufalini, J. J.; Arnts, R. R., eds. (1981) Atmospheric biogenic hydrocarbons, Vols. 1 and 2. Ann Arbor, MI; Ann Arbor Science, Bufalini, J. J.; Lonneman, W. A., eds. (1977) Proceedings of symposiu'm on 1975northeast oxidant transport study. Research Triangle Park, NC: U.S. Environmental Protection Agency. Report no.-EPA-600/3-77-017. Burns, R. C.; Hardy, R. W. F. (1975) Nitrogen fixation in bacteria and higher plants. New York: Springer-Verlag. California Bureau of Air Sanitation. (1966) The oxides of nitrogen in air pollution. Sacramento, CA: California Department of Public Health. Calvert, J. G.; Mohnen, V. (1983) The chemistry of acid formation. In: Acid deposition: atmospheric processes in eastern North America. Washington, DC: National Academy Press. Calvert, J. G.; Stockwell, W. R. (1983) Acid generation in the troposphere by gas-phase chemistry. Environ. Sci. Technol. 17: 428A-443A. Calvert, J. G.; Stockwell, W. R. (1984).The mechanism and rates of the gas-phase oxidation of sulfur dioxide and nitrogen oxides in the atmosphere. In: Calvert, J. G., ed., Acid precipitation: S0 2 , NO and 

### NO 2 oxidation mechanisms: atmospheric considerations., Boston, MA: Butterworth Publishers. (Teasley, J. I., ed., Acid precipitation series v. 3) Calvert, J. G.; Su, F.;:Bottenheim, J. W.; Strausz, 0. P. (1978) Mechanism of the homogeneous oxidation of sulfur dioxide in the troposphere. Atmos. Environ. 12: 197-226. Carter, W. P. L.; Darnall, K. R.; Lloyd, A. C.; Winer, A. M.; Pitts, J. N., Jr. (1976) Evidence for alkoxyl radical isomerization in photooxidation of C 4 -C6 alkanes under simulated atmospheric conditions. Chem. Phys. Lett. 42: 22-27. Carter, W. P. L.; Lloyd, A. C.; Sprung, J. L.; Pitts, J. N., Jr. (1979) Computermodeling of smog chamber data: progress in validation of detailed mechanism for the photooxidation of propene and 

### n -butane in photochemical smog. Int. J. Chem. Kinet. 11: 45-101.

3-148Carter, W. P. L.; Winer, A. M.; Darnall, K. R.; Pitts, J. N., Jr. (1979b) Smog

chamber studies of temperature effects in photochemical smog. Environ. Sci. Technol. 13: 1094-1100. Carter, W. P. L.; Winer, A. M.; Pitts, J. N., Jr. (1981a) Major atmospheric sink for phenol and the cresols: reaction with the nitrate radical. Environ. Sci. Technol. 15:.829-831. Carter, W. P. L.; Tuazon, E. C.; Winer, A. M.; Pitts, J. N., Jr. (1981b) Gas phase reactions of N,N-dimethylhydrazine with ozone and NO in simulated'atmospheres. In: Scanlan, R. A.; Tannenbaum, S. R., eds. &- nitroso compounds. Washington, DC: American Chemical Society; pp. 118-131. (American Chemical Society symposium series no. 174). Carter, W. P. L.; Winer, A. M.; Pitts, J. N., Jr. (1981c) Effect of peroxyacetyl nitrate on the initiation of photochemical smog. Environ. Sci. Technol. 15: 831-834. Carter, W. P. L.; Wjner, A. M.; Pitts, J. N., Jr. (1982) Effects of kinetic mechanisms and hydrocarbon composition on oxidant-precursor relationships predicted by the EKMA isopleth technique. Atmos. Environ. 16: 113-120. Cato, G. A.; Buening, H. J.; DeVivo, C. C.; Morton, B. G.; Robinson, J. M. (1974) Field testing: Application of combustion modifications to control pollutant emissions from industrial boilers--phase I. Research Triangle Park, NC: U.S. Environmental Protection Agency; EPA report no. EPA-650/2-74-078a. Available from: NTIS, Springfield, VA; PB-238920. Cato, G. A.; Muzio, L. J.; Shore, D. E. (1976) Field testing: Application of combustion modifications to control pollutant emissions from industrialboilers--phase II. Research Triangle Park, NC: U.S. Environmental Protection Agency; EPA report no. 600/2-76-086a.Chameides, W. L.; Davis, D. D. (1982) The free radical chemistry of cloud droplets and its impact upon the composition of rain. JGR J. Geophys. Res. 87: 4863-4877. Chameides, W. L.; Walker, J. C. G. (1976) A photochemical theory of tropospheric ozone. J. Geophys. Res. 78: 8751-8760.
Chameides, W. L:; Stedman, D. H.; Dickerson, R. R.; Rusch, D. W.; Cicerone, R. J. (1977) NO x production in lightning. J. Atmos. Sci. 34: 143-149. Chameides, W.L.; Tan, A. (1981). The two-demensional diagnostic modea for tropospheric OH: An uncertainty analysis. J. Geophys. Res. 86: 5209-5223. Cfiatfield, R. B.; Rasmussen, R. A. (1977) Hydrocarbons, fluorocarbons, and oxidants observed at a rural site near St. Louis. In: Proceedings from national AIChE meeting; March; Houston, TX. Chock, D. P.; Dunker, A. M.; Kumar, S.; Sloane, C. S. (1981) Effect of NO x emission rates on smog formation in the California South Coast Air Basin. Environ. Sci. Technol. 15: 933-939.

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of OH reactions with the atmospheric suifur compounds H 2 S, CH3SH, CH 3 SCH 3 and CH 3 SSCH 3 . 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.
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- 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.)

3-179Wolff, 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 O 2 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. Re search 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 x such 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 transportconditions 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 0 3 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 berequired 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 0 3 , hydrocarbons, and NO x , 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 0 3 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 0 3 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 froma 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 samplingstrategies. 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;meaningof 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 0 3 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 (NO 2 ). 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 inthe 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:

0 3 + 3I + H 2 0 -> I 3 + 0 2 + 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 0 3 (Chapter 5); and the other oxidants present, except NO 2 , 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 NO 2 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 NO 2 and correction of the corresponding total oxidant measurement. The interference from S0 2 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 S0 2 is simultaneously measured, the total oxidant may also be corrected for S0 2 . 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 H 2 S to S0 2 , and partial removal of 0 3 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 0 3 -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 0 3 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 0 3 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 0 3 is measured in the presence of water vapor, i.e., asignai enhancement of 3 to 12 percent in high humidity'as opposed to measurement of the same concentration of 0 3 in dry air (California Air ResourcesBoard, 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 0 3 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 0 3 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 0 3 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 molarabsorption coefficient at the mercury line has been measured by several investigators with good agreement and has an accepted value of 134 ± 2 M 1 cm -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 0 3 in fundamental chemical and physical studies. Some of the first atmospheric 0 3 measurements were, infact, 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 photometricinstruments 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 0 3 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 0 3 concentrations in the atmosphere are calibrated with standard 0 3 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 situ during 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 0 3 concentrations and chemical and physical means for absolute measurement of these concentrations. 4.2.4.1 

### Generation . 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 0 3 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.

0 2 + hy(x<200 nm) -> 20 (4

- 2)

0+ 0 2 + 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 0 3 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.

