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Chapter 4 Contents

Chapter 4. Exposure, Toxicokinetics, and Biomarkers _______________________________________________________ 4-1

4.1. Exposure Assessment________________________________________________________________________ 4-1

4.1.1. Pathways for Lead Exposure __________________________________________________________ 4-1

Figure 4-1. Conceptual model of multimedia Pb exposure. _________________________4-3

Table 4-1. Estimates of Pb measurements for EPA Region 5 from the NHEXAS 

study. _________________________________________________________4-4

4.1.1.1. Airborne Lead Exposure______________________________________________________ 4-4

Table 4-2. Estimates of fixed effects multivariate modeling of Pb levels measured 

during the NHEXAS-MD study _____________________________________4-5

Table 4-3. Comparison of personal, indoor, and outdoor Pb-bearing PM 

measurements from several studies. _________________________________4-7

4.1.1.2. Exposure to Lead in Soil and Dust ______________________________________________ 4-7

Table 4-4. Measurements of indoor Pb dust from various studies. ___________________4-8

4.1.1.3. Dietary Lead Exposure______________________________________________________ 4-10

Table 4-5. Pb bioaccumulation data for various plants.__________________________4-15

4.1.1.4. Occupational______________________________________________________________ 4-16

4.1.1.5. Exposure to Lead from Consumer Products ______________________________________ 4-17

Table 4-6. Pb content in various consumer products ____________________________4-18

4.2. Kinetics _________________________________________________________________________________ 4-19

4.2.1. Absorption _______________________________________________________________________ 4-19

4.2.1.1. Inhalation ________________________________________________________________ 4-19

4.2.1.2. Ingestion_________________________________________________________________ 4-21

Figure 4-2. Estimated relative bioavailability (RBA, compared to Pb-acetate) of 

ingested Pb in mineral groups, based on results from juvenile swine 

assays. _______________________________________________________4-24

4.2.2. Distribution ______________________________________________________________________ 4-26

4.2.2.1. Blood ___________________________________________________________________ 4-26

Figure 4-3. Plot of blood and plasma Pb concentrations in measured in adults and 

children. _____________________________________________________4-27

Figure 4-4. Relationship between Pb intake and blood Pb concentration in infants (n 

= 105, age 13 weeks, formula-fed). _________________________________4-28

Figure 4-5. Simulation of quasi-steady state blood and plasma Pb concentrations in 

a child (age 4 years) associated with varying Pb ingestion rates. 

Simulation based on ICRP Pb biokinetics model (Leggett, 1993).__________4-29

4.2.2.2. Bone ____________________________________________________________________ 4-30

4.2.2.3. Soft Tissues ______________________________________________________________ 4-31

4.2.2.4. Fetus ____________________________________________________________________ 4-32

4.2.2.5. Organic Lead _____________________________________________________________ 4-32

4.2.3. Elimination_______________________________________________________________________ 4-32

4.3. Lead Biomarkers __________________________________________________________________________ 4-34

4.3.1. Bone Lead Measurements ___________________________________________________________ 4-34

4.3.2. Blood Lead Measurements___________________________________________________________ 4-35

Figure 4-6. Simulation of temporal relationships between Pb exposure and blood Pb 

concentration in children. ________________________________________4-38

4.3.3. Urine Lead Measurements ___________________________________________________________ 4-39

Figure 4-7. Simulation of relationship between urinary Pb excretion and body 

burden in adults. _______________________________________________4-40

4.3.4. Lead in Other Potential Biomarkers ____________________________________________________ 4-41

4.3.4.1. Teeth____________________________________________________________________ 4-41

4.3.4.2. Hair_____________________________________________________________________ 4-41
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4.3.4.3. Saliva ___________________________________________________________________ 4-42

4.3.4.4. Serum δ-ALA and ALAD ___________________________________________________ 4-43

4.3.5. Relationship between Lead in Blood and Lead in Bone_____________________________________ 4-43

4.3.5.1. Children _________________________________________________________________ 4-44

Figure 4-8. Simulation of relationship between blood Pb concentration and body 

burden in children, with a constant Pb intake from age 2 to 5. ____________4-46

Figure 4-9. Simulation of relationship between time-integrated blood Pb 

concentration and cumulative Pb absorption in children. ________________4-47

4.3.5.2. Adults ___________________________________________________________________ 4-47

Figure 4-10. Simulation of relationship between blood Pb concentration, bone Pb 

and body burden in adults. _______________________________________4-49

4.3.6. Relationship Between Lead in Blood and Lead in Soft Tissues _______________________________ 4-52

Figure 4-11. Simulation of blood and soft tissue (including brain) Pb in children and 

adults who experience a period of increased Pb intake. Simulation based 

on ICRP Pb biokinetics model (Leggett, 1993).________________________4-53

Figure 4-12. Simulation of blood and brain Pb in children and adults who experience 

a period of increased Pb intake. ___________________________________4-55

4.3.7. Relationship Between Lead in Blood and Lead in Urine ____________________________________ 4-55

Figure 4-13. Top panel: Predicted relationship between plasma Pb concentration and 

urinary Pb excretion in an adult (age 40 years). Lower panel: 

Simulation of blood Pb, bone Pb and urinary excretion of Pb in an adult 

who experiences a period of increased Pb intake. Simulation based on 

ICRP Pb biokinetics model (Leggett, 1993). __________________________4-57

4.4. Observational Studies of Lead Exposure ________________________________________________________ 4-58

4.4.1. Lead in Blood_____________________________________________________________________ 4-58

Figure 4-14. Temporal trend in blood Pb concentration.___________________________4-59

Table 4-7. Blood Pb concentrations in the U.S. population _______________________4-60

Figure 4-15. Box plots of blood Pb levels among U.S. children (1-5 years old) from 

the NHANES survey, 1988-2008.___________________________________4-61

Figure 4-16. Percent distribution of blood Pb levels by race/ethnicity among U.S. 

children (1-5 years) from the NHANES survey, 1988-1991 (top) and 

1999-2004 (bottom). ____________________________________________4-62

Figure 4-17. a) Trends in 206Pb/204Pb isotope ratio in blood Pb and b) trends in blood 

Pb levels among Australian study populations during the period 1990-

2000. ________________________________________________________4-64

4.4.2. Lead in Bone _____________________________________________________________________ 4-64

Table 4-8. Epidemiologic studies that provide bone Pb measurements for nonoccupationally exposed populations ________________________________4-65

Table 4-9. Epidemiologic studies that provide bone Pb measurements for 

occupationally exposed populations ________________________________4-73

4.4.3. Lead in Urine _____________________________________________________________________ 4-76

Table 4-10. Urine Pb concentrations in the U.S. population _______________________4-77

4.4.4. Lead in Teeth _____________________________________________________________________ 4-77

Figure 4-18. Comparison of tooth enamel ______________________________________4-78

4.5. Empirical Models of Lead Exposure-Blood Lead Relationships ______________________________________ 4-78

4.5.1. Air Lead-Blood Lead Relationships____________________________________________________ 4-79

Table 4-11. Summary of estimated slopes for blood Pb to air Pb relationships in 

humans ______________________________________________________4-80

Figure 4-19. Predicted relationship between air Pb and blood Pb based on a meta 

analysis of 18 studies. ___________________________________________4-81

4.5.1.1. Children _________________________________________________________________ 4-81

Table 4-12. Environmental Pb levels and blood Pb levels in children in Trail, British 

Columbia_____________________________________________________4-82

Table 4-13. U.S. gasoline Pb consumption and air Pb levels _______________________4-84

Table 4-14. Air Pb levels and blood Pb levels in children in Mumbai, India ___________4-85

Figure 4-20. Predicted relationship between air Pb and blood Pb based on data from 

Chicago, IL (1974–1988). ________________________________________4-86
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4.5.1.2. Adults ___________________________________________________________________ 4-86

Table 4-15. Significant predictors of blood Pb concentration in bridge painters ________4-87

1. 4.5.2. ____________________________________Environmental Lead-Blood Lead Relationships4-87

Table 4-16. Linear model relating environmental Pb exposure and blood Pb 

concentration in childrena ________________________________________4-89

Figure 4-21. Predicted relationship between soil Pb concentration and blood Pb 

concentration in children based data collected in the New Orleans child 

Pb screening program (2000-2005) (Mielke et al., 2007).________________4-90

Table 4-17. General linear model relating blood Pb concentration in children and 

environmental Pb levels—Bunker Hill Superfund Site___________________4-92

4.6. Biokinetic Models of Lead Exposure-Blood Lead Relationships______________________________________ 4-92

4.7. Summary ________________________________________________________________________________ 4-93

4.7.1. Exposure_________________________________________________________________________ 4-93

4.7.2. Kinetics _________________________________________________________________________ 4-94

4.7.3. Lead Biomarkers __________________________________________________________________ 4-96

4.7.4. Air Lead-Blood Lead Relationships____________________________________________________ 4-97

Chapter 4. References ____________________________________________________________________________ 4-99
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Chapter 4. Exposure, Toxicokinetics, 

and Biomarkers

4.1. Exposure Assessment

1 The purpose of this section is to present recent studies that provide insight about human exposure 

2 to Pb through various pathways. The recent information provided here builds upon the conclusions of the 

3 2006 Pb AQCD (2006), which found that air Pb concentrations and blood Pb levels have decreased 

4 substantially following the 1996 ban on Pb in on-road vehicle gasoline, the 1978 ban on Pb in household 

5 paints, and the 1986 and 1995 restrictions on uses of Pb solder. At the same time, detectable quantities of 

6 Pb have still been observed to be bioaccessible in various media types. It was reported in the 2006 Pb 

7 AQCD (U.S. EPA, 2006) that airborne maximum quarterly Pb concentrations in the U.S. were in the 

range of 0.03-0.05 μg/m3 8 for non-source-oriented monitors for the years 2000-2004 and were 0.10-0.22 

μg/m3 9 for source-oriented monitors during that time period, while blood Pb levels reached a median of 

10 1.70 μg/dL among children ages 1-5 in 2001-2002. It was also observed that Pb exposures were 

11 associated with nearby industrial Pb sources, presence of Pb-based paint, and Pb deposited onto food in 

12 several of the studies described in the 2006 Pb AQCD. For the current review, Section 4.1. contains a 

13 description of studies related to pathways for human exposure to Pb. 

4.1.1. Pathways for Lead Exposure

14 Pathways of Pb exposure are difficult to assess because Pb has multiple sources in the environment 

15 and passes through various environmental media. These issues are described in detail in Sections 3.2 and 

16 3.3. Air-related pathways of Pb exposure are the focus of this ISA. Pb can be emitted to air or water. In 

17 addition to primary emission of particle-bound or gaseous Pb to the atmosphere, Pb can be resuspended to 

18 the air from soil or dust, and a fraction of that resuspended Pb may even originate from waters used to 

19 irrigate the soil. Additionally, Pb-bearing PM can be deposited from the air to soil or water through wet 

20 and dry deposition. In general, air-related pathways include those pathways where Pb passes through 

21 ambient air on its path from a source to human exposure. Air-related Pb exposures include inhalation and

22 ingestion of Pb-contaminated food, water or other materials including dust and soil. Non-air-related 

23 exposures include ingestion of indoor Pb paint, Pb in diet as a result of inadvertent additions during food 

 

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).
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1 processing, and Pb in drinking water attributable to Pb in distribution systems, as well as other generally 

2 less prevalent pathways. The complicated nature of Pb exposure is illustrated Figure 4-1, in which the 

3 Venn diagram depicts how Pb can cycle through multiple environmental media prior to human exposure. 

4 The “air/soil/water” arrows illustrate Pb exposures to plants, animals, and/or humans via contact with Pb5 containing media. The exposures are air-related if the Pb passed through the air compartment. When 

6 animals consume plant material exposed to Pb that has at some point passed through the air compartment, 

7 and when human diet includes animals and/or plants exposed to Pb that has passed through the air 

8 compartment, these are also considered air-related Pb exposures. As a result of the multitude of possible 

9 air-related exposure scenarios and the related difficulty of constructing Pb exposure histories, most 

10 studies of Pb exposure through air, water, and soil can be informative to this review. Figure 4-1 also 

11 illustrates other exposures, such as occupational exposures, contact with consumer goods in which Pb has 

12 been used, or ingestion of Pb in drinking water conveyed through Pb pipes. Most Pb biomarker studies do 

13 not indicate speciation or isotopic signature, and so exposures that are not related to Pb in ambient air are 

14 also reviewed in this section because they can contribute to Pb body burden. Many of the studies 

15 presented in the subsequent material focus on observations of Pb exposure via one medium: air, water, 

16 soil and dust, diet, or occupation.
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Figure 4-1. Conceptual model of multimedia Pb exposure.The Venn 

diagram is used to illustrate the passage of Pb through 

multiple environmental media compartments through which 

exposure can occur. 

1 The relative importance of different sources or pathways of potential exposure to Pb in the 

2 environment is often difficult to discern. Individual factors such as home environment, location, and 

3 susceptibility factors (described in more detail in Chapter 6) may influence exposures. The National 

4 Human Exposure Assessment Survey (NHEXAS) study sampled Pb, as well as other pollutants and 

5 VOCs, in multiple exposure media from subjects across six states in EPA Region 5 (Illinois, Indiana, 

6 Michigan, Minnesota, Ohio, and Wisconsin) (Clayton et al., 1999) as well as in Arizona (O'Rourke et al., 

7 1999) and Maryland (Egeghy et al., 2005). Results from NHEXAS indicate that personal exposure 

8 concentrations of Pb are higher than indoor or outdoor concentrations of Pb (Table 4-1). It is plausible 

9 that local resuspension of Pb-containing dust due to human activity increased personal exposure 

10 concentrations of airborne Pb relative to indoor or outdoor air Pb concentrations. Pb levels in windowsill 

11 dust were higher than Pb levels in surface dust collected from other surfaces. Clayton et al. (1999)

12 suggested that higher windowsill levels could be attributed to the presence of Pb-based paint and/or to 

13 accumulation of infiltrated outdoor Pb-bearing PM. Pb levels in food were higher than in beverages, and 

14 Pb levels in standing tap water (also referred to as “first flush” or “first draw”) were higher than Pb levels 

15 obtained after allowing water to run for three minutes to flush out pipes.
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Table 4-1. Estimates of Pb measurements for EPA Region 5 from the NHEXAS study.

Mediuma n Percentage 

measurableb (CLs)c Mean (CLs)c 50th (CLs)c 90th (CLs)c

Personal air (ng/m3)

d 167 81.6 (71.3; 92.0) 26.83 (17.60; 36.06) 13.01 (11.13; 18.13) 57.20 (31.18; 85.10)

Indoor air (ng/m3)

d 213 49.8 (37.2; 62.3) 14.37 (8.76; 19.98) 6.61 (4.99; 8.15) 18.50 (12.69; 30.31)

Outdoor air (ng/m3)

d 87 73.8 (56.3; 91.3) 11.32 (8.16; 14.47) 8.50 (7.14; 10.35) 20.36 (12.60; 34.91)

Surface dust (ng/cm2) 245 92.1 (87.4; 96.8) 514.43 (-336.6; 1365.5) 5.96 (3.37; 10.94) 84.23 (26.52; 442.63)

Surface dust (mg/kg) 244 92.1 (87.4; 96.8) 463.09 (188.15; 738.04) 120.12 (83.85; 160.59) 698.92 (411.84; 1,062.8)

Window sill dust (ng/cm2) 239 95.8 (92.5; 99.0) 1,822.6 (481.49; 3,163.6) 16.76 (10.44; 39.41) 439.73 (106.34; 4,436.2)

Window sill dust (mg/kg) 239 95.8 (92.5; 99.0) 954.07 (506.70; 1,401.4) 191.43 (140.48; 256.65) 1,842.8 (1,151.3; 2,782.5)

Standing tap water (μg/L) 444 98.8 (97.6; 100.0) 3.92 (3.06; 4.79) 1.92 (1.49; 2.74) 9.34 (7.87; 12.35)

Flushed tap water (μg/L) 443 78.7 (70.7; 86.7) 0.84 (0.60; 1.07) 0.33 (0.23; 0.49) 1.85 (1.21; 3.04)

Solid food (μg/kg) 159 100.0 (100.0; 100.0) 10.47 (6.87; 14.07) 6.88 (6.44; 8.04) 14.88 (10.78; 19.08)

Beverages (μg/kg) 160 91.5 (85.2; 97.8) 1.42 (1.13; 1.72) 0.99 (0.84; 1.21) 2.47 (2.06; 3.59)

Food+Beverages (μg/kg) 156 100.0 (100.0; 100.0) 4.48 (2.94; 6.02) 3.10 (2.66; 3.52) 6.37 (4.89; 8.00)

Food intake (μg/day) 159 100.0 (100.0; 100.0) 7.96 (4.25; 11.68) 4.56 (3.68; 5.36) 12.61 (9.27; 16.38)

Beverage intake (μg/day) 160 91.5 (85.2; 97.8) 2.15 (1.66; 2.64) 1.41 (1.18; 1.60) 4.45 (3.15; 5.65)

Food+Beverage intake 

(μg/day) 156 100.0 (100.0; 100.0) 10.20 (6.52; 13.89) 6.40 (5.21; 7.78) 16.05 (13.31; 18.85)

Blood (μg/dL) 165 94.2 (88.2; 100.0) 2.18 (1.78; 2.58) 1.61 (1.41; 2.17) 4.05 (3.24; 5.18)

Note: EPA Region 5 includes six states: Illinois, Indiana, Ohio, Michigan, Minnesota, and Wisconsin. Participants were enrolled using a stratified, 

four-stage probability sampling design, and submitted questionnaire and physical measurements data. Summary statistics (percentage

measurable, mean, median, 90th percentile) were computed using weighted sample data analysis. The estimates apply to the larger Region 5 

target population (all non-institutionalized residents residing in households).

a

Estimates for indoor air, outdoor air, dust media, and water media apply to the target population of Region 5 households; estimates for other media 

apply to the target population of Region 5 residents. 

b

Percentage measurable is the percentage of the target population of residents (or households) estimated to have Pb levels above limit of 

detection (LOD).

c

The lower and upper bounds of the 95% confidence limits (CL) are provided.

d

PM50.

Source: Used with permission from Nature Publishing Group, Clayton et al. (1999)

4.1.1.1. Airborne Lead Exposure

1 Limited personal exposure monitoring data for airborne Pb were available for the 2006 AQCD 

2 (U.S. EPA, 2006). As described above, the NHEXAS study showed personal air Pb concentrations to be 

3 significantly higher than indoor or outdoor air Pb concentrations (Clayton et al., 1999). Indoor air Pb 

4 concentration was shown to be moderately correlated with floor dust and residential yard soil Pb 

5 concentration (M. Rabinowitz et al., 1985). Egeghy et al. (2005) performed multivariate fixed effects 

6 analysis of the NHEXAS-Maryland data and found that Pb levels measured in indoor air were 

7 significantly associated with log-transformed outdoor air Pb levels, ambient temperature, number of hours 

8 in which windows were open, homes built before 1950, and frequency of fireplace usage (Table 4-2).
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Table 4-2. Estimates of fixed effects multivariate modeling of Pb levels measured during 

the NHEXAS-MD study

Pb in Indoor Air Pb in Dust Dermal Pb Blood Pb

Fixed Effect βa p-value βa p-value βa p-value βa p-value

Intercept -0.50 0.0051 6.22 <0.0001 6.23 <0.0001 0.02 0.91

Outdoor Pb concentrationb 0.51 <0.0001

Average weekly temperature (°F) 0.01 0.046

Open window periods (hr) 0.01 0.035 -0.03 0.0082

House pets (yes) -0.15 0.078

Air filter use (yes) -0.28 0.087 -0.12 0.088

Home age (<1950) 0.25 0.025 0.96 0.029

Fireplace (frequency of use) 0.11 0.045 0.46 0.0054

Pb concentration in soilb 0.27 0.037

Interior Pb paint chipping/peeling 

(yes) 0.43 0.091

Cement at primary entryway (yes) 1.97 0.0064

Indoor pesticide usage last 6 mo

(yes) -0.78 0.0003

Electrostatic air filter usage (yes) -0.91 0.062

Sex of participants (male) 0.41 0.0012 0.43 <0.0001

Ethnic minority participants (yes) 0.41 0.0063

Washing hands after lawn mowing

(no) 1.04 0.0010

Gasoline power- equipment usage 

(yes) 0.61 0.0072

Bathing or showering activities (yes) -0.43 0.019

Dust level indoors (scale: 1-3) 0.22 0.019

Residing near commercial areas 

(yes) 0.32 0.0087

Age of participants (yr) 0.02 <0.0001

Number cigarettes smoked (count) 0.03 <0.0001

Burning wood or trash (days) 0.58 0.0099

Showering frequency (avg # days) -0.29 0.0064

Work outside home (yes) -0.26 <0.0001

Health status (good) 0.23 0.0009

Adherence to high fiber diet (yes) -0.15 0.040

Gas or charcoal grill usage (yes) -0.17 0.0002

a

Estimates of fixed effects in final multiple regression analysis models for Pb in the Maryland investigation data in the National Human Exposure 

Assessment Survey (NHEXAS-MD) b

Log transform

Source: Used with permission from Nature Publishing Group, Egeghy et al. (2005).

1 Some recent studies have measured personal exposure to particle-bound Pb along with other trace 

2 metals. Adgate et al. (2007) measured the concentrations of several trace elements in personal, indoor, and 

3 outdoor samples of PM2.5 and found that average personal Pb-PM2.5 concentration was roughly three 

4 times higher than outdoor Pb-PM2.5 concentration and two times higher than indoor Pb-PM2.5

5 concentration (Table 4-3). Another study of indoor and outdoor air concentrations of Pb was carried out 

6 by Molnar et al. (2007). PM2.5 trace element concentrations were determined in homes, preschools and 

7 schools in Stockholm, Sweden. In all sampled locations, Pb-PM2.5 concentrations were higher in the 

8 outdoor environment than in the proximal indoor environment. The indoor/outdoor ratios for Pb-PM2.5

9 suggest an outdoor Pb-PM2.5 net infiltration of ~0.6 for these buildings. Outdoor air Pb concentrations did 

10 not differ between the central and more rural locations. Indoor air Pb concentrations were higher in spring 

11 than in winter, which the authors attributed to greater resuspension of elements that had accumulated in 

12 road dust over the winter period and increased roadwear on days with dry surfaces. In a pilot study in 

13 Windsor, Ontario, Rasmussen et al. (2007) observed that personal exposure to Pb measured using a PM2.5
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1 monitor was roughly 40% higher than outdoor Pb concentration and 150% higher than indoor Pb 

2 concentration. Pekey et al. (2010) measured indoor and outdoor trace element composition of PM2.5 and 

3 PM10 in Kocaeli, an industrial region of Turkey, and found that average airborne Pb concentrations were 

4 higher outdoors than indoors for both PM2.5 and PM10 during summer and for PM10 during winter, but that 

5 indoor Pb concentration was higher than outdoor Pb concentration for PM2.5 during winter. The indoor-to6 outdoor ratio of Pb in PM varied by environment; it tended to be less than one, but the ratio varied from 

7 one microenvironment to another. The three studies that included personal samples recorded 

8 measurements that were consistently higher than indoor or outdoor levels. It is likely that a number of 

9 factors influenced the indoor-to-outdoor ratio of Pb in PM for these studies. These factors may have 

10 included seasonal air exchange, which can vary as a function of window opening or air conditioning 

11 usage, prevalence and strength of outdoor and indoor Pb sources, and size distribution of airborne Pb12 bearing PM.

13 Several of the studies can be used to develop an understanding of how personal exposure to PM14 bound Pb varies with other exposures. Molnar et al. (2007) reported Spearman correlations of Pb with 

15 PM2.5 and NO2 in three outdoor microenvironments (residence, school, and preschool) and found that Pb 

16 and other trace metals were generally well correlated with PM2.5 (r = 0.72-0.85), but Pb was not always 

17 well-correlated with NO2 (r = 0.24-0.75). In the case where Pb and NO2 were well-correlated, it is 

18 possible that the Pb was traffic related from resuspended pulverized wheel weights or impurities in 

19 unleaded on-road gasoline. For the other two sites where the correlation between Pb and NO2 was low, it 

20 is possible that they were less affected by traffic. Table 3A-14 in the Appendix illustrates that Pb 

21 exposures are typically well below the level of the NAAQS. The higher exposures occurred in a heavily 

22 industrialized area with an incinerator and several industrial facilities including metal processing, cement, 

23 petroleum refining, agriculture processing. Otherwise, exposures were all between 0.002 and 0.006 

μg/m3 24 . The proportion of Pb compared with other trace metals varied with location and component. It was 

25 typically several times lower than S as well as crustal elements such as Ca and Fe. In the industrial area of 

26 Kocaeli, Pb comprised a greater proportion of the PM compared with other areas.
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Table 4-3. Comparison of personal, indoor, and outdoor Pb-bearing PM measurements 

from several studies.

Study Location Pb Metric Sampling 

Period Personal Pb Indoor Pb Outdoor Pb

Clayton et al. 

(1999)

IL, IN, MI, MN, 

OH, WI

Med. Pb-PM50

(ng/m3) 

July, 1995- May, 199713 6.6 8.5

Adgate et al. 

(2007)

Minneapolis-St. 

Paul, MN

Avg. Pb-PM2.5

(ng/m3)

Spring, 

Summer, Fall, 

1999

6.2 3.4 2.0

Molnar et al. 

(2007)

Stockholm, 

Sweden

Avg. Pb-PM2.5

(ng/m3)

December, 

2003-July, 

2004

Homes: 3.4

Schools: 2.5

Preschools: 1.8

Homes: 4.5

Schools: 4.6

Preschools: 2.6

Tovalin-Ahumada 

et al. (2007)

Mexico City, 

Mexico

Med. Pb-PM2.5

(ng/m3)

April-May, 

2002

26 56

Puebla, Mexico Med. Pb-PM2.5

(ng/m3)

April-May, 

2002

4 4

Rasmussen et al. 

(2007)

Windsor, Ontario, 

Canada

Med. Pb-PM2.5

(mg/kg)

April, 2004 311 124 221

Pekey et al. (2010) Kocaeli, Turkey Avg. Pb-PM2.5

(ng/m3)

May-June, 

2006, 

December, 

2006-January 

2007

Summer: 34

Winter: 85

Summer: 47

Winter: 72

Avg. Pb-PM10

(ng/m3)

May-June, 

2006, 

December, 

2006-January 

2007

Summer: 57

Winter: 125

Summer: 78

Winter: 159

4.1.1.2. Exposure to Lead in Soil and Dust

1 The 2006 AQCD (U.S. EPA, 2006) lists indoor Pb dust infiltrated from outdoors as a potential 

2 source of exposure to Pb soil and dust. Outdoor soil Pb concentration may present a direct inhalation 

3 exposure, or it can be tracked into homes to result in exposure to resuspended Pb PM or to Pb dust during 

4 hand-to-mouth contact. A detailed description of studies of outdoor Pb concentration is provided in 

5 section 3.6.1. Indoor measurements can reflect infiltrated Pb as well as Pb dust derived from debrided 

6 paint, consumer products, or soil that has been transported into the home via foot traffic. Table 4-4 

7 presents indoor Pb levels for several studies.
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Table 4-4. Measurements of indoor Pb dust from various studies.

Reference Study Location Metric (units) Sample Site Indoor Pb 

Concentration

Caravanos et al. 

(2006) New York City, New York Weekly dust loading (μg/m2) Glass plate Median: 52

Khoder et al. (2010) Giza, Egypt Weekly dust loading 

(μg/m2) Glass plate Median: 408

Yu et al. (2006) Syracuse, New York Dust concentration 

range (mg/kg) Floor Range: 209-1770

Turner and 

Simmonds (2006)

Birmingham, Plymouth, 

and 2 rural sites, UK

Dust concentration 

(mg/kg) Floor Median: 178

Gaitens et al. (2009) U.S. (nationwide) Dust loading (μg/m2) Smooth floor Median: 1.7

Avg.: 4.4

Rough floor Median: 5.6

Avg.: 16

Smooth windowsill Median: 2.5

Avg.: 190

Rough windowsill Median: 55

Avg.: 480

Mielke et al. (2001) New Orleans, Louisiana Dust concentration 

(mg/kg)

Multiple locations within homes 

prepared for painting; sanded house

Range: <3-28,000

Multiple locations within homes 

prepared for painting; scraped 

house

Range: 7-1,200

Spalinger et al. 

(2007) Rural towns, Idaho Dust concentration (mg/kg) Vacuum Median: 120Max: 830

Floor Median: 95

Max: 1,300

Bunker Hill, Idaho 

Superfund site Vacuum Median: 470Max: 2,000

Floor Median: 290

Max: 4,600

1 Several studies have demonstrated the infiltration of Pb dust into buildings. For example, 

2 Caravanos et al. (2006) collected dust on glass plates at an interior location near an open window, a 

3 sheltered exterior location, and an open exterior location for a two-year period in Manhattan, NY. Median 

weekly dust loading was reported to be 52 μg/m2 for the indoor site, 153 μg/m2 4 for the unsheltered 

outdoor site, and 347 μg/m2 5 for the sheltered outdoor site. This paper demonstrated the likely role of 

6 outdoor Pb in influencing indoor dust Pb loading and indicated that under quiescent conditions (e.g., no 

cleaning), the indoor dust Pb level might exceed EPA’s hazard level for interior floor dust of 430 μg/m2 7

(40 μg/ft2 8 ). Khoder et al. (2010) used the same methodology to study Pb dust deposition in Giza, Egypt 

and found a median weekly deposition rate of 408 μg/m2 9 and an exterior median deposition rate of 2,600 

μg/m2 10 . In the latter study, Pb deposition rate correlated with total dust deposition rate (R=0.92), Cd 

11 deposition rate (R=0.95), and Ni deposition rate (R=0.90). Statistically significant differences in Pb 

12 deposition rates were observed between old and new homes (p<0.01) in the Khoder et al. (2010) study, 

13 although the only quantitative information provided regarding home age stated that the oldest home was 

14 22 years old when the study was performed in 2007. Khoder et al. (2010) found no statistically significant 

15 difference between Pb loadings when segregating the data by proximity to roadways.

16 Residual Pb dust contamination following cleanup has been documented. For instance, Hunt et al. 

17 (2008) performed tests where moderately elevated soil Pb was tracked onto a tile test surface and then 

18 repeatedly cleaned with a moistened wipe and/or vacuumed until visual inspection of the tiles uncovered 
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1 no surface discoloration. The authors then used wet wipe samples to collect residual soil and estimate Pb 

2 deposition and concentration. Elevated Pb loadings and concentrations were observed even after multiple 

3 cleanings. Scanning electron microscopy (SEM) of the wipe samples revealed that most of the residual 

4 dust particles were in the range of 1-3 μm in area equivalent diameter. This indicates that Pb-bearing fine 

5 particles are not well captured by home cleaning. Johnson et al. (2009) surveyed the floors of 488 homes 

6 in Syracuse, NY census tracts and found that the variability in Pb dust within homes was greater than the 

7 variability between homes. A correlation between Pb dust loading on floors and the fraction of homes in 

census tracts that were renter-occupied (partial R2 8 = 0.48; where total number of homes is the sum of 

9 owner-occupied, renter-occupied, and vacant homes) was also observed in this study. Yu et al. (2006)

10 dissolved Pb dust, obtained from vacuuming carpet samples and found that Pb concentration in carpet 

11 ranged from 209 to 1,770 mg/kg dust.

12 Pb dust on floors, windowsills, and other accessible surfaces are potential exposure sources to 

13 small children who use touch to explore their environments. Gaitens et al. (2009) used National Health 

14 and Nutrition Examination Survey (NHANES) data from 1999 through 2004 to examine Pb dust in homes 

of children ages 12-60 months. The median value of Pb dust on floors was reported to be 1.7 μg/m2 15

(mean: 4.4 μg/m2 16 ), with floors that were not smooth and cleanable having a median Pb dust value of 5.6 

μg/m2 (mean: 16 μg/m2 17 ). Floor Pb dust level was modeled against several survey covariates and was 

18 significantly associated (p <0.05) with floor surface condition, windowsill Pb dust loading, race and 

19 ethnicity, poverty-to-income ratio, year of home construction, presence of smokers in the home, and year 

20 of survey. It was nearly significantly associated (p = 0.056) with renovations made to pre-1950 homes. 

