May 20 11 i DRAFT

– DO NOT CITE OR QUOTE 

# Chapter 3 Contents

Chapter 3. Ambient Lead: Source to Concentration

________________________________ _____________________________ 3

- 1

3.1. Introduction

________________________________ ________________________________ _______________ 3

- 1

3.2. Sources of Atmospheric Lead

________________________________ ________________________________ _ 3

- 1

3.2.1. National Emissions Inventory

________________________________ _________________________ 3

- 2

Figure 3

- 1. Trends in Pb emissions (thousand tons) from stationary and mobile sources
in the U.S., 1970

- 2008.

________________________________ ______________ 3

- 3

Figure 3

- 2. Trends in Pb emissions (thousand tons) from stationary and mobile sources
in the U.S., 1990

- 2008.

________________________________ ______________ 3

- 3

Figure 3

- 3. Nationwide stationary source Pb emissions (tons/yr) in the U.S. by source

sector in 2008.

________________________________ _____________________ 3

- 4

Figure 3

- 4. County

- level Pb emissions (tons) from stationary sources in the U.S. in

2008.

________________________________ ____________________________ 3

- 5
Figure 3

- 5. Pb emissions (tons) at airports in the U.S. in 2008. The size of the symbol indicates the magnitude of Pb emissions at each airport.

____________________ 3

- 6

3.2.2. Anthropogenic Sources

________________________________ ______________________________ 3

- 6

3.2.2.1. Lead Emissions from Piston Engine Aircraft Operating on Leaded Aviation Gasoline
and Other Non

- Road Sources

________________________________ __________________ 3

- 6

3.2.2.2. Fugitive Emissions from Metals Processing and Mining

_____________________________ 3

- 7

3.2.2.3. Fossil Fuel Combustion

________________________________ ______________________ 3

- 8

3.2.2.4. Other Industrial Sources

________________________________ ______________________ 3

- 9

3.2.2.5. Roadway

- Related Sources

________________________________ ___________________ 3

- 10

3.2.2.6. Deposited Lead

________________________________ ____________________________ 3

- 11

Figure 3

- 6. Total U.S. Pb additives in on

- road gasoline used in on

- road vehicles, 1927

- 1995.

________________________________ ___________________________ 3

- 13

Figure 3

- 7. Estimated Pb aerosol inputs from on

- road gasoline into 90 U.S. urbanized areas (UAs), from 1950 through 1982.

________________________________ _ 3

- 15

3.2.3. Source Attribution

________________________________ ________________________________ _ 3

- 15

3.2.3.1. Lead Speciation and Source Apportionment

________________________________ _____ 3

- 15

Table 3

- 1. Pb compounds observed in the environment

_____________________________ 3

- 16

3.2.3.2. Lead Isotop e Ratio Analysis

________________________________ __________________ 3

- 19

3.3. Fate and Transport of Lead

________________________________ ________________________________ __ 3

- 20

Figure 3

- 8. Fate of atmospheric lead.

________________________________ ___________ 3

- 21

3.3.1. Air

________________________________ ________________________________ _____________ 3

- 21

3.3.1.1. Transport

________________________________ ________________________________ 3

- 22

Figure 3

- 9. Scales of turbulence within an urban environment.

________________________ 3

- 24

3.3.1.2. Deposition

________________________________ _______________________________ 3

- 25

3.3.1.3. Resuspension of Lead from Soil to Air after Lead Deposition

________________________ 3

- 27

3.3.2. Water

________________________________ ________________________________ ___________ 3

- 28

3.3.2.1. Lead Transport in Water and Sediment

________________________________ _________ 3

- 29

3.3.2.2. Deposition of Lead within Bodies of Water and in Sediment

_________________________ 3

- 30

Table 3

- 2. Surface sediment Pb concentration s for various continental shelves; see Fang et al. (2009) and references therein.

______________________________ 3

- 31

3.3.2.3. Flux of Lead from Sediments

________________________________ _________________ 3

- 32

3.3.2.4. Lead in Runoff

________________________________ ____________________________ 3

- 33

3.3.3. Soil

________________________________ ________________________________ _____________ 3

- 40

3.3.3.1. Deposition of Lead onto Soil from Air

________________________________ __________ 3

- 41

3.3.3.2. Sequestration of Lead from Water to Soil

________________________________ _______ 3

- 43

3.3.3.3. Movement of Lead within the Soil Column

________________________________ ______ 3

- 45May 20 11 ii DRAFT

– DO NOT CITE OR QUOTE Figure 3

- 10. Schematic model summarizing the estimated flux of Pb within a typical podzol profile from northern Sweden using data from Klaminder et al. (2006).

________________________________ __________________________ 3

- 47

Figure 3

- 11. Eh

- pH diagram for Pb in shooting range soils, Jefferson National Forest,

VA.

________________________________ ____________________________ 3

- 51

3.4. Monitoring of Ambient Lead

________________________________ ________________________________ _ 3

- 52

3.4.1. Ambient Measurement Techniques

________________________________ ____________________ 3

- 52

3.4.1.1. Size

- Selective PM Monitoring for Lead Concentrations

____________________________ 3

- 52

3.4.1.2. Federal Reference Method and Federal Equivalenc e Method Evaluation

_______________ 3

- 53

Table 3

- 3. Specifications for Pb monitoring

________________________________ ______ 3

- 54

3.4.1.3. Chemical Speciation Network, IMPROVE, and National Air Toxics Trends Network
Monitors

________________________________ ________________________________ _ 3

- 55

3.4.1.4. Other Measurement Methods for Total Lead

________________________________ _____ 3

- 56

3.4.1.5. Seq uential Extraction

________________________________ _______________________ 3

- 57

3.4.1.6. Speciation Techniques

________________________________ ______________________ 3

- 58

3.4.1.7. Continuous Lead Monitoring

________________________________ _________________ 3

- 60

3.4.2. Ambient Network Design

________________________________ ____________________________ 3

- 60

3.4.2.1. Monitor Siting Requirements

________________________________ _________________ 3

- 62

3.4.2.2. Spatial and Temporal Coverage

________________________________ _______________ 3

- 63

Figure 3

- 12. Pb monitoring sites for SLAMS, CSN, NATTS, and IMPROVE networks,

2007

- 2009.

________________________________ ______________________ 3

- 63

3.5. Ambient Air Lead Concentrations

________________________________ _____________________________ 3

- 64

3.5.1. Spatial Distribution of A ir Lead

________________________________ _______________________ 3

- 64

3.5.1.1. Variability across the U.S.

________________________________ ___________________ 3

- 64

Table 3

- 4. Summary data for source oriented Pb monitors across the U.S.

______________ 3

- 65

Table 3

- 5. Summary data for sites at which sourc e oriented statistics are at a maxima

_____ 3

- 65

Figure 3

- 13. Highest county

- level source oriented Pb

- TSP concentrations (μg/m 3 ),

maximum 3

- month average, 2007

- 2009.

________________________________ 3

- 66

Figure 3

- 14. Highest county

- level Pb

- PM 10 concentrations (μg/m 3 ), maximum 3

- month average, 2007

- 20 09.

________________________________ _______________ 3

- 67

Figure 3

- 15. Highest county

- level Pb

- PM 2.5 concentrations (μg/m 3 ), maximum 3

- month average, 2007

- 2009.

________________________________ _______________ 3

- 68

3.5.1.2. Intra

- urban Variability

________________________________ ______________________ 3

- 68

3.5.2. Temporal Variability

________________________________ _______________________________ 3

- 70

3. 5.2.1. Multi

- year Trends

________________________________ __________________________ 3

- 70

Figure 3

- 16. National trends in Pb concentration (μg/m 3 ), all FRM monitors, 1980

- 2009.

____ 3

- 71

Figure 3

- 17. National trends in Pb concentration (μg/m 3 ), source oriented FRM
monitors, 1990

- 2009.

________________________________ ______________ 3

- 72

Figure 3

- 18. National trends in Pb concentration (μg/m 3 ), non

- source oriented FRM
monitors, 1990

- 2009.

________________________________ ______________ 3

- 72

3.5.2.2. Seasonal Variations

________________________________ ________________________ 3

- 73

Figure 3

- 19. Monthly source oriented Pb

- TSP average (μg/m 3 ) over 12 months of the year, 2007

- 2009.

________________________________ __________________ 3

- 74

Figure 3

- 20. Monthly non

- source oriented lead

- TS P average (μg/m 3 ) over 12 months of the year, 2007

- 2009.

________________________________ _______________ 3

- 75
Figure 3

- 21. Monthly lead

- PM 10 average (μg/m 3 ) over 12 months of the y ear, 2007

- 2009.

____ 3

- 75
Figure 3

- 22. Monthly lead

- PM 2.5 average (μg/m 3 ) over 12 months of the year, 2007

- 2009.

____ 3

- 76

3.5.3. Size Distribution of Lead

- Bearing PM

________________________________ __________________ 3

- 77

3.5.3.1. AQS Data Analysis

________________________________ ________________________ 3

- 77

Table 3

- 6. Summary of comparison data for co

- located lead

- TSP and lead

- PM 10 monitors.

________________________________ ________________________ 3

- 77

Table 3

- 7. Summary of comparison data for co

- located Pb

- TSP and Pb

- PM 2.5 monitors

____ 3

- 78

Table 3

- 8. Summary of comparison data for co

- located Pb

- PM 10 and Pb

- PM 2.5 monitors.

________________________________ ________________________ 3

- 78

3.5.3.2. Size Distribution Studies in the Literature

________________________________ _______ 3

- 79May 20 11 iii DRAFT

– DO NOT CITE OR QUOTE Figure 3

- 23. Comparison of urban background and near

- road size fractions of lead

bearing PM.

________________________________ _____________________ 3

- 81

3.5.4. Lead Concentrations in a Multipollutant Context

________________________________ __________ 3

- 82

Figure 3

- 24. Correlations of monitored Pb

- TSP concentration with copollutant
concentrations, 2007

- 2008.

________________________________ __________ 3

- 83

Figure 3

- 25. Correlations of monitored Pb

- TSP concentration with co

- pollutant
concentrations, 2009.

________________________________ ______________ 3

- 84

Figure 3

- 26. Correlations of monitored lead

- PM 2.5 concentration with copollutant
concentrations, 2007

- 2009.

________________________________ __________ 3

- 86

3.6. Ambient Lead Concentrations in Non

- Air Media and Biota

________________________________ _________ 3

- 87

3.6.1. Soils

________________________________ ________________________________ ____________ 3

- 88

Table 3

- 9. Outdoor soil Pb levels in various cities within the U.S.

_____________________ 3

- 88

Figure 3

- 27. Map of median Pb content in soil in New Orleans.

________________________ 3

- 90

3.6.2. Sediments

________________________________ ________________________________ ________ 3

- 92

Table 3

- 10. Sediment concentrations in vario us cities, prior to 2005

____________________ 3

- 93

Figure 3

- 28. Sediment core data (1992

- 1994) for the lakes and reservoirs along the Apalachicola, Chattahoochee, and Flint River Basin (ACF), which feeds from north of the Atlanta, GA metropolitan area into the Gulf of Mexico at Apalachicola Bay in the Florida panhandle.

_____________________________ 3

- 94

Figure 3

- 29. Sediment core data (1975

- 1995) for the lakes and reservoirs along the Apalachicola, Chattahoochee, and Flint River Basin (ACF), which feeds from north of the Atlanta, GA metropolitan area into th e Gulf of Mexico at Apalachicola Bay in the Florida panhandle.

_____________________________ 3

- 95

3.6.3. Rain

________________________________ ________________________________ ____________ 3

- 96

Figure 3

- 30. Trends in Pb concentration in precipitation from various sites in Norway over the period 1980

- 2005.

________________________________ __________ 3

- 97

3.6.4. Snowpack

________________________________ ________________________________ ________ 3

- 97

3.6.5. Natural Waters

________________________________ ________________________________ ____ 3

- 99

3.6.6. M oss

________________________________ ________________________________ ___________ 3

- 100

3.6.7. Grass, Foliage, and Tree Rings

________________________________ _______________________ 3

- 100

Figure 3

- 31. Trends in regional pollution near a copper smelter in Canada and Pb concentrations at the boundary of heartwood trees within roughly 75 km of

the smelter.

________________________________ _____________________ 3

- 101

3.6.8. Aquatic Bivalves

________________________________ ________________________________ _ 3

- 101

3.7. S ummary

________________________________ ________________________________ _______________ 3

- 102

3.7.1. Sources of Atmospheric Lead

________________________________ _______________________ 3

- 102

3.7.2. Fate and Transport of Lead

________________________________ _________________________ 3

- 102

3.7.3. Ambient Lead Monitoring

________________________________ __________________________ 3

- 104

3.7.4. Ambient Air Lead Concentrations

________________________________ ____________________ 3

- 104

3.7.5. Ambient Lead Concentrations in Non

- Air Media and Biota

________________________________ 3

- 105
Chapter 3 References

________________________________ ________________________________ ____________ 3

- 107

Chapter 3 Appendix

________________________________ ________________________________ ________________ 3

- 129

3.8. Variability across the U.S.

________________________________ ________________________________ __ 3

- 129

Table 3A

- 1

. Distribution of 1

- month average Pb

- TSP concentrations ( μ g/m 3 ) nationwide,

source

- oriented monitors, 2007

- 9.

________________________________ ___ 3

- 129

Table 3A

- 2. Distribution of 3

- month moving average Pb

- TSP concentrations (μg/m3)
nationwide, source

- oriented monitors, 2007

- 9. Sites listed in the bottom six

ro ws of the table fall in the upper 90th percentile of the data pooled by site.

___ 3

- 130

Table 3A

- 3

. Distribution of annual 1

- month site maxima TSP Pb concentrations ( μ g/m 3 ) nationwide, source

- oriented monitors, 2007

- 2009.

_______________________ 3

- 131

Table 3A

- 4

. Distribution of annual 3

- mon th site maxima Pb

- TSP concentrations ( μ g/m 3 )
nationwide, source

- oriented monitors, 2007

- 2009.

_______________________ 3

- 131

Table 3A

- 5

. One

- month average Pb

- TSP for individual county concentrations nationwide ( μg/m 3 ), non

- source

- oriented monitors, 2007

- 2009

_____________ 3

- 132May 20 11 iv DRAFT

– DO NOT CITE OR QUOTE Table 3A

- 6

. Three

- month moving average Pb

- TSP for individual county concentrations ( μg/m 3 ) nationwide, non

- source

- oriented monitors, 2007

- 2009

_____________ 3

- 133

3.8.1. Intra

- urban Variability

________________________________ _____________________________ 3

- 133

Figure 3A

- 1. Pb TSP monitor and source locations within Los Angeles County, CA (06

- 037), 2007

- 2009.

________________________________ _________________ 3

- 137

Figure 3A

- 2

. Wind roses for Los Angeles County, CA, from meteorological data at the Los Angeles International Airport, 1961

- 1990.

__________________________ 3

- 138

Figure 3A

- 3

. Box plots of annual and seasonal Pb TSP concentrations ( μg/m 3 ) from

source

- oriented and non

- source

- oriented monitors within Los Angeles County, CA (06

- 037), 2007

- 2009.

________________________________ ____ 3

- 139

Table 3A

- 7

. Correlations between Pb TSP concentrations from source

- oriented and non

source

- oriented monitors within Los Angele s County, CA (06

- 037), 2007

- 2009

________________________________ __________________________ 3

- 140

Figure 3A

- 4

. Pb TSP monitor locations within Hillsborough and Pinellas Counties, FL (12

- 057 and 12

- 103), 2007

- 2009.

________________________________ ____ 3

- 142

Figure 3A

- 5

. Wind roses for Hillsborough/Pinellas Counties, FL, obtained from meteorological data at Tampa International Airport, 1961

- 1990.

____________ 3

- 143

Figure 3A

- 6

. Box plots of annual and seasonal Pb TSP concentrations ( μg/m 3 ) from

sourc e

- oriented and non

- source

- oriented monitors within Hillsborough and
Pinellas Counties, FL (12

- 057 and 12

- 103), 2007

- 2009.

__________________ 3

- 144

Tabl e 3A

- 8

. Correlations between Pb TSP concentrations from source

- oriented and non

source

- oriented monitors within Hillsborough and Pinellas Counties, FL (12

- 057 and 12

- 103), 2007

- 2009.

________________________________ ____ 3

- 145

Figure 3A

- 7

. Pb TSP Monitor locations within Cook County, IL (17

- 031), 2007

- 2009.

______ 3

- 148

Figure 3A

- 8

. Wind roses for Cook County, IL, obtained from meteorological data at O’Hare International Airport, 1961

- 1990.

_____________________________ 3

- 149

Figure 3A

- 9

. Box plots of annual and seasonal Pb TSP concentrations ( μg/m 3 ) from

source

- oriented and non

- source

- oriented monitors within Cook County, IL (17

- 031), 2007

- 2009.

________________________________ ______________ 3

- 150

Table 3A

- 9

. Correlations between Pb TSP concentrations from source

- oriented and non

source

- oriented monitors within Cook County, IL (17

- 031), 2007

- 2009.

______ 3

- 151

Figure 3A

- 10

. Pb TSP Monitor locations within Jefferson County, MO (29

- 099), 2007

- 2009.

________________________________ __________________________ 3

- 153

Figure 3A

- 11

. Wind roses for Jefferson County, MO, obtained from meteorological

data at St. Louis/Lambert International Airport, 1961

- 1990.

_______________ 3

- 154

Figure 3A

- 12

. Box plots of annual and seasonal Pb TSP concentrations ( μg/m 3 ) from

source

- oriented and non

- source

- oriented monitors within Jefferson County, MO (29

- 099), 2007

- 2009.

________________________________ __________ 3

- 1 55

Table 3A

- 1 0

. Correlations between Pb TSP concent rations from source

- oriented and non

source

- oriented monitors within Jefferson County, MO (29

- 099), 2007

- 2009

__ 3

- 156

Figure 3A

- 13

. Pb TSP Monitor locations within Cuyahoga County, OH (39

- 035),

2007

- 2009.

________________________________ _____________________ 3

- 158

Figure 3A

- 14

. Wind roses for Cuyahoga County, OH, obtained from meteorological data at Cleveland/Hopkins International Airport, 1961

- 90.

________________ 3

- 159

Figure 3A

- 15

. Box plots of annual and seasonal Pb TSP concentrations ( μg/m 3 ) from

source

- oriented and non

- source

- oriented monitors within Cuyahoga County, OH (39

- 035), 2007

- 2009.

________________________________ ____ 3

- 160

Table 3A

- 1 1

. Correlations between Pb TSP concentrations from source

- oriented and non

source

- oriented monitors within Cuyahoga County, OH (39

- 035), 2007

- 2009

________________________________ __________________________ 3

- 161

Figure 3A

- 16

. Pb TSP Monitor locations within Sullivan County, TN (47

- 163), 2007

- 2009.

________________________________ __________________________ 3

- 163

Figure 3A

- 17

. Wind roses for Sullivan County, TN, obtained from meteorological data at Bristol/Tri City Airport, 1961

- 90.

________________________________ __ 3

- 164

Figure 3A

- 18

. Box plots of annual and seasonal Pb TSP concentrations ( μg/m 3 ) from

source

- oriented and non

- source

- oriented monitors within Sullivan County, TN (47

- 16 3), 2007

- 2009.

________________________________ ___________ 3

- 165

Table 3A

- 1 2

. Correlations between Pb TSP concentrations from source

- oriented and non

source

- oriented monitors within Sullivan County, TN (47

- 163), 2007

- 2009

____ 3

- 166May 20 11 v DRAFT

– DO NOT CITE OR QUOTE 3.8.2. Size Distribution of Pb

- Bearing PM

________________________________ ___________________ 3

- 166

Table 3A

- 1 3

. Correlations and average of the concentration ratios for co

- located
monitors, TSP versus PM 10

, TSP versus PM 2.5 , and PM 10 versus PM 2.5

.

_______ 3

- 166

3.8.3. Lead Concentration in a Multipollutant Context

________________________________ _________ 3

- 169

Figure 3A

- 19a. Seasonal correlations of monitored Pb

- TSP concentration with copollutant concentrations, 2007

- 2008. Top: winter; bottom: spring.

________ 3

- 169

Figure 3A

- 19b. Seasonal correlations of monitored Pb

- TSP concentration with copollutant concentrations, 2007

- 2008. Top: summer; bottom: fall.

__________ 3

- 170

Figure 3A

- 20a. Seasonal correlations of monitored Pb

- TSP concentration with copollutant concentrations, 2009. Top: winter; bot tom: spring.

_____________ 3

- 171

Figure 3A

- 20b. Seasonal correlations of monitored Pb

- TSP concentration with copollutant concentrations, 2009. Top: summer; bottom: fall.

______________ 3

- 172

Figure 3A

- 21a. Seasonal correlations of monitored Pb

- PM 2.5 concentration with copollutant concentr ations, 2007

- 2009. Left: winter; right: spring.

__________ 3

- 173

Figure 3A

- 21b. Seasonal correlations of monitored Pb

- PM 2.5 concentration with copollutant concentrations, 2007

- 2009. Left: summer; right: fall.

___________ 3

- 174

Table 3A

- 14. Copollutant exposures for various trace metal studies

____________________ 3

- 175May 20 11 3

- 1 DRAFT

– DO NOT CITE OR QUOTE 

# Chapter 3. Ambien t Lead: Source to Concentration

3.1. Introduction

This chapter reviews concepts and findings in atmospheric sciences that provide a foundation for 

## 1

the detailed presentation of evidence of Pb exposure and Pb

- related health and ecological effects in 

## 2

subsequent chapters. Information in this chapter builds on previous Pb AQCD s using new data and 

## 3

studies. This includes new knowledge of Pb fate and transport, the latest developments in monitoring 

## 4

methodologies, and recent data describing Pb concentrations as a function of size range

. Description of 

## 5

the chemical forms of Pb is not provided here, however, because this information is well established. The 

## 6

reader is referred to the 2006 Pb AQCD for a description of the chemical forms of Pb ( U.S. EPA, 2006 )

. 

## 7

Section 3.2 provides an overview of the primary a nd secondary sources of air Pb. Section 3.3 

## 8

provides a description of the fate and transport of Pb in air, soil, and aqueous media. Descriptions of Pb 

## 9

measurement methods, monitor siting requirements, and monitor locations are presented in Section 3.4. 

## 10

Amb ient Pb concentrations, their spatial and temporal variability, size distributions of Pb

- bearing 

## 11

particulate matter (PM), and associations with copollutants are characterized in Section 3.5. 

## 12

Concentrations of Pb in non

- air media and biota are described in Section 3.6. 

## 13



# 3.2. Sources of Atmospheric Lead

The following section reviews updated National Emissions Inventory (NEI) data from 2008 ( U.S. 

## 14

EPA, 2011 ) , which is the most recently available quality

- assured Pb emissions data. This section also 

## 15

reviews updated information from the peer

- reviewed literature regarding sources of ambient Pb. Detailed 

## 16

information about processes for primary and secondary anthropogenic emissions a nd naturally

- occurring 

## 17

emissions can be found in the 2006 Pb AQCD ( U.S. EPA, 2006 ) . The papers cited herein generally 

## 18

utilized PM sampling data, because ambient airborne Pb readily condenses to PM. The 2006 Pb AQCD 

## 19

( U.S. EPA, 2006 ) employed the 2002 NEI ( U.S. EPA, 2008a ) or source analysis and listed the largest 

## 20

sources to be (in order): industrial

- commercial

- institutional boilers and process heaters (17%), coal 

## 21

utilities boilers (12%), mobile sources (10%), iron and steel foundries (8%), and miscellaneous sources 

## 22

Note: Hyperlinks to the reference citations throughout this document will take you to the NCEA HERO database (Health and Environmental Research Online) at http://epa.gov/hero . HERO is a database of scientific literature used by U.S. EPA in the process of developing science
assessments such as the Integrated Science Assessments (ISA) and the Integrated Risk Information System (IRIS)

.May 20 11 3

- 2 DRAFT

– DO NOT CITE OR QUOTE 

### from in dustrial processes, incineration, and utilities, each contributing less than 5% (53%). Since that 

## 1

time, states have additional information to the inventory. The mobile source category included 

## 2

combustion products from organic Pb antiknock additives used in piston

- engine aircraft (hereafter 

## 3

referred to piston aircraft emissions). 

## 4



# 3.2.1. National Emissions Inventory

Emissions of Pb have dropped substantially over the past forty years, as shown in Figure 3

- 1 and 

## 5

Figure 3

- 2. The reduction before 1990 is largely due to the phase

- out of Pb as an anti

- knock agent in 

## 6

gasoline for on

- road automobiles, as discussed in the 2006 Pb AQCD ( U.S. EPA, 2006 ) . This action 

## 7

resulted in a 98% reduction in Pb emissions from 1970

- 1990. Total Pb emissions over the period 1990

- 

## 8

2008 decreased an additional 7 7 % , from 5,200 tons in 1990 to 1,200 tons in 2008 . Subsequent emissions 

## 9

reductions are related to enhanced control of the metals processing industry. In 1990, metals processing 

## 10

accounted for 42% (2,200 tons) of total Pb emissions. By 2008, metals processing accounted for 12% 

## 11

(150 tons) of total emissions. This represented more than an order of magnitude decrease in Pb emissions 

## 12

from metals processing. At the same time, emissions from piston engine aircraft varied only slightly over 

## 13

this time period. In 1990, off

- highway Pb emissions were 990 tons and represented 19% of total Pb 

## 14

emissions. In 2008, off

- highway Pb emissions from piston engine aircraft were slightly lower at 590 tons, 

## 15

which comprised 49% of all Pb emissions. “ Miscellaneous ” emis sions from other industrial processes, 

## 16

solvent utilization, agriculture, and construction comprised 20% of emissions (240 tons) in 2008. 

## 17May 20 11 3

- 3 DRAFT

– DO NOT CITE OR QUOTE Source: U. S. EPA ( 2011 ) 

### Figure 3

- 1. Trends in Pb emissions (thousand tons) from stationary and mobile sources in the U.S., 1970

- 2008.

Source: U. S. EPA ( 2011 ) 

### Figure 3

- 2. Trends in Pb emissions (thousand tons) from stationary and mobile sources in the U.S., 1990

- 2008.May 20 11 3

- 4 DRAFT

– DO NOT CITE OR QUOTE 

### Direct emissions of Pb into the atmosphere primarily come from piston engine aircraft, fuel 

## 1

combustion, and industrial activities. Direct Pb point source emissions estim ated by the 2008 NEI are 

## 2

shown in Figure 3

- 3. Emissions from piston engine aircraft emissions comprised nearly half of all 

## 3

emissions (590 tons). Industrial fuel combustion contributed 220 tons (18%) of Pb emissions in 2008, 

## 4

followed by metal working and mi ning (12%), other industry (9%), dusts from construction (7%) and 

## 5

miscellaneous contributions from agriculture, solvent utilization, and operation of commercial marine 

## 6

vessels and locomotives (4%) ( U.S. EPA, 2011 )

. Pb emissions from the “ metal working and mining ” 

## 7

category include the single primary Pb smelter in the U.S. , the Doe Run facility in Herculaneum, MO; 

## 8

secondary Pb smelters, mostly designed to reclaim Pb fo r use in Pb

- acid batteries; and smelters for other 

## 9

metals. 

## 10

Source: U. S. EPA ( 2011 ) 

### Figure 3

- 3. Nationwide stationary source Pb emissions (tons / y r ) in the U.S. by source sector in 200 8

.

There is substantial variability in Pb emissions from stationary sources across U.S. counties, as 

## 11

shown in Figure 3

- 4 for the continental U.S. The emissions levels, shown in units of tons, vary over 

## 12

several orders of magnitude. Ninety

- four percent of U.S. counties had 2008 emissions below 1 ton, and 

## 13

50% of counties had 2008 emissions below 0.044 tons. The upper 0.1% of stationary emissions came 

## 14

from thirty

- three counties. This category included all counties emitting more than 3.8 tons of Pb in 2008. 

## 15

Jefferson County, MO was the highest emitting county, with over 21 tons of airborne Pb emissions in 

## 16May 20 11 3

- 5 DRAFT

– DO NOT CITE OR QUOTE 

### 2008. Jefferson County is home to the Doe Run primary Pb smelting facility, which is the only remaining 

## 1

operational primary Pb smelter in the U.S. 

## 2

Sourc e: U.S. EPA ( 2011 ) 

### Figure 3

- 4. County

- level Pb emissions (tons) from stationary sources in the U.S. in 2008.

Pb emissions from piston engine aircraft operating on leaded fuel occur at approximately 20,000 

## 3

airports across the U.S. Figure 3

- 5 displays Pb emissions (tons) at airports around the continental U.S. 

## 4

The map illustrates that airport emissions tend to be e levated around highly populated metropolitan 

## 5

regions, which typically have multiple airports with varying activity levels. Clusters of airports around 

## 6

metropolitan areas and megapolitan regions (multiple contiguous metropolitan areas) are notable from 

## 7

Figu re 3

- 5. Among these sites, piston aircraft emissions at airports within twenty

- eight counties 

## 8

cumulatively emitted greater than one ton of Pb in 2008 U.S. EPA ( 2011 )

. Additionally, wi thin the 2008 

## 9

NEI, there were estimates of Pb emissions during flight from piston aircraft. These estimates are provided 

## 10

by state and cumulatively account for 296 tons in 2008. 

## 11May 20 11 3

- 6 DRAFT

– DO NOT CITE OR QUOTE Source: U.S. EPA ( 2011 ) 

### Figure 3

- 5

. Pb emissions (tons) at airports in the U.S. in 2008. The size of the symbol indicates the magnitude of Pb emissions at each airport.



# 3.2.2. Anthropogenic Sources

Anthropogenic Pb source categories are organized below in order of magnitude reported on the 

## 1

2008 NEI ( U.S. EPA, 2011 ) , with emissions from piston engine aircraft b eing the highest and resuspended 

## 2

dust from previously deposited Pb being substantially lower. 

## 3

3.2.2.1. Lead Emissions from Piston Engine Aircraft Operating on Leaded

Aviation Gasoline and Other Non

- Road Sources

The largest source of Pb in the NEI is from piston en gine aircraft operating on leaded aviation 

## 4

gasoline emissions ( U.S. EPA, 2011 ) . Murphy et al. ( 2007 ) cited fuel consumption estimates provided by 

## 5

the U.S. Department of Energy indicating that 575,000 kg/yr of Pb is used in piston engine aircraft fuel, 

## 6

and based on the 1999 NE I, that 490,000 kg/yr (85%) become airborne upon combustion. Levin et al. 

## 7

( 2008 ) point out that emissions from piston engine aircraft are exempt from reporting to the EPA Toxic 

## 8May 20 11 3

- 7 DRAFT

– DO NOT CITE OR QUOTE 

### Release Inventory. Levin et al. ( 2008 ) summarized findings from environmental protection departments of 

## 1

the State of Illinois, the U.S., and Canada regarding ambient Pb concentrations at and near airports. The 

## 2

Canadian report noted average and maximum air Pb levels were 0.030 and 0.30 μg/m 3 

### , respectively, 

## 3

compared with background levels of 0.007 and 0.018 μg/m 3 

### ( Conor Pacific Environmental Technologies 

## 4

Inc, 2000 )

. The Illinois report noted that air Pb concentrations were elevated downw ind of O’Hare airport 

## 5

compared with upwind levels ( Illinois Environmental Protection Agency, 2002 )

. Pb emission rates from 

## 6

piston aircraft vary with fuel consumption rates, which depend on the engine/airframe combination and 

## 7

the mode of operation of the aircraft. Fuel consumption rates can be obtained for some engine/aircraft 

## 8

combinations by running FAA’s Emissions and Dispersion Modeling System ( FAA, 2011 )

. The ASTM 

## 9

specification for the maximum Pb content is “100 Low Lead”, the most commonly used leaded avgas, is 

## 10

2.12 g of elemental Pb/gallon ( ASTM, 2007 )

. 

## 11

Dynamometer testing has indicated that Pb emissions from piston engine aircraft fuel combustion 

## 12

can occur in the particulate and gaseous form. For example, Gidney et al. ( 2010 ) performed dynamometer 

## 13

testing on automobiles operating on standard gasoline and on gasoline with low levels of organometallic 

## 14

additives. Tetraethyl Pb was included since it i s still used in avgas. The additives had trace levels of Pb 

## 15

that exist in gas phase when temperatures are higher than 650 C, below which they condense to 

## 16

particulate phase. Gidney et al. ( 2010 ) point out that, where tetraethyl Pb is used as an additive in piston 

## 17

engine aircraft fuel, the fuel also contains ethylene dibromide to act as a Pb “scavenging agent.” When 

## 18

ethylene dibromide reacts with Pb, it forms Pb bromi de and Pb oxybromides, which are more volatile. 

## 19

3.2.2.2. Fugitive Emissions from Metals Processing and Mining

Fugitive emissions from secondary Pb processing can be substantial over the course of a year, but 

## 20

they are difficult to estimate. Goyal et al. ( 2005 ) estimated fugitive emissions using concentration data 

## 21

obtained from samplers sited in close vicinity of secondary P b recovery facilities and meteorological data 

## 22

from nearby weather monitoring stations. Regression modeling and Bayesian hierarchical modeling were 

## 23

both used to estimate fugitive and stack emissions from facilities in Florida, Texas, and New York. 

## 24

Depending on the model used, median fugitive emissions were estimated to be 1.0 × 10 - 6 

### to 4.4 × 10 - 5 

### g 

## 25

Pb/ m 2 

### ·sec at the Florida site, 9.4 × 10 - 7 

### to 2.0 × 10 - 6 

### g/m 2 

### ·sec for the Texas site, and 8.8 x 10 - 7 

### to 1.1 × 10 - 

## 26

6 

### g/m 2 

### ·sec at the New York site. Median stack emissions estimates varied widely among the models, with 

## 27

the Florida site median ranging from 1.4 × 10 - 6 

### to 1.4 × 10 - 1 

### g Pb/sec, the Texas site median ranging from 

## 28

8.4 × 10 - 2 

### to 8.6 × 10 - 2 

### g/sec, and the New York site ranging from 8.4 × 10 - 3 

### to 1.0 × 10 - 2 

### g/sec. 

## 29

Additionally, the Bayesian hierarchical model was used to estimate fugitive Pb emissions nationwide 

## 30

using concentration data as prior information. Nationwide median fugitive emissions were estimated to be 

## 31

9.4 × 10 - 7 

### to 3.3 × 10 - 6 

### g/m 2 

### ·sec. 

## 32May 20 11 3

- 8 DRAFT

– DO NOT CITE OR QUOTE 

### Waste fro m current or defunct mines has been shown to present an additional fugitive source of Pb. 

## 1

For example, Zheng et al. ( 2009 ) a pplied source apportionment in three northeastern Oklahoma towns to 

## 2

identify the influence of “chat”, or waste piles from formerly operational Pb

- Zn mines, on PM 10

- 2.5 

### and 

## 3

PM 2.5 

### . They estimated that mine waste was responsible for 88% of Pb in PM 10

- 2.5 

### samples and 40% of Pb 

## 4

in PM 2.5 

### samples. 

## 5

3.2.2.3. Fossil Fuel Combustion

Fossil fuel combustion accounts for roughly 12% of Pb emissions in the U.S. Murphy et al. ( 2007 ) 

## 6

presented an es timated U.S. mass budget for Pb emitted from consumption of select fuels and crude oil. 

## 7

Fuel consumption estimates for 2005 were employed ( Freme, 2004 )

. Based on an annual cons umption of 

## 8

1.0 × 10 9 

### tons coal with an average Pb concentration of 20 mg/kg (range : 5 to 35 mg/kg ) and using an 

## 9

emission factor (airborne fraction) of approximately 0.01, coal contributed approximately 200 tons Pb / y r 

## 10

to the atmosphere. There were no emissi on factors for crude oil or residual oil but these represent 

## 11

potentially large sources (up to 100

- 500 tons/yr and up to 25

- 700 tons/ y r , respectively). The amounts of 

## 12

Pb emitted from these U.S. sources, however, are several orders magnitude smaller than tho se estimated 

## 13

to arise from coal combustion in China. 

## 14

Coal combustion is considered to be a major source of Pb in the atmosphere now that leaded 

## 15

gasoline has been phased out for use in on

- road vehicles ( Diaz

- Somoano et al., 2009 )

. Global Pb 

## 16

estimates are considered here to inform understanding of U.S. Pb emissions from coal combustion. 

## 17

Globally, Pb emissions from stationary sources have been increasing and the north

- south gradient in 

## 18

aerosol Pb concentrations over the Atlantic Ocean has disappeared as a result of industrialization of the 

## 19

s outhern h emisphere ( J. M. Pacyna & Pacyna, 2001 ; Witt et al., 2006 )

. The Pb isotope ratio v alues 

## 20

(mainly 206 

### Pb/ 207 

### Pb) for coals from around the world have been compared with those for atmospheric 

## 21

aerosols. In most parts of the world, there ha s been a difference between the signature for aerosols and 

## 22

that for coal, where the atmospheric 206 

### Pb/ 207 

### Pb ratio values are lower, indicative of additional 

## 23

contributions from other sources. 

## 24

Rauch and Pacyna ( 2009 ) constructed global metal cycles using anthropogenic data from 2000. 

## 25

They confirmed that the largest anthropogenic airborne Pb emissions arise from fossil fuel combustion, 

## 26

and they quantif ied Pb emissions at 85,000 tons/ y r worldwide. Using a separate global model, Niisoe et 

## 27

al. ( 2010 ) calculated emissions of Pb from coal combustion in J apan during 2000 to be 900 tons/ y r , based 

## 28

on 9 × 10 7 

### ton s/yr coal combustion and an emission factor of 10 g Pb / ton. The equivalent value for Pb 

## 29

emissions from China was 56,000 tons Pb / y r . Calculated Pb concentrations in surface air for China agreed 

## 30

with this value within a factor of two, although there was a systematic underestimation suggesting an 

## 31

incomplete knowledge of the Pb emissions ( Niisoe et al., 2010 ) . It was notable, however, that the 

## 32May 20 11 3

- 9 DRAFT

– DO NOT CITE OR QUOTE 

### calculated emissions from Chinese coal combustion make up a substantial proportion (~66%) of the total 

## 1

global Pb emissions from fossil fuels detailed in Rauch and Pacyna ( 2009 )

. 

## 2

Tan et al. ( 2006 ) compared several emissions sources in Shanghai, China . They estimated emission 

## 3

values for on

- road exhaust from use of Pb

- free gasoline ( 238 ± 5 mg/kg), vehicle exhaust from leaded on



## 4

road gasoline (7,804 ± 160 mg/kg ), coal combustion (1,788 ± 37 mg/kg ), metallurgic dust (6,140 ± 

## 5

130 mg/kg ), soil (11.7 ± 0.3 mg/kg ), and cement (103 ± 2 mg/kg ). Pb

- free automobile gasoline has been 

## 6

in use in Shanghai since 1997. The isotope ratios for each of these emission sources were determined. 

## 7

Based on the 4.4 × 10 7 

### tons of coal combusted annually in Shanghai, an average coal Pb concentration of 

## 8

13.6 ± 6.6 mg/kg , and an emission factor of 0.5, approximate ly 300 tons Pb was being emitted annually in 

## 9

association with fine PM . They concluded that a major priority should be to reduce Pb emissions from 

## 10

coal combustion now that the contribution from vehicle exhaust emissions has decreased. 

## 11

Seasonal effects of th e contributions of Pb emissions from coal combustion have been observed

. 

## 12

For example, in Tianjin, northern China, t he winter heating period starts in November, and the 

## 13

contribution from coal combustion to the Pb aerosol becomes high during the winter. This leads to both a 

## 14

high Pb content and a high 206 

### Pb/ 207 

### Pb ratio. Coal consumption and Pb

- bearing PM concentrations 

## 15

declined during the summer months, and Pb from other sources, mainly vehicle exhaust emissions, 

## 16

became relatively more pronounced ( W. Wang et al., 2006 ) . This season al relationship contrasts with 

## 17

observations for the U.S. described in the 2006 Pb AQCD ( U.S. EPA, 2006 ) which indicated that for West 

## 18

Virginia, higher emissions from power stations occurred in summer months. The increased energy use in 

## 19

summer periods in the U.S. may be attributable to increased requirements for air

- conditioning. 

## 20

3.2.2.4. Other Industrial Sources

Several Pb isotope studies have been used to distinguish contributions from industrial activities

. 

## 21

For example, in northern China, Wang et al. ( 2006 ) noted that, in the response to decre asing atmospheric 

## 22

Pb concentrations in total suspended particles ( TSP ) samples collected from 1994 to 1998, the 206 

### Pb/ 207 

### Pb 

## 23

isotope ratio showed a related trend with values of ~1.149 in 1994 increasing to ~1.161 in 1998. The Pb 

## 24

concentration and isotope ra tio values then remained approximately constant from 1998 through to 2001. 

## 25

Although this was consistent with a decreasing contribution of Pb from on

- road gasoline, the ratio values 

## 26

were still lower than those for Chinese coal [ 206 

### Pb/ 207 

### Pb ~1.18 in Mukai et al. ( 2001 ) ] , suggesting that 

## 27

local Pb ore sources typically have lower 206 

### Pb/ 207 

### Pb ratios. The range for Chinese ore s was 1.081 to 

## 28

1.176

, with lower values correspond ing to ores from northern China ( Mukai et al., 200 1 )

. 

## 29

Novak et al. ( 2008 ) evaluated changes in the amounts and sources of Pb emissions in t he U.K. and 

## 30

Czech Republic during the 19 th and 20 th centuries

. Deconvolution of sources was attempted using Pb 

## 31

isotopes, but one major area of uncertainty was the amount and the isotope composition of Pb emanating 

## 32

from incineration plants, particularly in the U.K . The isotopic signature of Pb recycled into the 

## 33May 20 11 3

- 10 DRAFT

– DO NOT CITE OR QUOTE 

### atmosph ere by incineration of various industrial wastes could have shifted from relatively high 206 

### Pb/ 207 

### Pb 

## 1

ratios consistent with local Variscan ores to lower values reflecting imported Precambrian ores. There 

## 2

have, however, been other environmental studies conc erning incineration, and these give a highly 

## 3

consistent value for the Pb isotope ratio for European incineration sources. For example, Cloquet et al. 

## 4

( 2006 ) showed that the Pb isotopic composition of urban waste incineration flue gases in northeastern 

## 5

France was ~1.155. de la Cruz et al. ( 2009 ) reported that waste incineration was an important source of 

## 6

Pb and showed that the 206 

### Pb/ 207 

### Pb and 208 

### Pb/ 207 

### Pb ratios for waste incineration Pb emitted in European 

## 7

countries were 1.1427

- 1.1576 and 2.4260

- 2.4346, respectively, i.e.

, quite a narrow range ( de la Cruz et 

## 8

al., 2009 and references therein )

. 

## 9

3.2.2.5. Roadway

- Related Sources

Contemporaneous Emissions from Vehicle Parts

Contemporaneous Pb emissions from motor vehicles may occur because several vehicle parts still 

## 10

contain Pb. Wheel weights , used to balance tires, are clipped to the rims of every automobile in the U.S. 

## 11

in order to balance the tires , and may be come loose and fall off . Ambient air Pb concentrations near 

## 12

heavily trafficked areas may be related to use of Pb

- based wheel weights that are prone to dislodgemen t. 

## 13

On pavement they may be ground into fine PM by the pounding forces of traffic ( Root, 2000 ) . For 

## 14

example, Schauer et al. ( 2006 ) measured Pb emissions in two traffic tunnels and f ound that the fraction of 

## 15

Pb in PM 2.5 

### was no more than 17% of Pb measured in PM 10 

### . Schauer et al. ( 2006 ) suggested that 

## 16

enrichment in the coarse fraction may have been related to wh eel weights. Additionally, Schauer et al. 

## 17

( 2006 ) measured PM 10 

### and PM 2.5 

### composition from brake dust and found low but substantial quantities of 

## 18

Pb in PM 10 

### (0.02 ± 0.01 mg/g) and PM 2.5 

### (0.01 ± 0.00 mg/g) for semi

- metallic brake pads and in PM 10 

## 19

(0.01 ± 0.00 mg/g) for low

- metallic brake pads. Additionally, Hjortenkrans et al. ( 2007 ) used material 

## 20

metal concentrations, traffic volume, emissions factors, and sales data to estimate the quantity of Pb 

## 21

emitted from brake wear and tires in Stockholm, Sweden in 2005. They observed that 24 kg Pb were 

## 22

emitted from brake wear each year, compared with 2.6 kg of Pb from tire tread wear; an estimated 549 kg 

## 23

was estimated to have been emitted from brake wear in 1998. McKenzie et al. ( 2009 ) determined the 

## 24

composition of various vehicle components including t i res and brakes and found that t i res were a possible 

## 25

source of Pb in stormwater, but no identification of Pb

- containing PM in stormwater was carried out. 

## 26

However, PM from t i re abrasion are usually found in coarser size ranges ( Chon et al., 2010 )

, whil e those 

## 27

in the submicron range are more typically associated with combustion and incineration sources. 

## 28May 20 11 3

- 11 DRAFT

– DO NOT CITE OR QUOTE 

### Unleaded Fuel

Unleaded fuel contains Pb as an impurity within crude oil ( E. G. Pacyna et al., 2007 ) . Schauer et al. 

## 1

( 2006 ) measured Pb in PM 2.5 

### from tailpipe emissions and observed significant quantities in on

- road 

## 2

gasoline emissions (83.5 ± 12.80 mg/kg) and higher but non

- significant quantities in diesel emissions 

## 3

(137 ± 133 mg/kg)

. Hu et al. ( 2009 ) investig ated the heavy metal content of diesel fuel and lubricating oil

. 

## 4

They found <1

- 3 ppm Pb in samples of lubricating oil. Hu et al. ( 2009 ) also measured the size distribution 

## 5

of Pb emissions during dynamometer testing of heavy duty diesel vehicles with different driving patterns 

## 6

and co ntrol technologies

. An urban dynamic driving schedule (UDDS) designed to mimic urban stop

- go 

## 7

driving conditions, was simulated in two cases to produce 80 and 241 ng Pb/km driven, depending on the 

## 8

control technology used

. Respectively, 54% and 33% of those emissions were smaller than 0.25 μm in 

## 9

MMAD

. The Tan et al. ( 2006 ) study cited in Section 3.2.2.3 for Shanghai, China, where unleaded fuel has 

## 10

been in use since 1997, illustrated that Pb emissions from unleaded on

- road gasoline were substantially 

## 11

lower than Pb emissions from coal combustion

. Assuming that the natural Pb content of unleaded fuels in 

## 12

the U.S. is similar, it is unlikely that road vehicle combustion of unleaded on

- road gasoline is currently a 

## 13

major contributor to total Pb emissions in the U.S. 

## 14

3.2.2.6. Deposited Lead

Soil Pb can serve as a reservoi r for deposited Pb. The following subsections describe studies of 

## 15

previously deposited Pb from industrial, historical leaded on

- road emissions, and urban sources such as 

## 16

paint and building materials. The 2006 Pb AQCD ( U.S. EPA, 2006 ) cited an estimate by Har ris and 

## 17

Davidson ( 2005 ) that more than 90% of airborne Pb emissions in the South Coast Basin of California 

## 18

were from soil resuspension. Since publication of the 2006 Pb AQCD ( U.S. EPA, 2006 ) , further a nalysis 

## 19

of the Harris and Davidson ( 2005 ) paper has revealed that the contributions of Pb from piston engine 

## 20

aircraft were underestimated compared with the 2002 NEI. Assumptions of spatial uniformity incurred by 

## 21

the “continuously stirred reactor” mass balance model and for mixing layer height used by Harris and 

## 22

Davidson ( 2005 ) were also not valid because Pb concentrations are spatially heterogen eous at the urban 

## 23

scale; see Section 3.5. Therefore, the estimate of 90% of airborne Pb from resuspension is not employed 

## 24

in the current assessment. Currently, data are not available with sufficient spatial resolution to discern the 

## 25

specific contribution o f soil Pb resuspension to air Pb concentration, but resuspended soil Pb cannot be 

## 26

eliminated as a potential source of airborne Pb. 

## 27

Lead from Industrial Sites

Several studies have indicated elevated levels of Pb in soil exposed to industrial emissions, 

## 28

incl uding brownfield sites ( Deng & Jennings, 2006 ; Dermont et al., 2010 ; Hofer et al., 2010 ; Jennings & 

## 29May 20 11 3

- 12 DRAFT

– DO NOT CITE OR QUOTE 

### Ma, 2007 ; Sriskandan et al., 2007 ; van Herwijnen et al., 2007 ; Verst raete & Van Meirvenne, 2008 )

. It is 

## 1

possible that P b in soil serves as a source of airborne Pb. Laidlaw and Filipelli ( 2008 ) reviewed the 

## 2

literature on Pb resuspension from soil and then analyzed IMPROVE data to explore conditions under 

## 3

which Pb may become resuspended. They observed a seasonal pattern in concentration of soil 

## 4

resuspended in the atmosphere, and they also found that 83% of the variability in concentrations of soil in 

## 5

the atmosphere was predicted by the variability in meteorology and soil mois ture content. The authors 

## 6

concluded that seasonality and climate parameters could not be eliminated in relation to ambient Pb 

## 7

concentrations. Such mechanisms are described in more detail in Section 3.3. 

## 8

Lead from Paint and Building Materials

Exterior painted structures have long been known to be a source of ambient Pb ( U.S. EPA, 2006 )

. 

## 9

Recent studies support older findings. Mielke and Gonzales ( 2008 ) sampled exterior paint chips from 25 

## 10

homes in New Orleans, LA, and they found elevated Pb levels in 24 of the 25 tested e xterior paints 

## 11

(median: 36,000 mg/kg). Weiss et al. ( 2006 ) studied the distribution of Pb concentration in roadway grit 

## 12

in the vicinity of steel structures in New York City and contrasted those d ata with roadway grit 

## 13

concentration data where no steel structure was nearby. In each case, the comparison was significant (p 

## 14

<0.006 at one site and p <0.0001 at 4 other sites), with median Pb concentrations under the steel 

## 15

structures being between 2.5 to 11 times higher than median Pb concentrations not near a structure. 

## 16

The studies described above considered paint as a source of Pb dust through gradual abrasion of the 

## 17

painted surfaces. However, ambient conditions may also affect the availability of Pb in paints. Edwards et 

## 18

al. ( 2009 ) performed experiments to simulate one week of exposure of Pb

- based paints to elevated levels 

## 19

of O 3 

### (11.3 ± 0.8 ppm or 150 times the level of the 8

- h NAAQS) and NO 2 

### (11.6 ± 0.9 ppm, or 220 times 

## 20

the level of the annual NAAQS). Following NO 2 

### exposure, the Pb availability in wipe samples increased 

## 21

by a median of 260% (p < 0.001), a nd following O 3 

### exposure, the Pb availability increased by a median of 

## 22

32% (p = 0.004). Edwards et al. ( 2009 ) state that the high O 3 

### and NO 2 

### concentrations simulated in the 

## 23

chamber were equivalent of 4.3 and 3.7 years of exposure at 50 and 60 ppb, respectively. 

## 24

Building demolition was listed as a source of urban Pb dust in the 2006 Pb AQCD ( U.S. EPA, 

## 25

2006 ) . In a follow

- up study to previous work cited therein, Farfel et al. ( 2005 ) observed that Pb dust 

## 26

surface loadings increased by 200% in streets, by 138% in alleys, and by 26% in sidewalks immediately 

## 27

following demolition of an old building. One month later, Pb dust loadings were still elevated in alleys 

## 28

(18%) and sidewalks (18%), although they had decreased in streets by 29%. However, Farfel et al. ( 2005 ) 

## 29

did not provide detai led time series samples from before or after demolition to judge whether the 

## 30

observations made one month following demolition were within the normal conditions of the urban area. 

## 31

These results suggest that building demolition may be a short

- term source of Pb in the environment, but it 

## 32

is unclear if demolition is related to long

- term Pb persistence in the environment. 

## 33May 20 11 3

- 13 DRAFT

– DO NOT CITE OR QUOTE 

### Lead from Historic Automobile Emissions

Historic Pb emissions, or Pb emitted from on

- road vehicles prior to the ban on use of leaded 

## 1

automobil e gasoline, deposited onto soil and still may be persistent in the environment as a potential 

## 2

source of airborne Pb. The historical combustion of leaded on

- road gasoline has been estimated from 

## 3

documents submitted by Ethyl Corporation to the U.S. Senate ( "Airborne Lead Reduction Act of 1984," 

## 4

1984 ) and a report by the U.S. Geological Survey ( USGS, 2005 ) ; see Mielke et al. ( 2010b ) . These 

## 5

estimates are presented in Figure 3

- 6. The pe ak U.S. use of Pb additives occurred between 1968 and 1972 

## 6

with an annual amount of over 200,000 metric tons. According to Ethyl Corporation, the 1970 use of Pb 

## 7

additives was 211,000 metric tons. By 1980, the annual use of Pb additives to on

- road gasoline decreased 

## 8

to about 91,000 metric tons or a 57% reduction from its 1970 peak. From 1970 to 1990 there was a 92% 

## 9

decline in Pb additive use. In 1990, the annual U.S. use of Pb additives decreased to 16,000 metric tons, a 

## 10

further 82% decline in Pb additive us e from 1980. The final U.S. ban on the use of Pb additives for 

## 11

highway use in on

- road gasoline occurred in 1996

. After that time, Pb additives were only allowed in non



## 12

highway applications, including piston engine aircraft fuel, racing fuels, farm tractors , snowmobiles, and 

## 13

boats. 

## 14

Source: Used with permission from Pergamon Press, Mielke et al. ( 2010b )

. 

### Figure 3

- 6. Total U.S. Pb additives in on

- road gasoline used in on

- road vehicles, 1927

- 1995. Estimates were derived from the proceedings of the U.S. Senate hearings on the Airborne Pb Reduction Act of 1984, S. 2609 ( "Airborne Lead Reduction Act of 1984," 1984 ) and the U.S. Geological Survey Pb end use
statistics ( USGS, 2005 )

. Lead Additives in U.S. Gasoline

0 50,000 100,000 150,000 200,000 250,000

1925 1930 1935 1940 1945 1950 1955 1960 1965 1970 1975 1980 1985 1990 1995

YearMetric TonsMay 20 11 3

- 14 DRAFT

– DO NOT CITE OR QUOTE 

### Th e particle sizes of Pb emissions from on

- road sources were estimated by the U.S. EPA ( 1986 )

, 

## 1

which indicated that 75% of Pb additives were emitted as exhaust. The tonnages of relatively large 

## 2

>10 μm mass m edian aerodynamic diameter (MMAD) Pb

- PM probably settled locally, especially in high 

## 3

traffic urbanized areas where soil Pb, from historic emissions as well as contemporaneous sources, are 

## 4

elevated adjacent to roadways and decrease with distance away from t he roadway ( Laidlaw & Filippelli, 

## 5

2008 ) . EPA ( 1986 ) indicated that 35% of the PM were < 0.25 μm in MMAD. The majority of PM, by 

## 6

number, emitted from automobiles was in the ultrafine size range ( Londahl et al., 2009 )

. However, 

## 7

Londahl et al. ( 2009 ) did not include ethylene dibromide to the fuel in these experiments. As described in 

## 8

the Gidney et al. ( 2010 ) study referenced in Section 3.2.2.1, it was found that ethylene dibromide acted to 

## 9

scavenge Pb in PM to form more volatile Pb bromide and Pb oxybromide to produce gaseous Pb 

## 10

emissions. 

## 11

The use of Pb additives also resulted in a national scale of influence. For example, variou s sized 

## 12

urbanized areas of the U.S. have different amounts of vehicle traffic associated with Pb ( Mielke et al., 

## 13

2010a )

. Figure 3

- 7 illustrates the national scale of the estimated vehicle

- derived Pb aerosol emissions

. 

## 14

Note that the estimated 1950

- 1982 Pb aerosol emissions in the 90 cities below vary from 606 metric tons 

## 15

for Laredo, Texa s, to nearly 150,000 metric tons for the Los Angeles

- Long Beach

- Santa Anna urbanized 

## 16

area

. The implication of this figure is that the soil Pb concentration will be proportional to the magnitude 

## 17

of historic on

- road emissions in each city

. It is recognized t hat the amount of soil turnover since 1982 may 

## 18

have varied substantially among the cities illustrated in Figure 3

- 7, depending on the amount of highway 

## 19

construction in those cities

. Hence, the map may overestimate potential amounts of Pb in soil, and 

## 20

conse quently of airborne Pb from resuspended soil, in some of the cities illustrated. 

## 21May 20 11 3

- 15 DRAFT

– DO NOT CITE OR QUOTE Source: Used with permission from Pergamon Press, Mielke et al. ( 2010b ) 

### Figure 3

- 7

. Estimated Pb aerosol inputs fr om on
- road gasoline into 90 U.S. urbanized areas (UAs), from 1950 through 1982. The numbers on the map are rankings of each UA

. The size of each dot refers to the magnitude of motor vehicle gasoline

- related emissions for each group of UAs. The extremes are , Los Angeles UA (ranked #1) and Laredo, Texas (ranked #90). Some of the UAs have been used as sites in soil Pb studies, as
indicated in Table 3

- 9.



# 3.2.3. S ource Attribution

3.2.3.1. Lead Speciation and Source Apportionment

The following section describes new findings wi th respect to speciation of Pb content in aerosols. 

## 1

Analytic techniques for speciation are explicated in Section 3.4. Forms of Pb commonly observed in the 

## 2

environment are presented in Table 3

- 1 to serve as a reference for the categories of Pb sources described 

## 3

in Sections 3.2.1 and 3.2.2. Detailed descriptions of the related chemistry were presented in the 2006 Pb 

## 4

AQCD ( U.S. E PA, 2006 )

. 

## 5May 20 11 3

- 16 DRAFT

– DO NOT CITE OR QUOTE 

### Table 3

- 1. Pb compounds observed in the environment

Emission Source Observed Pb Compounds

Minerals PbS (Galena) PbO (Litharge, Massicot) Pb 3 O 4 (Mini m um or “ Red Pb ” ) PbSO 4 (Anglesite) Smelting Aerosols Pb 0 , PbS PbSO 4 , PbO PbC O 3 Pb silicates Coal Combustion Aerosols PbS PbSe Coal Combustion Flue Gases Pb 0 , PbO, PbO 2 (Above 1150 K) PbC l 2 (Low rank coals, above 1150 K) PbSO 4 (Below 1150 K) Wood Combustion PbC O 3 Waste Incineration Aerosols PbC l 2 , PbO Soils Near Mining Operations PbC O 3 PbSO 4 [PbFe 6 (SO 4 ) 4 (OH) 12 ] [Pb 5 (PO 4 ) 3 Cl] [Pb 4 SO 4 (C O 3 ) 2 (OH) 3 ] PbS

- Bi 2 S 3 Pb oxides, silicates Motor Vehicle Exhaust (Combustion of Leaded Fuel) 1 PbBrCl PbBrCl

- 2NH 4 Cl PbBrCl

- NH 4 Cl Roadside Dust PbSO 4 , Pb 0, PbSO 4 (NH 4 )SO 4 , Pb 3 O 4 , PbO

- PbSO 4 and 2PbC O 3 - Pb(OH) 2 Other mobile sources: Brake wear, wheel weights Pb 0 Racing vehicle emissions Pb halides Aircraft Engine Wear Pb 0

Lawn Mowers Pb halides (battery leakage)

Source: Biggins and Harrison ( 1979

, 1980 ) ; U.S. EPA ( 2006 )

.



### Chemical speciation of Pb by source was reviewed in the 2006 Pb AQCD ( U.S. EPA, 2006 ) , and is 

## 1

fairly well understood and varies considerably between sources. Pb components from Pb smelters are 

## 2

mainly elemental Pb (Pb 0 

### ), Pb sulfide (PbS), Pb sulfates (PbSO 4 

### , PbO

- PbSO 4 

### ), and Pb oxide (PbO), with 

## 3

Pb carbonate (PbCO 3 

### ) and Pb silicates. Other smelters also produce Pb, mainly as Pb oxide (PbO). Pb in 

## 4

coal combustion emissions is mainly in the form of Pb sulfide (PbS) and Pb sulfate (PbSO 4 

### ) with some Pb 

## 5

selenide (PbSe) as well as sulfates, oxides and chlorides. In wood combustion emissions PbCO 3 

### and Pb 

## 6

oxides are important. Waste incineration produces mainly Pb chloride (PbCl 2 

### ) and PbO. Resuspended 

## 7

mining soils were reported to be abundant in PbCO 3 

### and PbSO 4 

### , as well as a variety of complex salts 

## 8

containing phosphates, hydroxides, chlorides, oxides, and silicates. On

- road engine exhaust from leaded 

## 9

gasoline use contained mostly Pb bromide and Pb chloride salts, including PbBrC l, PbBrCl

- NH 4 

### Cl, and 

## 10

PbBrCl

- 2NH 4 

### Cl. Road dust is rich in PbSO 4 

### , Pb 0 

### , oxides, carbonates, and hydroxides. Motor vehicles also 

## 11

contribute Pb 0 

### from brake wear and wheel weights. 

## 12

The 2006 Pb AQCD ( U.S. EPA, 2006 ) describe d the atmosphere as the major environme nt al 

## 13

transport pathway for Pb, with Pb primarily present in submicron aerosols . Although not directly 

## 14

addressed in the 2006 Pb AQCD ( U.S. EPA, 2006 ) , organolead vapor emissions were extensively 

## 15

discussed in the 1986 Pb AQCD ( U.S. EPA, 1986 ) which concluded that they were primarily emitted 

## 16May 20 11 3

- 17 DRAFT

– DO NOT CITE OR QUOTE 

### during manufacture, transport, and handling of leaded on

- road gasoline. Organolead vapors contributed 

## 1

less than 10% of vehicular Pb tailpipe emissions when leaded on

- road gasoline w as still in use. Studies of 

## 2

Pb emissions within enclosed microenvironments where automobiles were the dominant Pb source cited 

## 3

within the 1986 Pb AQCD ( U.S. EPA, 1986 ) , reported that organic Pb vapors contribute d less than 20% 

## 4

of total vehicular Pb emissions. More recent studies support this ( Shotyk et al., 2002 )

. The 20% estimate 

## 5

of organic Pb vapors from the previous studies of on

- road Pb emissions may potentially provide an upper 

## 6

bound for organic Pb vapors from current piston engine aircraft emissions. 

## 7

Several recent studies have used speciation techniques either simply to determine the chemical 

## 8

composition of emissions or for sou rce attribution. In urban environments, airborne Pb concentrations are 

## 9

likely a mix of various sources. Ogulei et al. ( 2006 ) performed source apportionment of PM 2.5 

### samples in 

## 10

Baltimore, MD and found multiple industrial, urban, and background influences. Sixty

- three percent of 

## 11

the Pb was associated with incineration, while 20% was associated with a wildfire episode from which 

## 12

PM was transported from Quebec, 8% was associated with secondary nitrate, 6% was associated with 

## 13

operations at a steel plant, and 3% was associated with local gasoline traffic. Dillner et al. ( 2006 ) 

## 14

analyzed the composition of PM 2.5 

### and TSP samples in Beijing, China and found that Pb comprised 

## 15

roughly 0.2% of TSP and 1.4% of PM 2.5 

### during industrial pollution events, 0.1% of TSP and 0.5% of 

## 16

PM 2.5 

### during urba n pollution events, and 0.05% of TSP and 0.4% of PM 2.5 

### during dust storms. During 

## 17

industrial pollution events, the authors note that the amount of Pb in PM 2.5 

### can be substantial. 

## 18

Speciation of emissions from a battery recycling facility indicated that PbS was most abundant, 

## 19

followed by Pb sulfates (PbSO 4 

### and PbSO 4 

### - PbO), PbO and Pb 0 

### ( Uzu et al., 2009 ) . Pb speciation 

## 20

emissions from a sintering plant, a major component of the steel making process, were reported for the 

## 21

first time, with cerussite, a Pb carbonate (PbCO 3 

### - 2H 2 

### O), emerging as the most abundant species ( Sammut 

## 22

et al., 2010 ) . The predominance of carbonates as major Pb species in industrial emissions is unusual. 

## 23

Choel et al. ( 2006 ) confirmed that Pb was strongly associated with sulfur in smelter emission PM, and 

## 24

that Pb sulfates and Pb oxy

- sulfates were the most abundant species, with important contributions from 

## 25

Pb oxides. Zhang et al. ( 2009 ) used single particle aerosol mass spectrometry (ATOFMS) to speciate Pb



## 26

bearing PM in Shanghai, China in 2007. PM containing Pb along with OC and/or EC was attributed to 

## 27

coal combustion processes; this accounted for roughly 45% of Pb

- bearing PM. PM producing high 

## 28

correlations between Cl and Pb were ascribed to waste incineration, while Pb

- bearing PM with a st rong 

## 29

phosphate signal was attributed to the phosphate industry. 

## 30

A few recent studies have used speciation techniques to characterize Pb and other components of 

## 31

PM 10 

### , PM 2.5 

### , and PM 1


### . Reinard et al. ( 2007 ) used a real

- time single particle mass spectrometer to 

## 32

characterize the composition of PM 1 

### collected in Wilmington, Delaware in 2005 and 2006. 

## 33

Approximately two

- thirds of PM 1 

### consisted of secondary aerosols, e.g. mainly sulfate, nitrate and 

## 34

primary/secondary organics. The remaining third included PM from biomass burning, fossil fuel 

## 35

combustion and various industrial sources. For the latter group, strong Pb

- Zn

- K

- Na associations were 

## 36May 20 11 3

- 18 DRAFT

– DO NOT CITE OR QUOTE 

### observed. Comparison with stack emissions revealed that a nearby steel manufacturing facility was an 

## 1

important source of Pb. Ambient PM classes containing only a subset of such elements, e.g.