4.2.4.2 Calibration 

### 4.2.4.2.1 KI procedures: original EPA reference method



### . The output of photolytic 0 3 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 0 3 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 purporseof 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 photometricmeasurements. 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 0.96 a

a 

### Correction 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. 4.2.4.2.2 Ultraviolet 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 0 3 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 0 3 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 0 3 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 0 3 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 NO 2 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 measuringthe 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 0 3 concentration may then be calculated directly from the Beer-Lambert equation:

### _ 10 6 T 760 Io (4-4) 

### [03

~ 

### ppm aQ X ~ X ~ ln 

### T_

where a= 0 3 absorption coefficient at 254 nm, 1 atm, and 0°C,



### = 308 ± 4 atm l cm

-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 0 3 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 0 3 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 0 3 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 thatit 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 ofcommercial 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 0 3 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 0 3 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 2 O 2 ) 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 2 O 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 significantbias 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 10 4 M 1 cm

-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 reducedby 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 sincebeen 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 NO 2 (Rehme et al., 1974):

NO + 0 3 -> NO 2 + 0 2 (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 0 3 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 NO X (NO + NO 2 ) 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 collaborativestudy, 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 oxidantsas 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 0 3 . Instrumental methods for the specific measurement of atmospheric 0 3 became commercially available in 1970. These had several practical advantagesover total oxidant KI-based instruments. These advantages were greater sensitivity, precision, specificity-

- no interferences from ambient S0 2 and NO 2 --and improved reliability in routine monitoring. Second, the data available showed that 0 3 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 0 3 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 S0 2 . 'The presence of such species could cause the total oxidant measurements to be either higher or lower than 0 3 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 0 3 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 0 3

. 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- i b i [Ox] i + c[NO 2 ] (4-6)

- d[S0 2 ] - 7- i e i [Red] i
In this equation, [Ox] i and [Red] i represent the concentrations of otheroxidizing and reducing species in the atmosphere. The atmospheric concentrations of other reducing species, such as H 2 S, are normally quite low compared to S0 2 concentrations (Stevens et al., 1972b, and references therein) and these species will not be considered further here. If the concentrationsabove 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 3 concentration 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)

= [0 3 ] + 7i b' i [Ox]i + c'[NO 2 ] - d'[S0 2 ]

Other atmospheric oxidants that have been identified and that may contribute to the total oxidant reading are hydrogen peroxide (H 2 0 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). Anestimation 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 A a E
Peracetic acid A NA b 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 NO 2 ) D(10%) Peroxyacetyl nitrate (PAN) D N Peroxypropionyl nitrate (PPN) D D

a 

### A = 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 NO 2 and S0 2 interferences can be expressed "as below, assuming that no other significant reducing interferences are present:

[Total Ox] corr. = [Total Ox] - c[NO 2 ] + [S0 2 ] ( 4-8)

= [0 3 ] + F i b i [Ox] i

Thus, a total oxidant measurement for which legitimate corrections or compensations for NO 2 and S0 2 have been made should always be higher than a simultaneous 0 3

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 0 3 losses within sample inlet systems are insignificant. On the other hand, total oxidant data uncorrected for S0 2 and NO 2 interferences 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 3 measurements during evening and early morning hours, when 0 3 concentrations 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 ozonizedair stream that has been analyzed by a common refer.ence method. In a simultaneous comparison, daily calibrations should be made with an 0 3 generator 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 NO 2 and S0 2 should be made. If chromium trioxide scrubbers are used to remove S0 2 in the inlet to the oxidant instrument, these must be frequently tested to ensure that 0 3 is not also removed during continued use, particularly under very humid conditions. These scrubbers may cause some additional bias by oxidation of NO to NO 2

.

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 0 3 absorption wavelengths. The only meaningful measurements for consideration here are the KI oxidant and UV 0 3 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,


## x 1 = 265 nm, X 2 = 313 nm, and X 3 = 280 nm. Intensity ratios at the three wavelengths were used to minimize the effects of other UV absorbers and of particulate scattering. 'Some non-0 3 absorption may still have been present, and, if so, the measured values would be higher than the true 03 values. Thepublished 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 0 3 occurred at the same time, but the 0 3 maximum was usually less than the total oxidant maximum. The UV-0 3 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 0 3 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 calibrationbias. Since interferences may be present in the UV measurement and simultaneous measurements of NO 2 and S0 2 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).



### 4-26was 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 0 3 generator calibrated by the 1 percent NBKI method. Simultaneous NO 2 and S0 2measurementswere 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 3measurements. The instruments were not calibrated by the same procedure, however, nor were any corrections attempted for S0 2 and

4-27

### NO 2 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 0 3 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 oxidantsfor this area by making simultaneous measurements with a gas-phase chemiluminescence 0 3 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 S0 2 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 0 3 . 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 daysshows considerably more variation among the methods, with total oxidant measurements both higher and lower than 0 3 measurements. Intercomparisons of only UV photometric and chemiluminescence 0 3 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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wdd 'N011Vli1N30N00



## 4-30Mast 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 0 3 , 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 NO 2 and S0 2 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 0 3 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 0 3 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



### omolo 

# c

~ 

### 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),



### CH 3 0(0 2 )NO 2 E 

### CH 3 C0(0 2 )• + 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 obtainedpredominantly 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 0 3 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

### 4.2.6.2 

### Analytical 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 25 0 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 (Stephensand 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 urbanenvironments, 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

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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 beof 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 orlower 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 percentrelative 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 CH 3 CH0-C1 2 -NO 2 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 3 0NO 2 , 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 , HNO 3 , HCOOH, and HCHO have been made by long-path FTIR spectrometry during smog episodes in the Los Angeles Basin. Tuazon et al. (1978) havedescribed an FTIR system operable at pathlengths up to 2 km for ambient measurements of PAN and their trace constituents. This system employed an eight-mirrormultiple 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
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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 m

3 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 0 3 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 NO 2 . No'further work has been reported on the development of this technique, and there have been no atmospheric applications. 4.2.6.3 Generation 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 anddiscussed 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' + 0 2 + M RC-0 2 ' + M (4-13)

0 0 RC-0 2 ' + NO 2 

## = RC-0 2 -NO 2 (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 , NO 2 , 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 

## situ with 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 N 2 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 cellsand 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 andSalas (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 -

- 

### CH 3 COONO 2 + 20H

~ 

### CH3C0 + 0 2 + NO 2 + H 2 0 (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: NO 2 ) 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 thermaldecomposition of PAN in the presence of excess N0, and measurement by chemiluminescence of the NO consumed (Lonneman et al., 1982). The acetylperoxy radical, CH 3 C(0)0 21 and its decomposition products rapidly oxidize NO to NO 2 . 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 1 region. 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 2 NO 2 

## 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,.7 Methods for Sampling and Analysis of Hydrogen Peroxide 

### 4.2.7.1 Introduction 

### . Hydrogen peroxide (H 2 O 2 ) 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 2 O 2 is the recombination of the hydroperoxyl radical (Graedel et al., 1976):

H02 + H02 + M-> H 2 O 2 + 0 2 + M (4-16)

4-43Recent studies have implicated atmospheric H 2 O 2 in the aqueous-phase oxidation

of S0 2 to SO 4 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 in assessing the role of atmospheric H 2 O 2 has been a lack of adequate measurement methodology. Earlier measurements for atmospheric H 2 O 2 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 2 O 2 concentrations (0.01 to 0.18 ppm) are now believed to be far too high, primarily as a result of artifact H 2 O 2formation from reactions of absorbed 0 3 (Zika and Saltzman, 1982; Heikes et al., 1982; Heikes, 1984). Furthermore, maximum tropospheric 'H 2 O 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 0 3 interference, and a maximum H 2 O 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 H 2 O 2 are discussed below, as well as methods for sampling, generation, and standardization'of 'H 2 O 2 samples. 4.2.7.2 Sampling. Almost all of the methods used for the measurement of atmospheric H 2 O 2 have used aqueous traps for sampling. Midget impingers (Gayand 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 2 O 2from gas streams (Zika and Saltzman, 1982). Atmospheric 0 3 , however, which is also absorbed at concentrations much higher than H 2 O 2 , 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 2 O 2 production is relatively slow (Heikes, 1984), the removal of absorbed 0 3 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 2 O 2 in