Median Pb dust on smooth windowsills was 25 μg/m2 (mean: 190 μg/m2 21 ). When windowsills were not 

smooth, the median Pb dust level was 55 μg/m2 (mean: 480 μg/m2 22 ). Windowsill Pb dust level was also 

23 found to be significantly associated (p <0.05) with race and ethnicity, year of home construction, window 

24 surface condition, presence of smokers in the home, deterioration of indoor paint, and year of survey. It 

25 was nearly significantly associated (p = 0.076) with deterioration of outdoor paint when homes were built 

26 prior to 1950. Dust Pb levels were found by Egeghy et al. (2005) to be significantly associated with the 

27 log-transform of soil Pb levels, cement content in the home entryway, indoor pesticide use, frequency of 

28 fireplace usage, number of hours in which windows were open, and homes built before 1950 (Table 4-2).

29 Building demolition and renovation activities can create dust from interior and exterior paints with 

30 Pb content. Mielke and Gonzales (2008) measured Pb content in paint chips from paint applied prior to 

31 1992 and found that median Pb levels were 420 mg/kg for interior paint and 77,000 mg/kg for exterior 

32 paint. Maximum levels were 63,000 mg/kg and 120,000 mg/kg for interior and exterior paint, 

33 respectively. Mielke et al. (2001) compared dust samples from two New Orleans houses that were 

34 prepared for painting. One home was power sanded, while the other was hand-scraped. Immediately after 

35 sanding, Pb dust samples ranged from <3 to 28,000 mg/kg at the sanded house. Pb dust samples from the 

36 scraped house ranged from 7 to 1,200 mg/kg. Pb in dust or paint samples was not quantified.
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1 Dust Pb concentrations have also been reported for homes in the vicinity of historical metals 

mining and smelting sources. For example, Spalinger et al. (2007) measured Pb dust in homes in a 34 km2 2

3 area surrounding a designated Superfund site where formerly a Pb and Zn smelter operated at Bunker 

Hill, ID. Vacuum and floor mat samples were taken from homes in three towns within the 34 km2 4 area and 

5 five “background” towns further from the Superfund site. For the background towns, Pb concentration in 

6 vacuum dust had a median of 120 mg/kg and a maximum of 830 mg/kg, and Pb concentration in floor 

7 dust had a median of 95 mg/kg and a maximum of 1,300 mg/kg. The median Pb dust loading rate was 

measured to be 40 μg/m2 8 per day. Among the background homes, median vacuum and floor mat Pb dust 

9 samples were 3 and 2.5 times higher, respectively, when comparing homes built before 1960 with those 

10 built after 1960. Deposition rate of Pb dust was 5 times higher in the older homes. In contrast, Pb in 

11 vacuum dust and floor mats for the towns contained within the Bunker Hill Superfund site had a median 

12 of 470 mg/kg with a maximum of 2,000 mg/kg and a median of 290 mg/kg with a maximum of 4,600 

mg/kg, respectively. The median Pb loading rate for these towns was 300 μg/m2 13 per day, and the 

maximum Pb dust loading rate was 51,000 μg/m2 14 per day. These results suggest that those living in close 

15 proximity to an industrial site are at much greater risk of exposure to Pb dust compared to the general 

16 population.

4.1.1.3. Dietary Lead Exposure

17 This subsection covers several dietary Pb exposures from a diverse set of sources. Included among 

18 those are drinking water, fish and meat, pesticides via vegetables, urban gardening, dietary supplements, 

19 tobacco, cultural food sources, and breastfeeding.

Drinking Water

20 Differences in sources and transport of drinking water may cause variation in Pb levels. For 

21 example, Shotyk and Krachler (2009) measured the Pb concentration in tap water, commercially bottled 

22 tap water and bottled natural water. They found that, in many cases, tap water contained less Pb than 

23 bottled water. Excluding bottled water in glass containers that have higher Pb concentrations due to 

24 leaching from the glass, the median Pb concentration was 8.5 ng/L (range ≤ 1 to 761 ng/L). This was 

25 significantly less than the EU, Health Canada and WHO drinking water maximum admissible 

26 concentration of 10 μg/L.It is now recognized that environmental nanoparticles (NPs) (~1-100 nm) can 

27 play a key role in determining the chemical characteristics of engineered as well as natural waters

28 (Wigginton et al., 2007). An important question is whether or not NPs from source waters affect the 

29 quality of drinking water. For example, if Fe-oxide NPs are not removed during the 

30 flocculation/coagulation stage of the treatment process, they may become effective transporters of 
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1 contaminants such as Pb, particularly if these contaminants are leached from piping in the distribution 

2 system. Edwards and Dudi (2004) observed a red-brown particle-bound Pb in Washington, DC water that 

3 could be confused with innocuous Fe. The source of the particle-bound Pb was not known but was 

4 thought to originate from the source water.

5 Corrosion byproducts can lead to Pb exposures in drinking water. Schock et al. (2008)

6 characterized Pb pipe scales from 91 pipes made available from 26 different municipal water systems 

7 from across the northern U.S. They found a wide range of elements including Cu, Zn and V as well as Al, 

8 Fe and Mn. Interestingly, V was present at nearly one percent levels in pipes from many geographically 

9 diverse systems. In a separate study, Gerke et al. (2009) identified the corrosion product, vanadinite 

10 (Pb5(VO4)3Cl) in Pb pipe corrosion byproducts collected from 15 Pb or Pb-lined pipes representing 8

11 different municipal drinking water distribution systems in the Northeastern and Midwest regions of the

12 U.S. Vanadinite was most frequently found in the surface layers of the corrosion products. The vanadate 

ion, VO4

3- 13 , essentially replaces the phosphate ion in pyromorphite and hydroxyapatite structures. It is not 

14 known whether the application of orthophosphate as a corrosion inhibitor would destabilize vanadinite,

15 but this would have implications for V release into drinking water. The stability of vanadinite in the 

16 presence of monochloramine is also not known, and this might have implications for both Pb and V 

17 release into drinking water.

18 In recent years, drinking water treatment plants in many municipalities have switched from using 

19 chlorine to other disinfecting agents because their disinfection byproducts may be less carcinogenic. 

20 However, chloramines are more acidic than chlorine and can increase Pb solubility (Raab et al., 1991) and 

21 increase Pb concentrations in tap water. For example, after observing elevated Pb concentrations in 

22 drinking water samples, Kim and Herrera (2010) observed Pb oxide corrosion scales potentially occurring

23 after using acidic alum as a disinfection agent. High Pb concentrations in Washington, DC drinking water 

24 were attributed to leaching of Pb from Pb-bearing pipes promoted by breakdown products of disinfection 

25 agents (Edwards & Dudi, 2004). Maas et al. (2007) tested the effect of fluoridation and chlorine-based 

26 (chlorine and chloramines) disinfection agents on Pb leaching from plumbing soldered with Pb. When 

27 using chlorine disinfection agents alone, the Pb concentration in water samples doubled during the first 

28 week of application (from 100 to 200 ppb) but then decreased over time. When adding fluorosilic acid 

29 and ammonia, the Pb concentration spiked to 900 ppb and increased further over time. Similarly, Lasheen 

30 et al. (2008) observed leaching from pipes in Egypt. In this study, the authors tested polyvinyl chloride 

31 (PVC), polypropylene (PP), and galvanized iron pipes and observed leaching from both the PVC and PP 

32 pipes when exposed to an acid of pH = 6, with PVC having greatest amount of leaching. Exposure to 

33 basic solutions actually resulted in reduction of Pb concentration in the drinking water. 

34 Miranda et al. (2007) modeled blood Pb levels among children living in Wayne County, NC as a 

35 function of household age, use of chloramines and other covariates. It was found that blood Pb level was 

36 significantly associated with the year the home was built (p <0.001), use of chloramines (p <0.001), and 
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1 the interaction between these two variables (p <0.001). When year in which the home was built was 

2 broken into categories for the independent variables and interaction terms, Miranda et al. (2007) found 

3 that significance increased with the age of the home, based on the assumption that older homes will have 

4 more Pb pipes and Pb solder connecting the pipes. However, the study did not control for the presence of 

5 Pb paint in the dwellings, so it is difficult to distinguish the effect of Pb pipes from the presence of paint 

6 from that variable.

7 Several chemical mechanisms may contribute to release of Pb during use of chloramine 

8 disinfection agents. Edwards and Dudi (2004) hypothesized that Pb leaching through chloramines 

9 exposure through the breakdown of brass alloys and solder containing Pb. They also proposed that

10 chloramines may trigger nitrification and hence cause decreasing pH, alkalinity and dissolved oxygen that 

11 lead to corrosion after observing that nitrification also leads to increased Pb concentrations in water. 

12 However, Zhang et al. (2009) found no evidence that nitrification brought about significant leaching of Pb 

13 from Pb pipes. Lytle et al. (2009) suggested that a lack of increased Pb(II) concentrations in drinking 

14 water following a change from free chlorine to chloramines disinfection is attributed to the formation of 

15 the Pb(II) mineral hydroxypyromorphite (Pb5(PO4)3OH) instead of Pb(IV) oxide. Xie et al. (2010) further 

16 investigated the mechanisms by which Pb(II) release is affected by chloramines. Two opposing 

17 mechanisms were proposed: Pb(IV)O2 reduction by an intermediate species from decomposition of 

18 monochloramine; and increasing redox potential which decreases the thermodynamic driving force for 

19 reduction. They suggest that the contact time of monochloramine with PbO2 and the Cl2:N ratio in 

20 monochloramine formation will determine which mechanism is more important. Free chlorine can control 

21 Pb concentrations from dissolution under flowing conditions but for long stagnation periods, Pb 

22 concentrations can exceed the action level: 4-10 days were required for Pb concentrations to exceed 15 

23 μg/L (for relatively high loadings of PbO2 of 1 g/L). Thus, under less extreme conditions, it was 

24 concluded that chloramination was unlikely to have a major effect on the release of Pb into drinking 

25 water.

Agriculture

26 Dietary Pb has the potential to emanate from soil Pb used for agricultural purposes. For example, 

27 Jin et al. (2005) tested soil Pb, bioaccessibility of soil Pb (determined by CaCl2 extraction), and Pb in tea 

28 samples from tea gardens. They observed that the Pb concentration in tea leaves was proportional to the 

29 bioaccessible Pb in soil. Fernandez et al. (2007; 2010; 2008) measured Pb from atmospheric deposition in 

30 two adjacent plots of land having the same soil composition but different uses: one was pasture land and 

31 one was agricultural. In the arable land, size distributions of soil particle-bound Pb, were uniformly 

32 distributed. In pasture land, size distributions of Pb were distributed bimodally with peaks around 2-20 

33 μm and 50-100 μm (Fernandez et al., 2010). For the agricultural plot, Pb concentration was constant 
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1 around 70 mg/kg in samples taken over the first 30 cm of soil, at which time it dropped below 10 mg/kg 

2 at soil depths between 35 and 100 cm. In contrast, Pb concentration in pasture land peaked at a depth of 

3 10 cm at a concentration of roughly 70 mg/kg and then dropped off gradually to approach zero 

4 concentration at a depth of approximately 50 cm. The sharp change in concentration for the arable land 

5 was attributed to a combination of plowing the soil and use of fertilizers to change the acidity of the soil 

6 and hence the bioaccessibility of the Pb within the soil (Fernandez et al., 2007). They found that the 

7 surface layer was acidic (pH: 3.37-4.09), as was the subsurface layer (pH: 3.65-4.38). 

8 There is some evidence that Pb contamination of crops can originate with treatment of crops. For 

9 example, compost produced from wastewater sludge has the potential to add Pb to crops. Cai et al. (2007)

10 demonstrated that production of compost from sludge enriched the Pb content by 15-43% prior to its 

11 application. Chen et al. (2008) observed that the median concentration of Pb in California crop soil 

12 samples was 16.2 mg/kg (range: 6.0-62.2 mg/kg). Factor analysis suggested that the soil was enriched 

13 with Pb in crop soils in the Oxnard/Ventura region. Chen et al. (2008) further observed that in three of the 

14 seven California agricultural regions sampled, concentrations of Pb increased following addition of 

15 fertilizer, but sub-proportionally to the increase in P and Zn indicators of fertilizer. In four regions, there 

16 was no increase of Pb at all. Furthermore, Tu et al. (2000) observed a decrease in Pb fraction with 

17 increasing P application. Nziguheba and Smolders (2008) also surveyed phosphate-based fertilizers sold 

18 in European markets to determine the contribution of these fertilizers to heavy metal concentrations in 

19 agricultural products. They observed a median Pb concentration of 2.1 mg/kg based on total weight of the 

20 fertilizer, with a 95th percentile concentration of 7.5 mg/kg. Across Europe, Nziguheba and Smolders 

21 (2008) observed that the amount of Pb applied via fertilizers was only 2.6% of that from atmospheric 

22 deposition. Although Pb in on-road vehicle gasoline has been phased out in the U.S., this remains a 

23 relevant issue in the U.S. because some imported crops that are produced in countries that still use Pb 

24 antiknock agents in on-road gasoline. For example, high concentrations of Pb have been found in 

25 chocolate from beans grown in Nigeria, where leaded gasoline is legal. Rankin et al. (2005) observed that 

the ratios of 207Pb to 206Pb and 208Pb to 207 26 Pb were found to be similar to those of Pb in gasoline. Although 

27 this study showed that Pb concentration in the shelled cocoa beans was low (~1 ng/g), manufactured 

28 cocoa powder and baking chocolate was observed to have Pb concentrations similar to those of the cocoa 

29 bean shells, on the order of 200 ng/g, and Pb concentration in chocolate products was roughly 50 ng/g 

30 (Rankin et al., 2005). It is possible that the increases were attributed to contamination of the cocoa by the 

31 shells during storage or manufacture, but the authors note that more research is needed to verify the 

32 source of contamination. Likewise, it is possible that resuspended Pb that originated from legacy mobile 

33 and industrial sources could deposit on crops. 

34 Uptake of Pb by plants growing in contaminated soil has been demonstrated in some species during 

35 controlled potted plant experiments (Del Río-Celestino et al., 2006). In this study, most species retained 

36 Pb in the roots with little mobilization to the shoots of the plants. However, certain species Cichorium 
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1 intybus [chicory], Cynodon dactylon [Bermuda grass], Amaranthus blitoides [matweed or mat amaranth], 

2 and Silybum marianum [milk thistle]) were able to mobilize Pb from the roots to the shoots of the plant; 

3 these specific species could lead to human exposures through consumption of grazing animals. Lima et al. 

4 (2009) conducted similar greenhouse experiments with several vegetable crops grown in soil 

5 contaminated by Pb-containing residue from battery recycling waste. In this study, carrots were 

6 demonstrated to have high bioaccumulation, measured as the percent of Pb concentration measured in the 

7 plant compared with the Pb concentration in the soil, with little translocation of the Pb to the shoots, 

8 measured as the percent of Pb mass in the shoots compared to the Pb mass within the entire plant, of the 

9 Pb to the shoots. Conversely, beets, cabbages, sweet peppers, and collard greens had low bioaccumulation 

10 but moderate to high translocation. Okra, tomatoes, and eggplants had moderate bioaccumulation and 

11 moderate to high translocation. Sesli et al. (2008) also noted uptake of Pb within wild mushrooms. 

12 Vandenhove et al. (2009) compiled bioaccumulation data for plant groupings from various references; 

13 these data are reproduced in Table 4-5. Based on this review, grasses had the highest uptake, followed by 

14 leafy vegetables and root crops grown in sandy soils; these references also suggested high transfer from 

15 roots to shoots among root crops, with shoots having roughly four times higher Pb bioaccumulation than 

16 roots.
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Table 4-5. Pb bioaccumulation data for various plants.Bioaccumulation is expressed as 

percent of Pb concentration in the plant to the Pb concentration in 

the soil. 

Plant Group Plant 

Compartment Soil n GM GSD AM SD Min Max

All 210 2.0% 14 63% 290% 0.015% 2500%

Cereals Grain All 9 1.0% 3.6 1.8% 1.6% 0.19% 4.8%

Straw All 4 2.3% 3.5 3.8% 4.0% 0.51% 9.6%

Maize Grain All 9 0.12% 2.3 0.17% 0.14% 0.052% 0.38%

Straw All 3 0.28% 6.6 0.85% 1.3% 0.060% 2.3%

Rice Grain All 2 2.2% 1.4% 1.2% 3.2%

Leafy Vegetables All 31 8.0% 13 210% 610% 0.32% 2500%

Sand 4 7.3% 1.5 7.8% 3.3% 4.9% 11%

Loam 3 82% 1.0 82% 3.5% 79% 86%

Clay 7 2.8% 4.1 5.1% 4.8% 0.41% 12%

Non-Leafy Vegetables Fruits All 5 1.5% 26 78% 170% 0.15% 390%

Shoots All 2 0.88% 0.42% 0.58% 1.17%

Legumes Pods All 17 0.53% 12 34% 120% 0.046% 490%

Sand 3 0.27% 3.2 0.42% 0.34% 0.065% 0.89%

Loam 5 0.14% 4.4 0.42% 0.34% 0.065% 0.89%

Clay 4 0.080% 1.0 0.33% 0.47% 0.046% 1.0%

Shoots All 1 0.080%

Root Crops Roots All 27 1.5% 16 41% 98% 0.024% 330%

Sand 5 6.4% 1.6 7.0% 3.4% 4.2% 12%

Loam 5 2.3% 4.7 0.50% 0.68% 0.024% 1.7%

Shoots All 12 6.3% 15 250% 570% 0.30% 16%

Tubers Tubers All 30 0.15% 7.4 9.1% 48% 0.015% 260%

Sand 5 0.64% 3.5 1.2% 1.6% 0.16% 3.9%

Loam 17 0.052% 2.4 0.073% 0.062% 0.015% 0.23%

Fruits Fruits All 5 0.77% 2.6 1.0% 0.60% 0.15% 1.7%

Leaves All 1 25%

Grasses All 17 31% 1.8 36% 22% 11% 100%

Natural Pastures All 34 92% 4.8 23% 29% 0.22% 100%

Leguminous Fodder All 1 1.6%

All Cereals All 20 0.43% 4.7 1.1% 1.4% 0.052% 4.8%

Sand 5 0.61% 5.3 1.3% 1.3% 0.052% 3.2%

Loam 8 0.17% 3.9 0.53% 1.1% 0.059% 3.2%

Clay 6 0.90% 4.0 1.8% 1.8% 0.22% 4.8%

Pastures/Grasses All 51 14% 4.2 27% 27% 0.22% 100%

Fodder All 24 2.5% 12 130% 420% 0.060% 1600%

Sand 4 4.5% 2.3 5.6% 4.0% 1.6% 11%

Clay 4 0.82% 5.7 2.7% 4.6% 0.16% 9.6%

Source: Used with permission from Elsevier Publishers, Vandenhove et al. (2009).

1 Findings from Pb uptake studies have implications for urban gardening if urban soils may be 

2 contaminated with Pb, as described in Section 4.1.1.2. For instance, Clark et al. (2006) tested the soil in 

3 103 urban gardens in two Boston neighborhoods. They found that Pb-based paint contributed 40-80% of 

4 measured Pb in the urban garden soil samples, with the rest coming from historical gasoline emissions. 

5 Furthermore, Clark et al. (2006) estimated that Pb consumption from urban gardens can be responsible for 

6 up to 25% of exposure to Pb in drinking water for children living in the Boston neighborhoods studied. 

7 Because soil Pb levels in urban areas will depend on surrounding sources (Pruvot et al., 2006), Pb 

8 exposures in urban garden vegetables will vary.

9 In addition to uptake of Pb through the roots of a plant, deposition of airborne Pb-bearing PM can 

10 also contribute to human dietary Pb exposures, as described in the 2006 Pb AQCD (2006). In a recent 

11 study, Uzu et al. (2010) found that Pb deposition from smelter emissions caused a linear increase in Pb 

concentrations of 7.0 mg/kg per day (R2 12 =0.96) in lettuce plants cultivated in the courtyard of a smelter. 
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1 They reported that lettuce grown 250-400 m from the smelter had concentrations that were 10-20 times 

2 lower, which is consistent with findings described in Section 3.3 that deposition of Pb containing material 

3 drops off with distance from a source. 

Fish

4 Accumulation in fish could also lead to human exposure to Pb. Ghosh et al. (2007) demonstrated in 

5 laboratory experiments that exposure to Pb in water can lead to linearly increasing accumulation in fish. 

6 Several studies have documented the potential for human exposure through fish and seafood. Welt et al. 

7 (2003) conducted a survey of individuals who fished in Bayou St. John, Louisiana in conjunction with 

8 sampling Pb content in sediment. They found that median sediment Pb concentrations ranged from 43 to 

9 330 mg/kg in different locations, while maximum Pb concentrations ranged from 580 to 6,500 mg/kg. In 

10 total, 65% of those surveyed fished for food from the Bayou, with 86% consuming fish from the Bayou 

11 each week. In a study of the effect of coal mining on levels of metals in fish (measured as blood Pb) in 

12 northeastern Oklahoma, Schmitt et al. (2005) found that Pb content varied with respect to species of fish 

13 but were found to be elevated in some species. Pb concentrations in fish were higher in areas close to 

14 mining activities. Similarly, Besser et al. (2008) observed higher levels of blood Pb in fish close to mining 

15 activities in southeastern Missouri. In a related study of fish species in the same region of Missouri, blood 

16 Pb levels in fish were found to be significantly higher in sites within 10 km downstream of active Pb-Zn 

17 mines (p <0.01) compared with fish located further from the mines (Schmitt et al., 2007), and elevated 

18 blood Pb levels in fish were again noted near a Pb-Zn mine (Schmitt et al., 2009). It was noted that the 

19 Ozark streams where these studies were performed were often used for recreational fishing. There has 

20 also been evidence of elevated Pb concentration within large game from mining areas (Reglero et al., 

21 2009). 

4.1.1.4. Occupational

22 Occupational environments have the potential to expose individuals to Pb. Some modern day 

23 occupational exposures are briefly discussed below in the context of understanding potential exposures 

24 that are not attributed to ambient air. For example, Miller et al. (2010) obtained personal and area samples 

25 of particle-borne Pb in a precious metals refinery. It was not stated explicitly, but it is likely that Miller et 

26 al. (2010) measured the PM as TSP because the Occupational Safety and Health Administration (OSHA) 

27 permissible exposure limit (PEL) for Pb is based on TSP rather than a smaller size cut, and the OSHA 

PEL was used for comparison. Concentrations measured by personal samples ranged from 2 to 6 μg/m3 28 , 

and concentrations from area samples ranged from 4 to 14 μg/m3. The OSHA PEL is 5 μg/m3 29 . In steel 

30 production, sintering was found to be the largest source of airborne Pb exposure in a survey of operations 
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1 (Sammut et al., 2010), with Pb enrichment in PM reported to be 20,000 mg/kg, although total PM 

2 concentration, reported to have 75% below 2.5 μm diameter, was not reported.

3 Operations involving PM in various industries are a source of occupational Pb exposure, in 

4 addition to a residential exposure. Rodrigues et al. (2010) reported exposures to airborne Pb among New 

5 England painters, who regularly use electric grinders to prepare surfaces for painting. Two-week averaged 

6 airborne Pb concentrations, sampled with an Institute of Medicine inhalable PM sampler designed to 

capture PM smaller than 100 μm, were reported to be 59 μg/m3 7 , with a maximum daily value of 210 

μg/m3 8 . The Pb concentrations reported here were corrected by the National Institute for Occupational 

9 Safety and Health (NIOSH) respirator protection factors, although the respirator protection factors were 

10 not reported by Rodrigues et al. (2010). Information on the air Pb-blood Pb relationship can be found in 

11 Section 4.5.1. Nwajei and Iwegbue (2007) measured Pb contamination in sawdust; such contamination 

12 has been reported to occur when trees are grown in soil contaminated with Pb (Andrews et al., 1989). 

13 Sawdust samples from fifteen locations in Nigerian sawmills were reported to have Pb concentrations 

14 ranging from 2.0 to 250 mg/kg.

4.1.1.5. Exposure to Lead from Consumer Products

15 Pb is present in varying amounts in several consumer products including alternative medicines, 

16 candies, cosmetics, pottery, tobacco, toys, and vitamins (Table 4-6). Several of these categories suggest 

17 children may incur regular exposures. Pb concentrations were reported to range from non-detectable 

18 levels up to 77% by mass, for the case of one medicinal product. Exposure to these products, which 

19 originate in a range of different countries, can account for substantial influence on Pb body burden (Levin 

20 et al., 2008; Miodovnik & Landrigan, 2009).
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Table 4-6. Pb content in various consumer products

Product 

Category Product Location of Purchase Pb Content (units) Reference

Alternative 

and 

Traditional 

Medicines

Cissus quadrangularis, Caulophyllum thalictroides, 

Turnera diffusa, Centella asiatica, Hoodia gordonii, 

Sutherlandia frutescens, Curcuma longa, fucoxanthin, 

Euterpe oleracea

(dietary supplements claimed to be from Hoodia 

gordonii)

U.S. 

(Mississippi)a Not detected (N.D.) -4.21 mg/kg Avula et al (2010)

Malva sylvestris Turkey 1.1-2.0 mg/kg Hiçsömnez et al 

(2009)

Yugmijihwang-tang, Bojungigki-tang, Sibjeondaebotang, Kuibi-tang, Ojeogsan Korea 7.9×10-6 to 2.5×10-5 mg/kg 

body weight/day Kim et al (2009)

Lemongrass, licorice, holy basil, cloves, ginger India

Average:

Lemongrass & Holy Basil 

Leaves: 6.1 mg/kg; Licorice 

Stolons: 6.1 mg/kg, Clove 

Dried Flower Buds: 7.8 

mg/kg, Ginger Rhizome: 

5.8 mg/kg

Naithani and Kakkar 

(2006)

B-Success 28, Operation Sweep, Aloe Vera Plus 

Bitter Aloes, Zarausmacine, Virgy-Virgy Computer 

Worm-Expeller, Dorasine Powder, Sexual Energy, 

U&DEE Infection Cleansing Powder, U&DEE Sweet 

Bitter, Natural Power Stone, Charma Black Stone, 

Portugal Antiseptic Soap, Edysol Antiseptic Soap, H- Nal, M-Reg, Veins Flocher, Diabor, C-Candi, C-Cysta, 

Firas, D-Diab, P-Pile, Infecta, Ribacin Forte, Aloe 

Vera Cure Formula

Nigeria 925-27,000 μg Obi et al (2006)

Shell of Hen’s Egg India 14 mg/kg Sharma et al (2009)

Berberis (B. aristata, B. chitria, B. asiatica, B. 

lyceum), Daruharidra India

Berberis:

Roots: 3.1-24.7 mg/kg

Stems: 8.0-23.8 mg/kg

Daruharidra:

16.9-49.8 mg/kg

Srivastava et al 

(2006)

Greta powder U.S. (California) 770,000 ppm CDC (2002)

Candy

Tamarind Candy U.S. 

(Oklahoma)

Product: 0.15-3.61 mg/kg

Stems: 0.36-2.5 mg/kg

Wrappers: 459-27,125 

mg/kg

Lynch et al (2000)

Tamarind Candy U.S. (California)

Product: 0.2-0.3 mg/kg

Stems: 400 mg/kg

Wrappers: 16,000- 21,000 mg/kg

CDC (2002)

Cosmetics

Lipsticks U.S. Average: 1.07 mg/kg Hepp et al (2009)

Eye Shadows Nigeria N.D.-55 mg/kg Omolaoye et al 

(2010a)

Pottery Foods prepared in Pb-glazed pottery Mexico N.D.-3,100 mg/kg Villalobos et al (2009)

Tobacco

Smokeless Tobacco United Kingdom 0.15-1.56 mg/kg McNeill et al (2006)

Cigarette Tobacco (210Pb concentrations) Pakistan Activity conc.: 7-20 Bq/kg Tahir and Alaamer 

(2008)

Toys

Red and yellow painted toy vehicles and tracks Brazil 500-6,000 mg/kg Godoi et al (2009)

535 PVC and non-PVC toys from day care centers U.S. (Nevada)

PVC: avg. 325 mg/kg

Non-PVC: avg. 89 mg/kg 

Yellow: 216 mg/kg

Non-yellow: 94 mg/kg

Greenway and 

Gerstenberger (2010)

Soft plastic toys India Average (by city): 21-280 

mg/kg

Kumar and Pastore 

(2007)

Toy necklace U.S. 388,000 mg/kg Meyer et al (2008)

Soft plastic toys Nigeria 2.5-1,445 mg/kg Omolaoye et al 

(2010b)

Vitamins Vitamins for young children, older children, and 

pregnant or lactating women U.S.

Average:

Young children: 2.9 μg/day

Older children: 1.8 μg/day

Pregnant and lactating 

women: 4.9 μg/day

Mindak et al. (2008)

a

Hoodia gordonii, from Eastern Cape, South Africa Euterpe oleracea from Ninole Orchard, Ninole, Hawaii
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4.2. Kinetics

1 This section summarizes the empirical bases for our understanding of Pb toxicokinetics in humans. 

2 The large amount of empirical information on Pb biokinetics in humans and animal models has been 

3 integrated into mechanistic biokinetics models (U.S. EPA, 2006). These models support predictions about 

4 the kinetics of Pb in blood and other selected tissues based on the empirically-based information about Pb 

5 biokinetics. In Section 4.3, Pb biokinetics is described from the context of model predictions.

4.2.1. Absorption

6 The focus of the following sections within absorption is on inhalation and ingestion because these 

7 are the major exposure routes of Pb in humans. The 2006 AQCD also presented dermal absorption of 

8 inorganic and organic Pb compounds, which is generally considered to be much less than by inhalation or 

9 ingestion. No recent literature has advanced our knowledge of dermal absorption of Pb beyond that which 

10 was included in the 2006 AQCD. No additional information provides evidence of dermal absorption being 

11 a major exposure route of environmental Pb.

12 The term absorption refers to the fraction of the amount of Pb ingested or inhaled that is absorbed 

13 from the respiratory or gastrointestinal tract. The term bioavailability, as it is used in this section, refers to 

14 the fraction of the amount of Pb ingested or inhaled that enters the systemic circulation. If properly 

15 measured (e.g., time-integrated blood Pb), under most conditions Pb bioavailability is equivalent (or 

16 nearly equivalent) to Pb absorption. Bioaccessibility is a measure of the physiological solubility of Pb in 

17 the respiratory or gastrointestinal tract. Pb must become bioaccessible in order for absorption to occur. 

18 Processes that contribute to bioaccessibility include physical transformation of Pb particles and 

dissolution of Pb compounds into forms that can be absorbed (e.g., Pb2+ 19 ).

4.2.1.1. Inhalation

20 Systemic absorption of Pb deposited in the respiratory tract is influenced by particle size and 

21 solubility, as well as by the pattern of regional deposition within the respiratory tract. Fine particles (<1 

22 μm) deposited in the bronchiolar and alveolar region can be absorbed after extracellular dissolution or can 

23 be ingested by phagocytic cells and transported from the respiratory tract (Bailey & Roy, 1994). Larger 

24 particles (>2.5 μm) that are primarily deposited in the ciliated airways (nasopharyngeal and 

25 tracheobronchial regions) can be transferred by mucociliary transport into the esophagus and swallowed, 

26 thus being absorbed via the gut. 

27 Inhaled Pb lodging deep in the respiratory tract seems to be absorbed equally and totally, regardless 

28 of chemical form (Chamberlain et al., 1978; Morrow et al., 1980; M. B. Rabinowitz et al., 1977). 

29 Absorption half-times (t1/2) have been estimated for radon decay progeny in adults who inhaled aerosols 
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of Pb and bismuth isotopes generated from decay of 220Rn or 222 1 Rn. The absorption half-time for Pb from 

2 the respiratory tract to blood was estimated to be approximately 10 hours in subjects who inhaled aerosols 

3 having an activity median particle diameter of approximately 160 nm (range 50-500 nm) (Marsh & 

4 Birchall, 1999), and approximately 68 min for aerosols having diameters of approximately 0.3–3 nm 

5 (Butterweck et al., 2002). Given the submicron particle size of the exposure, these rates are thought to 

6 represent, primarily, absorption from the bronchiolar and alveolar regions of the respiratory tract.

7 Several studies have attempted to quantify the bioavailability of Pb in atmospheric PM, although 

8 different laboratory techniques are used throughout the literature, as described in Section 3.4. Unlike the 

9 bioavailability methods described in Section 4.2.1.2, many of these in vitro assays have not been 

10 validated with in vivo models. Therefore, any in vitro method can only be a simplistic surrogate of the 

11 complex uptake of Pb from the gastrointestinal (GI) tract. Despite this limitation, the studies mentioned 

12 below are the only ones to evaluate atmospheric Pb.

