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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 

2 exposure to Pb through various pathways. Pb is considered to be a multimedia 

3 contaminant with multiple pathways of exposure. The relative importance of various 

4 media in affecting Pb exposure changes with source strength and location, location and 

5 time activity of the exposed individuals, behavior of the exposed individuals, and risk 

6 factors such as age and socioeconomic factors (risk factors are discussed in detail in 

7 Chapter 6). Blood Pb and bone Pb biomarkers (discussed in Section 4.3, Section 4.4, 

8 Section 4.5, and Section 4.6), are often used to indicate composite Pb exposure resulting 

9 from multiple media and pathways of exposure. 

10 The recent information provided here builds upon the conclusions of the 2006 Pb AQCD

11 (2006b), which found that air Pb concentrations and blood Pb levels have decreased 

12 substantially following the restrictions on Pb in on-road vehicle gasoline, Pb in household 

13 paints, the use of Pb solder, and reductions in industrial Pb emissions that have occurred 

14 since the late 1970s. Nevertheless, detectable quantities of Pb have still been observed to 

15 be bioaccessible in various media types. It was reported in the 2006 Pb AQCD (U.S. 

16 EPA, 2006b) that airborne maximum quarterly Pb concentrations in the U.S. were in the 

range of 0.03-0.05 μg/m3 17 for non-source-oriented monitors for the years 2000-2004 and 

were 0.10-0.22 μg/m3 18 for source-oriented monitors during that time period, while blood 

19 Pb levels reached a median of 1.70 μg/dL among children (1-5 years of age) in 

20 2001-2002. It was also observed that Pb exposures were associated with nearby industrial 

21 Pb sources, presence of Pb-based paint, and Pb deposited onto food in several of the 

22 studies described in the 2006 Pb AQCD.

4.1.1 Pathways for Lead Exposure 

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

24 environment and passes through various environmental media. These issues are described 

25 in detail in Sections 3.2 and 3.3. Air-related pathways of Pb exposure are the focus of this 

26 ISA. Pb can be emitted to air, soil, or water and then cycle through any or all of these 

27 media. In addition to primary emission of particle-bound or gaseous Pb to the 

28 atmosphere, Pb can be resuspended to the air from soil or dust. Additionally, Pb-bearing 

29 PM can be deposited from the air to soil or water through wet and dry deposition. Air-
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1 related Pb exposures also include inhalation and ingestion of Pb-contaminated food, 

2 water or other materials including dust and soil via hand-to-mouth contact. In general, 

3 air-related pathways include those pathways where Pb passes through ambient air on its 

4 path from a source to human exposure. Some non-air-related exposures of Pb include 

5 ingestion of indoor Pb paint, Pb in diet as a result of inadvertent additions during food 

6 processing, and Pb in drinking water attributable to Pb in distribution systems, as well as 

7 other generally less prevalent pathways. 

8 Particle size of Pb-PM is relevant to transport through various media leading to exposure. 

9 The inhalability of airborne particles is a gradually decreasing function of particle size. 

10 Inhalability criteria established from experimental data obtained at wind speeds of 1-8

11 m/s describe PM inhalability of 77% for particles <10 μm (dae, aerodynamic diameter). 

12 Inhalability of particles ranging in size from 40 to 100 μm dae is 50%; above 100 μm, 

13 inhalability data are lacking (Soderholm, 1989; ACGIH, 1985). Of the particles that are 

14 not inhaled, their settling to surfaces makes them available for subsequent ingestion. The 

15 main pathway for Pb ingestion by children is by hand to mouth (Lanphear et al., 1998). In 

16 a playground environment in London, U.K., Duggan et al. (1985) reported that hand to 

17 mouth transfer was effectively limited to particles smaller than 10 μm, even when the soil 

18 itself exhibited a much larger particle size distribution. More recently, Yamamoto et al.

19 (2006) reported for a cohort of children in Kanagawa Prefecture, Japan (greater Tokyo 

20 area) that the mode of size distributions of particles adhering to children’s hands was 39 

21 ± 26 μm, with the upper tail ranging from 200-300 μm. Differences in the size 

22 distribution results may be related to differences in the soil between the two locations 

23 and/or to differences between the analytical methods used to measure size distribution; 

24 Duggan and Inskip (1985) used optical microscopy of the dust wipes, while Yamamoto et 

25 al. (2006) used a laser scattering device measuring sampled particles suspended in an 

26 aqueous solution. Similar studies focusing on particle size distributions of ingestion of 

27 house dust are lacking. Ingestion of house dust has been reported to be the major source 

28 of lead intake during early childhood (Lanphear et al., 2002). If a similar particle size 

29 distribution holds for household dust, then ingestion of indoor Pb of atmospheric origin 

30 could also be strongly dependent on dust particle size. Therefore, larger particles of 

31 atmospheric origin, which may not be considered relevant for exposure by inhalation 