### 4-44

### aqueous traps and also react with collected H 2 O 2 to reduce the apparent-H 2 O 2

concentration measured (Heikes et al., 1982). 4.2.7.3 Measurement. A number of inethods for measuring low levels of H 2 O 2 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. H 2 0 2 -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 appliedin 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 foratmospheric measurements (unpublished work in progress, Schiff, 1985). Iodometric techniques are useful only for calibration of H 2 0 2 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 H 2 0 2 (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


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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 S0 2 discussed above for aqueous traps. Das et ai. (1982) employed a static version of the method of Kok et al. (1978a) to measure H 2 O 2 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 H 2 O 2 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. 4.2.7.3.2 

### Enzyme-Catalyzed Methods (Peroxidase) . This general method involves three components: a substrate that is oxidizable; the enzyme, horseradish peroxidase (HRP); and H 2 O 2 . The production or decay of the fluorescence intensity of the substrate or reaction product is measured as it is oxidized by H 2 O 2 , 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 H 2 O 2 , 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 2 O 2 from a calibratibn curve. 'The most significant advantages of this method of analysis are the specificity for H 2 O 2 , the sensitivity, and the stoichiometry of the scopoletin:H 2 O 2 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 2 O 2 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 -11 M(Perschke and Broda, 1961) to 2 x 10 -10 M(Andreae, 1955). This technique has been applied to measurements of H 2 0 

# 2 in rainwater by Zika and Saltzman (1982).

### With the leuco crystal violet (LCV) substrate, intensely colored crystal

violet is formed from the reaction of H 2 0 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 10 5 M 1 cm 1 , a very high value and an inherent advantage of this method. The concentration of H 2 O 2 is a linear function of the concentration of crystal violet produced. The detection limit reported is about 10 -8 M H 2 O 2 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 H 2 0 2 . 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 H 2 O 2 . When HPPA reagent solution is mixed with HRP solution and a test solution containing H 2 O 2 , 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 2 O 2 concentration. The detection limit was reported to be 10 -10 mole H 2 O 2 with a linear range extending to 10 -8 mole H 2 O 2 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 H 2 0 2 concentration.The enzymatic methods appear to be the most promising aqueous, colorimetric methods for H 2 O 2 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 H 2 O 2 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 H 2 

# O2 in water may be generated by the irradiation of water with 60 00 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 2 0 2 is simply the serial dilution of commercial-grade=30 percent H 2 O 2 (Fisher Analytical Reagent). Samples prepared in this mannermust be standardized and the method usually employed is the iodometric technique discussed below. Stock standard solutions of H 2 O 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 2 O 2 , there is an obvious need for an H 2 O 2 generator comparable to the photolytic 0 3 .gener,atordiscussed in Section 4.2.4.1. A technique that has often been used for generating ppm concentration levels of H 2 O 2 in air has been described by Cohen and Purcell (1967). Microliter quantities of 30 percent H 2 O 2 solution are injected into a metered stream of air that flows into a Teflon bag.. The concentration of H 2 O 2 in the bag is then determined by the iodometric method discussed below. 4.2.7.4.2 

### Calibration . By far the most common method for standardizing low concentrations of H 2 O 2 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 10 4 by measuring 10 -5 to 10 -4 M H 2 O 2 solutions prepared from 0.2 M stock H 2 O 2 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 0 2

. 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 OXIDANTSDuring 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) asthe 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 nonmethaneorganic compounds, which was promulgated in.1971, invoives the gas chromatographic separation of inethane (CH 4 ) 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 CH 4 are 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 CH 4 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 (CO 2 ) 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 CO 2 and H 2 O. Since carbon monoxide (CO) will also oxidize to CO 2 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, moreextensive 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 N 2 cannot be used since it will also condense methane and oxygen). At this temperature all NMOCs are condensed into the trap. After the residual CH 4 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 foraromatic compounds, however, were less than expected. The researchers recommended additional testing and instrument refinement in order to resolve this problem. 4.3.1.1.2 

### Speciation 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 C 2 -C 3 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 broaderapplicability. 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 C 5 ) 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 retentiontimes 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 techniquesutilized 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 C 2 to C 5 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 2and 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 >C 2 organic gases, but it was impossible to completely desorb these specieswithout partially decomposing the XAD-4. Good collection and recovery efficiencies were provided by XE-340 only for organics of C 4 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 C 5 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 stepoccasionally 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 collectionstep. 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
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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 mostcommonly 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 otherhydrocarbons (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 factoris 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 

### situ concentrations of the individual compound of concern and include devices such as permeation and diffusion tubes and syringe delivery systems. 4.3.1.1.4 

### Comparison of non-speciation versus speciation methods . Speciationmethods 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 topretreated 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. 4.3.1.2 Aldehydes . 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 measurethe 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 r 2 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 NA a 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

a 

### Data 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. 4.3.1.2.1 Chromotropic 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 sulfuricacid (Altshuller et al., 1961; Olansky and Deming, 1976). Trapping formaldehyde in a 1 percent bisulfite solution before adding the CA solution increasedcollection 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 byEegriwe (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, NO 2 , 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 NO 2 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). vMiksch 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 

### 4-63employ 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. Theadvantages 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 2 NHC 6 H 3 (NO 2 )2 ~ H 2 

### 0 + RR'C=NNHC 6 H 3 (NO 2 ) 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 theextract 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) havereported 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 knownamounts 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, andGC-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 silylationprocedures 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). 4.3.2.1 

### Measurement 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, NO 2 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 NO 2 ) measurement to yield the NO 2 concentration. Typical commercial chemiluminescence 7nstruments can detect levels as low as 2.5 pg/m 3 (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 inpositive 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 HNO 3 on afternoons of high oxidant concentrations can exceed 30 percent of the NO 2 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 twomethods 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 NO 2 , 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 S0 2

. 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

~ 

### .:Na2S0 3 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 NO 2in 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,' 0 3 , HNO 3 , S0 2 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). Tuckeret 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 photoly

## sis of NO 2 (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 NO 2 . 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 0 3 to cause significant errors in obtaining NO x or 0 3 measurements. Conversely, when 0 3 is at a minimum, as in the early morning or possibly even in the late afternoon, NO (and NO 2 ) 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 preparedfrom 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 N 2 (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 NO 2 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 atconstant temperature, must also be accurately measured. An alternative procedur.e commonly employed for the calibration of N0, NO 2 , and NO x 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 + 0 3 -> NO 2 + 0 2 (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 4.4.1 Sampling 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 providequality 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. 4.4.1.1 

### Quality 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 prescribedin 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 0 3 monitoring stations, to help ensure the quality of aerometric data, have been established (U.S. Environmental Protection Agency, 1977a; National Research Council, 1977). 4.4.1.2 

### Measurement Methods for Total Oxidants and Ozone . Techniques for the continuous monitoring of total oxidants and 0 3 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,

0 3 + 3I - + H 2 0 a 1 3 + 0 2 + 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 atdifferent 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
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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, forevaluating 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 (Table4-13). 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 becalibrated 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, ozonegenerator, 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
4-74controlled 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 theseprior 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

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U C ~ O L U 4

~ - E

O O

~ O l7 t 0 .7 ' E tp O t--f 6--4 7 +) ro+ ) O O O L •r ri G. 'O ]G Y .G O t 10 bG L O.0 .c 4- o m m c r-+ L a i Q