13 In a study of PM10 and PM2.5 samples from downtown Vienna, Austria, Falta et al. (2008) used 

14 synthetic gastric juice to investigate the bioavailability of heavy metals including Pb. The rationale was 

15 that inhaled PM in the 2.5-10 μm size range are mostly deposited in the tracheal and bronchial regions of 

16 the lung from where they are transported within hours by mucociliary clearance, i.e., they are mainly 

17 swallowed. In contrast, the <2.5 μm particles are deposited in the pulmonary alveoli where they can stay 

18 for months to years. The study aimed to determine the bioavailability of the 2.5-10 μm PM. It is important 

19 to note that they do not isolate the 2.5-10 μm size range; instead, they infer the characteristics from the 

20 difference between the PM2.5 and PM10 fractions. The Pb concentrations associated with the two fractions 

21 were almost identical, as was the percentage extracted by synthetic gastric juice (86% and 83% Pb for 

22 PM2.5 and PM10 fractions, respectively). The mean daily bioavailable mass was calculated to be 16 ng for 

23 the PM2.5-10 size range. Since the quantitative clearance of these particles to the stomach was assumed, 

24 this value represents an upper estimate for the amount of bioavailable Pb. Niu et al. (2010) determined the 

25 bioavailability of Pb fine (100-1,000 nm) and ultrafine-sized (<100 nm) urban airborne PM from two sites 

26 within the city of Ottawa, Canada. For all size fractions, the median Pb concentrations for PM smaller 

27 than 10 μm were 8,800 and 7,800 mg/kg for the two different locations. The bioavailability was based on 

28 ammonium acetate extractability and it was found that, within the fine and nano-size ranges, 13-28% Pb 

29 was extracted. The Falta et al. (2008) and Niu et al. (2010) results illustrate that different extraction 

30 techniques result in different bioavailable fractions. The main finding from Niu et al. (2010) was that the

31 highest values (~28% and ~19% for the two different locations) were found for the <57 nm PM, with 

32 percent bioavailability decreasing with increasing PM size. This indicated that Pb was potentially most 

33 bioavailable in the nano-size range.

34 A recent study by Barrett et al. (2010) investigated the solid phase speciation of Pb in urban road 

35 dust in Manchester, UK, and considered the health implications of inhalation and ingestion of such 

36 material. Human exposure via inhalation is likely to involve only the finest grained fractions (up to 10 
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1 μm) and unfortunately this study characterized only the <38 μm fraction. Pb-goethite and PbCrO4

2 comprised the largest fractions, 45% and 21% respectively, of Pb in the <38 μm fraction. These forms 

3 tend to be less bioavailable if ingested compared with PbO or Pb-acetate because they are less soluble.

Organic Lead

4 Alkyl Pb compounds can exist in ambient air as vapors. Inhaled tetraalkyl Pb vapor is nearly 

5 completely absorbed following deposition in the respiratory tract. As reported in the 2006 AQCD, a single 

exposure to vapors of radioactive (203 6 Pb) tetraethyl Pb resulted in 37% initially deposited in the 

7 respiratory tract, of which ~20% was exhaled in the subsequent 48 hours (Heard et al., 1979). In a similar 

experiment conducted with 203Pb tetramethyl Pb, 51% of the inhaled 203 8 Pb dose was initially deposited in 

9 the respiratory tract, of which ~40% was exhaled in 48 hours (Heard et al., 1979). 

10 Estimation of bioavailability of organic Pb is relevant to some aviation fuel exposures (i.e., piston11 engine aircraft). Mahaffey (1977) estimated that 40% of inhaled Pb is bioavailable to adults. Chamberlain 

et al. (1975) suggested that 35% of inhaled combustion products of tetraethyl 203 12 Pb fuel are deposited and 

then retained in adult lungs with a half-life of 6 hours. Fifty percent of that 203 13 Pb was detectable in the 

14 blood within 50 hours of inhalation, and the rest was found to deposit in bone or tissue. Chamberlain et al. 

(1975) estimated that continuous inhalation of Pb at a concentration of 0.001 μg/m3 15 could produce a 1 

16 μg/dL increment in blood Pb.

4.2.1.2. Ingestion

17 The extent and rate of GI absorption of ingested inorganic Pb are influenced by physiological states 

18 of the exposed individual (e.g., age, fasting, nutritional calcium and iron status, pregnancy) and 

19 physicochemical characteristics of the Pb-bearing material ingested (e.g., particle size, mineralogy, 

20 solubility). Pb absorption in humans may be a capacity-limited process, in which case the percentage of 

21 ingested Pb that is absorbed may decrease with increasing rate of Pb intake. Numerous observations of 

22 nonlinear relationships between blood Pb concentration and Pb intake in humans provide support for the 

23 likely existence of a saturable absorption mechanism or some other capacity-limited process in the 

24 distribution of Pb in humans (Pocock et al., 1983; J. Sherlock et al., 1982; J. C. Sherlock et al., 1984; J. C. 

25 Sherlock & Quinn, 1986). While evidence for capacity-limited processes at the level of the intestinal 

26 epithelium is compelling, the dose at which absorption becomes appreciably limited in humans is not 

27 known.

28 In adults, estimates of absorption of ingested water-soluble Pb compounds (e.g., Pb chloride, Pb 

29 nitrate, Pb-acetate) range from 3 to 10% in fed subjects (Heard & Chamberlain, 1982; James et al., 1985; 

30 Maddaloni et al., 1998; M. B. Rabinowitz et al., 1980; Watson et al., 1986). The absence of food in the GI 

31 tract increases absorption of water-soluble Pb in adults. Reported estimates of soluble Pb absorption range 
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1 from 26 to 70% in fasted adults (Blake et al., 1983; Heard & Chamberlain, 1982; James et al., 1985; 

2 Maddaloni et al., 1998; M. B. Rabinowitz et al., 1980). Reported fed:fasted ratios for soluble Pb 

3 absorption in adults range from 0.04 to 0.2 (Blake et al., 1983; Heard & Chamberlain, 1982; James et al., 

4 1985; M. B. Rabinowitz et al., 1980).

5 Limited evidence demonstrates that GI absorption of water-soluble Pb is higher in children than in 

6 adults. Estimates derived from dietary balance studies conducted in infants and children (ages 2 weeks to 

7 8 years) indicate that ~ 40-50% of ingested Pb is absorbed (Alexander et al., 1974; Ziegler et al., 1978). 

8 Experimental studies provide further evidence for greater absorption of Pb from the gut in young animals 

9 compared to adult animals (Aungst et al., 1981; Forbes & Reina, 1972; Kostial et al., 1978; Pounds et al., 

10 1978). The mechanisms for an apparent age difference in GI absorption of Pb have not been completely 

11 elucidated and may include both physiological and dietary factors (Mushak, 1991).

12 Nutritional deficiencies have also been linked to Pb absorption in the GI tract, particularly in 

13 children. Children who are iron-deficient have higher blood Pb concentrations than similarly exposed 

14 iron-replete children, suggesting that iron deficiency may result in higher Pb absorption or, possibly, other 

15 changes in Pb biokinetics that contribute to altered blood Pb concentrations (Mahaffey & Annest, 1986; 

16 Marcus & Schwartz, 1987; Schell et al., 2004). Studies conducted in animal models have provided direct 

17 evidence for interactions between iron deficiency and increased Pb absorption, perhaps by enhancing 

18 binding of Pb to iron-binding proteins in the intestine (Bannon et al., 2003; Barton, Conrad, Nuby, et al., 

19 1978; Morrison & Quarterman, 1987). 

20 The effects of iron nutritional status on blood Pb include changes in blood Pb concentrations in 

21 association with genetic variation in genes involved in iron metabolism. For example, genetic variants in 

22 the hemochromatosis (HFE) and transferrin genes are associated with higher blood Pb concentrations in 

23 children (Hopkins et al., 2008). In contrast, HFE gene variants are associated with lower bone and blood 

24 Pb levels in elderly men (Wright et al., 2004). 

25 Several studies have suggested that dietary calcium may have a protective role against Pb by 

26 decreasing absorption of Pb in the GI tract and by decreasing the mobilization of Pb from bone stores to 

27 blood. In experimental studies of adults, absorption of a single dose of Pb (100-300 μg Pb chloride) was 

28 lower when the Pb was ingested together with calcium carbonate (0.2 g calcium carbonate) than when the 

29 Pb was ingested without additional calcium (Blake & Mann, 1983; Heard & Chamberlain, 1982). A 

30 similar effect of calcium occurs in rats (Barton, Conrad, Harrison, et al., 1978). Similarly, an inverse 

31 relationship was observed between dietary calcium intake and blood Pb concentration in children, 

32 suggesting that children who are calcium-deficient may absorb more Pb than calcium-replete children 

33 (Elias et al., 2007; Mahaffey et al., 1986; Schell et al., 2004; Ziegler et al., 1978). These observations 

34 suggest that calcium and Pb share and may compete for common binding and transport mechanisms in the 

35 small intestine which are regulated in response to dietary calcium and calcium body stores (Bronner et al., 

36 1986; Fullmer & Rosen, 1990). However, animal studies have also shown that multiple aspects of Pb 
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1 toxicokinetics are affected by calcium nutritional status. For example, feeding rats a calcium deficient diet

2 is associated with increased Pb absorption, decreased whole body Pb clearance, and increased volume of 

3 distribution of Pb (Aungst & Fung, 1985). These studies suggest that associations between calcium 

4 nutrition and blood Pb that have been observed in human populations may not be solely attributable to 

5 effects of calcium nutrition on Pb absorption. Other potential mechanisms by which calcium nutrition 

6 may affect blood Pb and Pb biokinetics include effects on bone mineral metabolism and renal function.

7 Blood Pb concentrations in young children have also been shown to increase in association with 

8 lower dietary Zn levels (Schell et al., 2004). Mechanisms for how Zn affects blood Pb concentration, i.e., 

9 whether it involves changes in absorption or changes in distribution and/or elimination of Pb, have not 

10 been determined.

11 Dissolution of Pb from the soil/mineralogical matrix in the stomach appears to be the major process 

12 that renders soil Pb bioaccessible for absorption in the GI tract. Relative bioavailability (RBA) of Pb in 

13 soils and dust has been most extensively studied in animal models. Relative bioavailability is the ratio of 

14 the absorbed fraction (AF) of ingested dose of soil Pb to that of a water-soluble form of Pb (e.g., Pb15 acetate) that is considered to be completely bioaccessible (e.g., RBA = AFSoil Pb/AFPb-acetate). In typical 

16 studies, the absorbed fraction of the Pb dose is estimated based measurements of Pb measured in blood 

17 and/or other tissues (e.g., kidney, liver, bone) after dosing . Gastric function of swine is thought to be 

18 sufficiently similar to that of humans to justify use of swine as a model for assessing RBA of Pb in soils 

19 (Casteel et al., 1997; Casteel et al., 2006; Juhasz et al., 2009; U.S. EPA, 2007a; Weis & Lavelle, 1991). 

20 Other practical advantages of the swine model over rodent models have been described, and include:

21 absence of coprophagia; ease with which Pb dosing can be administered and controlled; and higher 

22 bioavailability of soluble Pb (e.g., Pb-acetate) in swine, which is more similar to humans than rats (D. M. 

23 Smith et al., 2009). Relative bioavailability of Pb has been shown to vary depending upon the Pb 

24 mineralogy and physical characteristics of the Pb in the soil (e.g., encapsulated or exposed) and size of the 

25 Pb-bearing grains. GI absorption of larger Pb-containing particles (>100 μm) tends to be lower than 

26 smaller particles (Barltrop & Meek, 1979; Healy et al., 1992).

27 Collectively, published studies conducted in swine have provided 39 estimates of Pb RBA for 37 

28 different soil or “soil-like” test materials (Bannon et al., 2009; Casteel et al., 2006; Marschner et al., 2006; 

29 D. M. Smith et al., 2009). The mean of RBA estimates from 25 soils is 49% (± 29[SD]), median is 51%, 

and 5th 30 to 95th percentile range is 12 to -89%. RBA estimates for soils collected from 8 firing ranges were 

31 approximately 100% (Bannon et al., 2009). The relatively high RBA for the firing range soils may reflect 

32 the high abundance of relatively un-encapsulated Pb carbonate (30-90% abundance) and Pb oxide (1-

33 60%) in these soils. Similarly, a soil sample (low Pb concentration) mixed with a NIST paint standard 

34 (55% Pb carbonate, 44% Pb oxide) also had a relatively high bioavailability (72%)(Casteel et al., 2006). 

35 Samples of smelter slag, or soils in which the dominant source of Pb was smelter slag, had relatively low 
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1 RBA (14-40%, n = 3), as did a sample from a mine tailings pile (RBA = 6%), and a sample of finely 

2 ground galena mixed with soil (Casteel et al., 2006).

3 Based on data for 18 soil materials assayed in swine, RBA of Pb mineral phases were categorized 

4 into “low” (<0.25 [25%]), “medium” (0.25-0.75 [25 to 75%]), and “high” (>0.75 [75%]) categories 

5 (Casteel et al., 2006). Figure 4-2 shows some of the materials that fall into these three categories. Mineral 

6 phases observed in mineralogical wastes can be expected to change over time (i.e., weathering), which 

7 could change the RBA over time. The above observations in swine are supported by various studies 

8 conducted in rats that have found RBA of Pb in soils to vary considerably and to be less that 100% 

9 (Freeman et al., 1996; Freeman et al., 1992; Freeman et al., 1994; D. M. Smith et al., 2008, 2009).

Source: Casteel et al. (2006).

Figure 4-2. Estimated relative bioavailability (RBA, compared to Pbacetate) of ingested Pb in mineral groups, based on results 

from juvenile swine assays.

10 Drexler and Brattin (2007) developed an in vitro bioaccessibility (IVBA) assay for soil Pb that 

11 utilizes extraction fluid comprised of glycine, deionized water, and hydrochloric acid at a pH of 1.50 that 

12 is combined with sieved test material (<250 μm) for 1 hour. The assay was tested for predicting in vivo

13 RBA of 18 soil-like test materials that were assayed in a juvenile swine assay (Casteel et al., 2006). A 

14 regression model relating IVBA and RBA was derived based on these data (Equation 4-1):
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1 RBA = (0.878 • IVBA) – 0.028

Equation 4-1

where RBA and IVBA are expressed as fractions (i.e., not as percent). The weighted r2 2 for the relationship 

3 (weighted for error in the IVBA and RBA estimates) was 0.924 (p<0.001). The IVBA assay reported in 

4 Drexler and Brattin (2007) has been identified by the U.S. EPA as a validated method for predicting RBA 

5 of Pb in soils for use in risk assessment (U.S. EPA, 2007b). A review of soil Pb RBA estimates made 

6 using the IVBA assay described above and Equation 4-1 identified 270 estimates of Pb RBA in soils 

7 obtained from 11 hazardous waste sites. The mean for the site-wide RBA estimates (n = 11 sites) was 57% 

8 (SD 15), median was 63%, and 5th to 95th percentile range was 34 to 71%.

9 Equation 4-1 cannot be reliably extrapolated to other in vitro assays that have been developed for 

10 estimating Pb bioaccessibility without validation against in vivo RBA measurements made on the same 

11 test materials. Comparisons of outcomes of in vitro assays applied to the same soil test materials have 

12 found considerable variability in IVBA estimates (Saikat et al., 2007; Van de Wiele et al., 2007). This 

13 variability has been attributed to differences in assay conditions, including pH, liquid:soil ratios, inclusion 

14 or absence of food material, and differences in methods used to separate dissolved and particle-bound Pb 

15 (e.g., centrifugation versus filtration). Given the dependence of IVBA outcomes on assay conditions, in 

16 vitro assays used to predict in vivo RBA should be evaluated against in vivo RBA estimates to 

17 quantitatively assess uncertainty in RBA predictions (U.S. EPA, 2007b). 

18 Absorption of Pb in house dusts has not been rigorously evaluated quantitatively in humans or in 

19 experimental animal models. The RBA for paint Pb mixed with soil has been reported to be 

20 approximately 72% (95% CI: 44, 98) in juvenile swine, suggesting that paint Pb dust maybe highly 

21 bioavailable (Casteel et al., 2006). The same material yielded a bioaccessibility value (based on IVBA 

22 assay) of 75% (Drexler & Brattin, 2007), which corresponds to a predicted RBA of 63%, based on 

23 Equation 4-1. A review of indoor Pb RBA estimates made using the IVBA assay and Equation 4-1

24 identified 100 estimates of Pb RBA in dusts obtained from two hazardous waste sites. Mean Pb RBAs for 

25 the Herculaneum site were 47% (SD 7, 10 samples) for indoor dust and 69% (SD 3, 12 samples) for soil. 

26 At the Omaha site, mean Pb RBAs were 73% (SD 10, 90 samples) for indoor dust and 70% (SD 10, 45 

27 samples) for soil. Yu et al. (2006) applied an IVBA method to estimate bioaccessibility of Pb in house 

28 dust samples collected from 15 urban homes. Homes were selected for inclusion in this study based on 

29 reporting to the state department of health of at least on child with a blood Pb concentration >15 μg/dL 

30 and Pb paint dust may have contributed to indoor dust Pb. The mean IVBA was 64.8% (SD 8.2, age: 52.5

31 to 77.2 months). 

32 The above results, and the IVBA assays used in studies of interior dust, have not been evaluated 

33 against in vivo RBA estimates for dust samples. Although, expectations would be that a validated IVBA 

34 methodology for soil would perform well for predicting RBA of interior dust, this has not actually been 
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1 experimentally confirmed. Factors that may affect in vitro predictions of RBA of interior dust Pb could 

2 include particle size distribution of interior dust Pb and the composition of the dust matrix, which may be 

3 quite different from that of soil.

4 Other estimates of “bioavailability” of Pb exposure samples are derived from less validated in vitro 

5 methods. Roussel et al. (2010) estimated that 63% of soil Pb is bioavailable in the stomach, while 39% is 

6 bioavailable in the intestines, using different acidities of solutions to simulate acids in the digestive 

7 system. Yu et al. (2006) dissolved Pb dust, obtained from vacuuming carpet samples, into simulated 

8 gastric and intestinal acids. The carpet samples were obtained from homes located in northern New 

9 Jersey. Pb concentration in carpet ranged from 209 to 1,770 mg/kg dust, with 52-77% of Pb dissolving in 

10 simulated gastric acid and 5-32% dissolving in simulated intestinal acids. In a similar test in the U.K., 

11 Turner and Simmonds (2006) observed median Pb dust concentrations of 178 mg/kg with approximately 

12 80% bioavailability in simulated gastric acid. Jin et al. (2005) observed that bioaccessibility of Pb in soil 

13 was proportional to the soil acidity and organic matter content of the soil.

4.2.2. Distribution

14 A simple conceptual representation of Pb distribution is that it contains a fast turnover pool, 

15 comprising mainly soft tissue, and a slow pool, comprising mainly skeletal tissues (M. B. Rabinowitz et 

16 al., 1976). The highest soft tissue concentrations in adults occur in liver and kidney cortex (Barry, 1975; 

17 Gerhardsson et al., 1986; Gerhardsson et al., 1995; Gross et al., 1975; Oldereid et al., 1993). Pb in blood 

18 (i.e., plasma) exchanges with both of these compartments.

4.2.2.1. Blood

19 Blood comprises ~1% of total Pb body burden. Pb in blood is found primarily (>99%) in the RBCs

20 (Bergdahl, Grubb, et al., 1997; Bergdahl et al., 1998; Bergdahl et al., 1999; Hernandez-Avila et al., 1998; 

21 Manton et al., 2001; Schutz et al., 1996; D. Smith et al., 2002). δ-aminolevulinic acid dehydratase 

22 (ALAD) is the primary binding ligand for Pb in erythrocytes (Bergdahl, Grubb, et al., 1997; Bergdahl et 

23 al., 1998; Sakai et al., 1982; Xie et al., 1998). Two other Pb-binding proteins have been identified in the 

24 RBC, a 45 kDa protein (Kmax 700 μg/dL; Kd 5.5 μg/L) and a smaller protein(s) having a molecular weight 

25 <10 kDa (Bergdahl, Grubb, et al., 1997; Bergdahl et al., 1996; Bergdahl et al., 1998). Of the three 

26 principal Pb-binding proteins identified in RBCs, ALAD has the strongest affinity for Pb (Bergdahl et al., 

27 1998) and appears to dominate the ligand distribution of Pb (35 to 84% of total erythrocyte Pb) at blood 

28 Pb levels below 40 μg/dL (Bergdahl et al., 1996; Bergdahl et al., 1998; Sakai et al., 1982). Pb binding to 

29 ALAD is saturable; the binding capacity has been estimated to be ~850 μg/dL RBCs (or ~40 μg/dL whole 

30 blood) and the apparent dissociation constant has been estimated to be ~1.5 μg/L (Bergdahl et al., 1998).
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1 Saturable binding to RBC proteins contributes to an increase in the plasma/blood Pb ratio with 

2 increasing blood Pb concentration and curvature to the blood Pb–plasma Pb relationship (Barbosa, 

3 Ramires, et al., 2006; Bergdahl, Schutz, et al., 1997; Bergdahl et al., 1998; Bergdahl et al., 1999; DeSilva, 

4 1981; C. Jin et al., 2008; Kang et al., 2009; Manton et al., 2001; D. Smith et al., 2002). An example of this 

5 is shown in Figure 4-3. Saturable binding of Pb to RBC proteins has several important consequences. As 

6 blood Pb increases and the higher affinity binding sites for Pb in RBCs become saturated at 

7 approximately 40 μg/dL blood, a larger fraction of the blood Pb is available in plasma to distribute to 

8 brain and other Pb-responsive tissues. This change in distribution of Pb contributes to a curvature in the 

9 relationship between Pb intake (at constant absorption fraction) and blood Pb concentration. 

10 Typically, at blood Pb concentrations <100 μg/dL, only a small fraction (<1%) of blood Pb is found 

11 in plasma (DeSilva, 1981; Manton & Cook, 1984; Marcus, 1985). However, as previously noted, plasma 

12 Pb may be the more biologically labile and toxicologically active fraction of the circulating Pb. 

13 Approximately 40-75% of Pb in the plasma is bound to proteins, of which albumin appears to be the 

14 dominant ligand (Al-Modhefer et al., 1991; Ong & Lee, 1980). Pb in serum that is not bound to protein 

15 exists largely as complexes with low molecular weight sulfhydryl compounds (e.g., cysteine, 

16 homocysteine) and other ligands (Al-Modhefer et al., 1991).

Source: Adapted, with permission from Elsevier Publishing and the Finland Institute of Occupational Health, from Bergdahl et al. (1997; 1999).

Figure 4-3. Plot of blood and plasma Pb concentrations in measured in 

adults and children.

17 As shown in Figure 4-3, the limited binding capacity of Pb binding proteins in RBCs produces a 

18 curvilinear relationship between blood and plasma Pb concentration. The limited binding capacity of RBC

0.0

0.5

1.0

1.5

2.0

0 20 40 60 80 100

Plasma Pb (

μg/dL)

Blood Pb (μg/dL)

Adults Children
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1 binding proteins also confers, or at least contributes, to a curvilinear relationship between Pb intake and 

2 blood Pb concentration. A curvilinear relationship between Pb intake and blood Pb concentration has been 

3 observed in children (Lacey et al., 1985; Ryu et al., 1983; J. C. Sherlock & Quinn, 1986). As shown in 

4 Figure 4-4, the relationship becomes pseudo-linear at relatively low daily Pb intakes (i.e., <10 μg/day/kg) 

5 and at blood Pb concentrations <25 μg/dL. 

Source: Adapted, with permission from Taylor & Francis Publishing, from Sherlock and Quinn (1986).

Figure 4-4. Relationship between Pb intake and blood Pb concentration in 

infants (n = 105, age 13 weeks, formula-fed). Data represent 

mean and standard errors for intake; the line is the regression 

model (blood Pb = 3.9 + 2.43 (Pb intake [μg/week]1/3). 

6 Figure 4-5 shows the predicted relationship between quasi-steady state blood and plasma Pb 

7 concentrations in a 4-year old child using the ICRP model (ICRP, 1994; Leggett, 1993; Pounds & Leggett, 

8 1998). The abrupt inflection point that occurs at approximately 25 μg/dL blood Pb is an artifact of the 

9 numerical approach to simulate the saturation of binding using discontinuous first-order rate constants for 

10 uptake and exit of Pb from the RBC. A continuous function of binding sites and affinity, using empirical 

11 estimates of both parameters, yield a similar but continuous curvature in the relationship (Bergdahl et al., 

12 1998; O'Flaherty, 1995). Nevertheless, either approach predicts a pseudo-linear relationship at blood Pb 

13 concentrations below approximately 25 μg/dL which, in this model, corresponds to an intake of 

14 approximately 100 μg/day (absorption rate ≈ 30 μg/day) (upper panel). An important consequence of the 

15 limited Pb binding capacity of RBC proteins is that the plasma Pb concentration will continue to grow at 

16 a linear rate above the saturation point for RBC protein binding. One implication of this is that a larger 

0

5

10

15

20

25

30

35

40

45

0 100 200 300 400

Blood Pb (

μg/dL)

Pb Intake (μg/day)
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1 fraction of the Pb in blood will become available to distribute to brain and other Pb-responsive tissues as 

2 blood Pb increases. This could potentially contribute to non-linearity in dose-response relationships in 

3 studies in which blood Pb is the used as the internal dose metric.

Figure 4-5. Simulation of quasi-steady state blood and plasma Pb 

concentrations in a child (age 4 years) associated with varying 

Pb ingestion rates. Simulation based on ICRP Pb biokinetics 

model (Leggett, 1993).

0.00

0.10

0.20

0.30

0.40

0.50

0.60

0 10 20 30 40 50

Plasma Pb (

μg/dL)

Blood Pb (μg/dL)

0.0

0.2

0.4

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1.0

0

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0 100 200 300 400

Plasma Pb (μg/dL)

Blood Pb (

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Blood Plasma
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1 Studies conducted in swine provide additional evidence in support of RBC binding kinetics 

2 influencing distribution of Pb to tissues. In these studies, the relationship between the ingested dose of Pb 

3 and tissue Pb concentrations (e.g., liver, kidney, bone) was linear, whereas, the relationship between dose 

4 and blood Pb was curvilinear with the slope decreasing as the dose increased (Casteel et al., 2006). 

5 Saturable binding of Pb to RBC proteins also contributes to a curvilinear relationship between urinary Pb 

6 excretion and plasma Pb concentration (Section 4.2.3) (Bergdahl, Schutz, et al., 1997; Besser et al., 2008).

4.2.2.2. Bone

7 The dominant compartment for Pb in the body is in bone. In human adults, 94% of the total body 

8 burden of Pb is found in the bones, whereas bone Pb accounts for 73% of the body burden in children 

9 (Barry, 1975). Bone is comprised of two main types, cortical (or compact) and trabecular (or spongy or 

10 calcellous). The proportion of cortical to trabecular bone in the human body varies by age, but on average 

11 is about 80 to 20 (ICRP, 1973; Leggett, 1993; O'Flaherty, 1998).

12 The exchange of Pb from plasma to the bone surface is a rapid process (i.e., adult t1/2 =0.19 and 

13 0.23 hours for trabecular and cortical bone, respectively) (Leggett, 1993). Some Pb diffuses from the bone 

14 surface to deeper bone regions (adult t1/2=150 days) where it is relatively inert (in adults) and part of a 

15 “nonexchangeable” pool of Pb in bone (Leggett, 1993).

16 Pb distribution in bone includes uptake into cells that populate bone (e.g., osteoblasts, osteoclasts, 

17 osteocytes) and exchanges with proteins and minerals in the extracellular matrix (Pounds et al., 1991). Pb

18 forms highly stable complexes with phosphate and can replace calcium in the calcium-phosphate salt, 

19 hydroxyapatite, which comprises the primary crystalline matrix of bone (Brès et al., 1986; Miyake, 1986; 

20 Verbeeck et al., 1981). Several intracellular kinetic pools of Pb have been described in isolated cultures of 

21 osteoblasts and osteoclasts which appear to reflect physiological compartmentalization within the cell, 

22 including membranes, mitochondria, soluble intracellular binding proteins, mineralized Pb (i.e., 

23 hydroxyapatite) and inclusion bodies (Long et al., 1990; Pounds & Rosen, 1986; Rosen, 1983). 

24 Approximately 70-80% of Pb taken up into isolated primary cultures of osteoblasts or osteocytes is 

25 associated with mitochondria and mineralized Pb (Pounds et al., 1991).

26 Pb accumulates in bone regions having the most active calcification at the time of exposure. Pb

27 accumulation is thought to occur predominantly in trabecular bone during childhood and in both cortical 

28 and trabecular bone in adulthood (Aufderheide & Wittmers, 1992). Early Pb uptake in children is greater 

29 in trabecular bone due to its larger surface area and higher metabolic rate. With continued exposure, Pb

30 concentrations in bone may increase with age throughout the lifetime beginning in childhood, indicative 

31 of a relatively slow turnover of Pb in adult bone (Barry, 1975; Barry & Connolly, 1981; Gross et al., 

32 1975; Park, Mukherjee, et al., 2009; Schroeder & Tipton, 1968). The cortical and trabecular bones have 
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1 distinct rates of turnover and Pb release. For example, tibia has a turnover rate of about 2% per year 

2 whereas trabecular bone has a turnover rate of more than 8% per year in adults (M. B. Rabinowitz, 1991).

3 A high bone formation rate in early childhood results in the rapid uptake of circulating Pb into 

4 mineralizing bone; however, bone Pb is also recycled to other tissue compartments or excreted in 

5 accordance with a high bone resorption rate (O'Flaherty, 1995). Thus, most of the Pb acquired early in life 

6 is not permanently fixed in the bone (60-65%) (ICRP, 1973; Leggett, 1993; O'Flaherty, 1995). However, 

7 some Pb accumulated in bone does persist into later life. McNeill et al. (2000) compared tibia Pb levels 

8 and cumulative blood Pb indices in a population of 19- to 29-year-olds who had been highly exposed to 

9 Pb in childhood from the Bunker Hill, Idaho smelter; they concluded that Pb from exposure in early 

10 childhood had persisted in the bone matrix until adulthood.

11 A key factor affecting Pb uptake into bone is the fraction of bone surface in trabecular and cortical 

12 bone adjacent to active bone marrow. Of the total bone surface against red marrow, 76% is trabecular and 

13 24% is cortical endosteal (Salmon et al., 1999). The fraction of total bone marrow that is red and active 

14 decreases from 100% at birth to about 32% in adulthood (Cristy, 1981). However, bone marrow has much 

15 lower Pb concentrations than bone matrix (Skerfving et al., 1983). 

4.2.2.3. Soft Tissues

16 Most of the Pb in soft tissue is in liver and kidney (Barry, 1975; Gerhardsson et al., 1986; 

17 Gerhardsson et al., 1995; Gross et al., 1975; Oldereid et al., 1993). Pb in these soft tissues (i.e., kidney, 

18 liver, and brain) exists predominantly bound to protein. High affinity cytosolic Pb-binding proteins have 

19 been identified in rat kidney and brain (DuVal & Fowler, 1989; Fowler, 1989). The Pb-binding proteins in 

20 rat are cleavage products of α2μ globulin, a member of the protein superfamily known as retinol-binding 

21 proteins that are generally observed only in male rats (Fowler & DuVal, 1991). Other high-affinity 

22 Pb-binding proteins (Kd ~14 nM) have been isolated in human kidney, two of which have been identified 

23 as a 5 kDa peptide, thymosin 4 and a 9 kDa peptide, acyl-CoA binding protein (D. R. Smith et al., 1998). 

24 Pb also binds to metallothionein, but does not appear to be a significant inducer of the protein in 

25 comparison with the inducers Cd and Zn (Eaton et al., 1980; Waalkes & Klaassen, 1985).

26 The liver and kidneys rapidly accumulate systemic Pb (t1/2=0.21 and 0.41 hours, respectively), 

27 which amounts to 10-15% and 15-20% of intravenously injected Pb, respectively (Leggett, 1993). A 

28 linear relationship in dose-tissue Pb concentrations for kidney and liver has been demonstrated in swine, 

29 dogs, and rats (Azar et al., 1973; Casteel et al., 1997; Casteel et al., 2006; D. M. Smith et al., 2008). In 

30 contrast to Pb in bone, which accumulates Pb with continued exposure in adulthood, concentrations in 

31 soft tissues (e.g., liver and kidney) are relatively constant in adults (Barry, 1975; Treble & Thompson, 

32 1997), reflecting a faster turnover of Pb in soft tissue relative to bone.
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4.2.2.4. Fetus

1 Evidence for maternal-to-fetal transfer of Pb in humans is derived from cord blood to maternal 

2 blood Pb ratios. These ratios range from about 0.6 to 1.0 at the time of delivery (Carbone et al., 1998; 

3 Goyer, 1990; Graziano et al., 1990; B. Gulson, Jameson, et al., 1998; Kordas et al., 2009; Manton, 1985). 

4 In addition, the similarity of isotopic ratios in maternal blood and in blood and urine of newly-born 

5 infants provide further evidence for placental transfer of Pb to the fetus (B. Gulson et al., 1999). 