32 exposure, are still relevant for Pb exposure by ingestion. However, no studies in the 

33 literature have presented information on the relative contributions of Pb from different 

34 PM size fractions to blood Pb concentrations. 

35 The complicated nature of Pb exposure is illustrated Figure 4-1, in which the Venn 

36 diagram depicts how Pb can cycle through multiple environmental media prior to human 

37 exposure. The “air/soil/water” arrows illustrate Pb exposures to plants, animals, and/or 

38 humans via contact with Pb-containing media. The exposures are air-related if the Pb 
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1 passed through the air compartment. When animals consume plant material exposed to 

2 Pb that has at some point passed through the air compartment, and when human diet 

3 includes animals and/or plants exposed to Pb that has passed through the air 

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

5 multitude of possible air-related exposure scenarios and the related difficulty of 

6 constructing Pb exposure histories, most studies of Pb exposure through air, water, and 

7 soil can be informative to this review. Figure 4-1 also illustrates other exposures, such as 

8 occupational exposures, contact with consumer goods in which Pb has been used, or 

9 ingestion of Pb in drinking water conveyed through Pb pipes. Most Pb biomarker studies 

10 do not indicate speciation or isotopic signature, and so exposures that are not related to 

11 Pb in ambient air are also reviewed in this section because they can contribute to Pb body 

12 burden. Many of the studies presented in the subsequent material focus on observations 

13 of Pb exposure via one medium: air, water, soil and dust, diet, or occupation.

Note: The Venn diagram is used to illustrate the passage of Pb through multiple environmental media compartments through which 

exposure can occur.

Figure 4-1 Conceptual model of multimedia Pb exposure.

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

15 the environment is often difficult to discern. Individual factors such as home 
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1 environment, location, and risk factors (described in more detail in Chapter 6) may 

2 influence exposures. The National Human Exposure Assessment Survey (NHEXAS) 

3 study sampled Pb, as well as other pollutants and VOCs, in multiple exposure media from 

4 subjects across six states in EPA Region 5 (Illinois, Indiana, Michigan, Minnesota, Ohio, 

5 and Wisconsin) (Clayton et al., 1999) as well as in Arizona (O'Rourke et al., 1999) and 

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

7 concentrations of Pb are higher than indoor or outdoor concentrations of Pb (Table 4-1). 

8 Pb levels in windowsill dust were higher than Pb levels in surface dust collected from 

9 other surfaces. Clayton et al. (1999) suggested that higher windowsill levels could be 

10 attributed to the presence of Pb-based paint and/or to accumulation of infiltrated outdoor 

11 Pb-bearing PM. Pb levels in food were higher than in beverages, and Pb levels in 

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

13 levels 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: Reprinted with permission of Nature Publishing Group, Clayton et al. (1999)

1 Several studies have used a combination of measured values and default model values to 

2 represent exposures and determine their relative contributions to blood Pb. Cornelis et al.

3 (2006) used the Integrated Exposure Uptake Biokinetic model (IEUBK), described in 

4 detail in the 2006 Pb AQCD (U.S. EPA, 2006b) to model children’s exposures to Pb 

5 emissions from a non-ferrous smelter in Hoboken, Belgium. They point out that ambient 

6 air Pb concentrations decrease with distance from the smelter while IEUBK is fairly 

7 insensitive to changes in ambient air concentration. They input the average ambient air 

8 Pb concentration of 0.81μg/m3 and assumed that the indoor air Pb concentration was 

9 30% of the outdoor Pb. In the absence of indoor air Pb samples, the default assumption 

10 was adopted for the Cornelis et al. (2006) study. Similarly, Carrizales et al. (2006)

11 analyzed exposures to children living near a copper smelter in San Luis Potosi, Mexico. 

12 They employed the IEUBK default options for assignment of Pb dust concentration as 
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1 70% of the soil Pb concentration, while air Pb concentration was assigned based on 

2 measurements by the Mexican government. Based on these assumptions, they attributed 

3 87% of blood Pb to soil and dust exposure. Input soil and dust Pb levels were obtained 

4 from an average of 28 sites. Using defaults for all IEUBK values, the soil/dust Pb 

5 ingestion pathway has been predicted to contribute 63-75% (depending of a child’s age) 

6 of blood Pb, whereas the air Pb inhalation pathway contributed <1% (SRC, 2007). This 

7 analysis did not specifically estimated the portion of the soil/dust Pb ingestion pathway 

8 that derives from air Pb, such as recently airborne Pb deposited to soil and dust which 

9 remains available for inhalation and ingestion. In the 1986 Pb AQCD, blood Pb-air Pb 

10 slopes were examined that included both direct (inhalation) and indirect (via soil, dust, 

11 etc.) air Pb contributions, resulting in the conclusion that slopes for the total air Pb 

12 contribution were roughly twice that of the slope due to inhaled air Pb alone. This “recent 

13 air” Pb was described in the 2007 Pb Risk Assessment (U.S. EPA, 2007f) to include 

14 those pathways involving Pb that is or has recently been in the outdoor ambient air, 

15 including inhalation and ingestion of indoor dust Pb derived from recent ambient air 

16 (e.g., air Pb that has penetrated into the residence recently and loaded indoor dust). 