. +• 1 1 >1 Ll tn O (n .0 Z Z - =D Y C. 1+) m 41 61 r• Q •r 4.1 •r i-) - uf •r C

~ r G. t .0 +' 4 ro ~ U •o ° LL ' a)

~ -- ° v cm
~ cz 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 0 3 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 adata 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 S0 2 interferences prevent the establishment of adefinite 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 NO 2 and S0 2 interfei^ences, and in such corrected totaloxidants measurements ozone is the dominant contributor. Indeed, it is diffi

### cul 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, onany 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 analysesare 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) permitthe quantitative analysis of PAN without calibration standards. The.absorptivity of the 990 cm 1 band of PBzN, a higher homologue of PAN, has been reported 

# V S'tephens (1969). Tuazon et al. (1978) describes an FTIR system operable atpathlengths 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 availableand 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 byinfrared 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 3 C(0)0 21 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 (H 2 O 2 ) 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 S0 2 to SO 4 -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 H 2 O 2 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 H 2 O 2 concentrations of 0.01 to 0.18 ppm are now believed to be far too high, and to be the result of artifact H 2 O 2 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 2 0 2 have used aqueous traps for sampling. Atmospheric 0 3 , however, which is also absorbed at concentrations much higher than H 2 O 2 , reacts in the bulk aqueous phase and at surfaces to produce H 2 O 2 and thus is a serious interference (Zika and Saltzman, 1982; Heikes et al., 1982; Heikes, 1984). The removal of°absorbed 0 3 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, S0 2 ) may interfere with the generation of H 2 O 2 in aqueous,traps and also react with collected H 2 O 2 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 2 O 2 , 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 H 2 0 2 (Armstrong and Humphreys, 1955; Kok et al., 1978a,b). Th7s technique as initially employed suffered the interferences from 0 3 and S0 2 discussed above for aqueous traps. Das et al. (1982) employed a static version of the.method of Kok et al. (1978a) to measure H 2 O 2 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 beenreported which used hemin, a blood component, as a catalyst for the luminolbased H 2 O 2 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 H 2 O 2 . The production or decay of the fluorescence intensity of the substrate or reaction product is measured as it is oxidized by H2O 2 , 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 H 2 0 2 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 H 2 O 2

. As with 0 3 , H 2 0 2 calibration standards are not commercially available andare usually prepared at the time of use. The most convenient method for preparing aqueous samples containing micromolar concentrations of H 2 O 2 is simply the serial di1ution of commercia1 grade 30 percent H 2 O 2 (Fisher Analytical Reagent). Techniques for the convenient generation of gas-phase standards are not available. A technique often used for generating ppm concentrations of H 2 O 2 in air involves the injection of microliter quantities of 30 percent H 2 0 2 solution into a metered stream of air that flows into a Teflon bag. Aqueousand 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-speciationanalyzers 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 etpart-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 ofambient 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 todirect emissions, it is also expected that photochemical reactions of hydrocarbons with oxides of nitrogen, 0 3 , 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 (N 2 0), 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 0 3 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 andBufalini, 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 NO 2 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 twomethods 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,andthe 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 N 2 ; the mi xture i s both accurate and stable (Hughes, 1975). The SRM for NO 2 is the NO 2 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 0 2 . 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.

4.5 REFERENCES

Adamson, P.; Guenthard, Hs. H. (1980) Reinvestigation of the matrix i.r. spectrum of peroxyacetylnitrate (PAN). Spectrochim. Acta Part A 36: 473-475.Allen, A. 0.; Hochanadel, C. J.; Ghormley, J. A.; Davis, T. W. (1952) Decomposition of water and aqueous solutions under mixed fast neutron and gamma-radiation. J. Phys. Chem. 56: 575-586.Altshuller, A. P. (1983) Measurements of the products of atmospheric photochemical reactions in laboratory studies and in ambient air--relationships between ozone and other products. Atmos. Environ. 17: 2383-2427.Altshuller, A. P.; Leng, L. J. (1963) Application of the 3-methyl-2-benzothiazolone hydrazone method for atmospheric analysis of aliphatic aldehydes. Anal. Chem. 35: 1541-1542. Altshuller, A. P.; McPherson, S. P. (1963) Spectrophotometric analysis of aldehydes in the Los Angeles atmosphere. J. Air Pollut. Control Assoc. 13: 109-111. Altshuller, A. P. ; Cohen, I. R. ; Meyer, M. E. ; Wartburg, A. F. , Jr. (1961) Analysis of aliphatic aldehydes in source effluents and in the atmosphere. Anal. Chim. Acta 25: 101-117. Andreae, W. A. (1955) A sensitive method for the estimation of hydrogen peroxide in biological materials. Nature (London) 175: 859-860.Appel, B. R. (1973) A new and more sensitive procedure for analysis of peroxybenzoyl nitrate. J. Air Pollut. Control Assoc. 23: 1042-1044. Armstrong, W. A.; Humphreys, W. G. (1965) A L.E.T. independent dosimeter based on the chemiluminescent determination of H 2 O 2 . Can. J. Chem. 43: 2576-2584. Ballard, L. F.; Tommerdahl, J. B.; Decker, C. E.; Royal, T. M.; Nifong, D. R. (1971a) Field evaluation of new air pollution monitoring systems: the LosAngeles Study: interim report. Research Triangle Park, NC: U.S. Environmental Protection Agency; report no. APTD-0775. Available from: NTIS, Springfield, VA; PB-204444. Ballard, L. F.; Tommerdahl, J. B.; Decker, C. E.; Royal, T. M.; Matus, L. K. (1971b) Field evaluation of new air pollution monitoring systems: St.Louis Study, phase 1. Contract no. CPA 70-101, U.S. Environmental Protection Agency, Research Triangle Park, NC. Available from: NTIS, Springfield, VA; PB-264232.

, •

### Bass; A. M.; Ledford, A. E., Jr.; Whittaker, J. K. (1977) Ultraviolet photometer

for ozone calibration. In: Dimitriades, B., ed. International conference on photochemical oxidant pollution and its control; September 1976; Raleigh, NC. Proceedings: v. I. Research Triangle Park, NC: U.S. Environ - mental Protection Agency; pp. 13-17; EPA report no. EPA-600/3-77-001a. Available from: NTIS, Springfield, VA; PB-264232. Beard, M. E.; Margeson, J. H.; Ellis, E. C. (1977) Evaluation of 1 percent neutral buffered potassium iodide procedure for calibration of ozone monitors. Research Triangle Park, NC: U.S. Environmental Protection Agency; EPA report no. EPA-600/4-77-005. Available from: NTIS, Springfield, VA; PB-267985. Bergshoeff, G. (1970) Ozone determination in air. In: Pittsburgh conference: analytical chemistry and applied spectroscopy; publication no. 342. Delft, The Netherlands: IG-TNO. Bersis, D.; Vassiliou, E. (1966) A chemiluminescence method for determining ozone. Analyst 91: 499-505. Bowman, L. D.; Horak, R. F. (1972) A continuous ultraviolet absorption ozone photometer. Anal. Instrum. 10: 103-108. Bowman, R. L.; Alexander, N. (1966) Ozone-induced chemiluminescence of organic compounds. Science (Washington, DC) 154: 1454-1456. Brewer, A. W.; Milford, J. R. (1960) The Oxford-Kew ozone sonde. Proc. R. Soc. London Ser. A 256: 470-495. Bricker, C. E.; Johnson, H. R. (1945) Spectrophotometric method for deter= mining formaldehyde. Ind. Eng. Chem. Anal. Ed. 17: 400-402.Bruckmann, P. W.; Willner, H. (1983) Infrared spectroscopic study of peroxyacetyl nitrate (PAN) and its decomposition products. Environ. Sci. Technol. 17: 352-357. Bryan, R. J.; Romanovsky, J. C. (1956) Instrumentation for air pollution , . Instrum. Autom. 29: 2432-2438. Bufalini, J. J.; Brubaker, K. L. (1969) Symposium on chemical reactions inurban atmospheres; October. Warren, MI: General Motors Research Laboratories. Bufal i ni , J. J. ; Gay, B. W. , Jr. ; Brubaker, K. L. (1972) Hydrogen peroxi de formation from formaldehyde photooxidation and its presence in urban atmospheres. Environ. Sci. Technol. 6: 816-821. Burton, C. S.; Bekowies, P. J.; Pollack, R. I.; Connell, P. (1976) Oxidant/ ozone ambient measurement methods: an assessment and evaluation. San Rafael, CA: Systems Applications, Inc.; EF76-111R. Butcher, S. S.; Ruff, R. E. (1971) Effect of inlet residence time on analysis of atmospheric nitrogen oxides and ozone. Anal. Chem. 43: 1890-1892.