6 Transplacental transfer of Pb may be facilitated by an increase in the plasma/blood Pb 

7 concentration ratio during pregnancy (Lamadrid-Figueroa et al., 2006; Montenegro et al., 2008). 

8 Maternal-to-fetal transfer of Pb appears to be related partly to the mobilization of Pb from the maternal 

9 skeleton. Evidence for transfer of maternal bone Pb to the fetus has been provided by stable Pb isotope 

10 studies in cynomolgus monkeys exposed during pregnancy. Approximately 7-39% of the maternal Pb 

11 burden transferred to the fetus was derived from the maternal skeleton, with the remainder derived from 

12 contemporaneous exposure (Franklin et al., 1997; O'Flaherty, 1998).

4.2.2.5. Organic Lead

13 Information on the distribution of Pb in humans following exposures to organic Pb is extremely 

14 limited. However, as reported in the 2006 AQCD, the available evidence demonstrates near complete 

15 absorption following inhalation of tetraalkyl Pb vapor and subsequent transformation to trialkyl Pb 

metabolites. One hour following brief inhalation exposures to 203 16 Pb tetraethyl or tetramethyl Pb (1 

mg/m3), ~50% of the 203 17 Pb body burden was associated with liver and 5% with kidney; the remaining 

203Pb was widely distributed throughout the body (Heard et al., 1979). The kinetics of 203 18 Pb in blood 

19 showed an initial declining phase during the first 4 hours (tetramethyl Pb) or 10 hours (tetraethyl Pb) after 

20 the exposure, followed by a reappearance of radioactivity back into the blood after ~20 hours. The high 

21 level of radioactivity initially in the plasma indicates the presence of tetraalkyl/trialkyl Pb. The 

22 subsequent rise in blood radioactivity, however, probably represents water-soluble inorganic Pb and 

23 trialkyl and dialkyl Pb compounds that were formed from the metabolic conversion of the volatile parent 

24 compounds (Heard et al., 1979).

25 Alkyl Pb compounds undergo oxidative dealkylation catalyzed by cytochrome P450 in liver and, 

26 possibly, in other tissues. Trialkyl Pb metabolites have been found in the liver, kidney, and brain 

27 following exposure to the tetraalkyl compounds in workers (Bolanowska et al., 1967); these metabolites 

28 have also been detected in brain tissue of nonoccupational subjects (Nielsen et al., 1978).

4.2.3. Elimination

29 The rapid-phase (30-40 days) of Pb excretion amounts to 50-60% of the absorbed fraction 

30 (Chamberlain et al., 1978; Kehoe, 1961a, 1961b, 1961c; M. B. Rabinowitz et al., 1976). Absorbed Pb is 
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1 excreted primarily in urine and feces, with sweat, saliva, hair, nails, and breast milk being minor routes of 

2 excretion (Chamberlain et al., 1978; Griffin et al., 1975; Hursh et al., 1969; Hursh & Suomela, 1968; 

3 Kehoe, 1987; M. B. Rabinowitz et al., 1976).

4 Approximately 30% of Pb excretion during the first 20 days after exposure is due to urinary and 

5 fecal losses (Leggett, 1993). The kinetics of urinary excretion following a single dose of Pb is similar to 

6 that of blood (Chamberlain et al., 1978), likely due to the fact that Pb in urine derives largely from Pb in 

7 plasma. Evidence for this is the observation that urinary Pb excretion is strongly correlated with the 

8 rate of glomerular filtration of Pb (Araki et al., 1986) and plasma Pb concentration (Bergdahl, Schutz, et 

9 al., 1997) (i.e., glomerular filtration rate × plasma Pb concentration), and both relationships are linear. 

10 While the relationship between urinary Pb excretion and plasma Pb concentration has been shown to be 

11 linear, the plasma Pb relationship to blood Pb concentration is curvilinear (as described in Section 4.2.2.1 

12 and demonstrated in Figure 4-5). This contributes to an increase in the renal clearance of Pb from blood

13 with increasing blood Pb concentrations (Chamberlain, 1983). Similarly, a linear relationship between 

14 plasma Pb concentration and urinary excretion rate predicts a linear relationship between Pb intake (at 

15 constant absorption fraction) and urinary Pb excretion rate, whereas the relationship with blood Pb 

16 concentration would be expected to be curvilinear (Section 4.2.7). 

17 Estimates of urinary filtration of Pb from serum (or plasma) range from 13-22 L/day, with a mean 

18 of 18 L/day (Araki et al., 1986; Chamberlain et al., 1978; Manton & Cook, 1984; Manton & Malloy, 

19 1983), which corresponds to half-time for transfer of Pb from plasma to urine of 0.1-0.16 days for a 70-kg 

20 adult who has a plasma volume of ~3 L. The rate of urinary excretion of Pb was less than the rate of 

21 glomerular filtration of ultrafilterable Pb, suggesting that urinary Pb is the result of incomplete renal 

22 tubular reabsorption of Pb in the glomerular filtrate (Araki et al., 1986); although, net tubular secretion of 

23 Pb has been demonstrated in animals (Vander et al., 1977; Victery et al., 1979). On the other hand, 

24 estimates of blood-to-urine clearance range from 0.03-0.3 L/day with a mean of 0.18 L/day (Araki et al., 

25 1990; Berger et al., 1990; Chamberlain et al., 1978; Koster et al., 1989; Manton & Malloy, 1983; M. B. 

26 Rabinowitz et al., 1973; Ryu et al., 1983) (Diamond, 1992), consistent with a plasma Pb to blood Pb 

27 concentration ratio of ~0.005–0.01 L/day (Klotzback et al., 2003). Based on the above differences, 

28 urinary excretion of Pb can be expected to reflect the concentration of Pb in plasma and variables that 

29 affect delivery of Pb from plasma to urine (e.g., glomerular filtration and other transfer processes in the 

30 kidney).

31 The value for fecal:urinary excretion ratio (~0.5) was observed during days 2-14 following 

32 intravenous injection of Pb in humans (Booker et al., 1969; Chamberlain et al., 1978; Hursh et al., 1969). 

33 This ratio is slightly higher (0.7-0.8) with inhalation of submicron Pb-bearing PM due to ciliary clearance 

34 and subsequent ingestion. The transfer of Pb from blood plasma to the small intestine by biliary secretion 

35 in the liver is rapid (adult t1/2 = 10 days), and accounts for 70% of the total plasma clearance (O'Flaherty, 

36 1995).
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Organic Lead 

1 Pb absorbed after inhalation of tetraethyl and tetramethyl Pb is excreted in exhaled air, urine, and 

2 feces (Heard et al., 1979). Fecal:urinary excretion ratios were 1.8 following exposure to tetraethyl Pb and 

3 1.0 following exposure to tetramethyl Pb (Heard et al., 1979). Occupational monitoring studies of 

4 workers who were exposed to tetraethyl Pb have shown that tetraethyl Pb is excreted in the urine as 

5 diethyl Pb, ethyl Pb, and inorganic Pb (Turlakiewicz & Chmielnicka, 1985; Vural & Duydu, 1995; W. 

6 Zhang et al., 1994).

4.3. Lead Biomarkers

7 This section describes the biological measurements of Pb and their interpretation as indicators of 

8 exposure or body burden. The focus is on blood Pb and bone Pb and the interplay between them, as these 

9 are the most commonly measured biomarkers in recent epidemiologic and toxicological studies. 

10 Mechanistic models are used throughout the section as a means to describe basic concepts that 

11 derive from the wealth of information on Pb toxicokinetics. Although predictions from models are 

12 inherently uncertain, models can serve to illustrate expected interrelationships between Pb intake and

13 tissue distribution that are important in interpreting human clinical and epidemiologic studies. Thus, 

14 models serve as the only means we have for synthesizing our extensive, but incomplete, knowledge of Pb 

15 biokinetics into a holistic representation of Pb biokinetics. Furthermore, models can also be used to make 

16 predictions about biokinetics relationships that have not been thoroughly evaluated in experiments or 

17 epidemiologic studies. In this way, models can serve as heuristic tools for shaping data collection to 

18 improve our understanding of Pb biokinetics.

19 Numerous mechanistic models of Pb biokinetics in humans have been proposed, and these are 

20 described in the 2006 Pb AQCD (U.S. EPA, 2006) and in the supporting literature cited in that report. In 

21 this section, for simplicity and for internal consistency, we have limited the discussion to predictions from 

22 a single model, the ICRP Pb biokinetics model (ICRP, 1994; Leggett, 1993; Pounds & Leggett, 1998). 

23 This model was originally developed for the purpose of supporting radiation dosimetry predictions; 

24 however, it has also been applied in Pb risk assessment (Abrahams et al., 2006; Khoury & Diamond, 

25 2003; Lorenzana et al., 2005). Portions of the model have been incorporated into an All Ages Lead Model 

26 (AALM) that is being developed by EPA (2005).

4.3.1. Bone Lead Measurements

27 For Pb measurements in bone, the most commonly examined bones are the tibia, calcaneus, patella, 

28 and finger bone. For cortical bone, the midpoint of the tibia is measured. For trabecular bone, both the 
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1 patella and calcaneus are measured. The tibia consists of more than 95% cortical bone, the calcaneus and 

2 patella comprise more than 95% trabecular bone, and finger bone is a mixed cortical and trabecular bone 

3 although the second phalanx is dominantly cortical. Recent studies favor measurement of the patella, 

4 because it has more bone mass and may afford better measurement precision than the calcaneus. Bone 

5 analysis methods have included flame atomic absorption spectrometry (AAS), anode stripping 

6 voltammetry (ASV), inductively coupled plasma atomic emission spectroscopy (ICP-AES), inductively 

7 coupled plasma mass spectrometry (ICP-MS), laser ablation inductively coupled plasma mass 

8 spectrometry (LA-ICP-MS), thermal ionization mass spectrometry (TIMS), synchrotron radiation induced 

9 X-ray emission (SRIXE), particle induced X-ray emission (PIXE), and X-ray fluorescence (XRF). The 

10 upsurge in popularity of the XRF method has paralleled a decline in the use of the other methods. More 

11 information on the precision, accuracy, and variability in bone Pb measurements can be found in the 2006 

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

13 Two main approaches for XRF measurements have been used to measure Pb concentrations in 

14 bone, the K-shell and L-shell methods. The K-shell method is the most widely used, as there have been 

15 relatively few developments in L-shell devices since the early 1990s. However, Nie et al. (2011) recently 

16 reported on the use of a new portable L-shell device for human in vivo Pb measurements. Advances in L17 shell device technology resulted in much higher sensitivity than previous L-shell devices. The new L18 shell device showed sensitivity similar to that of K-shell methods and a high correlation with results 

19 obtained from K-shell methods (intraclass correlation = 0.65).

20 Bone Pb measurements are typically expressed in units of Pb/g bone mineral. This may potentially 

21 introduce variability into the bone Pb measurements related to variation in bone density. Typically, 

22 potential associations between bone density and bone Pb concentration are not evaluated in epidemiologic

23 studies (Hu et al., 2007; Theppeang et al., 2008).

4.3.2. Blood Lead Measurements

24 Analytical methods for measuring Pb in blood include AAS, graphite furnace atomic absorption 

25 spectrometry (GFAAS), ASV, ICP-AES, and ICP-MS. GFAAS and ASV are generally considered to be 

26 the methods of choice (Flegal & Smith, 1995). Limits of detection for Pb using AAS are on the order of 

27 5-10 μg/dL for flame AAS measurements and approximately 0.1 μg/dL for flameless AAS measurements 

28 (Flegal & Smith, 1995; NIOSH, 1994). A detection limit of 0.005 μg/dL has been achieved for Pb in 

29 blood samples analyzed by GFAAS.

30 For measurement of Pb in plasma, ICP-MS provides sufficient sensitivity (Schutz et al., 1996). 

31 While the technique has been applied to assessing Pb exposures in adults, it has not received widespread 

32 use in epidemiologic studies.
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1 The primary binding ligand for Pb in RBC, ALAD, is encoded by a single gene in humans that is 

2 polymorphic in two alleles (ALAD1 and ALAD2) (Scinicariello et al., 2007). Since the ALAD1 and 

3 ALAD2 alleles can be codominantly expressed, 3 different genotypes (ALAD 1-1, ALAD 1-2, and ALAD 

4 2-2) are possible. The ALAD 1-1 genotype is the most common. Scinicariello et al. (2010) tested 

5 genotypes in civilian, noninstitutionalized U.S. individuals that participated as part of NHANES III from 

6 1988–1994 and found that 15.6% of non-Hispanic whites, 2.6% non-Hispanic blacks, and 8.8% Mexican 

7 Americans carried the ALAD2 allele.

8 The 2006 AQCD document reports that many studies have shown that, with similar exposures to 

9 Pb, individuals with the ALAD-2 allele have higher blood Pb levels than those without (Astrin et al., 

10 1987; Bergdahl, Schutz, et al., 1997; H.-S. Kim et al., 2004; Pérez-Bravo et al., 2004; C. M. Smith et al., 

11 1995; Wetmur, 1994; Wetmur, Lehnert, et al., 1991). More recent meta analyses provide further support 

12 for ALAD2 carriers having higher blood Pb levels than ALAD1-1 homozygotes (Scinicariello et al., 

13 2007; Zhao et al., 2007). The mechanism for this association may be higher Pb binding affinity of 

14 ALAD2. Although, this would be consistent with the structural differences that result in greater 

15 electronegativity of ALAD1 compared to ALAD2 (Wetmur, 1994; Wetmur, Kaya, et al., 1991), 

measurements of Pb binding affinity to ALAD1 and ALAD2 (i.e., Pb2+ displacement of Zn2+ 16 binding to 

17 recombinant ALAD1 and ALAD2) have not revealed differences in Pb binding affinity (Jaffe et al., 2000). 

18 Both analyses reported the greatest differences for ALAD2 compared to ALAD1 in highly exposed adults 

19 and little difference among environmentally-exposed adults; large differences were also observed for 

20 children at low exposures. However, there were few studies that evaluated children and the largest study 

21 contributing to the meta analysis may have been influenced by selection bias (Scinicariello et al., 2007). 

22 Individual studies find similar results, with blood Pb levels being higher in individuals with ALAD2 

23 alleles (Miyaki et al., 2009; Shaik & Jamil, 2009). A subsequent meta analysis of adult data from 

24 NHANES III did not find any differences in blood Pb level between all carriers of either the ALAD 1-1 or 

25 ALAD 1-2/2-2 allele (Scinicariello et al., 2010). Other studies provide further support for no blood Pb 

26 differences among ALAD1 and ALAD2 carriers (Montenegro et al., 2006; Rabstein et al., 2008; 

27 Wananukul et al., 2006) or lower blood Pb levels for individuals with ALAD 1-2/2-2 (Chia et al., 2006).

28 Analyses of serial blood Pb concentrations measured in longitudinal epidemiologic studies have 

29 found relatively strong correlations (e.g., r = 0.5-0.8) between individual blood Pb concentrations 

30 measured after 6-12 months of age (Dietrich et al., 1993; McMichael et al., 1988; Otto et al., 1985; M. 

31 Rabinowitz et al., 1984; Schnaas et al., 2000). These observations suggest that, in general, exposure 

32 characteristics of an individual child (e.g., exposure levels and/or exposure behaviors) tend to be 

33 relatively constant across age. However, a single blood Pb measurement may not distinguish between a 

34 history of long-term lower-level Pb exposure from a history that includes higher acute exposures 

35 (Mushak, 1998). This is illustrated in Figure 4-6. Two hypothetical children are simulated. Child A has a 

36 relatively constant Pb intake from birth, whereas Child B has the same long-term Pb intake as Child A, 
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1 but with a 1-year elevated intake beginning at age 24 months (Figure 4-6, upper panel). The absorption 

2 fraction is assumed to be the same for both children. Blood Pb samples 1 and 5 for Child A and B, or 2 

3 and 4 for Child B, will yield similar blood Pb concentrations (~3 or 10 μg/dL, respectively), yet the 

4 exposure contexts for these samples are very different. Two samples (e.g., 1 and 2, or 4 and 5), at a 

5 minimum, are needed to ascertain if the blood Pb concentration is changing over time. The rate of change 

6 can provide information about the magnitude of change in exposure, but not necessarily about the time 

7 history of the change (Figure 4-6, lower panel). Time-integrated measurements of Pb concentration may 

8 provide a means for accounting for some of these factors and, thereby, provide a better measure of long9 term Pb exposure.
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Figure 4-6. Simulation of temporal relationships between Pb exposure 

and blood Pb concentration in children.Child A and Child B 

have a relatively constant basal Pb intake (μg/day/kg body 

weight) from birth; Child B experiences 1-year elevated intake 

beginning at age 24 months (upper panel). Blood Pb samples 

1 and 5 for Child A and B, or 2 and 4 for Child B, will yield 

similar blood Pb concentrations (~3 or 10 μg/dL, respectively), 

yet the exposure scenarios for these samples are very 

different. As shown in the example of Child C and Child D, two 

samples can provide information about the magnitude of 

change in exposure, but not necessarily the temporal history 

of the change (lower panel). Simulation based on ICRP Pb 

biokinetics model (Leggett, 1993).
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4.3.3. Urine Lead Measurements

1 Standard methods that have been reported for urine Pb analysis are, in general, the same as those 

2 analyses noted for determination of Pb in blood. Reported detection limits are ~50 μg/L for AAS, 5-10 

3 μg/L for ICP AES, and 4 μg/L for ASV for urine Pb analyses.

4 The concentration of Pb in urine is a function of the urinary Pb excretion (Section 4.2.3) and the 

5 urine flow rate. Urine flow rate requires collection of a timed urine sample, which is often problematic in 

6 epidemiologic studies. Collection of untimed (“spot”) urine samples, a common alternative to timed 

7 samples, requires adjustment of the Pb measurement in urine to account for variation in urine flow 

8 (Diamond, 1988). Several approaches to this adjustment have been explored, including adjusting the 

9 measured urine Pb concentration by the urine creatinine concentration, urine osmolality, or specific 

10 gravity (Araki et al., 1990; Fukui et al., 1999). Urine flow rate can vary by a factor or more than 10, 

11 depending on the state of hydration and other factors that affect glomerular filtration rate and renal tubular 

12 reabsorption of the glomerular filtrate. All of these factors can be affected by Pb exposure at levels that 

13 produce nephrotoxicity (i.e., decreased glomerular filtration rate, impaired renal tubular transport 

14 function). Therefore, urine Pb concentration measurements provide little reliable information about 

15 exposure (or Pb body burden), unless they can be adjusted to account for unmeasured variability in urine 

16 flow rate (Araki et al., 1990).

17 Urinary Pb concentration reflects, mainly, the exposure history of the previous few months; thus, a 

18 single urinary Pb measurement cannot distinguish between a long-term low level of exposure or a higher 

19 acute exposure. Urinary Pb measurements would be expected to correlate with concurrent blood Pb. 

20 Chiang et al. (2008) reported a significant, but relatively weak correlation between urinary Pb levels 

21 (μg/dg creatinine) and individual Pb intakes (μg/day) estimated in a group of 10- to 12-year-old children 

22 (β: 0.053, R = 0.320, p = 0.02, n = 57). A contributing factor to the relatively weak correlation may have 

23 been the temporal displacement between the urine sampling and measurements used to estimate intake, 

24 which may have been as long as six months for some children.

25 Thus, a single urine Pb measurement, or a series of measurements taken over short-time span, is 

26 likely a relatively poor index of Pb body burden for the same reasons that blood Pb is not a good indicator 

27 of body burden. On the other hand, long-term average measurements of urinary Pb can be expected to be 

28 a better index of body burden (Figure 4-7).
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Figure 4-7. Simulation of relationship between urinary Pb excretion and 

body burden in adults.A change in Pb uptake results in a 

relatively rapid change in urinary excretion of Pb, to a new 

quasi-steady state, and a relatively small change in body 

burden (upper panel). The long-term average urinary Pb 

excretion more closely tracks the pattern of change in body 

burden (lower panel). Simulation based on ICRP Pb 

biokinetics model (Leggett, 1993).

0

5

10

15

20

0

2

4

6

8

10

25 30 35 40 45 50 55 60 65 70

Body Burden (mg)

Urine Pb (

μg/day)

Age (year)

Urine

Body

0

5

10

15

20

0

1

2

3

4

5

6

25 30 35 40 45 50 55 60 65 70

Body Burden (mg)

Time-Averaged Urine Pb

(μg/day)

Age (year)

Urine

Body
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4.3.4. Lead in Other Potential Biomarkers

1 There was extensive discussion in the 2006 Pb AQCD regarding the utility of other Pb biomarkers 

2 as indicators of exposure or body burden. Due to the fact that most epidemiologic studies continue to use 

3 blood Pb or bone Pb as biomarkers of exposure or body burden, and other potential biomarkers (i.e., teeth, 

4 hair, and saliva) have not been established to the same extent as blood or bone Pb, only summaries are 

5 provided below.

4.3.4.1. Teeth

6 Tooth Pb is a minor contributor to the total body burden of Pb. As teeth accumulate Pb, tooth Pb 

7 levels are generally considered an estimate of cumulative Pb exposure. The tooth Pb-blood Pb 

8 relationship is more complex than the bone Pb-blood Pb relationship because of differences in tooth type, 

9 location, and analytical method. Although mobilization of Pb from bone appears well established, this is 

10 not the case for Pb in teeth. Conventional wisdom has Pb fixed once it enters the tooth. Although that may 

11 be the case for the bulk of enamel, it is not true for the surface of the enamel and dentine (B. L. Gulson et 

12 al., 1997; M. B. Rabinowitz et al., 1993). Limited studies have demonstrated moderate-to-high 

13 correlations between tooth Pb levels and blood Pb levels (M. B. Rabinowitz, 1995; M. B. Rabinowitz et 

14 al., 1989).

15 Teeth are composed of several tissues formed pre- and postnatal. Therefore, if a child’s Pb exposure 

16 during the years of tooth formation varied widely, different amounts of Pb would be deposited at different 

17 rates (M. B. Rabinowitz et al., 1993). This may allow investigators to elucidate the history of Pb exposure 

18 in a child. Robbins et al. (2010) found a significant association between environmental Pb measures that 

19 correlated with leaded gasoline use and tooth enamel Pb in permanent teeth. Costa de Almeida et al. 

20 (2007) was able to discern differences between tooth enamel Pb concentration in biopsy samples from 

21 children who lived in areas having higher or lower levels of Pb contamination. Gulson and Wilson (1994)

22 advocated the use of sections of enamel and dentine to obtain additional information compared with 

23 analysis of the whole tooth (e.g., (Fosse et al., 1995; Tvinnereim et al., 1997). For example, deciduous 

24 tooth Pb in the enamel provides information about in utero exposure whereas that in dentine from the 

25 same tooth provides information about postnatal exposure until the tooth exfoliates at about 6-7 years of 

26 age.

4.3.4.2. Hair

27 The 2006 Pb AQCD discussed applications of hair Pb measurements for assessing Pb body burden 

28 or exposure and noted methodological limitations (e.g., external contamination) and lack of a strong 

29 empirical basis for relating hair Pb levels to body burden or exposure. No new methodological or 
May 2011 4-42 DRAFT – DO NOT CITE OR QUOTE

1 conceptual advances regarding hair Pb measurements have occurred since 2006, and widespread 

2 application of hair Pb measurements in epidemiologic studies has not occurred.

3 Pb is incorporated into human hair and hair roots (Bos et al., 1985; M. B. Rabinowitz et al., 1976)

4 and has been explored as a possibly noninvasive approach for estimating Pb body burden (Gerhardsson et 

5 al., 1995; Wilhelm et al., 1989; Wilhelm et al., 2002). Hair Pb measurements are subject to error from 

6 contamination of the surface with environmental Pb and contaminants in artificial hair treatments (i.e., 

7 dyeing, bleaching, permanents) and are a relatively poor predictor of blood Pb concentrations, particularly 

8 at low levels (<10-12 μg/dL) (Campbell & Toribara, 2001; G. Drasch et al., 1997; Esteban et al., 1999; J. 

9 L. Rodrigues et al., 2008). Temporal relationships between Pb exposure and hair Pb levels, and kinetics of 

10 deposition and retention of Pb in hair have not been evaluated. Although hair Pb measurements have been 

11 used in some epidemiologic studies, an empirical basis for interpreting hair Pb measurements in terms of 

12 body burden or exposure has not been firmly established.

4.3.4.3. Saliva

13 A growing body of literature on the utility of measurements of salivary Pb has developed since the 

14 completion of the 2006 AQCD (U.S. EPA, 2006). Earlier reports suggested a relatively strong correlation 

15 between salivary Pb concentration and blood Pb concentration (Brodeur et al., 1983; Omokhodion & 

16 Crockford, 1991; P'an, 1981); however, more recent assessments have shown relatively weak or 

17 inconsistent associations (Barbosa, Heloisa, et al., 2006; Costa de Almeida et al., 2009; Nriagu et al., 

18 2006). The differences in these outcomes may reflect differences in blood Pb concentrations, exposure 

19 history and/or dental health (i.e., transfer of Pb between dentin and saliva) and possibly methods for 

20 determining Pb in saliva. Barbosa et al. (2006) found a significant but relatively weak correlation 

21 (log[blood PB] versus log[saliva Pb], r = 0.277, p = 0.008) in a sample of adults, ages 18-60 years (n =

22 88). The correlation was similar for salivary and plasma Pb. Nriagu et al. (2006) found also found a 

relatively weak association (R2 23 = 0.026) between blood Pb (μg/dL) and salivary Pb (μg/L) in a sample of 

24 adults who resided in Detroit, MI (n = 904). Costa de Almeida et al. (2009) found a significant correlation 

25 between salivary and blood Pb concentrations in children in a Pb-contaminated region in Sao Paulo State, 

26 Brazil (r = 0.76. p = 0.04, n = 7) prior to site remediation; however, the correlation degenerated (r = 0.03, 

27 p = 0.94, n = 9) following remediation. Nevertheless, salivary Pb concentrations in the group of children 

28 who lived in the contaminated area were significantly elevated compared to a reference population. It is 

29 possible, that salivary Pb measurements may be a useful non-invasive biomarker for detecting elevated Pb 

30 exposure; however, it is not clear based on currently available data, if salivary Pb measurements would be 

31 a more reliable measure of exposure than blood Pb measurements.
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4.3.4.4. Serum δ-ALA and ALAD

1 The association between blood Pb and blood ALAD activity and serum δ-aminolevulinc acid (δ2 ALA) levels was recognized decades ago as having potential use as a biomarker of Pb exposure 

3 (Hernberg et al., 1970; Mitchell et al., 1977). More recently reference values for blood ALAD activity 

4 ratio (the ratio of ALAD activity in the blood sample to that measured after fully activating the enzyme in 

5 the sample) have been reported (Gultepe et al., 2009). Inhibition of erythrocyte ALAD by Pb results in a 

6 rise in the plasma concentration of the ALAD substrate δ-ALA. The δ-ALA biomarker can be measured 

7 in serum and has been used as a surrogate for Pb measurements in studies in which whole blood samples 

8 or adequately prepared plasma or serum samples were not available for Pb measurements (Opler et al., 

9 2004; Opler et al., 2008).

4.3.5. Relationship between Lead in Blood and Lead in Bone

10 The kinetics of elimination of Pb from the body reflects the existence of fast and slow pools of Pb 

11 in the body. The dominant phase of Pb kinetics in the blood, exhibited shortly after a change in exposure 

12 occurs, has a half-life of ~20-30 days (Leggett, 1993). A slower phase becomes evident with longer 

13 observation periods following a decrease in exposure. Slow transfer rates for the movement of Pb from 

14 nonexchangeable bone pools to the plasma are the dominant transfer process determining long-term 

15 accumulation and elimination of bone Pb burden. 

16 Bone Pb stores can contribute 40-70% to blood Pb (B. Gulson et al., 1995; Manton, 1985; D. R. 

17 Smith et al., 1996). Bone Pb burdens in adults are slowly lost by diffusion (heteroionic exchange) as well 

18 as by resorption (O'Flaherty, 1995). Half-times for the release of Pb in bone is dependent on age and 

19 intensity of exposure. Slow bone volume compartments are much more labile in infants and children than 

20 in adults as reflected by half-times for movement into the plasma (e.g., cortical t1/2 = 0.23 years at birth, 

21 1.2 years at 5 years of age, 3.7 years at 15 years of age, and 23 years in adults; trabecular t1/2 = 0.23 years 

22 at birth, 1.0 years at 5 years of age, 2.0 years at 15 years of age, and 3.9 years in adults) (Leggett, 1993). 

23 Children who have been removed from a relatively brief exposure to elevated environmental Pb exhibit 

24 faster slow-phase kinetics than children removed from exposures that lasted several years, with half-times 

25 of 10 and 20-38 months, respectively (Manton et al., 2000). The longer half-times measured under the 

26 latter conditions reflect the contribution of bone Pb stores to blood Pb following a change in exposure.

27 The longer half-life of Pb in bone compared to blood Pb, allows a more cumulative measure of Pb 

28 dose. Pb in adult bone can serve to maintain blood Pb levels long after external exposure has ceased 

29 (Fleming et al., 1997; Inskip et al., 1996; Kehoe, 1987; O'Flaherty et al., 1982; D. R. Smith et al., 1996), 

30 even for exposures that occurred during childhood (F. E. McNeill et al., 2000). The more widespread use 

31 of in vivo XRF Pb measurements in bone and indirect measurements of bone processes with stable Pb 

32 isotopes have enhanced the use of bone Pb as a biomarker of Pb body burden.
May 2011 4-44 DRAFT – DO NOT CITE OR QUOTE

1 There is a stronger relationship between patella Pb and blood Pb than tibia Pb and blood Pb 

2 (Hernandez-Avila et al., 1996; Hu et al., 1996; Hu et al., 1998; Park, Mukherjee, et al., 2009). Hu et al. 

3 (1998) suggest that trabecular bone is the predominant bone type providing Pb back into circulation under 

4 steady-state and pathologic conditions. The stronger relationships between blood Pb and trabecular Pb 

5 compared with cortical bone is probably associated with the larger surface area of trabecular bone 

6 allowing for more Pb to bind via ion exchange mechanisms and more rapid turnover making it more 

7 sensitive to changing patterns of exposure.

4.3.5.1. Children

8 As mentioned in Section 4.2.2.2, bone growth in children will contribute to accumulation of Pb in 

9 bone, which will comprise most of the Pb body burden. As a result, Pb in bone will more closely reflect 

10 Pb body burden than blood Pb. However, changes in blood Pb concentration in children (i.e., associated 

11 with changing exposures) are thought to more closely parallel changes in total body burden. Figure 4-8

12 shows a simulation of the temporal profile of Pb in blood and bone in a child who experiences a period of

13 constant Pb intake (from age 2-5) via ingestion (μg Pb/day) followed by an abrupt decline in intake. The 

14 figure illustrates several important general concepts about the relationship between Pb in blood and bone. 

15 While blood Pb approaches a quasi-steady state after a period of a few months with a constant rate of Pb 

16 intake (as demonstrated by the vertical dashed line), Pb continues to accumulate in bone with continued 

17 Pb intake after the quasi-steady state is achieved in blood. The model also predicts that the rate of release 

18 of Pb from bone after cessation of exposure is faster than in adults. This difference has been attributed to 

19 accelerated growth-related bone mineral turnover in children, which is the primary mechanism for release 

20 of Pb that has been incorporated in to the bone mineral matrix.

21 Empirical evidence in support of this comes from longitudinal studies in which relatively high 

22 correlations (r = 0.85) were found between concurrent (r = 0.75) or lifetime average blood Pb 

23 concentrations (r = 0.85) and tibia bone Pb concentrations (measured by XRF) in a sample of children in 

24 which average blood Pb concentrations exceeded 20 μg/dL; the correlations was much weaker (r <0.15) 

25 among children who had average blood Pb concentration <10 μg/dL (Wasserman et al., 1994).