, Zn only, Pb



## 2

K only were non

- specific and so could not be mapped to individual sources. Wojas and Almquist ( 2007 ) 

## 3

used ICPMS to characterize trace metals in PM 2.5 

### , PM 10 

### , and TSP samples obtai ned in Oxford, OH. They 

## 4

observed that Pb was highly correlated with several other elements (Ca, Co, Cu, Fe, K, Mg, Mn, Mo, Ni, 

## 5

Pb, Sb, Si, and Zn), suggesting that Pb and copollutants emanated from a variety of sources including 

## 6

road dust and fuel combusti on. Similarly, a study by Moffet et al. ( 2008 ) found that Pb

- Zn

- Cl particles in 

## 7

P M 2.5 

### samples collected from an industrial area in Mexico City represented as much as 73% of fine PM. 

## 8

These were mainly in the submicron size range and were typically mixed with elemental carbon (EC), 

## 9

suggesting a combustion source. The unique single partic le chemical associations closely matched 

## 10

signatures indicative of waste incineration. A study of PM 10 

### and PM 2.5 

### collected in Shanghai, China and 

## 11

analyzed using extended X

- ray absorption fine structure spectroscopy (EXAFS) , found that the main 

## 12

chemical form s of Pb were PbCl 2 

### (41 ± 4%), PbSO 4 

### (37 ± 2%) and PbO (22 ± 3%) ( Tan et al., 2006 )

. 

## 13

There was no significant difference in the chemical forms of Pb between the PM 10 

### and PM 2.5 

### samples. 

## 14

The main sources of these forms of Pb, based on Pb isotopic composition, were coal combustion, 

## 15

metallu rgic dust and vehicle exhaust emissions (none though from leaded on

- road gasoline). 

## 16

Approximately 83% Pb was in the <2.5 μm size range . Murphy et al. ( 2007 ) found that the volatility of 

## 17

Pb and its compounds such as PbO results in its presence at high concentration in the submicron fraction 

## 18

of PM emitted from coal emissions

. PbSO 4 

### , also deri ved from coal combustion, has low solubility 

## 19

( Barrett et al., 2010 )

. Variations in the relative proportions of Pb

- containing compounds may acco unt for 

## 20

the difference in Pb solubility in aerosols ( Fernández Espinosa & Ternero

- Rodríguez, 2004 ; Tan et al., 

## 21

2006 ; von Schneidemesser et al., 2010 )

. 

## 22

Murphy et al. ( 2007 ) also carried out a detailed study of the distribution of Pb in single atmospheric 

## 23

particles. During the fifth Cloud and Aerosol Characterization Experim ent in the Free Troposphere 

## 24

( CLACE 5) campaign conducted at the Jungfraujoch research station, Switzerland, about 5% of analy z ed 

## 25

aerosol particles in PM 1 

### contained Pb. Of these, 35% had a relative signal for Pb greater than 5% of the 

## 26

total mass spectrum me asured by an aerosol time of flight mass spectrometer (ATOFMS). These “ high 

## 27

Pb ” particles also contained one or more positive ions (e.g.

, of Na, Mg, Al, K, Fe, Zn, Mo, Ag, Ba). 

## 28

Sulfate fragments were present in 99% of the negative ion spectra associated wi th high Pb particles and 

## 29

50% also contained nitrite and nitrate. About 80% contained positive and/or negative polarity organic 

## 30

fragments. The average aerodynamic diameter of the Pb

- rich particles (500 nm) was larger than the 

## 31

background aerosol (350 nm) but none had a diameter less than 300 nm. For urban aerosols collected in 

## 32

the U.S. , two types of Pb

- PM were found; in the main class, Pb was found together with K and usually 

## 33

also Zn. There were also minor amounts of Na, EC and organic carbon (OC) including a mines. The 

## 34

second, minor class contained Pb together with Na, K, Zn, smaller amounts of Fe, EC and OC . The size 

## 35

distribution of the first group of Pb

- PM usually peaked around 200 nm. 

## 36May 20 11 3

- 19 DRAFT

– DO NOT CITE OR QUOTE 

### Sample solubility can inform speciation efforts

. For example, a study inv olving weak acid leaching 

## 1

was carried out by Erel et al. ( 2006 ) . The transport of anthropogenic pollution by desert dust in the e astern 

## 2

Mediterranean region was studied by determining major and trace element concentrations, organic 

## 3

pollutants, and Pb isotope ratios. PM 10 

### samples were collected during 10 dust storms in 2001

- 2003. Most 

## 4

samples were po lluted to some extent with pollutants released by weak acid (0.5 M HN O 3 

### ) extraction 

## 5

(includ ing carbonates, oxides and surface

- bound fractions). From the Pb isotope data, most of the Pb 

## 6

came from recently north African emissions and from past Israeli emissi ons (Pb now residing in Israeli 

## 7

soils). 

## 8

3.2.3.2. Lead Isotope Ratio Analysis

Classifying Pb by its relative isotopic abundance has also proved useful for a variety of purposes, 

## 9

including the determination of its geochemical origins in natural samples and the relat ive contributions of 

## 10

coal burning, mining, smelting, and motor vehicle emissions in polluted samples ( Far mer et al., 1996 )

. 

## 11

Typically, isotopes of Pb ( 208 

### Pb, 207 

### Pb, 206 

### Pb, and 204 

### Pb) are measured in a sample using mass 

## 12

spectrometry, and then ratios of the isotopes are calculated to obtain a “signature.” Isotopes of 208 

### Pb, 

## 13

207 

### Pb, and 206 

### Pb are substantially more abundant than 204 

### Pb, but they vary depending on the geologic 

## 14

conditions under which the ore was produced through decay of different isotopes of uranium and thorium 

## 15

( Cheng & Hu, 2010 ) . Isotope ratio analysis was fir st applied to airborne PM in 1965 to identify the 

## 16

impact of motor vehicle exhaust on marine and terrestrial Pb deposition in the Los Angeles area ( Chow & 

## 17

Johnstone, 1965 ) . More recently, high resolution ICPMS has also proved to be a sensitive tool for isotope 

## 18

ratio analysis. High resolution ICPMS was first applied to geological samples ( Walder & Freedman, 

## 19

1992 ) , and has since been widely used for determination of Pb isotope ratios in airborne PM samples. Pb 

## 20

isotope ratios have been measured in a number of recent studies in a variety of loca tions to investigate the 

## 21

origin of airborne Pb ( Hsu et al., 2006 ; Knowlton & Moran, 2010 ; Noble et al., 2 008 ; Widory, 2006 )

. 

## 22

Shotyk and Krachler ( 2010 ) also used Pb isotopes to demonstrate that the fate of Pb from runoff can be 

## 23

different from Pb with different origins

. They observed that humus PM i mpacted by leaded on

- road 

## 24

gasoline that are derived from soil surfaces are likely to be more easily transferred to sediments th an Pb of 

## 25

other origins, with substantial amounts retained by lakes. 

## 26

Recent studies have examined the use of Pb isotope ratios as a tool for source apportionment. 

## 27

Duzgoren

- Aydin and Weiss ( 2008 ) provide caveats for using isotope ratio analyses. They point out that Pb 

## 28

isotope ratios may vary when Pb from several sources of different geological origins are introduced to the 

## 29

same location. Duzgoren

- Aydin ( 2007 ) warned that the presence of a complex mixture of contaminants 

## 30

containing common Pb isotopes can lead to an overestimation of the contribution of one source (e.g., soil 

## 31

contaminated by Pb emissions from on

- road gasoline) and an underestimate of another sou rce, such as 

## 32

that from industry. For this reason, Cheng and Hu ( 2010 ) suggest that Pb isotope analysis only be used 

## 33May 20 11 3

- 20 DRAFT

– DO NOT CITE OR QUOTE 

### when the investigators are confident that the isotopic signatures of various sources differ substantially. Pb 

## 1

recycling and in ternational trading may cause more blending of Pb from various sources, so that there is 

## 2

less heterogeneity in the Pb isotopic signatures sampled. Additionally, Cheng and Hu ( 2010 ) point out that 

## 3

the isotopic signature of Pb in air or soil ma y change over time with changing source contributions, but 

## 4

historical Pb isotope data are lacking. Duzgoren

- Aydin and Weiss ( 2008 ) suggest the use of GIS mapping 

## 5

of Pb isotopic information to help distinguish potential sources based on location of sources in addition to 

## 6

the sources’ isotopic signature. 

## 7

Gulson et al. ( 2007 ) examined the relationships between Pb isotope ratios and source 

## 8

apportionment metrics at urban and rural sites in New South Wales, Australia. In this study, Gulson et al. 

## 9

( 2007 ) performed source apportionment with both principal component analysis (PCA) and a neural 

## 10

ne twork technique called the self

- organizing map (SOM) and compared results from each method with 

## 11

206 

### Pb/ 204 

### Pb, 207 

### Pb/ 206 

### Pb, and 208 

### Pb/ 206 

### Pb obtained from PM samples, although only 206 

### Pb/ 204 

### Pb results were 

## 12

presented in detail. Wintertime “fingerprints” from b oth the PCA and SOM methods produced similarly 

## 13

linear relationships with 206 

### Pb/ 204 

### Pb, with linearly decreasing relationships between the isotope ratios and 

## 14

the “secondary industry,” “smoke,” “soil,” and “seaspray” source categories. However, the relationsh ips 

## 15

of the isotope ratios with the SOM fingerprints and PCA factors, respectively, were very similar. This 

## 16

finding may have been due to the presence of elements such as black carbon and sulfur in several SOM 

## 17

fingerprints and PCA factors. The authors sugges t that this might be related to the presence of several 

## 18

sources, which in combination result in a weak atmospheric signal. Additionally, both PM 2.5 

### and TSP 

## 19

samples were utilized for this study, and it was found that similar results were obtained for either size cut. 

## 20

At the urban site, they observed that the 206 

### Pb/ 204 

### Pb ratio decreased over time with increasing 

## 21

contributions of industrial, soil, smoke, and sea spray sources. For the most part, these sources were not 

## 22

substantial contributions to Pb

- PM 2.5 

### for the rural site. As for the Tan et al. ( 2006 ) speciation study 

## 23

described above, no notable differences were observed between the size fractions with regard to isotopic 

## 24

signature. 

## 25



# 3.3. Fate and Transport of Lead

There are multiple routes of exposure to Pb, including direct exposure to atmospheric Pb, e xposure 

## 26

to Pb deposited in other media after atmospheric transport, and exposure to Pb in other media that does 

## 27

not originate from atmospheric deposition. As a result, an understanding of transport within and between 

## 28

media such as air, surface water, soil, and sediment is necessary for understanding direct and indirect 

## 29

impacts of atmospheric Pb as well the contribution of atmospheric Pb to total Pb exposure. Figure 3

- 8 

## 30

describes relevant Pb transport pathways through environmental media discussed in this ch apter and their 

## 31

relationship to key environmental exposure pathways for which some or all of the Pb is processed through 

## 32May 20 11 3

- 21 DRAFT

– DO NOT CITE OR QUOTE 

### the atmosphere. This discussion includes new research on atmospheric transport of Pb, atmospheric 

## 1

deposition and resuspension of Pb, Pb transport in surface waters and sediments, and Pb transport in soil. 

## 2

Figure 3

- 8. Fate of atmospheric lead. Media through which Pb is transported and deposited are shown in bold.



# 3.3.1. Air

The 2006 Pb AQCD ( U.S. EPA, 2006 ) concluded that Pb was primarily prese nt in submicron 

## 3

aerosols, but that bimodal size distributions were frequently observed. Pb

- PM in the fine fraction is 

## 4

transported long distances

, found in remote areas

, and can be modeled using Gaussian plume models and 

## 5

Lagrangian or Eulerian continental t ransport models a s reported by several studies . G ood agreement 

## 6

between measurements and these models have been reported. Historical records of atmospheric 

## 7

deposition to soil, sediments, peat, plants, snowpacks, and ice cores have provided valuable informat ion 

## 8

on trends and characteristics of atmospheric Pb transport

. Numerous studies using a variety of 

## 9

environmental media indicated a consistent pattern of Pb deposition peaking in the 1970s, followed by a 

## 10

more recent decline . These findings indicated that th e elimination of leaded gasoline for motor vehicles 

## 11

has not only led to lower atmospheric concentrations in areas impacted by vehicles (Section 3.5) , but a 

## 12

pervasive pattern of decreasing atmospheric Pb deposition and decreasing concentrations in other 

## 13

environmental media even at great distances from sources. 

## 14

The 2006 Pb AQCD ( U.S. EPA, 2006 ) documented that soluble Pb was mostly removed by wet 

## 15

deposition

, and most of the insoluble Pb was mostly removed b y dry deposition . As a result, dry 

## 16May 20 11 3

- 22 DRAFT

– DO NOT CITE OR QUOTE 

### deposition was the major removal mechanism for Pb in coarse PM ( which is mainly insoluble ) and wet 

## 1

deposition as the most important removal mechanism for fine PM and Pb halides ( which were more 

## 2

soluble ). Numerous studies reported that Pb dry deposition velocities in the U

. S

. were mostly within a 

## 3

range of 0.05 to 1.0 cm/sec and dry deposition f luxes ranging from 0.04 to 4 mg/ m 2



### · y r. Precipitation 

## 4

concentrations ranged mostly from 0.5 to 60 μg/ L, but with considerably lower c oncentrations in remote 

## 5

areas, and wet deposition fluxes in the United States ranged from 0.3 to 1.0 mg/ m 2



### · y r. Wet deposition was 

## 6

linked to precipitation intensity, with slow even rainfalls usually depositing more Pb than intense rain 

## 7

showers

. Rain concent rations decreased dramatically between the early 1980s and the 1990s, reflecting 

## 8

the overall decreasing trend in Pb emissions due to elimination of leaded motor vehicle gasoline. A 

## 9

summary of studies investigating total deposition including both wet and dr y deposition indicated typical 

## 10

deposition fluxes of 2

- 3 mg/ m 2



### · y r and dry to wet deposition ratios ranging from 0.25 to 2.5

. Seasonal 

## 11

deposition patterns can be affected by both variations in local source emissions and vegetation cover, and 

## 12

as a result a consistent seasonal pattern across studies has not been observed, although there have been 

## 13

only a few investigations. 

## 14

The 2006 Pb AQCD ( U.S. EPA, 2006 ) concluded that resuspension by wind and traffic contribute 

## 15

to airborne Pb near sources

. Pb in resuspended road dust exhibited a bimodal size distribution, but mass 

## 16

was predominantly associated with coarse PM. The Pb fraction in resu spended dust ranged from 0 .002 to 

## 17

0.3%, with the highest fractions observed for paved road dust and lowest for agricultural soil. 

## 18

3.3.1.1. Transport

New research on long range transport as well as transport of Pb in urban areas has advanced the 

## 19

understanding of Pb transport in the atmosphere. While the 2006 Pb AQCD described long range Pb 

## 20

transport as essentially a process of submicron PM transport ( U.S. EPA, 2006 ) , much of the recent 

## 21

research on Pb transport has focused on interactions between anthropogenic and coars er geogenic PM that 

## 22

leads to incorporation of Pb into coarse PM as well as subsequent transformation on exposure to mineral 

## 23

components of coarse PM. Using scanning electron microscopy (SEM), Schleicher et al. ( 2010 ) observed 

## 24

interactions of anthropogenic soot and fly ash particles on the surfa ces of coarse geogenic mineral 

## 25

particles and concluded that toxic metals were often associated with coarse PM. Murphy et al. ( 2007 ) 

## 26

found that PM released from wild fires and transported over long distances scavenge d and accumulate d 

## 27

Pb and sul f ate through coagulation with small Pb rich PM during transport and that Pb was associated 

## 28

with PM over a wide size range . Erel et al. ( 2006 ) also found that Pb enrichment factors calculated for 

## 29

PM from dust storms collected in Israel were much greater than those sam pled at their north African 

## 30

source , suggesting that the dust samples had picked up pollutant Pb in transit between the Saharan desert 

## 31

and Israel. Marx et al. ( 2008 ) characteri z ed dust samples collected from the surface of glaciers and in dust 

## 32May 20 11 3

- 23 DRAFT

– DO NOT CITE OR QUOTE 

### traps on the remote west coast of New Zealand ’ s South Island and observed that m ost of the dust samples 

## 1

were enriched in metals

, including Pb

, compared w ith their source area sediments. 

## 2

Pb accumulated on mineral dusts is also subject to atmospheric transformations. PbSO 4 is one of 

## 3

the main constituents of Pb

- containing aerosols resulting from coal combustion ( Gieré et al., 2006 ) and it 

## 4

has been shown to react with calcite, CaC O 3 

### , a PM mineral component, to form Pb 3 

### (CO 3 

### ) 2 

### (OH) 2 

### , Pb(CO 3 

### ) 

## 5

and Ca( SO 4 

### ) 2 

### · H 2 

### O on the surface of the calcite ( Falgayrac et al., 2006 ) . In laboratory experiments, 

## 6

( Ishizaka et al., 2009 ) also showed that PbSO 4 

### could be converted to PbC O 3 

### in the presence of water. 

## 7

Approximately 60

- 80% was co nverted after only 24 hours for test samples immersed in a water droplet. 

## 8

This compared with only 4% conversion for particles that had not been immersed . As a result of recent 

## 9

research, there is considerable evidence that appreciable amounts of Pb can accu mulate on coarse PM 

## 10

during transport, and that the physical and chemical characteristics of Pb can be altered by this process 

## 11

due to accompanying transformations. 

## 12

Transport and Dispersion Mechanisms in Urban Environments

Several major U.S. sources of Pb e missions are located in urban areas. The urban environment can 

## 13

be considered unique because it has been highly modified by human activity

, includ ing above

- and below



## 14

ground infrastructure, buildings, and pavement

, and a high density of motorized transporta tion. This 

## 15

section focus es on special features of urban environments and upon processes that influence the 

## 16

distribution and redistribution of Pb

- bearing PM. 

## 17May 20 11 3

- 24 DRAFT

– DO NOT CITE OR QUOTE Source: Used with permission from Annual Reviews, Fernando ( 2010 ) 

### Figure 3

- 9. Scales of turbulence within an urban environment. Top: multiple scales within the atmospheric boundary layer. Bottom: illustration of airflow recirculation within a single street canyon located in the urban canopy layer.May 20 11 3

- 25 DRAFT

– DO NOT CITE OR QUOTE 

### As shown in Figure 3

- 9, urban turbulence occurs on several scales. Transport and dispersion of 

## 1

urban grit is subject to air movement within the urban canopy layer, where air movement is driven by air 

## 2

velocity within the urban boundary layer and urban topographical conditions such as bu ilding shape, 

## 3

building façade, and street canyon aspect ratio ( Fernando, 2010 ) . Within a street canyon, air circulates and 

## 4

tends to form counter

- rotating eddies along the height of the canyon (see Figure 3

- 9), which result in 

## 5

low er mean components of air movement, higher turbulence components, and higher sheer stress within 

## 6

the canyon compared with open field conditions ( Bri tter & Hanna, 2003 ; Kastner

- Klein & Rotach, 2004 )

. 

## 7

Recirculation around intersection corners and two

- way traffic conditions can also enhance turbulence 

## 8

levels, while one

- way traffic conditions increase air velocity along the street ( Kastner

- Klein et al., 2003 ; 

## 9

Kastner

- Klein et al., 2001 ; Soulhac et al., 2009 ) . All of these factors have the potential to influence human 

## 10

exposure to atmospheric Pb in urban areas with substantial Pb emissions. 

## 11

3.3.1.2. Deposition

Wet Deposition

T he 2006 Pb AQCD ( U.S. EPA, 2006 ) documented that dry deposition was the major removal 

## 12

mechanism for Pb in coarse PM and wet deposition as the most important removal mechanism for fine 

## 13

PM. Which process is most important for atmospheric removal of metals by deposition is largely 

## 14

controlled by solubility in rain water. Metal solubility in natural waters is determined by a complex 

## 15

multicomponent equilibrium between metals and their soluble complexes and insolu ble ionic solids 

## 16

formed with hydroxide, oxide, and carbonate ions. This equilibrium is strongly dependent on pH and 

## 17

ionic composition of the rain water. Recent research confirms the general trend described in the 2006 Pb 

## 18

AQCD ( U.S. EPA, 2006 ) that Pb associa ted with fine PM is usually more soluble in rain water than Pb 

## 19

associated with coarse PM , leading to a relatively greater importance of wet deposition for fine Pb and of 

## 20

dry deposition for coarse Pb . This could also explain the greater importance of dry de position near 

## 21

sources because coarse mode PM makes a greater contribution to PM mass. Although recent observations 

## 22

are consistent with these trends they also indicate considerable spatial and seasonal variability. Birmili et 

## 23

al. ( 2006 ) found that Pb solubility varied between the two main Pb

- containing size fractions, <0.5 μm 

## 24

(~40%) and 1.5

- 3.0 μm (~10%), indica tive of a different chemical speciation . However, t he observation 

## 25

that the amount of soluble Pb was higher in the ir U.K. samples than in an analytically identical study 

## 26

carried out in Seville, Spain ( Fernandez Espinosa et al., 2004 ) , led them to conclude that Pb solubility in 

## 27

fine PM may vary on a regional basis ( Birmili et al., 2006 ) . For PM 10 

### from Antarctica, 90 to 100% of the 

## 28

Pb was insoluble at the beginning of the summer season (November), but by the end of the summer 

## 29

(January), approximately 50% was soluble. Most of the Pb was from long range transport ( Annibaldi et 

## 30

al., 2007 ) . These studies illustrate the variable nature of a tmospheric Pb solubility. 

## 31May 20 11 3

- 26 DRAFT

– DO NOT CITE OR QUOTE 

### Dry Deposition

New measurements of Pb dry deposition fluxes are similar to those reported in the 2006 Pb AQCD 

## 1

( U.S. EPA, 2006 ) , except in industrialized urban areas, where it is considerably greater. Yi et al. ( 2006 ) 

## 2

calculated dry deposition fluxes for trace elements including Pb i n New York

- New Jersey harbor and 

## 3

observed much greater dry deposition fluxes for an urban industrial site than for New Brunswick. This is 

## 4

consistent with similar observations of dry deposition fluxes that were more than ten times greater in 

## 5

urban Chicago t han in rural South Haven, Michigan ( Paode et al., 1998 ) . These results illustrate the 

## 6

strongly localized nature of atmospheric Pb deposition in source rich areas. Elements from anthropogenic 

## 7

sources, including Pb, were generally associated with fine PM

. In a study of Tokyo Bay ( Sakata & 

## 8

Asakura, 2008 )

, reported an average dry deposition velocity of 1.06 cm/sec, which is at the upper end of 

## 9

dry deposition velocities reported in the 2006 AQCD ( U.S. EPA, 2006 ) . Th ey also reported that dry 

## 10

deposition fluxes were greater in industrially impacted urban areas, ranging from 12

- 17 μg/m 2



### · yr, more 

## 11

than 10 times the upper bound of the range reported in the 2006 Pb AQCD ( U.S. EPA, 2006 )

. 

## 12

Recent results also confirmed the tren d of decreasing overall deposition fluxes after removal of Pb 

## 13

from on

- road gasoline, as described in the 2006 Pb AQCD ( U.S. EPA, 2006 ) . Watmough and Dillon 

## 14

( 2007 ) found that the bulk annual deposition of Pb in a central Ontario forested watershed during 2002

- 

## 15

20 03 was 0.49 mg/ m 2



### · y r ; this was lower than the value of 1.30

- 1.90 mg/ m 2



### · y r for 1989

- 91 and 

## 16

represented a 75% decline in Pb deposition. It was consistent with the decline more generally observed 

## 17

for the N ortheastern U.S. as a consequence of the restrictions to alkyl

- Pb additives in on

- road gasoline. 

## 18

From previously published work , and in agreement with the precipitation data described above, most of 

## 19

the decline took place before the start of the Watmough and Dillon ( 2007 ) study. 

## 20

Several important observations can be highlighted from the few studies of atmospheric Pb 

## 21

deposition carried out in the past several years. Deposition fluxes have greatly declined since the removal 

## 22

of P b additives from on

- road gasoline. However, new results in industrial areas indicate that local 

## 23

deposition fluxes there are much higher than under more typical conditions. In general, wet deposition 

## 24

appears to be more important for Pb in fine PM, which is relatively soluble; and dry deposition appears to 

## 25

be generally more important for Pb in coarse PM, which is relatively insoluble. However, the relative 

## 26

importance of wet and dry deposition is highly variable with respect to location and season, probably 

## 27

re flecting both variations in Pb speciation and variations in external factors such as pH and rain water 

## 28

composition. Although industrial Pb emissions are mainly associated with fine PM, and wet deposition is 

## 29

likely to be more important for this size range, a substantial amount of Pb is apparently removed near 

## 30

industrial sources. 

## 31May 20 11 3

- 27 DRAFT

– DO NOT CITE OR QUOTE 

## 3.3.1.3. Resuspension of Lead from Soil to Air after Lead Deposition

As described in Section 3.2, the greatest Pb emissions in the United States occur in locations near 

## 1

major specific point s ources, including airports, secondary smelters, and other industrial operations 

## 2

involving large scale metal processing or fuel combustion. However, in the absence of such sources and in 

## 3

the vicinity of previous major sources, t he 2006 Pb AQCD ( U.S. EPA, 2006 ) concluded that resuspension 

## 4

by wind and traffic can be a substantial source of airborne Pb above background levels near sources , with 

## 5

resuspended dust accounting for between 0 .002 to 0.3 % of PM mass. Since then, results from several 

## 6

studies have provided support for a substantial contribution from resuspension by indicating a smoothed 

## 7

soil Pb concentration profile that decreases with distance from various sources, including city centers 

## 8

( Laidlaw & Filippelli, 2008 )

, major freeways ( Sabin, Lim, Venezia, et al., 2006 )

, and steel structures with 

## 9

abrading paint ( Weiss et al., 2006 )

. The smoothed profile is consisten t with continual Pb resuspension and 

## 10

deposition due to atmospheric turbulence. Recent Pb speciation results also indicate a substantial 

## 11

contribution from resuspended soils in areas with previous major emission sources, but without current 

## 12

major sources. Da ta from airborne PM in the vicinity of an inactive smelter in El Paso, TX were 

## 13

consistent with Pb

- humate as the major form of Pb in airborne PM, suggestive of soil resuspension since 

## 14

the local near

- surface soils had high humic content ( Pingitore et al., 2009 )

. 

## 15

Recent research on urban PM transport is also highly relevant to Pb transport and dispersion 

## 16

because Pb is most prevalently particle

- bound. Relevant results for Pb exposure in these areas include 

## 17

observations that PM concentration peaks dissipate more rapidly on wider streets than in narrow street 

## 18

canyons ( Buonanno et al., 2011 ) ; concent ration s are typically low next to a building because either less 

## 19

source material is available or less material penetrates the boundary layer of the building ( Buonanno et 

## 20

al., 2011 ) ; and there are stronger inverse relationship between mean wind speed and PM concentration 

## 21

fluctuation intensities at middle sections of urban street blocks compared with intersections ( Hahn et al., 

## 22

2009 )

. Patra et al. ( 2008 ) conducted experiments in London, U.K. in which a “tracer” grit was applied to a 

## 23

road and then the grit’s dispersion by traffic was measured over time to simulate resuspension and 

## 24

transport of a trace metal such as Pb. During the experiments, 0.039% of the tracer grit was measured to 

## 25

move down the road with each passing vehicle, 0.0050% was estimated t o be swept across the road with 

## 26

each passing vehicle, and 0.031% was estimated to become airborne when a vehicle passed. 

## 27

New resuspension studies complement previous research indicating street dust half

- lives on the 

## 28

order of one

- hundred days ( Allott et al., 1989 ) , with resuspension and street run

- off as major sinks 

## 29

( Vermette et al., 1991 ) as well as observations of a strong influence of street surface pollution on 

## 30

resuspension ( Bukowiecki et al., 20 10 )

, observations of greater resuspension of smaller PM than coarser 

## 31

PM ( Lara

- Cazenave et al., 1994 ) , leading to enrichment of metal concentrations in resuspended PM 

## 32

relative to street dust ( Wong et al., 2006 ) and observations of wind speed , wind direction

, vehicular 

## 33

traffic, pedestrian traffic, agricultural activities, st reet sweeping and construction operations as important 

## 34May 20 11 3

- 28 DRAFT

– DO NOT CITE OR QUOTE 

### factors determining resuspension. Together these results demonstrate that under in the vicinity of previous 

## 1

major emission sources and the absence of current major sources, resuspension can make a subs tantial 

## 2

contribution to atmospheric Pb concentrations. 

## 3



# 3.3.2. Water

As described in the 2006 Pb AQCD ( U.S. EPA, 2006 ) , atmospheric deposition has been identified 

## 4

as the largest source of Pb in surface waters, but urban runoff and industrial discharge are also impor tant. 

## 5

Water columns have been described as transient reservoirs with Pb residence times in lakes typically 

## 6

several months long , and shorter residence times expected in turbulent waterways

. Because dispersal in 

## 7

waterways is a relatively rapid process, conce ntrations in surface waters are highest near sources of 

## 8

pollution before substantial Pb by flushing, evaporation and sedimentation

. Transport in surface water is 

## 9

largely controlled by exchange with sediments, and the cycling of Pb between water and sedimen ts is 

## 10

governed by chemical, biological, and mechanical processes that are affected by many factors, including 

## 11

salinity, organic complexation, oxidation

- reduction potential, and pH . As described in the 2006 Pb AQCD 

## 12

( U.S. EPA, 2006 ) , m etals in waterways are tr ansported primarily as soluble chelates and ions, or adsorbed 

## 13

on colloidal surfaces, including secondary clay minerals, iron and manganese oxides or hydroxides, and 

## 14

organic matter , and adsorption on organic or inorganic colloids is particularly important f or Pb. The extent 

## 15

of sorption is strongly depends on particle size as smaller particles have larger collective surface areas

. 

## 16

Aqueous Pb concentrations also increase with increasing salinity

. Pb is found predominantly as PbO or 

## 17

PbC O 3 

### in aqueous ecosystems

. Pb is relatively stable in sediments, with long residence times and limited 

## 18

mobility

. However, Pb

- containing sediment particles can be remobilized into the water column

. As a 

## 19

result trends in sediment concentration tend to follow those in overlying waters . Fe and Mn oxides are 

## 20

especially susceptible to recycling with the overlying water column

. Although resuspension of sediments 

## 21

into overlying waters is generally small compared to sedimentation, resuspension of contaminated 

## 22

sediments is often a more import ant source than atmospheric deposition

. Organic matter (OM) in 

## 23

sediments has a high capacity for accumulating trace elements

. In an anoxic environmental removal by 

## 24

sulfides is particularly important

. 

## 25

Although atmospheric deposition was identified as the l argest source of Pb in surface waters in the 

## 26

2006 Pb AQCD ( U.S. EPA, 2006 )

, runoff from storms was also identified as an important source. A 

## 27

substantial portion of Pb susceptible to runoff is originates from atmospheric deposition

. The 2006 Pb 

## 28

AQCD ( U.S. EPA, 200 6 ) concluded that important contributors to Pb in dust on roadways included 

## 29

vehicle wear, vehicle emissions, road wear, fluid leakage, and atmospheric deposition

. Runoff from 

## 30

buildings due to paint, gutters, roofing materials and other housing materials were also identified as major 

## 31

contributors to Pb in runoff waters

. Investigations of building material contributions indicated runoff 

## 32

concent rations ranging from 2 to 88 mg/ L, with the highest concentrations observed from more than 10

- 

## 33May 20 11 3

- 29 DRAFT

– DO NOT CITE OR QUOTE 

### year

- old paint and t he lowest concentrations from residential roofs

. There was some indication that Pb 

## 1

from roofing materials, siding, and piping could be due to dissolution of Pb carbonate ( cerussite ) or 

## 2

related compounds

. In several studies Pb in runoff was consistently mos tly PM , with a relatively small 

## 3

dissolved fraction

. Runoff release was dependent on storm intensity and length of dry periods between 

## 4

rain events, with greater runoff of Pb associated with more intense storms and with longer periods 

## 5

between rain events

. Several studies indicated a “ first flush effect ,” with highest runoff concentrations 

## 6

observed at the beginning of a rain event. 

## 7

3.3.2.1. Lead Transport in Water and Sediment

Recent publications provide additional detail regarding Pb adsorption on iron rich and orga nic rich 

## 8

colloids. Correlation between Pb concentration in unfiltered water with total Fe was observed ( Hassellov 

## 9

& von der Kammer, 2008 ) , which is consistent with previous research using cross flow filtration 

## 10

( Pokrovsky & Schott, 2002 ; Ross & Sherrell, 1999 ) and SEM examination of single particles ( Taillefert et 

## 11

al., 2000 )

. 

## 12

Two distinct colloidal phases , one o rganic

- rich (0.5

- 3 nm in dia meter) and the other Fe

- rich (>3 nm 

## 13

in diameter) , have been observed to coexist in both soil isolates and river water ( Stolpe & Hassellov, 

## 14

2007 ) . Pb was observ ed to be predominantly associated with Fe

- oxide PM in river water but also 

## 15

associated with the organic colloids in the soil isolates ( Hassellov & von der Kammer, 2008 )

. 

## 16

Investigation of Pb binding onto ferrihydrite showed Pb binding data were consistent with Pb being held 

## 17

at the surface by sorption processes , rather than enclosed within the particle structure ( Hassellov & von 

## 18

der Kammer, 2008 )

. 

## 19

Observations in boreal rivers and soil pore waters in permafrost dominated areas of Central Siberia 

## 20

indicated t hat Pb was transported w ith colloids in Fe

- rich waters. T race elements that normally exhibited 

## 21

limited mobility ( including Pb ) had 40

- 80% of their annual flux in the nominal dissolved phase, 

## 22

operationally defined as material that passes through a 0.45 μm p ore

- size filter, and that these metals had 

## 23

a higher affinity for organo

- mineral Fe

- Al colloids ( Pokrovsky et al., 20 06 ) . Pokrovsky et al. ( 2006 ) 

## 24

postulated that during the summer, rainwater interacts with degrading plant litter in the top soil leading to 

## 25

the formation of Fe

- Al

- organic colloids with incorporated trace elements

. Migration of trace element

- Fe



## 26

Al

- OM colloids may result in export of Pb and other elements to riverine systems. Most of the transport 

## 27

occurred after thawing had commenced. This cont rasts with permafrost free areas where trace elements 

## 28

such as Pb are incorporated into iron colloids during OM

- stabili z ed Fe

- oxyhydroxide formation at the 

## 29

redox boundary of Fe(II)

- rich waters and surficial DOC

- rich horizons. Similarly, d uring a spring floo d 

## 30

(May) that exported 30

- 60% of total annual dissolved and suspended flux of elements including Pb

, Pb 

## 31

was mainly in the nominal dissolved phase, operationally defined as material that passes through a 

## 32

0.45 μm pore

- size filter ( Pokrovsky et al., 2010 ) . This was likely due to the presence of organic

- bound 

## 33May 20 11 3

- 30 DRAFT

– DO NOT CITE OR QUOTE 

### colloids smaller than 0.45 μm rather than true Pb dissolution ( Pokrovsky e t al., 2010 ) . Pb adsorbed on 

## 1

colloidal surfaces rather than incorporated into particle structure is likely to be more readily dissolved 

## 2

because dissolution of the entire particle is not required. 

## 3

Recent research on retention of Pb in water bodies and se diments has focused on the estuarine and 

## 4

marine environment, where considerable retention of Pb was observed in estuarine sediments. For a large 

## 5

riparian system, the Trinity River, Texas, Warnken and Santschi ( 2009 ) found that 80% of riverine Pb was 

## 6

retained in Lake Livingston , an estuarine region, while an additional 16% was removed to estuarine 

## 7

sediments , and only about 4% eventually reached the ocean

. Geochemical (sorption by Fe 

## 8

oxyhydroxides), biological (seasonal uptake by sinking algae in Lake Livingston) and hydrological 

## 9

(dilution effects by increasing flow rates) processes were mainly responsible for controlling dissolved 

## 10

trace metal concentrations rather than pollution sources. 

## 11

Overall, recent resea rch on Pb transport in aquatic systems has provided a large body of 

## 12

observations confirming that Pb transport is dominated by iron and organic rich colloids. In addition, new 

## 13

results indicated that although the 2006 Pb AQCD ( U.S. EPA, 2006 ) described rivers and lakes as 

## 14

temporary reservoirs with Pb lifetimes of months or less, estuaries can present a substantial barrier to 

## 15

transport into the open ocean. 

## 16

3.3.2.2. Deposition of Lead within Bodies of Water and in Sediment

As described in the 2006 Pb AQCD ( U.S. EPA, 2006 )

, in general Pb is relatively stable in 

## 17

sediments, with long residence times and limited mobility. As described in previous sections, Pb enters 

## 18

and is distributed in bodies of water largely in PM form. In rivers, particle

- bound metals can often 

## 19

account for ≥ 75% of the total load , e.g. ( Horowitz & Stephens, 2008 ) . Urbani z ed areas tend to have 

## 20

greater aquatic Pb loads, as several studies have shown the strong positive correlation between population 

## 21

density and river or lake sediment Pb concentrations ( Chalmers et al., 2007 ; Horowi tz et al., 2008 )

. 

## 22

Indeed, Chalmers et al. ( 2007 ) revealed tha t in river and lake sediments in New England, there was an 

## 23

order of magnitude difference between Pb sediment concentrations in rural versus urbani z ed areas

. 

## 24

The fate of Pb in the water column is determined by the chemical and physical properties of the 

## 25

water (pH, salinity, oxidation status, flow rate and the suspended sediment load and its constituents, etc). 

## 26

Desorption, dissolution, precipitation, sorption an d complexation processes can all occur concurrently and 

## 27

continuously, leading to transformations and redistribution of Pb. The pH of water is of primary 

## 28

importance in determining the likely chemical fate of Pb in terms of solubility, precipitation or organ ic 

## 29

complexation. In peatland areas, such as those in upland areas of the U.K. , organic acids draining from 

## 30

the surrounding peat land s can lower stream water pH to below 4. Under these conditions, Pb

- PM can be 

## 31

desorbed and released into solution, leading to elevated dissolved Pb concentrations ( Rothwell et al., 

## 32

2008 ) . At the other end of the pH scale, Pb tends to remain or become complexed, prec ipitated or sorbed 

## 33May 20 11 3

- 31 DRAFT

– DO NOT CITE OR QUOTE 

### to TSP , as observed by Das et al. ( 2008 ) who studied trace metal geochemistry in a South African lake 

## 1

with water pH of 9. They als o found marked differences in Pb concentrations associated with increasing 

## 2

depth in the water column [ e.g. , the surface Pb

- PM concentration of 2 μg / L increased to 60 μg/ L at depth 

## 3

and the Pb concentration in the <0.45 μm fraction increased from 2 μg/ L at the surface to 19 μg/ L at 

## 4

depth ( Das et al., 2008 ) ] . This is suggestive of a settlement process in action. 

## 5

In estuarine and wider ma rine environments the processes may be more complex because of the 

## 6

additional perturbation caused by tidal action and the strong effects of salinity. Again, PM forms of Pb are 

## 7

important in determining Pb distribution and behavior. Li et al. ( 2010 ) reported that PM Pb accounted for 

## 8

85 ± 15% and 50 ± 22% in Boston Harbor and Massachus etts Bay, respectively, while Lai et al. ( 2008 ) 

## 9

reported a solid (acid soluble):dissolved Pb ratio of 2.6 for areas of t he Australian sector of the Southern 

## 10

Ocean

. 

## 11

The a ccurate modeling of Pb behavio r in marine waters (including estuaries) requires consideration 

## 12

of many parameters such as hydrodynamics, salinity, pH, suspended PM , fluxes between PM and 

## 13

dissolved phases ( Hartnett & Berry, 2010 )

. Several new advances in the study of Pb cycling in these 

## 14

complex environments have been described in recent publicat ions. Li et al. ( 2010 ) used partic le organic 

## 15

carbon (POC) as a surrogate for the primary sorption phase in the water column to describe and model the 

## 16

partitioning of Pb between PM and dissolved forms . Huang and Conte ( 2009 ) observed that considerable 

## 17

cha nge in the composition of PM occurs as they sink in the marine environment of the Sargasso Sea, with 

## 18

mineraliz ation of OM resulting in increased PM

- Pb concentration with increased depth

. As a result of this 

## 19

depletion of OM in sinking particles, geochemical behavio r at depth was dominated by inorganic 

## 20

processes, e.g. adsorption onto surfaces, which were largely independent of Pb source. Sinking rates in 

## 21

marine environments can vary, but a rate approximating 1 m / d ay has been used in some models of Pb 

## 22

transpor t and distribution in aquatic

- sediment systems ( L. Li et al., 20 10 )

. Surface sediment Pb 

## 23

concentrations for various continental shelves were collated and compared by Fang et al. ( 2009 ) ; see 

## 24

Table 3

- 2

. 

## 25

Table 3

- 2

. Surface sediment Pb concentrations for various continental shelves ; see Fang et al. ( 2009 ) and references therein.

Location Digestion solution Pb ( mg/kg)

East China Sea HCl/HN O 3 /HF 10

- 49 (27) a Mediterranean, Israel coast HN O 3 9.9

- 20 Aegean Sea HCl/HN O 3 /HF 21

- 44 (34) Banc d ’ Arguin, Mauritania HCl/HN O 3 /HF 2.8

- 8.9 Campeche shelf, Gulf of Mexico HCl/HN O 3 0.22

- 20 (4.3) Laptev Sea, Siberia HCl/HN O 3 /HF 12

- 22 Pechora Sea, Russia Not reported 9.0

- 22 (14)

a Values in parentheses are the average, where calculableMay 20 11 3

- 32 DRAFT

– DO NOT CITE OR QUOTE 

## 3.3.2.3. Flux of Lead from Sediments

Sediments can be either a source or a sink for metals in the aquatic environment. Release can be via 

## 1

re

- suspension of the sediment bed via wind, wave and tidal action or by dissolution from sediment to the 

## 2

water column. When external Pb inputs to bodies of water are decreased by environmental improvement 

## 3

actions or regulations, contributions of Pb to the water column from the existing sediments can become an 

## 4

increasingly important source. ( J. L. Roulier et al., 2010 ) determined that Pb flux from sediments 

## 5

originated mostly from organic fractions

, but also partially from Mn and Fe components undergoing 

## 6

reductive dissolution. The rate of release was controlled by OM conten t, particle size, clay type and 

## 7

content

, and silt fraction ( J. L. Roulier et al., 2010 ) . The importance of sediment particle size, OM content 

## 8

and acid volatile sulf ide concentration in relation to metal release was similarly identified ( Cantwell et al., 

## 9

2008 ) . The effect of pH change on Pb release from lake sedi ments has also been examined, revealing that 

## 10

1.8 protons (H + 

### ) were exchanged per divalent metal cation released ( G. Lee et al., 2008 )

. P rocesses 

## 11

governing Pb release from lake sediments, including microbial reductive dissolution of Fe, biogenic 

## 12

sul f ide production and metal sorption

- desorption, have been investigated and results indicated that release 

## 13

of Pb from sub

- oxic an d anoxic zones of the sediment act as a Pb source to the overlying water of the lake 

## 14

( Sengor et al., 2007 )

. 

## 15

Sediment r esuspension from marine environments is similarly important, with disturbance of bed 

## 16

sediments by tidal action in estuarine areas resulting in a general greater capacity for re

- suspension of 

## 17

PM . Benthic fluxes of dissolved metals released from sediments me asured in Boston Bay were calculated 

## 18

as strong enough that in the absence of Pb inputs such benthic flux would reduce sediment Pb 

## 19

concentrations in Boston Bay to background levels in 30

- 60 years ( L. Li et al., 2010 ) . In a related way, a 

## 20

half

- life for sediment Pb (considering benthic flux alone as the loss mechanism) of 5.3 years was 

## 21

estimated for marine sediments off the Belgian coast ( Gao et al., 2009 )

. 

## 22

Radakovitch et al. ( 2008 ) investigated the riverine transport of PM including Pb to the Gulf of 

## 23

Lion, France, and also concluded that a major part of annual fluxes could be delivered over a short time 

## 24

period. From budget calculations, riverine inp uts were more important than atmospheric deposition and 

## 25

Pb concentrations in the prodelta sediments showed a strong correlation with OM content. These 

## 26

sediments, however, were not considered to be a permanent sink

, as resuspension in these shallow areas 

## 27

wa s an important process. OM, Pb and other metals were enriched in resuspended PM compared with the 

## 28

sediment. 

## 29

Birch and O’Hea ( 2007 ) reported h igher total suspended solid s, turbidity and total water metal 

## 30

concentration in surface compared with bottom water as well as a difference in suspended PM metal 

## 31

concentrations between surface water and bottom sediments , demonstrating that stormwater discharge 

## 32

was the dominant process of metal transfer during high rainfall events. T otal suspended sediments ( and 

## 33

total water metals ) in bottom water were higher than in the surface water plume

, indicat ing that 

## 34May 20 11 3

- 33 DRAFT

– DO NOT CITE OR QUOTE 

### resuspension of bottom sedime nt i s a greater contributor of total suspended sediments than stormwater 

## 1

during such events, especially in shallower regions of the bay. Soto

- Jimenez and Páez

- Osuna ( 2010 ) 

## 2

determined diffusive and advective fluxes, geochemical partitioning of Pb and Pb

- isotopic signatures in a 

## 3

study of mobility and behavior of Pb in hypersaline salt marsh sediments. They determined that sulfides 

## 4

were the main scavengers for Pb that was diagenetically released Pb

. 

## 5

Overall, recent research on Pb flux from sediments in natural waters provided greater detail on 

## 6

resuspension processes than was available in the 2006 Pb AQCD ( U.S. EPA , 2006 ) , and has demonstrated 

## 7

that resuspended Pb is largely associated with OM or Fe and Mn particles, but that anoxic or depleted 

## 8

oxygen environments in sediments play an important role in Pb cycling. This newer research indicated 

## 9

that resuspension an d release from sediments largely occurs during discrete events related to storms. It 

## 10

has also confirmed that resuspension is an important process that strongly influences the lifetime of Pb in 

## 11

bodies of water. Finally, there have been important advances in understanding and modeling of Pb 

## 12

partitioning in complex aquatic environments. 

## 13

3.3.2.4. Lead in Runoff

Runoff is a major source of Pb in surface waters. This complicates any evaluation of the 

## 14

contribution of atmospheric Pb to surfaces waters, which must take into account direct atmospheric 

## 15

deposition, runoff of atmospherically deposited Pb, and runoff of Pb from sources such as mine tailings or 

## 16

paint chips that are deposited from the atmosphere. The 2006 Pb AQCD ( U.S. EPA, 2006 ) identified 

## 17

important contributors to P b pollution in dust associated with roadways , such as vehicle wear, vehicle 

## 18

emissions, road wear, fluid leakage, and atmospheric deposition That review identified contributors to 

## 19

r unoff from buildings , such as paint, gutters, roofing materials and other ho using materials. The 2006 Pb 

## 20

AQCD ( U.S. EPA, 2006 ) also concluded that runoff was consistently mostly PM , with a relatively small 

## 21

dissolved fraction , and that dissolution of carbonate and related compounds were important contributors 

## 22

to Pb polluti on in runoff waters. It also described runoff Pb release into runoff as dependent on storm 

## 23

intensity and length of dry periods between rain events, and a “ first flush effect ,” with highest runoff 

## 24

concentrations observed at the beginning of a rain event . Su bsequent research has provided considerable 

## 25

new information about roadway and urban runoff and snow melt. 

## 26

Severe contamination due to export of anthropogenic Pb to adjacent ecosystem s via sewage 

## 27

systems (urban runoff and domestic wastewater) and to a lesse r extent by direct atmospheric deposition 

## 28

has been documented ( Soto

- Jiménez & Flegal, 2009 )

. Recent investigations also confirm roof runoff as an 

## 29

imp ortant contributor to Pb pollution. Huston et al. ( 2009 ) measured Pb concentrations in water from 

## 30

urban rainwater tanks and found Pb concentrations in bulk deposition were consistently lower than in 

## 31

water in the rainwater tanks , but that s ludge in the tanks had a high Pb content, indicating that not all 

## 32

major sources of Pb are from atmospheric deposition . Pb levels frequently exceeded drinking water 

## 33May 20 11 3

- 34 DRAFT

– DO NOT CITE OR QUOTE 

### standards. Pb flashing on the roofs was implicated as the source of Pb in the rainwater tanks although 

## 1

other possible sources include old paint and Pb stabili z ed PVC drain pipes ( Al

- Malack, 2001 ; Lasheen et 

## 2

al., 2008 ; Weiss et al., 2006 )

. 

## 3

New research has improved the understanding of sus pended PM size ranges, speciation, and 

## 4

impacts of Pb runoff from urban soil and road dust. Soil and road dust have been identified as major 

## 5

sources of Pb pollution to near

- coastal waters , leading to high Pb concentrations in stormwater runoff that 

## 6

became a ssociated with dissolved and suspended PM phases as well as bedload, material moved by 

## 7

rolling, sliding, and saltating along the bottom of a stream ( Birch & McCready, 2009 )

. 

## 8

Several new studies reported that the size distribution of PM transported in runoff is relatively 

## 9

uniform. Characteri z ation of the roadside dust in Australia showed that Pb in PM was approximately 

## 10

uniformly distributed among PM size fractions of up to 250 μm . The Pb

- containing particles had the 

## 11

potential to be dispersed to some distance into sensitive ecosystems ( Pratt & Lottermoser, 2007 )

. Pb in 

## 12

roadside dusts in Thessaloniki, Greece was characteri z ed by Ewen et al. ( 2009 ) and no difference in Pb 

## 13

concentration was found between <75 μm and 75

- 125 μm PM size ranges, although a difference in the 

## 14

chemical form of Pb between slightly versus highly contaminated areas was observed

. 

## 15

Ewen et al. ( 2009 ) reported that Pb was mainly in a more exchangeable form (similar to that in an 

## 16

old auto

- catalyst reference mat erial) in small particles, but in the residual , or least mobile fraction in 

## 17

larger particles

. In urban road dust from Manchester U.K., Pb

- bearing Fe

- oxides were observed to be 

## 18

dominant in most of the size fractions, and PbCrO 4 

### comprised 8

- 34% of total Pb w ith the highest 

## 19

concentrations being found in the largest and smallest size fractions. Pb(C O 3 

### ) 2 

### and Pb(OH) 2 

### were 

## 20

measured in the two middle size fractions whilst PbO and PbSO 4 

### w ere present in the largest and smallest 

## 21

size fractions ( Barrett et al., 2010 )

. 

## 22

Murakami et al. ( 2007 ) also emphasized the importance of PbCrO 4 

### as an important species of Pb 

## 23

from road surfaces using , identified i ndividual particles containing high levels of Pb and Cr (≥ 0.2%)

, 

## 24

most likely from the yellow road line markings. The identifi ed PM constituted 46% of Cr and Pb in heavy 

## 25

traffic dust and 7

- 28% in dust from residential areas and soakaway sediments. The presence of such 

## 26

particles in soakaway sediments is consistent with their low environmental solubility. 

## 27

Recent research also continues to document the first flush effect described in the 2006 Pb AQCD. 

## 28

Flint and Davis ( 2007 ) reported that i n 13% of runoff events, more than 50% of Pb was flushed in the first 

## 29

25% of event water. A second flush occurred less frequently (4% of runoff events for Pb ). In agreement 

## 30

with the 2006 Pb AQCD ( U.S. EPA, 2006 ) , most recent studies have concluded that, during storm events, 

## 31

Pb is transpo rted together with large PM . Some studies, however, found that Pb was concentrated in the 

## 32

fine PM fraction and, occasionally, Pb was found predominantly in the dissolved fraction. Tuccillo ( 2006 ) 

## 33

found that Pb was almost entirely in the >5 μ m size range and, indeed, may be associated with PM larger 

## 34

than 20 μm

. ( J. Sansalone et al., 2010 ) compared Pb

- containing PM siz e distributions from New Orleans, 

## 35

LA; Little Rock, AR; North Little Rock , AR; and Cincinnati , OH and found no common distribution 

## 36May 20 11 3

- 35 DRAFT

– DO NOT CITE OR QUOTE 

### pattern

. Pb was associated with Cincinnati PM mainly in the <75 μm fractions, at Baton Rouge and Little 

## 1

Rock Pb mainly in the 75

- 425 μm PM fractions, and at North Little Rock Pb predominantly in the 

## 2

>425 μm PM fractions

. New Orleans Pb was almost uniformly distributed among the smaller size PM 

## 3

fractions. McKenzie et al. ( 2008 ) found that Pb was enriched in the finest PM (0.1

- 0.3 μm ) in stormwater 

## 4

samples collected in California, particularly for storms that occurred during and after an exten ded dry 

## 5

period. 

## 6

Guo et al. ( 2006 ) investigated the effect of engineered partial exfiltration reactor (PER) systems on 

## 7

the partitioning and speciation of Pb in rainfall

- runoff at the upstream end of an urban source area 

## 8

catchment that is part of the much larger urbani z ed and industrial Mill Creek watershed in Hamilton 

## 9

County, Ohio. The catchment is paved to a large extent with asphalt and is used for transportation. Guo et 

## 10

al. ( 2006 ) investigated a catchment that drain ed towards a wide grassy area and found that Pb was mainly 

## 11

associated with dissolved organic matter ( DOM ). The study su ggested that interaction of the rainfall



## 12

runoff with the grassy area may have resulted in removal of PM

- bound Pb and hence in the association of 

## 13

Pb with DOM . PM amount and size can also be influenced by the runoff surface. Guo et al. ( 2006 ) found 

## 14

that Pb entering the engineered PER system was mainly in the dissolved fraction with ~76%

. 

## 15

There were several recent observations of a relationship between road traffic volume and runoff Pb 

## 16

concentration, although a clear relationship was not always observed. At a relatively clean location, Desta 

## 17

et al. ( 2007 ) studied highway runoff characteristics in Ireland and found that although as expected, Pb was 

## 18

strongly correlated with TSP, no relationship between total suspended solids and rainfall, rain intensity, 

## 19

antecedent dry days or runoff event duration were observed, and t raffic volume also did not appear to 

## 20

have an effect They conclud ed that runoff composition from site to site could be highly variable. Most 

## 21

other studies, however, did find a relationship between traffic volume and Pb concentration. A California 

## 22

study of highway runoff by Kayhanian et al. ( 2007 ) reported that 70

- 80% Pb was in PM form for both 

## 23

non

- urban and urban highways , and that t he concentration of Pb in runoff from low traffic flow (30,000

- 

## 24

100,00 0 vehicles / day) urban highways was 50% higher than that from non

- urban highways (mean = 

## 25

16.6 μg/ L). Additionally, t he concentrations in runoff from high traffic flow (>100,000 vehicles / day) 

## 26

urban areas were five times higher than those from non

- urban highw ays. Helmreich et al. ( 2010 ) 

## 27

characteri z ed road runoff in Munich, Germany, with an average daily traffic load of 57,000 vehicles. The 

## 28

mean Pb concentration, 56 μg/ L (maximum value = 405 μg/ L), lay in between the values for low traffic 

## 29

flow and high traffic flow runoff from urban areas in California, i.e.

, there was good agreement with 

## 30

Kayhanian et al. ( 2007 ) . There was no detectable dissolved Pb, i.e. 100% in PM form . Seasonal effects of 

## 31

highway runoff have also been observed recently. Hallberg et al. ( 2007 ) found that summer Pb 

## 32

concentrations in runoff water in Stockholm ranged from 1.37

- 47.5 μg/ L whil e , in winter, the range was 

## 33

1.06

- ~296 μg/ L

. T here was a strong correla tion between Pb (and most other elements) and total 

## 34

suspended solids ( R 2



### = 0.89) . Helmreich et al. ( 2010 ) also found higher metal concentrations during cold 

## 35

seasons in Stockholm but Pb concentrations increased only slightly during the snowmelt season. There 

## 36May 20 11 3

- 36 DRAFT

– DO NOT CITE OR QUOTE 

### was no change in the distribution of Pb between dissolved and PM forms for the rain and snowmelt 

## 1

periods. Runoff from urban snowmelt has been intensively investigated since the 2006 Pb AQCD was 

## 2

pub lished ( U.S. EPA, 2006 )

. The relocation of snow means that the area receiving the snowmelt is not 

## 3

necessarily the same area that which received the snowfall . Magill and Sansalone ( 2010 ) also noted that 

## 4

plowed snowbanks alongside roadways form a temporary linear reservoir for traffic generated 

## 5

constituents such as metals and PM

. Snowm elt concentrations of metals such as Pb can therefore be 

## 6

several orders of magnitude higher than those in rainfall runoff ( J. J. Sansalone & Buchberger, 1996 ) . The 

## 7

melt process usually occurs in a sequence: pavement melt, followed by roadside (impervious) and finally 

## 8

pervious area melt. As part of this sequence, rain

- on

- snow can transport high loads of PM

- associated 

## 9

pollutants ( Oberts, 2000 )

. Westerlund and Viklander ( 2006 ) investigated differences in PM and Pb 

## 10

concentrations between rainfall events occurring during snowmelt and rain periods. Runoff e vents 

## 11

occurring during the snowmelt period (i.e. rain

- on

- snow) had about five time s higher numbers of particles 

## 12

(in the size range 4 to 120 μm ) /liter of runoff. The first rain

- on

- snow event was characteri z ed by an 

## 13

increase in the number of particles in the 4 to 25 μm size range. The rain

- on

- snow gave a “ flush ” through 

## 14

the snow but this was still not sufficient to transport the larger sized particles. Only the highest energy 

## 15

rain

- on

- snow events increased transport of PM across the entire size spectrum. T here was no difference in 

## 16

particle size distributions between snowmelt and rain on sno w events, although more was transported 

## 17

during snowmelt . Pb concentrations were most strongly associated with the smaller PM size fractions

. 

## 18

Overall, there was a significant difference between the melt period and the rain period in terms of 

## 19

concentrations, loads, transportation and association of heavy metals with particles in different size 

## 20

fractions ( Westerlund & Viklander, 2006 )

. Over a 4

- year period, Mag ill and Sansalone ( 2010 ) analy z ed 

## 21

the distribution of metal in snow plowed to the edge of roads in the Lake Tahoe catchment in Nevada, and 

## 22

concluded that metals including Pb were m ainly associated with coarser PM (179

- 542 μm ). The PM



## 23

associated metal could be readily separated from runoff water ( e.g. , in urban drainage systems ) , but there 

## 24

is potential for leaching of metals from the PM within storage basins ( Ying & Sansalone, 2008 ) . For 

## 25

adsorbed species that form outer sphere complexes, a decrease in adsorption and an increase in aqueous 

## 26

complexes for pollutant metals is a likely consequence of higher deicing salt concentrations. If metals 

## 27

form inn er

- sphere complexes directly co ordinated to adsorbent surfaces, background deicing salt ions 

## 28

would have less impact. It is thought that physical and outer

- sphere compl exes predominate for coarse 

## 29

PM, as was the case in Nevada, and so leaching would be likely to cause an increase in dissolved phase 

## 30

Pb concentrations. 

## 31

Rural runoff has also been extensively studied since publication of the 2006 Pb AQCD ( U.S. EPA, 

## 32

2006 ) , inclu ding s everal recent publications on a forested watershed (Lake Plastic) in central Ontario 

## 33

( Landre et al., 2009

, 2010 ; Wa tmough & Dillon, 2007 ) and nearby Kawagama Lake, Canada ( Shotyk & 

## 34

Krachler, 2010 )

. Results indicated that bulk deposition substantially decreased to 0.49 mg/ m 2 

### in 2002 

## 35

from 1.30

- 1.90 mg/ m 2 

### in 1989

- 91. The upland soils retained > 95% of the Pb in bulk deposition, i.e. 

## 36May 20 11 3

- 37 DRAFT

– DO NOT CITE OR QUOTE 

### leaching losses to stream water were small. The wetland area was, however, a net source of Pb with 

## 1

annual Pb concentrations in stream water ranging from 0.38 to 0.77 μg/ L. L ake sediments were efficient 

## 2

sinks for atmos pherically deposited Pb with 80

- 91% of the Pb input being retained. Up to 68% of the Pb 

## 3

entering the lake was derived from the terrestrial catchment. Overall, the watershed effectively retain ed 

## 4

atmospherically deposited Pb

, but some Pb was then redistribut ed from the catchment to the lake 

## 5

sediments ; and the Pb in the near

- surface lake sediments reflect ed terrestrially transported soil material

, 

## 6

rather Pb being deposited from the atmosphere. The h ighest concentrations of dissolved organic carbon 

## 7

(DOC) , Fe an d Pb in the wetland draining stream occurred in summer when it frequently exceeded 1 μg/ L 

## 8

( Landre et al., 2009 )

. 

## 9

Graham et al. ( 2006 ) observed two temporally separated mechanisms occurring during storm 

## 10

events in a rural organic rich upland catchment . At the beginning of an event, Pb was transported together 

## 11

with large particles in the >25 μm size range , but after several hours Pb was mainly transported with 

## 12

colloidal or DOM (<0.45 μm ) , and t he remaining 30

- 40% of storm related Pb was transported in this 

## 13

form . This indicated that rapid overland flow rapidly transported Pb

- PM into the receiving streams at the 

## 14

very beginning of the event , and this was followed within a few hours by transport of organic

- colloidal Pb 

## 15

via near

- surface throughflow. The authors used a conservative estimate of Pb removal, based on their 

## 16

observations that the catchment was continuing to act as a sink for Pb. These observations about th e 

## 17

transport and fate of Pb agree well with those of Watmough and Dillon ( 2007 ) and Shotyk et al. ( 2010 )

. 

## 18

Soil type was also found to have a strong influence on runoff contributions. Dawson et al. ( 2010 ) 

## 19

found that for organic

- rich soils, Pb was mobili z ed from near

- surface soils together with DOC but for 

## 20

more minerogenic soils, percolation of water allowed Pb, bound to DOC, to be retai ned in mineral 

## 21

horizons and combine with other groundwater sources. The resulting Pb in stream water that had been 

## 22

transported from throughout the soil profile and had a more geogenic signature ( Dawson et al., 2010 ) . The 

## 23

findings of both Graham et al. ( 2006 ) and Dawson et al. ( 2010 ) were important because the provenance 

## 24

and transport mechanisms of Pb may greatly aff ect the net export to receiving waters, particularly since 

## 25

higher concentrations of previously deposited anthropogenic Pb are usually found in the near

- surface 

## 26

sections of upland U.K. soils ( e.g. , ( Farmer et al., 2005 ) )

. 

## 27

In another study Rothwell et al. ( 2007 ) observed stormflow Pb concentrations almost three times 

## 28

higher than those reported by Graham et al. ( 2006 ) for northeastern Scotland. The generally high 

## 29

dissolved Pb stores and high stream water DOC concentrations ( Rothwell, Evans, Daniels, et al., 2007 )

. 

## 30

In a separate study, Rothwell et al. ( 2007 ) showed that OM was the main vector for Pb transport in the 

## 31

fluvial system. Some seasonal variability was observed: declining Pb concentrations in autumn stormflow 

## 32

may indicate the exhaustion of DOC from the acrotelm (the hydrologically active upper layer of peat 

## 33

which i s subject to a fluctuating water table and is generally aerobic) or a dilution effect from an 

## 34

increasing importance of overland flow. 

## 35May 20 11 3

- 38 DRAFT

– DO NOT CITE OR QUOTE 

### Erosion of agricultural soils and the effects of different types of storm event s on soil particle and Pb 

## 1

losses from thes e soils was characteri z ed by Quinton and Catt ( 2007 ) . A close link between metal 

## 2

concentration and the silt, or clay and organic content of stream sediments was consistent with enrichment 

## 3

of m etals as a consequence of small erosion events. They also noted that short intense events could 

## 4

produce the same amount of sediment as longer low

- intensity events. More intense events, however, 

## 5

could mobili z e a wider range of particle sizes whereas low int ensity events mobili z ed finer but more 

## 6

metal

- rich material. Smaller events accounted for 52% of Pb losses from the agricultural soil. 

## 7

The Tinto River in Spain drains one of the largest polymetallic massive sul f ide regions in the 

## 8

world : the Iberian Pyrite B elt. Evaporitic sul f ate salts, formed as a result of acid mine drainage processes, 

## 9

are considered to be a temporary sink for many heavy metals. Upon the arrival of rainfall, however, they 

## 10

rapidly dissolve, releasing acidity and contaminant metals into rece iving waters. Thus rivers in semi

- arid 

## 11

climate regions such as the Tinto R iver which alternate between long periods of drought and short but 

## 12

intense rainfall events, can experience quick acidification and increases in metal concentration. In a study 

## 13

of suc h events, Cánovas et al. ( 2010 ) found that whil e many element concentrations decreased during 

## 14

events, the concentrations of Fe, Cr, Pb and As increased. This was attribute d to the redissolution and 

## 15

transformation of Fe oxyhydroxysulfates and/or desorption processes. 

## 16

Several investigators considered a Pb isotope study by Dunlap et al. ( 2008 ) of a large 

## 17

(>160,000 km 2 

### ) riparian system ( the Sacramento River, CA), which showed that the present day flux of 

## 18

Pb was dominated by Pb from historical anthropogenic sources, which included a mixture of high

- ratio 

## 19

hydraulic Au mining

- derived Pb and persistent historically

- derived Pb from leaded on

- road gasoline. 

## 20

Outside of the mining region, 57

- 67% Pb was derived from past on

- road gasoline emissions and 33

- 43% 

## 21

was from hydraulic Au mining sediment. The flow into the Sacramento River from these sources is an 

## 22

ongoing process. Periods of high surface runoff, however, mobili z e additional fluxes of Pb from these two 

## 23

sources and carry them into the river. These pulses of Pb, driven by rainfall events, suggest a direct link 

## 24

between local climate change and transport of toxic metals in surface waters ( Dunlap et al., 2008 )

. 