# r ;

### California Air Resources Board. (1976) A study of the effect of atmospheric

humidity on analytical oxidant,measurement methods. Presented
, ':;at: 15th conference [on] methods in air pollution studies; January; Long:Beach, CA. Berkeley, CA: California Air- Resources Board,. Air and Industrial Hygiene Laboratory. California Air Resources Board. (1978) California air quality data:,
, JulyAugust-September. 10(13): 2-3. Available from: Technical Servi.de Div., California Air Resources Board, Sacramento, CA. Carroll, H.; Dimitriades, B.; Ray, R. M. (1972) A chemiluminescenceadetector for ozone measurement. Washington, DC: U.S. Department of the Interior, Bureau of Mines; report of investigations no. RI7650. C.F.R. (1971) 40:§50. Reference methods. C.F.R. (1975) 40:§53. Specifications for automated methods. Chameides, W.L.; Tan, A. (1981). The two-demensional diagnostic model for tropospheric OH: An uncertainty analysis. J. Geophys. Res. 86: 5209-5223. Cherniack, I.; Bryan, R.. J. (1965)_ A- comparison study of various types of ozone and oxidant detectors which are used for atmospheric air sampling. J. Air Pollut. Control Assoc. 15: 351-354. Chock, D. P.; Kumar, S.; Herrmann, R. W. (1982) An analysis of trends in oxidant air quality in the South Coast Air Basin of California. Atmos. Environ. 16: 2615-2624.Clark, T. A.; Baumgardner, R..E.; Stevens, R. K.; Krost, K.. J. (1974) Evaluation of new ozone monitor7ng instruments by measuring in nonurban atmospheres. In: Instrumentation for monitoring air quality: a symposium; August 1973; Boulder, C0. Philadelphia, PA: American.Society for Testing and Materials; pp. 101-7.11. Clements, J. B. (1975) Workshop on ozone measurement by the potassi.um iodide method summary. Research Triangle Park, NC: U.S. Environmental Protection Agency; EPA summary report no. EPA-650/4-75-007. Available from: NTIS, Springfield, VA; PB-240939. Clyne, M. A. A.; Coxon, J. A. (.1968) Kinetic studies of oxy-halogen radical systems. Proc. R. Soc. London Ser. A 303: 207-231.Cohen, I. R.; Purcell, T. C. (1967) Spectrophotometric determination of hydrogen peroxide with 8-quinolinol. Anal. Chem. 39: 131-132. Cohen, I. R.; Purcell, T. C.; Altshuller, A. P. (1967) Analysis of the oxidant in photooxidation reactions. Environ. Sci. Technol. 1: 247-252. Corse, E. W. (1981) Aldehyde ana7yses for the 1981 Atlanta study. EPA contract no. 68-02-2566. Research Triangle Park, NC: Northrop Services, Inc.- Environmental Sciences.

, ..

### Cox,:R:.'D.; Balfour, W. D.; Langley, G. J. (1982) Quality control for ambient

- 

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, 

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## w. ; Atkinson, R. ; Pitts, J. N. , Jr: (1984) Measurement of nitrate.radical concentrations in continental air. Environ. Sci. Technol.: in press. Poli, 

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## r

. .

### Rehme, K. A. ; Puzak, J. C. ; Beard, M. E. ; Smi th, C. F. ; Paur, R. J. (1981) Evaluation of ozone calibration procedures. Research Triangle Park, NC: U.S. Environmental Protection Agency, Environmental Monitoring Systems Laboratory; EPA final report no. EPA-600/4-80-050. Available from: NTIS, Springfield, VA; PB81-118911. Renzetti, N. A.; Romanovsky, J. C. (1956) Comparative study of oxidants in the Los Angeles atmosphere. J. Air Pollut. Control Assoc. 6: 154-159. Richter, H. G. (1983) Analysis of organic compound data gathered during 1980in northeast corridor cities. Research Triangle Park, NC: U.S..Environmental Protection Agency, Office of Air Quality Planning and Standards; EPA report no. EPA-450/4-83-017. Availabiz from: NTIS, Springfield, VA; PB84-116052.Ridley, B. A.; Howlett, L. C. (1974) An instrument for nitric oxide measurements in the stratosphere. Rev. Sci. Instrum. 45: 742-746. Riggin, R. M. (1983) Technical assistance document for sampling and analysis of toxic organic compounds in ambient air. Research Triangle Park, NC: U.S. Environmental Protection Agency, Office of Air Quality Planning and Standards; EPA report no. EPA-600/4-83-027. Available from: NTIS, Springfield, VA; PB83-239020. Salas, L. J.; Singh, H. B. (1982) Measurements of formaldehyde and acetaldehyde in the ambient air. Environ. Sci. Technol.: submitted. Salo, A. E.; Whitz, S.; MacPhee, R. D. (1975) Determination of solvent vapor concentrations by total combustion analysis: a comparison of infrared with flame ionization detectors. Presented at: the 68th annual meeting of the Air Pollution Control Association; June; Boston, MA. Pittsburgh, PA: Air Pollution Control Association; paper no. 75-33.2. Saltzman, B. E.; Wartburg, A. F., Jr. (1965) Absorption tube for removal of interfering sulfur dioxide in analysis of atmospheric oxidant. Anal. Chem. 37: 779-782. Sawicki, E.; Hauser, T. R.; Stanley, T. W.; Elbert, W. (1961). The 3-methyl-2- benzothiazolone hydrazone test: sensitive new methods for the detection, rapid estimation, and determination of aliphatic aldehydes. Anal. Chem. 33: 93-96.Scaringelli, F. P.; Rosenberg, P. E.; Rehme, K. A. (1970) Comparison of permeation devices and nitrite ion as standards for the colorimetric determination of nitrogen dioxide. Environ. Sci. Technol. 4: 924-929. Schiff, H. (1985). York University, Toronto, Canada. Personal communication: work in progress. Schiff, H. I.; Pepper, D.; Ridley, B. A. (1979) Tropospheric NO measurements up to 7 km. JGR J. Geophys. Res. 84: 7895-7897.Scott, W.E.; Stephens, E.R.; Hanst, P.L.; Doerr, R.C. (1957). Further developments in the chemistry of the atmosphere. Proc. Amer. Petrol. Inst. 37 (III): 171-183.
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### Singh, H. B.; Salas, L. J. (1983a) Peroxyacetyl nitrate in the free troposphere. Nature (London) 302: 326-328. 