26 Two alternative blood Pb metrics depicted in Figure 4-8 include the time-averaged and time27 integrated blood Pb concentrations. Both the time-averaged and time-integrated blood Pb metrics display 

28 rates of change in response to the exposure event that more closely approximate the slower kinetics of 

29 bone Pb and body burden, than the kinetics of blood Pb concentration, with notable differences. The time30 averaged blood Pb concentration increases during the exposure event and decays following the event, 

31 consistent with the changing body burden. The time-integrated blood Pb concentration (conceptually 

32 identical to cumulative blood lead index [CBLI] used in epidemiologic studies) is a cumulative function 

33 and increases throughout childhood; however, the slope of the increase is higher during the exposure 
May 2011 4-45 DRAFT – DO NOT CITE OR QUOTE

1 event than prior to or following the event. Following cessation of exposure, the time-integrated blood Pb 

2 and body burden diverge. This is expected, as the time-integrated blood Pb curve is a cumulative function 

3 which cannot decrease over time and bone Pb levels will decrease with cessation of exposure.

4 The time-integrated blood Pb concentration will be a better reflection of the total amount of Pb that 

5 has been absorbed, than the body burden at any given time. The time-integrated blood Pb concentration 

6 will also reflect cumulative Pb absorption, and cumulative exposure if the absorption fraction is constant. 

7 This is illustrated in the hypothetical simulations of an exposure event experienced by a child (Figure 4-

8 9). This pattern is similar for adults.
May 2011 

Figure 4-8. Simulation of relationship between blood Pb concentration 

and body burden in c

age 2 to 5.

burden more closely in children than in adults, due to more 

rapid turnover of bone and bone

panel). Nevertheless, the time

concentration more closely tracks t

body burden (middle panel). The time

concentration increases over time (lower panel). Simulation 

based on ICRP Pb biokinetics model

4-46 DRAFT 

Simulation of relationship between blood Pb concentration 

burden in children, with a constant Pb intake from 

age 2 to 5.Blood Pb concentration is thought to parallel body 

burden more closely in children than in adults, due to more 

rapid turnover of bone and bone-Pb stores in children (upper 

panel). Nevertheless, the time-averaged blood Pb 

concentration more closely tracks the pattern of change in 

body burden (middle panel). The time-integrated blood Pb 

concentration increases over time (lower panel). Simulation 

based on ICRP Pb biokinetics model (Leggett, 1993). 

DO NOT CITE OR QUOTE 

Simulation of relationship between blood Pb concentration 

hildren, with a constant Pb intake from 

Blood Pb concentration is thought to parallel body 

burden more closely in children than in adults, due to more 

Pb stores in children (upper 

he pattern of change in 

integrated blood Pb 

concentration increases over time (lower panel). Simulation 
May 2011 4-47 DRAFT – DO NOT CITE OR QUOTE

Figure 4-9. Simulation of relationship between time-integrated blood Pb 

concentration and cumulative Pb absorption in children. The 

simulations include a 3-year period of elevated Pb intake 

during ages 2-5 years. The time-integrated blood Pb 

concentration closely parallels cumulative Pb absorption. 

Simulation based on ICRP Pb biokinetics model (Leggett, 

1993). 

4.3.5.2. Adults

1 In adults, where a relatively large fraction of the body burden residing in bone has a slower 

2 turnover compared to blood, a constant Pb uptake (or constant intake and fractional absorption) gives rise 

3 to a quasi-steady state blood Pb concentration, while the body burden continues to increase over a much 

4 longer period, largely as a consequence of continued accumulation of Pb in bone. This pattern is 

5 illustrated in Figure 4-10 which depicts a hypothetical simulation of an exposure event consisting of a 20-

6 year period of daily ingestion of Pb in an adult. The exposure event shown in the simulations gives rise to 

7 a relatively rapid increase in blood Pb concentration, to a new quasi-steady state, achieved in ~75-100 

8 days (i.e., approximately 3-4 times the blood elimination half-life). In contrast, the body burden continues 

9 to increase during this period. Following cessation of the exposure, blood Pb concentration declines 

10 relatively rapidly compared to the slower decline in body burden. Careful examination of the simulation 

11 shown in Figure 4-10 reveals that the accumulation and elimination phases of blood Pb kinetics are not 

12 symmetrical; elimination is slower than accumulation as a result of the gradual release of bone Pb stores 

0

5

10

15

20

0

10

20

30

40

50

0 2 4 6 8 10

Cumulative Absorption (mg)

Time-Integrated Blood Lead

(μg/dL x yr )

Age (year)

Blood

Absorption
May 2011 4-48 DRAFT – DO NOT CITE OR QUOTE

1 to blood. This response, known as the prolonged terminal elimination phase of Pb from blood, has been 

2 observed in retired Pb workers and in workers who continued to work after improved industrial hygiene 

3 standards reduced their exposures. In the adult simulation shown in Figure 4-10, the initial phase of 

4 elimination (the first 5 years following cessation of exposure at 50 years) has a half-time of approximately 

5 14 years; however, the half-time increases to approximately 60 years during the period 5-30 years after 

6 cessation of exposure. These model predictions are consistent with the slow elimination of Pb from blood 

7 and elimination half-times of several decades for bone Pb (e.g., 16-98 years) that have been estimated 

8 from observations made on Pb workers (Fleming et al., 1997; Gerhardsson et al., 1995). Based on this 

9 hypothetical simulation, a blood Pb concentration measured 1 year following cessation of a period of 

10 increased Pb uptake would show little or no appreciable change from prior to the exposure event, 

11 whereas, the body burden would remain elevated. This illustrates how a single blood Pb concentration 

12 measurement, or a series of measurements taken over a short-time span, could be a relatively poor index 

13 of Pb body burden. 

14 One important potential implication of the profoundly different kinetics of Pb in blood and bone is

15 that, for a constant Pb exposure, bone Pb will increase with increasing duration of exposure and, 

16 therefore, with age. In contrast, blood Pb will achieve a quasi-steady state. As a result, the relationship 

17 between blood Pb and bone Pb will diverge with increasing exposure duration and age. This divergence 

18 can impart different degrees of age-confounding when either blood Pb or bone Pb is used as an internal 

19 dose metric in dose-response models. In a review of epidemiologic studies that evaluated the associations 

20 between blood Pb, bone Pb and cognitive function, the effects of bone Pb were more pronounced than 

21 blood Pb (particularly for longitudinal studies) for older individuals with environmental Pb exposures and 

22 low blood Pb levels (Shih et al., 2007). In contrast, occupational workers with high current Pb exposures 

23 had the strongest associations for blood Pb levels with cognitive function, thus providing evidence for this 

24 divergence (Shih et al., 2007).
May 2011 

Figure 4-10. Simulation of relationship between blood Pb concentration, 

bone Pb and body burden in adults.

results in a quasi

body burden (upper panel). A change in Pb uptake gives rise 

to a relatively rapid change in blood Pb, to a new quasi

state, and a slower change in body burden. The long

time-averaged blood Pb co

slower pattern of change in body burden (middle panel). The 

time-integrated blood Pb concentration increases over the 

lifetime, with a greater rate of increase during periods of 

higher Pb 

biokinetics model

1 Tibia bone Pb has been shown to be correlated with time

2 CBLI). McNeill et al. (2000) compared tibia Pb levels and cumulative blood Pb indices in a population of 

3 19- to 29-year-olds who had been highly exposed 

4 They concluded that Pb from exposure in early childhood had persisted in the bone matrix until 

5 adulthood. The bone Pb/CBLI slopes from vario

4-49 

Simulation of relationship between blood Pb concentration, 

e Pb and body burden in adults.A constant baseline intake 

n a quasi-steady state blood Pb concentration and 

body burden (upper panel). A change in Pb uptake gives rise 

to a relatively rapid change in blood Pb, to a new quasi

state, and a slower change in body burden. The long-term 

averaged blood Pb concentration more closely tracks the 

slower pattern of change in body burden (middle panel). The 

integrated blood Pb concentration increases over the 

lifetime, with a greater rate of increase during periods of 

 uptake (lower panel). Simulation based on ICRP Pb 

biokinetics model (Leggett, 1993). 

Tibia bone Pb has been shown to be correlated with time-integrated blood Pb concentration (i.e., 

compared tibia Pb levels and cumulative blood Pb indices in a population of 

olds who had been highly exposed to Pb in childhood from the Bunker Hill, Idaho smelter. 

They concluded that Pb from exposure in early childhood had persisted in the bone matrix until 

adulthood. The bone Pb/CBLI slopes from various studies range from 0.022 to 0.067 μg/g bone mineral 

DO NOT CITE OR QUOTE 

Simulation of relationship between blood Pb concentration, 

A constant baseline intake 

steady state blood Pb concentration and 

body burden (upper panel). A change in Pb uptake gives rise 

to a relatively rapid change in blood Pb, to a new quasi-steady 

term 

ncentration more closely tracks the 

slower pattern of change in body burden (middle panel). The 

integrated blood Pb concentration increases over the 

lifetime, with a greater rate of increase during periods of 

based on ICRP Pb 

integrated blood Pb concentration (i.e., 

compared tibia Pb levels and cumulative blood Pb indices in a population of 

to Pb in childhood from the Bunker Hill, Idaho smelter. 

They concluded that Pb from exposure in early childhood had persisted in the bone matrix until 

μg/g bone mineral 
May 2011 4-50 DRAFT – DO NOT CITE OR QUOTE

1 per μg-year/dL (Healey et al., 2008; Hu et al., 2007). Because the CBLI is a cumulative function which 

2 cannot decrease over time, CBLI and bone Pb would be expected to diverge following cessation of 

3 exposure, as bone Pb levels decrease. This has been observed as a lower bone Pb/CBLI slope in retired Pb 

4 workers compared to active workers and in worker populations whose exposures declined over time as a 

5 result of improved industrial hygiene (Fleming et al., 1997; Gerhardsson et al., 1993). 

6 Although, differences in kinetics of blood and bone Pb degrade the predictive value of blood Pb as 

7 a metric of Pb body burden, within a population that has similar exposure histories and age demographics, 

8 blood and bone Pb may show relatively strong associations. A recent analysis of a subset of data from the 

9 Normative Aging Study showed that cross-sectional measurements of blood Pb concentration accounted 

10 for approximately 9% (tibia) to 13% (patella) of the variability in bone Pb levels. Inclusion of age in the 

11 regression model accounted for an additional 7-10% of the variability in bone Pb (Park, Mukherjee, et al., 

12 2009). 

Mobilization of Lead from Bone in Adulthood

13 Potential mobilization of Pb from the skeleton increases this contribution of bone Pb to blood Pb, 

14 which occurs at times of physiological stress associated with enhanced bone remodeling such as during 

15 pregnancy and lactation (Hertz-Picciotto et al., 2000; Manton, 1985; Silbergeld, 1991), menopause or in 

16 the elderly (Silbergeld et al., 1988), extended bed rest (Markowitz & Weinberger, 1990), 

17 hyperparathyroidism (Kessler et al., 1999) and severe weight loss (Riedt et al., 2009). 

18 During pregnancy, bone Pb can serve as a Pb source as maternal bone is resorbed for the 

19 production of the fetal skeleton (Franklin et al., 1997; B. Gulson et al., 1999; B. Gulson et al., 2003; B. L. 

20 Gulson et al., 1997). Increased blood Pb during pregnancy has been demonstrated in numerous studies 

21 and these changes have been characterized as a “U-shaped” pattern of lower blood Pb concentrations 

22 during the second trimester compared to the first and third trimesters (B. Gulson et al., 2004; B. L. Gulson 

23 et al., 1997; Hertz-Picciotto et al., 2000; Lagerkvist et al., 1996; Lamadrid-Figueroa et al., 2006; 

24 Rothenberg et al., 1994; Schuhmacher et al., 1996). The U-shaped relationship reflects the relatively 

25 higher impact of hemodilution in the second trimester versus the rate of bone Pb resorption accompanying 

26 calcium releases for establishing the fetal skeleton. In the third trimester, fetal skeletal growth on calcium 

27 demand is greater, and Pb released from maternal skeleton offsets hemodilution. Gulson et al. (1998)

28 reported that, during pregnancy, blood Pb concentrations in the first immigrant Australian cohort (n = 15) 

29 increased by an average of about 20% compared to non-pregnant migrant controls (n = 7). Skeletal 

30 contribution to blood Pb, based on the isotopic composition for the immigrant subjects, increased in an 

31 approximately linear manner during pregnancy. The mean increases for each individual during pregnancy 

32 varied from 26% to 99%. Interestingly, the percent change in blood Pb concentration was significantly 

33 greater during the post-pregnancy period than during the second and third trimesters. The contribution of 
May 2011 4-51 DRAFT – DO NOT CITE OR QUOTE

1 skeletal Pb to blood Pb during the post-pregnancy period remained essentially constant at the increased 

2 level of Pb mobilization.

3 Gulson et al. (2004) observed that calcium supplementation was found to delay increased 

4 mobilization of Pb from bone during pregnancy and halved the flux of Pb release from bone during late 

5 pregnancy and postpartum. In another study, women whose daily calcium intake was 850 mg per day 

6 showed lower amounts of bone resorption during late pregnancy and postpartum than those whose intake 

7 was 560 mg calcium per day (Manton et al., 2003). Similarly, calcium supplementation (1200 mg/day) in 

8 pregnant Mexican women resulted in an 11% reduction in blood Pb level compared to placebo and a 24%

9 average reduction for the most compliant women (Ettinger et al., 2009). When considering baseline blood 

10 Pb levels in women who were more compliant in taking calcium supplementation, the reductions were 

11 similar for those <5 μg/dL and those ≥ 5 μg/dL (14% and 17%, respectively). This is in contrast to a study 

12 of women who had blood Pb concentrations <5 μg/dL, where calcium supplementation had no effect on 

13 blood Pb concentrations (B. Gulson et al., 2006b). These investigators attributed their results to changes 

14 in bone resorption with decoupling of trabecular and cortical bone sites.

15 Miranda et al. (2010) studied blood Pb level among pregnant women aged 18-44 years old. The 

16 older age segments in the study presumably had greater historic Pb exposures and associated stored Pb 

17 than the younger age segments. Compared with the blood Pb levels of a reference group in the 25-29 

18 years old age category, women ≥ 30 years old had significant odds of having higher blood Pb levels (aged 

19 30-34: OR = 2.39, p <0.001; aged 35-39: OR = 2.98, p <0.001; aged 40-44: OR = 7.69, p <0.001). 

20 Similarly, younger women had less chance of having higher blood Pb levels compared with the reference 

21 group (aged 18-19: OR = 0.60, p = 0.179; aged 20-24: OR = 0.54, p = 0.015). These findings indicate that 

22 maternal blood Pb levels are more likely the result of Pb remobilization from bone stores from historic 

23 exposures as opposed to contemporaneous exposures.

24 Lactation can affect the endogenous bone Pb release rate. After adjusting for patella Pb 

25 concentration, an increase in blood Pb levels of 12.7% (95% CI: 6.2, 19.6) was observed for women who 

26 practiced partial lactation and an increase of 18.6% (95% CI: 7.1, 31.4) for women who practiced 

27 exclusive lactation compared to those who stopped lactation (Tellez-Rojo et al., 2002). In another Mexico 

28 City study, Ettinger et al. (2004; 2006) concluded that an interquartile increase in patella Pb was 

29 associated with a 14% increase in breast milk Pb, whereas for tibia Pb the increase was ~5%. Breast 

30 milk:maternal blood Pb concentration ratios are generally <0.1, although values of 0.9 have been reported 

31 (Ettinger et al., 2006; B. Gulson, Jameson, et al., 1998; Koyashiki et al., 2010). Dietary intake of 

32 polyunsaturated fatty acids (PUFA) has been shown to weaken the association between Pb levels in 

33 patella and breast milk, perhaps indicating decreased transfer of Pb from bone to breast milk with PUFA 

34 consumption (Arora et al., 2008).

35 The Pb content in some bones (i.e., mid femur and pelvic bone) plateau at middle age and then 

36 decreases at older ages (G. A. Drasch et al., 1987). This decrease is most pronounced in females and may 
May 2011 4-52 DRAFT – DO NOT CITE OR QUOTE

1 be due to osteoporosis and release of Pb from resorbed bone to blood (B. Gulson et al., 2002). Two 

2 studies indicate that the endogenous release rate in postmenopausal women ranges from 0.13-0.14 μg/dL 

3 in blood per μg/g bone and is nearly double the rate found in premenopausal women (0.07-0.08 μg/dL per 

4 μg/g bone) (Garrido Latorre et al., 2003; Popovic et al., 2005). 

5 Studies of the effect of hormone replacement therapy on bone Pb mobilization have yielded 

6 conflicting results (Berkowitz et al., 2004; Garrido Latorre et al., 2003; Korrick et al., 2002; Popovic et 

7 al., 2005; Webber et al., 1995). In women with severe weight loss (28% of BMI in 6 months) sufficient to 

8 increase bone turnover, increased blood Pb levels of approximately 2.1 μg/dL (250%) were reported, and 

9 these blood Pb increases were associated with biomarkers of increased bone turnover (e.g., urinary 

10 pyridinoline cross-links) (Riedt et al., 2009).

4.3.6. Relationship Between Lead in Blood and Lead in Soft 

Tissues

11 Figure 4-11 shows simulations of blood and soft tissues Pb (including brain) for the same exposure 

12 scenarios previously displayed. Pb uptake and elimination in soft tissues is much faster than bone. As a 

13 result, following cessation of a period of elevated exposure, Pb in soft tissues is more quickly returned to 

14 blood. The terminal elimination phase from soft tissue mimics that of blood, and it is similarly influenced 

15 by the contribution of bone Pb returned to blood and being redistributed to soft tissue. 
May 2011 4-53 DRAFT – DO NOT CITE OR QUOTE

Figure 4-11. Simulation of blood and soft tissue (including brain) Pb in 

children and adults who experience a period of increased Pb 

intake. Simulation based on ICRP Pb biokinetics model

(Leggett, 1993).

1 Information on Pb levels in human brain are limited to autopsy data and the simulation of brain Pb 

2 shown in Figure 4-12 reflects general concepts derived from observations made in non-human primates, 

3 dogs and rodents. These observations suggest that peak Pb levels in the brain are reached 6 months 

4 following a bolus exposure and within two months approximately 80% of steady state brain Pb levels are 

5 reached (Leggett, 1993). There is a relatively slow elimination of Pb from brain (t1/2 ≈ 2 years) compared 

6 to other soft tissues (Leggett, 1993). This slow elimination rate is reflected in the slower elimination 

0.0

0.2

0.4

0.6

0.8

1.0

0

2

4

6

8

10

0 2 4 6 8 10

Total Soft Tissue Pb (mg)

Blood Pb (

μg/dL)

Age (year)

Blood

Soft Tissue

0.0

0.4

0.8

1.2

1.6

2.0

2.4

0

2

4

6

8

10

25 30 35 40 45 50 55 60 65 70

Total Soft Tissue Pb (mg)

Blood Pb (μg/dL)

Age (year)

Blood

Soft Tissue
May 2011 4-54 DRAFT – DO NOT CITE OR QUOTE

1 phase kinetics in shown Figure 4-12. Although in this model, brain Pb to blood Pb transfer half-times are 

2 assumed to be the same in children and adults, uptake kinetics are assumed to be faster during infancy and 

3 childhood, which achieves a higher fraction of the soft tissue burden in brain, consistent with higher 

4 brain/body mass relationships. This is reflected in the simulation as slower brain Pb accumulation in 

5 children. The uptake half times predicted by Leggett (1993) vary from 0.9 to 3.7 days, depending on age. 

6 Brain Pb kinetics represented in the simulations are simple outcomes of modeling assumptions and cannot 

7 currently be verified with available observations in humans. 
May 2011 4-55 DRAFT – DO NOT CITE OR QUOTE

Figure 4-12. Simulation of blood and brain Pb in children and adults who 

experience a period of increased Pb intake.Simulation based 

on ICRP Pb biokinetics model (Leggett, 1993).

4.3.7. Relationship Between Lead in Blood and Lead in 

Urine

1 Urinary filtering and excretion of Pb is associated with plasma Pb concentrations. Given the 

2 curvilinear relationship between blood Pb and plasma Pb, a secondary expectation is for a curvilinear 

3 relationship between blood Pb and urinary Pb excretion that may become evident only at relatively high 

0

8

16

24

32

40

48

0

2

4

6

8

10

25 30 35 40 45 50 55 60 65 70

Total Brain Pb (

μg)

Blood Pb (

μg/dL)

Age (year)

Blood

Brain
May 2011 4-56 DRAFT – DO NOT CITE OR QUOTE

1 blood Pb concentrations (e.g., >25 μg/dL). Figure 4-13 shows these relationships predicted from the 

2 model. In this case, the exposure scenario shown is for an adult (age 40 years) at a quasi-steady state 

3 blood Pb concentration; the same relationships hold for children. At low blood Pb concentrations (<25 

4 μg/dL), urinary Pb excretion is predicted to closely parallel plasma Pb concentration for any given blood 

5 Pb level (Figure 4-13, top panel). It follows from this that, similar to blood Pb, urinary Pb will respond 

6 much more rapidly to an abrupt change in Pb exposure than will bone Pb. One important implication of 

7 this relationship is that, as described previously for blood Pb, the relationships between urinary Pb and 

8 bone Pb will diverge with increasing exposure duration and age, even if exposure remains constant. 

9 Furthermore, following an abrupt cessation of exposure, urine Pb (i.e., not provoked by administration of 

10 chelating agents) will quickly decrease while bone Pb will remain elevated (Figure 4-13, lower panel).
May 2011 4-57 DRAFT – DO NOT CITE OR QUOTE

Figure 4-13. Top panel: Predicted relationship between plasma Pb 

concentration and urinary Pb excretion in an adult (age 40 

years). Lower panel: Simulation of blood Pb, bone Pb and 

urinary excretion of Pb in an adult who experiences a period 

of increased Pb intake. Simulation based on ICRP Pb

biokinetics model (Leggett, 1993).

0.00

0.05

0.10

0.15

0.20

0.25

0.30

0

20

40

60

80

0 10 20 30 40 50

Plasma Pb (μg/dL)

Urine Pb (

μg/day)

Blood Pb (μg/dL)

Urine Plasma

0

2

4

6

8

10

0

4

8

12

16

25 30 35 40 45 50 55 60 65 70

Bone Pb (mg)

Urine Pb (μg/day)

Blood Pb (μg/dL)

Age (year)

Bone Urine Blood
May 2011 4-58 DRAFT – DO NOT CITE OR QUOTE

4.4. Observational Studies of Lead Exposure

4.4.1. Lead in Blood

1 Overall, trends in blood Pb levels have been decreasing among U.S. residents over the past twenty 

2 years. Blood Pb concentrations in the U.S. general population have been monitored in the NHANES. 

3 Analyses of these data have shown a progressive downward trend in blood Pb concentrations during the 

4 period 1976-2008, with the most dramatic declines coincident with the phase out of leaded gasoline 

5 (Brody et al., 1994; Pirkle et al., 1994; Pirkle et al., 1998; J. Schwartz & Pitcher, 1989). The temporal 

6 trend for the period 1988-2008 is shown in Figure 4-14. Summary statistics from the most recent 

7 publically available data (1999-2008) are presented in Table 4-7 (CDC, 2011). The geometric mean Pb 

8 concentration among children 1-5 years of age, based on the sample collected during the period 2007-

9 2008, was 1.51 μg/dL (95% CI: 1.37, 1.66), which was a slight increase from the previous year (1.46 

10 μg/dL, 95% CI: 1.36, 1.57). Figure 4-15 uses NHANES data to illustrate the distribution of blood Pb 

11 levels among U.S. children aged 12-60 months. The median blood Pb in this age group was 1.4 μg/dL 

12 with a 95th percentile value of 4.1 μg/dL (NCHS, 2010). For 2005-2008, 95% of childhood blood Pb 

13 levels were less than 5 μg/dL. 

14 When data were aggregated for years 1999-2004, Pb concentrations in children were highest in the 

15 ethnicity category non-Hispanic black (GM 2.8, 95% CI: 2.5, 3.0) compared to the categories Mexican16 American (GM 1.9, 95% CI: 1.7, 2.0) and non-Hispanic white (GM 1.7, 95% CI: 1.6, 1.8) (Jones et al., 

17 2009). Figure 4-16 demonstrates the change in percent of children with various blood Pb levels by 

18 race/ethnicity from 1988-1991 and 1999-2004. When these data were aggregated for the years 1988-2004, 

19 residence in older housing, poverty, age, and being non-Hispanic black were significant risk factors for 

20 higher Pb levels (Jones et al., 2009). The geometric mean blood Pb concentration among adults ≥ 20 years 

21 of age was 1.38 μg/dL (95% CI: 1.31, 1.46) based on the sample collected during the period 2007-2008 

22 (CDC, 2011). Based on these same data, the geometric mean for all males was 1.47 μg/dL (95% CI: 1.39, 

23 1.56), and for females was 1.11 μg/dL (95% CI: 1.06, 1.16).
May 2011 4-59 DRAFT – DO NOT CITE OR QUOTE

Figure 4-14. Temporal trend in blood Pb concentration.Shown are 

geometric means and 95% CIs based on data from NHANES III 

Phase 1 (Brody et al., 1994; Pirkle et al., 1994); NHANES III 

Phase 2 (Pirkle et al., 1998); and NHANES IV (CDC, 2011). Data 

for adults during the period 1988-1994 are for ages 20-49 

years, and ≥ 20 years for the period 1999-2008.

0

1

2

3

4

5

88-91 91-94 99-00 01-02 03-04 05-06 07-08

Blood Lead (μg/dL)

Survey Period 

Children 1-5 yrs

Adults ≥ 20 yrs
May 2011 4-60 DRAFT – DO NOT CITE OR QUOTE

Table 4-7. Blood Pb concentrations in the U.S. population

Survey Stratum Period Geometric Mean (μg/dL) 95% Confidence Interval Number of Subjects

All 1999-2000 1.66 1.60, 1.72 7970

2001-2002 1.45 1.39, 1.51 8945

2003-2004 1.43 1.36, 1.50 8373

2005-2006 1.29 1.23, 1.36 8407

2007-2008 1.27 1.21, 1.34 8266

1-5 yr 1999-2000 2.23 1.96, 2.53 723

2001-2002 1.70 1.55, 1.87 898

2003-2004 1.77 1.60, 1.95 911

2005-2006 1.46 1.36, 1.57 968

2007-2008 1.51 1.37, 1.66 817

6-11 yr 1999-2000 1.51 1.36, 1.66 905

2001-2002 1.25 1.14, 1.36 1044

2003-2004 1.25 1.12, 1.39 856

2005-2006 1.02 0.95, 1.01 934

2007-2008 0.99 0.91, 1.07 1011

12-19 yr 1999-2000 1.10 1.04, 1.17 2135

2001-2002 0.94 0.90, 0.99 2231

2003-2004 0.95 0.88, 1.02 2081

2005-2006 0.80 0.75, 0.85 1996

2007-2008 0.80 0.74, 0.86 1074

≥ 20 yr 1999-2000 1.75 1.68, 1.81 4207

2001-2002 1.56 1.49, 1.62 4772

2003-2004 1.52 1.45, 1.60 4525

2005-2006 1.41 1.34, 1.48 4509

2007-2008 1.38 1.31, 1.46 5364

Males 1999-2000 2.01 1.93, 2.09 3913

2001-2002 1.78 1.71, 1.86 4339

2003-2004 1.69 1.62, 1.75 4132

2005-2006 1.52 1.42, 1.62 4092

2007-2008 1.47 1.39, 1.56 4147

Females 1999-2000 1.37 1.32, 1.43 4057

2001-2002 1.19 1.14, 1.25 4606

2003-2004 1.22 1.14, 1.31 4241

2005-2006 1.11 1.05, 1.17 4315

2007-2008 1.11 1.06, 1.16 4119

Mexican - Americans 1999-2000 1.83 1.75, 1.91 2742

2001-2002 1.46 1.34, 1.60 2268

2003-2004 1.55 1.43, 1.69 2085

2005-2006 1.29 1.21, 1.38 2236

2007-2008 1.25 1.15, 1.36 1712

Non-Hispanic blacks 1999-2000 1.87 1.75, 2.00 1842

2001-2002 1.65 1.52, 1.80 2219

2003-2004 1.69 1.52, 1.89 2293

2005-2006 1.39 1.26, 1.53 2193

2007-2008 1.39 1.30, 1.48 1746

Non-Hispanic whites 1999-2000 1.62 1.55, 1.69 2716

2001-2002 1.43 1.37, 1.48 3806

2003-2004 1.37 1.32, 1.43 3478

2005-2006 1.28 1.19, 1.37 3310

2007-2008 1.24 1.16, 1.33 3461

Source: Adapted from data from the NHANES (CDC, 2011)
May 2011 4-61 DRAFT – DO NOT CITE OR QUOTE

Source: Adapted from data from the NHANES (NCHS, 2010)

Figure 4-15. Box plots of blood Pb levels among U.S. children (1-5 years 

old) from the NHANES survey, 1988-2008.Top: all data.

Bottom: data for subjects having blood Pb levels less than 15 

μg/dL. 
May 2011 4-62 DRAFT – DO NOT CITE OR QUOTE

Data used with permission from the American Academy of Pediatrics, Jones et al. (2009)

Figure 4-16. Percent distribution of blood Pb levels by race/ethnicity 

among U.S. children (1-5 years) from the NHANES survey, 

1988-1991 (top) and 1999-2004 (bottom).

1 Several studies have shown seasonal variation in blood Pb concentrations in children (B. Gulson et 

2 al., 2008; D. L. Johnson et al., 1996; Laidlaw et al., 2005). Seasonal variation in blood Pb concentration

3 was also evident in individual children when repeated blood Pb measurements were made over a 5-year 

0

10

20

30

40

50

60

70

< 1 1 - < 2.5 2.5 - < 5 5 - < 7.5 7.5 - <10 ≥ 10

Percent of Children

1988-1991

0

10

20

30

40

50

60

70

< 1 1 - < 2.5 2.5 - < 5 5 - < 7.5 7.5 - <10 ≥ 10

Percent of Children

Blood Pb Level (ug/dL)

Non-Hispanic black Mexican American Non-Hispanic white

1999 - 2004
May 2011 4-63 DRAFT – DO NOT CITE OR QUOTE

1 period (B. Gulson et al., 2008). Meteorological factors appear to contribute to blood Pb seasonality. 

2 Laidlaw et al. (2005) analyzed the temporal relationships between child blood Pb concentrations and 

3 various atmospheric variables in three cites (Indianapolis, IN; Syracuse, NY; New Orleans, LA). Blood 

4 Pb data was obtained from public health screening programs conducted in the three cities during 

5 1998-2003. Blood Pb samples were dominated by children <5 years of age and age distribution varied 

6 across the three cities. The number of blood Pb measurements included in the analyses were as follows: 

7 Indianapolis, 15,969; Syracuse, 14,457 (D. Johnson & Bretsch, 2002; D. L. Johnson et al., 1996); New 

8 Orleans, 2,295 (Mielke et al., 2007). The temporal variation in blood Pb concentrations in each city were 

9 predicted by multivariate regression models that included the following significant variables: PM10, wind 

10 speed, air temperature, and soil moisture; as well as dummy variables accounting for temporal 

displacement of the effects of each independent variable on blood Pb. Laidlaw et al. (2005) reported R2 11

values for the regression models, but did not report the actual regression coefficients. The R2 12 values were 

13 as follows: Indianapolis 0.87 (p = 0.004); Syracuse 0.61 (p = 0.0012); New Orleans 0.59 (p <0.00001). 

14 This analysis provides a possible explanation for the seasonal patterns of blood Pb concentrations in 

15 children that involves weather-dependent entrainment and air transport of surface dusts.

16 Trends in blood Pb levels have been accompanied by changes in Pb isotope ratios within blood Pb. 

17 Isotopic ratios, described in Sections 3.2 and 3.3 as a tool for source apportionment, have been used to 

18 associate blood Pb measurements with anthropogenic sources of Pb in the environment. Changes in Pb 

19 isotopic ratios in blood samples reflect the changing influence of sources of Pb following the phase-out of 

20 tetraethyl Pb antiknock agents in automotive gasoline and changes in Pb usage in paints and other 

21 industrial and consumer products (B. Gulson et al., 2006a; B. Gulson et al., 2008; Ranft et al., 2006, 

2008). Gulson et al. (2006a) illustrated how a linear increase in the isotopic ratio 206Pb/204 22 Pb occurred in 

23 concert with a decrease in blood Pb levels among various study populations in Australia during the period 

1990-2000 (Figure 4-17). Gulson et al. (2006a) point out that the isotopic signature of 206Pb/204 24 Pb derived

from Australian mines (median ~16.8) differs from that of European and Asian mines, where 206Pb/204 25 Pb 

26 varies between ~17.4 and ~18.1. Liang et al. (2010) also examined the trends in blood Pb level over the 

27 period 1990 to 2006 in Shanghai and saw a reduction corresponding to the phase out of Pb in gasoline. A 

plot of 208Pb/206Pb to 207Pb/206 28 Pb for blood and environmental samples showed overlap between the 

29 isotopic signature for coal combustion ash and that measured in blood. This result suggests a growing 

30 influence of Pb from coal ash in Shanghai in the absence of Pb in automobile emissions.
May 2011 4-64 DRAFT – DO NOT CITE OR QUOTE

Source: Used with permission from Academic Press, Gulson et al. (2006a)

Figure 4-17. a) Trends in 206Pb/204Pb isotope ratio in blood Pb and b) trends 

in blood Pb levels among Australian study populations during 

the period 1990-2000.