## 25

Rothwell et al. ( 2007 ) commented that although there have been substantial reductions in sul f ur 

## 26

deposition to U.K. uplands over the last few decades ( Fowler et al., 2005 ) , anthropogenic acidification of 

## 27

upland waters is likely to continue due to nitrogen leaching from the surrounding catchment and this may 

## 28

inc rease with nitrogen saturation ( Curtis et al., 2005 )

. Rothwell et al. ( 2007 ) predicted that if an increase 

## 29

in surface water acidification is coupled with further increases in DOC export from organic

- rich 

## 30

catchments, metal export from peatland systems will increase. The de terioration of peat soils by erosion is 

## 31

considered to be exacerbated by climatic change. Rothwell et al. ( 2010 ) used digital terrain analysis to 

## 32

model sus pended Pb concentrations in contaminated peatland catchments. The peat soils of the Peak 

## 33

District are characteri z ed by extensive eroding gullies and so they were combined in an empirical 

## 34

relationship between sediment

- associated Pb concentrations and mean u pslope gully depth with fine

- 

## 35May 20 11 3

- 39 DRAFT

– DO NOT CITE OR QUOTE 

### resolution mapping of the gully areas. This model will enable prediction of metal contamination in 

## 1

receiving waters. 

## 2

Klaminder et al. ( 2010 ) investigated the environmental recovery of sub

- arctic lakes in response to 

## 3

reduced atmospheric deposition over the last few decades. They found that there had been no 

## 4

improvement in surface sediments and indeed the reduction in Pb contamination had been much less than 

## 5

the 90% reduction in emissions over the last four decades. The weak improvement in the 206 

### Pb / 207 

### Pb ratio 

## 6

together with the Pb contaminant concentrations suggests that catchment export processes of previously



## 7

deposited atmospheric contaminants have had a considerable impact on the recent contaminant burden of 

## 8

sub

- arctic lakes. In Arctic regions, s oil export of contaminants to surface waters may dramatically 

## 9

increase in response to climate change if it triggers thawing of frozen soil layers. It is thought that 

## 10

thawing may generate accelerated soil erosion, altered hydrological flow paths, increased runoff and 

## 11

exposure of soluble compounds that had previously been in the frozen layers. At this stage, however, the 

## 12

links between catchment export and climate change have not yet been clearly established

. 

## 13

Coynel et al. ( 2007 ) also considered the effects of climate change on heavy metal transport. In this 

## 14

case, the scenario of flood

- related transport of PM in the Garonne

- Gironde fluvial

- estuarine syst em was 

## 15

investigated. Export of suspended PM during a five

- day flood in December 2003 was estimated at 

## 16

~440,000 t ons , accounting for ~75% of the annual suspended PM fluxes. Sediment remobili z ation 

## 17

accounted for ~42% of the total SPM flux during the flood ev ent (~185,000 t ons suspended PM) and 

## 18

accounted for 61% of the 51 t ons Pb that was exported. Coynel et al. ( 2007 ) postulate that flood hazards 

## 19

and t ransport of highly polluted sediment may increase as a result of climate change and/or other 

## 20

anthropogenic impacts (flood management, reservoir removal). 

## 21

In heavily contaminated catchments ( e.g. , that of the Litavka River, Czech Republic ( Zak et al., 

## 22

2009 ) ) , the flux of heavy metals to the river during storm events can be substantial. Even during a minor 

## 23

4

- day event, 2

, 954 kg of Pb was transported , and the majority was associated with suspended PM

. For the 

## 24

Adour River in a mountainous area of France, Pb pollution predominantly originated from mining 

## 25

activities, and Point et al. ( 2007 ) showed that 75% of annual soil fluxes into the river were transported in 

## 26

30

- 40 days. 

## 27

The consequences of flood management (dam flushing) practi c es on suspended PM and heavy 

## 28

metal fluxes in a fluvial

- estuarine system (Garonne

- Gironde, France) were considered by Coynel et al. 

## 29

( 2007 ) . Dam flushing enhanced mobili z ation of up to 30

- year

- old polluted sediment from reservoir lakes. 

## 30

Sediment remobili z ation accounted for ~42% of the total suspended PM fluxes during the flood and 

## 31

strongly contributed to PM

- bound metal transport (61% for Pb). They concluded that flood management 

## 32

will need to be taken into consideration in future models for erosion and pollutant transport. 

## 33

Bur et al. ( 2009 ) investigated the associations of Pb in stream

- bed sediments of the French 

## 34

Gasco ny region. They found that Pb enrichment in stream sediments was positively correlated with 

## 35

catchment cover and increasing organic content whereas Pb concentration was strongly linked with Fe

- 

## 36May 20 11 3

- 40 DRAFT

– DO NOT CITE OR QUOTE 

### oxide content in cultivated catchments. For the low

- OM , anthropogenic Pb was associated with 

## 1

carbonates and Fe

- oxides (preferentially, the amorphous fraction). Fe

- oxides became the most efficient 

## 2

anthropogenic Pb trapping component as soon as the carbonate content is reduced. They noted, however, 

## 3

that OM was always weakly involved. N’Guessan et al. ( 2009 ) also studied trace elements in stream

- bed 

## 4

sediments of the French Gasco ny region. They used enrichment factors to show that only ~20

- 22% of Pb 

## 5

was from anthropogenic sources with the remainder ori ginating from natural weathering processes. 

## 6

Overall, research results from the last several years have greatly expanded the extent of the 

## 7

knowledge concerning Pb from runoff

. Substantial Pb input to estuarine and marine ecosystems ha s been 

## 8

well documented

. More detail concerning the origin of Pb from roof runoff has led to the conclusion that 

## 9

roof flashing could be especially important. Research on road runoff has provided valuable insight into 

## 10

PM size and composition, indicating that size distributions fo r Pb

- containing PM in runoff water varies 

## 11

from study to study and from location to location, and that Pb is frequently associated with chromate near 

## 12

roads, probably from paint used to mark road lines. Recent studies confi rmed the “first flush” effect, 

## 13

rele asing more Pb at the beginning of rainfall than subsequently, and documented size distributions of Pb



## 14

containing PM also vary considerably when water from the first flush is isolated

. Influence of road traffic 

## 15

volume on runoff has also been more fully docu mented in recent years. The role of urban snowmelt and 

## 16

rain

- on

- snow events is also better understood, and it has been observed that greater runoff occurs from 

## 17

snowmelt and in rain on snow events than when snow is not present, and that metals, including Pb, are 

## 18

often associated with coarse PM under these circumstances

. Runoff in rural areas is strongly controlled by 

## 19

soil type and the presence of vegetation, with less runoff and greater retention in mineral soils or when 

## 20

grass is present, and more runoff for soils high in OM. Runoff also follows a two

- step process of transport 

## 21

of larger particles at the beginning of an event, followed within hours by transport of finer colloidal 

## 22

material

. Some initial research on the effects of climate change on runoff ha s foc used on documenting the 

## 23

association between increased runoff and more intense rain events and greater thawing

. Overall, recent 

## 24

research has provided greater detail on amounts, particle size distributions, composition, and important 

## 25

processes involving Pb t ransport, and the understanding of Pb runoff has become more complete since 

## 26

publication of the 2006 Pb AQCD ( U.S. EPA, 2006 )

. 

## 27



# 3.3.3. Soil

The 2006 Pb AQCD ( U.S. EPA, 2006 ) summarized that Pb h as a relative ly long retention time in 

## 28

the organic soil horizon, although its movement through the soil column also suggests potential for 

## 29

contaminat ion of groundwater

. Leaching was consistently observed to be a slower process for Pb than for 

## 30

other contaminants because Pb was only weakly soluble in pore water, but anthropogenic Pb is more 

## 31

available for leaching than natural Pb in soil. Pb can bind to many different surfaces and Pb sorption 

## 32

capacity is influenced by hydraulic conductivity, solid composition, OM content, clay mineral content, 

## 33May 20 11 3

- 41 DRAFT

– DO NOT CITE OR QUOTE 

### microbial activity, plant root channel s , animal holes, and geochemical reactions

. As a result of Pb binding 

## 1

to soil components, leaching is retarded by partitioning to soils, which is not only influenced by sorption 

## 2

capacity, but leaching also increases with proximity to source, increasing pH, and increasing metal 

## 3

concentrations

. Leaching is also strongly influenced by pore water flow rates, with more complete 

## 4

sorption contributing to slower leaching at lighter flows

. Leaching rates are especially high in soils with a 

## 5

high Cl content, but typic ally the most labile Pb fraction is adsorbed to colloidal particles that include 

## 6

OM , clay, and carbonates

. Transport through soils is enhanced by increasing amount of colloidal 

## 7

suspensions, increasing colloidal surface charge, increasing organic content of colloids, increasing 

## 8

colloidal macroporosity, and decreasing colloidal size . Acidity and alkalinity ha ve a more complex 

## 9

influence, with sorption maximized at neutral pH between pH = 5 and pH = 8.2, and greater mobility at 

## 10

higher and lower pH

. High Pb leve ls have been observed in leachates from some contaminated soils, but 

## 11

this effect appears to be pH dependent

. In several studies of contaminated soils a substantial fraction of Pb 

## 12

was associated was associated with Mn and Fe oxides or carbonate

. 

## 13

3.3.3.1. Deposition of Lead onto Soil from Air

As described in the 2006 Pb AQCD ( U.S. EPA, 2006 ) , a considerable amount of Pb has been 

## 14

deposited from air onto soils in urban areas and near stationary sources and mines, and soil Pb 

## 15

concentrations can reach several thousand mg/k g. Major sources in urban soil were identified as 

## 16

automotive traffic (prior to when leaded on

- road gasoline was phased out), and deteriorating Pb

- based 

## 17

paint, with the highest Pb concentrations observed where traffic and population density were the greates t. 

## 18

Highest concentrations were found in city centers and near roadways, and several studies reported 

## 19

concentrations falling off rapidly with distance and depth of soil layers near roads. High Pb soil 

## 20

concentrations were also observed near stationary source s such as smelters and battery disposal 

## 21

operations, also decreasing rapidly with distance from the source. Several recent studies continue to 

## 22

document high concentrations of Pb in soil. A study of soil Pb concentrations in Queensland, Australia 

## 23

described atmospheric transport and deposition of Pb in urban soils due to ongoing emissions from nearby 

## 24

mining and smelting activities are continuing to impact on the urban environment ( Taylor et al., 2010 )

. 

## 25

Similarly, s ediment cores from four remote Canadian Shield headwater lakes located along a transect 

## 26

extending 300 km from a non

- ferrous metal smelter generated useful in formation about distance of Pb 

## 27

transport from the smelter prior to deposition ( Gallon et al., 2006 )

. Shotyk and Krachler ( 2010 ) postulated 

## 28

that long

- range transport of Pb from a smelter at Rouyn

- Noranda may still contribute to deposition on 

## 29

these lakes. Recent measurements of deposition fluxes to soil in rural and remote areas have ranged from 

## 30

approximately 0.5 mg/m 2



### · yr to about 3 mg/m 2



### · yr with fair agreement between locations in Canada, 

## 31

Scandinavia, and Scotland and showed a substantial decrease compared to when leaded on

- road gasoline 

## 32May 20 11 3

- 42 DRAFT

– DO NOT CITE OR QUOTE 

### was in widespread use ( Fowler et al., 2006 ; Graham et al., 2006 ; Shotbolt et al., 2008 ; Watmough & 

## 1

Dillon, 2007 )

. 

## 2

Differences between throughfall and litterfall in forested areas have also been investigated i n 

## 3

forested areas, and the combined input of Pb to the forest floor from throughfall and litterfall was 

## 4

approximately twice th at measured in bulk deposition ( Landre et al., 2010 ) . The difference was attributed 

## 5

to a substantial contribution from internal forest cycling and indicates that bulk deposition collectors may 

## 6

underestimate the amount of Pb reaching th e forest floor by about 50% ( Landre et al., 2010 )

. 

## 7

There has been considerable interest in the response of soils to the decreasing aerosol Pb 

## 8

concentrations and Pb deposition rates that have been recorded in recent years. Kaste et al. ( 2006 ) 

## 9

resampled soils at 26 locations in the Northeast U.S. ( during a 2001

- 20 02 survey of soil sites originally 

## 10

sampled in 1980 )

, and found no significant change in the amount of Pb in the O

- horizon at high altit ude 

## 11

sites

. However, t he amount of Pb in the O

- horizon had decreased at some locations in the southern part of 

## 12

the survey region (Connecticut, New York, Pennsylvania), where the forest soils have typically thinner 

## 13

O

- horizon s , the reasons for which are discu ssed further in Section 3.3.3.2

. Higher Pb concentrations at 

## 14

greater altitudes were also found in Japan, especially above 600 m ( Takamatsu et al., 2010 )

. 

## 15

There is wide agreement that atmospheric deposition due to long

- range transport from industrial 

## 16

areas has been the major source of Pb to re mote surface soils over the past decades , e.g. ( Steinnes et al., 

## 17

2005 ) . However, another hypothes i s proposes that the gradual increase in Pb content in O

- horizon soil 

## 18

with changing latitude is attributable to “ plant pumping and organic binding ” rather than to atmospheric 

## 19

deposition ( Rasmussen, 1997 ; Reimann et al., 2008 ; Reimann et al., 2001 )

. 

## 20

Further support for the use of mosses as bioindicators o r monitors for atmospheric Pb inputs to peat 

## 21

bogs have recently been published by Kempter et al. ( 2010 ) who found that high moss productivity did 

## 22

not cause a diluti on of Pb concentrations , and that p roductive plants were able to accumulate more 

## 23

particles from the air and that r ates of net Pb accumulation by the mosses were in excellent agreement 

## 24

with the fluxes obtained by direct atmospheric measurements at nearby mo nitoring stations. In addition, 

## 25

Bindler et al. ( 2008 ) used Pb isotopes to compare the distribution of Pb in the forest soils with that of lake 

## 26

sediments where no “ plant pumping ” processes could be invoked , and u sed Pb isotope ratios to 

## 27

demonstrate that observations were consistent with anthropogenic Pb deposition to the soils rather than 

## 28

intermixing of natural Pb from underlying mineral soil horizons. 

## 29

Overall, recent studies provided deposition data that was con sistent with deposition fluxes reported 

## 30

in the 2006 Pb AQCD ( U.S. EPA, 2006 ) , and demonstrated consistently that Pb deposition to soils has 

## 31

decreased since the phase

- out of leaded on

- road gasoline. Additional research highlighted the importance 

## 32

of taking for est cycling and litter throughput account in estimating input by deposition. Follow

- up studies 

## 33

in several locations indicated little change in soil Pb concentrations since the phase

- out of leaded on

- road 

## 34

gasoline, consistent with the high retention reporte d for Pb in soils. Finally, although there has been 

## 35

considerable discussion of plant pumping as an alternative hypothesis for explaining the increases in soil 

## 36May 20 11 3

- 43 DRAFT

– DO NOT CITE OR QUOTE 

### concentrations, much evidence is more consistent with atmospheric deposition as the explanation f or 

## 1

observed increases in soil Pb concentration. 

## 2

3.3.3.2. Sequestration of Lead from Water to Soil

The 2006 Pb AQCD described Pb as being more strongly retained in soil than other metals because 

## 3

of its weak solubility in pore water, but that anthropogenic Pb was mor e available for leaching than 

## 4

natural Pb ( U.S. EPA, 2006 ) . It also described a complex variety of factors that influence Pb retention, 

## 5

including hydraulic conductivity, solid composition, OM content, clay mineral content, microbial activity, 

## 6

plant root chann els, geochemical reactions , colloid amounts, colloidal surface charge, and pH. 

## 7

Recent research in this area has provided more insight into the details of the Pb sequestration 

## 8

process. Importance of leaf litter was further investigated, and it was observed that t he absolute Pb 

## 9

content can be substantial because rain events cause splashing of the leaf litter with soil thus placing the 

## 10

litter in direct contact with soil metals. The resulting increase in leaf litter metal concentrations suggest s 

## 11

that the litter can act as a temporary sink for metals from the soil around and below leaves on the ground

. 

## 12

The low solubility of Pb in the leaf litter indicates that the Pb is tightly bound to the decomposing litter

, 

## 13

making the decomposing leaves act as an efficient metal storage pool ( Scheid et al., 2009 )

. 

## 14

New research has also provided details about the complexity of Pb sequestration during soil OM 

## 15

decomposition. Schroth et al. ( 2008 ) investigated Pb sequestration in the surface layer of forest soils and 

## 16

the transformation of Pb speciation during soil OM decomposition. The pH range for forest floor soils in 

## 17

the Northeast U.S. is typically 3.5

- 5 and, under these conditions, dissolved Pb would adsorb strongly to 

## 18

soluble OM and to Fe/Al/Mn oxides and oxyhydroxides

. It had been thought that the high affinity of Pb 

## 19

for organic ligands meant that sequestered atmospheric Pb would be preferentially bound to soluble OM. 

## 20

As a consequence, decomposition of OM would lead to Pb migrati on to the underlying mineral layers 

## 21

where it would be precipitated with the dissolved OC or adsorbed to pedogenic mineral phases. However, 

## 22

recent research has revealed a more complicated picture of gasoline

- derived Pb associations in the forest 

## 23

floor . More recent research indicates that, a s decomposition progresse s , Pb and Fe bec o me more 

## 24

concentrated in “ hotspots ” and Pb likely bec omes increasingly distributed on surfaces associated with Fe 

## 25

and Mn (and to some extent Ca)

. It was postulated that Pb was initi ally bound to labile organic but, 

## 26

following decomposition, the Pb was adsorbed at reactive si tes on pedogenic mineral phases ( Schroth et 

## 27

al., 2008 )

. Differences in litter types were also reported, with m ore rapid decomposition of OM in high 

## 28

quality deciduous litter mobilizing more Pb initially bound to labile OM than coniferous litter, and 

## 29

produc ing more pedog enic minerals that could readily sequester the released Pb ( Schroth et al., 2008 ) . In 

## 30

the next stage of the study, the speciation of Pb in the O

- horizon soils of Northern Hardwood, Norway 

## 31

s pruce and r ed p ine forest soils w ere compared. In general there was good agreement between the Pb 

## 32

speciation results for the soils and those for the laboratory decompositi on experiments. Specifically, for 

## 33May 20 11 3

- 44 DRAFT

– DO NOT CITE OR QUOTE 

### the Northern Hardwood forest soil , a little more than 60% of the Pb was bound to SOM and this 

## 1

percentage increased to ~70% and ~80% for the Norway s pruce and r ed p ine soils, respectively. In all 

## 2

three cases, however, most of the remainder of the Pb was bound to ferrihydrite rather than to birnessite. 

## 3

This was not considered to be surprising because of the well

- known leaching and cycling behavior of Mn 

## 4

that would be expected in the natural system. Thus the prevalence of Mn p hases in the field based 

## 5

samples would be lessened ( Schroth et al., 2008 )

. 

## 6

More generally, other studies have observed Pb sorption to Mn and Fe phases in soils. For example, 

## 7

Boonfueng et al. ( 2006 ) investigated Pb sequestration on Mn oxide

- coated montmorillonite. Pb formed 

## 8

bidentate corner

- sharing complexes. It was found that Pb sorption to MnO 2 

### occurred even when MnO 2 

## 9

was present as a coating on other minerals, e.g.

, montmorillonite. Although their importance in the near



## 10

surface p hases has clearly been demonstrated by Schroth et al. ( 2008 ) , ferrihydrite surfaces may not be a 

## 11

long

- term sink for Pb since reduc tive dissolution of this Fe ( III ) phase may release the surface

- bound Pb 

## 12

into the soil solution. Sturm et al. ( 2008 ) explored the fate of Pb during dissimilatory Fe reduction. Pb 

## 13

was indeed released but was then incorporated into less reactive ph ases. These phases could not, however, 

## 14

be identified. Even so, it was asserted that Pb should be largely immobile under Fe

- reducing conditions 

## 15

due to its incorporation into refractory secondary minerals. 

## 16

Kaste et al. ( 2006 ) found that Pb species currently in the O soil horizons in the Northeast U.S. 

## 17

differed considerably from those that were originally deposited from fossil fuel combustion (including on



## 18

road gasoline). PbSO 4 

### was considered to be the main form of Pb that had been delivered from the 

## 19

atmosphere to the surface of the Earth and it was post ulated that the presence of sul f ate may have 

## 20

facilitated the adsorption of Pb to colloidal Fe phases within the organic

- rich horizons. 

## 21

Altogether, these new results enhance the understanding of Pb sequestration in forest soils. First, 

## 22

the role of leaf lit ter as a major Pb reservoir is better understood. Second, the effect of decomposition on 

## 23

Pb distribution and sequestration on minerals during OM decomposition has been further characterized, 

## 24

and finally, the relative importance of Mn and Fe in sequestratio n is better understood. 

## 25

Recent research also addressed roadsides soils . Jensen et al. ( 2006 ) found that Pb was retained by 

## 26

an organic

- rich blackish deposit with a high OM content and elevated soil Pb concentrations, originating 

## 27

from total suspended solids in road runoff and from aerial deposition. Hossain et al. ( 2007 ) observed that 

## 28

after long dry periods, OM oxidation may potentially result in the release of Pb. Microbial activity may 

## 29

also breakdown OM and have similar consequences (i.e., Pb release). Bouvet et al. ( 2007 ) investigated the 

## 30

effect of pH on retention of Pb by roadside soils where municipal solid waste incineration (MSWI) 

## 31

bottom ash had been used for road construction. They found that the Pb that had leached from the road 

## 32

construction materials was retained by the proximal soils under the prevailing enviro nmental conditions 

## 33

(at pH = 7, <2% was released

, but a t pH = 4, slightly more Pb (4

- 47%) was released ) and the authors 

## 34

speculated that the phase from which Pb had been relea sed may have been Pb(CO 3 

### ) 2 

### (OH) 2 

### , indicating that 

## 35May 20 11 3

- 45 DRAFT

– DO NOT CITE OR QUOTE 

### sequestration of Pb via formation of oxycarbonat e minerals is only effective at near

- neutral to alkaline pH 

## 1

values ( Figure 3

- 10 in Section 3.3.3.3)

. 

## 2

Other recent research on Pb sequestration focused on microbial impacts and soil amendments. 

## 3

There have been few if any previous observations of microbial sequestration of Pb in soil. Perdrial et al. 

## 4

( 2008 ) observed bacterial Pb sequestration and proposed a mechanism of Pb complexation by 

## 5

polyphosphate. They also postulated that bacterial tr ansport of Pb could be important in sub

- surface soil 

## 6

environments. Wu et al. ( 2006 ) also and concluded that Pb adsorption to the bacterial cell walls may be 

## 7

imp ortant with respect to Pb transport in soils . This new area of research provides important evidence that 

## 8

bacteria can play an important role in both sequestration and transport of Pb. Phosphate addition to 

## 9

immobilize Pb

- contaminated soils has often been us ed to immobili z e Pb in situ through the formation of 

## 10

Pb phosphate minerals such as chloropyromorphite. Recent research investigated factors affecting the 

## 11

long

- term stability of such products , which depends on the equilibrium solubility and the dissolution rate 

## 12

of the mineral , trace impurities, such as Pb(OH) 2 

### , the presence of complexing agents, and pH ( Xie & 

## 13

Giammar, 2007 ) . Overall, in agreement with the 2006 Pb AQCD ( U.S. EPA, 2006 ) , the addition of 

## 14

phosphate can enhance immobili z ation of Pb under certain conditions in the field but may cause 

## 15

desorption and mobili z ation of anionic species of As, Cr and Se. 

## 16

3.3.3.3. Movement of Lead within the Soil Column

The 2006 Pb AQCD summarized studies that demonstrated that Pb has a long retention time in the 

## 17

organic soil horizon, it also has some capacity to leach through the soil column and contaminate 

## 18

groundwater more than other contaminants do, because Pb is only weakly soluble in pore water ( U.S. 

## 19

EPA, 2006 )

. The fate of any metal transport in soil is in response to a complex set of parameters including 

## 20

soil texture, mineralogy, pH and redox potential, hydraulic conductivity, abundance of OM and 

## 21

oxyhydroxides of Al, Fe, and Mn, in addition to climate, situation and nature of the parent material. As a 

## 22

consequence, it is impossible to make general conclusions about the final fate of anthropogenic Pb in 

## 23

soils. Indeed, Shotyk and LeRoux ( 2005 ) contend that the fate of Pb in soils may have to be evaluated on 

## 24

the basis of soil type. Some generalizations are, however, possible: Pb migration is likely to be g reater 

## 25

under acidic soil conditions ( Shotyk & Le Roux, 2005 ) . In this respect, it would be expected that th ere 

## 26

should be considerable mobility of Pb in the surface layers of c ertain types of forest soils. This section 

## 27

reviews recent research on movement of Pb through soil types by first focusing on forest soils, followed 

## 28

by a broader tre atment of a more diverse range of soils. 

## 29

Forest S oils and W etlands

Several studies confirmed the slow downward movement of Pb within the soil column. Kaste et al. 

## 30

( 2006 ) found that the amount of Pb in O

- horizon soils had remained constant at 15 of 26 sites in remote 

## 31May 20 11 3

- 46 DRAFT

– DO NOT CITE OR QUOTE 

### forested areas of the Northeast U.S. that had been re

- sampled after a 21

- year time period had elapsed , but 

## 1

that measured soil Pb concentrations were lower than predicted concentrations from total deposition, 

## 2

strongly suggesting that the O

- horizon had not retained all of the atmospheric Pb

, and that a proportion of 

## 3

the atmospheric deposition must have leac hed into the underlying mineral layers. At some sites, mainly 

## 4

those at the southern latitude s and lower altitude sites, the proportion of Pb that had been leached 

## 5

downward from the O

- horizon was quite considerable . Relative retention of Pb was influenced b y the rate 

## 6

of OM decomposition, depth of soil O

- horizon, and pH. For soils where Pb was strongly retained by the 

## 7

O

- horizon

, a relationship between Pb and Fe

- rich phase was observed, but Pb was also significantly 

## 8

correlated with other metals. XANES data sug gested a possible interaction with an amorphous Fe oxide

, 

## 9

but spectra were n o t entirely explained by Fe and oxygen and an additional spectral feature suggested the 

## 10

presence of a S or P atom , which could result if OM functional groups were binding to Pb. Ka ste et al. 

## 11

( 2006 ) concluded that a substantial fraction of Pb was associated with amorph ous Fe

- hydroxides. The 

## 12

strong binding of Pb coupled with the low solubility of Fe phases under oxic conditions

, helped to explain 

## 13

the relatively long residence time of gasoline

- derived Pb in forest floors which had thick O

- horizon s and 

## 14

were well

- drained. I n the situations where Pb was leached downward to a large extent, mobility was 

## 15

likely governed by OM decomposition and colloidal transport of Pb associated with colloidal Fe and OM. 

## 16

Klaminder et al. ( 2006 ) also considered the transfer of Pb from the O

- horizon to the underlying 

## 17

mineral horizons (including the C

- horizon ). They concluded that atmospheric pollution

- derived Pb 

## 18

migrated at a rate about 10

- 1

, 000 times slower than water. They assumed that Pb was mainly transported 

## 19

by dissolved OM and so the mean resid ence time of Pb in the O

- horizon depended on OM transport and 

## 20

turnover. The retardation rate was a reflection of the slow minerali z ation and slow downward transport 

## 21

rates of organic

- Pb complexes

, due to sorption and desorption reactions involving mineral s urfaces. 

## 22

In a study involving stable Pb isotopes, Bindler et al. ( 2008 ) showed that Pb with a different 

## 23

isotopic composition could be detected in the soil down to a depth of at least 30 cm and sometimes down 

## 24

to 80 cm in Swedish soils. In comparison, in N orth American podzols, pollution Pb is typically only 

## 25

identified to a depth of 10

- 20 cm (even with the aid of isotopes). This difference is attributed to the longer 

## 26

history of metal pollution in Europe (as has bee n traced using lake sediments )

. 

## 27

Several research groups have attempted to determine the mean residence time of Pb in the 

## 28

O

- horizon of forest soils. Klaminder et al. ( 2006 ) used three independent methods to estimate a mean 

## 29

residence time of about 250 years for Pb in the O

- horizon of boreal forests in Sweden , indicating that 

## 30

deposited atmospheric Pb pollution is stored in the near

- surface layers for a considerable period and, 

## 31

consequently, will respond only slowly to the reduction in atmospheric inputs. It should be noted, 

## 32

however, the OM in the upper parts of the O

- horizon is continually being replaced by fresh litter and the 

## 33

mean residence time of Pb in these horizons is only 1

- 2 years. Thus, the uppermost layer will respond 

## 34

more quickly than the rest of the O

- horizon to the decreases in Pb inputs. 

## 35May 20 11 3

- 47 DRAFT

– DO NOT CITE OR QUOTE 

### Klaminder et al. ( 2008 ) considered the biogeochemical behavio r of atmospherically derived Pb in 

## 1

boreal forest soils in Sweden (Figure 3

- 10) . The estimated annual losses via percolating soil water were 

## 2

~2 mg/ m 2



### · yr ( Klaminder, Bindler , & Renberg, 2008 ) and so the annual loss, assumed to be from the mor 

## 3

layer, was greater than the atmospheric input of ~0.5 mg/ m 2



### · yr . The upward transport of Pb did not 

## 4

compensate for the losses either. In contrast, the amount of Pb being stored in the mineral soil layers was 

## 5

increasing. The mean residence time of Pb in the mor layer was estimated to be ~300 years, in reasonable 

## 6

agreement with their earlier work ( Klaminder, Bindler, Emteryd, et al., 2006 ) . These values were greater 

## 7

than the values of 2

- 150 years determined for U.S. forest soils , e.g. ( J. Kaste et al., 2003 ; Watmough et 

## 8

al., 2004 ) but the difference was attributed to the lower decomposition r ates of OM within the northern 

## 9

boreal forests of Sweden. They concluded that more research was needed to determine the processes 

## 10

occurring within the mor layer that control the release of Pb from this horizon. 

## 11

Figure 3

- 10. Schematic model summarizing the estimated flux of Pb within a typical podzol profile from northern Sweden using data from
Klaminder et al. ( 2006 )

. The atmospheric deposition rate is
from ( Klaminder, Bindler, Emteryd, et al., 2006 ) , the plant

up take rates from ( Klaminder et al., 2005 ) and estimated soil

water fluxes from ( Klaminder, Bindler, Laudon, et al., 2006 )

.May 20 11 3

- 48 DRAFT

– DO NOT CITE OR QUOTE 

### Klaminder et al. ( 2008 ) investigated in more detail the distribution and isotopic signature of Pb 

## 1

persisted wit hin the O

- horizon (mor layer) of boreal forest soils. They found that the mor layer preserved 

## 2

a record of past Pb emissions from a nearby smelter. Minimal animal burrowing activity and low leaching 

## 3

rates observed at the sampling location were important fac tors contributing to the preservation of this 

## 4

record. They concluded that temporal changes in atmospheric fallout in addition to adsorption processes 

## 5

need to be considered when interpreting Pb concentrations changes within the mor layer. 

## 6

S ignificantly high er O

- horizon Pb concentrations have been observed in coniferous than deciduous 

## 7

forest soils ( McGee et al., 2007 )

. Steinne s et al. ( 2005 ) noted evidence for downward migration of Pb 

## 8

from the O

- horizon to the E

- horizon of most soils and in some cases the upper B horizon. They found that 

## 9

the downward transport of Pb differed considerably between the sites, e.g.

, from almost no anthropogenic 

## 10

Pb in the B

- horizon at some sites to ~70% at other sites. The greater downwards tr ansport in some 

## 11

locations was attributed to climatic variations, with more extensive leaching and possibly a greater 

## 12

turnover of OM at sites where higher mean annual temperatures were experienced. H igher atmospheric 

## 13

deposition of acidifying substances in t hese locations was considered the most important factor in Pb 

## 14

transport, causing release of Pb from exchange sites in the humus layer and promot ing downward 

## 15

leaching. 

## 16

Seasonal variation in Pb mobility has also been observed in forest soil. Other research indicated 

## 17

that Pb concentrations correlated with DOC concentrations in the soil solution from the O

- horizon, and 

## 18

were lower during late winter and spring compared with summer months ( Landre et al., 2009 ) . The 

## 19

degradation of OM in the O

- horizon produced high DOC concentrations in the soil solution. It was also 

## 20

shown that Pb was associated with the DOC, and concluded that DOC pr oduction is a primary factor 

## 21

enhancing metal mobility in this horizon. In the underlying mineral horizons, DOC concentrations 

## 22

declined due to adsorption and cation exchange processes. The B

- horizon retained most of the DOC 

## 23

leached from the O

- horizon and it has also been observed that Pb is similarly retained

. 

## 24

Non

- forested S oils

In contrast with forest soils, most non

- forested soils are less acidic an d so most studies of Pb 

## 25

behavio r in non

- forested soils have focu sed on Pb immobility . However, there are aci d soils in some 

## 26

locations that are not forested. For these soils, as for forest soils, Pb mobility is weak but correlated with 

## 27

OM. For example, Schwab et al. ( 2008 ) observed that low molecular weight organic acids added to soil 

## 28

enhanced Pb movement only slightly . C itric acid and tartaric acid enhanced Pb transport to the greatest 

## 29

degree but the extent of mobili z ation was only slightly higher than that attained using deioni z ed water 

## 30

even at high concentrations . Whil e the formation of stable solution comple xes and more acidic conditions 

## 31

favo red mobili z ation of Zn and Cd, Pb remained strongly sorbed to soil particles and so the presence of 

## 32

complexing agents and low pH (2.8

- 3.8) did not substantial ly enhance Pb mobility. Similarly, limited 

## 33May 20 11 3

- 49 DRAFT

– DO NOT CITE OR QUOTE 

### penetration and leac hing was observed in an extremely complex temperate soil profile , with h ighes t 

## 1

concentrations of Pb (~80 mg/k g) found in the top 0

- 5 cm section of soil. F or this uppermost soil section, 

## 2

there was a strong correlation between Pb concentration and OC content , both for the total soil fraction 

## 3

and the acid

- extractable fraction. The Pb migration rate was calculated to be 0.01 cm / y r and it was 

## 4

estimated that Pb would be retained in the soil column for 20,000 years , with no evidence of rapid 

## 5

movement of anthropoge nic Pb from the top 0

- 5 cm soil section into the soil profile Kylander et al. 

## 6

( 2008 )

. 

## 7

Other recent studies also reported strong retention on non

- forest soil s and enhanced mobility on Fe 

## 8

and OM colloids. Pb was strongly retained on acidic Mediterranean soil columns, with a ssociation of Pb 

## 9

with the exchangeable, OM and crystalline Fe oxide fractions appear ing to favo r mobility whil e 

## 10

association with Mn oxides a nd amorphous Fe oxides was linked with semi

- irreversible retention of Pb in 

## 11

the solid phase ( Garrido et al., 2008 ) . Pedrot et al. ( 2008 ) studied colloid

- mediated trace element release 

## 12

at the soil/water interface and showed that Pb was mobili z ed by Fe nanoparticles that were bound to 

## 13

humic acids. 

## 14

Soil pH value i s pr obably the single most important factor affecting solubility, mobility and 

## 15

phytoavailability but reducing conditions also result s in increased Pb mobility , with t he release of Pb into 

## 16

an anoxic soil solution due to the combined effect of Fe ( III ) reductive dissolution and dissolved OM 

## 17

release . Dissolved OM is more important than Fe oxyhydroxides in determining Pb mobility. Under oxic 

## 18

conditions, Fe

- Mn

- hydroxides often play an important role in the sorptio n of Pb to the solid phase soil 

## 19

( Schulz

- Zunkel & Krueger, 2009 ) . In an agricultural soil, fate of Pb in soils i s related to agricultural 

## 20

management. Although Pb was found to be strongly sorbed to the soil, downward migration was observed 

## 21

and t he movement of Pb to deeper soils was due to the soil mixing activities of earthworms ( Fernandez et 

## 22

al., 2007 ) . Thus in relatively unpolluted non

- forested soils, as in forested soils, colloidal Fe and OM, pH, 

## 23

and biophysical transport all enhance Pb mobility in soil. Pb transport in more highl y contaminated soils 

## 24

has also been the subject of recent research. I n a vegetated roadside soil

, Pb was leached from the upper 

## 25

50 cm of the soil even though the pH was 7.2. Pb was transported on mobile particles and colloids in the 

## 26

soil solution. Some of t he colloids may have formed from OM produced by roots and decaying shoots. 

## 27

The transport process was enhanced by preferential flow triggered by intense rainfall events. This study 

## 28

suggested that the value of the effective sorption coefficient estimated und er dynamic conditions was 

## 29

unrelated to values measured in conventional batch studies. This indicates that the use of batch studies to 

## 30

derive input values for sorption coefficients in transport models requires caution. It was concluded that 

## 31

the primary cont rol of Pb transport in the long term was the degree of preferential flow in the system ( S. 

## 32

Roulier et al., 2008 )

. 

## 33

Other studies also noted similarly low Pb mobility, but with substantial variation between soil types 

## 34

and locations. A decline in O

- horizon Pb concentrations and Pb accumulation in mineral horizons was 

## 35

also observed for forest soils by Watmough and Dillon ( 2007 ) , but did not hold for nearby wetland areas 

## 36May 20 11 3

- 50 DRAFT

– DO NOT CITE OR QUOTE 

### from which a large amount of DOC is exported, with approximately 10 times more Pb being associated 

## 1

with a given amount of DOC in the leachate from the LFH

- horizon of the wetland soil than with the DOC 

## 2

in the stream water draining the wetland. This may reflect greater retention of Pb by the wetland and/or a 

## 3

change in structure of DOC leading to a change in compl exing capacity possibly because of changes in 

## 4

pH or competition with Al and Fe. 

## 5

Williams et al. ( 2006 ) characteri z ed Pb speciation in a mine waste

- derived fertili z er , ironite . It was 

## 6

thought that PbS would be the main form of Pb, but instead was the predominant form was PbSO 4 

### , which 

## 7

may move more easily through soil and enter proximal waters. In contrast, Courtin

- Noma de et al. ( 2008 ) 

## 8

showed that Pb was incorporated into barite rather than goethite in waste rock pile materials. The high



## 9

stability phase formed was an anglesite

- barite solid solution. 

## 10

I n weathering flotation residues of a Zn

- Pb sul f ide mine were more Pb was mobile in weathered 

## 11

topsoil than in the unweathered subsoil. The topsoil had a very high OM content and the Pb enrichment 

## 12

was att ributed to an interaction with soil OM . Overall, the results contrast strongly with most other 

## 13

studies but the interpretation was supported by the sequential extraction results which showed that there 

## 14

was a very large exchangeable Pb component in these sur face soils ( Schuwirth et al., 2007 ) . Scheetz and 

## 15

Rimstidt ( 2009 ) characteri z ed shooting range soils in Jefferson National Forest, VA

, in which t he metallic 

## 16

Pb shot rapidly became corroded and developed a coating of hydrocerussite , which dissolved at the pH 

## 17

values of 8

- 9; see Figure 3

- 11, which shows an Eh

- pH diagram indicating the solubility, equilibrium, and 

## 18

stability of these corroded Pb molecules in terms of the activity of hydrogen ions (pH) versus the activity 

## 19

of electrons (Eh [in volts]). T he solub i li z ed Pb was largely re

- adsorbed by the Fe and Mn oxides and 

## 20

ca rbonate soil fractions. The minimum solubility of hydro cerussite lies in the pH range 8

- 9 but solubility 

## 21

increases by several orde rs of magnitude at pH below 6 ( Scheetz & Rimstidt, 20 09 )

. 

## 22May 20 11 3

- 51 DRAFT

– DO NOT CITE OR QUOTE Source: Used with permission from Elsevier Publishing, Scheetz and Rimstidt ( 2009 ) 

### Figure 3

- 11. Eh

- pH diagram for Pb in shooting range soils, Jefferson National Forest, VA.

Rooney et al. ( 2007 ) also investigated the controls on Pb solubility in soils contaminated with Pb 

## 1

shot. Again, corrosion crusts were found to develop on Pb pellets. The concentrations of Pb in the soil 

## 2

solution were, however, mu ch lower than if they were controlled by the solubility of the dominant crustal 

## 3

Pb compounds (mainly hydro cerussite ). Instead it was suggested that the concentrations were being 

## 4

controlled by sorption of Pb by the soil solid phase

. The pH range in this stu dy was 4.5

- 6.5 and so again 

## 5

dissolution of hydroceru s site would be expected. Sorption to solid phases in the soil is also consistent 

## 6

with the findings of Scheetz and Rimstidt ( 2009 )

. Overall, in contrast to less polluted forested and non



## 7

forested soils, considerable mobility was often, but not always observed in soils near roadways and mines 

## 8

and on shooting ranges, with colloid transport and soil pH playing an important role in Pb mobility

. 

## 9

Although there hav e been steep declines in Pb deposition, surface soils in have been slow to recover 

## 10

( Bindler e t al., 2008 ; J. M. Kaste et al., 2006 ) . As was concluded in the 2006 Pb AQCD ( U.S. EPA, 2006 )

, 

## 11

soils continue to act as a predomi nant sink for Pb. 

## 12

While in some studies the flux of Pb

, from the soil through aquatic ecosystems to lakes has peaked 

## 13

and declined . In other studies, no recovery of lake sediments in response to emission reductions was 

## 14

observed ( Norton, 2007 )

. For example

, Klaminder et al. ( 2010 ) has shown tha t the Pb concentrations in 

## 15

sub

- Arctic lake sediments remain unchanged in recent years , with t he lack of recovery linked to the 

## 16

effects of soil warming

, which affect Pb

- OM transport from soil to the receiving lake systems. Shotyk and 

## 17

Krachler ( 2010 ) also reported a disconnect between atmospheric deposition and recent changes in Pb 

## 18

concentration and isotope ratios in the lake sediments

. Simu lations of future metal behavio r suggest that 

## 19May 20 11 3

- 52 DRAFT

– DO NOT CITE OR QUOTE 

### the more strongly sorbing metals such as Pb will respond to changes in metal inputs or acidification status 

## 1

only over centuries to millennia ( Tipping et al., 2006 )

. 

## 2

Overall, recent research confirms the generally low mobility of Pb in soil. This limited mobility is 

## 3

strongly dependent on both colloid amount and composition, as well as pH, and may be greater in some 

## 4

c ontaminated soils. Mobility is so low that soils continue to act as a sink for atmospheric Pb even though 

## 5

atmospheric Pb concentrations peaked several decades ago. 

## 6



# 3.4. Monitoring of Ambient Lead

3.4.1. Ambient Measurement Techniques

3.4.1.1. Size

- Selective PM Monitoring for Lead Concentrations

Ambient Pb is present in the atmosphere as PM and distributed over a wide range of PM sizes. In 

## 7

recognition of the role of all PM sizes in ambient air Pb exposures, including the ingestion of PM 

## 8

deposited onto surfaces, the indicator fo r the Pb NAAQS is Pb in TSP. As described in Chapter 4, 

## 9

ingestion of deposited Pb can be a substantial contributor to total Pb exposure. Additionally, a substantial 

## 10

fraction of atmospheric Pb may be associated with PM larger than 10 μm (ultracoarse PM). Ho wever, the 

## 11

variability of capture efficiency for TSP using current TSP samplers is considerably greater than the 

## 12

capture efficiency for PM 10 

### using current PM 10 

### samplers. For example , the symmetrical design of Federal 

## 13

Reference Method ( FRM ) samplers for PM 1 0 

### makes their collection efficiency independent of wind 

## 14

direction, and collection efficiency is independent of wind speed under typical sampling conditions. 

## 15

While the collection efficiency of TSP samplers is nearly 100% for fine PM up to 5μm diameter, ther e is 

## 16

much greater variability associated with collection of larger PM ( Wedding et al., 1977 )

. For example, 

## 17

using the FRM for TSP

, a directional difference of 45 degrees can result in a nearly two

- fold difference in 

## 18

15 μm particle collection efficiency and a nearly five

- fold difference in 50 μm particle collection 

## 19

efficiency ( Rodes & Evans, 1985 )

. Effective D 50 

### (size at 50% efficien cy) was observed to decrease from 

## 20

50 μm at a 2 km/ h wind speed to 22 μm at 24 km/ h ( Wedding et al., 1977 )

. 

## 21

Recognizing the variability in capture efficiency associated with TSP samplers and the potential 

## 22

benefit of an indicator with lower measurement variability, the last NAAQS review considered whether 

## 23

the indicator for the Pb NAAQS should be revised from one based on Pb

- TSP to one based on Pb

- PM 10


### . 

## 24

The final decision in the review was to retain the Pb

- TSP indicator. The rationale for this decision 

## 25

included recognition of exposure due to Pb

- TSP that would not be captured by PM 10 

### sampling, the 

## 26

paucity of information documenting the relationship between Pb

- PM 10 

### and Pb

- TSP at th e broad range of 

## 27

Pb sources in the U.S., and uncertainty regarding the effectiveness of a Pb

- PM 10 

### - based NAAQS in 

## 28

controlling ultracoarse Pb

- PM near sources where Pb concentrations are highest (73 FR 66991). Changes 

## 29May 20 11 3

- 53 DRAFT

– DO NOT CITE OR QUOTE 

### were made to the monitoring and data hand ling provisions, however, to allow for the siting of Pb

- PM 10 

## 1

monitors for compliance purposes in locations remote from sources, where the evidence indicates that 

## 2

airborne Pb is predominantly in the PM 10 

### size fraction (73 FR 66964). For Pb

- PM 10 

### monitoring t o be 

## 3

allowed under these regulations, the maximum three

- month average Pb concentration at a site must not 

## 4

exceed 0.10 μg/m 3 

### over a three

- year time period. Additionally, if a Pb

- PM 10 

### monitor is sited near a source, 

## 5

the majority of the particle

- bound Pb must be smaller than 10 μm (40 CFR Part 58). 

## 6

3.4.1.2. Federal Reference Method and Federal Equivalence Method Evaluation

For enforcement of the air quality standards set forth under the Clean Air Act, EPA has established 

## 7

provisions in the Code of Federal Regulations un der which analytical methods can be designated as FRM 

## 8

or federal equivalence methods (FEM). Measurements for determinations of NAAQS compliance must be 

## 9

made with FRMs or FEMs. As of August 2010, 1 manual reference method and 24 manual equivalent 

## 10

methods ha d been approved for Pb ( http://www.epa.gov/ttn/amtic/criteria.html ). The FRM for Pb was 

## 11

promulgated in 1979 and is based on flame atomic absorption spectroscopy (AAS) (40 CFR Part 50)

. The 

## 12

FRM provid es for collection of PM by high volume sampling and analysis of the PM for Pb by atomic 

## 13

absorption spectrometry. Ambient air suspended in PM is collected on a glass fiber filter for 24 hours 

## 14

using a high volume air sampler. The analysis of the 24

- hour samp les may be performed for either 

## 15

individual samples or composites of the samples collected over a calendar month or quarter. Pb in PM is 

## 16

then solubilized by extraction with nitric acid (HNO 3 

### ), facilitated by heat, or by a mixture of HNO 3 

### and 

## 17

hydrochloric ac id (HCl) facilitated by ultrasonication. The Pb content of the sample is analyzed by atomic 

## 18

absorption spectrometry using an air

- acetylene flame, using the 283.3 or 217.0 nm Pb absorption line, and 

## 19

the optimum instrumental conditions recommended by the man ufacturer. Inductively

- coupled plasma 

## 20

mass spectrometry (ICPMS) is under consideration as the new FRM for Pb

- TSP. 

## 21

PM 10 

### monitoring can be used in limited circumstances to measure Pb concentration

. (40 CFR 58)

. 

## 22

The proposed method is based on sampling requir ements for an e xisting Federal Reference Method for 

## 23

the Determination of Coarse PM as PM 10 

### – PM 2.5 

### ( "Reference Method for the Determination of Coarse 

## 24

Particulate Matter as PM10

- 2.5 in the Atmosphere," 2010 )

, which requires a specially approved PM 10 C 

## 25

sampler that meets more demanding performance requirements than conventional PM 10 

### samplers. 

## 26

Ambient air is drawn through an inertial particle size separator for collection on a polytetrafluoroethylene 

## 27

(PTFE) filter. The analysis method for the FRM is based on x

- ray fluorescence spectrometry . In addition, 

## 28

several FEM have been approved based on a variety of principles of operation have been approved, 

## 29

including: inductively coupled plasma optical emission spectrometry, or ICPMS. Specifications for Pb 

## 30

monitoring are designed to help states demonstrate whether they have met compliance crite ria. 

## 31

Operational parameters required under Appendix G of 40 CFR Part 50 are listed in Table 3

- 3. 

## 32May 20 11 3

- 54 DRAFT

– DO NOT CITE OR QUOTE 

### Table 3

- 3

. Specifications for Pb monitoring

Parameter Specification a

Range 0.7

- 7.5 μg Pb/m 3 Sensitivity 0.2

- 0.5 μg Pb/mL for 1% change in absorbance Lower Detectable Limit 0.07 μg Pb/m 3

a Assumes sample volume of 2,400 m 3

.



### Atomic Absorption Spectrometry

AAS is the basis for the existing FRM. A tomic absorption spectrometry was first developed in the 

## 1

19th century, and became widely used in the 1950s

. More than 70 elements can be analyzed by AAS

. 

## 2

Typically a liquid sample is nebulized into a flame with sufficient heat for elements to be atomized

. The 

## 3

liquid specified by the FRM is a nitric acid extract of a glass fiber filter used for collection of suspended 

## 4

PM with a high volume sampler

. The atomized sample is then irradiated with visible light at a specific 

## 5

wavelength to promote an electronic transition to a short

- lived excited state, resulting in absorption of the 

## 6

light

. Elemental selectivity is achieved b ecause light absorption is specific to a particular electronic 

## 7

transition in a particular element

. As a result, absorption of light at a given wavelength generally 

## 8

corresponds to only one element

. The flame is irradiated with a known quantity of light and intensity of 

## 9

light is measured on the other side of the flame to determine the extent of light absorption in the flame

. 

## 10

Using the Beer

- Lambert law the concentration of the element is determined from the decrease in light 

## 11

intensity due to sa mple absorption. 

## 12

A more sensitive variation of atomic absorption spectrometry for most elements is graphite furnace 

## 13

atomic absorption spectrometry (GFAAS). Instead of introducing the sample into a flame, the liquid 

## 14

sample is deposited in a graphite tube that is then heate d to vaporize and atomize the sample. 

## 15

Inductively

- Coupled Plasma Mass Spectrometry

Inductively coupled plasma mass spectrometry (ICPMS) is a sensitive method of elemental 

## 16

analysis developed in the late 1980s. Argon (Ar) plasma (ionized gas) is produced by transmitting radio 

## 17

frequency electromagnetic radiation into hot argon gas with a coupling coil. Temperatures on the order of 

## 18

10,000 K are achieved, which is sufficient for ionization of elements. Liquid samples can be introduced 

## 19

into the plasma by extracti ng samples in an acid solution or water, and nebulizing dissolved elements. 

## 20

Resulting ions are then separated by their mass to charge ratio with a quadrupole and signals are 

## 21

quantified by comparison to calibration standards. While solid samples can be intr oduced by laser 

## 22

ablation, nebulization of liquid extracts of PM collected on Teflon filters is more typical. One major 

## 23

advantage of ICPMS over AAS is the ability to analyze a suite of elements simultaneously. An additional 

## 24

advantage is low detection limits of 50

- 100 parts/trillion for Pb. 

## 25May 20 11 3

- 55 DRAFT

– DO NOT CITE OR QUOTE 

### Inductively

- Coupled Atomic Emission Spectroscopy

Inductively coupled atomic emission spectroscopy (ICP

- AES) also generates ions from elements 

## 1

with a hot Ar plasma, similar to ICPMS. Excited atoms and ions are produced, and these emit 

## 2

electromagnetic radiation with frequencies characteristic of a particular element. Intensity of emission is 

## 3

used to determine the concentration of an element in the sample. Elements are extracted from filter 

## 4

samples and nebulized into the plasm a. 

## 5

Energy Dispersive X

- ray Fluorescence

In energy dispersive X

- ray fluorescence spectrometry a beam of X

- ray photons from an external 

## 6

excitation source is applied to a sample, causing ejection of inner shell electrons from elements in the 

## 7

sample. Because i nner shell electrons have higher electron binding energies than outer shell electrons, the 

## 8

ejection of the inner shell electron induces an energetically favorable electronic transition of an outer shell 

## 9

electron to replace the ejected electron. The energy released as a result of this transition is in the form of 

## 10

electromagnetic radiation, corresponding to the difference in electronic binding energies before and after 

## 11

the transition. The energy released is typically in the X

- ray portion of the electromagneti c spectrum. The 

## 12

release of electromagnetic radiation as a result of an electronic transition is defined as fluorescence. 

## 13

Fluorescence energies associated with electronic transitions depend on atomic structure, and vary between 

## 14

elements. As a result, X

- ray fluorescence energy is uniquely characteristic of an element, and X

- ray 

## 15

intensity at a given energy provides a quantitative measurement of elemental concentration in the sample. 

## 16

The X

- rays are detected by passing them through a semiconductor material, resu lting in an electrical 

## 17

current that depends on the energy of the X

- ray. 

## 18

3.4.1.3. Chemical Speciation Network, IMPROVE, and National Air Toxics

Trends Network Monitors

In addition to being monitored for regulatory purposes in the SLAMS network, Pb is also 

## 19

monitored in three other sampling networks. Pb is monitored at 53 monitoring sites as a part of the 

## 20

Chemical Speciation Network. Participating monitoring agencies responsible for site operation are given 

## 21

flexibility in sampler design, with filter collection media be st

- suited for the analysis of specific 

## 22

components ( U.S. EPA, 1999a ) . Several samplers are approved for CSN monitoring, all of which collect 

## 23

bulk PM sp ecies with multiple channels containing different types of filters appropriate for speciation 

## 24

sampling. Pb is one of 33 elements in PM 2.5 

### collected on Teflon filters every third day and analyzed by 

## 25

energy dispersive X

- ray fluorescence spectrometry. 

## 26

Pb is a lso monitored at 110 aerosol visibility

- monitoring sites as a part of the Interagency 

## 27

Monitoring of Protected Visual Environments (IMPROVE) program. An additional 59 aerosol samplers 

## 28

that are not directly operated through the program are operated following IMPROVE protocols. 

## 29May 20 11 3

- 56 DRAFT

– DO NOT CITE OR QUOTE 

### IMPROVE is a cooperative effort by Federal and state organizations to protect visibility in 156 national 

## 1

parks and wilderness areas as described in the 1977 amendments to the Clean Air Act. Objectives are: (1) 

## 2

to establish current visib ility and aerosol conditions in high priority (class I) areas for visibility protection; 

## 3

(2) to identify chemical species and emission sources responsible for existing man made visibility 

## 4

impairment (3) to document long

- term trends for assessing progress t owards visibility goals; and (4) to 

## 5

provide regional haze monitoring representing protected federal areas in accordance with the regional 

## 6

haze rule. The IMPROVE sampler operates with four sampling modules, three for PM 2.5 

### and one for 

## 7

PM 10 

### . Pb is not measur ed in PM 10 

### , but one of the three PM 2.5 

### modules contains a Teflon filter used for 

## 8

determination of gravimetric mass, absorbance, and elemental analysis by Particle Induced X

- Ray 

## 9

Emission (PIXE) and XRF. A total of 9 elements are determined by XRF, including Pb ( University of 

## 10

California Davis, 1995 )

. 

## 11

Pb in PM 10 

### is also monitored in the National Air Toxics Trends Station (NATTS) network ( ERG, 

## 12

2009 ) . PM is collecte d either by high volume sampling with a quartz fiber filter or low volume sampling 

## 13

with a PTFE filter following EPA Compendium Method IO

- 3.5 ( U.S. EPA, 1999b ) . Pb is one of seven 

## 14

core metals collected on Teflon filters and analyzed by ICPMS. The NATTS network was developed to 

## 15

fulfill the need for consistent data quality of long

- term monitoring data o n hazardous air pollutants of 

## 16

consistent data quality, for use in assessing trends and emission reduction program effectiveness, 

## 17

assessing and verifying air quality models, exposure assessments, emission control strategies, and as 

## 18

direct input for receptor modeling. As of December 2009, the network consisted of 27 monitoring 

## 19

stations, including 20 urban and 7 rural stations operating on a one in six day sampling frequency. 

## 20

Typically more than 100 pollutants are measured at each site, and monitoring is requi red for nineteen 

## 21

species, including Pb. Pb monitoring is also required for PM 2.5 

### samples at NCore monitoring sites 

## 22

beginning no later than January 1, 2011, and monitoring of Pb in coarse PM (PM 10 

### - PM 2.5 

### ) is also likely to 

## 23

be included ( U.S. EPA, 2006 )

. 

## 24

3.4.1.4. Other Measurement Methods for Total Lead

Several other methods that have not been designated as FRM or FEM methods have also been 

## 25

frequently used to obtain atmospheric Pb measurements. These include proton induced x

- ray emission 

## 26

(PIXE), X

- ray photoelectron spect roscopy (XPS), and other methods 

## 27

PIXE

Proton

- induced X

- ray emission (PIXE) spectroscopy has been widely used to measure Pb in 

## 28

atmospheric PM. It is the method used for Pb analysis in the IMPROVE network. In PIXE, a high

- energy 

## 29

proton beam passes through th e sample, causing electrons to be excited from inner shells. The x

- rays 

## 30

emitted when electronic transition occur to replace the inner shell electrons are characteristic of an 

## 31May 20 11 3

- 57 DRAFT

– DO NOT CITE OR QUOTE 

### element and can be used to identify it. Development of PIXE for analysis of airbo rne PM was reviewed 

## 1

by Cahill et al. ( 1981 )

. Numerous applications of PIXE to analysis of airborne Pb

- PM have been reported 

## 2

in the past five years ( Ariola et al., 2006 ; Chan et al., 2008 ; Cong et al., 2007 ; Johnson et al., 2006 ; 

## 3

Johnson et al., 2008 ; Sanchez

- Ccoyllo et al., 2009 ; Wåhlin et al., 2006 ) ( Cohen et al., 2010 ; Waheed et 

## 4

al., 2010 )

. 

## 5

XPS

X

- ray photoelectron spectroscopy (XPS), also called electron spectroscopy f or chemical analysis 

## 6

(ESCA) has been used to determine Pb concentrations on materials surfaces, including atmospheric PM 

## 7

( Finlayson

- Pitts & Pitts, 2000 ) . A fixed frequency X

- ray beam causes inner shell electrons to be emitted 

## 8

and kinetic energy of ejected electrons is measured. Binding energy characteristic of an element can be 

## 9

calculated from the measured kinetic energy, allowing identification of the element. XPS can also provide 

## 10

in formation about an element’s chemical environment or oxidation states because of chemical shifts in 

## 11

binding energy caused by differences in chemical environment. There have been some recent applications 

## 12

of XPS to airborne PM, concluding that Pb was mostly in the form in of Pb sulfate ( Qi et al., 2006 ) . XPS 

## 13

analysis is a surface technique that is suitable only to a depth of 20

- 50Å. 

## 14

Other Total Lead Methods

Anodic stripping voltammetry, a tomic emission spectroscopy, and colorimetry have also been used 

## 15

for measurement of atmospheric Pb ( Finlayson

- Pitts & Pitts, 2000 ) . In anodic stripping voltammetry, 

## 16

metal ions are reduced to metallic form and concentrated as an amalgam on a suitable electrode (e.g. a 

## 17

mercury amalgam on a mercury electrode). This is followed by re

- oxidation in solution, which requires 

## 18

“stripping” the reduced metal from the electrode. Emission spectro scopy can be compared to the existing 

## 19

FRM for Pb based on AAS. In atomic absorption spectroscopy radiation absorbed by non

- excited atoms 

## 20

in the vapor state is measured. In emission spectroscopy, radiation due to the transition of the electron 

## 21

back to groun d state after absorption is measured, and the energy of the transition is used to uniquely 

## 22

identify an element in a sample. Colorimetric methods are wet chemical methods based on addition of 

## 23

reagents to a Pb containing solution to generate measurable light absorbing products. These methods are 

## 24

less sensitive than ICPMS, XRF, and PIXE and their use is declining as more sensitive methods become 

## 25

more widely used, but have advantages regarding simplicity and cost. 

## 26

3.4.1.5. Sequential Extraction

Sequential extraction has been widely used to further classify Pb for various purposes, including 

## 27

bioavailability, mobility, and chemical speciation. In general the more easily extractable Pb is considered 

## 28May 20 11 3

- 58 DRAFT

– DO NOT CITE OR QUOTE 

### more mobile in soil and is more bioavailable to organisms. This approach ha s also been used widely in 

## 1

characterization of airborne PM. In its original application ( Tessier et al., 1979 ) metals extraction solvents 

## 2

were selected to correspond to common species present in soil, and metals were classified as 

## 3

exchangeable, b ound to carbonates, bound to iron and manganese oxides, bound to OM, and residual. 

## 4

Extraction was carried out with successively stronger solutions, starting with magnesium chloride for 

## 5

removal of exchangeable metals and ending with hydrofluoric and perchlo ric acids for removal of 

## 6

residual metals. Pb was one of the elements originally studied by Tessier et al. ( 1979 ) as well as one the 

## 7

elements analyzed when Tessier’s scheme was first applied to airborne PM ( Fraser & Lum, 1983 )

. 

## 8

Tessier’s scheme was modified and optimized for airborne PM over time ( Fernandez Espinosa et 

## 9

al., 2002 ) and addit ional extraction schemes were also developed ( Chester et al., 1989 ) , including the 

## 10

simplest case of two fractions c orresponding to soluble and insoluble fractions ( Canepari et al., 2006 ; 

## 11

Falta et al., 2008 ; Voutsa & Samara, 2002 ) . The variety of methods in current use was recently thoroughly 

## 12

reviewed by Smichowski et al. ( 2005 )

. With the recognition that biological processes involving deposited 

## 13

PM metals were related to their solubility ( U.S. EPA, 2009 ) , sequential extraction methods or simpler 

## 14

schemes to divide metals into water and acid soluble fractions were increasingl y applied to PM samples to 

## 15

obtain data not just on total metal concentration but also on water soluble concentration ( Graney et al., 

## 16

2004 ; Kyotani & Iwatsuki, 2002 ; Wang et al., 2002 ) . Compared to other elements, a large fraction of total 

## 17

Pb is soluble ( Graney et al., 2004 ) . Recent advances in this area have included application to size 

## 18

fractionated PM ( Birmili et al., 2006 ; Dos Santos et al., 2009 ) , time resolved measurements ( Perrino et 

## 19

al., 2010 ) , and an extensive comparison of different fractionatio n schemes ( Canepari et al., 2010 )

. 

## 20

Sequential extraction with two or more fractions is becoming more widely used for characterization of 

## 21

Pb

- PM in a variety of sources ( Canepari et al., 2008 ; Poykio et al., 2007 ; Sillanpaa et al., 2005 ; 

## 22

Smichowski et al., 2008 ) and locations ( Al

- Masri et al., 2006 ; Annibaldi et al., 2007 ; Canepari et al., 

## 23

2006 ; Cizmecioglu & Muezzinoglu, 2008 ; Dahl et al., 2008 ; Dos Santos et al., 2009 ; Fujiwara et al., 

## 24

2006 ; Gutierrez

- Castillo et al., 2005 ; Heal et al., 2005 ; Perrino et al., 2010 ; Richter et al., 2007 ; Sato et 

## 25

al., 2008 ; W. Wang et al., 2006 ; Yadav & Rajamani, 2006 ) , leading to a better understanding of mobility 

## 26

characteristics of Pb in airborne PM. 

## 27

3.4.1.6. Speciation Techniques

XAFS

There have been few attempts to speciate Pb in atmospheric PM into individual species. However, 

## 28

recently X

- ray absorption fine s tructure (XAFS) has been applied to PM and road dust to obtain Pb 

## 29

speciation data from direct analysis of particle surfaces. In XAFS the absolute position of the absorption 

## 30

edge can be used to determine the oxidation state of the absorbing atom, and scatte ring events that 

## 31May 20 11 3

- 59 DRAFT

– DO NOT CITE OR QUOTE 

### dominate in the near edge region provide data on vacant orbital energies, electronic configurations, and 

## 1

site symmetry of the absorbing atom that can be used to determine the geometry of the atoms surrounding 

## 2

the absorbing atom. XAFS can b e divided into two spectral regions. X

- ray absorption near edge structure 

## 3

(XANES) is the region of the x

- ray absorption spectrum up to 50 eV above the absorption edge observed 

## 4

when an inner shell electron is electronically excited into unoccupied states, a nd Extended X

- ray 

## 5

Absorption Fine Structure (EXAFS) up to 1 keV above the absorption edge. Both have been applied 

## 6

recently to Pb in PM. XANES spectra of Pb coordination complexes with a wide range of 

## 7

environmentally relevant ligands have been reported ( Swarbrick et al., 2009 ) . XANES has been used to 

## 8

show that several different Pb species are probably present in urban airborne PM ( Funasaka et al., 2008 ) 

## 9

and urban road dust ( Barrett et al., 2010 ) . XANES has been used to differentiate between Pb chromate, 

## 10

Pb

- sorbed minerals, Pb chloride, Pb oxide, Pb carbonate, Pb sulfide and Pb sulfate are probably present in 

## 11

urban PM and road dust samples ( Barrett et al., 2010 ; Funasaka et al., 2008 ; Tan et al., 2006 ) . XANES 

## 12

has also been used to quantify Pb complexed with humic substances from soil in road dust ( Pingitore et 

## 13

al., 2009 ) and to investigate the speciation of atmospheric Pb in soil after deposition ( X. Y. Guo et al., 

## 14

2006 ) . EXAFS has been applied to emission sources to show Pb from a sinter plant was mainly carbonate 

## 15

( Sammut et al., 2010 ) . XAFS has only been applied to airborne PM very recently and shows promise for 

## 16

chemical speciation of airborne metals, including Pb. 