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### Spicer, C. W.; Joseph, D. W.; Sticksel, R. R.; Ward, G. F. (1979) Ozone sources and transport in the northeastern United States. Environ. Sci. Technol. 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 pollution monitors by chemical and gravimetric techniques. In:''C-alibration in air monitoring. pp. 337-344. Philadelphia, PA: Amer. Soc. for Testing and Materiais. ASTM STP 598.Steinberg, M.; Dietz, R. N. (1969) Sr-90 ozone generator for sub-ppro'concentration range--instruction manual. Upton, Long Island, NY: Brookhaven National Laboratory, Department of Applied Science. Stephens, E. R. (1964) Absorptivities for infrared determination of peroxyacyl 

### nitrates. Anal. Chem. 36: 928-929. Stephens, E. R. (1967) The formation of molecular oxygen by alkaline hydrolysis of peroxyacetyl nitrate. Atmos. Environ. 1: 19-20. Stephens, E. R. (1969) The formation, reactions, and properties of peroxyacyl nitrates (PANs) in photochemical air pollution. In: Pitts, J. N., Jr.; Metcalf, R. L., eds. Advances in environmental sciences v. 1. New York, NY: Wiley-Interscience; pp. 119-146. Stephens, E. R.; Price, M. A. (1973) Analysis of an important air pollutant: peroxyacetyl nitrate. J. Chem. Educ. 50: 351-354. Stephens, E. R.; Hanst, P. L.; Doerr, R. C.; Scott, W. E. (1956a) Reactions of nitrogen dioxide and organic compounds in air. Ind. Eng. Chem. 48: 1498-1504. 

### Stephens, E. R. ; Scott, W. E. ; Hanst, P. L. ; Doerr, R. C. (1956b) Recent deve1opments in the study of the organic chemistry of the atmosphere.Presented at: 21st midyear meeting of the meeting of the American Petroleum Institute's Division of Refining; May; Montreal, PQ, Canada. Proc. 

### Am. Pet. Inst. 36 (III): 288-297.Stephens, E. R.; Darley, E. F.; Taylor, 0. C.; Scott, W. E. (1961) Photochemical reaction products in air pollution. Int. J. Air Water Pollut. 4: 

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### 4-102

### Sweeney,, M. P. ; Swartz, D. J. ; Rost, G. A. ; MacPhee, R. ; Chao, J. (1964)

Continuous measurements of oxides of nitrogen in auto exhaust. J. Air :Pollut. Control Assoc. 14: 249-254. Tanner, R. L.; Meng, Z. (1984) Seasonal variations in ambient atmospheric leveis of formaldehyde and acetaldehyde. Environ. Sci. Technol. 18: 723-726. Taube, H.; Bray, W. C. (1940) Chain reactions in aqueous solutions containing ozone, hydrogen peroxide and acid. J. Am. Chem. Soc. 62: 3357-3373. Tokiwa, Y.; Twiss, S.; de Vera, E. R.; Mueller, P. K. (1972) Atmospheric ozone determination by amperometry and colorimetry. In: Mamantov, G.; Shults, W. D., eds. Determination of air quality: proceedings of the ACS symposium on determination of air quality; April 1971; Los Angeles, CA. New York, NY: Plenum Press; pp. 109-130.Toyama, Y.; Kobayashi, J. (1966) Notes on an ozone generator and its calibrator. Pap. Meteorol. Geophys. 17: 65-75.
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## in: National Academy of Sciences. (1981) Formaldehyde and Other Aldehydes. Washington, DC: National Academy Press.
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### U.S. Environmental Protection Agency. (1975) Technical assistance document for

the chemiluminescence measurement of nitrogen dioxide. Researdh

~ 

### .Triangle Park, NC: U.S. Environmental Protection Agency; EPA report no. EPA-600/ 4-75-003. Available from: NTIS, Springfield, VA. U.S. Environmental Protection Agency, (1976a) Ambient air monitoring references and equivalent methods. Code of Federal Regulations, Title 40.;.:Part 53. U.S. Environmental Protection Agency, (1976b) Nitrogen dioxide measurement principle and calibration procedure: Notice of final rulemaking. Fed. Regist. 41: 52686-52694, December 1. 

### U.S. Environmental Protection Agency., (1977a) Air monitoring strategy

~ 

### :for state implementation plans. Research Triangle Park, NC: Office

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### U.S. Environmental Protection Agency. (1978b) Measurement of volatile organic compounds. Research Triangle Park, NC: U.S. Environmental Protection Agency; EPA report no. EPA-450/2-78-041. Available from: NTIS, Springfield, VA. U.S. Environmental Protection Agency. (1981) Technical assistance document for the calibration and operation of automated ambient non-methane organic compound analyzers. Research Triangle Park, NC: U.S. Environmental Protection Agency; EPA report no. EPA-600/4-81-015: Available from: NTIS, Springfield, VA; PB82-147406. U.S. Environmental Protection Agency: (1982) Air quality criteria for oxides of nitrogen. Research Triangle Park, NC: Environmental Criteria and Assessment Office; EPA report no. EPA-600/8-82-026. Available from: NTIS, Springfield, VA; PB83-131011. 

### Van de Wiel , H. J. ; Rei jnders, H. F. R. ; Sei fert, R. ; Gi es, H. ; Lanti ng, R.W.; Rudolf, W. (1979) International comparative programme on ozone reference methods. Atmos. Environ. 13: 555-557.

0 

### Vigroux, E. (1952) Spectrophotometry absorption of ozone at 18 ° C below 3130A. Compt. Rend: 2529-2530. Warren, G. J.; Babcock, G. (1970) Portable ethylene chemiluminescence ozone monitor. Rev. Sci.`Instrum. 41: 280-282.

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### Watanab:e, I.; Stephens, E. R. (1978) Reexamination of moisture anomaly in

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4-1055. CONCENTRATIONS OF OZONE AND

- OTHER PHOTOCHEMICAL OXIDANTS IN AMBIENT AIR

5.1 INTRODUCTION The data presented in this chapter on the concentrations of ozone andother photochemical oxidants in ambient air are intended to support and complement information presented in subsequent chapters on the effects of thesecompounds. 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 baseexists. Data for peroxyacetyl nitrate (PAN) and its homologues and for hydrogen peroxide (H 2 O 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.

5-1The 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 H 2 O 2 with sulfur dioxide (S0 2 ).

7. Physical scavenging in the atmosphere of ozone and other oxidants; e.g., the temperature-dependent decomposition of PAN, the precipitation scavenging of H 2 O 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 beenobserved 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

5-2night 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 relativelyhomogeneous 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. )

5-3

### 0.18

0.17

0.16 oA (27 Stations)

T / / / / / / / / / /

0.15 

### E rs a

~ O 

### 4
~ r Z w V Z O U 

### w 

### z O C 0.14

0.13

0.12 0 NAMS STATIONS (62) X 

# I 95a/o CONFIDENCE 1 .1111TERVALS O ALL STATIONS (176) . T 95°!o CONFIDENCE i INTERVALS p CALIFORNIA STATIONS (27)
0 ALL STATIONS EXCEPT CALIFORNIA (149) - 

# e`

` ~

%~ ~



### 1975 1976 1977 1978 1979 1980 1981 1982 1983

YEAR 

### 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).

5-4Ambient 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 Figure5-1, 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 dataheavily 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 through1982 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 methodwas 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. Increasedprecursor 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 x emissions 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 0 3 firbm 1982 to 1983 nationwide (U.S. Environmental Protection Agency, 1984a).