4.4.2. Lead in Bone

1 An extensive national database (i.e., NHANES) is available for blood Pb concentrations in children 

2 and adults, as described in Section 4.4.1. Bone Pb concentrations are less well characterized. Tables 4-8 

3 and 4-9 are compilations of data from epidemiologic studies that provided bone Pb concentrations and/or 

4 variability in concentrations among individuals without reported occupational exposure and those with 

5 occupational exposures, respectively. In non-occupationally exposed individuals, typical group mean tibia 

6 bone Pb concentrations ranged from 10 to 30 μg/g. Patella bone Pb levels are typically higher than tibia 

7 bone Pb levels in the studies considered (Table 4-8). For example, in the Normative Aging Study, patella 

8 bone Pb concentrations were approximately 32 μg/g, whereas tibia bone Pb concentrations were about 22 

9 μg/g. Occupationally exposed individuals generally had greater bone Pb concentrations than seen in 

10 control groups (i.e., unexposed). Bone Pb data in Table 4-9 for occupationally exposed individuals were 

11 also generally higher compared to non-occupationally exposed individuals (Table 4-8).
May 2011 4-65 DRAFT – DO NOT CITE OR QUOTE

Table 4-8. Epidemiologic studies that provide bone Pb measurements for nonoccupationally exposed populations

Reference Cohort Age (yrs) N Location Study Period Prior Pb 

Exposure

Bone Pb 

biomarker

Bone Pb 

Conc. (μg/g)

Distribution 

of Bone Pb 

(μg/g)

Bandeen- Roche et al. 

(2009)

Baltimore 

Memory 

Study cohort

50-70 1140 Baltimore, MD 2001-2005 Cumulative Tibia Mean±SD

Tibia: 18.8 ± 11.6

Not reported

Bellinger et 

al. (1994)

Not reported 5-8 

(recruited)

19-20 (follow- up)

79 Boston, MA 1989-1990 Cumulative Tibia

Patella

Mean 

(Range):

Tibia: 5.4 (3- 16)

Patella: 9.2 

(4-18)

High 

exposure: 

>24

Low 

exposure: 

<8.7

Cheng et al. 

(2001)

Normative 

Aging Study 

cohort

Mean±SD:

Normotensive

: 65.49 ± 7.17

Borderline 

hypertension: 

68.3 ± 7.79

Definite 

hypertension: 

67.93 ± 6.79

833 

males

Boston, MA 8/1/1991-

12/31/1997

Cumulative Tibia

Patella

Mean±SD

Tibia:

Normotensive

: 20.27 ± 11.55

Borderline 

hypertension: 

23.46 ± 15.02

Definite 

hypertension: 

22.69 ± 14.71

Patella:

Normotensive

: 28.95 ± 18.01

Borderline 

hypertension: 

33.73 ± 21.76

Definite 

hypertension: 

32.72 ± 19.55

Lowest 

quintile: Tibia: 

8.5

Patella: 12.0

Highest 

quintile:

Tibia: 36.0

Patella: 53.0

Coon et al. 

(2006)

Participants 

from Henry 

Ford Health 

System 

(HFHS)

≥ 50

Mean: 69.9

121 

cases

414 

controls

Southeastern 

Michigan

1995-1999 

(participants 

received 

primary health 

care services)

Cumulative Tibia

Calcaneu

s

Mean±SD:

Tibia: 12.5 ± 

7.8

Calcaneus: 

20.5 ± 10.2

Tibia

Q1: 0-5.91

Q2: 5.92- 10.40

Q3: 10.41-

15.50

Q4: ≥ 15.51

Calcaneus

Q1: 0-11.70

Q2: 11.71-

19.07

Q3: 19.08- 25.28

Q4: ≥ 25.29

Elmarsafaw

y et al.

(2006)

Normative 

Aging Study

Not reported 471 

elderly 

males

Greater 

Boston area, 

MA

6/1991-

12/1994

Not reported Tibia

Patella

Mean±SD:

Tibia: 21.6 ±1 

2.0

Patella: 31.7 

± 18.3

Not reported

Glass et 

al.,(2009)

Baltimore 

Memory 

Study

Mean: 59.4

Range: 50-70

1,001 Baltimore, MD 2001-2005 Cumulative 

(lifetime)

Tibia Mean±SD:

Tibia: 18.8 ± 11.1

NPH Scale:

Lowest 

tertile: Mean 

Tibia level: 

16.3 ± 11.0

Middle tertile: 

Mean Tibia 

level: 19.3 ± 

10.7

Highest 

tertile: Mean 

Tibia level: 

20.3 ± 11.4
May 2011 4-66 DRAFT – DO NOT CITE OR QUOTE

Reference Cohort Age (yrs) N Location Study Period Prior Pb 

Exposure

Bone Pb 

biomarker

Bone Pb 

Conc. (μg/g)

Distribution 

of Bone Pb 

(μg/g)

Hsieh et al. 

(2009)

Not reported Mean:

Control: 

46.06

18

controls

Not reported Not reported Control group for 

occupational 

exposure group

Tibia

Patella

Mean±SD

Tibia Control: 

18.51 ± 22.40

Patella 

Control: 7.14 

± 9.81

Not reported

Hu et al. 

(1996)

(As 

reported in 

NavasAcien et al., 

(2008))

Normative 

Aging Study

48-92

Mean ± SD: 

66.6 ± 7.2

590 

males

Boston, MA 8/1991- 12/1994Cumulative Tibia

Patella

Mean±SD: Tibia: 21.8 ± 

12.1

Patella: 32.1 

± 18.7

Range:

Tibia: <1-96

Patella: 1-142

Figures 1 and 

2 show both 

types of bone 

Pb levels 

increasing 

with age

Jain et al. 

(2007)

VANormative 

Aging Study

Not reported 837 

males

Greater 

Boston, MA

9/1/1991-

12/31/2001

Not reported Tibia

Patella

Mean ± SD

Tibia:

Non-Cases: 

21.4 ± 13.6

Cases: 24.2 ± 15.9

Patella: 

Non-cases:

30.6±19.7

Cases: 36.8 ± 20.8

Range:

Tibia:

Noncases:

-3-126

Cases: -5-75

Patella:

Noncases: -

10-165

Cases: 5-101

Mean ± SD 

(Range):

Tibia:

Non-cases: Tertile 1: 10.2 

± 3.8 (-3-15)

Tertile 2: 19.1 

± 2.3 (16-23)

Tertile 3: 35.5 

± 14.4 (24- 126)

Cases:

Tertile 1: 10.1 

± 5.3 (-5-15)

Tertile 2: 19.8 

± 2.2 (16-23)

Tertile 3: 39.5 

± 14.9 (25- 75)

Patella:

Non-cases: 

Tertile 1: 

13.9±4.9 (- 10-20)

Tertile 2: 

27.1±4.1 (21- 34)

Tertile 3: 

52.5± 20.7 

(35-165)

Cases:

Tertile 1: 

15.3±4.3 (5-

19)

Tertile 2: 25.7 

± 3.8 (21-33)

Tertile 3: 53.3 

± 17.3 (35-

101)

Kamel et al.

(2002); 

Kamel et al. 

(2005); 

Kamel et al. 

(2008)

Not reported 30-80 256 

controls

(Bone 

samples 

collecte

d from 

41 

controls

)

New England

(Boston, MA)

1993-1996 Cumulative

Control group for 

occupational 

exposure group

Tibia

Patella

Mean±SE

Tibia 

Controls: 11.1 

± 1.6

Patella 

Controls: 

16.7 ± 2.0

Controls

Tibia: N (%)

-7-7: 14 (34)

8-14: 12 (29)

15-61: 15 

(37)

Patella: N 

(%) 

-4-9: 14 (34)

10-20: 14 

(34)

21-107: 13 

(32)

Khalil et al. 

(2009)

1982 Pb 

Occupational 

Study

Control 

mean: 55

51 

controls

Eastern 

Pennsylvania

1982-2004 Control group for 

occupational 

exposure group

Tibia Median (IQR)

Tibia Control: 

12 (-8-32)

Not reported
May 2011 4-67 DRAFT – DO NOT CITE OR QUOTE

Reference Cohort Age (yrs) N Location Study Period Prior Pb 

Exposure

Bone Pb 

biomarker

Bone Pb 

Conc. (μg/g)

Distribution 

of Bone Pb 

(μg/g)

Korrick et 

al. (1999)

(As

reported in 

from NavasAcien et al., 

(2008)

Nurses’ 

Health Study

Combined: 

47-74

Mean±SD: 

Combined: 

58.7 ± 7.2

Cases: 61.1 ± 

7.1

High controls: 

61.1 ± 7.2

Low controls: 

58.7 ± 7.1

284 

females

(89 

cases; 

195 

controls

)

Boston, MA 7/1993- 7/1995Nonoccupationall

y exposed

Tibia

Patella

Mean ± SD

Tibia:

Combined: 

13.3 ± 9.0

Cases: 13.0 ± 

9.4

High controls: 

14.7 ± 10

Low controls: 

12.7 ± 8.1

Patella:

Combined: 

17.3 ± 11.1

Cases: 19.5 ± 12.9

High controls: 

17.2 ± 9

Low controls: 

15.8 ± 10.6

Range

Tibia 

Combined: -

5-69

Patella 

Combined: - 5-87

Patella:

10th

percentile: 6

90th

percentile: 31

Lee et al. 

(2001) (As 

reported in 

Navas- Acien et al., 

(2008))

Not reported 22.0-60.2

Mean ± SD:

Controls: 

34.5 ± 9.1

135 

controls

Republic of 

Korea

10/24/1997-

8/19/1999

Control group for 

occupational 

exposure group

Tibia Mean ± SD

Tibia 

Controls: 5.8 

± 7.0

Range

Tibia 

Controls:

-11-27

Not reported

Martin et al. 

(2006)

Baltimore 

Memory 

Study

50-70

Mean: 59.4

964 Baltimore, MD 5/2001- 9/2002 (1st

study visit)

8/2002- 3/2004 (2nd

study visit –

tibia Pb 

measured)

Cumulative 

(lifetime)

Tibia Mean ± SD

Tibia: 18.8 ± 

12.4

Tibia IQR: 

11.9-24.8

Needleman 

et al. (2002)

Not reported 12-18

Mean age ± SD:

African 

American 

cases: 15.8 ± 

1.4

African 

American 

controls: 15.5 

± 1.1

White cases: 

15.7 ± 1.3

White 

controls: 15.8 

± 1.1

194 

male 

youth 

cases

146 

male 

youth 

controls

Allegheny 

County, PA 

(cases); 

Pittsburgh, PA 

(controls)

4/1996- 8/1998Not reported Tibia Mean ± SD

Tibia Cases

(ppm):

All subjects: 

11.0 ± 32.7

African 

American: 9.0 

± 33.6

White: 20 ± 

27.5

Tibia Controls

(ppm): All subjects: 

1.5 ± 32.1

African 

American: -

1.4 ± 31.9

White: 3.5 ± 

32.6

Table 4 

distributes 

bone Pb by ≥ 

25 or <25 for 

race, two 

parental 

figures, and 

parent 

occupation

Osterberg 

et al. (1997)

(As 

reported in

Shih et al., 

(2007))

Not reported Median: 41.5 19 male 

controls

Not reported Not reported Control group for 

occupational 

exposure group

Finger 

bone

Median 

(range)

Finger Bone

Controls: 

4 (-19-18)

Not reported
May 2011 4-68 DRAFT – DO NOT CITE OR QUOTE

Reference Cohort Age (yrs) N Location Study Period Prior Pb 

Exposure

Bone Pb 

biomarker

Bone Pb 

Conc. (μg/g)

Distribution 

of Bone Pb 

(μg/g)

Park et al. 

(2006)

Normative 

Aging Study

Mean: 72.9 ± 6.5413 

males

Greater 

Boston, MA

11/14/2000- 12/22/2004

(HRV 

measurement

s taken)

1991-2002 

(bone Pb 

measurement

s taken)

Not reported Tibia

Patella

Median (IQR)

Tibia: 19.0 

(11-28)

Patella: 23.0 

(15-34)

Estimated 

Patellaa

: 16.3 

(10.4-25.8)

Median (IQR) 

for No. of 

metabolic 

abnormalities

:

Tibia:

0: 18.5 (10.5-

23)

1: 19 (11-28)

2: 19 (12-26)

Patella: 

0: 22 (13.5-

32)

1: 25 (16-36)

2: 20 (15-32)

Estimated 

Patella:

0: 16.3 (10.8-

24.8)

1: 17.1 (11-

29.3)

2: 15.1 (9.4- 22.1)

Park et al. 

(2009)

Normative 

Aging Study

Mean: 67.3 ± 7.2613 

males

Greater 

Boston, MA

8/1991 - 12/1995Not reported Tibia

Patella

Median (IQR)

Tibia: 19 (14-

27)

Patella: 26 

(18-37)

Table 1 

distributes 

tibia and 

patella Pb by 

genotype; 

Table 2 

distributes 

tibia and 

patella Pb by 

number of 

gene variants

Park et al. 

(2010)

VA Normative 

Aging Study 

cohort

Mean: 64.9 

(at bone Pb 

measurement

)

448 

males

Eastern 

Massachusett

s

1991-1996 Cumulative 

(chronic 

exposure)

Tibia

Patella

Mean±SD

Tibia: 22.5 ± 

14.2

Patella: 32.5 

± 20.4

Tibia IQR: 15

Patella IQR: 

21

Table 2 

provides ageadjusted 

mean bone 

Pb levels 

(age, race, 

education, 

smoking 

[pack-yr], 

occupational 

noise, noise 

notch, BMI, 

hypertension, 

diabetes)

Payton et 

al. (1998)

VA Normative 

Aging Study 

cohort

Mean: 66.8 141 

males

Boston, MA 4/1993- 3/1994Not reported Tibia

Patella

Mean ± SD

Tibia: 22.5 ± 

12.2

Patella: 31.7 

± 19.2

Not reported

Peters et al. 

(2007)

Normative 

Aging Study 

cohort

Mean: 66.9 513 

male 

cases

Boston, MA 1991-1996 Cumulative Tibia

Patella

Mean ± SD

Tibia: 21.5 ± 13.4

Patella: 31.5 

± 19.3

Not reported

Rajan et al. 

(2007)

VA Normative 

Aging Study 

Cohort

Mean: 67.5 

(at bone

scan)

1075 

males

Boston, MA 1991-2002 Not reported Tibia

Patella

Mean ± SD

Tibia: 22.1 ± 

13.8

Patella: 31.4 ± 19.6

Not reported
May 2011 4-69 DRAFT – DO NOT CITE OR QUOTE

Reference Cohort Age (yrs) N Location Study Period Prior Pb 

Exposure

Bone Pb 

biomarker

Bone Pb 

Conc. (μg/g)

Distribution 

of Bone Pb 

(μg/g)

Rajan et al. 

(2008)

VA Normative 

Aging Study 

Cohort

≥ 45 720 

males

Boston, MA 1993-2001 Current and 

cumulative

Tibia

Patella

Mean ± SD

ALAD 1-1

Tibia: 21.9 ± 

13.8

Patella: 29.3 

± 19.1

ALAD 1-2/2-2

Tibia: 21.2 ± 

11.6

Patella: 27.9 

± 17.3

Not reported

Rhodes et 

al. (2003)

VA Normative 

Aging Study 

Cohort

Mean: 67.1 526 

males

Boston, MA 1/1/1991- 12/31/1995Not reported Tibia

Patella

Mean ± SD

Tibia: 21.9 ± 13.5

Patella: 32.1 

± 19.8

No. of 

participants 

(%)

Tibia:

<1-15: 173 

(33)

16-24: 186 

(35)

25-126: 167 

(32)

Patella:

<1-22: 189 

(36)

23-35: 165 

(31)

36-165: 172 

(33)

Roels et al. 

(1994)

Not reported 30-60 68 

males

Belgium Not reported Control group for 

occupational 

exposure group

Tibia Geometric 

Mean 

(Range)

Tibia 

Controls:

Normotensive

: 21.7 (<15.2-

69.3)

Hypertensive: 

20.2 (<15.2-

52.9)

Total: 21.4 

(<15.2-69.3)

Not reported 

Rothenberg 

et al. (2002)

(As

reported in

NavasAcien et al., 

(2008))

Not reported 15-44

Mean ± SD: 

31.0 ± 7.7

720 

females

Los Angeles, 

CA

6/1995- 5/2001Not reported Tibia

Calcaneu

s

Mean ± SD

Tibia: 8.0 ± 

11.4

Calcaneus: 

10.7 ± 11.9

Tibia 

quartiles:

Q1: -33.7-0.9

Q2: 1.0-8.0

Q3: 8.1-16.1

Q4: 16.2-42.5

Calcaneus 

quartiles:

Q1: -30.6-3.0

Q2: 3.1-10.0

Q3: 10.1-18.7

Q4: 18.8-49.0

Shih et al., 

(2006)

Baltimore 

Memory 

Study cohort

Mean: 59.39 985 Baltimore, MD Not reported Not reported Tibia Mean ± SD: Tibia: 18.7 ± 

11.2

Not reported
May 2011 4-70 DRAFT – DO NOT CITE OR QUOTE

Reference Cohort Age (yrs) N Location Study Period Prior Pb 

Exposure

Bone Pb 

biomarker

Bone Pb 

Conc. (μg/g)

Distribution 

of Bone Pb 

(μg/g)

Stokes et 

al. (1998), 

as reported 

in Shih et 

al., (2007)

Not reported 19-29 (in 

1994)

Mean ± SD:

Cases: 24.3 ± 3.18

Control: 24.2 

± 3.02

Cases: 9 

months-9 yr 

(during 

1/1/1974-

12/31/1975)

257 

cases

276 

controls

Silver Valley, 

ID; Spokane, 

WA

7/10/1994- 8/7/1994Cumulative 

(lifelong)

Environmental 

(resided near Pb 

smelter during 

childhood)

Tibia Mean

(Range):

Tibia Cases: 

4.6 (-28.9-37)

Tibia 

Controls: 0.6 

(-46.4-17.4)

Tibia

No. of Cases:

<1 μg/g: 

31.5%

1-5 μg/g: 

24.4%

5-10 μg/g: 

22.3%

>10 μg/g: 

21.8%

No. of 

Controls:

<1 μg/g: 

50.4%

1-5 μg/g: 

25.6%

5-10 μg/g: 

19.4%

>10 μg/g: 

4.7%

Mean ± SD

Tibia 

concentration 

by age group:

Cases:

19-21: 1.47 ± 8.35

22-24: 4.48 ± 

7.45

25-27: 4.82 ± 8.92

28-30: 6.64 ± 9.53

Controls:

19-21: 1.27 ± 

6.60

22-24: -0.61 ± 6.19

25-27: 0.60 ± 

8.60

28-30: 1.74 ± 

6.42

Van 

Wijngaarde

n et al. 

(2009)

Not reported Mean: 61.5 47 Rochester, 

NY

Not reported Cumulative Tibia

Calcaneu

s

Mean ± SD

Tibia: 2.0 ± 

5.2

Calcaneus: 6.1 ± 8.5

Not reported

Wasserman 

et al. (2003)

Yugoslavia 

Prospective 

Study of 

Environment

al Pb 

Exposure

10-12 167 

children

Kosovska, 

Mitrovica, 

Kosovo, 

Yugoslavia; 

Pristina, 

Kosovo, 

Yugoslavia

5/1985- 12/1986 

(mother’s 

enrollment)

1986-1999 

(follow-up 

through age 

12 yr)

Tibia Pb 

measured 11-

13 yr old

Cumulative 

(lifetime)

Environmental

(Pb smelter, 

refinery, battery 

plant)

Tibia Mean ± SD:

Tibia

Pristina: 1.36 

± 6.5

Mitrovica: 

39.09 ± 24.55

Tibia 

quartiles:

Q1: -14.4- 1.85

Q2: 1.85-10.5

Q3: 10.5-35

Q4: 35-193.5

Table 3 

distributes 

tibia Pb by 

sex, ethnicity, 

address at 

birth relative 

to factory, 

and maternal 

education
May 2011 4-71 DRAFT – DO NOT CITE OR QUOTE

Reference Cohort Age (yrs) N Location Study Period Prior Pb 

Exposure

Bone Pb 

biomarker

Bone Pb 

Conc. (μg/g)

Distribution 

of Bone Pb 

(μg/g)

Weisskopf 

et al. 

(2004), as 

reported in 

Shih et al. 

194225 

(2007)

Normative 

Aging Study

Mean ± SD: 

67.4 ± 6.6

466 

males

Boston, MA 1991-2002 Environmental Tibia

Patella

Median (IQR)

Tibia: 19 

(12,26)

Patella: 23 

(15, 35)

Tibia IQR: 14

Patella IQR: 

20

Table 3 

shows mean 

Pb levels 

across categorical 

variables (yr 

of education, 

smoking 

status, 

computer 

experience, 

first language 

English)

Weisskopf 

et al. (2007)

VA Normative 

Aging Study 

cohort

Mean: 

Lowest 

Patella 

quintile: 73.2

Highest 

Patella 

quintile: 80.7

31 

males

Boston, MA Bone Pb 

measured: 

1994-1999

Scans 

performed: 

2002-2004

Not reported Tibia

Patella

Median (IQR)

Tibia

Lowest quintile: 13 (9-17)

Highest quin- tile: 41 (38-

59)

Patella

Lowest quintile: 9 (5-15)

Highest quintile: 63 (43- 86)

Not reported

Weisskopf 

et al. (2007)

VA Normative 

Aging Study 

cohort

Mean: 68.7 1,089 

males

Boston, MA 1993-2001 Concurrent and 

cumulative

Tibia

Patella

Median (IQR)

Tibia: 20 (13-

28)

Patella: 25 

(17-37)

Table 1 

shows the 

distribution of 

Pb

biomarkers 

by categories 

of covariates 

(age, education, smoking status, 

alcohol 

intake, physical activity, 

computer 

experience, 

first language 

English)

Weisskopf 

et al. (2009)

Normative 

Aging Study

(95% white)

Mean ± SD 

(at Patella 

baseline)

Tertile 1: 65.2 

± 7.1

Tertile 2: 66.5 

± 6.5

Tertile 3: 70.2 

± 7.2

868 

males

Greater 

Boston area, 

MA

1991-1999 Cumulative Tibia

Patella

Mean ± SD

Tibia: 21.8 ± 13.6

Patella: 31.2 

± 19.4

Patella

tertiles:

1: <22

2: 22-35

3: >35

Weisskopf 

et al. (2010)

BUMC, BWH, 

BIDMC, 

HVMA, 

Normative 

Aging Study 

(NAS), 

Harvard 

Cooperative 

Program on 

Aging 

(HCPOA)

Mean:

Cases: 66.5

Controls: 

69.4

330 

cases

308 

controls

Boston, MA 2003-2007

1991-1999 

(NAS patients 

bone Pb 

measured)

Cumulative Tibia

Patella

Mean ± SD:

Tibia: 10.7 ±

12.1

Patella: 13.6 

± 15.9

Tibia 

quartiles:

Q1: <3.1

Q2: 3.5-9.6

Q3: 10.0-17.0

Q4: >17.3

Patella 

quartiles:

Q1: <2.7

Q2: 3.5-11.0

Q3: 11.3-20.9

Q4: >20.9
May 2011 4-72 DRAFT – DO NOT CITE OR QUOTE

Reference Cohort Age (yrs) N Location Study Period Prior Pb 

Exposure

Bone Pb 

biomarker

Bone Pb 

Conc. (μg/g)

Distribution 

of Bone Pb 

(μg/g)

Weuve et 

al. (2006)

VA Normative 

Aging Study 

cohort

≥ 45 720 

males

Boston, MA 1991 

(measuring 

bone Pb 

levels)

End date not 

reported

Cumulative Tibia

Patella

Median (1st- 3rd quartile):

Tibia: 19 (13-

28)

Patella: 27 

(18-39)

Table 1 

shows 

distribution of 

mean Pb bio- marker levels 

by 

characteristic

s of 

participants 

(age, 

education, 

computer 

experience, 

smoking 

status, 

alcohol

consumption, 

tertile of calcium intake, 

tertile of 

physical 

activity, 

diabetes)

Weuve et 

al. (2009)

Nurses’ 

Health Study 

cohort

47-74 587 

females

Boston, MA 1995-2005 Recent and 

cumulative

Tibia

Patella

Mean ± SD:

Tibia: 10.5 ± 9.7

Patella: 12.6 

± 11.6

Not reported

Wright et al.

(2003), as 

reported in 

Shih et al. 

194225 

(2007)

Normative 

Aging Study

Mean ± SD: 

68.2 ± 6.9

736 

males

Boston, MA 1991-1997 Environmental Tibia

Patella

Mean ± SD:

Tibia: 22.4 ± 

15.3

Patella: 29.5 

± 21.2

Tibia:

Difference in 

mean from 

Lowest- highest 

quartile: 34.2

Patella:

Difference in 

mean from 

lowesthighest 

quartile: 47
May 2011 4-73 DRAFT – DO NOT CITE OR QUOTE

Table 4-9. Epidemiologic studies that provide bone Pb measurements for occupationally 

exposed populations

Reference Cohort Age 

(years)

Number 

of 

Subjects

Location Study 

Period

Prior Pb 

Exposure

Bone Pb 

biomarker

Bone Pb 

Concentration 

(μg/g)

Distribution 

of Bone Pb 

(μg/g)

Bleecker et 

al. (1997)

(As reported 

in Shih et al., 

(2007))

Canada 

Lead Study

Cumulative: 

24-64

Younger: 

24-43

Older: 44-64

Mean ± SD: 

Cumulative: 

44.1 ± 8.36

Younger: 

37.2 ± 4.57

Older: 50.9 

± 4.86

80 males Canada Not 

reported

Occupational

(Pb smelter 

workers)

Tibia Mean ± SD

(Tibia):

Cumulative: 41.0 

± 24.44

Younger: 35 ± 

24.11

Older: 46.9 ± 23.59

Range (Tibia): Cumulative: -12- 90

Younger: -12-80

Older: 3-90

Not reported

Bleecker et 

al. (2007)

Not reported Mean: 39.7 61 Northern 

Canada

Not 

reported

Occupational 

(primary Pb 

smelter 

workers)

Tibia Mean:

Tibia: 38.6

Not reported

Caffo et al. 

(2008)

Not reported Mean: 60.39 513 males Delaware 

and New 

Jersey, US

1994-1997 

(Phase 1 

recruitment

)

2001-2003 

(Phase 2 

recruitment

)

Cumulative

Occupational

(Former 

organolead 

manufacturin

g workers)

Tibia Mean ± SD:

Peak Tibia: 

23.99 ± 18.46

Not reported

Dorsey et al. 

(2006)

Not reported Mean: 43.4 652 Korea 10/24/1997

-8/19/1999 

(enrolled)

Occupational 

(Pb workers)

Tibia

Patella

Mean ± SD:

Tibia: 33.5 ± 

43.4

Patella: 75.1 ± 101.1

Not reported

Glenn et al. 

(2003), as 

reported in 

Navas-Acien 

et al., (2008)

Not reported 40-70

Mean: 55.8 

(baseline)

496 males Eastern U.S. 6/1994-

6/1996 

(enrolled)

6/1998 

(follow-up 

period 

ended)

Occupational

(Chemical 

manufacturin

g facility; 

inorganic and 

organic Pb)

Tibia Mean ± SD:

Tibia: 14.7 ± 9.4 

(at yr 3)

Peak Tibia: 24.3 

± 18.1

Range: 

Tibia: -1.6-52 (at 

year 3)

Peak Tibia: 

-2.2-118.8

Not reported

Glenn et al. 

(2006)

Not reported 0-36.2 

(baseline)

Mean ± SD: 

41.4 ± 9.5 

(baseline)

575

(76% 

male; 24% 

female)

South Korea 10/1997- 6/2001Cumulative 

and recent

Occupational 

(Pb-using 

facilities)

Tibia Mean ± SD:

Tibia: 38.4 ± 

42.9

Tibia-Women:

Visit 1: 

28.2±19.7

Visit 2: 

22.8±20.9

Tibia-Men:

Visit 1: 

41.7±47.6

Visit 2: 

37.1±48.1

Not reported
May 2011 4-74 DRAFT – DO NOT CITE OR QUOTE

Reference Cohort Age 

(years)

Number 

of 

Subjects

Location Study 

Period

Prior Pb 

Exposure

Bone Pb 

biomarker

Bone Pb 

Concentration 

(μg/g)

Distribution 

of Bone Pb 

(μg/g)

Hanninen et 

al. (1998), as 

reported in

Shih et al., 

(2007)

Not reported Mean±SD: Male: 43

Female: 48

BPb(max) 

≤ 2.4 

μmol/L: 41.7 

± 9.3

BPb (max)

>2.4 μmol/L: 

46.6 ± 6.2

54

(43 males, 

11 

females)

Helsinki, 

Finland

Not 

reported

Occupational 

(Pb acid 

battery 

factory 

workers)

Tibia

Calcaneus

Mean±SD:

Tibia:

BPb (max) 

≤ 2.4 μmol/L: 

19.8 ± 13.7

BPb (max) 

>2.4 μmol/L: 

35.3 ± 16.6

Calcaneus: 

BPb (max) 

≤ 2.4 μmol/L: 

78.6 ± 62.4

BPb (max) 

>2.4 μmol/L: 

100.4 ± 43.1

Not reported

Hsieh et al. 

2009 (2009)

Not reported Mean:

Cases: 

45.71

Controls: 

46.06

22 cases

18 controls

Location NR Not 

reported

Occupational

(Pb paint 

factory 

workers)

Tibia

Patella

Mean ± SD

Tibia

Case: 61.55 ± 30.21

Control: 18.51 ± 

22.40

Patella

Case: 66.29 ± 19.48

Control: 7.14 ± 

9.81

Not reported

Kamel et al.

(2002); 

Kamel et al. 

(2005); 

Kamel et al. 

(2008)

Not reported 30-80 109 cases

256 

controls

(Bone 

samples 

collected 

from 104 

cases and 

41 

controls)

New 

England

(Boston, 

MA)

1993-1996 Cumulative

Occupational 

(Pb fumes, 

dust, or 

particles)

Tibia

Patella

Mean ± SE

Tibia

Cases: 14.9 ± 1.6

Controls: 11.1 ± 

1.6

Patella

Cases: 20.5 ± 2.1

Controls: 16.7 ± 

2.0

Cases

Tibia Pb: N (%)

-7-7: 21 (20)

8-14: 35 (34)

15-61: 48 (46)

Patella Pb: N 

(%)

-4-9: 27 (26)

10-20: 40 (38)

21-107: 37 (36)

Controls

Tibia Pb: N (%)

-7-7: 14 (34)

8-14: 12 (29)

15-61: 15 (37)

Patella Pb: N 

(%) 

-4-9: 14 (34)

10-20: 14 (34)

21-107: 13 (32)

Khalil et al. 

(2009)

1982 Pb 

Occupationa

l Study 

cohort

Mean:

Cases: 54

Controls: 55

83 cases

51 controls

Eastern 

Pennsylvani

a

1982-2004 Occupational 

(Pb battery 

plant 

workers)

Tibia Median (IQR)

Tibia

Cases: 57 (20- 86)

Controls: 12 

(-8-32)

Not reported

Osterberg et 

al. (1997), as 

reported in

Shih et al.

(2007)

Not reported Median: 

41.5

38 male 

cases

19 male 

controls

Not reported Not 

reported

Occupational 

(secondary 

Pb smelter –

inorganic Pb)

Finger bone Median

Finger Bone:

High Cases: 32

Low cases: 16

Control: 4

Range

Finger Bone:

High Cases: 

17-101

Low cases: -7- 49

Control: -19-18

Not reported
May 2011 4-75 DRAFT – DO NOT CITE OR QUOTE

Reference Cohort Age 

(years)

Number 

of 

Subjects

Location Study 

Period

Prior Pb 

Exposure

Bone Pb 

biomarker

Bone Pb 

Concentration 

(μg/g)

Distribution 

of Bone Pb 

(μg/g)

Roels et al. 