## 17

GC

- and HPLC

- ICPMS

Environmental analytical methods for organolead compounds prior to 2000 were generally time 

## 18

consuming and costly, requiring extraction , derivatization, and detection ( Quevauviller, 2000 ) . These have 

## 19

been thoroughly reviewed ( Pyrzyńska, 1996 ) and method intercomparison studies have been conducted 

## 20

( Quevauviller, 2000 ) . More recently, speciation o f organometallic compounds in environmental samples 

## 21

has usually carried out by coupling a chromatographic separation step with a mass spectrometry

- based 

## 22

multi

- element detection systems capable of analyzing a wide range of elements along with Pb, and these 

## 23

approaches have also been recently reviewed ( Hirner, 2006 ) . Chromatographic systems in common use 

## 24

are gas chromatography and high performance liquid chromatography. Detection systems most commonly 

## 25

used are ICPMS, elec tron impact ionization mass spectrometry (EI

- MS), and electrospray ionization mass 

## 26

spectrometry (ESI

- MS) ( Hirner, 2006 ) . Using these techniques, organometallic species are separated from 

## 27

each other based on difference s in retention times on chromatographic columns, and elemental Pb is 

## 28

determined by the ICPMS used as a detector downstream of the column to measure elemental Pb in the 

## 29

pure compounds after chromatographic separation. Pb speciation analysis has benefited fr om the 

## 30

development of HPLC

- ICPMS in particular ( Quevauviller, 2000 ) . Recent advances in metal speciation 

## 31

analysis in environmental samples by HPLC

- ICPMS have been extensively reviewed ( Popp et al., 2010 )

. 

## 32

HPLC

- ICPMS has been used for analysis of Pb complexes with humic substances ( Vogl & Heumann, 

## 33May 20 11 3

- 60 DRAFT

– DO NOT CITE OR QUOTE 

### 1997 ) , which could be relevant for resuspended soil and road dust. GC

- ICPMS has been more widely 

## 1

used for separation and analysis of methyl and ethyl Pb species in atmospheric PM ( Jitaru et al., 2004 ; 

## 2

Leal

- Granadillo et al., 2000 ; Poperechna & Heumann, 2005 )

. 

## 3

3.4.1.7. Continuous Lead Monitoring

Development of high time resolution measurement capabilities has advantages for determining 

## 4

peak exposure concentrations and diurnal exposure trends. High time resolution samplers suitable for 

## 5

analysis after sampling by XRF and ICPMS have been developed an d applied. The eight

- stage Davis 

## 6

Rotating Unit for Monitoring (DRUM) impactor ( T. A. Cahill et al., 1987 ; Raabe et al., 1988 ) collects PM 

## 7

samples with a cascade impactor on Mylar film substrate on a slowly rotating drum, with sample s 

## 8

analyzed by XRF. It has been used to measure size and time resolved Pb and other elements with a time 

## 9

resolution of less than 6 hours using x

- ray fluorescence ( Bench et al., 2002 ; C. F. Cahill, 2003 ) . The 

## 10

University of Maryland Semi

- continuous Elements in Aerosol Sampler ( Kidwell & Ondov, 2001

, 2004 ) 

## 11

uses direct steam injection into promote condensational growth of sampled at a high flow rate, and 

## 12

accumulates resulting droplets in a slurry by impaction. It has been successfully applied to measurement 

## 13

of Pb and other elements by AAS ( Pancras et al., 2006 ; Pancr as et al., 2005 ) with a 30

- minute time 

## 14

resolution. This approach is also suitable for ICPMS analysis. A gas converter apparatus has also been 

## 15

developed to improve transfer of ions to the ICPMS, including Pb , and successfully tested with outdoor 

## 16

air ( Nishiguchi et al., 2008 ) . Other high time res olution methods suitable for Pb analysis in PM are under 

## 17

development, including near real

- time XRF analysis. 

## 18

Much of the recent progress in ambient aerosol instrumentation has been related to the 

## 19

development and improvement of single particle mass spectrom etry ( Prather et al., 1994 ) . This technique 

## 20

can also be considered as an effective method for real time Pb measurement in PM ( Silva & Prather, 

## 21

1997 ) . Pro gress has continued in the development of single particle mass spectrometry to quantify 

## 22

elements and organic ion fragments and a number of recent applications that included ( Bein et al., 2007 ; 

## 23

Johnson et al., 2008 ; Pekney et al., 2006 ; Reinard et al., 2007 ; Snyder et al., 2009 ) or specifically targeted 

## 24

( Moffet, de Foy, et al., 2008 ; Murphy et al., 2007 ; Salcedo et al., 2010 ) Pb measurements. 

## 25



# 3.4.2. Ambient Network Design

Ambient air Pb concentration is detected by FRM monitors at state and local monitoring stations 

## 26

(SLAMS) reporting data used for NAAQS compliance to the Air Quality System (AQS). Monitoring 

## 27

requirements for Pb have evolved over the past ten years. Monitorin g for ambient Pb levels has been 

## 28

required for all major urban areas where ambient air Pb measurements have been elevated near or beyond 

## 29

the level of the NAAQS

. Alternately, state and local agencies have located Pb monitoring stations in 

## 30

proximity to Pb poi nt source emissions. Prior to 2006, monitoring sites were established where sources 

## 31May 20 11 3

- 61 DRAFT

– DO NOT CITE OR QUOTE 

### emitted 5 or more tons/yr or where smaller stationary sources were located in proximity to populated 

## 1

neighborhoods. 

## 2

Pb monitoring requirements have experienced several cha nges since publication of the last Pb 

## 3

AQCD ( U.S. EPA, 2006 ) . In 2008 revisions for the Pb NAAQS were announced, and new monitoring 

## 4

requirements to support NAAQS revision called for expanded monitoring at sources that emit Pb at a rate 

## 5

of 1.0 or more tons/yr and non

- source oriented monitoring at each Core Based Statistical Area (CBSA) 

## 6

with a population of 500,000 or more ( "National Ambient Air Quality Standards for Lead (final rule)," 

## 7

2008 )

. This corresponded to approximately 100 non

- source oriented monitors. Some of the new monitors 

## 8

were required to become operational by January 1, 2010, with the remainder operational by January 1, 

## 9

2011. Subsequent revisions to these requirements have been pr omulgated, including reduction of the 

## 10

source oriented monitoring threshold from 1.0 tons/yr to 0.5 tons/yr and replacing the requirement for 

## 11

population

- based CBSA monitoring with a requirement for non

- source oriented Pb monitoring at National 

## 12

Core multipol lutant monitoring network (NCore) sites in CBSA’s with a population of 500,000 or more 

## 13

(75 FR 81126). NCore is a new network of multipollutant monitoring stations intended to meet multiple 

## 14

monitoring objectives. The NCore stations are a subset of the SLAMS network are intended to support 

## 15

long

- term trends analysis, model evaluation, health and ecosystem studies, as well as NAAQS 

## 16

compliance. The complete NCore network consists of approximately 60 urban and 20 rural stations, 

## 17

including some existing SLAMS site s that have been modified for additional measurements. Each state 

## 18

will contain at least one NCore station, and 46 of the states plus Washington, DC, will have at least one 

## 19

urban station. 

## 20

Data used in this chapter are from the period 2007

- 2009. The number of source oriented and non



## 21

source oriented monitors changed each analysis year because the monitoring requirements changed over 

## 22

this time. Monitors were designated to be source

- oriented if they were designated in AQS as source 

## 23

oriented, or they were locate d within one mile of a 0.5 ton/yr or greater source, as noted in the 2005 NEI 

## 24

( U.S. EPA, 2008a ) . Non

- source oriented monitors were those monitors not considered to be source 

## 25

oriented. This loo sening of the restrictions was intended to accommodate the 2007

- 2008 data that were 

## 26

obtained before the latest monitor designation requirements were implemented. 

## 27

In addition to FRM monitoring, Pb is also measured within the Chemical Speciation Network 

## 28

(CSN ), Interagency Monitoring of Protected Visual Environments (IMPROVE), and the National Air 

## 29

Toxics Trends Station (NATTS) networks. While monitoring in multiple networks improves geographic 

## 30

coverage, measurements between networks are not directly comparable in all cases because different PM 

## 31

size ranges are sampled in different networks. Depending on the monitoring network, Pb is monitored 

## 32

either in TSP, PM 10 

### , or PM 2.5



### . 

## 33May 20 11 3

- 62 DRAFT

– DO NOT CITE OR QUOTE 

## 3.4.2.1. Monitor Siting Requirements

Spatial scales defined for Pb monitoring range from microscale to regional scale ( 40 CFR Part 58 ) : 

## 1

 Microscale: Defines the concentrations in air volumes associated with area dimensions 

## 2

ranging from several meters up to about 100 m. This scale applies to areas in close proximity 

## 3

to Pb point sources. 

## 4

 Middle Scale: Def ines the concentration typical of areas up to several city blocks in size with 

## 5

dimensions ranging from about 100 m t o 0.5 km

. This scale generally represents Pb air 

## 6

quality levels in areas up to several city blocks. 

## 7

 Neighborhood Scale: Defines concentratio ns within some extended area of the city that has 

## 8

relatively uniform land use with dimensions in the 0.5 to 4.0 km range. Where a 

## 9

neighborhood site is located away from immediate Pb sources, the site may be very useful in 

## 10

representing typical air quality v alues for a larger residential area, and therefore suitable for 

## 11

population exposure and trends analyses. 

## 12

 Urban Scale: Defines concentrations within area of city

- like dimensions on the order of 4 to 

## 13

50 km. Such stations would be used to present ambient Pb c oncentrations over an entire 

## 14

metropolitan area with dimensions in the 4 to 50 k m range. An urban scale station would be 

## 15

useful for assessing trends in citywide air quality and the effectiveness of larger scale air 

## 16

pollution control strategies. 

## 17

 Regional Sca le: Defines usually an area of reasonably homogeneous geography without 

## 18

larges sources, and extends from tens to hundreds of kilometers. This large scale of 

## 19

representativeness has not been used widely for Pb monitoring, but provides important 

## 20

information o n background air quality and inter

- regional pollutant transport. 

## 21

Since the majority of Pb emissions mass comes from point sources, such as metals processing 

## 22

facilities, waste disposal and recycling, and fuel combustion, the SLAMS network is primarily used to 

## 23

assess the air quality impacts of Pb point sources

. A second purpose of the SLAMS network is to 

## 24

determine the broad population exposure from any Pb source

. The most important spatial scales to 

## 25

characterize the emissions from point sources are the micro, middle, and neighborhood scales

. 

## 26

Background information such as point source emissions inventories, climatological summaries, and 

## 27

local geographical characteristics are used to identify areas where monitoring is necessary. After siting 

## 28

each Pb station, s pecific siting criteria must be fulfilled for Pb monitoring in the SLAMS network

. Micro 

## 29

and middle scale monitors must be 2

- 7 m above ground

. All other scale monitors must be 2

- 15 m above 

## 30

ground

. Monitors must be more than 2 m from supporting structures

. Monitors must be more than 10 m 

## 31

from trees

. Microscale monitors designed for monitoring traffic related Pb must be 2

- 10 m from 

## 32May 20 11 3

- 63 DRAFT

– DO NOT CITE OR QUOTE 

### roadways

. Distance from roadways for other scales depends on the purpose and scale of the monitor and 

## 1

the level of traffic on the roadway

. 

## 2

3.4.2.2. Spatial and Temporal Coverage

Figure 3

- 12. Pb monitoring sites for SLAMS, CSN, NATTS, and IMPROVE networks, 2007

- 2009.

Figure 3

- 12 shows Pb monitoring sites for all networks

. The top left map shows the SLAMS 

## 3

monitors reporting to the AQS syste m from 2007 to 2009

. Monitors are indicated as Pb

- TSP source 

## 4May 20 11 3

- 64 DRAFT

– DO NOT CITE OR QUOTE 

### oriented , Pb

- TSP non

- source oriented, and PM 10 

### non

- source oriented . Source and non

- source designations 

## 1

used in the data analysis in Section 3.5 are indicated on this map. The top right map shows PM 2.5 

### monitors 

## 2

from the CSN

. The bottom left map shows National Air Toxics Trend Networks sites, and the bottom right 

## 3

map shows IMPROVE monitoring sites. There is a high density of FRM monitors in some cities 

## 4

containing Pb sources, including Los Angeles, S t. Louis, Pittsbu r gh, and Minneapolis

. As a result, 

## 5

population coverage varies across cities, with those cities with major Pb sources having greater coverage

. 

## 6

Coverage for PM 2.5 

### in the CSN and IMPROVE network is geographically comprehensive

. In comparison, 

## 7

the FRM sites are more representative of source effects. NATTS sites have less extensive national 

## 8

coverage. 

## 9



# 3.5. Ambient Air Lead Concentrations

The following section summarizes data on ambient air Pb concentrations during the years 2007

- 

## 10

2009. The section begi ns with a description of Pb concentrations observed in TSP, PM 10 

### , and PM 2.5 

### at 

## 11

source oriented and non

- source oriented monitors across the U.S. Next, seasonal patterns and multi

- year 

## 12

trends of Pb concentration are presented for the U.S. It is notable that Pb concentrations have declined 

## 13

substantially over the past 40 years; this is described further in Section 3.5.2. An examination of AQS 

## 14

data and the peer

- reviewed literature is provided to evaluate the size distribution of Pb

- bearing airborne 

## 15

PM under vari ed ambient conditions. Finally, the relationship between Pb concentration and 

## 16

concentrations of copollutants are presented. Summary information is presented within this section, and 

## 17

detailed data are included in an Appendix to this chapter. 

## 18



# 3.5.1. Spatial Distrib ution of Air Lead

3.5.1.1. Variability across the U.S.

This section presents nationwide Pb concentration data measured using source oriented and non



## 19

source oriented TSP FRM monitors, PM 10 

### monitors, and PM 2.5 

### monitors from the CSN for 2007

- 2009. 

## 20

This information is useful to develop a sense of variability in Pb concentrations at a national scale. For 

## 21

this analysis, source oriented monitors encompassed all those listed as “source oriented” in the AQS, 

## 22

based on state agency reporting, plus those within one mile of a f acility emitting 0.5 ton/yr or more. Non



## 23

source oriented monitors were those monitors in the system not designated to be source oriented (Figures 

## 24

3

- 13 though 3

- 16), the majority of U.S. counties do not have a Pb monitor. 

## 25May 20 11 3

- 65 DRAFT

– DO NOT CITE OR QUOTE 

### Concentrations of Pb Measured using TSP Monitors

Source oriented maximum 3

- month average Pb data were obtained for 22 counties across the U.S. 

## 1

during the period 2007

- 2009. Figure 3

- 13 illustrates that the level of the NAAQS was exceeded in 

## 2

fourteen counties where source oriented monitoring was performed. Summary data are presented below in 

## 3

Table 3

- 4, and detailed data for the one

- month and three

- month average and maxima source oriented Pb



## 4

TSP concentrations are provided in Tables 3A

- 1 through 3A

- 4 in the Appendix . The mean was skewed 

## 5

towards the 75 th perc entile of the distribution for both the monthly and three

- month data sets. The 

## 6

primary difference between the one

- month average and three

- month rolling average data sets occurs at the 

## 7

upper tails of the distribution. Data for sites at which national maxima were reached for 2007

- 2009 are 

## 8

presented in Table 3

- 5. The highest monthly and three

- month average concentrations occurred at the same 

## 9

sites: Herculaneum, MO followed by Los Angeles, CA. The highest annual site max 1

- month value 

## 10

occurred in Cook County, I L in 2008

, followed by Iron County, MO in 2008 and Hillsborough County, FL 

## 11

in 2007 . The Herculaneum and Los Angeles sites were also above the 90 th percentile annual 3

- month site 

## 12

max Pb concentrations. The highest annual site max 3

- month concentrations occu rred in Herculaneum in 

## 13

2008, Los Angeles in 2008, and Iron County, MO in 2008. The majority of data were below the level of 

## 14

the NAAQS over the 3

- year period, but high values at a subset of source oriented monitoring locations 

## 15

tended to skew the nationwide distribution of Pb concentration data upwards. 

## 16

Table 3

- 4

. Summary data for source oriented Pb monitors across the U.S.

Mean, μg/m 3 Median, μg/m 3 95th%, μg / m 3 99 th%, μg / m 3 Max, μg/m 3

Monthly 0.24 0.070 0.98 2.1 4.4 3

- mo r olling a vg 0.24 0.080 0.98 1.9 2.9



### Table 3

- 5

. Summary data for sites at which source oriented statistics are at a maxima

County AQS Highest Monthly Mean, μg/m 3 Highest 3

- mo Mean, μg/m 3 Highest Monthly Annual Site Max, μg/m 3 (Year) Highest 3

- mo Annual Site Max

, μg/ m 3 (Year)

Jefferson, MO 290990015 1.3 1.3 2.9 (2008)
Jefferson, MO 290990004 1.1 1.1 Los Angeles, CA 060371405 0.86 0.93 2.5 (2008) Chicago, IL 180350009 4.4 (2008)
Iron, MO 290930016 4.2 (2008) 2.5 (2008)
Hillsborough, FL 120571066 3.6 (2007)May 20 11 3

- 66 DRAFT

– DO NOT CITE OR QUOTE 

### Figure 3

- 13. Highest county

- level source oriented Pb

- TSP concentrations ( μ g/m 3 

### ), maximum 3

- month average, 2007

- 2009.

Non

- source oriented maximum 3

- month average Pb concentration data were obtained with TSP 

## 1

monitors for 36 counties across the U.S. during the period 2007

- 2009; non

- source PM 10 

### monitoring data 

## 2

are not included here because they were obtained at select sites in 2009 only . The median for monthly and 

## 3

3

- m onth rolling average data was 0.010 μg/m 3 

### . Detailed d ata for site

- specific Pb

- TSP concentrations are 

## 4

provided in Tables 3A

- 5 and 3A

- 6 in the Appendix to this chapter. Nationwide, the mean monthly average 

## 5

non

- source oriented Pb

- TSP concentration for 2007

- 20 09 was more than an order of magnitude lower 

## 6

than the source

- oriented data. Collectively, these data indicate that non

- source oriented monitors tend to 

## 7

measure concentrations an order of magnitude lower than the level of the NAAQS. 

## 8

Concentrations of Pb Mea sured using PM 10 

### Monitors

Figure 3

- 14 displays maximum 3

- month average county

- level data for Pb in PM 10 

### concentrations 

## 9

for 36 counties in which measurements were obtained. The data presented here are not compared to the 

## 10

level of the NAAQS because PM 10 

### moni tors were not incorporated into the non

- source oriented 

## 11

monitoring network until 2009. Among the 36 counties in which PM 10 

### monitoring was conducted, only 

## 12

one county, Gila County, AZ, reported concentrations above 0.076 μg / m 3 

### . Three other counties reported 

## 13May 20 11 3

- 67 DRAFT

– DO NOT CITE OR QUOTE 

### concentrations greater than 0.016 μg / m 3 

### : Wayne County, MI, Boyd County, KY, and the county of St. 

## 1

Louis City, MO. 

## 2

Figure 3

- 14

. Highest county

- level Pb

- PM 10 

### concentrations ( μg/ m 3 

### )

, maximum 3

- month average, 2007

- 2009.

Figure 3

- 15 displays maximum 3

- month average county

- level data for Pb in PM 2.5 

### concentrations 

## 3

for 323 counties in which PM 2.5 

### measurements were obtained for speciation in the CSN and IMPROVE 

## 4

networks. The data presented here are not compared to the level of the NAAQS because PM 2.5 

### mon itors 

## 5

are not deployed for that purpose. Among the 323 counties in which PM 2.5 

### monitoring was conducted, 

## 6

only eleven counties reported concentrations greater than 0.016 μg / m 3 

### : Jefferson, AL, San Bernardino, 

## 7

CA, Imperial, CA, Wayne, MI, Jefferson, MO, Erie, NY, Lorain, OH, Allegheny, PA, Berks, PA, 

## 8

Davidson, TN, and El Paso, TX. 

## 9May 20 11 3

- 68 DRAFT

– DO NOT CITE OR QUOTE 

### Figure 3

- 15. Highest county

- level Pb

- PM 2.5 

### concentrations ( μg/ m 3 

### ),
maximum 3

- month average, 2007

- 2009.



## 3.5.1.2. Intra

- urban Variability

Intra

- urban variability is defined as the variation in Pb concentration across an urban area. Because 

## 1

the source characteristics and size distribution of particle

- bound Pb can vary considerably, spatial 

## 2

variability of Pb concentrations in urban areas may also be high. Such variability may not be detected if 

## 3

one or a small number of central site monitors is in use, so cities with multiple monitors can be used to 

## 4

characterize intra

- urban variability. Intra

- urban variability in Pb concentrations reported to AQS at the 

## 5

individual county level was described in de tail in the Appendix in Section 3A.2 . County

- level data were 

## 6

used because PM

- bound airborne Pb tends to settle over short distances to produce large spatiotemporal 

## 7

variability, as described in Section 3.3.1 and revisited in Section 3.5.3. Los Angeles Count y, CA (Los 

## 8

Angeles), Hillsborough and Pinellas Counties, FL (Tampa), Cook County, IL (Chicago), Jefferson County, 

## 9

MO (Herculaneum), Cuyahoga County, OH (Cleveland), and Sullivan County, TN (Bristol) were selected 

## 10

for this assessment to illustrate the varia bility in Pb concentrations measured across different metropolitan 

## 11

regions with varying combinations of source and urban contributions of Pb. Maps and wind roses are 

## 12

presented in the Appendix for each of the six urban areas. Additionally, annual and season al box plots of 

## 13

the Pb concentration distributions and intra

- monitor correlation tables are presented to illustrate the level 

## 14

of variability throughout each urban area. 

## 15May 20 11 3

- 69 DRAFT

– DO NOT CITE OR QUOTE 

### When collectively reviewing the data from the six counties, it became apparent that sp atial and 

## 1

temporal variability of Pb concentrations were commonly high. Variability was high for areas that 

## 2

included a Pb source, with high concentrations downwind of the sources and low concentrations at areas 

## 3

far from sources. When no large sources of Pb were present, variability of Pb concentrations were lower, 

## 4

and more data were observed to lie below the MDL. For example, the Los Angeles County, CA data 

## 5

illustrated very high concentrations adjacent to a Pb recycling facility, but non

- source oriented 

## 6

con centrations were well below the level of the NAAQS at all times. As described in Section 3.3, PM size 

## 7

distribution influences how far the particle will travel upon initial emission or resuspension before being 

## 8

deposited. Meteorology, nature of the sources, distance from sources, and positioning of sources with 

## 9

respect to the monitors all appeared to influence the level of concentration variability across time and 

## 10

space for the monitoring data analyzed in the Appendix. Additionally, resuspension of wheel wei ghts and 

## 11

soils, emission of trace Pb during on

- road gasoline combustion, and urban background levels of Pb are 

## 12

uncertain influential factors in ambient Pb concentrations. This is consistent with field studies to 

## 13

characterize Pb concentrations that are desc ribed in the literature. 

## 14

A number of studies have characterized how Pb

- bearing PM is distributed over the neighborhood 

## 15

scale in the air. Martuzevicius et al. ( 200 4 ) examined the spatial variability of PM 2.5 

### samples obtained in 

## 16

the greater Cincinnati, OH area at 6 urban, 4 suburban, and 1 rural site and found that Pb concentration in 

## 17

PM 2.5 

### had a coefficient of variation (CV, defined as the standard deviation of site measurements divided 

## 18

by the average) of 33.8%, compared with a CV for PM 2.5 

### of 11.3% over all sites. Average Pb

- PM 2.5 

## 19

concentration among the sites varied from 1.79

- 28.4 ng/m 3 

### . Martuzevicius et a l. ( 2004 ) suggested that 

## 20

differences between mass and Pb spatial variability implied that Pb originated primarily from local 

## 21

sources. Sabin et al. ( 2006 ) measured Pb

- PM with an upper cutpoint of 29 μm and found that urban 

## 22

concentrations ranged from 2.2 to 7.4 ng/m 3 

### with a CV of 40%. In contrast, a rural location had a 

## 23

concentration of 0.62 ng/m 3 

### . Sabin et al. ( 2006 ) also reported deposit ion flux at the same sites, which 

## 24

ranged from 8.3 to 29 μg/m 2



### · day at the rural sites, with a CV of 48%, and was 1.4 ng/m 3 

### at the rural site. 

## 25

Han et al. ( 2007 ) found that Pb concentrations within resuspended road dust were higher at an inner

- city 

## 26

ring road and an industrial site compared with residential or construction sites. Li et al. ( 2009 ) observed 

## 27

that Pb concentration in PM 2.5 

### samples was 2.3

- 3.0 times higher near a bus depot than next to a rural



## 28

suburban road. Ondov et al. ( 2006 ) measured Pb

- PM 2.5 

### concentration at three Baltimore sites, one of 

## 29

which was industrial and the other two of which were considered “receptor” sites. Average Pb

- PM 2.5 

## 30

concentrations at the different sites were 8.3 ng/m 3 

### at the indus trial site and 1.9 ng/m 3 

### and 7.2 ng/m 3 

### at the 

## 31

receptor sites. As a group, these studies support the analysis of intra

- urban AQS data by illustrating that 

## 32

intra

- urban variability is most strongly related to type, strength, and location of sources. 

## 33May 20 11 3

- 70 DRAFT

– DO NOT CITE OR QUOTE 

# 3.5.2. Temporal Variability

The following section presents data for multi

- year trends and seasonal variability of Pb 

## 1

concentrations on a nationwide basis. The data presented here provide information on the success of Pb 

## 2

reduction efforts over past decades as well as on ar eas for continued attention with respect to Pb 

## 3

monitoring. The multi

- year trends illustrate changes in air Pb concentrations resulting from the phase

- out 

## 4

of leaded gasoline for automobiles and smaller reductions of industrial Pb usage. The seasonal variabi lity 

## 5

plots demonstrate changes in concentration within a given year, potentially related to climate or source 

## 6

variation. 

## 7

3.5.2.1. Multi

- year Trends

Figure 3

- 16 illustrates the trend in ambient air TSP

- Pb concentrations during the years 1980

- 2009. 

## 8

Over this time pe riod, average air Pb concentrations have declined by 91% from 1.3 μg / m 3 

### (in 1980) to 

## 9

0.12 μg / m 3 

### (in 2009). The median concentrations have declined by 97% from 0.87 μg / m 3 

### (in 1980) to 

## 10

0.025 μg / m 3 

### (in 2009). While the sharpest drop in Pb concentration occurred during 1980

- 1990 as a result 

## 11

of the phase

- out of Pb antiknock agents in on

- road fuel, a declining trend can also be observed between 

## 12

1990 and 2009 following reductions in industrial use and p rocessing of Pb, as described in Section 3.2.1. 

## 13

In 1990, the median Pb concentration was 0.19 μg / m 3 

### and the average Pb concentration was 0.55 μg / m 3 

### to 

## 14

yield 87% and 78% reductions, respectively, from 1990 to 2009. Average concentrations in these 

## 15

calculatio ns and in Figure 3

- 16, are heavily influenced by the source oriented monitors in the network. 

## 16

New Pb concentration data from expansion of the source oriented portion of the network in 2010 will 

## 17

allow for greater assessment of changes of Pb concentrations o n nationwide statistics and trends. 

## 18May 20 11 3

- 71 DRAFT

– DO NOT CITE OR QUOTE 

### Figure 3

- 1 6

. National trends in Pb concentration ( μg/m 3 

### ), all FRM monitors,
1980

- 2009. Average concentration is shown by the solid black line, median concentration is shown by the solid blue line, and 10th and 90th per centiles are plotted with dashed lines. The red lines on the plot illustrate former and current levels of the
Pb NAAQS.

Figure 3

- 1 7 and Figure 3

- 1 8 show ambient Pb concentrations from 1990 to 2009 for source 

## 1

oriented monitors and non

- source oriented monito rs, respectively. In both cases concentration data are 

## 2

consistent with a downward trend, and concentrations were considerably lower at the end of the period 

## 3

than at the beginning of the period. 

## 4 0 0.5

1 1.5

2 2.5

3 3.5

4TSP Concentration (μg/m3)

1978

- 2008 Level of the NAAQS

2009 Level of the NAAQSMay 20 11 3

- 72 DRAFT

– DO NOT CITE OR QUOTE 

### Figure 3

- 1 7

. National trends in Pb concentration ( μg/m 3 

### ), s ource oriented FRM monitors, 1990

- 2009. Average concentration is shown by the solid line, and the 10th and 90th percentiles are shown by
the dashed lines.

Figure 3

- 1 8

. National trends in Pb concentration ( μg/m 3 

### ), non

- source oriented FRM monitors, 1990

- 2009. Average concentration is shown by the solid line, and the 10th and 90th percentiles are shown by the dashed lines. 0 0.5

1 1.5

2 2.5

3TSP Pb Concentration (μg/m3)

1978

- 2008 Level of the NAAQS

2009 Level of the NAAQS

0 0.1 0.2 0.3 0.4 0.5 0.6TSP Pb Concentration (μg/m3)

2009 Level of the NAAQSMay 20 11 3

- 73 DRAFT

– DO NOT CITE OR QUOTE 

### For source oriented monitors, average concentrations decreased from 0.93 μg/m 3 

### to 0.23 μg/m 3 

## 1

(75% decline) and upper 90 th percent ile concentrations decreased from 2.5 μg/m 3 

### to 0.54 μg/m 3 

### (78% 

## 2

decline) over the 20

- year period. A portion of the decrease can be attributed to reductions in emissions 

## 3

from the Herculaneum, MO smelter between 2001 and 2002 ( U.S. EPA, 2010 ) . An abrupt decrease in 

## 4

average concentrations between these years is evident in Figure 3

- 17. Note that the number of monitors 

## 5

contributing to these statistics increased from 29 during 1990

- 1999 to 47 for 2000

- 2009. 

## 6

For non

- source oriented monitors, average con centrations decreased from 0.18 μg/m 3 

### to 

## 7

0.020 μg/m 3 

### (88% decline) and upper 90 th percentile concentrations decreased from approximately 0.49 

## 8

μg/m 3 

### to approximately 0.05 μg/m 3 

### (87% decline) over the 20

- year period. Average concentrations near 

## 9

stationary so urces were 5 to 12 times typical concentrations from non

- source oriented monitoring 

## 10

locations between 1990 and 1999; during the subsequent decade, average source oriented Pb 

## 11

concentrations were 8 to 22 times higher than non

- source concentrations ( U.S. EPA, 2010 ) . This 

## 12

differential likely reflects the absence of Pb emissions from automobiles during 2000

- 2009. The number 

## 13

of monitors contributing to these statistics increased from 29 during 1990

- 1999 to 53 for 2000

- 2009. 

## 14

When both source oriented and non

- source oriented monitoring sites are considered, average Pb 

## 15

concentrations decreased by 73% between 2001 and 2008 for maximum 3

- month average concentrations 

## 16

at 24 sites that are near large stationary sources and 101 sites that are not near stationary sources ( U.S. 

## 17

EPA, 2010 )

. 

## 18

3.5.2.2. Seasonal Variations

This section outlines seasonal variability among Pb monitors on a nationwide basis. Seasonal 

## 19

variation may provide insight related to differential influences of sources and climate throughout a year. 

## 20

Additionally, the magnitude of concentrations within the monthly data distributions and of variations 

## 21

between months sheds light on the influence of season as well as on differences between source oriented, 

## 22

non

- source orient ed, PM 10 

### , and PM 2.5 

### data. 

## 23

The average of Pb concentrations over all monitoring sites is higher in fall and lower in winter than 

## 24

other seasons. Monthly average Pb concentrations averaged over multiple sites and over 3 years from 

## 25

2007

- 2009 are shown in for P b

- TSP from source oriented monitors (Figure 3

- 19), Pb

- TSP from non



## 26

source oriented monitors (Figure 3

- 20), Pb

- PM 10 

### (Figure 3

- 21), and Pb

- PM 2.5 

### (Figure 3

- 22). For source 

## 27

oriented Pb

- TSP (Figure 3

- 19), monthly average concentrations were determined from betw een 146 and 

## 28

154 samples in each month. The highest monthly average concentrations were observed in March, April, 

## 29

and November, and exceeded 0.26 μg/m 3 

### . For non

- source oriented TSP (Figure 3

- 20), monthly average 

## 30

concentrations were determined from between 1 41 and 151 samples in each month. A winter minimum 

## 31

was observed with December, January, and February exhibiting the three lowest monthly average 

## 32May 20 11 3

- 74 DRAFT

– DO NOT CITE OR QUOTE 

### concentrations, each of which were below 0.015 μg/m 3 

### , but concentrations were similar between spring, 

## 1

summer an d fall months. 

## 2

Figure 3

- 19

. Monthly source oriented Pb

- TSP average ( μg/m 3 

### ) over 12 months of the year, 2007

- 2009. Box and whisker plots are used for each month, with the box comprising the interquartile

range of the data and the whiskers comprising the r ange within the 5th to 95th percentiles. The median is noted by the red line, and the blue star denotes the mean.concentration (μg/m3)

0.0 0.1 0.2 0.3 0.4 0.5 0.6 0.7 0.8 0.9 1.0 1.1 1.2 1.3 1.4 1.5

Month Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec 

##            May 20 11 3

- 75 DRAFT

– DO NOT CITE OR QUOTE 

### Figure 3

- 2 0

. Monthly non

- source oriented lead

- TSP average ( μg/m 3 

### ) over 12 months of the year, 2007

- 2009. Box and whisker plots are used
for each month, with the box comprising the interquartile range of the data and the whiskers comprising the range within the 5th to 95th percentiles. The median is noted by the red line, and the blue star denotes the mean.

Figure 3

- 2 1

. Monthly lead

- PM 10 

### aver age ( μg/m 3 

### ) over 12 months of the year,
2007

- 2009. Box and whisker plots are used for each month, with the box comprising the interquartile range of data and the whiskers comprising the range from 5th to 95th percentiles. The median is noted by the red lin e, and the blue star denotes
the mean.concentration (μg/m3)

0.00 0.01 0.02 0.03 0.04 0.05 0.06 0.07

Month Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec 

##           
concentration (μg/m3)

0.000 0.002 0.004 0.006 0.008 0.010 0.012 0.014 0.016 0.018 0.020 0.022 0.024 0.026 0.028

Month Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec 

##           

May 20 11 3

- 76 DRAFT

– DO NOT CITE OR QUOTE 

### Figure 3

- 2 2

. Monthly lead

- PM 2.5 

### average ( μg/m 3 

### ) over 12 months of the year

, 2007

- 2009

. Box and whisker plots are used for each month, with the box comprising the interquartile range of the data and
whiskers comprising the range from 5th to 95th percentiles. The median is noted by the red line, and the blue star denotes
the mean.

For both Pb

- PM 10 

### (Figure 3

- 21) and Pb

- PM 2.5 

### , (Figure 3

- 22) monthly average concentrations are 

## 1

considerably higher in the fall than in other seasons, with lowest the three highest monthly average 

## 2

concentrations observed in September, October, and November, and the average September concentration 

## 3

more than double the average December concentration. Pb

- PM 10 

### monthly average concentrations were 

## 4

determined from between 100 and 109 samples and Pb

- PM 2.5 

### from between 866 and 1,034 samples each 

## 5

month. Some of the Pb

- PM 2.5 

### concentrations were below Method Detection Limits and the absolute 

## 6

difference in monthly average concentrations between fall and other seasons for Pb

- PM 2.5 

### of 0.001 μg/m 3 

## 7

is extremely small. In spite of this the observed trends for PM 2.5 

### are consistent with the PM 10 

## 8

observations. Whether the seasonal trend for Pb

- PM 10 

### and Pb

- PM 2.5 

### differs from trends for both source 

## 9

oriented and non

- source oriented Pb

- TSP because of the difference in source proximity, sampling 

## 10

locations or difference in size range sampled is not clear. 

## 11

Although details of the seasonal trends varied with PM size range and source proximity, the data as 

## 12

a whole indi cate that average monthly concentrations in the fall were consistently higher than the lowest 

## 13

average monthly concentrations, and that average monthly concentrations in the winter were consistently 

## 14

lower than the highest average monthly concentration regar dless of PM size range or source proximity. 

## 15

Overall, there this indicates a clear tendency toward higher fall concentrations and lower winter 

## 16

concentrations. These results are consistent with observations at a single location by Melaku et al. ( 2008 ) 

## 17concentration (μg/m3)

0.000 0.001 0.002 0.003 0.004 0.005 0.006 0.007 0.008 0.009 0.010

Month Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec 

##            May 20 11 3

- 77 DRAFT

– DO NOT CITE OR QUOTE 

### that atmospheric Pb concentrations in fall were higher than in winter in an intensive study of urban 

## 1

Washington DC, but not consistent with their observations of higher s ummer concentrations than fall. 

## 2



# 3.5.3. Size Distribution of Lead

- Bearing PM

Size

- selective monitoring data from AQS and the literature is examined in this section to improve 

## 3

understanding of the size distribution of Pb

- bearing PM. This information informs monitor ing strategies 

## 4

because high content or correlation of PM 10 

### with TSP may allow for expanded usage of PM 10 

### within the 

## 5

FRM monitoring network. Additionally, size distribution data enhances understanding of the relationship 

## 6

between sources and characteristics of airborne Pb

- bearing PM. 

## 7

3.5.3.1. AQS Data Analysis

This section employs AQS data for Pb concentrations from co

- located TSP, PM 10 

### , and/or PM 2.5 

## 8

FRM monitors to analyze correlations and ratios of concentrations obtained from the different monitors. 

## 9

These data were used because relationships among the monitors provide information about the nature of 

## 10

Pb

- bearing PM at different locations (e.g., whether the mode is in the fine or coarse fraction). Correlations 

## 11

indicate the extent to which the size fractions vary togeth er in time, and the ratios signify the average 

## 12

proportion of the smaller fraction to the larger fraction (e.g., the ratio of PM 2.5 

### to PM 10 

### concentrations). 

## 13

Estimation of the size distribution of Pb

- bearing PM is possible at a limited number of monitoring 

## 14

sites where monitors having different size

- selective cut

- points are co

- located. Data for correlations 

## 15

between monitors and average concentration ratios are available per collocation site in Table 3A

- 13 in the 

## 16

Appendix . For the comparison between Pb

- TSP and Pb

- PM 10 

### , 27 sites were available for analysis. A 

## 17

summary of these data are provided in Table 3

- 6. Overall, the average correlation, ρ, was moderate, 

## 18

although the wide range across monitors indicates suggests site

- to

- site variability. The average ratio of 

## 19

Pb

- PM 10 

### to Pb

- TSP concentrations was relatively high at 0.88. When broken down by land type, the 

## 20

average ρ went up slightly for urban and city center land use, with a lower average ratio of concentrations. 

## 21

For suburban sites, the average ρ was reduced, but the average ratio of concentrations was near unity. 

## 22

Average ratio of concentrations greater than one suggest that some portion of the TSP was not collected 

## 23

due to instrument biases, as discussed in Section 3.4. 

## 24

Table 3

- 6

. Summary of comparison data for co
- located lead

- TSP and lead

- PM 10 

### monitors

.

Monitors Correlation Average Ratio PM 10 :TSP

Average Standard Deviation Range Average Standard Deviation Range

Overall 0.65 0.30 0.00

- 0.99 0.88 0.51 0.33

- 2.60 Urban and City Center 0.73 0.25 0.29

- 0.99 0.73 0.19 0.54

- 1.10

Suburban 0.55 0.33 0.00

- 0.98 0.99 0.74 0.33

- 2.60 Rural

- - - - - -May 20 11 3

- 78 DRAFT

– DO NOT CITE OR QUOTE 

### Forty

- five sites were available for analysis to compare Pb

- TSP with Pb

- PM 2.5 

### . A summary of these 

## 1

data are provided in Table 3

- 7. Overall, the average ρ was moderate. As for the TSP to PM 10 

### comparison, 

## 2

a wide range of correlations suggests that the tracking of the TSP and PM 2.5 

### time series was quite variable 

## 3

between sites and source influences. The average ratio of Pb

- PM 2.5 

### to Pb

- TSP concentrations was 

## 4

somewhat high, but with several monitors having ratios above one, sampler bias may have influenced 

## 5

these statistics. When broken down by land type, the average ρ was somewhat reduced for urban and city 

## 6

center land use. For suburba n sites the average ratio of concentrations was lower, with some of the same 

## 7

sampling bias issues. For rural sites, the average ρ was much higher

. Correlation between PM 2.5 

### and TSP 

## 8

may suggest commonality of sources or processes influencing both size fract ions at the same time. 

## 9

Table 3

- 7

. Summary of comparison data for co
- located Pb

- TSP and Pb

- PM 2.5 

### monitors

Monitors Correlation Average Ratio PM 2.5 :TSP

Average Standard Deviation Range Average Standard Deviation Range

Overall 0.50 0.25 0.03

- 0.95 0.79 0.49 0.11

- 2.34 Urban and City Center 0.43 0.26 0.03

- 0.90 0.75 0.54 0.11

- 2.34

Suburban 0.55 0.23 0.16

- 0.95 0.66 0.29 0.36

- 1.57 Rural 0.71 0.24 0.37

- 0.90 1.36 0.39 0.83

- 1.76



### Pb

- PM 10 

### and Pb

- PM 2.5 

### data were compared for 50 sites. A summary of these data are provided in 

## 10

Table 3

- 8. Overall, the average ρ was moderately high, with perfect correlation at a Providence, RI site 

## 11

(AQS ID: 440070022 ) and very high correlation at several other sites. The ave rage ratio of Pb

- PM 2.5 

### to 

## 12

Pb

- PM 10 

### concentrations was 1.01, suggesting that some bias existed in the data so that PM 2.5 

### was often 

## 13

higher than PM 10 

### concentration. Data for urban and city center and suburban sites were similar for 

## 14

correlations and average rat ios of concentration. For rural sites, the average ρ was slightly lower, and the 

## 15

range of ratios of concentrations showed that PM 2.5 

### concentrations were almost always higher than PM 10 

## 16

concentrations, suggesting bias among the rural monitors. 

## 17

Table 3

- 8

. Su mmary of comparison data for co
- located Pb

- PM 10 

### and Pb

- PM 2.5 

### monitors

.

Monitors Correlation A verage Ratio PM 2.5 : PM 10
Average Standard Deviation Range Average Standard Deviation Range

Overall 0.76 0.22 0.15

- 1.00 1.01 0.37 0.45

- 2.15 Urban and City Center 0.72 0.24 0.15

- 1.00 0.85 0.38 0.45

- 2.15

Suburban 0.81 0.17 0.34

- 0.98 1.08 0.35 0.61

- 1.89 Rural 0.66 0.28 0.22

- 0.99 1.23 0.29 0.97

- 1.63



### In a few cases, Pb

- TSP, Pb

- PM 10 

### , and Pb

- PM 2.5 

### monitors were co

- located simultaneously, so that 

## 18

more information regarding size distribution can be discerned. For example, in urban Jefferson County, 

## 19

AL (AQS: 010730023), the average ratio of PM 2.5 

### to TSP was 0.84, while the average ratio of PM 10 

### to 

## 20

TSP was 0.80 for the years 2005

- 2006. These reported values suggested that most of the PM was in the 

## 21May 20 11 3

- 79 DRAFT

– DO NOT CITE OR QUOTE 

### fine mode. But, the data also indicated instrumentation bias with either the PM 2.5 

### or PM 10 

### monitors or 

## 1

both, because the average concentration ratio was great er than one. In a suburban portion of Cook County, 

## 2

IL (AQS: 170314201), the PM 2.5 

### to TSP average ratio was reported to be 0.36, and the PM 10 

### to TSP 

## 3

average ratio was reported to be 0.39. This suggested that the majority of the sample mass was from 

## 4

particles larger than 10 μm, but that the distribution might have been bimodal since there was little 

## 5

difference between PM 2.5 

### and PM 10 

### . Note that o nly some of the years reported for all three monitors 

## 6

overlapped. In suburban Wayne, MI (AQS: 261630033), the PM 2.5 

### to TSP average ratio was reported to 

## 7

be 0.49, and the reported PM 10 

### to TSP ratio was reported to be 0.84. This suggests a smoother distribut ion 

## 8

with one mode likely in the large fine or small coarse region of the distribution. Note that the PM 2.5 

### to 

## 9

TSP collocation includes one year of data more than the PM 10 

### to TSP collocation. Taken together, these 

## 10

findings imply that the Pb

- bearing PM size distribution varies substantially by location. 

## 11

3.5.3.2. Size Distribution Studies in the Literature

Several studies in the literature have been designed to characterize the size distribution of Pb 

## 12

concentrations in the vicinity of sources. The following section des cribes studies that have measured more 

## 13

than one size fraction of Pb

- bearing PM in the vicinity of industrial, urban, and/or traffic

- related sources. 

## 14

Some discussion of the variability of size distribution over space and time is also provided. 

## 15

Size distribu tions of Pb

- bearing PM have been measured near several industrial sites. For example, 

## 16

Bein et al. ( 2006 ) measured the size distribution of Pb in PM from the Pittsburgh superfund site using 

## 17

rapid single particle mass spectrometry and a Multi

- Orifice Uniform Deposit Impacto r (MOUDI). The 

## 18

Pittsburgh, PA, superfund site had seventeen major PM sources within a 24

- km radius. Bein et al.’s 

## 19

( 2006 ) measurements yielded different results on different days, with a bimodal distribution with modes 

## 20

around 140 nm and 750 nm during an October, 2001 measu rement and a single dominant mode around 

## 21

800 nm during a March, 2002 measurement. Differences in the size distributions could have been related 

## 22

to differences among wind speed, wind direction, and source contributions on the respective dates. Singh 

## 23

et al. ( 2002 ) measured the mass distribution of Pb

- PM in the coarse and fine PM size ranges for the 

## 24

Downey industrial site along the Alameda industrial corridor in Los Angeles and a site approxi mately 

## 25

90 km downwind in Riverside, CA. At the industrial site, the Pb

- PM size distribution was unimodal with 

## 26

a concentration peak in the 100

- 350 nm size range with 34% of the particles in this size bin. At the 

## 27

downwind site, a bimodal distribution was obs erved with peaks in the 2.5

- 10 μm bin and the 350 nm

- 1 

## 28

μm bin, comprising 42% and 26% of the mass measured as PM 10 

### , respectively. Pb in the fine range only 

## 29

comprised 13% of the particles in the 100

- 350 nm bin. The authors suggested that higher wind speeds in 

## 30

Riverside compared with the Downey site are effective in resuspending larger particles from the ground to 

## 31

create a peak in the coarse mode of the distribution. 

## 32May 20 11 3

- 80 DRAFT

– DO NOT CITE OR QUOTE 

### Industrial operations associated with Pb emissions include metal works and incineration. Dall ’Osto 

## 1

et al. ( 2008 ) measured the size distribution of Pb emissions from a steel works facility in a coastal town 

## 2

within the United Kingdom (U.K.). A MOUDI was employed to measure concentrations in the coarse to 

## 3

fine PM size range. The size distribution was multimodal with a primary mode around 1 μm, a secondary 

## 4

mode around 300 nm, and a very small additional mode around 5 μm. This multimodal distribution was 

## 5

thought to be associated with sintering and steel working processes, from which Pb was emitted. 

## 6

Weitkamp et al. ( 2005 ) measured Pb

- bearing PM 2.5 

### concentrations across the river from a coke plant in 

## 7

the Pittsburgh, PA area. Pb was measured to comprise 0.088% of the PM 2.5 

### mass, and background



## 8

corrected Pb mass concentration was reasonably correlated with background

- corrected PM 2.5 

### mass 

## 9

concentration (R 2 

### = 0.55). Pekey et al. ( 2010 ) measured PM 2.5 

### and PM 10 

### concentrations in a heavily 

## 10

industrialized area of Kocaeli City, Turkey and obtained an average PM 2.5 

### concentration of 47 ng/m 3 

## 11

during summer and 72 ng/m 3 

### during winter. Average PM 10 

### concentration was 78 ng/m 3 

### during summer 

## 12

and 159 ng/m 3 

### during winter, to produce PM 2.5 

### /PM 10 

### ratios of 0.60 during summer and 0.45 during winter. 

## 13

Traffic can be a source of re suspension of Pb from deposited contemporaneous wheel weights or 

## 14

industrial emissions or historic sources via traffic

- induced turbulence. Sabin et al. ( 2006 ) compared the 

## 15

size distribution of coarse Pb

- PM captured at an urban background site and at a location 10 m from the I



## 16

405 Freeway in the southern California air basin; data from Sabin et al. ( 2006 ) are displayed in Figure 3

- 

## 17

23. For both the urban background and near

- road sites, the largest fraction was from PM sampled below 

## 18

the 6 μm cut p oint, but the near

- road Pb

- PM distribution appeared bimodal with a mode in the largest size 

## 19

fraction. Over all size fractions, the near

- road site had a Pb concentration of 17 ng/m 3 

### , compared with an 

## 20

urban background concentration of 9.7 ng/m 3 

### . Sabin et al. ( 2006 ) point out that the freeway tends to be a 

## 21

source of very large particles that are dispersed via the turbulent moti on of the vehicular traffic. In a near



## 22

road study conducted in Raleigh, NC, Hays et al. ( 2011 ) note that the concentration of Pb in ultraf ine, 

## 23

fine, and coarse size ranges was roughly constant at 50 mg/kg. The Pb samples from Hays et al. ( 2011 ) 

## 24

were highly correlated with As samples (ρ = 0.7, p < 0.0001); both Pb and As are found in wheel weights. 

## 25

Likewise, the Pb samples were not well correlated with crustal elements in the coarse size distribution, so 

## 26

it is more likely that resuspended Pb originated from wheel weights rather than historic Pb on

- road 

## 27

gasoline emissions. Pb dust was shown by Bukowiecki et al. ( 2010 ) to be significantly higher at roadside 

## 28

samples compared with urban background when the PM was in the coarse mode, measured as PM 10

- 2.5 

### , 

## 29

but not for fine modes measured as PM 2.5

- 1 

### and PM 1

- 0.1


### . Chen et al. ( 2010 ) measured Pb in PM 10 

### , PM 2.5 

### , 

## 30

and PM 0.1 

### at a roadside location and in a tunnel in Taipei, Taiwan in 2008. While roadside and tunnel 

## 31

concentra tions of PM 10 

### and PM 2.5 

### were roughly equivalent, Pb in PM 0.1 

### was approximately 15 times 

## 32

higher than by the roadside. The authors suggest that particle

- bound Pb was emitted from on

- road 

## 33

gasoline and diesel engines. This could possibly be attributed to resid ual Pb in unleaded gasoline. 

## 34May 20 11 3

- 81 DRAFT

– DO NOT CITE OR QUOTE 0 10 20 30 40 50 60 70 80

< 6 μ m 6

- 11 μ m 11

- 20 μ m 20

- 29 μ m > 29 μ mPercent of Total Pb-PM Sampled

Urban Background Near

- Road

Source: Adapted, with permission from Elsevier Publishing, from Sabin et al. ( 2006 )

. 

### Figure 3

- 2 3

. Comparison of urban background and near

- road size fractions
of lead

- bearing PM.

Several studies have suggested that near

- road ambient air Pb samples are derived from non

- road 

## 1

sources. Harrison et al. ( 2003 ) measured the distribution of Pb in PM 10 

### at a roadside sampler in 

## 2

Birmingham, U.K.. The size distribution was unimodal with approximately 2% of the Pb mass 

## 3

(0.5 ng/m 3 

### ) above the 10 μm cut point, 12% (3 ng/m 3 

### ) in the 2

- 10 μm bin, 9% (2 ng/m 3 

### ) in the 1

- 2 μm 

## 4

bin, 53% of the Pb mass (14 ng/m 3 

### ) in the 0.2

- 1.0 μm bin, and 24% (7 ng/m 3 

### ) collected below the 0.2 μm 

## 5

cut point. Regression analysis against NO X 

### concentration in the Harris on et al. ( 2003 ) paper provided a 

## 6

weak indication that Pb

- PM 0.2 

### was associated with traffic (β = 0.067, R 2 

### = 0.38) as well as PM 10 

### (β = 0.26, 

## 7

R 2



### = 0.35). Br üggemann et al. ( 2009 ) observed a unimodal Pb size distribution with 51% of the mass in 

## 8

the 0.42

- 1. 2 μm size bin. During winter, Pb concentrations were twice as high as during the summer, and 

## 9

they were also higher when winds blew from the east. Brüggemann et al. ( 2009 ) suggested that this 

## 10

finding reflected coal burning sources rather than road dust resuspension. Wang et al. ( 2006 ) observed a 

## 11

bimodal Pb distribution in a heavily trafficked area of Kanazawa, Japan with incineration and generation 

## 12

facilities also nearby. They observed a bimodal distribution with modes at the 0.65

- 1.1 μ m and the 3.3

- 4.7 

## 13

μ m size bins. Wang et al.’s ( 2006 ) source apportionment work in this study suggested that the fine mode 

## 14

derives from incineration and combustion of oil and coal. 

## 15

Spatial and temporal concentration variability is also reflected in varying Pb

- PM size distributions 

## 16

within and betwe en cities. Martuzevicius et al. ( 2004 ) measured the size distribution of Pb in Cincinnati, 

## 17

OH at the city center site using a MOUDI and showed it to be bimodal with a primary peak at 0.56 μm 

## 18May 20 11 3

- 82 DRAFT

– DO NOT CITE OR QUOTE 

### and a slightly smaller secondary peak at 5.6 μm. Moreno et al. ( 2008 ) measured Pb concentrations in 

## 1

PM 2.5 

### and PM 10 

### at urban, suburban, and rural s ites around Mexico City, Mexico to illustrate differences 

## 2

among the land use categories. At the urban site, average Pb

- PM 2.5 

### concentration was 30 ng/m 3 

### during the 

## 3

day and 92 ng/m 3 

### at night, and average Pb

- PM 10 

### concentration was 59 ng/m 3 

### during the day and 

## 4

162 ng/m 3 

### at night, to yield PM 2.5 

### /PM 10 

### ratios of 0.51 during the day and 0.57 at night. At the suburban 

## 5

site, average Pb

- PM 2.5 

### concentration was 15 ng/m 3 

### during the day and 34 ng / m 3 

### at night, and average Pb



## 6

PM 10 

### concentration was 24 ng/m 3 

### during the day and 42 ng / m 3 

### at night, to yield PM 2.5 

### /PM 10 

### ratios of 0.63 

## 7

during the day and 0.81 at night. Rural measurements were only made for Pb

- PM 10 

### and averaged 6 ng/m 3 

## 8

during the day and 5 ng/m 3 

### at night. Goforth et al. ( 2006 ) measured TSP and PM 2.5 

### in rural Georgia and 

## 9

observed a PM 2.5 

### concentration of 6 ng / m 3 

### and a TSP concentration of 15 ng/m 3 

### . Makkonen et al. ( 2010 ) 

## 10

measured concentrations of Pb in PM 1 

### , PM 2.5 

### , and PM 10 

### during a spate of wildfires in rural southeastern 

## 11

Finland. They found that the ratio of PM 1 

### /PM 10 

### varied substantially from day to day (examples provided 

## 12

of 64% on 8/14/07 and 35% on 8/25/07, with PM 2.5 

### /PM 10 

### ratio of 51% on 8/25/07), and they attributed the 

## 13

highest concentrations to long

- range transport of wildfire emissions via southerly winds; variability in 

## 14

concentration and ratios was related to shifting wind conditions. Birmili et al. ( 2006 ) compared 

## 15

concentrations of Pb in PM at various traffic and background sites in Birmingham, U.K.. captured at the 

## 16

stage below a 0.5 μm cutpoint and on the 1.5

- 3.0 μm stage for near

- road, in a traffic tunnel, and remote 

## 17

and urban background sites. The highe st concentrations were measured in the tunnel, at 3.3 ng/m 3 

### for Pb



## 18

PM 0.5 

### and 10 ng/m 3 

### for Pb

- PM 1.5

- 3.0 

### . Roadside concentrations were low. During the day, Birmili et al. 

## 19

( 2006 ) measured 0.4 ng/m 3 

### for Pb

- PM 0.5 

### and 1.2 ng/m 3 

### for Pb

- PM 1.5

- 3.0 

### . At night, roadside concentrations 

## 20

reduced to 0.17 ng/m 3 

### for Pb

- PM 0.5 

### and 0.6 ng/m 3 

### for Pb

- PM 1.5

- 3.0 

### . In contrast, urban bac kground was 

## 21

more enriched in the finer size fraction, with concentrations of 5.4 ng/m 3 

### for Pb

- PM 0.5 

### and 0.84 ng/m 3 

### for 

## 22

Pb

- PM 1.5

- 3.0 

### . Remote background concentrations were on 0.16 ng/m 3 

### for Pb

- PM 0.5 

### and 0.03 ng/m 3 

### for Pb



## 23

PM 1.5

- 3.0 

### . Brüggemann et al. ( 2009 ) measured roadside distribution of Pb in PM in Dresden, Germany to 

## 24

analyze the effect of season and direction of the air mass. For all data combined as well as for data broken 

## 25

down by season or by wind direction, it was found that the data followed a unimodal distribution with a 

## 26

peak at the 0.42

- 1.0 μm size bin. The distribution of data along the curves did not change substantially 

## 27

under the different conditions examined. When winds came from the east, the total concentration was 

## 28

approximately 22 ng/m 3 

### , compared with a concentration of approximately 13 ng/m 3 

### when winds came 

## 29

from the west. Total winter con centrations of Pb were approximately 26 ng/m 3 

### , while summertime 

## 30

concentrations were roughly 11 ng/m 3



### . 

## 31



# 3.5.4. Lead Concentrations in a Multipollutant Context

The correlations between Pb and copollutant concentrations were investigated because correlation 

## 32

may indicate commonality of sources among the pollutants. For example, correlation between Pb and SO 2 

## 33May 20 11 3

- 83 DRAFT

– DO NOT CITE OR QUOTE 

## US Overall

Spearman Correlation Coefficient -1.0 -0.5 0.0 0.5 1.0 

### may suggest common industrial sources. Correlation between Pb and NO 2 

### or CO may suggest roadway 

## 1

sources, such as trace Pb in unleaded on

- road gasoline or resuspension of Pb wheel weights or 

## 2

contaminated soil. Additionally, seasonality can influence correlations, potentially from differences 

## 3

among sources or the contaminants ’ responses to climate differences. 

## 4

Pb concentrations exhibit varying degrees of association with other criteria pollutant 

## 5

concentrations. Spearman correlations of monitored TSP

- Pb concentrations with concentrations of other 

## 6

criteria pollutants are summari zed in Figure 3

- 24 for 2007

- 2008 data from 129 monitoring sites, and in 

## 7

Figure 3

- 25 for 2009 data from 16 monitoring sites. At most sites, Pb monitors are co

- located with 

## 8

monitors for other criteria pollutants, but monitoring the full suite of criteria pol lutants at a single 

## 9

monitoring site is rare. As a result the number of observations for each copollutant varies, ranging from 

## 10

44 non

- source oriented sites for the association of Pb with SO 2 

### to 81 sites for the association of Pb with 

## 11

PM 10 

### . In Figure 3

- 24, a nd fewer for each copollutant in Figure 3

- 25. Each of these figures illustrates co



## 12

pollutant correlations across the U.S. Additionally, seasonal correlations between Pb and co

- pollutants are 

## 13

provided in Figures 3 A

- 19 through 3 A

- 21 in the Chapter 3 Appendix , with seasonal co

- pollutant 

## 14

measurement data from the literature (Table 3 A

- 14). As evident in each figure, there were considerably 

## 15

fewer source

- oriented sites available for co

- located comparisons. 

## 16

Source, SO 2

Non

- Source, SO 2

Source, PM 2.5

Non

- Source, PM 2.5

Source, PM 10

Non

- Source, PM 10

Source, O 3

Non

- Source, O 3

Source, NO 2

Non

- Source, NO 2

Source, CO

Non

- Source, CO



### Figure 3

- 2 4

. Correlations of monitored Pb

- TSP concentration with copollutant concentrations, 2007

- 2008.May 20 11 3

- 84 DRAFT

– DO NOT CITE OR QUOTE 

### US Overall

Spearman Correlation Coefficient -1.0 -0.5 0.0 0.5 1.0 Non

- Source, SO 2

Source, PM 2.5

Non

- Source, PM 2.5

Non

- Source, PM 10

Non

- Source, O 3

Non

- Source, NO 2

Non

- Source, CO



### Figure 3

- 2 5

. Correlations of monitored P b

- TSP concentration with co

pollutant concentrations, 2009.

Overall, Pb was most strongly associated with PM 2.5 

### , PM 10 

### and NO 2 

### (median R = 0.38 to 0.41), 

## 1

with positive Spearman correlation coefficients observed at nearly all sites. However, Pb was just as 

## 2

strongly associated with CO in fall and winter (median R = 0.48 to 0.58). Such cor relations may suggest 

## 3

common sources affecting the pollutants. Overall correlation coefficients between Pb and SO 2 

### and 

## 4

between Pb and CO were also positive at most sites, but associations were generally weaker (median R = 

## 5

0.29 for CO, 0.17 for SO 2 

### ). The po orest associations were observed between Pb and O 3 

### (median R = 

## 6

0.00). Although the overall associations of Pb concentration with PM 10 

### and PM 2.5 

### concentrations were 

## 7

similar, the association with PM 10 

### was stronger in the spring and the association with PM 2.5 

### stronger in 

## 8

summer and fall. The strongest associations between Pb and other criteria pollutants were observed in fall 

## 9

and winter, and the weakest in summer. 

## 10

The relationship between Pb and other species in PM 2.5 

### is explored in Figure 3

- 26, which describe s 

## 11

data from 3 years of CSN results. These data provide a national perspective on relationships between the 

## 12

various bulk and elemental species monitored in the CSN network. The strongest association was with Zn 

## 13

(median R = 0.51). Br, Cu, and K concentration s also exhibited moderately strong associations with Pb 

## 14

concentrations (median R = 0.40 to 0.41). Such correlations may suggest some common sources affecting 

## 15

the pollutants. Other species more useful for as diagnostic indicators of crustal, general combust ion, 

## 16

industrial emission, and coal combustion processes exhibited weaker, but still remarkable associations 

## 17

with Pb, including crustal elements (median R = 0.32), EC (median R = 0.32), Mn and Fe (median R = 

## 18

0.32 and 0.34, respectively), S (median R = 0.27) , and Se (median R = 0.27). Except for S, in summer 

## 19May 20 11 3

- 85 DRAFT

– DO NOT CITE OR QUOTE 

### associations with each of these species were much weaker than in other seasons, with a Spearman 

## 1

Correlation Coefficient greater than R = 0.3 observed only for Zn (median R = 0.37). The weakest 

## 2

association s were with As, Cl, Hg, Ni, NO 3 

### , and Na (median R =

- 0.03 to 0.10). 

## 3May 20 11 3

- 86 DRAFT

– DO NOT CITE OR QUOTE 

# US Overall

Spearman Correlation Coefficient -1.0 -0.5 0.0 0.5 1.0 r r Zinc

Volatile Nitrate

Vanadium

Titanium
Sulfur

Sulfate

Sodium Ion

Sodium
Silicon

Selenium

Potassium Ion

Potassium

Organic Carbon, Blank Adjusted

Non

- Volatile Nitrate

Nitrate

Nickel Mercury
Manganese Magnesium

Iron
Elemental Carbon
Crustal

Copper
Chromium

Chlorine
Carbonate Carbon
Calcium

Cadmium

Bromine

Arsenic

Ammonium

Aluminum



### Figure 3

- 2 6

. Correlations of monitored lead

- PM 2.5 

### concentration with copollutant concentrations, 2007

- 2009.May 20 11 3

- 87 DRAFT

– DO NOT CITE OR QUOTE 

# 3.6. Ambient Lead Concentrations in Non

- Air Media and Biota

There have been some major recent research efforts to characterize geographic and temporal trends 

## 1

in Pb concentrations in across a variety of environmental media and biota

. In general these concentrations 

## 2

reflect the decreases observed in atmospheric Pb concentrations due to reduced on

- road Pb emissions

. 

## 3

The 2006 Pb AQCD ( U.S. EPA, 2006 ) describes several studies showing highe r Pb concentrations 

## 4

in plants grown in Pb contaminated soil related to mine spoils, smelting operations, sludge amendment, 

## 5

contaminated irrigation water, and Pb containing agro

- chemicals. Pb accumulation occurs more readily 

## 6

for Pb salts applied to soils th an for sewage sludge or fly ash. Root uptake is the dominant means of 

## 7

accumulation, and it is strongly influenced by pH. Root vegetables are the most strongly affected, and 

## 8

fruits and grains are the least susceptible. More Pb is also generally found in roo ts than in other parts of 

## 9

the plant. 

## 10

The 2006 Pb AQCD ( U.S. EPA, 2006 ) identified ingestion and water intake as major routes of Pb 

## 11

exposure for aquatic organisms, and it identified food, drinking water, and inhalation as major routes of 

## 12

exposure for livestoc k and terrestrial wildlife

. The 2006 Pb AQCD ( U.S. EPA, 2006 ) reports data from the 

## 13

U.S. Geologic Service National Water

- Quality Assessment (NAWQA), which are updated every ten years

. 

## 14

In the NAWQA survey, maxima concentrations in surface waters, sediments, a nd fish tissues were 

## 15

30 μg/L, 12,000 mg /kg, and 23 mg /kg, respectively, compared with median values of 0.50 μg /L, 

## 16

28 mg /kg, and 0.59 mg /kg

. Some of the highest levels of Pb contamination occur near major sources, like 

## 17

smelters, and fatal doses have been me asured in tissue from sheep and horses near sources

. High levels in 

## 18

cattle have also been observed

. Wildlife in urban areas tend to contain higher Pb concentrations than in 

## 19

rural areas, and higher Pb accumulations have been observed for aquatic organisms l iving in polluted 

## 20

coastal zones than in the open sea

. Ingestion of deposited Pb

- PM on plant surfaces was consistently 

## 21

observed to be more important than Pb accumulated from soil

. Some important variations between 

## 22

animals have been observed, and ruminants a ppear to be less susceptible to Pb uptake than other animals

. 