5-6The 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

5-70.;

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v 

## 2 O U W 0.1 Z O N O Q

0.
EPA REGIONa

NO. OF SITES I If III IV V Vi Vil Vill iX X

6 19 26 14 40 15 8 7 36 5

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.re 0.12 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 percentof 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

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### TABLE 5-1. SECOND-HIGHEST 1-hr OZONE CONCENTRATIONSa' b REPORTED 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 0.16 c 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 0.18 c 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 0.12 c 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 ' b REPORTED

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 0.04 c ND d 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.11 c 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.IO c 0.11 Tulsa, OK 0.15 0.13 0.13 Akron, OH 0.24 e 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 ' b REPORTED

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 0.12 c 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

a 

### The 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. b 

### These 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. d 

### ND = no data.

e 

### Questionable data; third-highest value for 1981 was 0.125 ppm. Source: U.S. Environmental Protection Agency (1984a).

5-15median 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 appearsto 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 monitoringnetworks 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 inareas with relatively infrequent urban influences. Additional data on specific rural areas are presented in Sections 5.5.1 and 5.5.2.

5.4.1 

### National 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

5-16Figure 5-6. Locations of the eight national forest (NF) stations constituting the National Air Pollution Background Network (NAPBN). Source: Evans et al. (1983).



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### 5-181979 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, NorthCarolina (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

5-20MINNEAPOLIS •

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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 

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5-24ozone 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).

5.4.2 

### Sulfate 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 NO x patterns 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 orits 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, andexceptions 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.

5-25MN MI 

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### Figure 5-9. Location of Sulfate Regional Experiment (SURE) monitoring stations. Source: Martinez and Singh (1979).

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### 5-275.5.1 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 amongthe 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) atground 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 on2 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



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Figure 5-1 O. Diurnal pattern of 1-hr ozone concentrations on July 13, 1979, Philadelphia, PA. Source: SAROAD (1985b).



### 5-290.12

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Figure 5-11. Diurnal patterns of 1-hr ozone concentrations, September 20 and 21, 1980, Detroit, M I. Source: SAROAD (1985c).



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### Figure 5-12. Diurnal and 1-rnonth composite diurnal variations in ozone concentrations, . Washington, DC, July 1981. Source: SAROAD (1985d).

0.12

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### Figure 5-13. Diurnal and 1-month composite diurnal variations in ozone concentrations, St. Louis County, MO, September 1981. Source: SAROAD (1985d).



### 5-310.12

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### 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 diurnaldata) 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~ 

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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

5-33may 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 0 3 . To describe the quantitative relationships among the exposure statistics at sites in the two categories, a Pearson linear correlation matrix was calculated. For siteswith 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. Thecorrelation 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).

5-34No 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

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### 5-36the 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; whereasin 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 ofa 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.

5-37J H ¢ p W tJ) LU {-• H

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### 5-415.5.1.2 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 andidentified 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.

5-420.10

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### 5-43Composite 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 1981are 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. 

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0 

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### 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).

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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).



### 5-450.35

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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.

5-460.35 0.:
WASHINGTON, DC 

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## 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).

5-47levels 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 tobe 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 themost 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 NO x -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), Levittand 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=48attributed 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 inthe 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 ConcentrationsOzone 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. Numerousfactors 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

5-49data 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 SUREnetwork 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) ofthat 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 such a 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).

5-50New Haven, Connecticut, was the site of an epidemiological study in 1976

by Zagraniski et al. (1979). Symptoms recorded in diaries kept by studyparticipants 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, theyprobably 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).

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5-53SITES 1. Susan Wagner High School 2. Mabel Dean High School 3. Woolsey Post Office (Astoria) 4. Mamaroneck 4

5. Public School 321 

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### Northeast Corridor Monitoring Program (NECRMP). Source: Smith (1981).



### 5-54yet 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 sameday (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, theactual 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 ofactual 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

5-55significantly 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

5-56TABLE 5-13. SUMMARY OF REPORTED INDOOR-OUTDOOR OZONE RATIOS

Indoor-outdoor Structure ratio (I/0) Reference
Residence 0.60 a 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)

a Measured as total oxidants.

5-57to 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 reportedfor 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 thebuilding 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

5-58remains 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; Muelleret 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 turnoverrates 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

5-59roadways 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. 5.5.2.4 

### Altitudinal and Latitudinal Variations . Concentrations of ozone vary with altitude and with latitude. These variations occur for a number ofreasons, 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 andgeographic 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. 5-60Latitudinal 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 40percent lower concentrations at 2 km (MSL) in summer, and (3) similar concentrations in autumn and winter. Altitudinal profiles of ozone concentrations

5-61for 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 seasonaldistributions 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. 53 0 N to 59 0 N). 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 andRoy, 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 datacorroborate 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.

5-62Increases 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 ozonegradients 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 LosAngeles 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.

5-6370 Q O 50 `- 

### H 2 ~L~ 30 'x

10 0 

### 2745 90

I RIM FOREST IROCK

CAMP I i I MUD FLAT

~ DEEP I CITY CREEK CREEK I I —N—►

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## m m 

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5-64Total,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 andtemperature 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.

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## 5-67The 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 ozonemaxima 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 900m(2970 ft) would sustain exposures to relatively constant diurnal concentrations that would approximate the maximum ozone concentration found at thevalley 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. 5.5.2.5 

### Vertical 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 WrightCounty, 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

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~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 criteriadocument 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 5.6.1 

### Introduction 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 ofthe 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.



### 5-71Given 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 NO x precursors 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,

5-72i.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, PANconcentrations 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

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~ 3 a ~ +t 3 3 ~ W+) a L io •- O o o ~ o o v m x ++ + ~ vt at a o c ~ o~ ~ NU J t ~ J wIZ Z S to .0 U N 5-76over 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 inambient 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

5-770.07
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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/0 n 

### , % 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 NA a 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)

a llot available. Source: Adapted from Altshuller (1983).

5-79review 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 assampling 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 AreasData 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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# 5-82

mconcentration 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

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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 maximumoxidant 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 isborne 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

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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).

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### 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).



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### 5-885.6.5.2 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 PANwere 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 NO x 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 Nitrate5.6.6.1 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

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### 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 0 3concentration, 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+.

. . . . , ... .

. .
a These 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 inphotochemically 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.

5-910.080

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### 5-g25.6.6.2 

### Intracity 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: 1 00 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 148percent, 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 gradientsfor 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 oxidantsare presented in a recent criteria document on oxides of nitrogen (U.S. Environmental Protection Agency, 1982b) and are not given here. In addition, the

5-93TABLE 5-22. INTRACITY VARIATIONS IN PEAK OZONE AND PAN CONCENTRATIONS IN HOUSTON, OCTOBER 26 AND 27, 1977

0 3
, 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 2 O 2 ). The reader isreferred to these reviews for information on these oxidants. The two nonnitrogenous oxidants, formic acid and hydrogen peroxide, are appropriateconcerns 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.

5-94Data 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 (H 2 O 2 ) 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 H 2 O 2 (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 O 2 measurements. Laboratory experiments (Heikes et al., 1982) have indicated that 1 ppb H 2 O 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 H 2 O 2 levels exceed about 5 to'10 ppb; and Altshuller (1983) concluded that variations in H 2 O 2measurements there remain unexplained. Because of ineasurement probiems, the true levels of H 2 O 2 in ambient air are unknown, especially in polluted areas, where multiple interferences maypossibly occur. Attempts to detect hydrogen peroxide by means of FTIR spectroscopy have all been negative, even in polluted areas. The method can measure H 2 O 2 in ambient air with specificity at H 2 O 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 2 O 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

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### 5-96TABLE 5-23. CONCENTRATIONS a OF 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 NA b < 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)

a llote: 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. b Not available.

c See Chapter 4 for method used by Heikes et al. (1982). Source: Derived from data in Altshuller (1983).