(1994)

Not reported 30-60 76 male 

cases

68 male 

controls

Belgium Not 

reported

Occupational 

(Pb smelter 

workers)

Mean case 

exposure: 18 

yr (range: 6 to 

36 yr)

Tibia Geometric Mean

(Range)

Tibia Cases:

Normotensive: 

64.0 (19.6-

167.1)

Hypertensive: 

69.0 (21.7-

162.3)

Total: 65.8 (19.6-

167.1)

Tibia Controls:

Normotensive: 

21.7 (<15.2- 69.3)

Hypertensive: 

20.2 (<15.2- 52.9)

Total: 21.4 

(<15.2-69.3)

Not reported 

Schwartz 

(2000) et al., 

as reported 

in Shih et al., 

(2007)

U.S. 

Organolead 

Study

Mean ± SD:

Cases: 55.6 

± 7.4

Controls: 

58.6 ± 7.0

535 male 

cases

118 male 

controls

Eastern U.S. 6/1994- 10/1997 

(enrolled)

Completed 

2-4 annual 

follow-up 

visits

Tibia Pb 

taken in 

3rd year

Occupational

(tetraethyl 

and 

tetramethyl 

Pb 

manufacturin

g facility)

Tibia Mean ± SD

Current Tibia:

Cases: 14.4 ± 9.3

Peak Tibia:

Cases: 22.6 ± 16.5

Not reported

Schwartz et 

al. (2000), as 

reported in

Navas-Acien 

et al.(2008)

Not reported 41.7-73.7 

(Combined)

Mean ± SD: 

Combined: 

57.6 ± 7.6

Hypertensiv

e: 60.2 ± 6.9

Nonhyperte

nsive: 56.6 

± 7.5

543 males Eastern U.S. 1995 

(recruited)

1996-1997

(Tibia Pb 

taken 

during the 

3rd yr)

Occupational

(former 

organolead 

manufacturin

g workers)

Tibia Mean ± SD

Tibia:

Combined: 14.4 

± 9.3

Hypertensive: 

15.4 ± 9.1

Nonhypertensive

: 14.0 ± 9.3 

Range Tibia:

Combined: -1.6-

52

Not reported

Schwartz et 

al. (2001); 

Lee et al. 

(2001)

Not reported Mean:

Exposed: 

40.4

Control: 

34.5

803 cases

135 

controls

South Korea 10/24/1997

-8/19/1999

Occupational 

(battery 

manufacturin

g, secondary 

smelting, Pb 

oxide 

manufacturin

g, car radiator 

manufacturin

g)

Tibia Mean ± SD

Tibia

Cases: 37.1 ± 40.3

Control: 5.8 ± 

7.0

Range:

Tibia

Cases: -7-338

Controls: -11-27

Not reported

Schwartz et 

al. (2005)

Not reported Mean at 1st

visit: 41.4

576 South Korea 10/1997-

6/2001

Occupational 

(current and 

former Pb 

workers)

Tibia Mean ± SD

Tibia: 38.4 ± 43

Tibia:

25th percentile 

at V1: 14.4

75th percentile 

at V1: 47.1
May 2011 4-76 DRAFT – DO NOT CITE OR QUOTE

Reference Cohort Age 

(years)

Number 

of 

Subjects

Location Study 

Period

Prior Pb 

Exposure

Bone Pb 

biomarker

Bone Pb 

Concentration 

(μg/g)

Distribution 

of Bone Pb 

(μg/g)

Stewart et al.

(1999), as 

reported in

Shih et al., 

(2007)

U.S. 

Organolead 

Study

40-70 (in 

1995)

38% ≥ 60 

yrs

Mean: 58

534 males Eastern U.S. Not 

reported

Occupational

(tetraethyl 

and 

tetramethyl 

Pb 

manufacturin

g facility)

Tibia Mean ± SD

Tibia:

Current: 14.4 ± 

9.3

Peak: 23.7 ± 

17.4

Range: Tibia

Current: -1.6-52

Peak: -2.2-105.9

Current Tibia 

Pb: N (%)

<5: 77 (14.2)

5-9.99: 113 

(20.8)

10-14.99: 119 

(21.9)

15-19.99: 117 

(21.5)

≥ 20: 118 

(21.7)

Peak Tibia Pb: 

N (%)

<5: 49 (9.1)

5-9.99: 64 

(11.8)

10-14.99: 70 

(12.9)

15-19.99: 87 

(16.1)

20-24.99: 79 

(14.6)

25-29.99: 55 

(10.2)

≥ 30: 137 

(26.1)

Stewart et al. 

(2006)

Not reported Mean: 56.1 532 males Eastern U.S. 1994-

1997; 

2001-2003

Cumulative

Occupational

(Organolead 

workers - not 

occupationall

y exposed to 

Pb at time of 

enrollment)

Tibia Mean ± SD

Current Tibia: 

14.5 ± 9.6

Peak Tibia: 23.9 

± 18.3

Not reported

Weaver et al. 

(2008)

Not reported Mean ± SD: 

43.3 ± 9.8

652 South Korea 12/1999-

6/2001

Occupational

(Current and 

former Pb 

workers; 

plants 

produced Pb 

batteries, Pb 

oxide, Pb 

crystal, or 

radiators, or 

were 

secondary Pb 

smelters)

Patella Mean±SD

Patella: 37.5 ± 41.8

Not reported

4.4.3. Lead in Urine

1 Urine Pb concentrations in the U.S. general population have been monitored in the NHANES. Data 

2 from the most recent survey (CDC, 2011) are shown in Table 4-10. The geometric mean for the entire 

3 sample for the period 2007-2008 (n = 2,627) was 0.52 μg/g creatinine (95% CI: 0.48, 0.55). The 

4 geometric means for males (n = 1,327) and females (n = 1,300) were 0.50 μg/g creatinine (95% CI: 0.47, 

5 0.53) and 0.53 μg/g creatinine (95% CI: 0.49, 0.57), respectively.
May 2011 4-77 DRAFT – DO NOT CITE OR QUOTE

Table 4-10. Urine Pb concentrations in the U.S. population

Survey Stratum Period Geometric Mean (μg/g CR) 95% Confidence Interval Number of Subjects

All 1999-2000 0.721 0.700, 0.742 2465

2001-2002 0.639 0.603, 0.677 2689

2003-2004 0.632 0.603, 0.662 2558

2005-2006 0.546 0.502, 0.573 2576

2007-2008 0.515 0.483, 0.549 2627

6-11 yr 1999-2000 1.170 0.975, 1.41 340

2001-2002 0.918 0.841, 1.00 368

2003-2004 0.926 0.812, 1.06 290

2005-2006 0.628 0.563, 0.701 355

2007-2008 0.644 0.543, 0.763 394

12-19 yr 1999-2000 0.496 0.460, 0.535 719

2001-2002 0.404 0.380, 0.428 762

2003-2004 0.432 0.404, 0.461 725

2005-2006 0.363 0.333, 0.395 701

2007-2008 0.301 0.270, 0.336 376

≥ 20 yr 1999-2000 0.720 0.683, 0.758 1406

2001-2002 0.658 0.617, 0.703 1559

2003-2004 0.641 0.606, 0.679 1543

2005-2006 0.573 0.548, 0.600 1520

2007-2008 0.546 0.513, 0.580 1857

Males 1999-2000 0.720 0.679, 0.763 1227

2001-2002 0.639 0.607, 0.673 1334

2003-2004 0.615 0.588, 0.644 1281

2005-2006 0.551 0.522, 0.582 1271

2007-2008 0.502 0.471, 0.534 1327

Females 1999-2000 0.722 0.681, 0.765 1238

2001-2002 0.639 0.594, 0.688 1355

2003-2004 0.648 0.601, 0.698 1277

2005-2006 0.541 0.507, 0.577 1305

2007-2008 0.527 0.489, 0.568 1300

Mexican - Americans 1999-2000 0.940 0.876, 1.01 884

2001-2002 0.810 0.731, 0.898 682

2003-2004 0.755 0.681, 0.838 618

2005-2006 0.686 0.638, 0.737 652

2007-2008 0.614 0.521, 0.722 515

Non-Hispanic blacks 1999-2000 0.722 0.659, 0.790 568

2001-2002 0.644 0.559, 0.742 667

2003-2004 0.609 0.529, 0.701 723

2005-2006 0.483 0.459, 0.508 692

2007-2008 0.452 0.414, 0.492 589

Non-Hispanic whites 1999-2000 0.696 0.668, 0.725 822

2001-2002 0.615 0.579, 0.654 1132

2003-2004 0.623 0.592, 0.655 1074

2005-2006 0.541 0.500, 0.585 1041

2007-2008 0.506 0.466, 0.550 1095

Values are μg Pb/g creatinine

Source: Based on data from the NHANES (CDC, 2011)

4.4.4. Lead in Teeth

1 The influence of historical Pb exposures was recently studied by Robbins et al. (2010). Tooth 

2 enamel samples from 127 subjects born between 1936 and 1993 were analyzed for Pb concentration and 

3 Pb isotope ratios of the tooth enamel and compared with those parameters for sediment cores and 

4 estimates of Pb emissions from gasoline during the years when 50% enamel formation was estimated to 

5 occur. They found that the log-transform of tooth enamel concentration was significantly predicted by the 

6 log-transform of Lake Erie sediment core data obtained by Graney et al. (1995) (p <0.00001) and by the 

7 log-transform of U.S. consumption of Pb in gasoline (p <0.00001); Figure 4-18. Additionally, Robbins et 
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al. (2010) found that 207Pb/206Pb was significantly predicted by the 207Pb/206 1 Pb observed in the Lake Erie 

2 sediment cores obtained by Graney et al. (1995) (p <0.0001) and for this study (p <0.0002).

Source: Used with permission from Elsevier Publishing, Robbins et al. (2010).

Figure 4-18. Comparison of tooth enamel50% formation (solid line) with 

newly obtained Pb sediment Lake Erie cores (open triangles), 

previously obtained Lake Erie sediment (open circles, Graney 

et al. (1995)), and U.S. gasoline usage (closed circles). All 

values are normalized by the peak observation for that 

parameter.

4.5. Empirical Models of Lead Exposure-Blood 

Lead Relationships

3 Multivariate regression models, commonly used in epidemiology, provide estimates of the 

4 contribution of variance in the internal dose metric to various determinants or control variables (e.g., air 

5 Pb concentration, surface dust Pb concentration). Structural equation modeling links several regression 

6 models together to estimate the influence of determinants on the internal dose metric. Regression models 

7 can provide estimates of the rate of change of blood or bone Pb concentration in response to an 

8 incremental change in exposure level (i.e., slope factor). One strength of regression models is that they are 

9 empirically verified within the domain of observation and have quantitative estimates of uncertainty 

10 imbedded in the model structure. However, regression models are based on (and require) paired predictor11 outcome data, and, therefore, the resulting predictions are confined to the domain of observations and are 

12 typically not generalizable to other populations. Regression models also frequently exclude numerous

13 parameters that are known to influence human Pb exposures (e.g., soil and dust ingestion rates) and the 
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1 relationship between human exposure and tissue Pb levels, parameters which are expected to vary 

2 spatially and temporally. Thus, extrapolation of regression models to other spatial or temporal contexts, 

3 which is often necessary for regulatory applications of the models, can be problematic.

4.5.1. Air Lead-Blood Lead Relationships

4 The 2006 AQCD and its 1986 predecessor (U.S. EPA, 1986, 2006) described epidemiological 

5 studies of relationships between air Pb and blood Pb. Much of the pertinent earlier literature (e.g., prior to 

6 1984) was summarized by Brunekreef (1984). Based on meta-analysis of 18 studies of urban or industrial7 urban populations, Brunekreef (1984) estimated the blood Pb-air Pb slope for children to be 0.3485 

ln[μg/dl blood Pb] per ln[μg/m3 air Pb] (R2 8 = 0.69; Figure 4-19). This corresponds to an increase of 6.3

μg/dL blood Pb for an increase in air Pb concentration from 0.15 to1.5 μg/m3 9 . When the analysis was 

10 limited to children whose blood Pb concentrations were <20 μg/dL, the regression coefficient was 0.2159 

(R2 11 =0.33), which corresponds to an increase of 3.2 μg/dL blood Pb for an increase in air Pb from 0.15 

to1.5 μg/m3 12 . Blood Pb-air Pb slopes are presented for recent studies in the following paragraphs. These 

13 data are summarized in Table 4-11.
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Table 4-11. Summary of estimated slopes for blood Pb to air Pb relationships in humans

Reference Study Methods Model Description Blood Pb–Air Pb 

Slope (μg/dL/μg/m3)

Children Populations

Brunekreef et al. (1984) Location: Various countries

Years: 1974-1983

Subjects: Children (varying age ranges)

Analysis: Meta analysis of 18 studies

Model: Log-Log

Blood Pb: 5-41 μg/dL (mean 

range for studies)

Air Pb: 0.2-10 μg/m3 (mean 

range for studies)

All children: 18a, 6.1b

Children <20 μg/dl: 13a, 

3.0b

Hayes et al. (1994) Location: Chicago, IL

Years: 1974-1988

Subjects: 0.5-6 yr (9,604 blood Pb measurements)

Analysis: Regression of blood Pb screening and quarterly 

average air Pb 

Model: Log-Log

Blood Pb: 12-30 μg/dL 

(annual GM range)

Air Pb: 0.5-1.2 μg/m3 (annual 

GM range)

24a, 5.7b

Hilts et al. (2003) Location: Trail, BC

Years: 1989-2001

Subjects: 0.5-6 yr (292-536 blood Pb measurements/yr)

Analysis: Regression of blood Pb screening and community air 

Pb following upgrading of a local smelter

Model: Linear

Blood Pb: 4.7-11.5 μg/dL 

(annual median range)

Air Pb: 0.03-1.1 μg/m3

(annual median range)

6.5

Ranft et al. (2008) Location: Germany

Years: 1983-2000

Subjects: 6-11 yr (n = 843) 

Analysis: Pooled regression 5 cross-sectional studies

Model: Log-Linear

Blood Pb: 2.2-13.6 μg/dL (5th95th percentile)

Air Pb: 0.03-0.47 μg/m3 (5th95th percentile)

3.2c

Schnaas et al. (2004)

Location: Mexico City

Years: 1987-2002

Subjects: 0.5-10 yr (n = 321)

Analysis: Regression of longitudinal blood Pb measurements 

and annual average air Pb data

Model: Log-Log

Blood Pb: 5-12 μg/dL (annual 

GM range)

Air Pb: 0.7-2.8 μg/m3 (annual 

mean range)

4.8a, 1.1b

Schwartz and Pitcher 

(1989), U.S. EPA (1986)

Location: US

Years: 1976-1980

Subjects: 0.5-7 yr (n = 7,000)

Analysis: NHANES blood Pb, gasoline consumption data and 

Pb concentrations in gasoline

Model: Linear

Blood Pb: 11-18 μg/dL (mean 

range)

Air Pb: 0.36-1.22 μg/m3

(annual maximum quarterly 

mean) 

9.3

Schwartz and Pitcher 

(1989), U.S. EPA (1986)

Location: Chicago, IL

Years: 1976-1980

Subjects: 0-5 yr (n = 7,000)

Analysis: Chicago blood Pb screening, gasoline consumption 

data, and Pb concentrations in gasoline

Model: Linear

Blood Pb: 18-27 μg/dL (mean 

range)

Air Pb: 0.36-1.22 μg/m3

(annual maximum quarterly 

mean)

7.7

Tripathi et al. (2001) Location: Mumbai, India

Years: 1984-1996

Subjects: 6-10 yr (n = 544)

Analysis: Regression of blood Pb and air Pb data

Model: Linear

Blood Pb: 8.6-14.4 μg/dL 

(regional GM range)

Air Pb: 0.11-1.18 μg/m3

(regional GM range)

3.6

Adult Populations

Rodrigues et al. (2010) Location: New England, U.S.

Years: 1994-1995

Subjects: Adult bridge painters (n-84, 1 female)

Analysis: Regression analysis of blood Pb and air Pb data 

(personal monitors) collected during work performing various 

job-related tasks

Model: Log-log

Blood Pb: 16.1 μg/dL (GM, 

1,7 GSD)

Air Pb: 58 μg/m3 (GM, 2.8 

GSD)

5.4a, 4.2b

a

At an air concentration of 0.15 μg/m3

b

At an air concentration of 1 μg/m3

c

For a change in air Pb concentration from 0.025 to 0.465 μg/m3

GM, geometric mean; GSD, geometric standard deviation
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Data provided from Brunekreef (1984).

Figure 4-19. Predicted relationship between air Pb and blood Pb based on

a meta analysis of 18 studies. The regression model is: 

ln[μg/dL blood Pb] = 0.3485⋅ln[μg/m3 air Pb] + 2.85 for all 

children and ln[μg/dL blood Pb] = 0.2159⋅ln[μg/m3 air Pb] + 

2.62 when the sample was restricted to populations that had 

blood Pb concentrations <20 μg/dL.

4.5.1.1. Children

1 Hilts et al. (2003) reported child blood Pb and air Pb trends for the city of Trail, British Columbia, 

2 over a period preceding and following installation of a new smelter process in 1997 which resulted in 

3 lower air Pb concentrations. Blood Pb data were obtained from annual (1989-2001) surveys of children 6-

4 60 months of age (n: 292-536 per year) who lived within 4 km from the smelter. Air Pb concentrations

5 were obtained from high volume suspended particulate samplers placed within 2 km of the smelter that 

6 operated 24 hours every 6th day. Data on Pb levels in air, residential soil, interior dust, and blood for three 

7 sampling periods are summarized in Table 4-12. Based on these data, blood Pb decreased 6.5 μg/dL per 1 

μg/m3 8 air Pb and by 0.068 μg/dL per mg/kg soil Pb (based on linear regression with air or soil Pb as the 

9 sole independent variable). Several uncertainties apply to these estimates. Potential mismatching of air Pb 

10 concentrations (often termed misclassification) with individual blood Pb levels may have occurred as a 

11 result of air Pb being measured within 2 km of the smelter, whereas, the blood Pb data included children 

12 who resided >2 km from the smelter. The regression estimates were based on group mean estimates for 

13 three sampling dates, rather than on the individual blood Pb estimates, which included repeated measures 

0

10

20

30

40

50

60

0 5 10 15 20 25

Blood Pb (

μg/dL)

Air Pb (μg/m3)

All children < 20 ug/dL
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1 on an unreported fraction of the sample. The limited number of data pairs (three) constrained parameter 

2 estimates to simple regression coefficients. Other important factors probably contributed to blood Pb 

3 declines in this population that may have been correlated with air, soil and dust Pb levels. These include 

4 aggressive public education and exposure intervention programs (Hilts, 1996; Hilts et al., 1998). 

5 Therefore, the coefficients shown in Table 4-12 are likely to overestimate the influence of air, dust, or soil 

6 Pb on blood Pb concentrations at this site.

Table 4-12. Environmental Pb levels and blood Pb levels in children in Trail, British 

Columbia

Date 1996 1999 2001 Regression Coefficient

(μg/dL per μg/m3)

Blood Pb (μg/dL) 11.5 5.9 4.7 NA

Air Pb (μg/m3) 1.1 0.3 0.03 6.5 ± 0.52 (R2=0.99, p=0.050)

Soil Pb (mg/kg) 844 756 750 0.068 ± 0.008 (R2=0.99, p=0.069)

Interior Dust Pb (mg/kg) 758 583 580 0.035 ± 0.005 (R2=0.98, p=0.097)

A new smelter process began operation in 1997. Values for air, soil and dust Pb are annual averages; values for blood Pb are annual geometric means. Regression coefficients 

are for simple linear regression of each exposure variable on blood Pb.

Source: Data from Hilts et al. (2003).

7 Ranft et al. (2008) reported a meta-analysis of five cross-sectional surveys of air and soil Pb levels 

8 and blood Pb concentrations in children living in Duisburg, Germany. The analysis included observations 

9 on 843 children (6-11 years of age) made during the period 1983-2000. Pb was measured in PM10 samples 

10 collected in a 200 meter by 200 meter grid that encompassed the city. Pb in surface soil (0-10 cm) was 

11 measured at 145 locations in the city. Air and soil Pb concentrations were assigned to each participant by 

spatial interpolation from the sampling grid data to each home residence. The 5th 12 -95th percentile ranges 

were 0.025-0.465 μg Pb/m3 13 for air and 72-877 mg Pb/kg for soil. The results of multivariate regression 

14 analyses were reported in terms of the relative increase (the geometric mean blood Pb ratio, GMR) for an 

increase in air or soil Pb from the 5th to 95th percentile value. In a multivariate linear regression model (R2 15

16 = 0.586) that included air and soil Pb in the same model and adjusted for covariates, the GMR values 

were: 2.55 per 0.44 μg/m3 increase in air Pb (95% CI: 2.40, 2.71, R2 17 =0.484, p<0.001) and 1.30 per 800 

mg/kg soil Pb (95% CI: 1.19, 1.43, R2 = 0.017, p <0.001). Based on the values for R2 18 , the regression 

19 model accounted for approximately 59% of the total variance in blood Pb and, of this, 83% was attributed 

20 to air Pb. Values for GMR for soil Pb varied depending on the sampling data and ranged from 1.41 to 

21 2.89, with most recent data (2000) yielding a value of 1.63 per 800 mg/kg increase in soil Pb. The GMR 

22 values can be converted to regression slopes (slope = starting blood Pb×ln(GMR)/5th-95th percentile air 

23 or soil Pb) for calculating equivalent air:blood Pb ratios. The model predicts an increase of 3.2 μg/dL 

blood Pb per 1 μg/m3 24 increase in air Pb. Based on the GMR estimate of 1.63 for soil Pb, a 1,000 mg/kg 

25 increase in soil Pb would be associated with an increase in blood Pb of 0.6 μg/dL per mg/kg soil. The 

26 degree of confounding of the GMR and estimates resulting from the air and soil Pb correlation was not 
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1 reported, although the correlation coefficient for the two variables was 0.136 for the whole data set and 

2 0.703 when data collected in 1983 was omitted. The Ranft et al. (2008) model is log-linear, with the 

3 natural logarithm of blood Pb being a function of linear increase in air Pb. This results an upward 

4 curvature of the blood Pb-air Pb relationship (i.e., in linear scale, the blood Pb-air Pb slope increases with 

5 increasing air Pb concentration). By comparison, log-log models predict an increase in the blood Pb-air 

6 Pb slope with decreasing air Pb concentration, whereas linear models predict a constant blood Pb-air Pb 

7 slope across all air Pb concentrations.

8 Schnaas et al. (2004) analyzed data on blood Pb and air Pb concentrations during and after the 

9 phase out of leaded gasoline use in Mexico (1986-1997) in children as part of a prospective study 

10 conducted in Mexico City. The sample included 321 children born during the period 1987 through 1992. 

11 Repeated blood Pb measurements were made on each child at 6-month intervals up to age 10 years. Air 

12 Pb measurements in PM10 (annual average of quarterly means) were derived from three area monitors 

13 which represented distinct study zones. Children were assigned to study zones based on their current 

14 address and were assigned the corresponding annual average air Pb concentrations for appropriate air 

15 monitoring zones. Associations between blood Pb concentration, air Pb concentration and other variables 

16 (e.g., age, year of birth, family use of glazed pottery) were evaluated using multivariate regression 

models. The regression model (r2 17 = 0.96) predicted blood Pb-air Pb slopes that decreased with year of 

18 birth. The largest slope occurred in the cohort born in 1987, who experienced the largest decline in air Pb 

(from 2.8 to <0.1 μg/m3 19 ); the predicted slope for this group of children was 0.213 (95% CI: 0.114-0.312) 

ln [μg/dL blood] per ln[μg/m3 20 air]. This slope corresponds to an increase of 2.1 μg/dL blood Pb for an 

increase in air Pb from 0.15 to1.5 μg/m3 21 .

22 Schwartz and Pitcher (1989) reported a multivariate regression analysis of associations between 

23 U.S. gasoline Pb consumption (i.e., sales) and blood Pb concentrations in the U.S. population during the 

24 period 1976-1980 when use of Pb in gasoline was being phased out. Although this analysis did not 

25 directly derive a slope for the air Pb-blood Pb relationships, other analyses have shown a strong 

26 correlation between U.S. gasoline Pb consumption and ambient air Pb levels during this same period 

27 (U.S. EPA, 1986). Therefore, it is possible to infer an air Pb-blood Pb relationship from these data. Two 

28 sources of blood Pb data were used in Schwartz and Pitcher (1989): NHANES II provided measurements 

29 for U.S. children 6 months to 7 years of age (n = 9,996) during 1976-1980, and the City of Chicago blood 

30 Pb screening program provided approximately 7,000 blood Pb measurements in black children during 

31 1976-1980. Gasoline Pb consumption was estimated as the product of monthly gasoline sales in the U.S. 

32 and quarterly estimates of Pb concentrations in gasoline reported to U.S. EPA. Based on the NHANES 

33 blood Pb data for white children, the regression coefficient was 2.14 μg/dL blood per 100 metric tons of 

34 gasoline Pb/day (SE=0.19, p=0.0000); results for black children were essentially identical. Based on the 

35 Chicago blood Pb data the regression coefficient was 16.12 (μg/dL per 1,000 metric tons gasoline 

36 Pb/quarter (SE=1.37, p=0.0001), which is roughly equivalent to 1.79 μg/dL blood per 100 metric tons of 
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1 gasoline Pb/day. U.S. EPA (1986) reported data on gasoline Pb consumption (sales) and ambient Pb levels 

2 in the U.S. during the period 1976-1984 (Table 4-13). Based on these data, air Pb concentrations

3 decreased in association with gasoline Pb consumption. The linear regression coefficient for the air Pb 

decrease was 0.23 μg/m3 per 100 metric tons gasoline Pb/day (SE = 0.02, R2 4 = 0.95, p <0.0001). If this 

5 regression coefficient is used to convert the blood Pb slopes from Schwartz and Pitcher (1989), the 

corresponding air Pb-blood Pb slopes would be 9.3 and 7.8 μg/dL per μg/m3 6 , based on the NHANES and 

7 Chicago data, respectively (e.g., 2.14/0.23 = 9.3).

Table 4-13. U.S. gasoline Pb consumption and air Pb levels

Date Total Gasoline Pb

(103 metric tons/yr)

Total Gasoline Pb

(102 metric tons/day)a Air Pb(μg/m3)

1976 171.4 4.70 1.22

1977 168.9 4.63 1.20

1978 153 4.19 1.13

1979 129 3.53 0.74

1980 78.8 2.16 0.66

1981 60.7 1.66 0.51

1982 59.9 1.64 0.53

1983 52.3 1.43 0.40

1984 46 1.26 0.36

The linear regression coefficient is 0.23 μg/m3 air per 100 metric tons/day (SE= 0.020, R2 = 0.95, p<0.0001). a

Conversion factor is 10/365 days/year.

Source: U.S. EPA (1986).

8 Tripathi et al. (2001) reported child blood Pb and air Pb trends for the city and suburbs of Mumbai, 

9 India over the period 1984-1996. Blood Pb data were obtained from children 6-10 years of age (n = 544) 

10 who lived in 13 locations within the Mumbai area. Air Pb concentrations were measured from high 

11 volume PM samplers (with the majority of Pb in the respirable size range) placed at a height of 1.6 meters 

12 that operated 24 hours. Data on Pb concentrations in air, residential soil, interior dust, and blood for three 

13 sampling periods are summarized in Table 4-14. Based on these data, blood Pb increased 3.6 μg/dL per 1 

μg/m3 14 air Pb (based on linear regression with air or soil Pb as the sole independent variable). Several 

15 uncertainties apply to these estimates, including potential exposure misclassification since the mean air Pb 

16 concentration was used for each suburb over the entire study period. The regression estimates were based 

17 on group mean blood Pb estimates for the 13 sampling locations, rather than on the individual blood Pb 

18 estimates, which included repeated measures on an unreported fraction of the sample.
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Table 4-14. Air Pb levels and blood Pb levels in children in Mumbai, India

Blood Pb (μg/dL) Air Pb (μg/m3

)

Location N GM GSD N GM GSD

Borivilli 12 10.4 1.67 10 0.32 1.51

Byculla 117 11.0 1.99 30 0.99 1.73

Deonar 46 9.5 2.29 93 0.11 3.21

Goregaon 21 9.1 1.30 24 0.35 1.77

Govandi 20 8.9 1.42 10 0.10 1.52

Jogeshwari 20 8.6 1.32 24 0.11 2.47

Khar 17 9.0 1.53 22 0.18 3.15

Parel 168 10.4 1.91 37 0.44 1.48

Sion 34 9.6 1.49 96 0.39 1.75

Thans (SS) 37 12.0 1.86 4 1.18 1.04

Vile Parle 19 9.1 1.46 7 0.37 1.34

Colaba 12 9.2 1.86 9 0.14 1.63

Vakola 21 14.4 1.64 7 1.12 1.12

The linear regression coefficient is 3.62 μg/dL blood per μg/m3 air (SE= 0.61, R2

= 0.76, p<0.001).

GM, geometric mean; GSD, geometric standard deviation; N, number of subjects.

Source: Data are from Tripathi et al. (2001). 

1 Hayes et al. (1994) analyzed data collected as part of the Chicago, IL blood Pb screening program 

2 for the period 1974-1988, following the phase-out of leaded gasoline. The data included 9,604 blood Pb 

3 measurements in children (age: 6 months to 6 years) and quarterly average air Pb concentrations

4 measured at 12 monitoring stations in Cook County, IL. Quarterly median blood Pb levels declined in 

5 association with quarterly mean air Pb concentrations. The regression model predicted a slope of 0.24 ln 

[μg/dL blood] per ln[μg/m3 6 air], as illustrated in Figure 4-20. This corresponds to an increase of 11.1 

μg/dL blood Pb for an increase in air Pb from 0.15 to1.5 μg/m3 7 .
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Modified from Hayes et al. (1994).

Figure 4-20. Predicted relationship between air Pb and blood Pb based on 

data from Chicago, IL (1974–1988).The regression model is: 

ln[μg/dL blood Pb] = 0.24⋅ln[μg/m3 air Pb] + 3.17.

4.5.1.2. Adults

1 Rodrigues et al. (2010) examined factors contributing to variability in blood Pb concentration in 

2 New England bridge painters, who regularly use electric grinders to prepare surfaces for painting. The 

3 study included 84 adults (1 female) who were observed during a 2-week period in 1994 or 1995. Subjects 

4 wore personal inhalable PM samplers designed to capture PM smaller than 100 μm, while performing 

various job-related tasks. The geometric mean air Pb concentration for the 2-week period was 58 μg/m3 5

(GSD 2.8), with a maximum daily value of 210 μg/m3 6 . The Pb concentrations reported were corrected by 

7 the National Institute for Occupational Safety and Health (NIOSH) respirator protection factors, which 

8 were not reported by the authors. Hand wipe samples were collected at the mid-shift break and at the end 

9 of the shift (after the subjects had reportedly cleaned up for the day; GM = 793 μg, GSD 3.7). Blood Pb 

10 samples were collected at the beginning of the 2-week period (GM = 16.1 μg/dL, GSD 1.7). Associations 

11 between exposure variables and blood Pb concentrations were explored with multivariate regression 

12 models (Table 4-15). When the model excluded hand-wipe data (not all participants who wore the 

13 personal air samplers agreed to provide hand-wipes), the regression coefficient for the relationship 

between ln[blood Pb concentration (μg/dL)] and ln[air Pb (μg/m3 14 )] was 0.11 (SE = 0.05, p = 0.03). This 

15 corresponds to a 1.3-fold increase in blood Pb concentration for a 10-fold increase in air Pb concentration. 

0

5

10

15

20

25

30

0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6

Blood Pb (

μg/dL)

Air Pb (μg/m3)
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1 A second regression model included hand wipe Pb (n = 54) and yielded a regression coefficient of 0.05 

2 (SE = 0.07, p = 0.45), which corresponds to a 1.12-fold increase in blood Pb concentration per 10-fold 

3 increase in air Pb concentration.