## 23

Uptake of Pb by lowest trophic levels, including invertebrates, phytoplankton, krill, were described as the 

## 24

most important means of introduction into food chains

. Elevated Pb levels have been ob served in aquatic 

## 25

organisms that feed from sediments when the sediments contain appreciable Pb

. In shrimp, a substantial 

## 26

fraction of Pb can be absorbed from prey, and considerably more accumulated Pb from food has been 

## 27

observed to be irreversibly retained than is the case for dissolved Pb from water

. These examples all 

## 28

illustrated that substantial Pb uptake by livestock and wildlife readily occurs in Pb contaminated 

## 29

environments. 

## 30May 20 11 3

- 88 DRAFT

– DO NOT CITE OR QUOTE 

# 3.6.1. Soils

Several studies suggest that soil can act as a reservoir for contemporan eous and historical Pb 

## 1

emissions. In a recent review of soil data collected from 90 U.S. cities, Mielke et al. ( 2010b ) cited studies, 

## 2

some of which were 35 years old but many from the last 15 years, reporting that median soil Pb 

## 3

concentrations ranged from 16 to 262 mg/kg, with maximum levels ranging from 461 to 348,000 mg/kg 

## 4

(see Table 3

- 9). Soil Pb was thought to origina te from present

- day sources, such as industry, debrided 

## 5

paint, and piston engine aircraft fuel, as well as historic sources, such as on

- road gasoline emissions, as 

## 6

described in Section 3.2. 

## 7

Emissions trends have shown that industrial activities are now on e of the largest sources of Pb 

## 8

following phase out of Pb in on

- road gasoline. Pruvot et al. ( 2006 ) compared urban and agricultural soils 

## 9

near a closed Pb smelter with soils in similar environments not exposed to smelter emissions in northern 

## 10

France. For samples near the smelter, Pruvot et al. ( 2006 ) observed that median soil Pb levels in lawns 

## 11

were roughly 2 times higher, while kitchen garden soil Pb concentrations were 10 times higher and 

## 12

agricultural soil Pb was almost 15 times highe r than soil not exposed to smelter emissions. In soil samples 

## 13

obtained near a defunct smelter in El Paso, TX, in 1999 and 2005, Pingatore et al. ( 2009 ) found that TSP 

## 14

concentration was predicted s trongly by concentrations of Pb

- humate, which is created by sorption of Pb 

## 15

onto humic substances in soil. Spalinger et al. ( 2007 ) compared soil Pb samples from surrounding towns 

## 16

with those from the Bunker Hill Superfund remediation site in Idaho. Median background soil Pb 

## 17

concentrations was 48 mg/kg, while the median soil Pb concentration at Bunker Hill was 245 mg/kg. 

## 18

Table 3

- 9

. Outdoor soil Pb levels in various cities within the U.S.

Study Year n Min a Med a Max a

Background soil Pb, U.S. 2001 1

, 319 16.5 City

- State Los Angeles, California 2010 550 9 216

, 174 Los Angeles, California 1995 343 Chicago Illinois 2008 57 Chicago Illinois 1987 667 Chicago

- Urban Parks 1986 255 12 262 1

, 312 Chicago

- Suburban Parks 1986 245 12 87 1

, 637 Illinois, Rural Parks 1986 177 12 37 937 Detroit, Michigan 2003 59 13 189 1

, 345 Detroit

- Suburbs, Michigan 2003 76 4 16 810 Pontiac, Michigan 2003 38 15 86 495 Oakland, California 1995 358 7 347

, 900 Alameda, California 1993 138 22 3

, 187 Boston, Massachusetts 1988 195 7 13

, 240 Miami, Florida 2004 240 2 1

, 060 Seattle, Washington 1991 51 150 74

, 000 Washington, D.C. 1995 240 12 6

, 015 M inneapolis/St. Paul, Minnesota 1984 90 5 7

, 650 Minneapolis, Minnesota 1988 898 1 230 20

, 136 St. Paul, Minnesota 1988 832 1 170 7

, 994 Duluth, Minnesota 1988 229 1 144 11

, 110 St. Cloud, Minnesota 1988 124 1 41 1

, 952 Rochester, Minnesota 1988 165 1 25 1

, 930 Outstate farms Minnesota 1988 781 1 31 7

, 111 Cleveland, Ohio 2006 50 19 811 Baltimore, Maryland 2008 122 0.01 5

, 620May 20 11 3

- 89 DRAFT

– DO NOT CITE OR QUOTE Study Year n Min a Med a Max a

Baltimore, Maryland 1983 422 1 100 10

, 900 Indianapolis, Indiana 2005 116 46 565 Milwaukee, Wisconsin summary data 1994 372 1 160 880

Milwaukee , Wisconsin cent city + North and South Side 1994 256 1 240 7

, 220

Milwaukee, W isconsin suburbs 1994 122 1 50 1

, 780 Cincinnati, Ohio 1990 60 2 3

, 166 Cincinnati, Ohio

- Childcare centers 2008 69 17 4

, 636 Cincinnati, Ohio 1990 60 Sacramento, California 1995 232 57 229 320

, 834 Tampa, Florida 1994 146 <1 100 9

, 160 New Orleans, Louisiana Survey 1 2005 4

, 026 18 134 183

, 588 New Orleans, L ouisiana Survey 2 2005 5

, 467 3 100 52

, 798 Louisiana Orleans Parish 1997 ~1

, 540 <25 >200 <1

, 000 L ouisiana Lafourche Parish 1997 ~190 <25 Albuquerque, New Mexico 1981 43 3 5

, 280 Omaha, Nebraska 1979 176 Dayton, Ohio 1983 22 22 461 El Paso, Texas 2009 94 <20 8700 Honolulu, H awaii 1988 18 Charleston, So uth Carolina 1975 164 9 7

, 890 New Haven, Connecticut 1982 487 30 7

, 000 Corpus Christi, Texas 1987 485 21 2

, 969 Pueblo, Colorado 2006 33 Connecticut 2008 174 <10 2

, 200 Gainesville, Florida 2004 202 2.13 1

, 091 Champaign Illinois 1976 116 20 1

, 061 Louisiana and Minnesota 1993 6

, 342 Urban

- rural comparisons S outheastern. Michigan 2004 171 3 7

, 400 Mt. Pleasant, Michigan 1992 189 100 16

, 839 Syracuse, New York 2009 2

, 998 45 Syracuse, New York 2002 162 Syracuse, New York 2002 194 80 (GM) Lubbock, Texas 2008 52 35 Maine urban soils 1989 100

a Minimum, median, and maximum values are reported in units of mg/kg for individual cities cited in the review paper.

Source: Used with permission from Elsevier Publishing, Mielke et al. ( Mielke et al., 2010b )

.



### Several studies explore the relationship between soil Pb concentration and land use. For example, 

## 1

the Mielke et al. ( 2010b ) review also found that soil Pb concentrations tended to be higher within inner



## 2

city communities compared with neighborhoods surrounding city outskirts. Laidlaw and Filippelli ( 2008 ) 

## 3

displayed data for Indianapolis, IN showing the Pb concentration at the soil surface had a smoot hed 

## 4

“bull’s eye” pattern, which suggested that the Pb in soil is continually resuspended and deposited within 

## 5

the urban area so that smooth air and soil concentration gradients emanating from the city center could be 

## 6

created over time. Cities generally hav e a similar pattern consisting of larger quantities of Pb accumulated 

## 7

within the inner city and smaller quantities of Pb in outer cities (i.e. near the outskirts or suburban areas) 

## 8

( Filippelli & Laidlaw, 2010 ) . Similarly, Filippelli et al. ( 2005 ) reported soil Pb concentration distribution 

## 9

to have a maximum at the center of Indianapolis, IN, around the location where two interstate highways 

## 10

intersect, and to decrease w ith distance away from the center. However, the spatial distribution of Pb was 

## 11

presumed to be smoothed over time from resuspension and deposition with contributions from historic 

## 12

sources of on

- road gasoline and Pb paint. In this paper, soil Pb concentratio ns were also shown to 

## 13

decrease with distance from roadways, but the levels were roughly four times higher in urban areas 

## 14

compared with suburban areas. This is also illustrated for urban scale Pb accumulation in New Orleans, 

## 15May 20 11 3

- 90 DRAFT

– DO NOT CITE OR QUOTE 

### LA in Figure 3

- 27. Brown et al. ( 2008 ) also measured soil Pb concentration along three transects of 

## 1

Lubbock, TX and observed that soil Pb decreased with increasing distance from the city center, which 

## 2

was the oldest part of the city. 

## 3

Source: Used with permission from Elsevier Publishing, Mielke et al. ( 2007 ) 

### Figure 3

- 2 7

. Map of median Pb content in soil in New Orleans. At the urban scale, Pb quantities are largest within the inner

- city residential communities that surround the Central Business District
where pavement and concrete cover the soil. Note the several orders of magnitude difference between the interior and the exterior areas of the city.

The amount of Pb within the inner

- city is likely not from a single source but instead composed of 

## 4

all modern an d historic sources of Pb dust that have been released in the city including Pb from several 

## 5

local industries, Pb dust from pulverized wheel weights, deteriorated Pb

- based paint, Pb additives to on



## 6

road gasoline, and defunct incinerators that once dotted Ne w Orleans prior to being shut down by EPA in 

## 7

the early 1970’s. Similarly, Mielke et al. ( 2008 ) compared soil Pb concentrations for public and private 

## 8

housing at the center and outer sections of New Orleans and found that median and maximum soil Pb 

## 9

concentrations were substantially higher in the city center compared with the outer portions of th e city. 

## 10

This study also found that private residences had higher soil Pb compared with public housing. In a 

## 11

separate study to examine surface soil Pb loading and concentration on 25 properties in New Orleans, 

## 12

Mielke et al. ( 2007 ) observed median and maxima deposition values of roughly 25,000 and 

## 13May 20 11 3

- 91 DRAFT

– DO NOT CITE OR QUOTE 

### 265,000 μg/m 2 

### , respectively. Median and maxima surface soil Pb concentrations were observed to be 

## 1

1,000 and 20,000 mg/kg, respectively. Clark et al. ( 2006 ) performed isotopic analysis on urban garden 

## 2

soils in Boston, M A and estimated that 60% of the soil Pb could be attributed to historic Pb on

- road 

## 3

gasoline emissions in an urban area of the city while 14% could be attributed to historic Pb on

- road 

## 4

gasoline emissions in a suburban area. The remainder of the Pb was attri buted to paint degradation. 

## 5

Several studies have examined the effects of roadways on Pb content in roadside dust. In an 

## 6

analysis of the relationship between land use parameters and Pb concentration in soil in Los Angeles, Wu 

## 7

et al. ( 2010 ) observed that soil Pb concentration was higher near freeways and major traffic arteries 

## 8

compared with other locations. The age of the land parcel (square

- root transformed), length of highw ay 

## 9

within a 1,000 m buffer, and length of local road within a 20 m buffer in which the sample was obtained 

## 10

were significant predictors of Pb. Home age within 30 m of a soil sample and road length within 3,000 m 

## 11

of a road sample were also shown to be signif icant predictors of soil Pb concentration in areas not 

## 12

designated to be near a freeway or major traffic artery. Wu et al. ( 2010 ) concluded that both historical 

## 13

traffic and leaded paint contributed to Pb contamination in soils. Amato et al. ( 2009 ) o bserved that 

## 14

deposited PM 10 

### onto roadways, measured as dust samples, in Barcelona, Spain was differentially enriched 

## 15

with Pb. Pb concentration in PM 10 

### was highest at ring roads (229 ppm) and in the city center (225 ppm), 

## 16

followed by demolition and construction sites (177 ppm) and near a harbor (100 ppm). J oshi et al. ( 2008 ) 

## 17

also observed Pb dust concentrations to be highest at in dustrial sites followed by commercial and 

## 18

residential sites in Singapore. 

## 19

Two recent studies focused on Pb from paint degradation by examining Pb dust loading to hard 

## 20

surfaces located along transects of each of the five boroughs of New York City ( Caravanos et al., 2006 ; 

## 21

Weiss et al., 2006 )

. Caravanos et al. ( 2006 ) used GIS to examined Pb dust loadings on top of pedestrian 

## 22

traffic signals and observed “hot spots,” defined by the authors as at least twice the Pb dust loading at 

## 23

adjacent samples near major elevated bridges in upper Manhattan, the Bronx, and Queens. In Brooklyn 

## 24

and Staten Island, areas with high dust loading were not clearly attributed to a source. “Low spots,” 

## 25

defined by the authors as at least two times lower Pb dust loading compared with adjacent samples were 

## 26

observed in lower Manhattan, were thought to correspond with intensive cleaning efforts that followed 

## 27

the September 11, 2001 World Trade Center attack. Weiss et al. ( 2006 ) studied Pb concentrations of grit 

## 28

(granules of mixed composition found to accumulate alongside street curbs) along the transects and found 

## 29

that median Pb concentrations in grit under the elevated steel structures were 2.5

- 11.5 ti mes higher than 

## 30

those obtained away from steel structures; 90 th percentile values were up to 30 times higher near steel 

## 31

structures compared with those further from these structures. 

## 32

Outdoor Pb dust has been also associated with demolition activities. Farfe l et al. ( 2003

, 2005 ) 

## 33

measured Pb dust within 100 m of a demolition site before, immediately after, and 1 month following the 

## 34

demolition. They found that the rate of Pb dust fall increased by a factor of more than 40 during 

## 35

demolition ( Farfel et al., 2003 ) . Immediately after demolition, one demolition site had dust loadings 

## 36May 20 11 3

- 92 DRAFT

– DO NOT CITE OR QUOTE 

### increase by a factor of 200% for streets (87,000 μg/m 2 

### ), 138% for alleys (65,000 μg/m 2 

### ), and 26% for 

## 1

sidewalks (23,000 μg/m 2 

### ) compared with pre

- demolition Pb dust levels. At another demolition site, 

## 2

smaller increases were observed: 29% for streets (29,000 μg/m 2 

### ), 18% for alleys (19,000 μg/m 2 

### ) and 18% 

## 3

for sidewalks (22,000 μg/m 2 

### ) ( Farfel et al., 2005 )

. 

## 4

Pb can be present in soils located where ammunition is used for military or hunting purposes. In a 

## 5

study of Pb content in sand used to cover a firing range, Lewis et al. ( 2010 ) found that 93% of bullet mass 

## 6

was recovered in the top 0.3 m of the sand, and 6.4% was recovered at a depth of 0.3

- 0.45 m. Pb oxides 

## 7

were observed to be the dominant species in the contaminated sand. Berthelot et al. ( 2008 ) studied soil Pb 

## 8

concentrations in grounds used for testing military tanks and munitions and measured soil Pb levels to 

## 9

range from 250 to 2,000 mg/kg. 

## 10

Soil Pb variability depends on the stre ngth and prevalence of nearby sources. Griffith et al. ( 2002 ) 

## 11

investigated spatial autocorrelation of soil Pb concentration at three sites: urban Syracuse, NY, rural Geul 

## 12

River, The Netherlands, and an abandoned Pb Superfund sit e in Murray, UT. In both Syracuse and Geul 

## 13

River, the soil Pb concentrations were not strongly correlated in space, with the exception of soil obtained 

## 14

near roads, which exhibited less variability. The smelting and shooting areas of the Superfund site were 

## 15

both demonstrated to have spatial clusters that were well correlated. These results suggest that soil Pb 

## 16

concentration tends to be spatially heterogeneous in the absence of a source. Later work on the spatial 

## 17

distribution of metals in Syracuse produced si milar results for that city ( Griffith et al., 2009 ) . These 

## 18

studies did not adjust for age of housing, although Griffith et al. ( 2009 ) did find that housing age and Pb 

## 19

co

- vary. An association between housing age and soil Pb would likely be enhanced by such co

- variation. 

## 20



# 3.6.2. Sediments

The recently completed Western Airborne Contaminants Assessment Project (WACAP) is the most 

## 21

comprehensive database, to date, on contaminant transport and depositional effects on sensitive 

## 22

ecosystems in the U.S

. ( Landers et al., 2010 )

. The transport, fate, and ecological impacts of semi

- volatile 

## 23

compounds and metals from atmospheric sources were assessed on ecosystem components collected from 

## 24

2002

- 2007 in watersh eds of eight core national parks ( Landers et al., 2008 ) . The goals of the study were 

## 25

to assess where these contaminants were accumulating in remote ecosystems in the Western U.S. , identify 

## 26

ecological receptors for the pollutants, and to determine the source of the air masses most likely to have 

## 27

transported the contaminants to the parks

. Although, Pb was measured in snow, water, sediment, lichen 

## 28

and fish during the multiyear project

, this metal was not quantified in air samples. 

## 29

In the WACAP study, bioaccumulation of airborne contaminants was demonstrated on a regional 

## 30

scale in remote ecosystems in the Western U.S . Contaminants were shown to accumulate geographically 

## 31

based on proximity to individual sources or source areas, primarily agriculture and industry. This finding 

## 32May 20 11 3

- 93 DRAFT

– DO NOT CITE OR QUOTE 

### was counter to the original working h ypothesis that most of the contaminants found in western parks 

## 1

would originate from e astern Europe and Asia. 

## 2

Pb concentrations in sediments from all lakes in which Pb was measured in the conterminous 48 

## 3

states exhibited higher Pb concentrations near the s urface relative to preindustrial Pb levels measured at 

## 4

greater depth

. This was not the case for other metals measured, except for cadmium (Cd) and mercury 

## 5

(Hg)

. Sediments in most lakes exhibited maximum concentrations between 1960 and 1980, followed by a 

## 6

d ecrease

. A clear decline in Pb concentrations in sediments after the discontinued use of leaded on

- road 

## 7

gasoline was observed at almost all WACAP locations in the for nearly all WACAP sites in the 

## 8

conterminous 48 states. Pb concentrations in sediments were much lower in Alaska, and no such decline 

## 9

was observed. Pb in sediments was mainly attributed to on

- road gasoline use, but for some lakes a strong 

## 10

influence from other local sources of Pb to lake sediments was shown to be important, including Pb 

## 11

mining, s melting, logging, and other industrial activities

. Pb was also consistently observed in WACAP 

## 12

fish and wildlife samples. 

## 13

Data from select regions of the U.S. illustrate that Pb concentrations in surface waters and sediment 

## 14

are likely to be higher in urban ized areas compared with rural locations. Table 3

- 10 presents data from 

## 15

seven metropolitan areas ( Cobb et al., 2006 ) . Differences among the intraurban co ncentration ranges 

## 16

illustrate a high level of spatial variability within individual cities as well as high interurban variability. 

## 17

The rural New Orleans site reported relatively low Pb sediment concentrations, and the highest Pb 

## 18

sediment concentrations wer e reported for the city of New Orleans. Figure 3

- 28 and Figure 3

- 29 illustrate 

## 19

such variability within a single watershed for the Apalachicola, Chattahoochee, and Flint River Basin, 

## 20

which runs south from north of the greater Atlanta, GA metropolitan area a nd drains into the Gulf of 

## 21

Mexico at the Apalachicola Bay in the Florida panhandle. Sediment concentrations peaked near the 

## 22

Atlanta area and diminished as distance from the Apalachicola Bay decreased. This observation suggests 

## 23

that rural areas have lower P b sediment levels compared with urban areas. The data also illustrated that Pb 

## 24

concentrations in sediment have declined in the U.S. since 1975 (Figure 3

- 29), prior to the phase

- out of 

## 25

on

- road leaded gasoline. 

## 26

Table 3

- 10. Sediment concentrations in various cities, prior to 2005

City Av g Pb Concentration (mg/kg) a Pb Concentration Range (mg/kg) a

Baltimore, MD 1

- 10,900 Miami, FL 275 25

- 1612 Mt. Pleasant, MI 320 100

- 840 New Orleans, LA 784 31.7

- 5195 New Orleans, LA (rural outskirts) 11 4.8

- 17.3 St. Louis, MO 427 35

- 1860 Syracuse, NY 80 20

- 800

a Dry weight basis.

Source: Used with permission from the American Chemical Society, Cobb et al. ( 2006 )

.May 20 11 3

- 94 DRAFT

– DO NOT CITE OR QUOTE Source: Use d with permission from the American Chemical Society, Callender and Rice ( 2000 )

. 

### Figure 3

- 28. Sediment core data (1992

- 1994) for the lakes and reservoirs along the Apalachicola, Chattahoochee, and Flint River Basin (ACF), which feeds from north of the Atlanta, GA metropolitan area into the Gulf of Mexico at Apalachicola Bay in the Florida panhandle. Note that background refers to concentrations from
undevel oped geographic regions and baseline samples are
obtained from the bottom of the sediment core to minimize anthropogenic effects on the sample.May 20 11 3

- 95 DRAFT

– DO NOT CITE OR QUOTE Source: Used with permission from the American Chemical Society, Callender and Rice ( 2000 )

. 

### Figure 3

- 29. Sediment core data (1975

- 1995) for the lakes and reservoirs along the Apalachicola, Chattahoochee, and Flint River Basin (ACF), which feeds from north of the Atlanta, GA metropolitan area into the Gulf of Mexico at Apalachicol a Bay in the Florida panhandle. Note that background refers to concentrations from undeveloped geographic regions and baseline samples are
obtained from the bottom of the sediment core to minimize anthropogenic effects on the sample.

Many recent studies have illustrated the effects of natural disasters on Pb concentrations in surface 

## 1

water and sediment in the wake of Hurricane Katrina, which made landfall on August 29, 2005 in New 

## 2

Orleans, LA, and Hurricane Rita, which made landfall west of New Orleans on September 23, 2005. 

## 3

Pardue et al. ( 2005 ) sampled floodwaters on September 3 and September 7, 2005 following the hurricanes 

## 4

and observed that elevated concentrations of P b along with other trace elements and contaminants were 

## 5

not irregular for stormwater but were important because human exposure to the stormwater was more 

## 6

substantial for Hurricane Katrina than for a typical storm. Floodwater samples obtained throughout the 

## 7

city on September 18, 2005 and analyzed for Pb by Presley et al. ( 2006 ) were below the limit of detection. 

## 8

Likewise, Hou et al. ( 2006 ) measured trace metal concentration in the water column of Lake 

## 9

Pontchartrain and at various locations within New Orleans during the period September 19 through 

## 10

October 9, 2005 and found that almost all Pb concentrations were b elow the limit of detection. However, 

## 11May 20 11 3

- 96 DRAFT

– DO NOT CITE OR QUOTE 

### several studies noted no appreciable increase in Pb concentration within Lake Pontchartrain soils and 

## 1

sediments ( Abel et al., 2010 ; Abel et al., 2007 ; Cobb et al., 2 006 ; Presley et al., 2006 ; K. J. Schwab et al., 

## 2

2007 ) . Shi et al. ( 2010 ) analyzed Lake Pont chartrain sediment samples using a factored approach and 

## 3

found that most Pb was sequestered in carbonate

- rich, iron oxide

- rich, and magnesium oxide

- rich 

## 4

sediments in which it can be more readily mobilized and potentially more bioaccessible. Zahran et al. 

## 5

( 2010 ) and Presley et al. ( 2010 ) noted that soil Pb samples obtained outside schools also tended to 

## 6

decrease in the wake of Hurricanes Katrina and Rita, with some sites observing substantial increases and 

## 7

others noting dramatic reductions. T hese studies suggest that floodwaters can change the spatial 

## 8

distribution of Pb in soil and sediments to result in increased or reduced concentrations. 

## 9



# 3.6.3. Rain

Recent results from locations outside the United States were consistent with decreasing rain water 

## 10

concentrations described in the 2006 Pb AQCD , reflecting the elimination of Pb from on

- road gasoline in 

## 11

most countries

. From the 2006 Pb AQCD ( U.S. EPA, 2006 ) , volume weighted Pb concentrations in 

## 12

precipitation collected in 1993

- 94 from Lake Superior, Lake M ichigan and Lake Erie ranged from ~0.7 to 

## 13

~1.1 μg/ L ( Sweet et al., 1998 ) . These values fit well with the temporal trend reported in Watmough and 

## 14

Dillon ( 2007 ) , who calculated annual volume

- weighted Pb conce ntrations to be 2.12, 1.17 and 0.58 μg/ L 

## 15

for 1989

- 90, 1990

- 91 and 2002

- 03, respectively, in precipitation from a central Ontario, Canada, forested 

## 16

watershed. A similar value of 0.41 μg/ L for 2002

- 03 for Plastic Lake, Ontario, was reported in Landre et 

## 17

al. ( 2009 ) . For the nearby Kawagama Lake, Shotyk and Krachler ( 2010 ) gave Pb concentrations in 

## 18

unfiltered rainwater collected in 2008. For August and September 2008, the values wer e 0.45 and 

## 19

0.2 2 μg/ L, respectively, and so there had been little discernible change over the post

- 2000 period. I n 

## 20

support, Pb concentrations in snow pit samples collected in 2005 and 2009 collected 45 km n ortheast of 

## 21

Kawagama Lake had not changed to any noticeable extent (0.13, 0.17, and 0.2 8 μg/ L in 2005; 0. 15 and 

## 22

0.26 μg/ L in 2009) ( Shotyk & Krachler, 2010 )

. 

## 23

There have also been a few recently published, long

- term European studies of Pb concentration in 

## 24

precipitation including Berg et al. ( 2008 ) and Farmer et al. ( 2010 ) . Berg et al. ( 2008 ) compared the trends 

## 25

in Pb concentration in precipitation at Norwegian background sites in relation to the decreasing European 

## 26

emissions of Pb over the period 1980

- 2005. The Birkenes site a t the southern tip of Norway is most 

## 27

affected by long

- range transport of Pb from mainland Europe but there had been a 97% reduction in the 

## 28

concentration of Pb in precipitation over the 26

- year time period. This was similar to the reductions of 

## 29

95% and 92% found for the more northerly sites, Karvatn and Jergul/Karasjok, respectively ( Figure 3

- 30 )

. 

## 30

A decline of ~95% in Pb concentrations in moss (often used as a biomonitor of Pb pollution) from the 

## 31

southernmost part of Norway, collected every 5 years over the period 1977

- 2005, agreed well with the 

## 32

Birkenes precipitation results ( Berg et al., 2008 ) . The reductions in Pb concentration in both precipitation 

## 33May 20 11 3

- 97 DRAFT

– DO NOT CITE OR QUOTE 

### and moss appear to agree well with the reductions in emissions in Europe (~85%) and Norway (~99%). 

## 1

However, similarly to the situation in the U.S. , the greatest reductions occurred by the late 1990 s and only 

## 2

relatively minor reduc tions have occurred thereafter; see Figure 3

- 30

. 

## 3

Source: Used with permission from Pergamon Press, Berg et al. ( 2008 ) 

### Figure 3

- 30

. Trends in Pb concentration in precipitation from various sites in Norway over the period 1980

- 2005

.

Farmer et al. ( 2010 ) showed the trends in concentration of Pb in precipitation collected in a remote 

## 4

part of northeastern Scotl and over the period 1989

- 2007. The 2.6

- and 3.0

- fold decline in mean 

## 5

concentration from 4.92 μg/ L (1989

- 1991) to 1.88 μg/ L (1999) and then to 0.63 μg/ L (2006

- 2007) is 

## 6

qualitatively but not quantitatively in line with the sixfold decline in annual total U.K

. emissions of Pb to 

## 7

the atmosphere over each of these time periods. Since the outright ban on the use of leaded on

- road 

## 8

gasoline in 2000, however, the ratio of Pb concentrations in rainwater to U.K. Pb emissions ( metric tons) 

## 9

appears to have stabili z ed to a near

- constant value of 0.009 μg/ L per metric ton

. The concentrations in 

## 10

precipitation reported in these studies are all at the lower end of the range reported in the 2006 Pb AQCD 

## 11

( U.S. EPA, 2006 ) , and similar to concentrations reported for those studies conducted after the removal of 

## 12

Pb from on

- road gasoline

. Overall, recent studies of wet deposition tended to confirm the conclusions of 

## 13

the 2006 Pb AQCD ( U.S. EPA, 2006 ) that wet deposition fluxes have greatly decreased since the removal 

## 14

of Pb from on

- road g asoline. 

## 15



# 3.6.4. Snowpack

The location of Pb deposition impacts its further environmental transport

. For example, Pb 

## 16

deposited to some types of soil may be relatively immobile, while Pb deposited to snow is likely to 

## 17May 20 11 3

- 98 DRAFT

– DO NOT CITE OR QUOTE 

### undergo further transport more easily when snow melts

. Deposition to snow was investigated in several 

## 1

studies

. Seasonal patterns of heavy metal deposition to snow on Lambert Glacier basin, east Antarctica, 

## 2

were determined by Hur et al. ( 2007 ) . The snow pit samples covered the period from austral spring 1998 

## 3

to summer 2002 and Pb concentrations ranged from 1. 29

- 9.6 pg/g with a mean value of 4.0 pg/g . This 

## 4

was similar to a mean value of 4.7 pg/g (1965

- 1986) obtained by Planchon et al. ( 2003 ) for Coats Land, 

## 5

northwest Antarctica. Estimated contributions to the Pb in Lambert Glacier basin snow were ~1% from 

## 6

rock and soil dust (based on Al concentrations) and ~4.6% from volcanoes (based on the concentra tions 

## 7

of nss

- sul f ate). There was almost negligible contribution from seaspray (based on Na concentrations)

, and 

## 8

so it was suggested that a substantial part of the measured Pb concentration must originate from 

## 9

anthropogenic sources. Highest Pb concentration s were generally observed in spring/summer with an 

## 10

occasional peak in winter. This contrasts with data for the Antarctic Peninsula, where highest 

## 11

concentrations occurred during autumn/winter, and again with Coats Land, where high concentrations 

## 12

were observ ed throughout the winter. These differences were attributed to spatial changes in input 

## 13

mechanism of Pb aerosols arriving at different sites over Antarctica

, which could be due to their different 

## 14

source areas and transport pathways. Hur et al. ( 2007 ) , however, suggested that the good correlation 

## 15

betw een Pb and crustal metals in snow samples shows that Pb pollutants and crustal PM are transported 

## 16

and deposited in Lambert Glacier basin snow in a similar manner

. 

## 17

Lee et al. ( 2008 ) collected 42 snow samples during the period autumn 2004

- summer 2005 from a 

## 18

2.1 m snow pit at a high

- altitude site on the northeast slope of Mount Everest, Himalayas. Pb 

## 19

concentrations ranged from 5

- 530 pg/g with a mean valu e of 77 pg/g . The mean value is clearly higher 

## 20

than the Hur et al. ( 2007 ) value for Antarctica but is substantial ly lower than a mean concentration of 573 

## 21

pg/g for snow from Mont Blanc, France (1990

- 1991) ( collated in 2008 ) . The mean Pb concentration for 

## 22

Mount Everest snow was lower during the monsoon (28 pg/g ) compared with the non

- monsoon periods 

## 23

(137 pg/g ). From calculated enrichment factors (Pb/Al snow :Pb/Al crust ), anthropogenic inputs of Pb were 

## 24

partly important but soil and rock dust also contributed. The low Pb concentrations during monsoon 

## 25

periods are thought to be attributable to low levels of atmospheric loadings of crustal dusts. K. Lee et al. 

## 26

( 2008 ) noted that their conclusions differ from those in Kang et al. ( 2007 ) , who stated that anthropogenic 

## 27

contributions of Pb to Mount Everest snow were negligible because the Everest concentrations were 

## 28

similar to those in Antarctica. Kang et al. ( 2007 ) did not take account of the difference in accumulation 

## 29

rates at the two sites and had also used Pb concentrations for Antarctic snow from a study by Ikegawa et 

## 30

al. ( 1999 )

. Lee et al. ( 2008 ) suggested that these Pb concentrations were much higher than expected and 

## 31

that their snow samples may have suffered from contamina tion during sampling and analysis. 

## 32May 20 11 3

- 99 DRAFT

– DO NOT CITE OR QUOTE 

# 3.6.5. Natural Waters

Shotyk and Krachler ( 2007 ) measured Pb concentrations in six artesian flows in Simcoe County, 

## 1

near Elmvale, Ontario, Canada. The values ranged fr om 0.9 to 18 ng/ L with a median (n = 18) of 

## 2

5.1 ng/ L. These are comparable with reports of a range of 0.3

- 8 ng/ L for Lake Superior water samples 

## 3

( Field & Sherrell, 2003 )

. Shotyk and Krachler ( 2007 ) also commented that such low concentrations for 

## 4

ground and surface waters are not significantly different from those (5.1 ± 1.4 ng/ L) reported for Arctic 

## 5

ice from Devon Island, Canada, dating from 4,000

- 6,000 year s ago. In a separate study, Shotyk and 

## 6

Krachler ( 2009 ) reported concentrations of Pb in groundwater (from two locations, Johnson and Parnell), 

## 7

surface water (Kawagama Lake) and contemporary snow (Johnson and Parnell). The lowest mean 

## 8

dissolved Pb concentrations were found for groundwater: 5.9 (Johnson, n = 11) and 3.4 (Parnell, n = 

## 9

12) ng/ L. For lake water the mean Pb concentration was 57 (Kawagama Lake, n = 12) ng/ L and that for 

## 10

contemporary snow was 672 (Johnson, n = 6; Parnell, n = 3) ng/ L. Shotyk et al. ( 2010 ) gave additional 

## 11

values for Pb in contemporary snow samples and these were again higher than for ground and surface 

## 12

w aters. Luther Bog and Sifton Bog snow had mean Pb concentrations of 747 and 798 ng/ L, respectively. 

## 13

The relatively high concentrations in snow were attributed to contamination with predominantly 

## 14

anthropogenic Pb, although it was noted that the extent of co ntamination was considerably lower than in 

## 15

past decades. The extremely low concentrations of Pb in the groundwaters were attributed to natural 

## 16

removal processes. Specifically, at the sampling location in Canada, there is an abundance of clay 

## 17

minerals with high surface area and high cation exchange capacity and these, combined with the elevated 

## 18

pH values (pH=8.0) resulting from flow through a terr ai n rich in limestone and dolostone, provide 

## 19

optimal circumstances for the removal of trace elements such as Pb f rom groundwater. Although such 

## 20

removal mechanisms have not been demonstrated, the vast difference between Pb concentration in snow 

## 21

and that in the groundwaters indicate that the removal process is very effective. Shotyk and Krachler 

## 22

( 2010 ) speculate that even at these very low Pb concentrations, much if not most of the Pb is likely to be 

## 23

colloidal, as suggested by the 2006 Pb AQCD ( U.S. EPA, 2006 ) . F inally, Shotyk et al. ( 2010 ) suggest that 

## 24

the pristine groundwaters from Simcoe County, Canada, provide a useful reference level against which 

## 25

other water samples can be compared. 

## 26

Although Pb concentrations in Kawagama Lake water were approaching “ natural values ,” the 

## 27

20 6 

### Pb/ 207 

### Pb ra tios for the samples that had the lowest dissolved Pb concentrations of 10, 10 and 6 ng/ L 

## 28

were 1.16, 1.1 5 and 1.16, respectively. These values are far from those expected for natural Pb (the clay 

## 29

fraction from the lake sediments dating from the pre

- industr ial period had values of 1.19

- 1.21) and it was 

## 30

concluded that most of the dissolved Pb in the lake water was of industrial origin. Shotyk and Krachler 

## 31

( 2010 ) found that the full range of isotope ratios for Kawagama Lake water samples (Ontario, Canada) 

## 32

was 1.09 to 1.15; this was not only much lower than the stream water values entering the lake but also 

## 33

lower than the values attributed to leaded on

- road gasoline in Canada (~1.15). The streamwater ratio 

## 34May 20 11 3

- 100 DRAFT

– DO NOT CITE OR QUOTE 

### values were ~1.16 to 1.17 whilst those for rainwater were as low as 1.09, in good agreement with the 

## 1

lower lake water values. This means that there must be an additional atmospheric source of Pb

, which has 

## 2

a lower 206 

### Pb/ 207 

### Pb ratio than leaded on

- road gasoline. Supporting evidence came from contemporary 

## 3

samples such as near surface peat, rainwater and snow, all of which confirmed a trend away from natural 

## 4

Pb (1.191 to 1.201) to lower 206 

### Pb/ 207 

### Pb ratios. The local smelting activities (Sudbury) were unlikely to be 

## 5

the source of anthropogenic Pb as Sudbury

- derived emissions exhibit a typical 206 

### Pb/ 207 

### Pb ratio of ~1.15, 

## 6

similar to leaded on

- road gasoline. Instead, it was suggested that long

- range transport o f Pb from the 

## 7

smelter at Rouyn

- Noranda (known as the “C apital of M etal ,” NW Quebec) may still be impacting on 

## 8

Kawagama Lake but no Pb isotope data was quoted to support this supposition. 

## 9



# 3.6.6. Moss

Mosses can be used effectively for monitoring trends in Pb depos ition as demonstrated in many 

## 10

studies ( Harmens et al., 2008 ; Harmens et al., 2010 ) . For example, Harmens et al. ( 2008 ) showed that a 

## 11

52% decrease in deposited Pb concentrations corresponded to a 57% decrease in Pb concentrations in 

## 12

moss. Farmer et al. ( 2010 ) showed that there was good agreement between the 206 

### Pb/ 207 

### Pb ratio for 

## 13

precipitation and mosses collected in northeast Scotland. A study in the Vosges mountains also found a 

## 14

ratio value of 1.158 for a moss sample and a surface soil litter value of 1.167 and concluded that 1.158 to 

## 15

1.167 represented the current atmospheric baseline ( Geagea et al., 2008 ) . For rural northeast Scotland, a 

## 16

combination of sources is giving ri se to a 206 

### Pb/ 207 

### Pb ratio of ~1.15 in recent precipitation and mosses 

## 17

( Farmer et al., 2010 ) . Clearly, sources with a lower ratio than coal (~1.20) must be contributing 

## 18

substantial ly to the overall emissions. Pb from waste incineration has been implicated as a possible 

## 19

current source (cf. typical 206 

### Pb/ 207 

### Pb ratios for Pb from European incineration plants are ~1.14 to 1.15 (de 

## 20

la Cruz et al. ( 2009 ) and references therein)

. 

## 21



# 3.6.7. Grass , Foliage, and Tree Rings

Trends in Pb concentration among flora have decreased in recent years. For example, Franzaring et 

## 22

al. ( 2010 ) evaluated data from a 20

- y ear biological monitoring study of Pb concentration in permanent 

## 23

forest and grassland plots in Baden

- Württemberg, southwest Germany. Grassland and tree foliage samples 

## 24

were collected from 1985

- 2006. The samples were not washed and so atmospheric deposition rather than 

## 25

uptake from the soil probably predominates. For all foliage (beech and spruce), Pb concentrations have 

## 26

shown large reductions over time, particularly in the early 1990s. The Pb concentrations in the grassland 

## 27

vegetation also decreased from the late 1980s to the early 1990s but the trend thereafter was found to be 

## 28

statistically non

- significant. The reduction corresponded to the phase

- out of leaded on

- road gasoline in 

## 29

Germany . Similarly, Aznar et al. ( 2008 ) observed that the decline in Pb concentrations in the outer level of 

## 30

tree rings corresponded with the decline in Cu smelter emissions in Gaspé Peninsula in Can ada; see 

## 31May 20 11 3

- 101 DRAFT

– DO NOT CITE OR QUOTE 

### Figure 3

- 31. Both Pb concentrations and Pb isotope ratios declined with distance from the smelter ( Aznar, 

## 1

Richer

- Lafleche, et al., 2008 ; Aznar, Richer

- Laflèche, et al., 2008 )

. 

## 2

Source: Used with permission from Elsevier Publishing, Aznar et al. ( 2008 ) 

### Figure 3

- 3 1

. Trends in regional pollution near a copper smelter in Canada and Pb concentrations at the boundary of heartwood trees within roughly 75 km of the smelter.



# 3.6.8. Aquatic Bivalves

Data from invertebrate waterborne populations can serve as in indicator of Pb contamination 

## 3

because animals such as mussels an d oysters take in contaminants during filter feeding. Kimbrough et al. 

## 4

( 2008 ) surveyed Pb concentrations in mussels, zebra mussels, and oysters along the coastlines of the 

## 5

continental U.S

. In general, they observed the highest concentrations of Pb in the vicinity of urban and 

## 6

industrial areas

. Company et al. ( 2008 ) measured Pb concentrations and Pb isotope ratios in bivalves 

## 7

along the Guadiana River separating Spain and Portugal

. Analysis of Pb isotope ratio data suggested that 

## 8

high Pb concentrations we re related to historical mining activities in the region

. Elevated Pb 

## 9May 20 11 3

- 102 DRAFT

– DO NOT CITE OR QUOTE 

### concentrations were also observed by Company et al. ( 2008 ) in the vicinity of more populated areas

. 

## 1

Couture et al. ( 2010 ) report data from a survey of the isotopic ratios of Pb in Mytilus edulis blue mussel, 

## 2

collected off the coast of France from 1985

- 2005

. The results indicated that the likely source of Pb in 

## 3

mussel tissue is from resuspension of contaminated sediments enriched with Pb runoff from wastewater 

## 4

treatment plants, municipal waste incinerators, smelters and refineries rather than from atmospheric 

## 5

deposition ( Couture et al., 2010 )

. 

## 6



# 3.7. Summary

3.7.1. Sources of Atmospheric Lead

The 2006 Pb AQCD ( U.S. EPA, 2006 ) documented the decline in ambient air Pb emissions 

## 7

following the ban on alkyl

- Pb additives for on

- road gasoline. Pb emissions declined by 98% from 1970 to 

## 8

1990 and then by an additional 77% from 1990 to 2008, at which time emissions were 1,200 tons/yr. 

## 9

Industrial processes, including metals processing and industrial fuel combustion, had replaced mobile 

## 10

sources as the primary source of Pb to the atmosphere by the 2006 Pb A QCD ( U.S. EPA, 2006 ) . More 

## 11

recent data from the 2008 NEI ( U.S. EPA, 2011 ) illustrate that piston engine aircraft emissions now 

## 12

comprise the largest share (~49%) of to tal atmospheric Pb emissions; the 2008 NEI ( U.S. EPA, 2011 ) 

## 13

estimated that 590 tons of Pb were emitted from aircraft point sources. Other sources of ambient air Pb, in 

## 14

approximate order of importance, include metals processing, fossil fuel combustion, other industrial 

## 15

sources, roadway related sources, and historic Pb. 

## 16

Chemical speciation of Pb had been fairly well characterized in the 2006 Pb AQCD ( U.S. EPA, 

## 17

2006 ) . Est imates from the 1986 Pb AQCD ( U.S. EPA, 1986 ) for organic on

- road Pb emissions provides 

## 18

an upper bound for organic vapor emissions of 20% of total Pb dibromide and Pb bromide emissions from 

## 19

piston engine aircraf t. Recent speciation studies of smelting and battery recycling operations have shown 

## 20

that PbS and Pb sulfates are abundant within the emissions mixture for such industrial operations. 

## 21



# 3.7.2. Fate and Transport of Lead

T he atmosphere is the ma in environme ntal tra nsport pathway for Pb, and on a global scale 

## 22

atmospheric Pb is primarily associated with fine PM. On a global scale, Pb associated with fine PM is 

## 23

transported long distances and found in remote areas . Global atmospheric Pb deposition peak ed in the 

## 24

1970s, f ollowed by a more recent decline . On a local scale, Pb concentrations in soils (including urban 

## 25

areas where historic use was widespread) can be substantial, and coarse Pb

- bearing PM experiences 

## 26

cycles of deposition and resuspension that serve to distribute it. Both wet and dry deposition are important 

## 27

removal mechanisms for atmospheric Pb. Because Pb in fine PM is typically fairly soluble, wet 

## 28May 20 11 3

- 103 DRAFT

– DO NOT CITE OR QUOTE 

### deposition is more important for fine Pb. In contrast, Pb associated with coarse PM is usually insoluble, 

## 1

and remov ed by dry deposition. However, local deposition fluxes are much higher near local industrial 

## 2

sources, and a substantial amount of emitted Pb is deposited near sources, leading to high soil Pb 

## 3

concentrations. R esuspension by wind and traffic can be a n impor tant source of airborne Pb near sources 

## 4

where Pb occurs in substantial amounts in surface dust. 

## 5

In water, Pb is transported as free ions, soluble chelates, or on surfaces of iron and organic rich 

## 6

colloids, and w ater columns behave as important reservoirs o f Pb . In surface waters, atmospheric 

## 7

deposition is the largest source of Pb, but urban runoff and industrial discharge are also considerable. A 

## 8

substantial portion of Pb in runoff ultimately originates from atmospheric deposition , but substantial 

## 9

amounts o f Pb from vehicle wear and building materials can also be transported by runoff waters without 

## 10

becoming airborne. Often a disproportionate amount of Pb is removed by runoff at the beginning of a 

## 11

rainfall event. Pb is rapidly dispersed in water, and highest concentrations of Pb are observed near sources 

## 12

where Pb is deposited. 

## 13

Transport in surface water s is largely controlled by exchange with sediments

. T he cycling of Pb 

## 14

between water and sediments is governed by chemical, biological, and mechanical processes

, which are 

## 15

affected by many factors

. Organic matter in sediments has a high capacity for accumulating trace 

## 16

elements like Pb. In anoxic sediments removal by sulfides is particularly important

. Pb is relatively stable 

## 17

in sediments, with long residence time s and limited mobility

. However, Pb containing sediment particles 

## 18

can be remobilized into the water column , and sediment concentration s tend to follow those in overlying 

## 19

waters . Resuspended Pb is largely associated with OM or iron and manganese particles. This resuspension 

## 20

of contaminated sediments strongly influences the lifetime of Pb in water bodies and can be a more 

## 21

important Pb source than atmospheric deposition . Resuspension and release from sediments largely 

## 22

occurs during discrete events related to s torms. 

## 23

A complex variety of factors that influence Pb retention in soil, including hydraulic conductivity, 

## 24

solid composition, OM content, clay mineral content, microbia l activity, plant root channels, animal 

## 25

holes, geochemical reactions , colloid amounts, c olloidal surface charge, and pH. Leaf litter c an be an 

## 26

important temporary sink for metals from the soil around and below leaves , and decomposition of leaf 

## 27

litter can reintroduce substantial amounts of Pb into soil “hot spots,” where re

- adsorption of Pb is favored. 

## 28

A small fraction of Pb in soil is present as the free Pb 2+ 

### ion. The fraction of Pb in this form is strongly 

## 29

dependent on soil pH. 

## 30

In summary, environmental distribution of Pb occurs mainly through the atmosphere, from where it 

## 31

is deposited into s urface waters and soil. Pb associated with coarse PM deposits to a great extent near 

## 32

sources, leading to high soil concentrations in those locations, while fine Pb

- PM can be transported long 

## 33

distances, leading to contamination of remote areas. Surface wate rs act as an important reservoir, with Pb 

## 34

lifetimes largely controlled by deposition and resuspension of Pb in sediments. Pb retention in soil 

## 35

depends on Pb speciation and a variety of factors intrinsic to the soil. 

## 36May 20 11 3

- 104 DRAFT

– DO NOT CITE OR QUOTE 

# 3.7.3. Ambient Lead Monitoring

In recognition of the role of all PM sizes in ambient air Pb exposures, including the ingestion of 

## 1

particles deposited onto surfaces, the indicator for the Pb NAAQS is Pb in Pb

- TSP. Although there is a 

## 2

lower rate of error in estimating ambient Pb from Pb

- PM 10 

### monitoring than from Pb

- TSP monitoring, the 

## 3

Pb

- TSP indicator was retained in 2008 because ingestion after deposition in the upper respiratory tract 

## 4

was considered an important component of Pb exposure. A new FRM for Pb

- PM 10 

### has been implemented 

## 5

in which ambient air i s drawn through an inertial PM size separator for collection on a PTFE filter. 

## 6

Several FEMs have also been approved. The FRM is based on flame AAS. ICPMS is under consideration 

## 7

as a new FRM for Pb

- TSP. 

## 8

Monitoring for ambient Pb levels is required for all a reas where Pb levels have been shown or are 

## 9

expected to contribute to maximum concentrations of 0.10 μg/m 3 

### or greater over a three

- year time period. 

## 10

Pb is monitored routinely at SLAMS that report data used for NAAQS compliance to the AQS database. 

## 11

Pb monit oring requirements have experienced several changes since publication of the 2006 Pb AQCD 

## 12

( U.S. EPA, 2006 ) . In addition to FRM monitoring, Pb is also routinely measured in smaller PM fractions 

## 13

in the CSN, IMPROVE, and the NATTS networks, and is planned for t he NCore network. While 

## 14

monitoring in multiple networks provides extensive geographic coverage, measurements between 

## 15

networks are not directly comparable in all cases because different PM size ranges are sampled in 

## 16

different networks. Depending on monitori ng network, Pb is monitored in TSP, PM 10 

### , or PM 2.5 

### . Monitors 

## 17

reporting to the AQS were considered for the purpose of this ISA to be source oriented if they were 

## 18

designated in AQS as source oriented, or they were located within 1 mile of a 0.5 ton/yr or gre ater source, 

## 19

as noted in the 2005 NEI ( U.S. EPA, 2008a ) . Non

- source oriented monitors were those monitors not 

## 20

considered to be source oriented based on these two criteria. 

## 21



# 3.7.4. Ambient Air Lead Co ncentrations

Ambient air Pb concentrations have declined drastically over the period 1980

- 2009. The median 

## 22

annual concentrations for all monitors have dropped by 97% from 0.87 μg / m 3 

### in 1980 to 0.025 μg / m 3 

### in 

## 23

2009. While the sharpest drop in Pb concentratio n occurred during 1980

- 1990, a declining trend was 

## 24

observed between 1990 and 2009. A smaller reduction was observable among source oriented Pb 

## 25

concentration (56%) and non

- source oriented Pb data (51%) for 2000

- 2009. 

## 26

AQS data for source oriented and non

- sou rce oriented monitoring were analyzed for 2007

- 2009. 

## 27

For source oriented monitoring, the 3

- month rolling average was measured to be above the level of the 

## 28

NAAQS in 14 counties across the U.S. Fourteen monitoring sites had maximum 3

- month rolling average 

## 29

va lues that exceeded the level of the NAAQS. The maximum 3

- month rolling average concentrations 

## 30

ranged from 0.17

- 2.9 μg / m 3



### . 

## 31May 20 11 3

- 105 DRAFT

– DO NOT CITE OR QUOTE 

### Pb concentrations, seasonal variations, inter

- monitor correlations, and wind data were analyzed for 

## 1

six counties: Los Angeles County, CA, Hillsborough County, FL, Cook County, IL, Jefferson County, 

## 2

MO, Cuyahoga County, OH, and Sullivan County, TN. These sites were selected for analysis because 

## 3

they contained a mix of source oriented and non

- source oriented monitors in urban areas. Spati al and 

## 4

temporal variability of Pb concentrations in each county were commonly high. Meteorology, distance 

## 5

from sources with respect to the monitors, and source strength all appeared to influence the level of 

## 6

concentration variability across time and space. PM size distribution also influenced how far the particle 

## 7

will travel upon initial emission or resuspension before being deposited. 

## 8

Size distribution of Pb

- bearing PM was demonstrated to vary substantially for several studies 

## 9

presented, depending on the nature of Pb sources and proximity of the monitors to the Pb sources. AQS 

## 10

data were also used to estimate the size distribution of Pb

- bearing PM at several sites with co

- located Pb 

## 11

monitors. On average, Pb

- TSP and Pb

- PM 10 

### were moderately correlated, but th e correlation improved for 

## 12

urban and city center land use types compared with all data. When comparing Pb

- TSP with Pb

- PM 2.5 

### , 

## 13

correlations were lower in urban and city center areas compared with suburban and rural sites. A 

## 14

relationship between land use type and correlation was less obvious when comparing Pb

- PM 10 

### with Pb



## 15

PM 2.5 

### . For urban and city center types, average ρ was fairly high. Average ρ increased for suburban sites 

## 16

but decreased for rural sites. Variation in correlation of size

- fractionated Pb sampl es among different land 

## 17

use types may reflect differences among sources among land use types. 

## 18

Pb concentrations exhibit varying degrees of association with other criteria pollutant 

## 19

concentrations

. Overall, Pb was moderately associated with PM 2.5 

### , PM 10 

### and NO 2 

### , with positive 

## 20

Spearman correlation coefficients observed at nearly all sites

. However, Pb was just as strongly associated 

## 21

with CO in fall and winter The poorest associations were observed between Pb and O 3 

### . Among trace 

## 22

metals, the strongest associatio n was with Zn. Br, Cu, and K concentrations also exhibited moderate 

## 23

associations with Pb concentrations. Such correlations may suggest some common sources affecting the 

## 24

pollutants. 

## 25



# 3.7.5. Ambient Lead Concentrations in Non

- Air Media and

Biota

Atmospheric deposition has led to measurable Pb concentrations observed in rain, snowpack, soil, 

## 26

surface waters, sediments, agricultural plants, livestock, and wildlife across the world, with highest 

## 27

concentrations near Pb sources, such as metal smelters. After the phase

- out of Pb from on

- road gasoline, 

## 28

concentrations in these media decreased to varying degrees. In rain, snowpack, and surface waters, Pb 

## 29

concentrations have decreased considerably following elimination of leaded on

- road gasoline. Declining 

## 30

Pb concentrations in tree foliage, trunk sections, and grasses have also been observed. In contrast, Pb is 

## 31

retained in soils and sediments, where it provides a historical record of deposition and associated ambient 

## 32May 20 11 3

- 106 DRAFT

– DO NOT CITE OR QUOTE 

### concentrations. In remote lakes, sediment p rofiles indicate higher Pb concentrations in near surface 

## 1

sediment as compared to pre

- industrial era sediment from greater depth and indicate peak concentrations 

## 2

between 1960 and 1980, when leaded on

- road gasoline was at peak use. Concentrations of moss, l ichens, 

## 3

peat, and aquatic bivalves have been used to understand spatial and temporal distribution patterns of air 

## 4

Pb concentrations. Ingestion and water intake are the major routes of Pb exposure for aquatic organisms, 

## 5

and food, drinking water, and inhalat ion are major routes of exposure for livestock and terrestrial wildlife. 

## 6

Overall, Pb concentrations have decreased substantially in media through which Pb is rapidly transported, 

## 7

such as air and water. Substantial Pb remains in soil and sediment sinks. Alt hough in areas less affected 

## 8

by major local sources, the highest concentrations are below the surface layers and reflect the phase

- out 

## 9

of Pb from on

- road gasoline and emissions reductions from other sources. 

## 10May 20 11 3

- 107 DRAFT

– DO NOT CITE OR QUOTE 

# Chapter 3 References

Abel, M. T., Cobb, G. P., Presley, S. M., Ray, G. L., Rainwater, T. R., Austin, G. P., . . . Suedel, B. C. (2010). Lead distributions

and risks in New Orleans followin g hurricanes Katrina and Rita. Environmental Toxicology and Chemistry, 29 (7), 1429

- 1437. http://dx.doi.org/10.1002/Etc.205

Abel, M. T., Presley, S. M., Rainwater, T. R., Austin, G. P., Cox, S. B., McDaniel, L. N., . . . Cobb, G. P. (2007). Spatial and temporal evaluation of metal concentrations in soils and sediments from New Orleans, Louisiana, USA, following

hurr icanes Katrina and Rita. Environmental Toxicology and Chemistry, 26 (10), 2108

- 2114. http://dx.doi.org/10.1897/06

- 595R.1

Adgate, J. L., Mongin, S. J., Pratt, G. C., Zhang, J., Field, M. P., Ramachandran, G., & Sexton, K. (2007). Relationships between personal, indoor, and outdoor exposures to trace elements in PM2.5. Science of the Total Environment, 386 (1

- 3), 2 1

- 32. http://dx.doi.org/10.1016/j.scitotenv.2007.07.007

Airborne Lead Reduction Act of 1984, S., 2609 98th Cong. (1984).

Al

- Malack, M. H. (2001). Migration of lead from unplasticized polyvinyl chloride pipes. Journal of Hazardous Materials, 82 (3),

263

- 274. http://dx.doi.org/10.1016/S0304

- 3894(00)00366

- 6

Al

- Masri, M. S., Al

- Kharfan, K., & Al

- Shamali, K. (2006). Speciation of Pb, Cu and Zn determined by sequential extraction for identification of air pollution sources in Syria. Atmospheric Environment, 40 (4), 753

- 761. http://dx.doi.org/10 .1016/j.atmosenv.2005.10.008

Allott, R. W., Hewitt, C. N., & Kelly, M. R. (1989). The environmental half

- lives and mean residence times of contaminants in

dust for an urban environment: Barrow

- in

- Furness. Science of the Total Environment, 93 , 403

- 410. http://dx.doi.org/10.1016/0048

- 9697(90)90131

- D

Amato, F., Pandolfi, M., Viana, M., Querol, X., Alastuey, A., & Moreno, T. (2009). Spatial and chemical patterns of PM10 in road dust deposited in urban environment. Atmospheric Environment, 43 (9), 1650

- 1659. http://dx.doi.org/10.1016/j.atmosenv.2008.12.009

Annibaldi, A., Truzzi, C., Illuminati, S., Bassotti, E., & Scarponi, G. (2007). Determination of water

- soluble and insoluble (dilute

- HCl

- extractable) fractions of Cd, Pb and Cu in Antarctic aerosol by square wave anodic stripping voltammetry: Distribution and summer seasonal evolution at Terra Nova Bay (Vict oria Land). Analytical and Bioanalytical Chemistry,

387 (3), 977

- 998. http://dx.doi.org/10.1007/s00216

- 006

- 0994

- 0

Ariola, V., D'Alessandro, A., Lucarelli, F., Marcazzan, G., Mazzei, F., Nava, S., . . . Zucchiatti, A. (2006). Elemental characterization of PM10, PM2.5 and PM1 in the town of Genoa (Italy). Chemosphere, 62 (2), 226

- 232. http://dx.doi.org/10.1016/j.chemosphere.2005.05.004

ASTM. (American Society for Testing and Mat erials). (2007). Standard specification for aviation gasolines . (Report No. ASTM

D910

- 06). West Conshohocken, PA: American Society for Testing and Material. Retrieved from http://www.astm.org/Standards/D 910.htm

.
Aznar, J. C., Richer

- Lafleche, M., Begin, C., & Rodrigue, R. (2008). Spatiotemporal reconstruction of lead contamination using tree rings and organic soil layers. Science of the Total Environment, 407 (1), 233

- 241. http://dx.doi.org/10.1016/j.scitotenv.2008.09.044

Aznar, J. C., Richer

- Laflèche, M., & Cluis, D. (2008). Metal contamination in the lichen Alectoria sarmentosa near the copper smelter of Murdochville, Quebec. Environmental Pollution, 156 (1), 1. http://dx.doi.org/10.1016/j.envpol.2007.12.037

Barrett, J., Taylor, K., Hudson

- Edwards, K., & Charnock, J. (2010). Solid

- phase speciation of Pb in urban road dust sediment: A XANES and EXAFS study. Environmental Science and Technology, 44 (8), 2940

- 2946. http://dx.doi.org/10.1021/es903737k

Bein, K. J., Zhao, Y., Pekney, N. J., Davidson, C. I., Johnston, M. V., & Wexler, A. S. (2006). Identification of sources of atmospheric PM at the Pittsburgh supersite Part II: Quantitative comparisons of single par ticle, particle number, and particle mass measurements. Atmospheric Environment, 40 (Suppl 2), S424

- S444. http://dx.doi.org/10.1016/j.atmosenv.2007.01.039May 20 11 3

- 108 DRAFT

– DO NOT CITE OR QUOTE Bein, K. J., Zhao, Y. J., Johnston, M. V., & Wexler, A. S. (2007). Identification of sources of atmospheric PM at the Pittsburgh supersite, Part III: Source characterization. Atmospheric Environment , 41 (19), 3974

- 3992. http://dx.doi.org/10.1016/j.atmosenv.2007.01.039

Bench, G., Grant, P. G., Ueda, D., Cliff, S. S., Perry, K. D., & Cahill, T. A. (2002). The use of STIM and PESA to measure profiles of aerosol mass and hydrogen content, respectively, across Mylar rotating drums impactor samples. Aerosol Science and Technology, 36 (5), 642

- 651. http://dx.doi.org/10.1080/02786820252883874

Bennett, J. D. (2007). Memorandum regarding Doe Run

- Her culaneum State Implementation Plan (SIP) dispersion modeling

review . Jefferson City, MO: Missouri Department of Natural Resources. Retrieved from http://www.dnr.mo.gov/env/apcp/docs/drhs ipmodmemo1.pdf

. Berg, T., Aas, W., Pacyna, J., Uggerud, H. T., & Vadset, M. (2008). Atmospheric trace metal concentrations at Norwegian background sites during 25 years and its relation to European emissions. Atmospheric Environment, 42 (32), 7494

- 7501. http://dx.doi.org/10.1016/j.atmosenv.2008.05.020

Berthelot, Y., Valton, E., Auroy, A., Trottier, B., & Robidoux, P. Y. (2008). Integration of toxicological and chemical tools to assess the bioavailability of metals and energetic compounds in contaminated soi ls. Chemosphere, 74 (1), 166

- 177. http://dx.doi.org/10.1016/j.chemosphere.2008.07.056

Biggins, P. D. E., & Harrison, R. M. (1979). Atmospheric chemistry of automotive lead. Environmental Science and Technology,

13 , 558

- 565. http://dx.doi.org/10.1021/es60153a017

Biggins, P. D. E., & Harrison, R. M. (1980). Chemical speciation of lead compounds in street dusts. Environmental Science and Technology, 14 , 336

- 339. http://dx.doi.org/10.1021/es60163a005

Bindler, R., Renberg, I., & Klaminder, J. (2008). Bridging the gap betw een ancient metal pollution and contemporary biogeochemistry. Journal of Paleolimnology, 40 (3), 755

- 770. http://dx.doi.org/10.1007/s10933

- 008

- 9208

- 4

Birch, G., & McCready, S. (2009). Catchment condition as a major control on the quality of receiving basin sediments (Sydney Harbour, Australia). Science of the Total Environment, 407 (8), 2820

- 2835. http://dx.doi.org/10.1016/j.scitotenv.2008.12.051

Birch, G., & O'Hea, L. (2007). The chemistry of suspen ded particulate material in a highly contaminated embayment of Port Jackson (Australia) under quiescent, high

- wind and heavy

- rainfall conditions. Environmental Geology, 53 (3), 501

- 516. http://dx.doi .org/10.1007/s00254

- 007

- 0662

- 5

Birmili, W., Allen, A. G., Bary, F., & Harrison, R. M. (2006). Trace metal concentrations and water solubility in size

- fractiona ted atmospheric particles and influence of road traffic. Environmental Science and Technology, 40 (4), 1144

- 1153. http://dx.doi.org/10.1021/es0486925

Boonfueng, T., Axe, L., Xu, Y., & Tyson, T. A. (2006). Nickel and lead sequestration in manganese oxide

- coated montmorillonite. Journal of Colloid and Interface Science, 303 (1), 87

- 98. http://dx.doi.org/10.1016/j.jcis.2006.07.056

Bouvet, M., Francois, D., & Schwartz, C. (2007). Road soil retention Pb leached from MSWI bottom ash. Waste Management,

27 (6), 840

- 849. http://dx.doi.org/10.1016/j.wasman.2006.06.003

Britter, R. E., & Hanna, S. R. (2003). Flow and dispersion in urban areas. Annual Review of Fluid Mechanics, 35 , 469

- 496. http://dx.doi.org/10.1146/annurev.fluid.35.101101.161147

Brown, R. W., Gonzales, C., Hooper, M. J., Bayat, A. C., Fornerette, A. M., McBrid e, T. J., . . . Mielke, H. W. (2008). Soil lead (Pb) in residential transects through Lubbock, Texas: A preliminary assessment. Environmental Geochemistry and Health, 30 , 541

- 547.