5-97to 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 AreasIn 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 withpopulation, 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 0 3 ; 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 0 3 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

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### >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
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### >2 0.18

Minneapolis-St. Pau1,-MN/WI 

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### <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 0 3 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 0.17Anaheim-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

a Standard 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 possiblehourly 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

5-100from specified concentrations are location-specific. At a site in Dallas,

Texas, for example, daily maximum 1-hour concentrations were > 0.06 

### ppm for2 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 someurban 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, aswell 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 newcalibration 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~ . 0.16

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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).



### 5-102For 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 innationwide 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 (SUREstudy, Martinez and Singh, 1979). For example, maximum 1-hour ozone concentrations measured in 1980 at Kisatchie National Forest (NF), Louisiana; Custer

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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 0.035ppm. 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 fromurban 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 ConcentrationsSince 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

— — ~ 5-105photochemical 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 occurredearlier 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.

5-106Percentile 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 fromimportant 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

5-107on 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 meteorologicalfactors. 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 abovethe 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.

5-109The 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 theexistence 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 smalldifferences 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

5-1108 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 at817 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 hoursat >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

5-111open 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. 5.8.5.3 

### Other 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 concentrationsare 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.

5.8.6 

### Concentrations and Patterns of Other Photochemical Oxidants 5.8.6.1 

### Concentrations . 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 2 O 2 ). 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, PANand PPN concentrations are lower than even the lowest of the urban concentrations by a factor of 3 to 4. The concentrati ons of H 2 O 2 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 2 O 2 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 2 O 2 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 2 O 2 (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, H 2 O 2 , and HCOOH is important only if these non-ozone oxidants are

5-114shown to produce potentially adverse health or welfare effects,

~ singly, in

combination with each other, or in various combinations with ozone at concentra

### ti 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

5-115published 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, H 2 O 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 thathis conclusion is applicable to the non-ozone oxidants examined in this document.

5.9 REFERENCES

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5-116Berk, J. V.; Young, R. A.; Brown, S. R.; Hollowell, C. D. (1981) Impact of

energy-conserving retrofits on indoor air quality in residential housing.Presented at: 74th annual meeting of the Air Pollution Control Assocciation; June; Philadelphia, PA. Pittsburgh, PA: Air Pollution Control Association; paper no. 81-22.1 Berry, C. R. (1964) Differences in concentrations of surface oxidant between valley and mountaintop conditions in the southern Appalachians. J. Air. Pollut. Control Assoc. 14:238-239. Blumenthal, D. L.; White, W. H.; Peace, R. L.; Smith T. B. (1974) Determination of the feasibility of the long-range transport of ozone or ozone precursors. EPA-450/3-74-061, U.S. Environmental Protection Agency, Research Triangle Park, NC.
, Brennan, E. (1980) PAN concentrations in ambient air in New Brunswick, NJ.Final report. Prepared under grant 805827 for U.S. Environmental Protection Agency, Research Triangle Park, NC. Bufalini, J. J.; Gay, B. W., Jr.; Brubaker, K. L. (1972) Hydrogen peroxide formation from formaldehyde photooxidation and its presence in urban atmospheres. Environ. Sci. Technol. 6: 816-821. Cherniak, E. A.; Corkum, R. (1981) Continuous monitoring of PAN in the polluted troposphere. Brock University, St. Catherines, ON, Canada. Cited in Temple, P. J.; Taylor, O.C. (1983).Cleveland, W. S.; Graedel, T. E.; Kleiner, B. (1976a) Photochemical air pollution: transport from the New York City area into Connecticut and Massachusetts. Science (Washington, DC) 191: 179-181. Cleveland, W. S.; Guarino, R.; Kleiner, B.; McRae, J. E.; Warner, J. L. (1976b) The analysis of the ozone problem in the northeast United States. In: Specialty Conference on: Ozone/oxidants--interactions with the total environment. Pittsburgh, PA: Air Pollution Control Association; pp. 109-120. Coffey, P. E.; Stasiuk, W. N. (1975a) Evidence of atmospheric transport of ozone into urban areas. Environ. Sci. Technol. 9: 59-62. Coffey, P. E.; Stasiuk, W. N. (1975b) Urban ozone: its local and extraregional components. Presented at the 79th annual national meeting, American Institute of Chemical Engineers; March; Houston, TX. Paper no. 60C. Coffey, P. E. ; Stasi uk, W. N. ; Mohnen, V: (1977) Ozone i n rural and urbanareas of New York State.. In: Dimitriades, B., ed. International conference on photochemical oxidant pollution and its control. Proceedings: v. I. Research Triangle Park, NC: U.S. Environmental Protection Agency, Office of Research and Development; pp. 89-96; EPA report no. EPA-600/ 3-77-001a. Available from: NTIS, Springfield, VA; PB-624232.

5-117Contant, C. F.; Gehan, B. M.; Stock, T. H.; Holguin, A. H.; Buffler, P. A.

(1985) Estimation of individual ozone exposures using microenvironment measurements. In: APCA Specialty Conference: Scientific Basis for the Ozone Standard, Houston, TX, November 28-December 1,.1984. Pittsburgh; APCA. In Press. Cox, R. A.; Roffey, M. J. (1977) Thermal decomposition of peroxyacetyl nitrate' in the presence of nitric oxide. Environ. Sci. Technol. 11: 900-906. Darley, E. F.; Kettner, K. A.; Stephens, E. R. (1963) Analysis of peroxyacyl nitrates by gas chromatography with electron capture detection. Anal. Chem. 35: 589-591. Davies, T. D.; Ramer, B.; Kaspyzok, G.; Delany, A. C. (1984) Indoor/outdoor ozone concentrations at a contemporary art gallery. J. Air Pollut. Control Assoc. 31: 135-137. Edinger, J. G. (1973) Vertical distribution of photochemical smog in Los Angeles basin. Environ. Sci. Technol. 7: 247-252. Edinger, J. G.; McCutchan, M. H.; Miller, P. R.; Ryan, B. C.; Schroeder, M. J.; Behar, J. V. (1972) Penetration and duration of oxidant air pollution in the south coast air basin of California. J. Air Pollut. Control Assoc. 22: 882-886. Elkus, B.; Wilson, K. R. (1976) Photochemical oxidant: weekday-weekend differences. Draft report. Cited in U. S. Environmental Protection Agency (1978). Evans, G. F. (1985) The National Air Pollution Background Network: final project report. Research Triangle Park, NC: U. S. Environmental Protection Agency, Office of Research and Development; EPA report no. EPA-600/4-85-038. Available from: NTIS, Springfield, VA; PB85-212413. Evans, G.; Finkelstein, P.; Martin, B.; Possiel, N.; Graves, M. (1982) The National Air Pollution Background Network, 1976-1980. Research Triangle Park, NC: U.S. Environmental Protection Agency; EPA report no. EPA-600/ 4-82-058. Available from: NTIS, Springfield, VA; PB83-100412. Evans, G.; Finkelstein, P.; Martin, B.; Possiel, N.; Graves, M. (1983). Ozone measurements from a network of remote sites. J. Air Pollut. Control Assoc. 33: 291-296.Farmer, J. C.; Dawson, G. A. (1982) Condensation sampling of soluble atmospheric trace gases. JGR J. Geophys. Res. 87: 8931-8942: Galbally, I. E.; Roy, C. R. (1980) Destruction of ozone at the earth'-s surface. Quart. J. R. Met. Soc. 106: 599-620. Graedel, T. E.; Farrow, L. A.; Weber, T. A. (1977) Photochemistry of the "Sunday effect." Environ. Sci. Technol. 11: 690-694. Grosjean, D. (1981) Critical evaluation and comparison of ineasurement methods for nitrogenous compounds in the atmosphere. Final report, A 706-05, for Coordinating Research Council. Environmental Research and Technology. Westlake Village, CA.

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