Table 4-15. Significant predictors of blood Pb concentration in bridge painters

Parameters

Blood Pb (Air Only) Blood Pb (Air and Hand Wipe)

β(SE) p-value β(SE) p-value

Intercept 1.90 (0.24) <0.0001 2.12 (0.44) 0.0007

Time of blood Pb (end vs start of study) 0.16 (0.04) <0.0001 –0.31 (0.11) 0.005

Mean air Pb (μg/m3) 0.11 (0.05) 0.03 0.05 (0.07) 0.45

Hand wipe at break (μg Pb) — 0.007 (0.06) 0.91

Hand wipe at break * time of blood Pb — 0.07 (0.01) <0.0001

Months on bridge painting crews 0.001 (0.0004) 0.03 0.001 (0.0006) 0.04

Education

≤ High school

> High school

0.38 (0.10)

Reference 0.0002 0.29 (0.13)Reference 0.03

Respirator fit test

No

Yes

–0.14 (0.14)

Reference 0.32 –0.13 (0.21)Reference 0.53

Respirator fit test * time of blood Pb

No

Yes

0.18 (0.06)

Reference 0.003 0.17 (0.07)Reference 0.01

Smoke on site

No

Yes

0.14 (0.09) 

Reference 0.14 0.15 (0.10)Reference 0.14

Smoke on site * time of blood Pb

No

Yes

–0.15 (0.05)

Reference 0.002 –0.11 (0.04)Reference 0.009

Personal hygiene index

Low

High

0.27 (0.11)

Reference 0.02 0.29 (0.12)Reference 0.02

Site-level variables

Containment facility

Poor

Good

–0.59 (0.18)

Reference 0.001 –0.57 (0.22)Reference 0.01

Air Pb, hand wipe, and blood Pb levels are natural log-transformed.

Blood Pb concentration in units of μg/dL.

Source: Data from Rodrigues et al. (2010). 

4.5.2. Environmental Lead-Blood Lead Relationships

4 Empirically-based relationships between blood Pb levels and Pb intakes and/or Pb concentrations

5 in environmental media have provided the basis for what has become known as slope factor models. 

6 Slope factor models are highly simplified representations of empirically based regression models in which 

7 the slope parameter represents the change in blood Pb concentration projected to occur in association with 

8 a change in Pb intake or uptake. The slope parameter is factored by exposure parameters (e.g., exposure 

9 concentrations, environmental media intake rates) that relate exposure to blood Pb concentration (Abadin 

10 & Wheeler, 1997; Bowers et al., 1994; Carlisle & Wade, 1992; Maddaloni et al., 2005; Stern, 1994, 1996; 

11 U.S. EPA, 2003). In slope factor models, Pb biokinetics are represented as a linear function between the 

12 blood Pb concentration and either Pb uptake (uptake slope factor, USF) or Pb intake (intake slope factor, 

13 ISF). The models take the general mathematical forms:
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PbB = E ⋅ ISF

Equation 4-2

PbB = E ⋅ AF ⋅USF

Equation 4-3

1 where PbB is the blood Pb concentration, E is an expression for exposure (e.g., soil intake × soil Pb 

2 concentration) and AF is the absorption fraction for Pb in the specific exposure medium of interest. Intake 

3 slope factors are based on ingested rather than absorbed Pb and, therefore, integrate both absorption and 

4 biokinetics into a single slope factor, whereas models that utilize an uptake slope factor include a separate 

5 absorption parameter. In contrast to mechanistic models, slope factor models predict quasi-steady state 

6 blood Pb concentrations that correspond to time-averaged daily Pb intakes (or uptakes) that occur over 

7 sufficiently long periods to produce a quasi-steady state (i.e., >75 days, ~3 times the t1/2 for elimination of 

8 Pb in blood).

9 The U.S. EPA Adult Lead Methodology (ALM) is a example of a slope factor model that has had 

10 extensive regulatory use in the EPA Superfund program for assessing health risks to adults associated with 

11 non-residential exposures to Pb in contaminated soils (Maddaloni et al., 2005; U.S. EPA, 1996a). The 

12 model was developed to predict maternal and fetal blood Pb concentrations that might occur in relation to 

13 maternal exposures to contaminated soils. The model assumes an uptake slope factor of 0.4 μg/dL blood 

14 per μg/day Pb uptake. Additional discussion of slope factor models that have been used or proposed for 

15 regulatory use can be found in the 2006 AQCD (U.S. EPA, 2006).

16 Previous studies included in the 2006 AQCD (U.S. EPA, 2006) explored the relationship between 

17 blood Pb in children and environmental Pb concentrations. In a pooled analysis of 12 epidemiologic 

18 studies, interior dust Pb loading, exterior soil/dust Pb, age, mouthing behavior, and race were all 

19 statistically significant variables included in the regression model for blood Pb concentration (Lanphear et 

20 al., 1998). Significant interactions were found for age and dust Pb loading, mouthing behavior and 

21 exterior soil/dust level, and SES and water Pb level. In a meta-analysis of 11 epidemiologic studies, 

22 among children the most common exposure pathway influencing blood Pb concentration in structural 

23 equation modeling was exterior soil, operating through its effect on interior dust Pb and hand Pb (Succop 

24 et al., 1998). Similar to Lanphear et al. (1998), in the linear regression model, interior dust Pb loading had 

25 the strongest relationships with blood Pb concentration. Individual studies conducted in Rochester, NY, 

26 Cincinnati, OH, and Baltimore, MD report similar relationships between children’s blood Pb and interior 

27 dust concentrations (Bornschein et al., 1985; Lanphear & Roghmann, 1997; U.S. EPA, 1996b).

28 Dixon et al. (2009) reported a multivariate analysis of associations between environmental Pb 

29 concentrations and blood Pb concentrations, based on data collected in the NHANES (1999-2004). The 

30 analyses included 2,155 children, age 12-60 months. The population-weighted geometric mean blood Pb 

concentration was 2.03 μg/dL (GSD 1.03). A linear model applied to these data yielded an R2 31 of 40% 

32 (Table 4-16). The regression coefficient for the relationship between ln[blood Pb concentration (μg/dL)] 
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and ln[floor dust Pb concentration (μg/ft2 1 )] was 0.386 (SE 0.089) for “not smooth and cleanable” surfaces 

2 (e.g., high-pile carpets) and 0.205 (SE 0.032) for “smooth and cleanable” surfaces (e.g., uncarpeted or 

3 low-pile carpets). These coefficients correspond to a 2.4-fold or 1.6-fold increase in blood Pb 

4 concentration, respectively, for a 10-fold increase in floor dust Pb concentration. 

Table 4-16. Linear model relating environmental Pb exposure and blood Pb concentration 

in childrena

Variables Overall p-value Levels Estimate (SE) p-Value

Intercept 0.172 - 0.517 (0.373 0.172

Age (in yr) < 0.001

Age

Age2

Age3

Age4

2.620 (0.628)

–1.353 (0.354)

0.273 (0.083)

–0.019 (0.007)

< 0.001

< 0.001

0.002

0.008

Yr of construction 0.014

Intercept for missing

1990–present

1978–1989

1960–1977

1950–1959

1940–1949

Before 1940

–0.121 (0.052)

–0.198 (0.058)

–0.196 (0.060)

–0.174 (0.056)

–0.207 (0.065)

–0.012 (0.072)

0.000

0.024

0.001

0.002

0.003

0.003

0.870

-

PIR < 0.001 Intercept for missingSlope

0.053 (0.065

–0.053 (0.012)

0.420

< 0.001

Race/ethnicity < 0.001

Non-Hispanic white

Non-Hispanic black

Hispanic

Other

0.000

0.247 (0.035

–0.035 (0.030)

0.128 (0.070)

-

< 0.001

0.251

0.073

Country of birth 0.002

Missing

United Statesb

Mexico

Elsewhere

–0.077 (0.219)

0.000

0.353 (0.097)

0.154 (0.121

0.728

-

< 0.001

0.209

Floor surface/condition × log 

floor PbD < 0.001

Intercept for missing

Not smooth and cleanable

Smooth and cleanable or 

carpeted

0.178 (0.094)

0.386 (0.089)

0.205 (0.032)

0.065

< 0.001

< 0.001

Floor surface/condition × log 

floor PbD)2

Not smooth and cleanable

Smooth and cleanable or 

carpeted

0.023 (0.015)

0.027 (0.008)

0.124

0.001

Floor surface/condition × (log 

floor PbD)3

Uncarpeted not smooth and 

cleanable

Smooth and cleanable or 

carpeted

–0.020 (0.014)

–0.009 (0.004)

0.159

0.012

Log windowsill PbD 0.002 Intercept for missing

Slope

0.053 (0.040

0.041 (0.011

0.186

< 0.001

Home-apartment type < 0.001

Intercept for missing

Mobile home or trailer

One family house, detached

One family house, attached

Apartment (1–9 units)

Apartment (≥ 10 units)

–0.064 (0.097

0.127 (0.067)

–0.025 (0.046)

0.000

0.069 (0.060)

–0.133 (0.056)

0.511

0.066

0.596

-

0.256

0.022

Anyone smoke inside the home 0.015

Missing

Yes

No

0.138 (0.140)

0.100 (0.040)

0.000

0.331

0.015

-

Log cotinine concentration 

(ng/dL) 0.004 Intercept for missingSlope

–0.150 (0.063)

0.039 (0.012)

0.023

0.002

Window, cabinet, or wall 

renovation in a pre-1978 home 0.045

Missing

Yes

No

–0.008 (0.061)

0.097 (0.047)

0.000

0.896

0.045

-

a

n = 2,155 (age 10-60 mo); R2 = 40% b

Includes the 50 states and the District of Columbia

Source: Dixon et al. (2009).
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1 Mielke et al. (2007) analyzed data on blood Pb and soil Pb concentration collected as part of a 

2 universal blood Pb screening program in New Orleans (2000-2005). The data set included 55,551 blood 

3 Pb measurements for children 0-6 years of age and 5,467 soil Pb measurements. Blood Pb and soil Pb 

4 concentrations were matched at the level of census tracts. The association between blood Pb 

5 concentration and soil Pb concentration was evaluated using non-parametric permutation methods. The 

resulting best-fit model (R2 6 =0.528) was:

Equation 4-4

7 where PbB is the median blood Pb concentration and PbS is the median soil Pb concentration. The 

8 resulting curvilinear relationship predicts a twofold increase in blood Pb concentration for an increase in 

9 soil Pb concentration from 100 to 1000 ppm (Figure 4-21).

Figure 4-21. Predicted relationship between soil Pb concentration and 

blood Pb concentration in children based data collected in the 

New Orleans child Pb screening program (2000-2005) (Mielke 

et al., 2007). The data set included 55,551 blood Pb 

measurements for children 0-6 years of age and 5,467 soil Pb 

measurements. Blood Pb and soil Pb concentrations were 

matched at the level of census tracts. 

10 In a subsequent re-analysis of the New Orleans (2000-2005) data, individual child blood Pb 

11 observations were matched to census tract soil concentrations (Zahran et al., 2011). This analysis 

12 confirmed the association between blood Pb and both soil Pb and age reported in Mielke et al. (2007). 

0

2

4

6

8

10

0 400 800 1200 1600 2000

Blood Pb (μg/dL)

Soil Pb (ppm)

Blood=2.038+0.172 x Soil0.5
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1 Regression coefficients for soil Pb (random effects generalized least squares regression) ranged from 

0.217 to 0.214 (per soil Pb0.5 2 ), which is equivalent to approximately a 2-fold increase in blood Pb 

3 concentration for an increase in soil Pb concentration from 100 to 1000 ppm.

4 Several studies have linked elevated blood Pb levels to residential soil exposures for populations 

5 living nearby industrial or mining facilities. Gulson et al. (2009) studied the blood Pb and isotopic Pb 

6 ratios of children younger than 5 years old and adults older than 18 years old living in the vicinity of a 

7 mine producing Magellan Pb ore in western Australia. They observed a median blood Pb level of 6.6 

8 μg/dL for the children, with isotopic ratios indicating contributions from the mine ranging from 27 to 

9 93%. A weak but significant linear association between blood Pb level and percent Magellan Pb was 

observed (R2 10 = 0.12, p = 0.018). Among children with blood Pb levels over 9 μg/dL and among adults, the 

11 isotopic ratios revealed Pb exposures from a variety of sources. Garavan et al. (2008) measured soil Pb 

12 and blood Pb levels among children aged 1 month to 17.7 years old in an Irish town near a coal mine. The 

13 blood Pb measurements were instituted as part of a screening and community education program given 

14 that the presence of Pb had been documented in the environment. Garavan et al. (2008) found that over 3 

15 years of the screening period, median blood Pb levels reduced by roughly 22% from 2.7 to 2.1 μg/dL. 

16 An extensive discussion of the relationships between environmental Pb levels and blood Pb 

17 concentrations in children at the Bunker Hill Superfund Site, a former Pb mining and smelting site, was 

18 provided in the 2006 AQCD. In the most recent analysis (TerraGraphics Environmental Engineering, 

19 2004) of the data on environmental Pb levels and child blood Pb concentrations (1988-2002), blood Pb 

20 concentrations (annual GM) ranged from 2.6 to 9.9 μg/dL. Environmental Pb levels (e.g., dust, soil, paint 

21 Pb levels) data were collected at ~3,000 residences, with interior dust Pb concentrations (annual GM) 

22 ranging from ~400 to 4,200 mg/kg and yard soil Pb concentration (annual GM) ranging from ~150 to 

23 2,300 mg/kg. Several multivariate regression models relating environmental Pb levels and blood Pb 

concentration were explored; the model having the highest R2 24 (0.26) is shown in Table 4-17. The model 

25 predicts significant associations between blood Pb concentration, age, interior dust, yard soil, 

26 neighborhood soil (geometric mean soil Pb concentration for areas within 200 ft of the residence), and 

27 community soil Pb concentration (community GM). Based on the standardized regression coefficients, the 

28 community soil Pb concentration had the largest effect on blood Pb concentration, followed by 

29 neighborhood soil Pb concentration, interior dust Pb concentration, and yard soil Pb concentration (Table 

30 4-17). The model predicted a 1.8 μg/dL decrease in blood Pb concentration in association with a decrease 

31 in community soil Pb concentration from 2,000 to 1,000 mg/kg. The same decrease in neighborhood soil 

32 Pb concentration, interior dust Pb concentration, or yard soil Pb concentration was predicted to result in a 

33 0.8, 0.5, or 0.2 μg/dL decrease in blood Pb concentration, respectively.
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Table 4-17. General linear model relating blood Pb concentration in children and 

environmental Pb levels—Bunker Hill Superfund Site

Parameter Coefficient P-value Standardized

Coefficient

Intercept -0.1801 0.7916 0.00000

Age (yr) -0.4075 <0.0001 -0.2497

ln(interior dust Pb) (mg/kg) 0.7288 <0.0001 0.1515

ln(yard soil Pb) (mg/kg) 0.2555 0.0002 0.0777

GM soil Pb within 200 ft of residence (mg/kg) 0.0008 <0.0001 0.1380

GM community soil Pb (mg/kg) 0.0018 <0.0001 0.2250

R2 = 0.264; p <0.0001; based on data from Bunker Hill Superfund Site collected over the period 1988-2002.

GM: geometric mean; ln: natural log.

Source: TerraGraphics (2004).

1 Malcoe et al. (2002) analyzed 1997 data on blood Pb and environmental Pb concentrations in a 

2 representative sample of Native American and white children (n = 224, age 1-6 years) who resided in a 

3 former Pb mining region in Ottawa County, OK. The data set included measurements of blood Pb, yard 

4 soil Pb, residential interior dust Pb loading, first-draw water Pb, paint Pb assessment and other behavioral 

5 (i.e., hand-to-mouth activity, hygiene rating) and demographic variables (i.e., hand-to-mouth activity, 

6 hygiene rating, poverty level, caregiver education). A multivariate regression model accounted for 34% of 

7 the observed variability in blood Pb. Yard soil Pb and interior dust Pb loading accounted for 10% and 3% 

8 of the blood Pb variability, respectfully. The regression model predicted a slope of 0.74 μg/dL blood Pb 

per ln[μg/g soil Pb] and a slope of 0.45 μg/dL blood Pb per ln[μg/ft2 9 ] dust Pb loading.

4.6. Biokinetic Models of Lead Exposure-Blood 

Lead Relationships

10 An alternative to regression models are mechanistic models, which attempt to specify all 

11 parameters needed to describe the mechanisms (or processes) of transfer of Pb from the environment to 

12 human tissues. Such mechanistic models are more complex than regression models; this added 

13 complexity introduces challenges in terms of their mathematical solution and empirical verification. 

14 However, by incorporating parameters that can be expected to vary spatially or temporally, or across 

15 individuals or populations, mechanistic models can be extrapolated to a wide range of exposure scenarios, 

16 including those that may be outside of the domain of paired predictor-outcome data used to develop the 

17 model. Exposure-intake models, a type of mechanistic models, are highly simplified mathematical 

18 representations of relationships between levels of Pb in environmental media and human Pb intakes (e.g., 

19 μg Pb ingested per day). These models include parameters representing processes of Pb transfer between 

20 environmental media (e.g., air to surface dust) and to humans, including rates of human contact with the 

21 media and intakes of the media (e.g., g soil ingested per day). Intake-biokinetic models provide the 
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1 analogous mathematical representation of relationships between Pb intakes and Pb levels in body tissues

2 (e.g., blood Pb concentration). They include parameters that represent processes of Pb transfer (a) from 

3 portals of entry into the body and (b) from blood to tissues and excreta. Linked together, exposure-intake 

4 and intake-biokinetics models (i.e., integrated exposure-intake-biokinetics models) provide an approach 

5 for predicting blood Pb concentrations (or Pb concentrations in other tissues) that corresponds to a 

6 specified exposure (medium, concentration, and duration). Detailed information on exposure and internal 

7 dose can be obtained from controlled experiments, but almost never from epidemiological observations or 

8 from public health monitoring programs. Exposure intake-biokinetics models can provide these 

9 predictions in the absence of complete information on the exposure history and blood Pb concentrations 

10 for an individual (or population) of interest. Therefore, these models are critical to applying 

11 epidemiologic-based information on blood Pb-response relationships to the quantification and 

12 characterization of human health risk. They are also critical for assessing the potential impacts of public 

13 health programs directed at mitigation of Pb exposure or of remediation of contaminated sites. 

14 However, they are not without their limitations. Human exposure-biokinetics models include large 

15 numbers of parameters, which are required to describe the many processes that contribute to Pb intake, 

16 absorption, distribution, and elimination. The large number of parameters complicates the assessment of 

17 confidence in parameter values, many of which cannot be directly measured. Statistical procedures can be 

18 used to evaluate the degree to which model outputs conform to “real-world” observations and values of 

19 influential parameters can be statistically estimated to achieve good agreement with observations. Still, 

20 large uncertainty can be expected to remain about many, or even most, parameters in complex exposure21 biokinetic models. Such uncertainties need to be identified and their impacts on model predictions 

22 quantified (i.e., sensitivity analysis or probabilistic methods).

23 Modeling of human Pb exposures and biokinetics has advanced considerably during the past 

24 several decades, although there have been relatively few developments since the 2006 Pb AQCD was 

25 published. Still in use is the Integrated Exposure Uptake Biokinetic (IEUBK) Model for Lead in Children 

26 (U.S. EPA, 1994) and models that simulate Pb biokinetics in humans from birth through adulthood 

27 (Leggett, 1993; O'Flaherty, 1993, 1995). The EPA AALM is still in development. A complete and 

28 extensive discussion of these models can be found in the 2006 Pb AQCD (U.S. EPA, 2006).

4.7. Summary

4.7.1. Exposure

29 Exposure data considered in this assessment build upon the conclusions of the 2006 AQCD for Pb 

30 (2006), which found air Pb concentrations in the U.S. and associated biomarkers of exposure to Pb have 

31 decreased substantially following the ban on Pb in gasoline, house-hold paints, and solder. Pb exposure is 
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1 difficult to assess because Pb has multiple sources in the environment and passes through various media.

2 Air-related pathways of Pb exposure are the focus of this assessment. Pb can be emitted to air, water, or 

3 soil. In addition to primary emission of particle-bound or gaseous Pb to the atmosphere, Pb can be 

4 suspended to the air from disturbance of soil or dust, and a fraction of that suspended Pb may even 

5 originate from waters used to irrigate the soil. Pb-bearing PM can be deposited from the air to soil or 

6 water through wet and dry deposition. In general, air-related pathways include those pathways where Pb 

7 passes through ambient air on its path from a source to human exposure. Air-related Pb exposures include

8 inhalation and ingestion of Pb-contaminated food, water or other materials including dust and soil. Non9 air-related Pb exposures may include ingestion of indoor Pb paint, Pb in diet as a result of inadvertent 

10 additions during food processing, and Pb in drinking water attributable to Pb in distribution systems, as 

11 well as other generally less prevalent pathways. Pb can cycle through multiple media prior to human 

12 exposure. Given the multitude of possible air-related exposure scenarios and the related difficulty of 

13 constructing Pb exposure histories, most studies of Pb exposure through air, water, and soil can be 

14 informative to this review. Other exposures, such as occupational exposures, contact with consumer goods 

15 in which Pb has been used, or ingestion of Pb in drinking water conveyed through Pb pipes may also 

16 contribute to Pb body burden.

17 Section 4.1 presents data illustrating potential exposure mechanisms. Several studies suggested that 

18 soil can act as a reservoir for Pb emissions from industrial or and other activities. Exposure to soil 

19 contaminated with deposited Pb can occur through hand-to-mouth contact as well as inhalation of 

20 resuspended Pb-bearing PM. In general, soil Pb concentrations tended to be higher within inner-city 

21 communities compared with neighborhoods surrounding city outskirts. Infiltration of Pb dust has been 

22 demonstrated, and Pb dust has been shown to persist in indoor environments even after repeated 

23 cleanings. Measurements of particle-bound Pb exposures reported in this assessment have shown that 

24 personal exposure concentrations for Pb are typically higher than indoor or outdoor ambient Pb 

25 concentrations. These findings may be related to local resuspension with body movement.

4.7.2. Kinetics

26 The majority of Pb in the body is found in bone (roughly 90% in adults, 70% in children); only 

27 about 1% of Pb is found in the blood. Pb in blood is primarily (~99%) bound to RBCs. It has been 

28 suggested that the small fraction of Pb in plasma (<1%) may be the more biologically labile and 

29 toxicologically active fraction of the circulating Pb. Saturable binding to RBC proteins contributes to an 

30 increase in the plasma/blood Pb ratio with increasing blood Pb concentration and curvature to the blood 

31 Pb–plasma Pb relationship. As blood Pb increases and the higher affinity binding sites for Pb in RBCs 

32 become saturated at approximately 40 μg/dL blood, a larger fraction of the blood Pb is available in 

33 plasma to distribute to brain and other Pb-responsive tissues.
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1 The burden of Pb in the body may be viewed as divided between a dominant slow compartment 

2 (bone) and a smaller fast compartment (soft tissues). Pb uptake and elimination in soft tissues is much 

3 faster than in bone. Pb accumulates in bone regions undergoing the most active calcification at the time of 

4 exposure. During infancy and childhood, bone calcification is most active in trabecular bone (e.g. patella); 

5 whereas, in adulthood, calcification occurs at sites of remodeling in cortical (e.g. tibia) and trabecular 

6 bone (Aufderheide & Wittmers, 1992). A high bone formation rate in early childhood results in the rapid 

7 uptake of circulating Pb into mineralizing bone; however, bone Pb is also recycled to other tissue 

8 compartments or excreted in accordance with a high bone resorption rate (O'Flaherty, 1995). Thus, most 

9 of the Pb acquired early in life is not permanently fixed in the bone.

10 The exchange of Pb from plasma to the bone surface is a relatively rapid process. Pb in bone 

11 becomes distributed in trabecular and the more dense cortical bone. The proportion of cortical to 

12 trabecular bone in the human body varies by age, but on average is about 80 to 20. Of the bone types, 

13 trabecular bone is more reflective of recent exposures than is cortical bone due to the slow turnover rate 

14 and lower blood perfusion of cortical bone. Some Pb diffuses to deeper bone regions where it is relatively 

15 inert, particularly in adults. These bone compartments are much more labile in infants and children than in 

16 adults as reflected by half-times for movement to the plasma (e.g., cortical t1/2 = 0.23 years at birth, 3.7 

17 years at 15 years of age, and 23 years in adults; trabecular t1/2 = 0.23 years at birth, 2.0 years at 15 years of 

18 age, and 3.8 years in adults) (Leggett, 1993). Due to the more rapid turnover of bone mineral in children, 

19 changes in blood Pb concentration are thought to more closely parallel changes in total body burden. 

20 However, some Pb accumulated in bone during childhood does persist into later life. Potential 

21 mobilization of Pb from the skeleton could occur in adults at times of physiological stress associated with 

22 enhanced bone remodeling such as during pregnancy and lactation, menopause or in the elderly, extended 

23 bed rest, hyperparathyroidism, and weightlessness. Regardless of age, however, similar blood Pb 

24 concentrations in two individuals (or populations) do not necessarily translate to similar body burdens or 

25 similar exposure histories.

26 The kinetics of elimination of Pb from the body reflects the existence of fast and slow pools of Pb 

27 in the body. The dominant phase of Pb kinetics in the blood, exhibited shortly after a change in exposure 

28 occurs, has an elimination half-life of ~20-30 days. An abrupt change in Pb uptake gives rise to a 

29 relatively rapid change in blood Pb, to a new quasi-steady state, achieved in ~75-100 days (i.e., 3-4 times 

30 the blood elimination half-life). A slower phase may become evident with longer observation periods 

31 following a decrease in exposure due to the gradual redistribution of Pb among other compartments via 

32 the blood. Therefore, a single blood Pb concentration may reflect the near-term or longer-term history of 

33 the individual to varying degrees, depending on the relative contributions of internal (e.g., bone) and 

34 external sources of Pb to blood Pb, which in turn will depend on the exposure history and possibly age35 related and individual-specific (e.g., pregnancy, lactation) characteristics of bone turnover. In general, 
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1 higher blood Pb concentrations can be interpreted as indicating higher exposures (or Pb uptakes); 

2 however, they do not necessarily predict higher body burdens, especially in adults. 

4.7.3. Lead Biomarkers

3 Observational studies using biomarkers of Pb are included in Section 4.4. The median blood Pb 

4 level for the entire U.S. population is 1.2 μg/dL and the 95th percentile blood Pb level was 3.7 μg/dL, 

5 based on the 2007-2008 NHANES data (NCHS, 2010). Among children aged 1-5 years, the median and 

6 95th percentiles were slightly higher at 1.4 μg/dL and 4.1 μg/dL, respectively. Overall, trends in blood Pb 

7 levels have been decreasing among U.S. children and adults over the past twenty years. Concurrent 

8 changes in isotopic ratios of blood Pb samples reflect changes in source composition over the past several 

9 decades. Recent studies have observed a relationship between blood Pb and soil Pb concentration.

10 Additionally, studies have shown that blood Pb is also associated with non-air-related, non-policy11 relevant exposure to Pb paints in older homes, Pb released into drinking water, and occupational work 

12 with materials containing Pb.

13 Blood Pb is dependent on both the recent exposure history of the individual, as well as the long14 term exposure history that determines body burden and Pb in bone. The contribution of bone Pb to blood 

15 Pb changes depending on the duration and intensity of the exposure, age, and various other physiological 

16 variables that may affect bone remodeling (e.g., nutritional status, pregnancy, menopause). Blood Pb in 

17 adults is typically more an index of recent exposures than body burden, whereas bone Pb is an index of 

18 cumulative exposure and body burden. In children, due to faster exchange of Pb to and from bone, blood 

19 Pb is both an index of recent exposure and potentially an index of body burden. In some physiological 

20 circumstances (e.g., osteoporosis), bone Pb may contribute to blood Pb in adults. The disparity between 

21 blood Pb and body burden may have important implications for the interpretation of blood Pb 

22 measurements in some epidemiology studies. Conceptually, measurement of long-term Pb body burden 

23 (i.e., based on tibia Pb) may be the appropriate metric if the effects of Pb on a particular outcome are 

24 lasting and cumulative. However, if the effects of Pb on the outcome represent the acute effects of current 

25 exposure, then blood Pb may be the preferred metric. In the absence of clear evidence as to whether a 

26 particular outcome is an acute effect of recent Pb dose or a chronic effect of cumulative Pb exposure, both 

27 blood and bone metrics should be considered. 

28 Cross-sectional studies that sample blood Pb once generally provide an index of recent exposures. 

29 In contrast, cross-sectional studies of bone Pb and longitudinal samples of blood Pb concentrations over 

30 time provide an index of cumulative exposure and are more reflective of average Pb body burdens over 

31 time. The degree to which repeated sampling will reflect the actual long-term time-weighted average 

32 blood Pb concentration depends on the sampling frequency in relation to variability in exposure. High 
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1 variability in Pb exposures can produce episodic (or periodic) oscillations in blood Pb concentration that 

2 may not be captured with low sampling frequencies.

3 The concentration of Pb in urine is a function of the urinary Pb excretion and the urine flow rate. 

4 Urine flow rate requires collection of a timed urine sample, which is often problematic in epidemiologic 

5 studies. Collection of un-timed (“spot”) urine samples, a common alternative to timed samples, requires 

6 adjustment of the Pb measurement in urine to account for variation in urine flow (Diamond, 1988). 

7 Urinary Pb concentration reflects, mainly, the exposure history of the previous few months; thus, a single 

8 urinary Pb measurement cannot distinguish between a long-term low level of exposure or a higher acute 

9 exposure. Thus, a single urine Pb measurement, or a series of measurements taken over short-time span, is 

10 likely a relatively poor index of Pb body burden for the same reasons that blood Pb is not a good indicator 

11 of body burden. On the other hand, long-term average measurements of urinary Pb can be expected to 

12 better reflect body burden.

4.7.4. Air Lead-Blood Lead Relationships

13 The 1986 Pb AQCD described epidemiological studies of relationships between air Pb and blood 

14 Pb. Much of the pertinent earlier literature described in the AQCD was drawn from a meta-analysis by 

15 Brunekreef (1984). In addition to the meta-analysis of Brunekreef (1984), seven more recent studies have 

16 provided data from which estimates of the blood Pb-air Pb slope can be derived for children (Table 4-11). 

The range of estimates from these seven studies is 1-9 μg/dL per μg/m3 17 , which encompasses the estimate 

from the Brunekreef (1984) meta-analysis of (3-6 μg/dL per μg/m3 18 ). The Schnaas et al. (2004) had a 

19 particularly strong experimental design in that is the only longitudinal study in which blood Pb 

20 concentration was monitored repeatedly in individual children from age 6 months to 10 years. For 

children who experienced the largest declines in air Pb (i.e., from 2.8 to <0.1 μg/m3 21 ), the predicted blood 

22 Pb-air Pb slope (adjusted for age, year of birth, SES, and use of glazed pottery) was 0.213 ln[μg/dL 

blood] per ln[μg/m3 23 air]. The cross-sectional study done by Ranft et al. (2008) attempted to account for 

24 potential co-variates that influence blood Pb (e.g., soil Pb concentration, gender, environmental tobacco 

25 smoke, fossil heating system and parental education). It is the only study that reported a logarithmic blood 

26 Pb-linear air Pb relationship, which results in an upward curvature of the blood Pb-air Pb relationship 

27 (i.e., the blood Pb-air Pb slope increases with increasing air Pb concentration). In other studies (or based 

28 on other studies), the blood Pb-air Pb relationship was either log-log (Brunekreef, 1984; Hayes et al., 

29 1994; Schnaas et al., 2004), which predicts an increase in the blood Pb-air Pb slope with decreasing air Pb 

30 concentration or linear (Hilts, 2003; J. Schwartz & Pitcher, 1989; Tripathi et al., 2001), which predicts a 

31 constant blood Pb-air Pb slope across all air Pb concentrations. These differences may simply reflect 

32 model selection by the investigators; alternative models are not reported in these studies. Because air Pb 
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1 contributes to Pb in soil and indoor dusts, adjustment for the correlated co-variates such as soil Pb would 

2 introduce a downward bias in the slope estimate.
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Chapter 4. References

Abadin, H. G., & Wheeler, J. S. (1997). Guidance for risk assessment of exposure to lead: A site-specific, 

multi-media approach. In J. S. Andrews, H. Frumkin, B. L. Johnson, M. A. Mehlman, C. Xintaras 

& J. A. Bucsela (Eds.), Hazardous waste and public health: International congress on the health 

effects of hazardous waste (pp. 477-485). Princeton, NJ: Princeton Scientific Publishing.

Abrahams, P. W., Follansbee, M. H., Hunt, A., Smith, B., & Wragg, J. (2006). Iron nutrition and possible 

lead toxicity: An appraisal of geophagy undertaken by pregnant women of UK Asian 

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