Bruggemann, E., Gerwig, H., Gnauk, T., Muller, K., & Herrmann, H. (2009). Influence of seasons, air mass origin and day of the

week on size

- segregated chemical composition of aerosol particles at a kerbside. Atmospheric Environment , 43 (15),

2456

- 2463. http://dx.doi.org/10.1016/j.atmosenv.2009.01.054

Bukowiecki, N., Lienemann, P. , Hill, M., Furger, M., Richard, A., Amato, F., . . . Gehrig, R. (2010). PM10 emission factors for

non

- exhaust particles generated by road traffic in an urban street canyon and along a freeway in Switzerland. Atmospheric Environment, 44 (19), 2330

- 2340. http://dx.doi.org/10.1016/j.atmosenv.2010.03.039

Buonanno, G., Fuoco, F. C., & Stabile, L. (2011 ). Influential parameters on particle exposure of pedestrians in urban

microenvironments. Atmospheric Environment, 45 (7), 1434

- 1443. http://dx.doi.org/10.1016/j.atmosenv.2010.12.015May 20 11 3

- 109 DRAFT

– DO NOT CITE OR QUOTE Bur, T., Probst, J. L., N'Guessan, M., & Probst, A. (2009). Distribution and origin of lead in stream sediments from small agricultural catchments draining Miocene molass ic deposits (SW France). Applied Geochemistry, 24 (7), 1324

- 1338. http://dx.doi.org/10.1016/j.apgeochem.2009.04.004

Cahill, C. F. (2003). Asian aerosol transport to Alaska during ACE

- Asia. Journal of Geophysical Research, 108 (D23), 8664. http://dx.doi.org/10.1029/2002JD003271

Cahill, T. A. (1981). Innovative aerosol sampling devices based upon PIXE capabilities. Nuclear Instruments and Methods,

181 (1

- 3), 473

- 480. http://dx.doi.org/10.1016/0029

- 554X(81)90652

- 2

Cahill, T. A., Feeney, P. J., & Eldred, R. A. (1987). Size

- time composition profil e of aerosols using the drum sampler. Nuclear Instruments and Methods in Physics Research Section B: Beam Interactions with Materials and Atoms, 22 (1

- 3), 344

- 348. http://dx.doi.org/10.1016/0168

- 5 83X(87)90355

- 7

Callender, E., & Rice, K. C. (2000). The urban environmental gradient: Anthropogenic influences on the spatial and temporal

distributions of lead and zinc in sediments. Environmental Science and Technology, 34 (2), 232

- 238. http://dx.doi.org/10.1021/es990380s

Canepari, S., Astolfi, M. L., Moretti, S., & Curini, R. (2010). Comparison of extracting solutions for elemental fractionation in airborne particulate matter. Talanta, 82 (2), 834

- 844. http: //dx.doi.org/10.1016/j.talanta.2010.05.068

Canepari, S., Cardarelli, E., Perrino, C., Catrambone, M., Pietrodangelo, A., & Strincone, M. (2006). Two

- stage chemical fractionation method for the analysis of elements and non

- volatile inorganic ions in PM10 samples: Application to ambient samples collected in Rome (Italy). Atmospheric Environment, 40 (40), 7908

- 7923. http://dx.doi.org/10.1016/j.atmosenv.2006.07.005

Canepari, S., Perrino, C., Olivieri, F., & Astolfi, M. L. (2008). Characterisation of the traffic sources of PM through size

segregated sampling, sequential leaching and ICP analysis. Atmospheric Environment, 42 (35), 8161

- 8175. http://dx.doi.org/10.1016/j.atmosenv.2008.07.052

Cánovas, C. R., Olías, M., Nieto, J. M., & Galván, L. (2010). Wash

- out processes of evaporitic sulfate salts in the Tinto river: Hydrogeochemical evolution and environme ntal impact. Applied Geochemistry, 25 , 288

- 301. http://dx.doi.org/10.1016/j.apgeochem.2009.11.014

Cantwell, M. G., Burgess, R. M., & King, J. W. (2008). Resuspension of contaminated field and formulated reference sediments Part I: Evaluation of metal release under controlled laboratory conditions. Chemosphere, 73 (11), 1824

- 1831. http://dx.doi.org/10.1016/j.chemosphere.2008.08.007

Caravanos, J., Weiss, A. L., Blaise, M. J., & Jaeger, R. J. ( 2006). A survey of spatially distributed exterior dust lead loadings in New York City. Environmental Research, 100 (2), 165

- 172. http://dx.doi.org/10.1016/j.envres.2005.05.001

Chalmers, A. T., Van Metre, P. C., & Callender, E. (2007). The chemical response of particle

- associated contaminants in aquatic sediments to urbanization in New England, USA. Jour nal of Contaminant Hydrology, 91 (1

- 2), 4

- 25. http://dx.doi.org/10.1016/j.jconhyd.2006.08.007

Chan, Y.

- C., Cohen, D. D., Hawas, O., Stelcer, E., Simpson, R., Denison, L., . . . Carswell, S. (2008). Apportionment of sources of fine and coarse particles in four major Australian cities by positive matrix factorisation. Atmospheric Environment, 42

, 374

- 389. http://dx.doi.org/10.1016/j.atmosenv.2007.09.030

Chen, S., Tsai, C., Huang, C., Chen, H., Lin, C., Tsai, J., & Chou, C. (2010). Chemical mass closure and chemical characteristics of ambient ultrafine particles and other PM fractions. Aerosol Science and Technology, 44 , 713

- 723.

Cheng, H., & Hu, Y. (2010). Lead (Pb) isotopic fingerprinting and its applications in lead pollution studies in China: A review. Environmental Pollution, 158 , 1134

- 1146. http://dx.doi.org/10.1016/j.envpol.2009.12.028

Chester, R., Lin, F. J., & Murphy, K. J. T. (1989). A three stage sequential leaching scheme for the characteri sation of the sources and environmental mobility of trace metals in the marine aerosol. Environmental Technology, 10 (10), 887

- 900. http://dx.doi.org/10.1080/09593338909384810

Choel, M., Deboudt, K., Flament, P., Lecornet, G., Perdrix, E., & Sobanska, S. (2006). Fast evolution of tropospheric Pb

- and Zn

rich particles in the vicinity of a lead smelter. A tmospheric Environment, 40 (24), 4439

- 4449. http://dx.doi.org/10.1016/j.atmosenv.2006.04.027

Chon, H

. - S., Ohandja, D.

- G., & Voulvoulis, N. (2010). Implementation of E.U. water framework directive: Source assessment of

metallic substances at catchment levels. Journal of Environmental Monitoring, 12 (1), 36

- 47. http://dx.doi.org/10.1039/b907851gMay 20 11 3

- 110 DRAFT

– DO NOT CITE OR QUOTE Chow, T. J., & Johnstone, M. S. (1965). Lead isotopes in gasoline and aerosols of Los Angeles Basin, California. Scien ce, 147 (3657), 502

- 503. http://dx.doi.org/10.1126/science.147.3657.502

Cizmecioglu, S. C., & Muezzino glu, A. (2008). Solubility of deposited airborne heavy metals. Atmospheric Research, 89 (4), 396

- 404. http://dx.doi.org/10.1016/j.atmosres.2008.03.012

Clark, H. F., Brabander, D. J., & Erdil, R. M. (2006). Sources, sinks, and exposure pathways of lead in urban garden soil. Journal of Environmental Quality, 35 (6), 2066

- 2074. http://dx.doi.org/10.2134/jeq2005.0464

Cloquet, C., Carignan, J., & Libourel, G. (2006). Isotopic compositio n of Zn and Pb atmospheric depositions in an urban/periurban area of northeastern France. Environmental Science and Technology, 40 (21), 6594

- 6600. http://dx.doi.org/10.1021/es0609654

Cobb, G. P., Abel, M. T., Rainwater, T. R., Austin, G. P., Cox, S. B., Kendall, R. J., . . . Presley, S. M. (2006). Metal distributions in New Orleans following hurrican es Katrina and Rita: A continuation study. Environmental Science and Technology,

40 (15), 4571

- 4577. http://dx.doi.org/10.1021/es060041g

Cohen, D. D., Crawford, J., Stelcer, E., & Bac, V. T. (2010). Characterisation and source apportionment of fine particulate

sources at Hanoi from 2001 to 2008. Atmospheric Environment, 44 , 320

- 328. http://dx.doi.org/10.1016/j.atmosenv.2009.10.037

Company, R., Serafim, A., Lopes, B., Cravo, A., Shepherd, T. J., Pearson, G., & B ebianno, M. J. (2008). Using biochemical and isotope geochemistry to understand the environmental and public health implications of lead pollution in the lower Guadiana River, Iberia: A freshwater bivalve study. Science of the Total Environment, 405 (1

- 3 ), 109

- 119. http://dx.doi.org/10.1016/j.scitotenv.2008.07.016

Cong, Z., Kang, S., Liu, X., & Wang, G. (2007). Elemental composition of aerosol in the Nam Co region, Tibetan Plateau, during

summer monsoon season. Atmospheric Environment, 41 , 1180

- 1187. http://dx.doi.org/10.1016/j.atmosenv .2006.09.046

Conor Pacific Environmental Technologies Inc. (2000). Airborne particulate matter, lead and manganese at Buttonville airport

. (Report No. 041

- 6710) . Toronto, Ontario: Environment Canada.

Courtin

- Nomade, A., Soubrand

- Colin, M., Marcus, M. A., & Fakra, S. C. (2008). Evidence for the incorporation of lead into barite from waste rock pile materials. Environmental Science and Technology, 42 (8), 2867

- 2872. http://dx.doi.org/10.1021/es702822k

Couture, R.

- M., J.

- F., C., Auger, D., Claisse, D., Gobeil, C., & Cossa, D. (2010). Seasonal and decadal variations in lead sources

to eastern North Atlantic mussels. Environmental S cience and Technology, 44 (4), 1211

- 1216. http://dx.doi.org/10.1021/es902352z

Coynel, A., Schafer, J., Blanc, G., & Bossy, C. (2007). Scenario of particulate trace metal and metalloid transport during a major flood event inferred from transient geochemical signals. Applied Geochemistry, 22 (4), 821

- 836. http://dx.doi.org/10.1016/j.apgeochem.2006.10.004

Curtis, C. J., Evans, C. D., Helliwell, R. C., & Monteith, D. T. (2005). Nitrate leaching as a confounding factor in chemical recovery from acidification in UK upland waters. Environmental Pollution, 137 (1), 73

- 82. http://dx.doi.org/10.1016/j.envpol.2004.12.032

Dahl, O., Poykio, R., & Nurmesniemi, H. (2008). Concentrations of heavy metals in fly ash from a coal

- fired power plant with respect to the new Finnish limit values. Journal of Material Cycles and Wa ste Management, 10 (1), 87

- 92. http://dx.doi.org/10.1007/s10163

- 007

- 0189

- 6

Dall'Osto, M., Booth, M. J., S mith, W., Fisher, R., & Harrison, R. M. (2008). A study of the size distributions and the chemical characterization of airborne particles in the vicinity of a large integrated steelworks. Aerosol Science and Technology,

42 (12), 981

- 991. http://dx.doi.org/10.1080/02786820802339587

Das, S. K., Routh, J., Roychoudhury, A. N., & Klump, J. V. (2008). Major and trace element geochemistry in Zeekoevlei, South Africa: A lacustrine record of present and past processes. Applied Geochemistry, 23 (8), 2496

- 2511. http://dx.doi.org/10.1016/ j.apgeochem.2008.02.011

Dawson, J. J., Tetzlaff, D., Carey, A. M., Raab, A., Soulsby, C., Killham, K., & Meharg, A. A. (2010). Characterizing Pb

mobilization f rom upland soils to streams using (206)Pb/(207)Pb isotopic ratios. Environmental Science and Technology,

44 (1), 243

- 249. http://dx.doi.org/10.1021/es902664dMay 20 11 3

- 111 DRAFT

– DO NOT CITE OR QUOTE de la Cruz, M. T., Laborda, F., Callen, M. S., Lopez, J. M., & Mastral, A. M. (2009). Study of Pb sources by Pb isotope ratios in the airborne PM of Zaragoza, Spain. Journal of Environmental Monitoring, 11 (11 ), 2052

- 2057. http://dx.doi.org/10.1039/b912274e

de la Cruz, M. T., Laborda, F., Callen, M. S., Lopez, J. M., & M astral, A. M. (2009). Study of Pb sources by Pb isotope ratios in the airborne PM10 of Zaragoza, Spain. Journal of Environmental Monitoring, 11 (11), 2052

- 2057. http://dx.doi.org/10.1039/b912274e

Deng, X., & Jennings, A. A. (2006). Evaluating an eletrokinetically driven extraction method for measuring heavy metal soil

contamination. Journal of Environmental Engi neering, 132 (4), 527

- 537. http://dx.doi.org/10.1061/(ASCE)0733

- 9372(2006)132:4(527 )

Dermont , G., Bergeron, M., Richer

- Laflèche, M., & Mercier, G. (2010). Remediation of metal

- contaminated urban soil using flotation technique. Science of the Total Environment, 408 (5), 1199

- 1211.

h ttp://dx.doi.org/10.1016/j.scitotenv.2009.11.036

Desta, M. B., Bruen, M., Higgins, N., & Johnston, P. (2007). Highway runoff quality in Ireland. Journal of Environmental Monitoring, 9 (4), 366

- 371. http://dx.doi.org/10.1039/b702327h

Diaz

- Somoano, M., Kylander, M. E., Lo pez

- Anton, M. A., Suarez

- Ruiz, I., Martinez

- Tarazona, M. R., Ferrat, M., . . . Weiss, D. J. (2009). Stable lead isotope compositions in selected coals from around the world and implications for present day aerosol source tracing. Environmental Science a nd Technology, 43 (4), 1078

- 1085. http://dx.doi.org/10.1021/es801818r

Dillner, A. M., Schauer, J. J., Zhang, Y. H ., Zeng, L. M., & Cass, G. R. (2006). Size

- resolved particulate matter composition in Beijing during pollution and dust events. Journal of Geophysical Research, 111 (D5), D05203. http://dx.doi.org/10.1 029/2005jd006400

Dos Santos, M., Gomez, D., Dawidowski, L., Gautier, E., & Smichowski, P. (2009). Determination of water

- soluble and insoluble compounds in siz e classified airborne particulate matter. Microchemical Journal, 91 (1), 133

- 139. http://dx.doi.org/10.1016/j.microc.2008.09.001

Dunlap, C. E., Alpers, C. N., Bouse, R., Taylor, H. E., Unruh, D. M., & Flegal, A. R. (2008). The persistence of lead from past gasoline emissions and mining drainage in a large riparian system: Evidence from lead isotopes in the Sacramento River,

California. Geochimica et Cosmochimica Acta, 72 (24), 5935

- 5948. http://dx.doi.org/10.1016/j.gca.2008.10.006

Duzgoren

- Aydin, N. S. (2007). Sources and characteristics of lead pollution in the urban environment of Guangzhou. Science of the Total Environment, 385 (1

- 3), 182

- 195. http://dx.doi.org/10.1016/j.scitotenv.2007.06.047

Duzgoren

- Aydin, N. S., & Weiss, A. L. (2008). Use and abuse of Pb

- isotope fingerprinting technique and G IS mapping data to

assess lead in environmental studies. Environmental Geochemistry and Health, 30 , 577

- 588. http://dx.doi.org/10.1007/s10653

- 008

- 9179

- 4

Edwards, R. D., Lam, N. L., Zhang, L., Johnson, M. A., & Kleinman, M. T. (2009). Nitrogen dioxide and ozone as factors in the availability of lead from lead

- based paints

. Environmental Science and Technology, 43 (22), 8516

- 8521. http://dx.doi.org/10.1021/es901077m

Erel, Y., Dayan, U., Rabi, R., Rudich, Y., & Stein, M. (2006). Trans boundary transport of pollutants by atmospheric mineral dust. Environmental Science and Technology, 40 (9), 2996

- 3005. http://dx.doi.org/10.1021/es05150 2l

Erel, Y., Listovsky, N., Matthews, A., Ilani, S., & Avni, Y. (2006). Tracing end

- member fluid sources in sub

- surface iron

mineralization and dolomitization along a proximal fault to the dead sea transform. Geochimica et Cosmochimica Acta,

70 (22), 5552

- 5570. http://dx.doi.org/10.1016/j.gca.2006.08.019

ERG. (Eastern Research Group Inc.). (2009). Technical assistance document for the national air toxics trends stations program:

Revision 2 . Research Triangle Park, NC: U.S. Environmental Protection Agency.

Ewen, C., Anagnostopoulou, M. A., & Ward, N. I. (2009). Monitoring of heavy metal levels in roadside dusts of Thessaloniki,

Greece in relation to motor vehicle traf fic density and flow. Environmental Monitoring and Assessment, 157 (1

- 4), 483

- 498. http://dx.doi.org/10.1007/s10661

- 008

- 0550

- 9

FAA. (U.S. Federal Aviation Administration). (2011). Emissions and Dispersion Modeling System (EDMS) (Version 5.1.3). Washington, DC: Author. Retrieved from http://www.faa.gov/about/office_org/headquarters_offices/apl/research/models/edms_model/May 20 11 3

- 112 DRAFT

– DO NOT CITE OR QUOTE Falgayrac, G., Sobanska, S., Laureyns, J., & Bremard, C. (2006). Heterogeneous chemistry between PbSO4 and calcite microparticles using Raman microimaging. Spectrochimica Acta Part A: Molecular and Biomolecular Spectroscopy,

64 (5), 1095

- 1101. http://dx.doi.org/10.1016/j.saa.2005.11.032

Falta, T., Limbeck, A., Koellensperger, G., & Hann, S. (2008). Bioaccessibility of selected trac e metals in urban PM2.5 and PM10 samples: A model study. Analytical and Bioanalytical Chemistry, 390 (4), 1149

- 1157. http://dx.doi.org/10.1007/s00216

- 007

- 1762

- 5

Fang, T. H., Li, J. Y., Feng, H. M., & Chen, H. Y. (2009). Distribution and contamination of trace metals in surface sediments of

the East China Sea. Marine Environmental Research, 68 (4), 178

- 187. http://dx.doi.org/10.1016/j.marenvres.2009.06.005

Farfel, M. R., Orlova, A. O., Lees, P. S. J., Roh de, C., Ashley, P. J., & Chisolm, J. J., Jr. (2003). A study of urban housing demolitions as sources of lead in ambient dust: demolition practices and exterior dust fall. Environmental Health Perspectives, 111 , 1228

- 1234.

Farfel, M. R., Orlova, A. O., Lees, P. S. J., Rohde, C., Ashley, P. J., & Chisolm, J. J., Jr. (2005). A study of urban housing demolition as a source of lead in ambient dust on sidewalks, s treets, and alleys. Environmental Research, 99 (2), 204

- 213. http://dx.doi.org/10.1016/j.envres.2004.10.005

Farmer, J. G., Eades, L. J., Graham, M. C., Cloy, J. M., & Bacon, J. R. (2010). A comparison of the isotopic composition of lead in rainwater, surface vegetation and tree bark at the long

- term monitoring site, Glensaugh, Scotland, in 2007. Science of the Total Environment, 408 (17), 3704

- 3710. http://dx.doi.org/10.1016/j.scitotenv.2010.03.050

Fa rmer, J. G., Eades, L. J., MacKenzie, A. B., Kirika, A., & Bailey

- Watts, T. E. (1996). Stable lead isotope record of lead pollution in Loch Lomond sediments since 1630 AD. Environmental Science and Technology, 30 , 3080

- 3083. http://dx.doi.org/10.1021/es960162o

Farmer, J. G., Graham, M. C., Bacon, J. R., Dunn, S. M., & Vinogradoff, S. I. (2005). Isotopic characterisation of the historical lead deposition record at Glensau, an organic

- rich, upland catchment in rural NE Scotland. Science of the Total Environment, 346 , 121

- 137. http://dx.doi.org/10.1016/j.scitotenv.2004.11.020

Fernandez, C., Labanowski, J., Cambier, P., Jongmans, A. G., & Van Oort, F. (2007). Fate of airborne metal p ollution in soils as related to agricultural management. 1. Zn and Pb distributions in soil profiles. European Journal of Soil Science, 58 (3),

547

- 559. http://dx.doi.org/10.1111/j.1365

- 2389.2 006.00827.x

Fernández Espinosa, A. J., & Ternero

- Rodríguez, M. (2004). Study of traffic pollution by metals in Seville (Spain) by physical and chemical speciati on methods. Analytical and Bioanalytical Chemistry, 379 , 684

- 699. http://dx.doi.org/10.1007/s00216

- 004

- 2640

- z

Fernandez Espinosa, A. J., Ternero Rodriguez, M., Barragan De La Rosa, F. J., & Jimenez Sanchez, J. C. (2002). A chemical speciation of trace metals for fine urban particles. Atmospheric Environment, 36 , 773

- 780. http://dx.doi.org/10.1016/S1352

- 2310(01)00534

- 9

Fernandez Espinosa, A. J., Ternero Rodriguez, M., & Fernandez Alvarez, F. (2004). So urce characterisation of fine urban particles by multivariate analysis of trace metals speciation. Atmospheric Environment, 38 , 873

- 886. http://dx.doi.org/10.1016/j.atmosenv.2003.10.046

Fernando, H. J. S. (2010). Fluid dynamics of urban atmospheres in complex terrain. Annual Review of Fluid Mechanics, 42 , 365

- 389. http://dx.doi.org/10.1146/annurev

- fluid

- 121108

- 145459

Field, P. M., & Sherrell, R. M. (2003). Direct determination of ultra

- trace levels of metals in fresh water using desolvating

micronebulization and HR

- ICP

- MS: Application to Lake Superior waters. Journal of Analytical Atomic Spectrometry, 18

, 254

- 259. http://dx.doi.org/10.1039/b210628k

Filippelli, G. M., & Laidlaw, M. A. S. (2010). The elephant in the playground: Confronting lead

- contaminated s oils as an important source of lead burdens to urban populations. Perspectives in Biology and Medicine, 53 (1), 31

- 45. http://dx.doi.org/10.1353/pbm.0.0136

Filippelli, G. M., Laidlaw, M. A. S., Latimer, J. C., & Raftis, R. (2005). Urban lead poisoning and medical geology: An unfinished story. Geological Society of America Today, 15 (1), 4

- 11. http://dx.doi.org/10.1130/1052

- 5173(2005)015<4:ULPAMG>2.0.CO;2

Finlayson

- Pitts, B. J., & Pitts, J. N., Jr. (2000). Chemistry of the upper and lower atmosphere: theory, experiments and applications . San Diego, CA: Academic Press.May 20 11 3

- 113 DRAFT

– DO NOT CITE OR QUOTE Flint, K. R., & Davis, A. P. (2007). Pollutant mass flushing characterization of highway stormwater runoff from an ultra

- urban

area. Journal of Environmental Engineering, 133 (6), 616

- 626. http://dx.doi.org/10.1061/ (ASCE)0733

- 9372(2007)133:6(616 )

Fowler, D., McDonald, A. G., Crossley, A., Nemitz, E., Leaver, D., Cape, J. N., . . . Harmens, H. (2006). UK heavy metal

monito ring network . (Report No. AS 06/07). Edinburgh, UK: NERC/Centre for Ecology and Hydrology. Retrieved from http://nora.nerc.ac.uk/3323

. Fowler, D., Smith, R. I., Muller, J. B., Hayman, G., & Vincent, K. J. (2005). Changes in the atmospheric deposition of acidifying compounds in the UK between 1986 and 2001. Environmental Pollution, 137 (1), 15

- 25. http://dx.doi.org/10.1016/j.envpol.2004.12.028

Franzaring, J., Holz, I., Zipperle, J., & Fangmeier, A. (2010). Twenty years of biological monitoring of element concentrations in permanent forest and grassland plots in Baden

- Württemberg (SW Germany). Environmental Science and Pollution Research, 17 (1), 4

- 12. http://dx.doi.org/10.1007/s11356

- 009

- 0181

- x

Fraser, J. L., & Lum, K. R. (1983). Availability of elements of environmental importance in incinerated s ludge ash. Environmental Science and Technology, 17 , 52

- 54. http://dx.doi.org/10.1021/es00107a013

Freme, F. (2004). U.S. coal supply and demand: 2004 Review . Washington, DC: U.S. Energy Information Administration.

Retrieved from http://tonto.eia.doe.gov/FTPROOT/features/feature04.pdf

. Fujiwara, F., Dos Santos, M., Marrero, J., Polla, G., Gomez, D., Dawidowskia, L., & Smichowski, P. (2006). Fractionation of eleven elements by chemical bonding from airborne particulate matter collected in an industrial city in Argentina. Journal of Environmental Monitoring, 8 (9), 913

- 922. http://dx.doi.org/10.1039/b604307k

Funasaka, K., Tojo, T., Katahira, K., Shinya, M., Miyazaki, T., Kamiura, T., . . . Takaoka, M. (2008). Detection of Pb

- LIII edge XANES spectra of urban atmospheric particles combined with simple acid extra ction. Science of the Total Environment,

403 (1

- 3), 230

- 234. http://dx.doi.org/10.1016/j.scitotenv.2008.05.020

Gallon, C., Tessier, A., & Gobeil, C. (2006). Historical perspective of industrial lead emissions to the atmosphere from a

Canadian smelter. Environmental Science and Technology, 40 (3), 741

- 747. http://dx.doi.org/10.1021/es051326g

Gao, Y., Lesven, L., Gillan, D., Sabbe, K., Billon, G., De Galan, S., . . . Le ermakers, M. (2009). Geochemical behavior of trace

elements in sub

- tidal marine sediments of the Belgian coast. Marine Chemistry, 117 (1

- 4), 88

- 96. http://dx.doi.org/10.1016/j.marchem.2009.05. 002

Garrido, F., Serrano, S., Campbell, C. G., Barrios, L., & Garcia

- Gonzalez, A. T. (2008). Evidence of physical and chemical nonequilibrium in lead and cadmi um transport and sorption in acid soils. Soil Science Society of America Journal, 72 (5),

1434

- 1444. http://dx.doi.org/10.2136/sssaj2007.0411

Geagea, M. L., Stille, P., Gauthier

- Lafaye, F., Perrone, T., & Aubert, D. (2008). Baseline determination of the atmospheric Pb, Sr and Nd isotopic compositions in the Rhine valley, Vosges mountains (France) and the Cent ral Swiss Alps. Applied Geochemistry, 23 (6), 1703

- 1714. http://dx.doi.org/10.1016/j.apgeochem.2008.02.004

Gidney, J., Twigg, M., & Kittelson, D. (2010). Effect of organometallic fuel additives on nanoparticle emissions from a gasoline

passenger car. Environmental Science and Technology, 44 (7), 2562

- 2569. http://dx.doi.org/10.1021/es901868c

Gieré, R., Blackford, M., & Smith, K. (2006). TEM study of PM2.5 emitted from coal and tire combustion in a thermal po wer station. Environmental Science and Technology, 40 (20), 6235

- 6240. http://dx.doi.org/10.1021/es060423m

Goforth , M. R., & Christoforou, C. S. (2006). Particle size distribution and atmospheric metals measurements in a rural area in

the South Eastern USA. Science of the Total Environment, 356 , 217

- 227. http://dx.doi.org/10.1016/j.scitotenv.2005.03.017

Goyal, A., Small, M. J., Von Stackelberg, K., & Bu rmistrov, D. (2005). Estimation of fugitive lead emission rates from secondary lead facilities using hierarchical Bayesian models. Environmental Science and Technology, 39 (13), 4929

- 4937. http://dx.doi.o rg/10.1021/es035465e

Graham, M. C., Vinogradoff, S. I., Chipchase, A. J., Dunn, S. M., Bacon, J. R., & Farmer, J. G. (2006). Using size fractionation

and Pb is otopes to study Pb transport in the waters of an organic

- rich upland catchment. Environmental Science and Technology, 40 (4), 1250

- 1256. http://dx.doi.org/10.1021/es0517670

Graney, J. R., Landis, M. S., & Norris, G. A. (2004). Concentrations and solubility of metals from indoor and personal exposure PM2.5 samples. Atmospheric Environment, 38 , 237

- 247. http://dx.doi.org/10.1016/j.atmosenv.2003.09.052May 20 11 3

- 114 DRAFT

– DO NOT CITE OR QUOTE Griffith, D. A. (2002). The geographic distribution of soil lead concentration: Description and concerns. URISA Journal, 14 (1), 5

- 16.

Griffith, D. A., Johnson, D. L., & Hunt, A. (2009). The geographic distribution of m etals in urban soils: The case of Syracuse,

NY. GeoJournal, 74 (4), 275

- 291. http://dx.doi.org/10.1007/s10708

- 008

- 9233

- x

Gulson, B., Korsch, M., Dickson, B., Cohen, D., Mizon, K., & Davis, J. M. (2007). Comparison of lead isotopes with source apportionment models, including SOM, for air particulates. Science of the Total Environment, 381 , 169

- 179. http://dx.doi.org/10.1016/j.scitotenv.2007.03.018

Guo, T., Sansalone, J., & Piro, P. (2006). The role of in situ unit operation/process infiltration treatment on partitioning and speciation of rainfall

- runoff. Water, Science and Technology, 54 (6

- 7), 255

- 261. http://dx.doi.org/10.2166/ws t.2006.589

Guo, X. Y., Zhang, S. Z., Shan, X. Q., Luo, L., Pei, Z. G., Zhu, Y. G., . . . Gault, A. (2006). Characterization of Pb, Cu, and Cd adsorption on par ticulate organic matter in soil. Environmental Toxicology and Chemistry, 25 (9), 2366

- 2373. http://dx.doi.org/10.1897/05

- 636R.1

Gutierrez

- Castillo, M. E., Olivos

- Ortiz, M., De Vizcaya

- Ruiz, A., & Cebrian, M. E. (2005). Chemical characterization of extractable water soluble matter associated with PM10 from Mexico City during 2000. Chemosphere, 61 (5), 701

- 710. http://dx.doi.org/10.1016/j.chemosphere.2005.03.063

H., K., M., K., & W., S. (2010). Atmospheric P b and Ti accumulation rates from sphagnum moss: Dependence upon plant productivity. Environmental Science and Technology, 44 (14), 5509

- 5515. http://dx.doi.org/10.1021/es100366d

Hahn, I., Brixey, L. A., Wiener, R. W., Henkle, S. W., & Baldauf, R. (2009). Characterization of traffic

- related PM concentration distribution and fluctuation patterns in near

- highway urba n residential street canyons. Journal of Environmental Monitoring, 11 (12), 2136

- 2145. http://dx.doi.org/10.1039/b907130j

Hallberg, M., Renman, G., & Lundbom, T. (2007). Seasonal variations of ten metals in highway runoff and their partition between dissolved and particulate matter. Water, Air, and Soil Pollution, 181 (1

- 4), 183

- 191. http://dx.doi.org/10.1007/s11270

- 006

- 9289

- 5

Han, L., Zhuang, G., Cheng, S., Wang, Y., & Li, J. (2007). Characteristics of re

- suspended road dust and its impact on the atmospheric environment in Beijing. Atmospheric Environment, 41 , 7485

- 7499.

Harmens, H., Norris, D. A., Koerber, G. R., Buse, A., Steinnes, E., & Ruhling, A. (2008). Temporal trends (1990

- 2000) in the concentration of cadmium, lead and mercury in mosses across Europe. Environmental Pollution, 151 (2), 368

- 376. http://dx.doi.org/10.1016/j.envpol.2007.06.043

Harmens, H., Norris, D. A., Steinnes, E., Kubin, E., Piispanen, J., Alber, R., . . . Zechmeister, H. G. (2010). Mosses as biomonitors of atmospheric heavy metal deposition: Spatial patterns and temporal trends in Europe. Environmental Pollution, 158 (10), 3144

- 3156. http://dx. doi.org/10.1016/j.envpol.2010.06.039

Harris, A. R., & Davidson, C. I. (2005). The role of resuspended soil in lead flows in the California South Coast Air Basin

. Environmental Science and Technology, 39 (19), 7410

- 7415. http://dx.doi.org/10.1021/es050642s

Harrison, R. M., T illing, R., Callen Romero, M. S., Harrad, S., & Jarvis, K. (2003). A study of trace metals and polycyclic aromatic hydrocarbons in the roadside environment. Atmospheric Environment, 37 , 2391

- 2402. http://dx.doi.org/10.1016/S1352

- 2310(03)00122

- 5

Hartnett, M., & Berry, A. (2010). Transport of lead in the Mersey Estuary: The development of a novel app roach to deriving partition coefficients. Advances in Engineering Software, 41 (1), 84

- 91. http://dx.doi.org/10.1016/j.advengsoft.2008.11.007

Hassellov, M., & von der Kammer, F. (2008). Iron oxides as geochemical nanovectors for metal transport in soil

- river systems. Elements, 4 (6), 401

- 406. http://dx.doi.org/10.2113/gselements.4.6.401

Hays, M. D., Cho, S.

- H., Baldauf, R., J., S. J., & Shafer, M. (2011). Particle size distributions of metal and non

- metal elements in

an urban near

- highway environment. Atmospheric Environment, 45 (4), 925

- 934. http://dx.doi.org/10.1016/j.atmosenv.2010.11. 010

Heal, M. R., Hibbs, L. R., Agius, R. M., & Beverland, I. J. (2005). Interpretation of variations in fine, coarse and black smoke particulate matter concent rations in a northern European city. Atmospheric Environment, 39 (20), 3711

- 3718. http://dx.doi.org/10.1016/j.atmosenv.2005.03.007

Helmreich, B., Hilliges, R., Schriewer, A., & Horn, H. (2010). Runoff pollutants of a highly trafficked urban road

-- correlation analysis and seasonal influences. Chemosphere, 80 (9), 991

- 997. http://dx.doi.org/10.1016/j.chemosphere.2010.05.037May 20 11 3

- 115 DRAFT

– DO NOT CITE OR QUOTE Hirner, A. V. (2006). Speciation of alkylated metals and metalloids in t he environment. Analytical and Bioanalytical Chemistry,

385 (3), 555

- 567. http://dx.doi.org/10.1007/s00216

- 006

- 0368

- 7

Hjortenkrans, D. S. T., Bergback, B. G., & Haggerud, A. V. (2007). Metal emissions from brake linings and tires: Case studies of Stockholm, Sweden 1995/1998 and 2005. Environmental Science and Technology, 41 (15), 5224

- 5230. http://dx.doi.org/10.1021/es070198o

Hofer, C., Gallagher, F. J., & Holzapfel, C. (2010). Metal accumulation and performance o f nestlings of passerine bird species at

an urban brownfield site. Environmental Pollution, 158 (5), 1207

- 1213. http://dx.doi.org/10.1016/j.envpol.2010.01.018

Horowitz, A. J., Elrick, K. A., & Smith, J. J. (2008). Monitoring urban impacts on suspended sediment, trace element, and nutrient fluxes within the City of Atlanta, Georgia, USA: Program design, methodological considerations, and initial

results. Hydrological Processes, 22 (10), 1473

- 1496. http://dx.doi.org/10.1002/hyp.6699

Horowitz, A. J., & Stephens, V. C. (2008). The effects of land use on fluvial sediment chemistry for the conterminous U.S. Results from the first cycle of the NAWQA Program: Trace and major elements, phosphorus, carbon, and sulfur. Science of the Total Environment, 400 (1

- 3), 290

- 314. http://dx.doi.org/10.1016/j.scitotenv.2008.04.027

Hossain, M. A., Furumai, H., Nakajima, F., & Aryal, R. K. (2007). Heavy metals speciation in soakaways sediment and evaluation of metal retention properties of surrounding soil. Water, Science and Technology, 56 (11), 81

- 89. http://dx.doi.org/10.2166/wst.2007.746

Hou, A., Laws, E., Gambrell, R., Tan, M., Delaune, R., Li, Y., & Roberts, H. (2006). Pathogen indicator microbes and heavy metals in Lake Pontchartrain following Hurricane Katrina. Environmental Scienc e and Technology, 40 (19), 5904

- 5910. http://dx.doi.org/10.1021/es060946u

Hsu, S. C., Chen Liu, S., Jeng, W. L., Chou, C. C., K. C., Hsu, R. T., Huang, Y. T., & Chen, Y. Y. (2006). Lead isotope ratios in ambient aerosols from Taipei, Taiwan: Identifying long

- range transport of airborne Pb from the Yangtze Delta. Atmospheric Environment, 40 (28), 5393

- 5404. http://dx.doi.org/10.1016/j.atmosenv.2006.05.003

Hu, S., Herner, J. D., Shafer, M., Robertson, W., Schauer, J. J., Dwyer, H., . . . Ayala, A. (2009). Metals emitted from heavy

- duty diesel vehicles equipped with advanced PM and NOx emission controls. Atmospheric Environment, 43 (18), 2950

- 2959.

htt p://dx.doi.org/10.1016/j.atmosenv.2009.02.052

Huang, S., & Conte, M. H. (2009). Source/process apportionment of major and trace elements in sinking particles in the Sargasso

sea. Geochimica et Cosmochimica Acta, 73 (1), 65

- 90. http://dx.doi.org/10.1016/j.gca.2008.08.023

Hur, S. D., Cunde, X., Hong, S., Barbante, C., Gabrielli, P., Lee, K., . . . Ming, Y. (2007). Seasonal patterns of heavy metal deposition to the snow on Lambe rt Glacier basin, East Antarctica. Atmospheric Environment, 41 (38), 8567

- 8578. http://dx.doi.org/10.1016/j.atmosenv.2007.07.012

Huston, R., Chan, Y. C., Gardner, T., Shaw, G., & Chapman, H. (2009). Characterisation of atmospheric deposition as a source of

contaminants in urban rainwater tanks. Water Research, 43 (6), 1630

- 1640. http://dx.doi.org/10.1016/j.watres.2008.12.045

Ikegawa, M., Kimura, M., Honda, K., Akabane, I., Makita, K., Motoyama, H., . . . Itokawa, Y. (1999). Geographical variations of major and trace elements in East Antarctica. Atmospheric Environment, 33 (9), 1457

- 1467. http://dx.doi.org/10.1016/S1352

- 2310(98)00243

- X

Illinois Environmental Protection Agency. (2002). Chicago O’Hare Airport air toxic monitoring program . Springfield, IL:

Author. Retrieved from http://www.epa.state.il.us/air/ohare/

.
Ishizaka, T., Tohno, S., Ma, C.

- J., Morikawa, A., Takaoka, M., Nishiyama, F., & Yamamoto, K. (2009). Reactivity between PbS04 and CaCO3 particles relevant to the modification of mineral particles and chemical forms of Pb in particles sampled at two remote sites during an Asian dust event. Atmospheric Environment, 43 (16), 2550

- 2560. http://dx.doi.org/10.1016/j.atmosenv.2009.02.041

Jennings, A. A., & Ma, J. (2007). Variation in North Ame rican regulatory guidance for heavy metal surface soil contamination at

commercial and industrial sites. Journal of Environmental Engineering and Science, 6 (5), 587

- 609. http://dx.doi.org/10.1139/s07

- 010

Jensen, M. B., Holm, P. E., Laursen, J., & Hansen, H. C. B. (2006). Contaminant aspects of blackish surface deposits on highway

roadsides. Water, Air, and Soil Pollution, 175 (1

- 4), 305

- 321. http://dx.doi.org/10.1007/s11270

- 006

- 9140

- zMay 20 11 3

- 116 DRAFT

– DO NOT CITE OR QUOTE Jitaru, P., Infante, H. G., & Adams, F. C. (2004). Simultaneous multi

- elemental speciation analysis of organometallic compounds by solid

- phase microextraction and multicapillary gas chromatography hyphenated to inductively coupled plasma

- time

of

- flight

- mass spectrometry. Journal of Analytical Atomic Spectrometry, 19 , 867

- 875. http://dx.doi.org/10.1039/B404106B

Johnson, K. S., de Foy, B., Zuber i, B., Molina, L. T., Molina, M. J., Xie, Y., . . . Shutthanandan, V. (2006). Aerosol composition and source apportionment in the Mexico City Metropolitan Area with PIXE/PESA/STIM and multivariate analysis. Atmospheric Chemistry and Physics, 6 , 4591

- 460 0. http://dx.doi.org/10.5194/acp

- 6

- 4591

- 2006

Johnson, K. S., Laskin, A., Jimenez, J. L., Shutthanandan, V. , Molina, L. T., Salcedo, D., . . . Molina, M. J. (2008). Comparative analysis of urban atmospheric aerosol by particle

- induced X

- ray emission (PIXE), proton elastic scattering analysis (PESA), and aerosol mass spectrometry (AMS). Environmental Science and Technology, 42 (17), 6619

- 6624. http://dx.doi.org/10.1021/es800393e

Joshi, S. R. (2008). Influence of road side pollution on the phylloplane microbial community of Alnus nepalensis (Betulaceae). Revista de Biologia Tropical, 56 (3), 1521

- 1529. http://www.ncbi.nlm.nih.gov/pubmed/19419061

Kang, S., Zhang, Q., Kaspari, S., Qin, D., Cong, Z., Ren, J., & Mayewski, P. A. (2007). Spatial and seasonal variations of elemental composition in Mt. Everest (Qomolangma) snow /firn. Atmospheric Environment, 41 (34), 7208

– 7218. http://dx.doi.org/10.1016/j.atmosenv.2007.05.024

Kaste, J., Friedland, A., & Sturup, S. (2003). Using stable and radioactive isotopes to trace atmospherically deposited Pb in

montane forest soils. Environmental Science and Technology, 37 (16), 3560

- 3567. http://dx.doi.org/10.1021/es026372k

Kaste, J. M., Bostick, B. C., Friedland, A. J., Schroth, A. W., & Siccama, T. G. (2006). Fate and speciation of gasoline

- d erived lead in organic horizons of the northeastern USA. Soil Science Society of America Journal, 70 (5), 1688

- 1698. http://dx.doi.org/10.2136/sssaj2005.0321

Kastner

- Klein, P., Fedorovich, E., Ketzel, M., R, B., & Britter, R. (2003). The modelling of turbulence from traffic in urban dispersion models

– part II: Evaluation against laboratory and full

- scale con centration measurements in street canyons. Environmental Fluid Mechanics, 3 (2), 145

- 172. http://dx.doi.org/10.1023/A:1022049224166

Kastner

- Klein, P., Fedorovich, E., & Rotach, M. W. (2001). A wind tunnel study of organised and turbulent air motions in urban
street canyons. Journal of Wind Engineering and Industrial Aerodynamics, 89 (9), 849

- 861. http://dx.doi.org/10.1016/S0167

- 6105(01)00074

- 5

Kastner

- Klein, P., & Rotach, M. W. (2004). Mean flow and turb ulence characteristics in an urban roughness sublayer. Boundary

Layer Meteorology, 111 (1), 55

- 84. http://dx.doi.org/10.1023/B:BOUN.0000010994.32240.b1

Kayhanian, M., Suverkropp, C., Ruby, A., & Tsay, K. (2007). Characterization and prediction of highway runoff constituent event

mean concentration. Journal of Environmental Management, 85 (2), 279

- 2 95. http://dx.doi.org/10.1016/j.jenvman.2006.09.024

Kidwell, C. B., & Ondov, J. M. (2001). Development and evaluation of a prototype system for collecting sub

- hourly ambient aerosol for chemical analysis. Aerosol Science and Technology, 35 , 596

- 601. http://dx.doi.org/10.1080/0278682 0118049

Kidwell, C. B., & Ondov, J. M. (2004). Elemental analysis of sub

- hourly ambient aerosol collections. Aerosol Science and Technology, 38 , 205

- 218. http://dx.doi.org/10.1080/02786820490261726

Kimbrough, K. L., Lauenstein, G. G., Christensen, J. D., & Apeti, D. A. (2008). An assessment of two decades of contaminant monitoring in the nation's coastal zone . Silver Spring, MD: National Centers for Coastal Ocean Science. Retrieved from http://aquaticcommons.org/22 32/

. Klaminder, J., Bindler, R., Emteryd, O., Appleby, P., & Grip, H. (2006). Estimating the mean residence time of lead in the organic

horizon of boreal forest soils using 210

- lead, stable lead and a soil chronosequence. Biogeochemistry, 78 , 31

- 49. http://dx.doi.org/10.1007/s10533

- 005

- 2230

- y

Klaminder, J., Bindler, R., Emteryd, O., & Renberg, I. (2005). Uptake and recycling of lead by boreal forest plants: Quantitative

estimates from a site in northern Sweden. Geochimica et Cosmochimica Acta, 69 , 2485

- 2496. http://dx.doi.org/10.1016/j.gca.2004.11.013

Klaminder, J., Bindler, R., Laudon, H., Bishop, K., Emteryd, O., & Renberg, I. (2006). Flux rates of atmospheric lead pollution within soils of a small catchment in northern Sweden and their implications for future stream water quality. Environmental Science and Technology, 40 (15), 4639

- 4645. http://dx.doi.org/10.1021/es0520666May 20 11 3

- 117 DRAFT

– DO NOT CITE OR QUOTE Klaminder, J., Bindler, R., & Renberg, I. (2008). The biogeochemistry of atmospherically derived Pb in the boreal forest of

Sweden. Applied Geochemistry, 23 (10), 2922

- 2931. http://dx.doi.org/10.1016/j.apgeochem.2008.04.007

Klaminder, J., Bindler, R., Rydberg, J., & Renberg, I. (2008). Is there a chronological record of atmospheric mercury and lead deposition preserved in the mor layer (O

- h orizon) of boreal forest soils? Geochimica et Cosmochimica Acta, 72 (3), 703

- 712. http://dx.doi.org/10.1016/j.gca.2007.10.030

Klaminder, J., Hammarlund, D., Kokfelt, U., Vonk, J. E., & Bigler, C. (2010). Lead contamination of subarctic lakes and its response to reduced atmospheric fallout: Can the recovery process be counteracted by the ongoing climate cha nge? Environmental Science and Technology, 44 (7), 2335

- 2340. http://dx.doi.org/10.1021/es903025z

Knowlton, S. W. , & Moran, S. B. (2010). Stable Pb isotope ratios in aerosols, precipitation, and size

- fractionated particulate matter in the Gulf of Maine, Scotian Shelf, and Labrador Sea. Marine Pollution Bulletin, 60 (7), 984

- 989. http://dx.doi.org/10.1016/j.marpolbul.2010.02.005

Kylander, M. E., Cortizas, A. M., Rauch, S., & Weiss, D. J. (2008). Lead penetrat ion and leaching in a complex temperate soil profile. Environmental Science and Technology, 42 (9), 3177

- 3184. http://dx.doi.org/10.1021/es702358e

Kyotani, T., & Iwatsuki, M. (2002). Characterization of soluble and insoluble components in PM25 and PM10 fractions of airborne particulate matter in Kofu city, Japan. Atmospheric Environment, 36 , 639

- 649. http://dx.doi.org/10.1016/S1352

- 2310(01)00494

- 0

Lai, X., Norisuye, K., Mikata, M., Minami, T., Bowie, A. R., & So hrin, Y. (2008). Spatial and temporal distribution of Fe, Ni, Cu and Pb along 140 degrees E in the Southern Ocean during austral summer 2001/02. Marine Chemistry, 111 (3

- 4), 171

- 183. http://dx.doi.org/10.1016/j.marchem.2008.05.001

Laidlaw, M. A. S., & Filippelli, G. M. (2008). Resuspension of urban soils as a persistent source of lead poiso ning in children: A

review and new directions. Applied Geochemistry, 23 (8), 2021

- 2039. http://dx.doi.org/10.1016/j.apgeochem.2008.05.009

Landers, D. H., Simonich, S. L., Jaffe, D. A., Geiser, L. H., Campbell, D. H., Schwindt, A. R., . . . Erway, M. M. (2008). The fate, transport, and ecological impacts of airborne contaminants in western national pa rks (USA) . (Report No. EPA/600/R

07/138). Corvallis, Oregon: U.S. Environmental Protection Agency, NHEERL, Western Ecology Division. Retrieved

from http://www.nature.nps.gov/air /studies/air_toxics/WACAPreport.cfm

. Landers, D. H., Simonich, S. M., Jaffe, D., Geiser, L., Campbell, D. H., Schwindt, A., . . . Erway, M. M. (2010). The West ern Airborne Contaminant Assessment Project (WACAP): An interdisciplinary evaluation of the impacts of airborne

contaminants in western U.S. National Parks. Environmental Science and Technology, 44 (3), 855

- 859. http://dx.doi.org/10.1021/es901866e

Landre, A. L., Watmough, S. A., & Dillon, P. J. (2009). The effects of dissolved organic carbon, acidity and seasonali ty on metal geochemistry within a forested catchment on the Precambrian Shield, central Ontario, Canada. Biogeochemistry, 93 (3),

271

- 289. http://dx.doi.org/10.1007/s10533

- 009

- 9305

- 0

Landre, A. L., Watmough, S. A., & Dillon, P. J. (2010). Metal pools, fluxes, and budgets in an acidified forested catchment on the Precambrian Shield, Central Ontario, Canada. Water, Air, and Soil Pollution, 209 (1

- 4), 209

- 228. http://dx.doi.org/10.1007/s11270

- 009

- 0193

- 7

Lara

- Caz enave, M. B., Levy, V., Castetbon, A., Potin

- Gauthier, M., Astruc, M., & Albert, E. (1994). Pollution of urban runoff waters by heavy metals: Part I. Total metal. Environmental Technology, 15 , 1135

- 1147. http://dx.doi.org/10.1080/09593339409385523

Lasheen, M. R., Sharaby, C. M., El

- Kholy, N. G., Elsherif, I. Y., & El

- Wakeel, S. T. (2008). Factors influencing lead and iron release from some Egyptian drinking water pipes. Journal of Hazardous Materials, 160 (2

- 3), 2

- 3. http://dx.doi.org/10.1016/j.jh azmat.2008.03.040

Leal

- Granadillo, I. A., Alonso, J. I. G., & Sanz

- Medel, A. (2000). Determination of the speciation of organolead compounds in airborne particu late matter by gas chromatography

- inductively coupled plasma mass spectrometry. Analytica Chimica
Acta, 423 , 21

- 29. http://dx.doi.org/10.1016/S0003

- 2670(00)01032

- 1

Lee, G., Faure, G., Bigham, J. M., & Williams, D. J. (2008). Metal release from bottom sediments of Ocoee Lake No. 3, a primary catchment area for the Ducktown Mining District. Journal of E nvironmental Quality, 37 (2), 344

- 352. http://dx.doi.org/10.2134/jeq2007.0223

Lee, K., Do Hur, S., Hou, S. G., Hong, S. M., Qin, X., Ren, J. W., . . . Boutron, C. F. (2008). Atmospheric pollution for trace elements in the remote high

- altitude atmosphere in central Asia as recorded in snow from Mt. Qomolangma (Everest) of the Himalayas. Science of the Total Envi ronment, 404 (1), 171

- 181. http://dx.doi.org/10.1016/j.scitotenv.2008.06.022May 20 11 3

- 118 DRAFT

– DO NOT CITE OR QUOTE Levin, R., Brown, M. J ., Kashtock, M. E., Jacobs, D. E., Whelan, E. A., Rodman, J., . . . Sinks, T. (2008). Lead exposures in US children, 2008: Implications for prevention. Environmental Health Perspectives, 116 (10), 1285

- 1293. http://dx.doi.org/10.1289/ehp.11241

Lewis, J., Sjostrom, J., Skyllberg, U., & Hagglund, L. (2010). Distribution, chemical speciation, and mobility of lead and antimony originating from small arms ammunition in a coarse

- grained unsaturated surface sand. Journal of Environmental Quality, 39 (3), 863

- 870. http://dx.doi.org/10.2134/jeq2009.0211

Li, C., Nguyen, Q., Ryan, P. H., LeMasters, G. K., Spitz, H., Lobaugh, M., . . . Grinshpun, S. A. (2009). School bus pollution and changes in the air quality at schools: A case study. Journal of Environmental Monitoring, 11 , 1037

- 1042. http://dx.doi.org/10.1039/b819458k

Li, L., Pal a, F., Jiang, M., Krahforst, C., & Wallace, G. T. (2010). Three

- dimensional modeling of Cu and Pb distributions in Boston Harbor, Massachusetts and Cape Cod Bays. Estuarine, Coastal and Shelf Science, 88 (4), 450

- 463. http://dx.doi.org/10.1016/j.ecss.2010.05.003

Londahl, J., Massling, A., Swietlicki, E., Brauner, E. V., Ketzel, M., Pagels, J., & Loft, S. (2009). Experimentally determined human respiratory tract deposition of airborne particles at a busy street. Environmental Science and Technology, 43

, 4659

- 4664. http://dx.doi.org/10.1021/es803029b

Magill, N., & Sansalone, J. (2010). Distribution of particulate

- bound metals for source area snow in the Lake Tahoe watershed. Journal of Environmental Engineering, 136 (2), 185

- 193. http://dx.doi.org/10.1061/(ASCE)EE.1943

- 7870.0000146

Makkonen, U., Hellen, H. , Anttila, P., & Ferm, M. (2010). Size distribution and chemical composition of airborne particles in

south

- eastern Finland during different seasons and wildfire episodes in 2006. Science of the Total Environment, 408 (3),

644

- 651. http://dx.doi.org/10.1016/j.scitotenv.2009.10.050

Martuzevicius, D., Grinshpun, S. A., Reponen, T., Gorny, R. L., Shukla, R., Lockey, J., . . . LeMasters, G. (2004). Spatial and temporal variations of PM2.5 concentration and composition throughout an urban area with high freeway density: The Greater Cincinnati study. Atmospheric Environment, 38 (8), 1091

- 1105. http://dx.doi.org/10.1016/j.atmosenv.2003.11.015

Marx, S. K., Kamber, B. S., & McGowan, H. A. (2008). Scavenging of atmospheric trace metal pollutants by mineral dusts: Inter

regional transport of Australian trace metal pollution to New Zealand. Atmospheric Environment, 42 (10), 2460

- 2478. http://dx.doi.org/10.1016/j.atmosenv.2007.12.014

McGee, C. J., Fernandez, I. J., Norton, S. A., & Stubbs, C. S. (2007). Cd, Ni, Pb, and Zn concentrations i n forest vegetation and

soils in Maine. Water, Air, and Soil Pollution, 180 (1

- 4), 141

- 153. http://dx.doi.org/10.1007/s11270

- 006

- 9257

- 0

McKenzie, E. R., Money, J. E., Green, P. G., & Young, T. M. (2009). Metals associated with stormwater

- relevant brake and tire samples. Science of the Total Environment, 407 (22), 5855

- 5860. http://dx.doi.org/10.1016/j.scitotenv.2009.07.018

McKenzie, E. R., Wong, C. M., Green, P. G., Kayhanian, M., & Young, T. M. ( 2008). Size dependent elemental composition of

road

- associated particles. Science of the Total Environment, 398 (1

- 3), 145

- 153. http://dx.doi.org/10.1016/j.scitotenv.2008.02.052

Melaku, S., Morris, V., Raghavan, D., & Hosten, C. (2008). Seasonal variation of heavy metals in ambient air and precipitation at a single site in Washington, DC. Environment al Pollution, 155 (1), 88

- 98. http://dx.doi.org/10.1016/j.envpol.2007.10.038

Mielke, H. W., & Gonzales , C. (2008). Mercury (Hg) and lead (Pb) in interior and exterior New Orleans house paint films. Chemosphere, 72 , 882

- 885. http://dx.doi.org/10.1016/j.chemosphere.2008.03.061

Mielke, H. W., Gonzales, C., Powell, E., & Mielke, P. W., Jr. (2008). Urban soil

- lead (Pb) footprint: Retrospective comparison of public and private properties in New Orleans. Environmental Geochemistry and Health, 30 (3), 231

- 242. http://dx.doi.org/10.1007/s1 0653

- 007

- 9111

- 3

Mielke, H. W., Gonzales, C. R., Powell, E., Jartun, M., & Mielke, P. W., Jr. (2007). Nonlinear association between soil lead and

blood lead of children in metropolitan New Orleans, Louisiana: 2000

- 2005. Science of the Total Environment, 388 (1

- 3),

43

- 53. http://dx.doi.org/10.1016/j.scitotenv.2007.08.012

Mielke, H. W., Laidlaw, M. A., & Gonzales, C. (2010a). Lead (Pb) legacy from vehicle traffic in eight California urbanized areas: Continuing influence of lead dust on children's health. Scie nce of the Total Environment, 408 (19), 3965

- 3975. http://dx.doi.org/10.1016/j.scitotenv.2010.05.017

Mielke, H. W., Laidlaw, M. A., & Gonzales, C. R. (2010b). Estimation of leaded (Pb) gasoline's continuing material and health impacts on 90 US urbanized areas. Environment International, 37 , 248

- 257. http://dx.doi.org/10.1016/j.envint.2010.08.006May 20 11 3

- 119 DRAFT

– DO NOT CITE OR QUOTE Mielke, H. W., Powell, E. T., Gonzales, C. R., & Mielke, P. W., Jr. (2007). Potential lead on play su rfaces: Evaluation of the "PLOPS" sampler as a new tool for primary lead prevention. Environmental Research, 103 , 154

- 159. http://dx.doi.org/10.1016/j.envres.2006.08.007

Moffet, R. C., de Foy, B., Molina, L. T., Molina, M. J., & Prather, K. A. (2008). Measurement of ambient aerosols in northern Mexico City by single particle mass spectrometry. Atmosphe ric Chemistry and Physics, 8 (16), 4499

- 4516. http://dx.doi.org/10.5194/acp

- 8

- 4499

- 2008

Moffet, R. C., Desy aterik, Y., Hopkins, R. J., Tivanski, A. V., Gilles, M. K., Wang, Y., . . . Prather, K. A. (2008). Characterization of aerosols containing Zn, Pb, and Cl from an industrial region of Mexico City. Environmental Science and Technology,

42 (19), 7091

- 7097. http://dx.doi.org/10.1021/es7030483

Molnár, P., Bellander, T., Sallsten, G., & Boman, J. (2007). Indoor and outdoor concentrations of PM2.5 trace elements at hom es, preschools and schools in Stockholm, Swed en. Journal of Environmental Monitoring, 9 (4), 348

- 357. http://dx.doi.org/10.1039/b616858b

Moreno, T., Querol, X. , Pey, J., Minguillon, M. C., Perez, N., Alastuey, A., . . . Gibbons, W. (2008). Spatial and temporal variations in inhalable CuZnPb aerosols within the Mexico City pollution plume. Journal of Environmental Monitoring,

10 (3), 370

- 378. http://dx.doi.org/10.1039/b716507b

Mukai, H., Tanaka, A., Fujii, T., Zeng, Y., Hong, Y., Tang, J., . . . Zhai, P. (2001). Region al characteristics of sulfur and lead isotope ratios in the atmosphere at several Chinese urban sites. Environmental Science and Technology, 35 (6), 1064

– 1071. http://dx.doi.org/10.1021/es001399u

Murakami, M., Nakajima, F., Furumai, H., Tomiyasu, B., & Owari, M. (2007). Identification of particles containing chromium and lead in road dust and soakaway sediment by electron probe microanalyser. Chemosphere, 67 (10), 2000

- 2010. http://dx.doi.org/10.1016/j.chemosphere.2006.11.044

Murphy, D. M., Hudson, P. K., Cziczo, D. J., Gallavardin, S., Froyd, K. D., Johnston, M. V., . . . Wexler, A. S. (2007). Distribution of lead in single atmospheric particles. Atmospheric Chemistry and Physics, 7 (12), 3195

- 3210.

N'guessan, Y. M., Probst, J. L., Bur, T., & Probst, A. (2009). Trace elements in stream bed sediments from agricultural catchments (Gascogne region, S

- W France): Whe re do they come from? Science of the Total Environment, 407 (8), 2939

- 2952. http://dx.doi.org/10.1016/j.scitotenv.2008.12.047

National Ambient Air Quality Standards for Lead (final rule), 73 66964

- 67062 (2008).

Niisoe, T., Nakamura, E., Harada, K., Ishikawa, H., Hitomi, T., Watanabe, T., . . . Koizumi, A. (2010). A global transport model of lead in the atmosphere. Atmospheric Environment, 44 (14), 1806

- 1814. http://dx.doi.org/10.1016/j.atmosenv. 2010.01.001

Nishiguchi, K., Utani, K., & Fujimori, E. (2008). Real

- time multielement monitoring of airborne particulate matter using ICP

- MS instrument equipped with gas converter apparatus. Journal of Analytical Atomic Spectrometry, 23 (8), 1125

- 1129. http://dx.doi.org/10.1039/b802302f

Noble, S. R., Horstwood, M. S. A., Davy, P., Pashley, V., Spiro, B., & Smith, S. (2008). Evolving Pb isotope signatures of London airborne particulate matter (PM10): Constraints from on

- filter and solution

- mode MC

- ICP

- MS. Journal of Enviro nmental Monitoring, 10 (7), 830

- 836. http://dx.doi.org/10.1039/b802151a

Norton, S. A. (2007). Atmospheric metal pollutants: Archives, methods, and history. Water, Air, and Soil Pollution, 7 , 93

- 98. http://dx.doi.org/10.1007/s11267

- 006

- 9089

- 3

Novak, M., Erel, Y., Zemanova, L., Bottrell, S. H., & Adamova, M. (2008). A comparison of lead pollution record in sphagnum peat with known historical Pb emission rates in the British Isles and the Czech Republic. Atmospheric Environment,

42 (40), 8997

- 9006. http://dx.doi.org/10.1016/j.atmosenv.2008.09.031

NRCS. (Natura l Resources Conservation Service). (2011). USDA National Water and Climate Center, from http://www.wcc.nrcs.usda.gov/ftpref/downloads/climate/windrose/

Oberts, G. L. (2000). Influence of snowmelt dynamics on stormwater runoff quality. Watershed Protection Techniques, 1 (2), 55

- 61.

Ogulei, D., Hopke, P. K., Zhou, L., Pancras, J. P., Nair, N., & Ondov, J. M. (2006). Source apportionment of Baltimore aerosol

from combined size distribution and chemic al composition data. Atmospheric Environment, 40 , 396

- 410. http://dx.doi.org/10.1016/j.atmosenv.2005.11.075May 20 11 3

- 120 DRAFT

– DO NOT CITE OR QUOTE Ondov, J. M., Buckley, T. J., Hopke, P. K., Ogulei, D., Parlange, M. B., Rogge, W. F., . . . Wexler, A. S. (2006). Baltimore Supersite: Highly time

- and size

- resolved concentrations of urban PM2.5 and its constituents for resolution of sources and immune responses. Atmospheric Environment, 40 , 224

- 237. http://dx.doi.org/10.1016/j.atmosenv.2005.11.072

Pacyna, E. G., Pacyna, J. M., Fudala, J., Strzelecka

- Jastrzab, E., Hlawiczka, S., Panasiuk, D., . . . Friedrich, R. (2007). Current and future emissions of selected heavy metals to the atmosphere from anthropogenic sources in Europe. Atmospher ic Environment, 41 (38), 8557

- 8566. http://dx.doi.org/10.1016/j.atmosenv.2007.07.040

Pacyna, J. M., & Pacyna, E. G. (2001). An assessment of global and regional emissions of trace metals to the atmosphere from anthropogenic sources worldwide. Environmental Reviews, 9 , 269

- 298. http://dx.doi.org/10.1139/a0 1

- 012

Pancras, J. P., Ondov, J. M., Poor, N., Landis, M. S., & Stevens, R. K. (2006). Identification of sources and estimation of emission profiles from highly time

- resolved pollutant measurements in Tampa, FL. Atmospheric Environment, 40 , 467

- 481. http://dx.doi.org/10. 1016/j.atmosenv.2005.12.036

Pancras, J. P., Ondov, J. M., & Zeisler, R. (2005). Multi

- element electrothermal AAS determination of 11 marker elements in fine

amb ient aerosol slurry samples collected with SEAS

- II. Analytica Chimica Acta, 538 (1

- 2), 303

- 312. http://dx.doi.org/10.1016/j.aca.2005.01.062

Paode, R. D., Sofuoglu, S. C., Sivadechathep, J., Noll, K. E., & Holsen, T. M. (1998). Dry deposition fluxes and mass size distributions of Pb, Cu, and Zn measured in southern Lake Michigan during AEOLOS. Environmental Science and Technology, 32 (11), 1629

– 1635. http://dx.doi.org/10.1021/es970892b

Pardue, J. H., Moe, W. M., McInn is, D., Thibodeaux, L. J., Valsaraj, K. T., Maciasz, E., . . . Yuan, Q. Z. (2005). Chemical and microbiological parameters in New Orleans floodwater following Hurricane Katrina. Environmental Science and Technology, 39 (22), 8591

- 8599. http://dx.doi.org/10.1021/es0518631

Patra, A., Colvile, R., Arnold, S., Bowen, E., Shallcross, D., Martin, D., . . . Robins, A. (2008). On street observations of particulate matter movement and dispersion due to traffic on an urban road. Atmospheric Environment, 42 , 3911

- 3926. http://dx.doi.org/10.1016/j.atmosenv.2006. 10.070

Pédrot, M., Dia, A., Davranche, M., Bouhnik

- Le Coz, M., Henin, O., & Gruau, G. (2008). Insights into colloid

- mediated trace

element release at the soil/ water interface. Journal of Colloid and Interface Science, 325 (1), 187

- 197. http://dx.doi.org/10.1016/j.jcis.2008.05.019

Pekey, B., Bozkurt, Z. B., Pekey, H., Doğan, G., Zararsiz, A., Efe, N., & Tuncel, G. (2010). Indoor/outdoor concentrations and elemental composition of PM10/PM2. 5 in urban/industrial areas of Kocaeli City, Turkey. Indoor Air, 20 (2), 112

- 125. http://dx.doi.org/10.1111/j.1600

- 0668.2009.00628.x

Pekney, N. J., Davidson, C. I., Bein, K. J., Wexler, A. S., & Johnston, M. V. (2006). Identification of sources of atmospheric PM at the Pittsburgh supersite, Part I: Single particle analysis and filter

- based positive matrix factorization. Atmospheric
Environment, 40 (Suppl. 2), 411

- 423. http://dx.doi.org/10.1016/j.atmosenv.2005.12.072

Perdrial, N., Liewig, N., Delphin, J.

- E., & Elsass, F. (2008). TEM evidence for intracellular accumulation of lead by bacteria in

subsurface environments. Chemical Geology, 253 (3

- 4), 196

- 204. http://dx.doi.org/10.1016/j.chemgeo.2008.05.008

Perrino, C., Canepari, S., Pappalardo, S., & Marconi, E. (2010). Time

- resolved measurements of water

- soluble ions and elements in atmospheric particulate matter f or the characterization of local and long

- range transport events. Chemosphere, 80 (11),

1291

- 1300. http://dx.doi.org/10.1016/j.chemosphere.2010.06.050

Pingitore, N. E., Jr., Clague, J. W., Amaya, M. A., Maciejewska, B., & Reynoso, J. J. (2009). Urban airborne lead: X

- ray absorption spectroscopy establishes soil as dominant source. PLoS ONE, 4 (4), e 5019. http://dx.doi.org/10.1371/journal.pone.0005019

Planchon, F. A. M., van de Velde, K., Rosman, K. J. R., Wolff, E. W., Ferrari, C. P., & Boutron, C. F. (2003). One hundred fifty

year record of lead isotopes in Antarctic snow from Coats Land. Geochimica et Cosmochimica Acta, 67 (4), 693

- 708. http://dx.doi.org/10.1016/S0016

- 7037(02)01136

- 5

Point, D., Monperrus, M., Tessier, E., Amouroux, D., Chauvaud, L., Thouzeau, G., . . . Donard, O. F. X. (2007 ). Biological control of trace metal and organometal benthic fluxes in a eutrophic lagoon (Thau Lagoon, Mediterranean Sea, France). Estuarine, Coastal and Shelf Science, 72 (3), 457

- 471. http://dx.d oi.org/10.1016/j.ecss.2006.11.013

Pokrovsky, O. S., & Schott, J. (2002). Iron colloids/organic matter associated transport of major and trace elements in small boreal rivers and their estuaries (NW Russia). Chemical Geology, 190 (1

- 4), 141

- 179. http://dx.doi.org/10.1016/S0009

- 2541(02)00115

- 8May 20 11 3

- 121 DRAFT

– DO NOT CITE OR QUOTE Pokrovsky, O. S., Schott, J., & Dupré, B. (2006). Trace element fractionation and transport in boreal rivers and soil porewaters of permafrost

- dominated basaltic terrain in Central Siberia. Geochimica et Cosmochimica Ac ta, 70 (13), 3239

- 3260. http://dx.doi.org/10.1016/j.gca.2006.04.008

Pokrovsky, O. S., Viers, J., Shirokov a, L. S., Shevchenko, V. P., Filipov, A. S., & Dupré, B. (2010). Dissolved, suspended, and colloidal fluxes of organic carbon, major and trace elements in the Severnaya Dvina River and its tributary. Chemical Geology, 273 (1

- 2), 136

- 149. http://dx.doi.org/10.1016/j.chemgeo.2010.02.018

Poperechna, N., & Heumann, K. G. (2005). Species

- specific GC/ICP

- I DMS for trimethyllead determinations in biological and environmental samples. Analytical Chemistry, 77 (2), 511

- 516. http://dx.doi.org/10.1021/ac048757m

Popp, M., Hann, S., & Koellensperger, G. (2010). Environmental application of elemental speciation analysis based on liquid or gas chromatography hyphenated to inductively coupled plasma mass spectrometry

-- a revie w. Analytica Chimica Acta,

668 (2), 114

- 129. http://dx.doi.org/10.1016/j.aca.2010.04.036

Poykio, R., Nurm esniemi, H., & Dahl, O. (2007). Concentrations of nutrients and heavy metals in cyclone fly ash from the grate

fired boiler at a small municipal district heating plant (6 MW). Journal of Residuals Science and Technology, 4 (3), 127

- 136.

Prather, K. A., Nordmeyer, T., & Salt, K. (1994). Real

- time characterization of individual aerosol particles using time

- of

- flight
mass spectrometry. Analytical Chemistry, 66 (9 ), 1403

- 1407. http://dx.doi.org/10.1021/ac00081a007

Pratt, C., & Lottermoser, B. G. (2007). Mobilisation of traffic

- derived trace metals from road corridors into coastal stream and estuarine sediments, Cairns, northern Australia. Environmental Geology, 52 (3), 437

- 448. http://dx.doi.org/10.1007/s00254

- 006

- 0471

- 2

Presley, S. M., Abel, M. T., Austin, G. P., Rainwater, T. R., Brown, R. W., McDaniel, L. N., . . . Cobb, G. P. (201 0). Metal concentrations in schoolyard soils from New Orleans, Louisiana before and after hurricanes Katrina and Rita. Chemosphere, 80 (1), 67

- 73. http://dx.doi.org/10.1016/j.chemosphere.2010 .03.031

Presley, S. M., Rainwater, T. R., Austin, G. P., Platt, S. G., Zak, J. C., Cobb, G. P., . . . Kendall, R. J. (2006). Assessment of pathogens and toxica nts in New Orleans, LA following hurricane Katrina. Environmental Science and Technology, 40 (2),

468

- 474. http://dx.doi.org/10.1021/es052219p

Pruvot, C., Douay, F., Hervé, F., & Waterlot, C. (2006). Heavy metals in soil, crops and grass as a source of human exposure in the former mining areas. Journal of Soils and Sediments, 6 (4), 215

- 220. http://dx.doi.org/10.1065/jss2006.10.186

Pyrzyńska, K. (1996). Organolead speciation in environmental samples: A review. Microch imica Acta, 122 (3

- 4), 279

- 293. http://dx.doi.org/10.1007/BF01245788

Qi, J., Feng, L., Li, X., & Zhang, M. (2 006). An X

- ray photoelectron spectroscopy study of elements on the surface of aerosol particles. Journal of Aerosol Science, 37 (2), 218

- 227. http://dx.doi.org/10.1016/j.jaerosci.2005.04.005

Quevauviller, P. (2000). Accuracy and traceability in environmental monitoring

— pitfalls in methylmercury determinations as a case study. Journal of Environmental Monitoring, 2 , 292

- 299. http://dx.doi.org/10.1039/b001852j

Quinton, J. N., & Catt, J. A. (2007). Enrichment of heavy metals in sediment resulting from soil erosion on agricultural fields. Environmental Science and Technology, 41 (10), 3495

- 3500. http://dx.doi.org/10.1021/es062147h

Raabe, O. G., Braaten, D. A., Axelbaum, R. L., Teague, S. V., & Cahill, T. A. (1988). Calibration studies of the drum impactor. Journal of Aerosol Science, 19 , 183

- 195. http://dx.doi.org/10.1016/0021

- 8502(88)90222

- 4

Radakovitch, O., Roussiez, V., Ollivier, P., Ludwig, W., Grenz, C., & Probst, J. L. (2008). Input of particulate heavy metals from rivers and associated sedimentary deposits on the Gulf of Lion continental shelf. Estuarine, Coastal and Shelf Science,

77 (2), 285

- 295. http://dx.doi.org/10.1016/j.ecss.2007.09.028

Rasmussen, P. E. (1997). Long

- range atmospheric transport of trace metals: The need for geoscience perspectives. Environmental Geology, 33 (23), 96

- 108. http://dx.doi.org/10.1007/s002540050229

Rauch, J. N., & Pacyna, J. M. (2009). Earth's global Ag, Al, Cr, Cu, Fe, Ni, Pb, and Zn cycles. Global Biogeochemical Cycles,

23 (2), GB2001. http://dx.doi.org/10.1029/2008gb003376

Reference Method for the Determination of Coarse Particulate Matter as PM10

- 2.5 in the Atmosphere, 40, C.F.R. § 50 Appendix

O (2010).May 20 11 3

- 122 DRAFT

– DO NOT CITE OR QUOTE Reimann, C., Flem, B., Arnoldussen, A., Englmaier, P., Finne, T. E., Koller, F., & Nordgulen, O. (2008). The biosphere: A homogeniser of Pb

- isotope signals. Applied Geochemistry, 23 (4), 705

- 722. http://dx.doi.org/10.1016/j.apgeochem.2007.12.002

Reimann, C., Kashulina, G., de Caritat, P., & Nisk avaara, H. (2001). Multi

- element, multi

- medium regional geochemistry in the European Arctic: Element concentration, variation and correlation. Applied Geochemistry, 16 (7

- 8), 759

- 780. http://d x.doi.org/10.1016/S0883

- 2927(00)00070

- 6

Reinard, M. S., Adoua, K., Martini, J. M., & Johnston, M. V. (2007). Source characterization and identification by real

- time single particle mass spectrometry. Atmospheric Environment, 41 (40), 9397

- 9409. http://dx.doi.org/10.1016/j.atmosenv.2007.09.001

Rice, J. (2007). Memorandum: Summary of method detection limits for ambient lead methods . Research Triangle Park, NC: United States Environmental Protection Agency.

Richter, P., Grino, P., Ahumada, I., & Giordano, A. (2007). Total element concentration and chemical fractionation in airborne particulate matter from Santiago, Chile. Atmospheric Environment, 41 , 6729

- 6738. http://dx.doi.org/10.1016/j.atmosenv.2007.04.053

Riediker, M., Williams, R., Devlin, R., Griggs, T. , & Bromberg, P. (2003). Exposure to particulate matter, volatile organic compounds, and other air pollutants inside patrol cars. Environmental Science and Technology, 37 , 2084

- 2093.

Rodes, C. E., & Evans, E. G. (1985). Preliminary assessment of 10 "mu"m particulate sampling at eight locations in the United

States. Atmospheric Environment, 19 , 293

- 303. http://dx.doi.org/10.1016/0004

- 6981(85)90097

- 6

Rooney, C. P., McLaren, R. G., & Condron, L. M. (2007). Control of lead solubility in soil contaminated with lead shot: Effect of soil pH. Environmental Pollution, 149 (2), 149

- 157. http://dx.doi.org/10.1016/j.envpol.2007.01.009

Root, R. A. (2000). Lead loading of urban streets by motor vehicle wheel weights. Environmental Health Perspectives, 108 , 937

- 940. http://dx.doi.org/10.1289/ehp.001089 37

Ross, J. M., & Sherrell, R. M. (1999). The role of colloids in tracemetal transport and adsorption behavior in New Jersey. Limnology and Oceanography, 44 (4) , 1019

- 1034.

Rothwell, J. J., Evans, M. G., & Allott, T. E. H. (2007). Lead contamination of fluvial sediments in an eroding blanket peat

catchment. Applied G eochemistry, 22 (2), 446

- 459. http://dx.doi.org/10.1016/j.apgeochem.2006.11.002

Rothwell, J. J., Ev ans, M. G., & Allott, T. E. H. (2008). In

- stream processing of sediment

- associated metals in peatland fluvial
systems. Water, Air, and Soil Pollution, 187 (1

- 4), 53

- 64. http://dx.doi.org/10.1007/s 11270

- 007

- 9496

- 8

Rothwell, J. J., Evans, M. G., Daniels, S. M., & Allott, T. E. H. (2007). Baseflow and stormflow metal concentrations in streams draining cont aminated peat moorlands in the Peak District National Park (UK). Journal of Hydrology, 341 (1

- 2), 90

- 104. http://dx.doi.org/10.1016/j.jhydrol.2007.05.004

Rothwell, J. J., Lindsay, J. B., Evans, M. G., & Allott, T. E. H. (2010). Modelling suspended sediment lead concentrations in contaminated peatland catchments using digital terrain analysis. Ecologica l Engineering, 36 (5), 623

- 630. http://dx.doi.org/10.1016/j.ecoleng.2008.10.010

Roulier, J. L., Belaud, S., & Coquery, M. (2010). Comparison of dynamic mobilization of Co, Cd and Pb in sediments using DGT and metal mobility assessed by sequential extraction. Chemosphere, 79 (8), 839

- 843. http://dx.doi.org/10.1016/j.chemosphere.2010.02.056

Roulier, S., Robinson, B., Kuster, E., & Schulin, R. (2008). Analysing the preferential transport of l ead in a vegetated roadside soil using lysimeter experiments and a dual

- porosity model. European Journal of Soil Science, 59 (1), 61

- 70. http://dx.doi.org/10.1111/j.1365

- 2389.2007.00954.x

Sabin, L. D., Lim, J. H., Stolzenbach, K. D., & Schiff, K. C. (2006). Atmospheric dry deposition of trace metals in the coastal region of Los Angeles, California, USA. Environmental Toxicology and Chemistry, 25 (9), 2334

- 2341. http://dx.doi.org/10.1897/05

- 300R.1

Sabin, L. D., Lim, J. H., Venezia, M. T., Winer, A. M., Schiff, K. C., & Stolzenbach, K. D. (2006). Dry deposition and resuspension of particle

- associated metals near a freeway in Los Angeles. Atmospheric Environment, 40 , 7528

- 7538. http://dx.doi.org/10.1016/j.atmosenv.2006.07.004

Sakata, M., & Asakura, K. (2008). Evaluating relative contribution of atmospheric mercury species to mercury dry deposition in Japan. Water, Air, and Soil Pollution, 193 (1

- 4), 51

- 63. http://dx.doi.org/10.1007/s11270

- 008

- 9667

- 2May 20 11 3

- 123 DRAFT

– DO NOT CITE OR QUOTE Salcedo, D., Onasch, T. B., Aiken, A. C., Williams, L. R., de Foy, B., Cubison, M. J., . . . Jimenez, J. L. (2010). Determination of particulate lead using aerosol mass spectrometry: MILAGRO/MCMA

- 2006 observatio ns. Atmospheric Chemistry and Physics, 10 , 5371

- 5389. http://dx.doi.org/10.5194/acp

- 10

- 5371

- 2010

Sammut, M. L., Noack, Y., Rose, J., Hazemann, J. L., Proux, O., Depoux, M., . . . Fiani, E. (2010). Speciation of Cd and Pb in dust emitted from sinter plant. Chemosphere, 78 (4), 445

- 450. http://dx.doi.org/10.1016/j.chemosphere.2009.10.039

Sanchez

- Ccoyllo, O. R., Ynoue, R. Y., Martins, L. D., Astolfo, R., Miranda, R. M., Freitas, E. D., . . . Andra de, M. F. (2009). Vehicular particulate matter emissions in road tunnels in Sao Paulo, Brazil. Environmental Monitoring and Assessment,

149 (1

- 4), 241

- 249. http://dx.doi.org/10.1007/s10661

- 008

- 019 8

- 5

Sansalone, J., Ying, G., & Lin, H. (2010). Distribution of metals for particulate matter transported in source area rainfall

- runoff. Journal of Environment al Engineering, 136 (2), 172

- 184. http://dx.doi.org/10.1061/(ASCE)EE.1943

- 7870.0000139

Sansalone , J. J., & Buchberger, S. G. (1996). Characterization of metals and solids in urban highway winter snow and spring

rainfall

- runoff. Transportation research record, 1523 , 147

- 159. http://dx.doi.org/10.3141/ 1523

- 18

Sato, K., Tamura, T., & Furuta, N. (2008). Partitioning between soluble and insoluble fractions of major and trace elements in

size

- classified airborne particulate matter collected in Tokyo. Journal of Environmental Monitoring, 10 (2), 211

- 218. http://dx.doi.org/10.1039/B709937A

Schauer, J. J., Lough, G. C., Shafer, M. M., Christensen, W. F., Arndt, M. F., DeMinter, J. T., & Park, J.

- S. (2006). Characterization of metals emitted from motor vehicles . (Report No. 133). Boston, MA: Health Effects Institute.

Retrieved from http://pubs.healtheffects.org/view.php?id=150

.
Scheetz, C. D., & Rimstidt, J. D. (2009). Diss olution, transport, and fate of lead on a shooting range in the Jefferson National Forest near Blacksburg, VA, USA. Environmental Geology, 58 (3), 655

- 665. http://dx.doi.org/10.1007/s00254

- 008

- 1540

- 5

Scheid, S., Gunthardt

- Goerg, M. S., Schulin, R., & Nowack, B. (2009). Accumulation and solubility of metals during leaf litter decomposition in non

- polluted a nd polluted soil. European Journal of Soil Science, 60 (4), 613

- 621. http://dx.doi.org/10.1111/j.1365

- 2389.2009.01153.x

Schleicher, N., Norra, S., Chai, F., Chen, Y., Wang, S., & Stuben, D. (2010). Anthropogenic versus geogenic contribution to total suspended atmospheric particulate matter and its variations during a two

- year sampling period in Beij ing, China. Journal of Environmental Monitoring, 12 (2), 434

- 441. http://dx.doi.org/10.1039/b914739j

Schroth, A. W ., Bostick, B. C., Kaste, J. M., & Friedland, A. J. (2008). Lead sequestration and species redistribution during soil organic matter decomposition. Environmental Science and Technology, 42 (10), 3627

- 3633. http://dx.doi.org/10.1021/es703002b

Schulz

- Zunkel, C., & Krueger, F. (2009). Trace metal dynamics in floodplain soils of the River Elbe: A review. Journal of Environmental Quality, 38 (4), 1349

- 1362. http://dx.doi.org/10.2134/jeq2008.0299

Schuwirth, N., Voegelin, A., Kretzschmar, R., & Hofmann, T. (2007). Vertical distribution and speciation of trace metals in weathering flotation residues of a zinc/lead sulfide mine. Journal of Environmental Quality, 36 (1), 61

- 69. http://dx.doi.org/10.2134/jeq2006.0148

Schwab, A. P., Zhu, D. S., & Banks, M. K. (2008). Influence of organic acids on the transport of heavy metals in soil

. Chemosphere, 72 (6), 986

- 994. http://dx.doi.org/10.1016/j.chemosphere.2008.02.047

Schwab, K. J., Gibson, K. E., Williams, D. L., Kulbicki, K. M., Lo, C. P., Mihalic, J. N., . . . Geyh, A. S. (2007). Microbial and chemical assessment of regions within New Orleans, LA impacted by hurricane Katrina. Environmental Science and Technology, 41 (7), 240 1

- 2406. http://dx.doi.org/10.1021/es062916x

Sengor, S. S., Spycher, N. F., Ginn, T. R., Sani, R. K., & Peyton, B

. (2007). Biogeochemical reactive

- diffusive transport of heavy

metals in Lake Coeur d'Alene sediments. Applied Geochemistry, 22 (12), 2569

- 2594. http://dx.doi.org/10.1016/j.apgeochem.2007.06 .011

Shi, H., Witt, E., III, Shu, S., Su, T., Wang, J., & Adams, C. (2010). Toxic trace element assessment for soils/sediments deposited during hurricanes Katr ina and Rita from southern Louisiana, USA: A sequential extraction analysis. Environmental Toxicology and Chemistry, 29 (7), 1419

- 1428. http://dx.doi.org/10.1002/etc.218

Shotbolt, L. A., Rothwell, J. J., & Lawlor, A. J. (2008). A mass balance approach to quantifying Pb storage and fluxes in an upland catchment of the Peak District, north

- central England. Earth Surfa ce Processes and Landforms, 33 (11), 1721

- 1741. http://dx.doi.org/10.1002/esp.1644May 20 11 3

- 124 DRAFT

– DO NOT CITE OR QUOTE Shotyk, W., & Krachler, M. ( 2007). Lead in bottled waters: Contamination from glass and comparison with pristine groundwater. Environmental Science and Technology, 41 (10), 3508

- 3513. http://dx.doi.org/10.1021/es062964h

Shotyk, W., & Krachler, M. (2009). Determination of trace element concentrations in natural freshwaters: How low is "low,', and how low do we need to go? Journal of Environme ntal Monitoring, 11 (10), 1747

- 1753. http://dx.doi.org/10.1039/b917090c

Shotyk, W., & Krachler, M. (2010). The isotopic evolution of atmospheric Pb in central Ontario since AD 1800, and its impacts on the soils, waters, and sediments of a forested watershed, Kawagama Lake. Geochimica et Cosmochimica Acta, 74 (7),

1963

- 1981. http://dx.doi.org/10.1016/j.gca.2010.01.009

Shotyk, W., Krachler, M., Aeschbach

- Hertig, W., Hillier, S., & Zheng, J. C. (2010). Tr ace elements in recent groundwater of an artesian flow system and comparison with snow: Enrichments, depletions, and chemical evolution of the water. Journal of Environmental Monitoring, 12 (1), 208

- 217. http ://dx.doi.org/10.1039/B909723F

Shotyk, W., & Le Roux, G. (2005). Biogeochemistry and cycling of lead. In A. Sigel, H. Sigel & R. K. O. Sigel (Eds.), Biogeochem ical cycles of elements (Vol. 43, pp. 239

- 275). Boca Raton, FL: Taylor & Francis.

Shotyk, W., Weiss, D., Heisterkamp, M., Cheburkin, A. K., Appleby, P. G., & Adam s, F. C. (2002). New peat bog record of atmospheric lead pollution in Switzerland: Pb concentrations, enrichment factors, isotopic composition, and organolead species. Environmental Science and Technology, 36 (18), 3893

- 3900. http://dx.doi.org/10.1021/es010196i

Sillanpaa, M., Saarikoski, S., Hillamo, R., Pennanen, A., Makkonen, U., Spolnik, Z., . . . Salonen, R. O

. (2005). Chemical composition, mass size distribution and source analysis of long

- range transported wildfire smokes in Helsinki. Science of the Total Environment, 350 (1

- 3), 119

- 135. http:/ /dx.doi.org/10.1016/j.scitotenv.2005.01.024

Silva, P. J., & Prather, K. A. (1997). On

- line characterization of individual particles from automobile emissions. E nvironmental Science and Technology, 31 , 3074

- 3080. http://dx.doi.org/10.1021/es961063d

Singh, M., Jaques, P. A., & Sioutas, C. (2002). Size distribution and diurnal characteristics of particle

- bound metals in source and receptor sites of the Los Angeles Basin. Atmospheric Environment, 36 , 1675

- 1689. http://dx.doi.org/10.1016/S1352

- 2310(02)00166

- 8

Smichowski, P., Polla, G., & Gomez, D. (2005). Metal fractionation of atmospheric aerosols via sequential chemi cal extraction: a

review. Analytical and Bioanalytical Chemistry, 381 (2), 302

- 316. http://dx.doi.org/10.1007/s00216

- 004

- 2849

- x

Smichowski, P., Polla, G., Gomez, D., Espinosa, A. J. F., & Lopez, A. C. (2008). A three

- step metal fractionation scheme for fly ashes collected in an Argentine thermal power plant. Fuel, 87 (7), 1249

- 1258. http://dx.doi.org/10.1016/j.fuel.2007.07.011

Snyder, D. C., Schauer, J. J., Gross, D. S., & Turner, J. R. (2009). Estimating the contribution of point sources to atmospheric metals using single

- particle mass spectrometry. Atmospheric Environment, 43 (26), 4033

- 4042. http://dx.doi.org/10.1016/j.atmosenv.2009.05.011

Soto

- Jiménez, M. F., & Flegal, A. R. (2009). Origin of lead in the Gulf of California Ecoregion using stable isotope analysis. Journal of Geochemical Exploration , 101 (3), 209

- 217. http://dx.doi.org/10.1016/j.gexplo.2008.07.003

Soto

- Jimenez, M. F., & Paez

- Osuna, F. (2010). A first approach to study the mobility and behavior of lead in hypersaline salt marsh sediments: Diffusive and advective fluxes, geochemical partitioning and Pb isotopes. Journal of Geochemical Exploration, 104 (3), 87

- 96. http://dx.doi.org/10.1016/j.gexplo.2009.12.006

Soulhac, L., Garbero, V., Salizzoni, P., Mejean, P., & Perkins, R. J. ( 2009). Flow and dispersion in street intersections. Atmospheric Environment, 43 (18), 2981

- 2996. http://dx.doi.org/10.1016/j.atmosenv.2009.02.061

Spalinger, S. M., von Braun, M. C., Petrosyan, V., & von Lindern, I. H. (2007). Northern Idaho house dust and soil lead levels compared to the Bunker Hill superfund site

. Environmental Monitoring and Assessment, 130 (1

- 3), 57

- 72. http://dx.doi.org/10.1007/s10661

- 006

- 9450

- z

Sriskandan, S., Faulkner, L., & Hopkins, P. (2007). Streptococcus pyogenes: Insight into the function of the streptococcal superantigens. International Journal of Biochemistry and Cell Biology, 39 (1), 12

- 19. http://dx.doi.org/10.1016/j.biocel.2006.08.009

Steinnes, E., Sjobakk, T. E., Donisa, C., & Brannvall, M. L. (2005). Quantification of pollutant lead in forest soils. Soil Science Society of America Journal, 69 (5), 1399

- 1404. http://dx.doi.org/10.2136/sssaj2004.0095

Stolpe, B., & Hassellov, M. (2007). Changes in size distribution of fresh water nanoscale colloidal matter and associated elements on mixing with seawater. Geochimica et Cosmochimica Acta, 71 (13), 3292

- 3301. http://dx.doi.org/10.1016/j.gca.2007.04.025May 20 11 3

- 125 DRAFT

– DO NOT CITE OR QUOTE Sturm, A., Crowe, S. A., & Fowle, D. A. (2008). Tr ace lead impacts biomineralization pathways during bacterial iron reduction. Chemical Geology, 249 (3

- 4), 282

- 293. http://dx.doi.org/10.1016/j.chemgeo.2008.01.017

Swarbrick, J. C., Skyllberg, U., Karlsson, T., & Glatzel, P. (2009). High energy resolution X

- ray absorption spectroscopy of environmentally relevant lead(II) compounds. Inorganic Chemistry, 48 (22), 10748

- 10756. http://dx.doi.org/10.1021/ic9015299

Sweet, C. W., Weiss, A., & Vermette, S. J. (1998). Atmospheric deposition of trace metals at three sites near the Great Lakes. Water, Air, and Soil Pollution, 103 , 423

- 439. http://dx.doi.org/10.1023/A:1004905832617

Taillefert, M., Lienemann, C.

- P., Gaillard, J.

- F., & Perret, D. (2000). Speciation, reactivity, and cycling of Fe and Pb in a

meromictic lake. Geochimica et Cosmochimica Acta, 64 (2), 169

- 183. http://dx.doi.org/10.1016/S0016

- 7037(99)00285

- 9

Takamatsu, T., Watanabe, M., Koshikawa, M. K., Murata, T., Yamamura, S., & Hayashi, S. (2010). Pollution of montane soil with Cu, Zn, As, Sb, Pb, and nitrate in Kanto, Japan. Science of the Total Environment, 408 (8), 1932

- 1942. http:// dx.doi.org/10.1016/j.scitotenv.2010.01.016

Tan, M. G., Zhang, G. L., Li, X. L., Zhang, Y. X., Yue, W. S., Chen, J. M., . . . Shan, Z. C. (2006). Comprehensive study of lead pollution in Shanghai by multiple techniques. Analytical Chemistry, 78 (23), 8044

- 8050. http://dx.doi.org/10.1021/ac061365q

Taylor, M. P., Mackay, A. K., Hudson

- Edwards, K. A., & Holz, E. (2010). Soil Cd, Cu, Pb and Zn contaminants around Mount Isa city, Queensland, Australia: Potential sources and risks to human health. Applied Geochemistry, 25 (6), 841

- 855. http://dx.doi.org/10.1016/j.apgeochem.2010.03.003

Tessier, A., Campbell, P. G. C., & Bisso n, M. (1979). Sequential extraction procedure for the speciation of particulate trace

metals. Analytical Chemistry, 51 , 844

- 851. http://dx.doi.org/10.1021/ac50043a017

Tipping, E., Lawlor, A. J., Lofts, S., & Shotbolt, L. (2006). Simulating the long

- term chemistry of an upland UK catchment: Heavy metals. Environmental Pollution, 141 (1), 139

- 150. http://dx.doi.org/10.1016/j.envpol.2005.08.019

Tuccillo, M. E. (2006). Size fractionation of metals i n runoff from residential and highway storm sewers. Science of the Total
Environment, 355 , 288

- 300. http://dx.doi.org/10.1016/j.scitotenv.2005.03.003

U.S. EPA. (U.S. Environmental Protection Agency). (1986). Air quality criteria for lead . (Report No. EPA/600/8

- 83/028aF

- dF). Washington, DC: Author.

U.S. EPA. (U.S. Environmental Protection Agency). (1999a). Compendium of methods for the determination of inorganic compounds in ambient air: Determination of metals in ambient particulate matter using inductive ly coupled plasma/mass spectrometry (ICP/MS) (Compendium Method IO

- 3.5) . (Report No. EPA/625/R

- 96/010a). Cincinnati, OH: Author.

Retrieved from http://www.epa.gov/ttn/amtic/fil es/ambient/inorganic/mthd

- 3

- 5.pdf

.

U.S. EPA. (U.S. Environmental Protection Agency). (1999b). Particulate matter (PM2.5) speciation guidance: Final draft (Editi on 1) . Research Triangle Park, NC: U.S. Environmental Protection Agency, Office of Air Quality Planning and

Standards.

U.S. EPA. (U.S. Environmental Protectio n Agency). (2006). Air quality criteria for lead . (Report No. EPA/600/R

- 05/144aF

- bF). Research Triangle Park, NC: U.S. Environmental Protection Agency, National Center for Environmental Assessment.

Retrieved from http://cfpub.epa.gov/ncea/CFM/recordisplay.cfm?deid=158823

.

U.S. EPA. (U.S. Environmental Protection Agency). (2008a). 2005 Nat ional Emissions Inventory data and documentation, from http://www.epa.gov/ttn/chief/net/2005inventory.html

U.S. EPA. (U.S. Environmental Protection Agency). (2008b). California area designations for the 2008 lead National Ambient Air Quality Standards . Research Triangle Park, NC: Author.

U.S. EPA. (U.S. Environmental Protection Agency). (2008c). Florida area designations for the 2008 lead National Ambient Air Quality Standards . Research Tria ngle Park, NC: Author.

U.S. EPA. (U.S. Environmental Protection Agency). (2008d). Missouri area designations for the 2008 lead National Ambient Air Quality S tandards . Research Triangle Park, NC: Author.

U.S. EPA. (U.S. Environmental Protection Agency). (2008e). Ohio area designations for the 2008 lead National A mbient Air Quality Standards . Research Triangle Park, NC: Author.

U.S. EPA. (U.S. Environmental Protection Agency). (2008f). Tennessee area designations for the 2008 lead National Ambient Air Quality Standards . Research Triangle Park, NC: Author.May 20 11 3

- 126 DRAFT

– DO NOT CITE OR QUOTE U.S. EPA. (U.S. Environmental Protection Agency). (2009). Integrate d science assessment for particulate matter . (Report No.

EPA/600/R

- 08/139F). Research Triangle Park, NC: U.S. Environmental Protection Agency, Office of Research and Development. Retrieved from http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=216546

.

U.S. EPA. (U.S. Environmental Protection Agency). (2010). Our nation's air: Status and t rends through 2008 . (Report No. EPA

454/R

- 09

- 002). Research Triangle Park, NC: U.S. Environmental Protection Agency, Office of Air Quality Planning and

Standards. Retrieved from http://w ww.epa.gov/airtrends/2010/report/fullreport.pdf

.

U.S. EPA. (U.S. Environmental Protection Agency). (2011). 2008 National Emissions Inventory data and documenta tion, from http://www.epa.gov/ttn/chief/net/2008inventory.html

University of California Davis. (1995). IMPROVE Data Guide . Davis, CA: Author.

USGS. (U.S. Geological Survey). (2005). Historical statistics for mineral and material commodities in the Uni ted States, from http://minerals.usgs.gov/ds/2005/140/

Uzu, G., Sobanska, S., Aliouane, Y., Pradere, P., & Dum at, C. (2009). Study of lead phytoavailability for atmospheric industrial

micronic and sub

- micronic particles in relation with lead speciation. Environmental Pollution, 157 (4), 1178

- 1185. http ://dx.doi.org/10.1016/j.envpol.2008.09.053

van Herwijnen, R., Hutchings, T. R., Ai

- Tabbaa, A., Moffat, A. J., Johns, M. L., & Ouki, S. K. (2007). Remediation o f metal

contaminated soil with mineral

- amended composts. Environmental Pollution, 150 (3), 347

- 354. http://dx.doi.org/10.1016/j.envpol.2007.01.023

Vermette, S., Irvine, K. N., & Drake, J. J. (1991). Temporal variability of the elemental composition in urban street dust. Environmental Monitoring and Assessment, 18 (1 ), 69

- 77. http://dx.doi.org/10.1007/BF00394478

Verstraete, S., & Van Meirvenne, M. (2008). A multi

- stage sam pling strategy for the delineation of soil pollution in a

contaminated brownfield. Environmental Pollution, 154 (2), 184

- 191. http://dx.doi.org/10.1016/j.envpol.2007.10.014

Vogl, J., & Heumann, K. G. (1997). Determination of heavy metal complexes with humic substances by HPLC/ICP

- MS coupling using on

- line isotope dilution technique. Fresenius' Journal o f Analytical Chemistry, 359 (4

- 5), 438

- 441. http://dx.doi.org/10.1007/s002160050606

von Schneidemesser, E., Stone, E. A., Quraishi, T. A., Shafer, M. M., & Schauer, J. J. (2010). Toxic metals in the atmosphere in Lahore, Pakistan. Science of the Total Environment, 408 (7), 1640

- 1648. http://dx.doi .org/10.1016/j.scitotenv.2009.12.022

Voutsa, D., & Samara, C. (2002). Labile and bioaccessible fractions of heavy metals in the airborne particulate matter from urban

and industrial areas. Atmospheric Environment, 36 , 3583

- 3590. http://dx.doi.org/10.1016/S1352

- 2310(02)00282

- 0

Waheed, A., Zhang, Y. X., Bao, L. M., Cao, Q. C., Zhang, G. L., Li, Y., & Li, X. L. (2010). Study of seasonal variation and source characteristic of PM10 of Shanghai urban atmosphere using PIXE. Journal of Radioanalytical and Nuclear Ch emistry,

283 (2), 427

- 432. http://dx.doi.org/10.1007/s10967

- 009

- 0390

- 2

Wåhlin, P., Berkowicz, R., & Palmg ren, F. (2006). Characterisation of traffic

- generated particulate matter in Copenhagen. Atmospheric Environment, 40 (12), 2151

- 2159. http://dx.doi.org/10.1016/j.atmosenv.2005.11.049

Walder, A. J., & Freedman, P. A. (1992). Isotopic ratio measurement using a double focusing magnetic sector mass analyser with an inductively coupled plasma as an ion s ource. Journal of Analytical Atomic Spectrometry, 7 , 571

- 575. http://dx.doi.org/10.1039/JA9920700571

Wang, G. , Huang, L., Gao, S., & Wang, L. (2002). Characterization of water

- soluble species of PM10 and PM2.5 aerosols in urban area in Nanjing, China. Atmospheric Environment, 36 (8), 1299

- 1307. http: //dx.doi.org/10.1016/S1352

- 2310(01)00550

- 7

Wang, W., Liu, X. D., Zhao, L. W., Guo, D. F., Tian, X. D., & Adams, F. (2006). Effectiveness of leaded petrol phase

- out in Tianjin, China based on the aerosol lead concentration and isotope abundance ratio. Science of the Total Environment,

364 (1

- 3), 175

- 187. ht tp://dx.doi.org/10.1016/j.scitotenv.2005.07.002

Wang, X. L., Sato, T., & Xing, B. S. (2006). Size distribution and anthropogenic sources apportionment of airbo rne trace metals in Kanazawa, Japan. Chemosphere, 65 (11), 2440

- 2448. http://dx.doi.org/10.1016/j.chemosphere.2006.04.050

Warnken, K. W., & Santschi, P. H. (2009). Delivery of trace metals (Al, Fe, Mn, V, Co, Ni, Cu, Cd, Ag, Pb) from the Trinity River

Watershed towards the ocean. Estuaries and Coasts, 32 (1), 158

- 172. http://dx.doi.org/10.1007/s12237

- 008

- 9088

- x

Watmough, S. A., & Dillon, P. J. (2007). Lead biogeochemistry in a central Ontario fores ted watershed. Biogeochemistry, 84 (2),

143

- 159. http://dx.doi.org/10.1007/s10533

- 007

- 9110

- 6May 20 11 3

- 127 DRAFT

– DO NOT CITE OR QUOTE Watmough, S. A., Hutchinson, T. C., & Dillon, P. J. (2004). Lead dynamics in the forest floor and mineral soil in south

- central

Ontari. Biogeochemistry, 71 (1), 43

- 68. http://dx.doi.org/10.1007/s10533

- 004

- 766 1

- 3

Wedding, J. B., McFarland, A. R., & Cermak, J. E. (1977). Large particle collection characteristics of ambient aerosol samplers.

Environmental Science and T echnology, 11 , 387

- 390. http://dx.doi.org/10.1021/es60127a005

Weiss, A. L., Caravanos, J., Blaise, M. J., & Ja eger, R. J. (2006). Distribution of lead in urban roadway grit and its association

with elevated steel structures. Chemosphere, 65 (10), 1762

- 1771. http://dx.doi.org/10.1016/j.chemosphere.2006.04.079

Weitkamp, E. A., Lipsky, E. M., Pancras, P. J., Ondov, J. M., Polidori, A., Turpin, B. J., & Robinson, A. L. (2005). Fine particle emission profile for a large coke production facility based on highly time

- resolved fence line measurements. Atmospheric
Environment, 39 , 6719

- 6733. http://dx.doi.org/10. 1016/j.atmosenv.2005.06.028

Westerlund, C., & Viklander, M. (2006). Particles and associated metals in road runoff during snowmelt and rainfall. Science of

the Total Environment, 362 (1

- 3), 143

- 156. http://dx.doi.org/10.1016/j.scitotenv.2005.06.031

Widory, D

. (2006). Lead isotopes decipher multiple origins within single PM10 samples in the atmosphere of Paris. Isotopes in Environmental and Health Studies, 42 (1), 97

- 105. http://dx.doi.org/P074 8WJ10.1080/10256010500502736

Williams, A. G. B., Scheckel, K. G., Tolaymat, T., & Impellitteri, C. A. (2006). Mineralogy and characterization of arsenic, iron,

and lead in a mine waste

- derived fertilizer. Environmental Science and Technology, 40 (16), 4874

- 4879. http://dx.doi.org/10.1021/es060853c

Witt, M., Baker, A. R., & Jickells, T. D. (2006). Atmospheric trace metals over the Atlantic and South Indian Oceans: Investigation of metal concentrations and lead isotope ratios in coastal and remote marine aerosols. Atmosph eric Environment, 40 (28), 5435

- 5451. http://dx.doi.org/10.1016/j.atmosenv.2006.04.041

Wojas, B., & Almquist, C. (2007). Mass concentrations and metals speciation of PM2.5, PM10, and total suspended solids in Oxford, Ohio and comparison with those from metropolitan sites in the Greater Cincinnati region. Atmospheric
Environment, 41 (39), 9064

- 9078. http://dx.doi.org/10.1016/j.atmosenv.2007.08.010

Wong, C. S. C., Li, X. D., & Thornton, I. (2006). Urban environmental geochemistry of trace metals. Environmental Pollution,

142 (1), 1

- 16. http://dx.doi.org/10.1016/j.envpol.2005.09.004

Wu, J., Edwards, R., He, X. E., Liu, Z., & Kleinman, M. (2010). Spatial analysis of bioavailable soil lead concentrations in Los Angeles, California. Environmental Research, 110 (4), 309

- 317. http://dx.doi.org/10.1016/j.envres.2010.02.004

Wu, S. C., Luo, Y. M., Cheung, K. C., & Wong, M. H. (2006). Influence of bact eria on Pb and Zn speciation, mobility and bioavailability in soil: A laboratory study. Environmental Pollution, 144 (3), 765

- 773. http://dx.doi.org/10.1016/j.envpol.2006.02.022

Xie, L., & Giammar, D. E. (2007). Equilibrium solubility and dissolution rate of the lead phosphate chloropyromorphite. Environmental Science and Technology, 41 (23), 8050

- 8055. http://dx.doi.org/10.1021/es071517e

Yadav, S., & Rajamani, V. (2006). Air quality and trace metal chemistry o f different size fractions of aerosols in N

- NW India

- implications for source diversity. Atmospheric Environment, 40 (4), 698

- 712. http://dx.doi.org/10.1016/j.atmosenv.2005.10.005

Yi, S.

- M., Totten, L. A., Thota, S., Yan, S., Offenberg, J. H., Eisenreich, S. J., . . . Holsen, T. M. (2006). Atmospheric dry deposition of trace elements measured around the urban and industrially impacted NY

- NJ harbor. Atmospheric
Environment, 40 , 6626

- 6637. http://dx.doi.org/10.1016/j.atmosenv.2006.05.062

Ying, G., & Sansalone, J. (2008). Granulometric relationships for urban source area runoff as a function of hydrologic event

classification and sedimentation. Water, Air , and Soil Pollution, 193 (1

- 4), 229

- 246. http://dx.doi.org/10.1007/s11270

- 008

- 9685

- 0

Zahran, S., Mielke, H. W., Gonzales, C. R., Powell, E. T., & Weiler, S. (2010). New Orleans before and after hurricanes Katrina/Rita: A quasi

- experiment of the association between soil lead and children's blood lead. Environmental Science and Technology, 44 (12), 4433

- 4440

. http://dx.doi.org/10.1021/es100572s

Zak, K., Rohovec, J., & Navratil, T. (2009). Fluxes of heavy metals fro m a highly polluted watershed during flood events: A case study of the Litavka River, Czech Republic. Water, Air, and Soil Pollution, 203 (1

- 4), 343

- 358. http://dx.doi.org/10.1007/s11270

- 009

- 0017

- 9

Zhang, L.

- F., Peng, S.

- Q., & Wang, S. (2009). Decreased aortic contractile reaction to 5

- hydroxytryptamine in rats with long

- term hypertension induced by lead ( Pb2+) exposure. Toxicology Letters, 186 (2), 78

- 83. http://dx.doi.org/10.1016/j.toxlet.2009.01.004May 20 11 3

- 128 DRAFT

– DO NOT CITE OR QUOTE Zhe ng, R. Q., & Li, C. Y. (2009). Effect of lead on survival, locomotion and sperm morphology of Asian earthworm, Pheretima guillelmi. Journal of Environmental Sciences, 21 (5), 691

- 695. http://d x.doi.org/10.1016/s1001

- 0742(08)62325

- 6May 20 11 3

- 129 DRAFT

– DO NOT CITE OR QUOTE 

# Chapter 3 Appendix

3.8. Variability across the U.S.

Table 3A

- 1

. Distribution of 1

- month average Pb

- TSP concentrations ( μ g/m 3 

### ) nationwide,

source
- oriented monitors, 2007

- 9. Sites listed in the bottom six rows of the table fall in the upper 90th percentile of the data pooled by site.

Year Season State/ County State County

name Site ID N monthly

means N sites Mean Min 1 5 10 25 50 75 90 95 99 max

Nationwide statistics

2007

- 2009 1,802 0.2353 0 0 0.007 0.012 0.028 0.070 0.254 0.670 0.976 2.061 4.440

2007 554 0.250 0 0 0 0.008 0.012 0.028 0.070 0.290 0.706 1.085 1.982 3.620

2008 622 0.2803 0 0 0.008 0.013 0.034 0.080 0.300 0.741 1.164 2.501 4.440 2009 626 0.1774 0 0 0.006 0.010 0.023 0.063 0.194 0.535 0.787 1.280 2.438 Winter 443 0.2324 0 0 0.006 0.010 0.025 0.064 0.224 0.575 1.111 2.440 3.103 Spring 447 0.2695 0 0 0.008 0.013 0.031 0.077 0.333 0.794 1.085 2.035 3.123 Summer 458 0.2142 0 0 0.008 0.013 0.032 0.074 0.265 0.652 0.852 1.384 4.440 Fall 454 0.2260 0 0 0.007 0.012 0.026 0.064 0.226 0.670 1.006 1.565 4.225 Nationwide statistics, pooled by site

2007

- 2009 56 0.2384 0.001 0.001 0.010 0.014 0.031 0.098 0.350 0.730 0.862 1.306 1.306

2007 48 0.2547 0.000 0.000 0.010 0.014 0.031 0.090 0.323 0.783 1.048 1.520 1.520 2008 53 0.2848 0.001 0.001 0.011 0.021 0.032 0.121 0.344 0.818 1.226 1.542 1.542 2009 53 0.1779 0.000 0.000 0.007 0.011 0.029 0.075 0.219 0.453 0.718 0.917 0.917 Winter 56 0.2353 0.001 0.001 0.008 0.010 0.033 0.088 0.290 0.654 1.205 1.389 1.389 Spring 56 0.2759 0.000 0.000 0.008 0.017 0.035 0.093 0.425 0.863 1.119 1.605 1.605 Summer 56 0.2136 0.000 0.000 0.008 0.016 0.034 0.095 0.289 0.654 0.802 1.170 1.170 Fall 55 0.2286 0.001 0.001 0.012 0.017 0.029 0.071 0.389 0.651 1.031 1.243 1.243 Statistics for individual counties (2007

- 2009)

01109 AL Pike 44 2 0.5110 0.070 0.070 0.083 0.094 0.190 0.355 0.710 1.277 1.356 1.434 1.434 06037 CA Los Angeles 108 5 0.2425 0.002 0.002 0.008 0.012 0.037 0.073 0.188 0.698 0.942 2.501 2.880 12057 FL Hillsborough 62 2 0.2939 0.014 0.014 0.020 0.040 0.080 0.139 0.333 0.600 0.840 3.620 3.620 17031 IL Cook 144 4 0.0241 0.010 0.010 0.010 0.010 0.014 0.022 0.030 0.040 0.046 0.070 0.084 17119 IL Madison 36 1 0.1088 0.022 0.022 0.024 0.028 0.037 0.058 0.138 0.300 0.325 0.454 0.454 17163 IL Saint Clair 36 1 0.0246 0.010 0.010 0.010 0.013 0.015 0.023 0.030 0.045 0.054 0.058 0.058 18035 IN Delaware 72 2 0.3318 0.034 0.034 0.045 0.056 0.094 0.202 0.341 0.639 0.980 4.440 4.440 18089 IN Lake 36 1 0.0297 0.004 0.004 0.006 0.007 0.017 0.025 0.045 0.060 0.063 0.065 0.065 18097 IN Marion 71 2 0.0198 0.003 0.003 0.005 0.007 0.010 0.016 0.028 0.040 0.046 0.057 0.057 27037 MN Dakota 35 1 0.2872 0.062 0.062 0.072 0.098 0.134 0.232 0.412 0.620 0.670 0.730 0.730 27163 MN Washington 31 1 0.0006 0.000 0.000 0.000 0.000 0.000 0.000 0.001 0.003 0.003 0.004 0.004 29093 MO Iron 144 4 0.4060 0.007 0.010 0.021 0.029 0.045 0.094 0.658 0.971 1.437 2.557 4.225 29099 MO Jefferson 296 10 0.6096 0.011 0.019 0.071 0.124 0.212 0.454 0.798 1.353 1.764 2.440 3.123 29189 MO Saint Louis 36 1 0.0361 0.005 0.005 0.005 0.005 0.008 0.050 0.050 0.050 0.050 0.066 0.066 34023 NJ Middlesex 12 1 0.0101 0.007 0.007 0.007 0.007 0.008 0.008 0.008 0.010 0.034 0.034 0.034 36071 NY Orange 106 3 0.0271 0.003 0.003 0.004 0.005 0.007 0.020 0.037 0.058 0.079 0.103 0.134 39035 OH Cuyahoga 108 3 0.0473 0.004 0.004 0.007 0.008 0.013 0.024 0.060 0.130 0.190 0.210 0.220 39051 OH Fulton 33 1 0.2434 0.015 0.015 0.055 0.066 0.093 0.190 0.360 0.510 0.620 0.690 0.690 39091 OH Logan 68 2 0.0763 0.020 0.020 0.020 0.030 0.040 0.070 0.090 0.120 0.190 0.290 0.290 42007 PA Beaver 32 1 0.1090 0.045 0.045 0.054 0.063 0.070 0.100 0.132 0.173 0.195 0.284 0.284 42011 PA Berks 108 3 0.1053 0.030 0.033 0.037 0.040 0.050 0.073 0.131 0.242 0.302 0.360 0.542 47163 TN Sullivan 108 3 0.0754 0.030 0.030 0.032 0.036 0.042 0.062 0.087 0.145 0.178 0.254 0.341 48085 TX Collin 108 3 0.2910 0.007 0.009 0.030 0.052 0.102 0.186 0.389 0.673 0.904 1.121 1.564 Statistics for individual sites where overall average monthly mean ≥ national 90th percentile (2007

- 2009)

060371405 24 0.8623 0.242 0.242 0.283 0.284 0.321 0.606 0.923 2.277 2.501 2.880 2.880 290930016 36 0.7302 0.166 0.166 0.186 0.235 0.347 0.545 0.837 1.295 2.435 4.225 4.225 290930021 36 0.7970 0.084 0.084 0.093 0.099 0.409 0.696 0.967 1.453 2.438 2.557 2.557 290990004 36 1.1177 0.242 0.242 0.285 0.518 0.764 1.005 1.519 1.905 2.101 2.416 2.416 290990015 34 1.3060 0.155 0.155 0.339 0.421 0.834 1.185 1.577 2.319 3.103 3.123 3.123 290990021 33 0.7498 0.084 0.084 0.141 0.423 0.550 0.681 0.901 1.164 1.553 2.162 2.162May 20 11 3

- 130 DRAFT

– DO NOT CITE OR QUOTE 

### Table 3A

- 2

. Distribution of 3

- month moving average Pb

- TSP concentrations (μg/m3) nationwide, s ource
- oriented monitors, 2007

- 9. Sites listed in the bottom six rows of the table fall in the upper 90th percentile of the data pooled by site

.

Year Season State/ County State County

name Site ID N monthly

means N sites Mean Min 1 5 10 25 50 75 90 95 99 max

Nationwide statistics

2007 9 1,737 0.2366 0.000 0.000 0.010 0.010 0.030 0.080 0.270 0.680 0.980 1.870 2.890 2007 507 0.2421 0.000 0.000 0.010 0.020 0.030 0.080 0.270 0.720 1.070 1.640 1.740 2008 612 0.2834 0.000 0.000 0.010 0.010 0.040 0.090 0.320 0.800 1.150 2.290 2.890 2009 618 0.1859 0.000 0.000 0.010 0.010 0.030 0.070 0.240 0.550 0.760 1.150 2.070 Winter 417 0.2434 0.000 0.000 0.010 0.010 0.030 0.070 0.270 0.660 1.120 2.080 2.620 Spring 436 0.2638 0.000 0.000 0.010 0.010 0.030 0.080 0.330 0.740 1.080 2.140 2.890 Summer 448 0.2269 0.000 0.000 0.010 0.020 0.030 0.080 0.290 0.670 0.900 1.730 2.140 Fall 436 0.2131 0.000 0.000 0.010 0.020 0.030 0.080 0.240 0.650 0.940 1.320 2.460 Nationwide statistics, pooled by site

2007 9 56 0.2386 0.000 0.000 0.011 0.017 0.032 0.099 0.333 0.733 0.930 1.332 1.332 2007 45 0.2435 0.003 0.003 0.016 0.019 0.033 0.084 0.277 0.687 0.831 1.524 1.524 2008 53 0.2903 0.000 0.000 0.010 0.022 0.034 0.119 0.383 0.806 1.301 1.641 1.641 2009 53 0.1848 0.000 0.000 0.009 0.014 0.029 0.075 0.229 0.617 0.855 0.863 0.863 Winter 56 0.2482 0.000 0.000 0.010 0.015 0.041 0.104 0.316 0.683 1.150 1.308 1.308 Spring 56 0.2736 0.000 0.000 0.010 0.014 0.032 0.096 0.369 0.820 1.320 1.679 1.679 Summer 56 0.2243 0.000 0.000 0.010 0.012 0.034 0.093 0.287 0.707 0.900 1.232 1.232 Fall 55 0.2132 0.000 0.000 0.011 0.017 0.032 0.085 0.341 0.592 0.960 1.164 1.164 Statistics for individual counties (2007

- 2009)

01109 AL Pike 44 2 0.5080 0.100 0.100 0.120 0.230 0.275 0.325 0.685 1.050 1.150 1.350 1.350 06037 CA Los Angeles 108 5 0.2594 0.010 0.010 0.010 0.010 0.045 0.075 0.185 0.710 1.260 2.450 2.490 12057 FL Hillsborough 48 2 0.2248 0.040 0.040 0.040 0.040 0.095 0.135 0.250 0.490 0.580 1.770 1.770 17031 IL Cook 144 4 0.0244 0.010 0.010 0.010 0.010 0.020 0.030 0.030 0.040 0.040 0.050 0.050 17119 IL Madison 36 1 0.1144 0.040 0.040 0.040 0.040 0.060 0.105 0.155 0.210 0.230 0.280 0.280 17163 IL Saint Clair 36 1 0.0250 0.010 0.010 0.010 0.020 0.020 0.020 0.030 0.040 0.040 0.050 0.050 18035 IN Delaware 72 2 0.3422 0.050 0.050 0.070 0.080 0.120 0.210 0.430 0.710 0.930 2.140 2.140 18089 IN Lake 36 1 0.0314 0.010 0.010 0.010 0.010 0.020 0.030 0.040 0.050 0.050 0.060 0.060 18097 IN Marion 69 2 0.0199 0.000 0.000 0.010 0.010 0.010 0.020 0.030 0.030 0.040 0.050 0.050 27037 MN Dakota 33 1 0.2742 0.100 0.100 0.120 0.140 0.210 0.240 0.350 0.400 0.470 0.570 0.570 27163 MN Washington 24 1 0.0000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 29093 MO Iron 144 4 0.4142 0.020 0.020 0.020 0.030 0.050 0.170 0.735 1.010 1.210 2.080 2.460 29099 MO Jefferson 280 10 0.6211 0.040 0.070 0.120 0.155 0.280 0.475 0.835 1.250 1.655 2.620 2.890 29189 MO Saint Louis 36 1 0.0383 0.010 0.010 0.010 0.010 0.010 0.050 0.050 0.050 0.050 0.060 0.060 34023 NJ Middlesex 12 1 0.0192 0.010 0.010 0.010 0.010 0.010 0.010 0.020 0.050 0.060 0.060 0.060 36071 NY Orange 102 3 0.0273 0.000 0.000 0.000 0.010 0.010 0.030 0.030 0.060 0.070 0.080 0.090 39035 OH Cuyahoga 102 3 0.0483 0.010 0.010 0.010 0.010 0.020 0.030 0.070 0.120 0.140 0.160 0.170 39051 OH Fulton 25 1 0.2420 0.050 0.050 0.050 0.080 0.130 0.240 0.320 0.450 0.520 0.570 0.570 39091 OH Logan 68 2 0.0809 0.030 0.030 0.030 0.030 0.050 0.070 0.090 0.160 0.190 0.260 0.260 42007 PA Beaver 32 1 0.1115 0.060 0.060 0.070 0.080 0.090 0.110 0.120 0.160 0.170 0.170 0.170 42011 PA Berks 108 3 0.1099 0.040 0.040 0.040 0.050 0.060 0.085 0.150 0.210 0.250 0.350 0.360 47163 TN Sullivan 102 3 0.0769 0.030 0.030 0.040 0.040 0.050 0.060 0.090 0.130 0.180 0.210 0.260 48085 TX Collin 108 3 0.2873 0.010 0.030 0.060 0.070 0.110 0.205 0.430 0.590 0.650 1.190 1.260 Statistics for individual sites where overall average monthly mean ≥ national 90th percentile (2007

- 2009)

060371405 24 0.9304 0.280 0.280 0.310 0.330 0.470 0.665 1.045 2.290 2.450 2.490 2.490 290930016 36 0.7328 0.210 0.210 0.260 0.380 0.445 0.570 0.770 1.170 2.080 2.460 2.460 290930021 36 0.8258 0.220 0.220 0.220 0.300 0.545 0.880 1.005 1.210 1.270 1.940 1.940 290990004 36 1.1383 0.630 0.630 0.680 0.700 0.810 1.085 1.340 1.710 2.130 2.140 2.140 290990015 34 1.3318 0.610 0.610 0.640 0.730 0.940 1.195 1.710 2.020 2.630 2.890 2.890 290990021 33 0.7682 0.430 0.430 0.440 0.520 0.610 0.750 0.900 1.090 1.310 1.320 1.320May 20 11 3

- 131 DRAFT

– DO NOT CITE OR QUOTE 

### Table 3 A

- 3

. Distribution of annual 1

- month site maxima TSP Pb concentrations ( μ g/m 3 

### ) nationwide, source

- oriented monitors, 2007

- 2009. Sites listed in the bottom eight rows of the table fall in the upper 9 0th percentile of the data pooled by site

.

Year Site ID

– year N (sites) Mean Min 1 5 10 25 50 75 90 95 99 max

Nationwide statistics

2007

- 2009 56 0.7961 0.004 0.004 0.025 0.040 0.076 0.289 0.747 2.557 3.620 4.440 4.440 2007 48 0.5517 0.003 0.003 0.025 0.035 0.054 0.192 0.735 1.565 2.162 3.620 3.620 2008 53 0.7157 0.004 0.004 0.014 0.039 0.059 0.247 0.754 2.416 3.123 4.440 4.440 2009 53 0.3758 0.003 0.003 0.016 0.019 0.065 0.141 0.536 1.124 1.357 2.438 2.438

Annual site max 3

- month means >= national 90th percentile (2007

- 2009)

180350009

- 2008 4.4400 290930016

- 2008 4.2252 120571066

- 2007 3.6200 290990015

- 2008 3.1228 060371405

- 2008 2.8800 290930021

- 2008 2.5566 180350009

- 2008 4.4400 290930016

- 2008 4.2252



### Table 3 A

- 4

. Distribution of annual 3

- month site maxima Pb

- TSP concentrations ( μ g/m 3 

### ) nationwide, source

- oriented monitors, 2007

- 2009. Sites listed in the bottom nine rows of the table fall in the upper 90th percen tile of the data pooled by site.

Year Site ID

– year N (sites) Mean Min 1 5 10 25 50 75 90 95 99 max

Nationwide statistics

2007

- 2009 56 0.5409 0.000 0.000 0.020 0.030 0.060 0.215 0.590 1.940 2.460 2.890 2.890

2007 45 0.3616 0.010 0.010 0.030 0.030 0.050 0.130 0.520 1.210 1.350 1.740 1.740

2008 53 0.5177 0.000 0.000 0.010 0.030 0.050 0.170 0.530 1.770 2.460 2.890 2.890

2009 53 0.3123 0.000 0.000 0.010 0.020 0.040 0.110 0.390 0.940 1.240 2.070 2.070

Annual site max 3

- month means >= national 90th percentile (2007

- 2009)

060371405

- 2008 2.4900

180350009

- 2008 2.1400

290930016

- 2008 2.4600

290930016

- 2009 2.0700

290930021

- 2009 1.9400

290990004

- 2008 2.1400

290990015

- 2008 2.8900

060371405

- 2008 2.4900May 20 11 3

- 132 DRAFT

– DO NOT CITE OR QUOTE 

### Table 3 A

- 5

. One

- month average Pb

- TSP for individual county concentrations nationwide ( μg/m 3 

### ), non

- source
- oriented monitors, 2007

- 2009

Stcou State County name N monthly

means N sites Mean Min 1 5 10 25 50 75 90 95 99 max

Statistics for individual counties (2007

- 2009)

06025 CA Imperial 33 1 0.0218 0.009 0.009 0.010 0.011 0.013 0.019 0.029 0.036 0.041 0.041 0.041

06037 CA Los Angeles 117 5 0.0107 0.000 0.000 0.000 0.002 0.006 0.010 0.015 0.020 0.024 0.032 0.038

06065 CA Riverside 48 2 0.0090 0.000 0.000 0.002 0.003 0.008 0.010 0.010 0.016 0.018 0.022 0.022

06071 CA San Bernardino 47 2 0.0112 0.000 0.000 0.002 0.003 0.008 0.010 0.014 0.020 0.022 0.036 0.036

12103 FL Pinellas 12 1 0.0000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000

17031 IL Cook 143 4 0.0144 0.010 0.010 0.010 0.010 0.010 0.014 0.017 0.020 0.024 0.028 0.032

17117 IL Macoupin 36 1 0.0102 0.010 0.010 0.010 0.010 0.010 0.010 0.010 0.010 0.013 0.014 0.014

17119 IL Madison 35 1 0.0187 0.010 0.010 0.010 0.010 0.010 0.015 0.022 0.030 0.038 0.066 0.066

17143 IL Peoria 36 1 0.0113 0.010 0.010 0.010 0.010 0.010 0.010 0.012 0.014 0.015 0.018 0.018

18097 IN Marion 35 1 0.0071 0.002 0.002 0.002 0.003 0.004 0.006 0.009 0.012 0.012 0.033 0.033

18163 IN Vanderburgh 21 2 0.0048 0.001 0.001 0.002 0.002 0.004 0.005 0.005 0.006 0.010 0.010 0.010

25025 MA Suffolk 24 1 0.0089 0.000 0.000 0.000 0.003 0.007 0.009 0.011 0.015 0.016 0.020 0.020

26163 MI Wayne 24 1 0.0187 0.007 0.007 0.009 0.012 0.013 0.018 0.022 0.027 0.032 0.045 0.045

27037 MN Dakota 129 4 0.0039 0.000 0.000 0.000 0.000 0.000 0.003 0.006 0.008 0.010 0.018 0.036

27053 MN Hennepin 128 4 0.0034 0.000 0.000 0.000 0.000 0.002 0.003 0.005 0.008 0.008 0.010 0.010

27123 MN Ramsey 71 3 0.0071 0.000 0.000 0.000 0.000 0.004 0.006 0.010 0.015 0.020 0.028 0.028

27137 MN Saint Louis 70 2 0.0017 0.000 0.000 0.000 0.000 0.000 0.002 0.003 0.005 0.006 0.010 0.010

27163 MN Washington 36 1 0.0035 0.000 0.000 0.000 0.000 0.003 0.004 0.004 0.006 0.006 0.008 0.008

36047 NY Kings 36 1 0.0142 0.010 0.010 0.010 0.010 0.011 0.013 0.016 0.020 0.021 0.026 0.026

39017 OH Butler 34 1 0.0052 0.000 0.000 0.000 0.003 0.004 0.005 0.007 0.008 0.009 0.009 0.009

39029 OH Columbiana 107 3 0.0145 0.000 0.000 0.004 0.006 0.008 0.012 0.019 0.031 0.034 0.045 0.050

39035 OH Cuyahoga 70 2 0.0138 0.004 0.004 0.006 0.006 0.009 0.013 0.016 0.022 0.028 0.041 0.041

39049 OH Franklin 36 1 0.0088 0.004 0.004 0.004 0.005 0.007 0.008 0.011 0.014 0.014 0.016 0.016

39143 OH Sandusky 12 1 0.0048 0.003 0.003 0.003 0.003 0.004 0.005 0.006 0.006 0.007 0.007 0.007

39167 OH Washington 43 2 0.0041 0.002 0.002 0.002 0.002 0.003 0.004 0.005 0.006 0.007 0.010 0.010

42003 PA Allegheny 36 1 0.0067 0.000 0.000 0.000 0.000 0.000 0.004 0.012 0.016 0.021 0.053 0.053

42045 PA Delaware 20 1 0.0400 0.040 0.040 0.040 0.040 0.040 0.040 0.040 0.040 0.040 0.040 0.040

42129 PA Westmoreland 36 1 0.0410 0.034 0.034 0.037 0.040 0.040 0.040 0.040 0.044 0.050 0.053 0.053

48061 TX Cameron 35 1 0.0038 0.002 0.002 0.002 0.002 0.003 0.003 0.005 0.006 0.007 0.009 0.009

48141 TX El Paso 22 2 0.0229 0.014 0.014 0.015 0.016 0.018 0.019 0.019 0.041 0.056 0.057 0.057

48201 TX Harris 31 1 0.0054 0.003 0.003 0.004 0.004 0.005 0.005 0.006 0.007 0.009 0.010 0.010

48479 TX Webb 31 1 0.0103 0.004 0.004 0.005 0.006 0.007 0.008 0.010 0.024 0.025 0.026 0.026May 20 11 3

- 133 DRAFT

– DO NOT CITE OR QUOTE 

### Table 3A

- 6

. Three

- month moving average Pb

- TSP for individual county concentrations ( μg/m 3 

### ) nationwide, non

- source
- oriented monitors, 2007

- 2009

Stcou State County name N monthly

means N sites Mean Min 1 5 10 25 50 75 90 95 99 max

Statistics for individual counties (2007

- 2009)

06025 CA Imperial 33 1 0.0216 0.010 0.010 0.010 0.020 0.020 0.020 0.030 0.030 0.030 0.030 0.030

06037 CA Los Angeles 117 5 0.0113 0.000 0.000 0.000 0.000 0.010 0.010 0.010 0.020 0.020 0.030 0.030

06065 CA Riverside 48 2 0.0102 0.010 0.010 0.010 0.010 0.010 0.010 0.010 0.010 0.010 0.020 0.020

06071 CA San Bernardino 47 2 0.0115 0.010 0.010 0.010 0.010 0.010 0.010 0.010 0.020 0.020 0.020 0.020

12103 FL Pinellas 12 1 0.0000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000

17031 IL Cook 143 4 0.0139 0.010 0.010 0.010 0.010 0.010 0.010 0.020 0.020 0.020 0.020 0.020

17117 IL Macoupin 36 1 0.0100 0.010 0.010 0.010 0.010 0.010 0.010 0.010 0.010 0.010 0.010 0.010

17119 IL Madison 35 1 0.0182 0.010 0.010 0.010 0.010 0.010 0.020 0.020 0.030 0.040 0.040 0.040

17143 IL Peoria 36 1 0.0100 0.010 0.010 0.010 0.010 0.010 0.010 0.010 0.010 0.010 0.010 0.010

18097 IN Marion 35 1 0.0076 0.000 0.000 0.000 0.000 0.000 0.010 0.010 0.010 0.020 0.020 0.020

18163 IN Vanderburgh 21 2 0.0037 0.000 0.000 0.000 0.000 0.000 0.000 0.010 0.010 0.010 0.010 0.010

25025 MA Suffolk 24 1 0.0092 0.000 0.000 0.000 0.000 0.010 0.010 0.010 0.010 0.020 0.020 0.020

26163 MI Wayne 24 1 0.0196 0.010 0.010 0.010 0.010 0.020 0.020 0.020 0.020 0.030 0.030 0.030

27037 MN Dakota 129 4 0.0029 0.000 0.000 0.000 0.000 0.000 0.000 0.010 0.010 0.010 0.010 0.010

27053 MN Hennepin 128 4 0.0024 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.010 0.010 0.010 0.010

27123 MN Ramsey 71 3 0.0071 0.000 0.000 0.000 0.000 0.000 0.010 0.010 0.010 0.020 0.020 0.020

27137 MN Saint Louis 70 2 0.0000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000

27163 MN Washington 36 1 0.0011 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.010 0.010 0.010 0.010

36047 NY Kings 36 1 0.0128 0.010 0.010 0.010 0.010 0.010 0.010 0.020 0.020 0.020 0.020 0.020

39017 OH Butler 34 1 0.0061 0.000 0.000 0.000 0.000 0.000 0.010 0.010 0.010 0.010 0.010 0.010

39029 OH Columbiana 107 3 0.0151 0.010 0.010 0.010 0.010 0.010 0.010 0.020 0.020 0.030 0.030 0.030

39035 OH Cuyahoga 70 2 0.0142 0.000 0.000 0.010 0.010 0.010 0.010 0.020 0.020 0.020 0.030 0.030

39049 OH Franklin 36 1 0.0094 0.000 0.000 0.000 0.010 0.010 0.010 0.010 0.010 0.010 0.010 0.010

39143 OH Sandusky 12 1 0.0060 0.000 0.000 0.000 0.000 0.000 0.010 0.010 0.010 0.010 0.010 0.010

39167 OH Washington 43 2 0.0008 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.010 0.010 0.010

42003 PA Allegheny 36 1 0.0058 0.000 0.000 0.000 0.000 0.000 0.000 0.010 0.010 0.020 0.030 0.030

42045 PA Delaware 20 1 0.0400 0.040 0.040 0.040 0.040 0.040 0.040 0.040 0.040 0.040 0.040 0.040

42129 PA Westmoreland 36 1 0.0414 0.040 0.040 0.040 0.040 0.040 0.040 0.040 0.050 0.050 0.050 0.050

48061 TX Cameron 35 1 0.0024 0.000 0.000 0.000 0.000 0.000 0.000 0.000 0.010 0.010 0.010 0.010

48141 TX El Paso 22 2 0.0256 0.020 0.020 0.020 0.020 0.020 0.020 0.030 0.040 0.040 0.040 0.040

48201 TX Harris 31 1 0.0063 0.000 0.000 0.000 0.000 0.000 0.010 0.010 0.010 0.010 0.010 0.010

48479 TX Webb 31 1 0.0100 0.010 0.010 0.010 0.010 0.010 0.010 0.010 0.010 0.010 0.010 0.010



# 3.8.1. Intra

- urban Variability

Maps of six areas (Los Angeles County, CA; Hillsborough/Pinellas Counties, FL; Cook County, IL; 

## 1

Jefferson County, MO; Cuyahoga County, OH; and Sullivan County, TN) are shown to illustrate the 

## 2

location of all Pb monitors meeting the inclusion criteria. Wind roses for e ach season are also provided to 

## 3

help put the source concentration data in context. Letters on the maps identify the individual monitor 

## 4

locations and correspond with the letters provided in the accompanying concentration box plots and pair



## 5

wise monitor comp arison tables. The box plots for each monitor include the annual and seasonal 

## 6

concentration median and interquartile range with whiskers extending from the 5th to the 95th percentile. 

## 7May 20 11 3

- 134 DRAFT

– DO NOT CITE OR QUOTE 

### Data from 2007

- 2009 were used to generate the box plots, which are strat ified by season as follows: 

## 1

1 = winter (December

- February), 2 = spring (March

- May), 3 = summer (June

- August), and 4 = fall 

## 2

(September

- November). The comparison tables include the Pearson correlation coefficient (r), the 90th 

## 3

percentile of the absolute diff erence in concentrations (P90) in ppm, the coefficient of divergence (COD) 

## 4

and the straight

- line distance between monitor pairs (d) in km. The COD provides an indication of the 

## 5

variability across the monitoring sites within each county and is defined as fo llows: 

## 6

Equation 3 A

- 1



### where 

## X ij 

### and 

## X ik 

### represent the observed hourly concentrations for time period 

## i at sites 

## j and 

## k , and 

## p is 

## 7

the number of paired hourly observations. A 

## COD of 0 indicates there are no differences between 

## 8

concentrations at paired sites (spatial homogeneity), while a 

## COD approaching 1 indicates extreme 

## 9

spatial heterogeneity. 

## 10

In certain cases, the information contained in these figures and tables should be use d with some 

## 11

caution since many of the reported concentrations for the years 2007

- 2009 are near or below the analysis 

## 12

method ’ s stated method detection limit (MDL). The MDL is generally taken as 0.01 because it is the 

## 13

upper value of the range of MDLs reporte d for AA and Emissions Spectra ICAP methods, which were the 

## 14

two methods reported in the AQS to have been used for analysis of FRM samples ( Rice, 2007 )

. Generally, 

## 15

data are reported to the hundredth place, so this assumption is reasonable. The approximate percentage of 

## 16

data below the MDL (to the nearest 5%) is provided for each site along with box plots of seasonal Pb 

## 17

concentration at monitors within each urban area studied. 

## 18

Figure 3 A

- 1 illustrates Pb monitor locations within Los Angeles County, CA. Ten monitors are 

## 19

located within Los Angeles County, five of which were source

- oriented and the other five were non



## 20

source

- oriented monitors. Monitor A was located immediately downwind of the Quemetco battery 

## 21

recycling facility in th e City of Industry, CA. This source was estimated to produce 0.32 tons of Pb/yr 

## 22

( U.S. EPA, 2008b ) . Monitor C was sited in a street canyon just upwind of th e Exide Pb recycling facility, 

## 23

which was estimated to produce 2.0 tons of Pb/yr ( U.S. EPA, 2008b )

. Monitor D was situated slightly 

## 24

northwest of the same Pb recycling facility. It is still in relatively close proximity but not downwind on 

## 25

most occasions. Monitor B was located 12 km downwind of the Exide facility. Monitor E was located 

## 26

nearby the Trojan Battery recycling facility, which emitted 0.79 tons Pb/yr ( U.S. EPA, 2008b )

. Location 

## 27

of the non

- source

- oriented mo nitors varied. Monitor F was positioned on a roof top 60 meters away from 

## 28

a 4

- lane arterial road and 100 m from of a railroad. Monitor G was located on a rooftop approximately 20 

## 29

m from an 8

- lane arterial road, and monitor H was positioned at the curbside of a four

- lane road roughly 

## 30

650 m north of that road ’ s junction with I

- 405. Monitor I was sited in a parking lot roughly 80 m from a 

## 31 2

1 

## 1 

# ∑

= 

##        

+ −

= p

i ik ij ik ij jk 

## X X X X

p CODMay 20 11 3

- 135 DRAFT

– DO NOT CITE OR QUOTE 

### four

- lane road, and monitor J was located approximately 130 m south of a 4

- lane highway. Figure 3A

- 2 

## 1

displays seasonal wind roses for Los Angeles County. In spring, summer, and fall, the predominant winds 

## 2

come from the west

- southwest. During winter, wind direction varies with a portion from the west



## 3

southwest and the remainder from the east. The highest winds during winter com e more frequently from 

## 4

the west

- southwest. 

## 5

The maps shown in Figure 3A

- 1 for source

- oriented monitors A

- E illustrate the different conditions 

## 6

captured by the monitors; this informs analysis of the seasonal and year

- round concentrations reported in 

## 7

Figure 3A

- 3. The average annual concentration at monitor A was 0.074 μg/m 3 

### . The 95th percentile 

## 8

exceeded the level of the NAAQS in the spring (0.16 μg/m 3 

### ) and summer (0.18 μg/m 3 

### ). Monitor C 

## 9

reported the highest concentrations in Los Angeles County, with a year

- ro und mean of 0.68 μg/m 3



### . Given 

## 10

the position of this monitor with respect to the Exide facility, there is the potential for recirculation of 

## 11

fugitive Pb emissions in the air sampled by that monitor. The average annual Pb concentration at monitor 

## 12

D was 0.12 μ g/m 3 

### , and the 75th percentile of year

- round data exceeded the level of the NAAQS; in 

## 13

spring, the 70th percentile exceeded 0.15 μg/m 3



### . Monitor B reported the lowest values among the source



## 14

oriented monitors with an average annual concentration of 0.013 μg/m 3



### . Note that 75% of reported values 

## 15

were below the MDL for this site, and no data from this site exceeded the level of the NAAQS. The 

## 16

annual average concentration at monitor E was 0.068 μg/m 3 

### , and the 95th percentile of concentration was 

## 17

0.17 μg/m 3



### . 

## 18

The non

- source

- oriented monitors located at sites F

- J all recorded low concentrations, with average 

## 19

values ranging from 0.004 to 0.018 μg/m 3 

### (Figure 3A

- 3). The highest average year

- round concentrations 

## 20

were recorded at site F. The 95th percentiles at these sit es ranged from 0.01 to 0.04 μg/m 3



### . There is much 

## 21

less certainty in the data recorded at the non

- source

- oriented sites, because 45

- 95% of the data from these 

## 22

monitors were below the MDL. Additionally, only one of the non

- source

- oriented monitors (monitor H) 

## 23

was positioned at roadside, and none of the non

- source

- oriented monitors were located at the side of a 

## 24

major highway. 

## 25

Intersampler correlations (Table 3A

- 7), illustrate that Pb has high intra

- urban spatial variability. For 

## 26

the source

- oriented monitors, th e highest correlation ρ = 0.57, occurred for monitors C and D, which 

## 27

covered the same site. Because monitor D was slightly farther from the Exide source and slightly 

## 28

upstream of the predominant wind direction, the signal it received from the source site wa s 

## 29

correspondingly lower. Hence, the correlation between these sites was moderate despite their relatively 

## 30

close proximity. In general, low or even negative correlations were observed between the source

- oriented 

## 31

and non

- source

- oriented monitors. The excepti on to this was the correlation between source

- oriented 

## 32

monitor B and non

- source

- oriented monitor F, with ρ = 0.74. Monitors B and F are roughly 16 km apart, 

## 33

whereas monitor B is only 12 km from monitors D and C, 8 km from monitor E, and 6 km from monitor 

## 34

A . It is possible that monitors B and F both captured a source that was either longer in range or more 

## 35

ubiquitous and so would have been obscured by the stronger source signals at sites A, C, D, and E. 

## 36May 20 11 3

- 136 DRAFT

– DO NOT CITE OR QUOTE 

### Comparisons between the non

- source

- oriented monitors re vealed moderate correlation between sites (G 

## 1

to J [ρ = 0.37 to 0.65]). Sites G, H, I and J are all located in the southwestern quadrant of Los Angeles. It 

## 2

is possible that they are also exposed to a ubiquitous source that produces a common signal at these four 

## 3

sites. 

## 4May 20 11 3

- 137 DRAFT

– DO NOT CITE OR QUOTE 

### Figure 3A

- 1. Pb TSP monitor and source locations within Los Angeles County, CA (06

- 037), 2007

- 2009. Note that monitor locations are denoted by green markers, and source locations are denoted by red markers. Top: view of all Pb FRM monitors in Los Angeles County. Bottom left: Close up of the industrial
site near monitors C and D. Bottom right: Close up of the populated area captured by monitor F.May 20 11 3

- 138 DRAFT

– DO NOT CITE OR QUOTE Source: NRCS ( 2011 )

. 

### Figure 3A

- 2

. Wind roses for Los Angeles County, CA, from meteorological data at the Los Angeles Int ernational Airport, 1961

- 1990. Clockwise from top left: January, A pril, July, and October. Note that the wind percentages vary from month to month.May 20 11 3

- 139 DRAFT

– DO NOT CITE OR QUOTE Site A B C D E F G H I J

SITE ID 06

- 037

- 1404 06

- 037

- 1602 06

- 037

- 1405 06

- 037

- 1406 06

- 037

- 1403 06

- 037

- 1103 06

- 037

- 1301 06

- 037

- 4002 06

- 037

- 4004 06

- 037

- 5005

MEAN 0.074 0.013 0.68 0.12 0.068 0.018 0.015 0.0083 0.0087 0.0040

SD 0.040 0.017 1.0 0.092 0.052 0.011 0.012 0.0068 0.0069 0.0064

OBS 66 112 617 242 128 121 108 120 117 109

% BELOW

MDL 0 75 0 0 0 45 65 85 85 95

Source
orientation Source Source Source Source Source Non

source Non

source Non

source Non

source Non

source



### Figure 3A

- 3

. Box plots of annual and seasonal Pb TSP concentrations ( μg/m 3 

### ) from source

- oriented and non

- source
- oriented monitors within Los Angeles County, CA (06

- 037), 2007

- 2009.May 20 11 3

- 140 DRAFT

– DO NOT CITE OR QUOTE 

### Table 3A

- 7

. Correlations between Pb TSP concentrations from source

- oriented and non

source
- oriented monitors within Los Angeles County, CA (06

- 037), 2007

- 2009

A B C D E F G H I J

Source Source Source Source Source Non

- Source Non

- Source Non

- Source Non

- Source Non

- Source

A Source 1.00 0.16 0.10 0.08 0.27

- 0.15 0.00 0.14

- 0.02

- 0.09

0.00 0.08 0.49 0.10 0.10 0.08 0.06 0.08 0.08 0.08

0.00 0.63 0.64 0.31 0.34 0.57 0.57 0.79 0.77 0.85

B Source 1.00 0.05 0.05 0.07 0.74 0.12 0.28 0.11 0.10

0.00 3.59 0.25 0.10 0.02 0.02 0.01 0.02 0.02

0.00 0.96 0.84 0.71 0.46 0.48 0.61 0.60 0.81

C Source 1.00 0.57 0.03

- 0.08 0.26 0.28 0.20 0.13

0.00 1.76 2.14 3.59 4.22 3.59 3.59 3.92

0.00 0.68 0.77 0.95 0.96 0.98 0.98 0.99

D Source 1.00 0.12 0.17 0.11 0.24 0.21 0.07

0.00 0.17 0.24 0.25 0.25 0.25 0.25

0.00 0.42 0.78 0.80 0.89 0.89 0.95

E Source 1.00 0.13 0.06 0.24 0.07 0.18

0.00 0.10 0.10 0.11 0.11 0.11

Legend

ρ P90 COD 0.00 0.61 0.64 0.78 0.79 0.90

F Non

- Source 1.00 0.02 0.19 0.09 0.09

0.00 0.02 0.02 0.02 0.02

0.00 0.39 0.61 0.58 0.82

G Non

- Source 1.00 0.65 0.39 0.38

0.00 0.01 0.02 0.02

0.00 0.54 0.61 0.85

H Non

- Source 1.00 0.51 0.40

0.00 0.01 0.01

0.00 0.55 0.77

I Non

- Source 1.00 0.37

0.00 0.01

0.00 0.78

J Non

- Source 1.00

0.00

0.00



### Figure 3A

- 4 illustrates Pb monitor locations within Hillsborough and Pinellas Counties in FL, 

## 1

which comprise the greater Tampa

- St. Petersburg metropolitan area. Two source

- oriented monitors (A and 

## 2

B) were located within Hillsborough County, and one non

- source

- oriented monitor (C) was located in 

## 3

Pinellas County. Monitor A was located 360 m nort h

- northeast of the EnviroFocus Technologies battery 

## 4

recycling facility, which produced 1.3 tons/yr ( U.S. EPA, 2008c ) , and monitor B was located 320 m 

## 5

southwes t of the same facility. Monitor C was located next to a two

- lane road in Pinellas Park, FL. 

## 6

Figure 3A

- 5 displays seasonal wind roses for the Tampa

- St. Petersburg metropolitan area. These 

## 7

wind roses suggest shifting wind directions throughout the winter, sp ring, and summer. During the winter, 

## 8May 20 11 3

- 141 DRAFT

– DO NOT CITE OR QUOTE 

### the highest winds came from the north and northeast with little influence from the west and southwest. 

## 1

During spring and summer, easterly and westerly winds were evident from the wind rose, with winds 

## 2

from the west bein g slightly higher in wind speed. During autumn, winds came predominantly from the 

## 3

northeast with little signal from the west or south. 

## 4

Seasonal and year

- round concentrations are reported for Hillsborough and Pinellas Counties in 

## 5

Figure 3A

- 6. The average annual concentration at monitor A was 0.15 μg/m 3 

### , and the 95th percentile was 

## 6

0.70 μg/m 3



### . During winter, the 60th percentile of the data met the level of the NAAQS. At this site, the 

## 7

highest concentrations occurred during summer, which corresponded to the time when westerly winds 

## 8

were stronger. Concentration data at monitor B were much higher, with an annual average of 0.45 μg/m 3 

## 9

and a 95th percentile of 1.9 μg/m 3



### . Annually, the 55th percentile exceeded the level of the NAAQS, and in 

## 10

autumn the 45th percent ile exceeded the NAAQS. The highest concentrations occurred in autumn, 

## 11

coinciding with the time when winds blew from the northeast, when monitor B was most often downwind 

## 12

of the battery recycling facility. The non

- source

- oriented monitor C always reported concentrations of 0.0 

## 13

μg/m 3



### . This is likely related to its location next to a quiet road in a small city. 

## 14

Intersampler correlations, shown in Table 3A

- 8, illustrate that Pb has high intra

- urban spatial 

## 15

variability. The source oriented monitors were antico rrelated (ρ =

- 0.08). This was likely related to the 

## 16

fact that they were designated to monitor the same source and were downwind of the source at different 

## 17

times. 

## 18May 20 11 3

- 142 DRAFT

– DO NOT CITE OR QUOTE 

### Figure 3A

- 4

. Pb TSP m onitor locations within Hillsborough and Pinellas Counties, FL (12

- 057 and 12

- 103), 2007

- 2009. Top: view of all Pb FRM monitors in Hillsborough and Pinellas Counties. Bottom: Close up of industrial site around monitors A and B.May 20 11 3

- 143 DRAFT

– DO NOT CITE OR QUOTE S ource: NRCS ( 2011 )

. 

### Figure 3A

- 5

. Wind roses for Hillsborough/Pinellas Counties, FL, obtained from meteorological data at Tampa International Airport, 1961

- 1990. Clockwise from top left: January, April, July, and October. Note that wind percentages vary from month to month.May 20 11 3

- 144 DRAFT

– DO NOT CITE OR QUOTE Site A B C

SITE ID 12

- 057

- 1073 12

- 057

- 1066 12

- 103

- 3005

MEAN 0.15 0.45 0.00

SD 0.27 1.08 0.00

OBS 154 155 58

% BELOW MDL 20 5 95

Source orientation Source Source Non

- source



### Figure 3A

- 6

. Box plots of annual and seasonal Pb TSP concentrations ( μg/m 3 

### ) from source

- oriented and non

- source
- oriented monitors within Hillsborough and Pinellas Counties, FL (12

- 057 and 12

- 103), 2007

- 2009.May 20 11 3

- 145 DRAFT

– DO NOT CITE OR QUOTE 

### Table 3A

- 8

. Correlations between Pb TSP concentrations from source

- oriented and non

source
- oriented monitors within Hillsborough and Pinellas Counties, FL (12

- 057 and 12

- 103), 2007

- 2009.

A B C

Source Source Non

- source

A Source 1.00

- 0.08

0.00 1.20 0.50

0.00 0.71 1.00

B Source 1.00

0.00 2.20

Legend 0.00 1.00

C Non

- source ρ 1.00

P90 0.00

COD 0.00



### Figure 3A

- 7 illustrates Pb monitor locations within Cook County, IL. Eight monitors were located 

## 1

within Cook County, four of which were designated by the Illinois Environmental Protection Agency 

## 2

(IEPA) in data reported to the AQS as source

- oriented and the other four were non

- source

- oriented 

## 3

monitors. Monitor A was situated within 10 km of 6 sources ranging in emissions from 0.14 to 1.08 

## 4

tons/yr ( U.S. EPA, 2008a )

. Monitor A was also sited in the median of I

- 90/I

- 94. Monitor B was located on 

## 5

the northern roadside of I

- 290 and was within 10 km of 2 Pb sources (0.41 and 1.08 tons/yr) ( U.S. EPA, 

## 6

2008a )

. Monitor C was also located within 10 km of 6 sources in Cook County and Lake County, IN; the 

## 7

largest of those sources was 2.99 tons/yr and was located 8 km southeast of monitor C ( U.S. EPA, 2008a )

. 

## 8

Monitor C was placed on the roof of a high school. Monitor D was located roughly 60 m west of I

- 294 

## 9

and adjacent to O ’ Hare International Airport. Monitor E was located on th e rooftop of a building rented 

## 10

for government offices in Alsip, IL, a suburb south of Chicago. This location was roughly 1 km north of I



## 11

294 but not located on an arterial road; it was 9 km southeast of a 0.56 tons/yr source ( U.S. EPA, 2008a )

. 

## 12

Monitor F was sited in the parking lot of a water pumping station, 100 m north of I

- 90 and 300 m 

## 13

northwest of the junction between I

- 90 and I

- 94. This site was 2 km north

- northwest of a 0.10 tons/yr 

## 14

sourc e ( U.S. EPA, 2008a )

. Monitor G was situated atop an elementary school in a residential 

## 15

neighborhood on the south side of Chicago, roughly 100 m south of a rail line and over 300 m west of the 

## 16

closest arterial road. Although not designated as a source monitor, monitor G was located 2 km southwest 

## 17

of facilities emitting 0.30 and 0.41 tons/yr ( U.S. EPA, 2008a ) . Monitor H was sited on the grounds of the 

## 18

Northbrook Water Plant. I

- 94 curves around this site and was approximately 700 m from the monitor to 

## 19

the east and around to the north. Figure 3A

- 8 displays seasonal wind roses for Cook County. Wind 

## 20

patterns were quite variable during eac h season for this area. During the winter, winds mostly came from 

## 21May 20 11 3

- 146 DRAFT

– DO NOT CITE OR QUOTE 

### the west, with smaller contributions from the northwest, southwest, and south. In spring, measurable 

## 1

winds were omni

- directional, with the highest winds coming from the south and northeast. Winds 

## 2

originated predominantly from the southwest and south during the summer, with measurable contributions 

## 3

from the northeast as well. In autumn, wind flow was predominantly from the south, but smaller 

## 4

contributions also came from the southwest, west, an d northwest. 

## 5

Figure 3A

- 9 presents seasonal box plots of Pb concentration at the eight monitors located within 

## 6

Cook County. The maximum 95th percentile concentration on this plot was 0.14 μg/m 3 

### , so the scale of 

## 7

this box plot makes the variability in these data appear wider than the data presented for Los Angeles 

## 8

County and Hillsborough/Pinellas Counties. 

## 9

Monitor C was in closest proximity to the industrial steel facilities located in Lake County, IN. The 

## 10

average of concentrations measured at monitor C was 0.031 μg/m 3 

### , with a median of 0.02 μg/m 3 

### and a 

## 11

maximum concentration of 0.31 μg/m 3



### . In winter, the 95th percentile of data was 0.14 μg/m 3



### . The higher 

## 12

values could potentially be attributed to transport of emissions; winds blow from the southeast roughly 

## 13

10

- 15% of the time throughout the year. No other monitors in Cook County reported values above the 

## 14

level of the NAAQS. 

## 15

Three “ near

- road ” monitors, A, B, and D can be compared with the other monitors to consider the 

## 16

possibility of roadside resuspension of Pb dust from contemporaneous sources, as discussed in Section 

## 17

3.2.2.5. It would be expected that resuspension would diminish with distance from the road. The 2 

## 18

roadside monitors, A and B, reported average concentrations of 0.030 μg/m 3 

### and 0.024 μg/m 3



### , 

## 19

respect ively. The median concentrations for monitors A and B were 0.02 μg/m 3



### . Fifteen percent of data 

## 20

were below the MDL for monitor A, and 25% were below the MDL for monitor B. Note that data 

## 21

obtained from monitor A may reflect industrial emissions as well. Moni tor D was located roughly 60 m 

## 22

from the closest interstate and 570 m from the closest runway at O ’ Hare International Airport. However, 

## 23

the average concentration at this site was 0.012 μg/m 3 

### , and 85% of data were below the MDL. In contrast, 

## 24

non

- source monit ors, E, F, G, and H had average concentrations of 0.011

- 0.017 μg/m 3



### . It is possible that 

## 25

the difference between Pb concentrations at monitors A and B and Pb concentrations at the other monitors 

## 26

was related to proximity to the roadway, although this cannot be stated with certainty without source 

## 27

apportionment data to confirm or refute the influence of industrial plumes from Lake County, IN or local 

## 28

sources at each of the monitors. 

## 29

Comparison among the monitor data demonstrates a high degree of spatial variab ility (Table 3A

- 9). 

## 30

None of the source

- oriented monitors were well correlated with each other. The highest correlation 

## 31

between source

- oriented monitors occurred for monitors (A and B [ρ = 0.26]). This might have reflected 

## 32

more substantial differences relat ed to the additional influence of industrial sources nearby monitor A. 

## 33

Monitors (C and D) were uncorrelated with each other and with monitors (A and B), likely because their 

## 34

exposure to sources was substantially different. The source

- oriented and non

- sourc e

- oriented monitors 

## 35

were generally not well correlated. The highest correlation occurred between monitors (D and H [ρ = 

## 36May 20 11 3

- 147 DRAFT

– DO NOT CITE OR QUOTE 

### 0.53]). Both were located on the north side of Cook County, but monitor H was roughly 20 km northeast 

## 1

of monitor D. Winds blew from the southwest roughly 20

- 30% of the time throughout the year and from 

## 2

the northeast 20

- 25% of the time between the months of March and July, so the correlation may have 

## 3

been related to a common signal transported across both sites. Monitors (B and F [ρ = 0.46] ) were also 

## 4

moderately correlated. Monitor F is roughly 12 km northeast of monitor B, so the same common wind 

## 5

influence for monitors D and H may have also caused the moderate correlation between monitors (B and 

## 6

F). Monitor F was also moderately correlated with the other 3 non

- source monitors (ρ = 0.36 to 0.45), and 

## 7

the correlation between monitors (E and G) was ρ = 0.40. The data from monitor H did not correlate well 

## 8

with those from monitors E and G. The non

- source monitors were oriented from north to south over a 

## 9

distance of roughly 50 km in the following order: monitor H, monitor F, monitor G, and monitor E. The 

## 10

correlation pattern may have been related to distance between samplers. H was located in the suburb of 

## 11

Northbrook, monitors F and G were sited wit hin the Chicago city limits, and monitor E was situated in a 

## 12

town near the south side of Chicago. Differences among land use may have been related to the lack of 

## 13

correlation of the monitor H data with those from monitors E and G. It is likely that data from monitor F 

## 14

was at times better correlated with monitors E and G and at other times with monitor H, since it had 

## 15

moderate correlation with all three other non

- source monitors. 

## 16May 20 11 3

- 148 DRAFT

– DO NOT CITE OR QUOTE 

### Figure 3A

- 7

. Pb TSP Monitor locations within Cook County, IL (17

- 031),
2007

- 2009. Top: view of all Pb FRM monitors in Cook County. Bottom left: Close up of the high traffic site around monitor A. Bottom right: Close up of O ’ Hare International Airport adjacent
to monitor D.May 20 11 3

- 149 DRAFT

– DO NOT CITE OR QUOTE Source: NRCS ( 2011 ) 

### Figure 3A

- 8

. Wind roses for Cook County, IL, obtained from meteorological
data at O ’ Hare Internationa l Airport, 1961

- 1990. Clockwise fr om the top left: January, April, July, and October. Note that the wind percentages vary from month to month.May 20 11 3

- 150 DRAFT

– DO NOT CITE OR QUOTE Site A B C D E F G H

SITE ID 17

- 031

- 0026 17

- 031

- 6003 17

- 031

- 0022 17

- 031

- 3103 17

- 031

- 0001 17

- 031

- 0052 17

- 031

- 3301 17

- 031

- 4201

MEAN 0.030 0.024 0.031 0.012 0.013 0.017 0.017 0.011

SD 0.020 0.013 0.036 0.0062 0.0078 0.0098 0.0097 0.0031

OBS 179 175 177 168 177 175 171 168

% BELOW
MDL 15 25 25 85 75 55 50 95

Source
orientation Source Source Source Source Non

source Non

source Non

source Non

source



### Figure 3A

- 9

. Box plots of annual and seasonal Pb TSP concentrations ( μg/m 3 

### ) from source

- oriented and non

- source
- oriented monitors within Cook County, IL (17

- 031), 2007

- 2009.May 20 11 3

- 151 DRAFT

– DO NOT CITE OR QUOTE 

### Table 3A

- 9

. Correlations between Pb TSP concentrations from source
- oriented and non

source
- oriented monitors within Cook County, IL (17

- 031), 2007

- 2009

.

A B C D E F G H

Source Source Source Source Non

Source Non

Source Non

Source Non

Source

A Source 1.00 0.26

- 0.01 0.08 0.06 0.32 0.18 0.06

0.00 0.03 0.06 0.04 0.04 0.03 0.03 0.04

0.00 0.29 0.38 0.43 0.41 0.36 0.36 0.45

B Source 1.00 0.05 0.10 0.32 0.46 0.35

- 0.01

0.00 0.04 0.03 0.03 0.02 0.02 0.03

0.00 0.33 0.36 0.34 0.29 0.30 0.40

C Source 1.00 0.04 0.16 0.10 0.17 0.06

0.00 0.05 0.05 0.04 0.05 0.05

0.00 0.40 0.39 0.35 0.35 0.42

D Source 1.00 0.21 0.37 0.07 0.53

0.00 0.01 0.01 0.02 0.01

0.00 0.19 0.24 0.28 0.15

E Non

- Source 1.00 0.36 0.40 0.07

0.00 0.02 0.01 0.01

Legend 0.00 0.24 0.24 0.20

F Non

- Source ρ 1.00 0.41 0.45

P90 0.00 0.01 0.02

COD 0.00 0.24 0.26

G Non

- Source 1.00 0.05

0.00 0.02

0.00 0.27

H Non

- Source 1.00

0.00

0.00



### Figure 3A

- 10 illustrates Pb monitor locations with Jefferson County, MO. Ten source

- oriented 

## 1

monitors surrounded the Doe Run primary Pb smelter in Herculaneum, MO on the west and northwestern 

## 2

sides. The largest distance between these monitors was approximately 1.5 km. Monitor E locat ed on the 

## 3

Doe Run facility roughly 20 m west of the nearest building. Monitors A, B, C, D, F, G, and H were all 

## 4

located approximately 200 m west of the facility. Monitors D, E, and H were situated alongside service 

## 5

roads to the facility. Monitor I was site d 100 m north of the smelter, and monitor J was located 

## 6

approximately 600 m northwest of the facility. The Doe Run smelter was the only active primary smelter 

## 7

in the U.S. at the time of this review (2007

- 2009), and the facility was estimated to have emitte d 

## 8

41.1 tons Pb/yr ( U.S. EPA, 2008d ) . Figure 3A

- 11 displays seasonal wind roses for Jefferson County. 

## 9

During winter, predominant winds originated from the northwest, with a smaller fraction of calmer winds 

## 10

originating in the south

- southeast. During the spring, the south

- southeasterly winds became more 

## 11

prevalent with a measurable fraction of stronger winds still originating in the north

- northwest. In the 

## 12May 20 11 3

- 152 DRAFT

– DO NOT CITE OR QUOTE 

### summ er, winds were omni

- directional and generally calmer. A slightly larger percentage came from the 

## 1

south compared with other wind directions. Autumn winds were most predominantly south

- southeastern, 

## 2

with a smaller fraction from the west and northwest. 

## 3

Figur e 3A

- 12 illustrates the seasonal distribution of concentrations at monitors A

- J in Jefferson 

## 4

County. The annual average concentrations ranged from 0.18 to 1.36 μg/m 3 

### across the monitors. The 

## 5

maximum concentration was measured at monitor C to be 21.6 μg/m 3



### , which was 144 times higher than 

## 6

the level of the standard. For this monitor, the 25th percentile of the data was at the level of the standard. 

## 7

In general, median and 75 th percentile concentrations were highest during the springtime and second 

## 8

highest duri ng the fall. These seasons coincide with periods when the southeastern winds were stronger 

## 9

and more prevalent. Because the Doe Run facility had two 30

- meter stacks ( Bennett, 2007 ) , it is possible 

## 10

that the emissions measured at the closer monitors were due to either fugitive emissions from the plant or, 

## 11

for the case where ground equipment or vehicles are ope rated nearby, that previously deposited emissions 

## 12

from the plant were resuspended. 

## 13

Spatial variability among the monitors is lower than at many sites, because the monitors are 

## 14

relatively close together and are located on one side of the same source (Table 3A

- 10). Correlations range 

## 15

from ρ =

- 0.04 to 0.96. High correlations (ρ ≥ 0.75) occurred for monitors (A and C), (A and D), (C and 

## 16

D), (D and F), (E and F), (G and H), and (I and J). Monitors (A and C), (A and D), (C and D), (D and F), 

## 17

(E and F), and (G an d H) are all within 250 m of each other. For the highest correlation (ρ = 0.96, for 

## 18

monitors (E and F), monitor F is 250 m directly east of monitor E. Low correlation (ρ ≤ 0.25) generally 

## 19

occurred when monitors B, I, and J were compared with monitors A, C, D, E, F, G, and H. Monitors B, I, 

## 20

and J were on the outskirts of the measurement area and so were likely oriented such that the 

## 21

southeasterly winds did not carry pollutant to these sites concurrently with the signal recorded by the 

## 22

other monitors. 

## 23May 20 11 3

- 153 DRAFT

– DO NOT CITE OR QUOTE 

### Figure 3A

- 10

. Pb TSP Monitor locations within Jefferson County, MO (29

- 099), 2007

- 2009. Note that all monitors surround the Doe Run industrial facility. Top: Map view of all monitors in Jefferson County. Bottom: Satellite view of the monitors and the Doe Run facility.May 20 11 3

- 154 DRAFT

– DO NOT CITE OR QUOTE Source: NRCS ( 2011 ) 

### Figure 3A

- 11

. Wind roses for Jefferson County, MO, obtained from meteorologic al data at St. Louis/Lambert International Airport,
1961

- 1990. Clockwise from top left: January, April, July, and
October. Note wind percentages vary from month to month.May 20 11 3

- 155 DRAFT

– DO NOT CITE OR QUOTE Site A B C D E F G H I J

SITE ID 29

- 099

- 0022 29

- 099

- 0024 29

- 099

- 0015 29

- 099

- 0023 29

- 099

- 0004 29

- 099

- 0020 29

- 099

- 0021 29

- 099

- 0005 29

- 099

- 0011 29

- 099

- 0013

MEAN 0.43 0.36 1.36 0.39 1.12 0.69 0.75 0.29 0.34 0.18

SD 0.54 0.49 1.97 0.54 1.67 1.01 1.25 0.59 0.85 0.33

OBS 622 2 09 1E3 632 1E3 575 953 351 366 177

% BELOW
MDL 0 5 0 0 5 0 5 25 5 15

Source

or ientation Source Source Source Source Source Source Source Source Source Source



### Figure 3A

- 12

. Box plots of annual and seasonal Pb TSP concentrations ( μg/m 3 

### ) from source

- oriented and non

- source
- oriented monitors within Jefferson County, MO (29

- 099), 2007

- 2009.May 20 11 3

- 156 DRAFT

– DO NOT CITE OR QUOTE 

### Ta ble 3A

- 

### 1 0

. Correlations between Pb TSP concentrations from source

- oriented and

non
- source
- oriented monitors within Jefferson County, MO (29

- 099), 2007

- 2009

A B C D E F G H I J

Source Source Source Source Source Source Source Source Source Source

A Source 1.00 0.59 0.80 0.83 0.57 0.64 0.33 0.35 0.07 0.05

0.00 0.71 1.55 0.42 1.93 1.14 1.41 0.74 0.92 0.78

0.00 0.46 0.48 0.30 0.55 0.45 0.57 0.64 0.67 0.69

B Source 1.00 0.53 0.43 0.10 0.14 0.07 0.22 0.10 0.09

0.00 1.86 0.87 2.77 1.96 2.08 0.94 1.04 0.91

0.00 0.58 0.51 0.69 0.62 0.68 0.68 0.65 0.65

C Source 1.00 0.86 0.59 0.72 0.26 0.27

- 0.04 0.04

0.00 1.56 2.26 1.26 2.94 2.65 3.18 2.60

0.00 0.50 0.50 0.46 0.60 0.74 0.73 0.73

D Source 1.00 0.71 0.80 0.41 0.56 0.14 0.18

0.00 1.83 1.02 1.38 0.76 0.88 0.70

0.00 0.50 0.36 0.53 0.61 0.63 0.66

E Source 1.00 0.96 0.54 0.46 0.06 0.16

0.00 0.86 2.16 2.50 3.09 2.57

Legend 0.00 0.35 0.49 0.66 0.70 0.72

F Source ρ 1.00 0.56 0.54 0.10 0.19

P90 0.00 1.13 1.51 1.74 1.40

COD 0.00 0.47 0.63 0.65 0.70

G Source 1.00 0.87 0.28 0.38

0.00 1.53 2.10 2.08

0.00 0.61 0.63 0.66

H Source 1.00 0.20 0.30

0.00 0.89 0.56

0.00 0.67 0.65

I Source 1.00 0.79

0.00 0.62

0.00 0.48

J Source 1.00

0.00

0.00



### Figure 3A

- 13 illustrates Pb monitor locations with Cuyahoga County, OH. Five monitors are 

## 1

located within Cuyahoga County, three of which were designated by the Ohio EPA (OEPA) as source



## 2

oriented and the other two were non

- source

- oriented monitors. Monitors A, B, and C were all located 

## 3

within 1

- 10 km of six 0.1 tons/yr source facilities and one 0.2 tons/yr source ( U.S. EPA, 2008e )

. 

## 4

Additionally, monitor B was located 30 m north of the Ferro Corporation headquarters. This facility was 

## 5

stated in the 2005 NEI to have no emissions, but it was thou ght by the OEPA to be the source of 

## 6

exceedances at this monitor ( U.S. EPA, 2008e )

. Monitor A was sited roughly 300 m south of the Ferro 

## 7

Corporation facility. Mon itor C was located 2.2 km west

- northwest of the 0.5 ton/yr Victory White Metal 

## 8May 20 11 3

- 157 DRAFT

– DO NOT CITE OR QUOTE 

### Co. facility. Monitor C was also roughly 250 m southeast of I

- 490. Monitors D and E were designated as 

## 1

non

- source

- oriented monitors, although monitor D was just 600 m further fr om the Victory White Metal 

## 2

facility than was monitor C. Monitor D was sited on a residential street located 50 m north of I

- 490. 

## 3

Monitor E was located on the rooftop of a building within 20 m of a four

- lane arterial road. Figure 3A

- 14 

## 4

displays seasonal win d roses for Cuyahoga County. During winter, summer, and autumn, the predominant 

## 5

winds were from the southwest, with stronger winds recorded during the winter. In the spring, the 

## 6

strongest winds still emanated from the south

- southwest, but measurable winds were also scattered from 

## 7

the northeast to the northwest. 

## 8

Figure 3A

- 15 illustrates the seasonal distribution of Pb concentration data at the five monitoring 

## 9

sites. The influence of southern winds, along with close proximity to a potentially

- emitting facili ty, could 

## 10

have caused the elevated concentrations observed at monitor B (av era g e : 0.10 μg/m 3 

### ). The 80th percentile 

## 11

of data was at the level of the NAAQS at this monitor, and during autumn the 60th percentile of data met 

## 12

the level of the NAAQS. The maximum concentration during fall and for the monitor year

- round was 

## 13

0.22 μg/m 3



### . Concentration data from all other monitors were below the level of the NAAQS. For monitor 

## 14

A, the average concentration was 0.025 μg/m 3 

### , and the median reached 0.04 μg/m 3 

### during the su mmer. 

## 15

Maximum concentration at this monitor was 0.07 μg/m 3



### . Concentrations at monitor C averaged 0.017 

## 16

μg/m 3 

### , and those at monitors D and E averaged 0.014 μg/m 3 

### and 0.013 μg/m 3 

### , respectively. Maximum 

## 17

concentrations reached 0.04 μg/m 3 

### at all three monitors. 

## 18

The level of spatial variability is illustrated by the intersampler correlations presented in Table 3A



## 19

11. Monitors A and B appear to be anticorrelated (ρ =

- 0.13). If the Ferro site was the dominant source in 

## 20

this area, then the anticorrelation was likel y caused by the positioning of monitors A and B on opposite 

## 21

sides of that facility. At any given time, potential emissions from the Ferro plant may have affected 

## 22

monitors A and B at distinct times. Monitors C, D, and E correlated well with each other (ρ = 0.67 to 

## 23

0.77). Given that all 3 monitors are separated by roughly 2.8 km, it is possible that the relatively high 

## 24

correlations related to common sources, as suggested in the previous paragraph. Little correlation was 

## 25

observed between the source

- oriented an d non

- source

- oriented monitors. 

## 26May 20 11 3

- 158 DRAFT

– DO NOT CITE OR QUOTE 

### Figure 3A

- 13

. Pb TSP Monitor locations within Cuyahoga County, OH (39

- 035), 2007

- 2009. Top: view of all Pb FRM monitors in Cuyahoga County. Bottom left: Close up of industrial site around
monitors A and B. Bottom right: Close up of monitor D north
of I

- 490.May 20 11 3

- 159 DRAFT

– DO NOT CITE OR QUOTE Source: NRCS ( 2011 ) 

### Figure 3A

- 14

. Wind roses for Cuyahoga County, OH, obtained from meteorological data at Cleveland/Hopkins International Airport, 1961

- 90. Clockwise from top left: Jan, April, July, and October. Note wind percentages vary from month to month.May 20 11 3

- 160 DRAFT

– DO NOT CITE OR QUOTE Site A B C D E

SITE ID 39

- 035

- 0050 39

- 035

- 0049 39

- 035

- 0061 39

- 035

- 0038 39

- 035

- 0042

MEAN 0.025 0.10 0.017 0.014 0.013

SD 0.018 0.060 0.010 0.0072 0.0076

OBS 36 36 36 35 36

% BELOW MDL 20 0 30 45 45

Source orientation Source Source Source Non

- source Non

- source



### Figure 3A

- 15

. Box plots of annual and seasonal Pb TSP concentrations ( μg/m 3 

### ) from source

- oriented and non

- source
- oriented monitors within Cuyahoga County, OH (39

- 035), 2007

- 2009.May 20 11 3

- 161 DRAFT

– DO NOT CITE OR QUOTE 

### Table 3A

- 1 1

. Correlations between Pb TSP concentrations from source

- oriented and

non
- source
- oriented monitors within Cuyahog a County, OH (39

- 035), 2007

- 2009

A B C D E

Source Source Source Non

- Source Non

- Source

A Source 1.00

- 0.13 0.24 0.19 0.21

0.00 0.18 0.05 0.04 0.05

0.00 0.64 0.33 0.35 0.37

B Source 1.00 0.31 0.24 0.34

0.00 0.18 0.19 0.19

0.00 0.69 0.71 0.73

C Source 1.00 0.77 0.67

0.00 0.01 0.01

Legend 0.00 0.17 0.18

D Non

- Source ρ 1.00 0.67

P90 0.00 0.01

COD 0.00 0.17

E Non

- Source 1.00

0.00

0.00



### Figure 3A

- 16 illustrates Pb monitor locations within Sullivan County, TN. Three source

- oriented 

## 1

monitors were situated around an Exide Pb recycling facility emitting 0.78 tons/yr ( U.S. EPA, 2008f )

. 

## 2

Monitors A and C are positioned along the facility ’ s service road and are approximately 100 m and 200 m 

## 3

away from the facility, respectively. Monitor A is directly next to the road, and monitor C is roughly 15 m 

## 4

from the road. Monitor B is located in the facility ’ s p arking lot roughly 50 m from the closest building. 

## 5

The facility and all three monitors are approximately 1.5 km northwest of the Bristol Motor Speedway 

## 6

and Dragway racetracks, which hosts a variety of auto races each year, including NASCAR, KART, and 

## 7

drag racing. Although the NASCAR circuit no longer uses tetraethyl Pb as an anti

- knock agent in its fuel, 

## 8

some of the smaller rac ing circuits continue to do so. However, the speedway is rarely upwind of the 

## 9

monitoring sites and so likely had minimal influence on the reported concentrations. Figure 3A

- 17 

## 10

displays seasonal w ind roses for Sullivan County. During winter and spring, the predominant winds com e 

## 11

from the southwest and west. In the summer, the percentage of wind coming from the we st and southwest 

## 12

is roughly equal to that for wind coming from the east and northeast, although the easterly winds are 

## 13

calmer. During autumn, winds come predominantly from the northeast and east, although these winds 

## 14

tend to be calmer than those originatin g from the southwest and west. 

## 15

The data presented in Figure 3A

- 18 illustrates that concentrations above the level of the NAAQS 

## 16

occurr ed frequently at the monitors. The average concentrations at monitors A, B, and C were 0.11 μg/m 3



### , 

## 17

0.051 μg/m 3 

### , and 0.059 μg/m 3 

### , respectively. Median concentrations were 0.08 μg/m 3 

### , 0.03 μg/m 3 

### , and 

## 18

0.04 μg/m 3 

### , respectively. The 75 th 

### percentile of year

- round data at monitor A was at the level of the 

## 19May 20 11 3

- 162 DRAFT

– DO NOT CITE OR QUOTE 

### NAAQS, while the 95th percentile of data were below the NAAQS level for monito rs B and C. The 

## 1

maxima at each monitor were 0.76 μg/m 3 

### , 0.26 μg/m 3 

### , and 0.43 μg/m 3 

### for monitors A, B, and C. It was 

## 2

surprising that the concentrations measured at monitor A tended to be higher because the predominant and 

## 3

stronger winds came from the southw est, so in many cases monitor A was upwind of the facility. It is 

## 4

possible that Pb that had either deposited or was stored in waste piles became readily resuspended by 

## 5

traffic

- related turbulence and was measured at monitor A since that monitor was closest to the road. The 

## 6

slightly higher concentrations at monitor C compared with those from monitor C are consistent with the 

## 7

southwestern winds. 

## 8

Not surprisingly, the correlations of monitor A with monitors B and C were quite low (Table 3A



## 9

12). The correlation between monitors B and C was ρ = 0.45. It makes sense that the correlation for these 

## 10

monitors would be somewhat higher because they are both oriented to the east of the Pb recycling facility, 

## 11

although monitor C is to the northeast and monitor B to the east

- southeast. 

## 12May 20 11 3

- 163 DRAFT

– DO NOT CITE OR QUOTE 

### Figure 3A

- 16

. Pb TSP Monitor locations within Sullivan County, TN (47

- 163),
2007

- 2009. Top: Map, bottom: Satellite image. Monitors A, B, and C surround the Exide Pb recycling facility. Just to the southeast is the Bristol motor speedway.May 20 11 3

- 164 DRAFT

– DO NOT CITE OR QUOTE Source: NRCS ( 2011 ) 

### Figure 3A

- 17

. Wind roses for Sullivan County, TN, obtained from meteorological data at B ri stol/Tri City Airport, 1961

- 90. Clockwise from top left: January, April, July, and October. Note that the wind percentages vary from month to month.May 20 11 3

- 165 DRAFT

– DO NOT CITE OR QUOTE Site A B C

SITE ID 47

- 163

- 3001 47

- 163

- 3002 47

- 163

- 3003

MEAN 0.11 0.051 0.059

SD 0.11 0.036 0.047

OBS 334 362 345

% BELOW MDL 0 0 0

Source orientation Source Source Source



### Figure 3A

- 18

. Box plots of annual and seasonal Pb TSP concentrations ( μg/m 3 

### ) from source

- oriented and non

- source
- oriented monitors within Sullivan County, TN (47

- 163), 2007

- 2009.May 20 11 3

- 166 DRAFT

– DO NOT CITE OR QUOTE 

### Table 3A

- 1 2

. Correlations between Pb TSP concentrations from source

- oriented and

non
- source
- oriented monitors within Sullivan County, TN (47

- 163), 2007

- 2009

A B C

Source Source Source

A Source 1.00

- 0.04 0.13

0.00 0.21 0.19

0.00 0.47 0.43

B Source 1.00 0.45

0.00 0.06

Legend 0.00 0.23

C Source ρ 1.00

P90 0.00

COD 0.00



# 3.8.2. Size Distribution of Pb

- Bearing PM

Table 3A

- 1 3

. Correlations and average of the concentration ratios for co

- located monitors, TSP v ersu s PM 10 

### , TSP v ersu s PM 2.5 

### , and PM 10 

### v ersu s PM 2.5 

### . Data are bolded for sites where the TSP, PM 10 

### , and PM 2.5 

### monitors were co

located for at least one sampling year.

Site County Land Type Years Corr Avg Ratio Years Corr Avg Ratio Years Corr Avg Ratio

PM 10 :TSP PM 2.5 :TSP PM 2.5 : PM 10

010730023 Jefferson, AL Urban and center city 2005

- 2006 0.87 0.80 2005

- 2006 0.74 0.84 2005

- 2006 0.81 1.02

040139997 Maricopa, AZ Urban and center city 2006

- 2009 0.40 0.59

0600110 01 Alameda, CA Suburban 1994

- 1998 0.00 2.31

060130002 Contra Costa, CA Suburban 1994

- 1998 0.27 2.60 060130003 Contra Costa, CA Suburban 1994

- 1997 0.31 1.04 060190008 Fresno, CA Suburban 1995

- 2001 0.92 0.92 2000

- 2009 0.58 0.66 2000

- 2001 0.83 1.16 060250005 Imperial, CA Suburban 1996

- 2001 0.77 0.91 2002

- 2009 0.95 0.74

060290014 Kern, CA Urban and center city 1995

- 2000 0.92 0.76 2001

- 2009 0.22 0.72

060371103 Los Angeles, CA Urban and center city 2002

- 2009 0.22 0.28

060374002 Los Angeles, CA Suburban 1995

- 2000 0.72 0.39

060550003 Napa, CA Urban and center city 1994

- 1994 0.42 0.93

060658001 Riverside, CA Suburban 1995

- 1997 0.27 0.39 2001

- 2009 0.59 0.40 060730003 San Diego, CA Suburban 2001

- 2009 0.35 0.48

060750005 San Francisco, CA Urban and center city 1994

- 1998 0.29 0.98

060771002 San Joaquin, CA Urban and center city 1995

- 2000 0.73 0.82

060850004 Santa Clara, CA Urban and center city 1994

- 2000 0.32 0.95 2000

- 2002 0.63 0.63

060850004 Santa Clara, CA Urban and center city 2002

- 2009 0.11 0.78 2008

- 2009 0.76 1.41

060853001 Santa Clara, CA Suburban 1994

- 1998 0.16 1.00

060990002 Stanislaus, CA Urban and center city 1995

- 1998 0.94 0.82

060990005 Stanislaus, CA Urban and center city 1998

- 2000 0.79 1.02

061112002 Ventura, CA Suburban 2001

- 2009 0.47 0.74May 20 11 3

- 167 DRAFT

– DO NOT CITE OR QUOTE Site County Land Type Years Corr Avg Ratio Years Corr Avg Ratio Years Corr Avg Ratio

PM 10 :TSP PM 2.5 :TSP PM 2.5 : PM 10
080010006 Adams, CO Suburban 2001

- 2006 0.16 0.43

080410011 El Paso, CO Urban and center city 2001

- 2006 0.03 0.39

080770017 Mesa, CO Urban and center city 2004

- 2008 0.42 0.87

100032004 New Castle, DE Urban and center city 2003

- 2007 0.47 0.44

110010043 District of Columbia, DC Urban and center city 2004

- 2009 0.47 0.73

120571075 Hillsborough, FL Urban and center city 2001

- 2003 0.83 1.86

120573002 Hillsborough, FL Rural 2004

- 2009 0.66 1.55 121030026 Pinellas, FL Suburban 2004

- 2009 0.56 1.89 130690002 Coffee, GA Rural 2003

- 2009 0.82 1.46 130890002 DeKalb, GA Suburban 2003

- 2009 0.92 1.69

150032004 Honolulu, HI Urban and center city 2003

- 2009 0.77 2.34

170314201 Cook, IL Suburban 2006

- 2009 0.53 0.39 2002

- 2009 0.25 0.36 2005

- 2009 0.65 1.01 180970078 Marion, IN Suburban 2002

- 2009 0.75 0.78 201731012 Sedgwick, KS Suburban 1993

- 1997 0.63 0.33

202090015 Wyandotte, KS Urban and center city 1993

- 1997 0.66 0.55

202090020 Wyandotte, KS Urban and center city 1993

- 1997 0.99 0.56

210430500 Carter, KY Rural 2008

- 2009 0.22 1.39 211930003 Perry, KY Suburban 2003

- 2008 0.76 1.47 220511001 Jefferson, LA Suburban 2005

- 2006 0.96 0.79

220710010 Orleans, LA Urban and center city 2005

- 2006 0.94 0.64

220710012 Orleans, LA Urban and center city 2005

- 2006 0.91 0.53

220718105 Orleans, LA Urban and center city 2005

- 2006 0.98 0.80

220718401 Orleans, LA Urban and center city 2005

- 2006 0.89 0.45

220758400 Plaquemines, LA Urban and center city 2005

- 2005 0.98 0.83

220870004 St. Bernard, LA Suburban 2005

- 2006 0.94 0.78

220878103 St. Bernard, LA Urban and center city 2005

- 2006 0.90 0.66

250250042 Suffolk, MA Urban and center city 2009

- 2009 0.43 0.37 2003

- 2009 0.38 0.78

260770905 Kalamazoo, MI Urban and center city 1993

- 1996 0.98 0.77

260810020 Kent, MI Urban and center city 2005

- 2007 0.58 0.70

261130001 Missaukee, MI Rural 2002

- 2007 0.73 1.38

261250010 Oakland, MI Urban and center city 2001

- 2002 0.40 1.60

261390009 Ottawa, MI Urban and center city 2000

- 2001 0.90 0.67

261610008 Washtenaw, MI Urban and center city 2003

- 2007 0.58 0.71

261630001 Wayne, MI Suburban 2001

- 2007 0.68 0.58 261630019 Wayne, MI Suburban 2001

- 2002 0.45 0.51 261630033 Wayne, MI Suburban 2003

- 2009 0.87 0.84 2002

- 2009 0.84 0.49 2003

- 2009 0.86 0.61

270530053 Hennepin, MN Urban and center city 1996

- 2001 0.47 0.55

270530963 Hennepin, MN Urban and center city 2002

- 2009 0.64 0.42

280458104 Hancock, MS Suburban 2005

- 2006 0.93 1.51 280458105 Hancock, MS Rural 2005

- 2006 0.99 1.02 280458201 Hancock, MS Suburban 2005

- 2006 0.94 1.05 280470008 Harrison, MS Rural 2005

- 2006 0.87 0.99 280478101 Harrison, MS Suburban 2005

- 2006 0.98 0.99 280478102 Harrison, MS Suburban 2005

- 2006 0.80 1.10 280478103 Harrison, MS Suburban 2005

- 2006 0.93 1.09

280590006 Jackson, MS Urban and center city 2005

- 2006 0.97 0.78

295100085 St. Louis City, MO Urban and center city 2004

- 2004 0.96 1.10 2004

- 2004 0.26 1.08 2003

- 2009 0.76 0.76

330110020 Hillsborough, NH Urban and center city 2002

- 2005 0.42 0.81May 20 11 3

- 168 DRAFT

– DO NOT CITE OR QUOTE Site County Land Type Years Corr Avg Ratio Years Corr Avg Ratio Years Corr Avg Ratio

PM 10 :TSP PM 2.5 :TSP PM 2.5 : PM 10

330150014 Rockingham, NH Urban and center city 2003

- 2005 0.80 0.72

350010023 Bernalillo, NM Urban and center city 2004

- 2009 0.04 0.73

360050110 Bronx, NY Urban and
center city 2007

- 2009 0.80 0.60

360551007 Monroe, NY Urban and center city 2007

- 2009 0.77 1.30

360610062 New York, NY Urban and center city 2005

- 2005 0.96 0.57

360632008 Niagara, NY Suburban 2005

- 2005 0.74 0.84 401091037 Oklahoma, OK Suburban 2009

- 2009 0.25 1.57

401431127 Tulsa, OK Urban and center city 2009

- 2009 0.34 0.74

410390060 Lane, OR Urban and center city 2002

- 2004 0.33 1.65

410510246 Multnomah, OR Urban and center city 2002

- 2003 0.90 0.90

410290133 Jackson, OR Urban and center city 2009

- 2009 0.54 1.10

410390060 Lane, OR Urban and center city 2004

- 2009 0.76 2.15

410510246 Multnomah, OR Urban and center city 2003

- 2006 0.98 0.99

410610119 Union, OR Urban and
center city 2004

- 2007 0.15 1.60

420450002 Delaware, PA Urban and center city 2002

- 2008 0.04 0.11

420710007 Lancaster, PA Suburban 2004

- 2007 0.78 0.80

421010004 Philadelphia, PA Urban and center city 2000

- 2007 0.37 0.38

421010055 Philadelphia, PA Urban and center city 2005

- 2007 0.36 0.32

421010136 Philadelphia, PA Urban and center city 2004

- 2005 0.75 0.38

440070022 Providence, RI Urban and center city 2001

- 2002 0.78 0.54 2002

- 2009 1.00 0.70

440071010 Providence, RI Suburban 2001

- 2002 0.98 0.71 450250001 Chesterfield, SC Rural 2002

- 2009 0.37 1.76 2004

- 2009 0.41 1.63

450790019 Richland, SC Urban and center city 2001

- 2006 0.74 0.58

470370023 Davidson, TN Urban and center city 2003

- 2004 0.39 0.83

482011034 Harris, TX Urban and center city 2002

- 2005 0.41 0.40

482011039 Harris, TX Suburban 2000

- 2009 0.34 0.63 490110004 Davis, UT Suburban 2003

- 2009 0.68 0.91 500070007 Chittenden, VT Rural 2004

- 2009 0.89 0.97 510870014 Henrico, VA Suburban 2004

- 2008 0.54 0.79 2008

- 2009 0.91 1.08 530330038 King, WA Suburban 2001

- 2002 0.64 0.53

530330080 King, WA Urban and center city 2003

- 2009 0.91 0.81

530630016 Spokane, WA Suburban 2005

- 2005 0.77 0.91 550270007 Dodge, WI Rural 2004

- 2005 0.90 0.83 2005

- 2009 0.59 1.02



## 1May 20 11 3

- 169 DRAFT

– DO NOT CITE OR QUOTE 

# 3.8.3. Lead Concentration in a Multipollutant Context

Source, SO 2

Non

- Source, SO 2
Source, PM 2.5

Non

- Source, PM 2.5
Source, PM 10

Non

- Source, PM 10
Source, O 3

Non

- Source, O 3
Source, NO 2

Non

- Source, NO 2
Source, CO

Non

- Source, CO

Source, SO 2

Non

- Source, SO 2
Source, PM 2.5

Non

- Source, PM 2.5
Source, PM 10

Non

- Source, PM 10
Source, O 3

Non

- Source, O 3
Source, NO 2

Non

- Source, NO 2
Source, CO

Non

- Source, CO



### Figure 3A

- 19a. Seasonal correlations of monitored Pb

- TSP concentration with copollutant concentrations, 2007

- 2008. Top: winter; bottom: spring.May 20 11 3

- 170 DRAFT

– DO NOT CITE OR QUOTE 

### Figure 3A

- 19b. Seasonal correlations of monitored Pb

- TSP concentration with copollutant concentrations, 2007

- 2008. Top: summer;
bottom: fall. Source, SO 2

Non

- Source, SO 2
Source, PM 2.5

Non

- Source, PM 2.5
Source, PM 10

Non

- Source, PM 10
Source, O 3

Non

- Source, O 3
Source, NO 2

Non

- Source, NO 2
Source, CO

Non

- Source, CO

Source, SO 2

Non

- Source, SO 2
Source, PM 2.5

Non

- Source, PM 2.5
Source, PM 10

Non

- Source, PM 10
Source, O 3

Non

- Source, O 3
Source, NO 2

Non

- Source, NO 2
Source, CO

Non

- Source, COMay 20 11 3

- 171 DRAFT

– DO NOT CITE OR QUOTE Non

- Source, SO 2

Source, PM 2.5

Non

- Source, PM 2.5

Non

- Source, PM 10

Non

- Source, O 3

Non

- Source, NO 2

Non

- Source, CO

Non

- Source, SO 2

Source, PM 2.5

Non

- Source, PM 2.5

Non

- Source, PM 10

Non

- Source, O 3

Non

- Source, NO 2

Non

- Source, CO



### Figure 3A

- 20a. Seasonal correlations of monitored Pb

- TSP concentration with copollutant concentrations, 2009. Top: winter; bottom: spring. US Winter

Spearman Correlation Coefficient -1.0 -0.5 0.0 0.5 1.0 e e e e e e eMay 20 11 3

- 172 DRAFT

– DO NOT CITE OR QUOTE 

### Figure 3A

- 20b. Seasonal correlations of monitored Pb

- TSP concentration
with copollutant concentrations, 2009. Top: summer;
bottom: fall. Non

- Source, SO 2

Source, PM 2.5

Non

- Source, PM 2.5

Non

- Source, PM 10

Non

- Source, O 3

Non

- Source, NO 2

Non

- Source, CO

Non

- Source, SO 2

Source, PM 2.5

Non

- Source, PM 2.5

Non

- Source, PM 10

Non

- Source, O 3

Non

- Source, NO 2

Non

- Source, COMay 20 11 3

- 173 DRAFT

– DO NOT CITE OR QUOTE 

### Figure 3A

- 21a. Seasonal correlations of monitored Pb

- PM 2.5 

### concentration with copollutant concentrations, 2007

- 2009. Left: winter; right: spring. Zinc

Volatile Nitrate

Vanadium

Titanium

Sulfur

Sulfate

Sodium Ion

Sodium

Silicon

Selenium

Potassium Ion

Potassium Organic Carbon, Blank Adjusted

Non

- Volatile Nitrate

Nitrate

Nickel
Mercury Manganese

Magnesium

Iron

Elemental Carbon

Crustal
Copper

Chromium

Chlorine

Carbonate Carbon

Calcium

Cadmium

Bromine

Arsenic

Ammonium

Aluminum Zinc

Volatile Nitrate

Vanadium

Titanium

Sulfur

Sulfate

Sodium Ion

Sodium

Silicon

Selenium

Potassium Ion

Potassium Organic Carbon, Blank Adjusted

Non

- Volatile Nitrate

Nitrate

Nickel
Mercury Manganese

Magnesium

Iron

Elemental Carbon

Crustal
Copper

Chromium

Chlorine

Carbonate Carbon

Calcium

Cadmium

Bromine

Arsenic

Ammonium

AluminumMay 20 11 3

- 174 DRAFT

– DO NOT CITE OR QUOTE 

### Figure 3A

- 21b. Seasonal correlations of moni tored Pb

- PM 2.5 

### concentration with copollutant concentrations, 2007

- 2009. Left: summer; right: fall. Zinc

Volatile Nitrate

Vanadium

Titanium

Sulfur

Sulfate

Sodium Ion

Sodium

Silicon

Selenium

Potassium Ion

Potassium Organic Carbon, Blank Adjusted

Non

- Volatile Nitrate

Nitrate

Nickel
Mercury
Manganese
Magnesium

Iron

Elemental Carbon

Crustal
Copper

Chromium

Chlorine

Carbonate Carbon

Calcium

Cadmium

Bromine

Arsenic

Ammonium

Aluminum Zinc

Volatile Nitrate

Vanadium

Titanium

Sulfur

Sulfate

Sodium Ion

Sodium

Silicon

Selenium

Potassium Ion

Potassium Organic Carbon, Blank Adjusted

Non

- Volatile Nitrate

Nitrate

Nickel

Mercury Manganese

Magnesium

Iron

Elemental Carbon

Crustal
Copper

Chromium

Chlorine

Carbonate Carbon

Calcium

Cadmium

Bromine

Arsenic

Ammonium

AluminumMay 20 11 3

- 175 DRAFT

– DO NOT CITE OR QUOTE 

### Table 3A

- 14

. Copollutant exposures for various trace metal studies

Adgate et al. ( 2007 ) Riediker et al. ( 2003 ) Pekey et al. ( 2010 ) Molnar et al. ( 2007 )

I

- R (med) a,b Personal (median) c Vehicle (range) c Roadside (range) c I

– near industry (range) a I

- R (median) a,b I

- Scho ol (median) a I

- Pre

- School (median) a

Location Minnesota New Jersey Kocaeli, Turkey Stockholm, Sweden

PM 2.5 24,000 31,579 24,400

- 29,800

Pb 1.5 3.2 2

- 3 4

- 6 34

- 85 2.8 2.5 1.7

S 272.1 351.6 905

- 1592 1416

- 2231 435

- 489 330 290 220

Ca 85.0 174.1 31

- 44 18

- 40 309

- 452 70 110 58

Al 23.3 58.6 53

- 60

Na 20.6 31.9

Fe 43.1 78.6 307

- 332 82

- 163 44

- 58 57 100 71

Mg 16.3 27.5

K 38.4 47.5 6

- 75 23

- 57 160

- 215 120 96 67

Ti 0.8 1.4 9

- 10 6

- 10 29

- 39 8.0 13 8.7

Zn 6.5 9.6 5

- 10 14

- 17 51

- 88 14 17 11

Cu 1.

- 0.15 4.9 18

- 32 8

- 16 21

- 58 9.3 1.7 2.1

Ni 2.4 1.8 0 0 2

- 3 0.99 1.0 0.72

Mn 0.21 2.3 3

- 4 3 28

- 32 2.2 2.5 2.1

Sb 0.12 0.30

Cd 0.12 0.14 4

- 6 4

- 7

V 0.05 0.16 1 1 3

- 5 2.5 2.7 1.8

La 0.00 0.11

Cs 0.00 0.00

Th 0.00 0.00

Sc 0.00 0.01

Ag 0.07 0.08

Co 0.02 0.07

Cr 1.2 2.6 2 1 3

- 8 <1.1 1.3 1.1

Si 198

- 464 338

- 672 387

- 401

Cl 7

- 32 3

- 9

Se 1 1

- 2

Rb 1 1

Sr 5

- 28 1

As 1 1 1

- 2

Mo

Br 2.1 1.3 1.3

a I: Indoor; Units: ng/m 3

b R: Residential; Units: ng/m 3

c Units: ng/m 3