17 Beyond the direct inhalation pathway (outdoor and indoor ambient air Pb), Appendix I of 

18 the 2007 Pb Risk Assessment (U.S. EPA, 2007f) provides estimates of the contribution of 

19 various other pathways (e.g., diet, drinking water, outdoor soil ingestion) as well as those 

20 involving recent air Pb to blood Pb. Estimates of recent air Pb contributions were highly 

21 dependent on the exposure scenario evaluated. For the General Urban Case Study in that 

22 assessment, “current conditions” at the time of the risk assessment (0.14 μg/m3, mean of 

23 maximum quarterly average concentration of Pb in TSP) were modeled as remaining 

24 constant throughout the 7 years of the exposure modeled in the biokinetic model. A 

25 median concurrent blood Pb concentration of 1.8-2.8 μg/dL (depending on the exposure 

26 and risk model) was estimated as the average of the results at 75 and 81 months of age. 

27 The recent air Pb ingestion pathways contributed a median of 12-28% (depending on the 

28 exposure model) to blood Pb beyond the 0.5-0.6% attributable to inhalation of recent 

29 airborne Pb. In simulations for the 95th percentile of current conditions (0.87 μg/m3, 

30 95th percentile of maximum quarterly average concentration of Pb in TSP), median 

31 children’s blood Pb increased to 2.0-3.1 μg/dL and the contribution of recent air Pb 

32 ingestion pathways increased to 22-37% of blood Pb. Results of other exposure scenarios 

33 for which greater or lesser contributions of recent air Pb to blood Pb were predicted are 

34 available in Appendix I of the 2007 Pb Risk Assessment (U.S. EPA, 2007f).
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4.1.2 Environmental Exposure Assessment Methodologies

1 A number of monitoring and modeling techniques have been employed for exposure 

2 assessment. These are detailed in either the 2006 Pb AQCD (U.S. EPA, 2006b) or in the 

3 subsequent Risk and Exposure Assessment performed as part of the same NAAQS 

4 review (U.S. EPA, 2007g). Some of these methods are briefly described here to provide a 

5 context for the exposure studies described in Section 4.1.3. Blood Pb sampling is 

6 described in detail in Section 4.3.2.

7 Data collection to assess Pb exposure pathways may involve air, soil, and dust samples. 

8 Methods used for digesting air Pb samples are described in Section 3.4, as are ambient air 

9 Pb monitoring techniques. Factors affecting collection of ambient air Pb samples are 

10 described in detail in Section 3.4. For the monitors in the FRM network, the primary role 

11 is compliance assessment. Accordingly, this network includes monitors in locations near 

12 sources of air Pb emissions which are expected to or have been shown to contribute to 

13 ambient air Pb concentrations in excess of the Pb NAAQS. In such locations, Pb may be 

14 associated with relatively larger size particles, contributing to air Pb concentration 

15 gradients with distance from the source and greater deposition in the near-source 

16 locations. The FRM network also includes non-source-oriented monitors for which the 

17 main objective is to gather information on neighborhood-scale lead concentrations that 

18 are typical in urban areas so to better understand ambient air-related Pb exposures for 

19 populations in these areas. This part of the Pb NAAQS network, is required to be 

20 operational as of December 27, 2011. These monitor locations are distributed across a 

21 broad geographic area, representing approximately 63 large urban areas which contain 

22 approximately half of the total U.S. population (based on recently published 2010 Census 

23 Bureau data). In lieu of more detailed analysis of population proximity for these newly 

24 established monitors, population counts were calculated near previously existing 

25 monitors for which data are presented in Section 3.5. For the monitors in that limited 

26 dataset, among the total population of 311,127,619 people in the 2010 Census (ESRI, 

27 2011), 181,100 (0.06%) lived within 1 km of a source-oriented monitor, while 918,351 

28 (0.30%) lived within 1 km of a non-source-oriented monitor. 

29 Dust sampling has not changed drastically since it was first proposed by Sayre et al.

30 (1974), in which a disposable paper towel was soaked in 20% denatured alcohol and 

1:750 benzalkonium chloride and then used to wipe a 1 ft2 31 sampling area in a systematic 

32 fashion. Que Hee et al. (1985) and Sterling et al. (1999) compared wipe testing with 

33 vacuum methods. Sampling efficiency for the first attempt varied between 53-76% with 

34 vacuum pump flow rate and tube type and was 52% for the wipe method for the Que Hee 

35 et al. (1985) study, with 100% efficiency after five consecutive samples were obtained. 

36 Sterling et al. (1999) observed that two of three vacuuming methods had significantly 
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higher geometric mean collection (vacuum 1: 94.3 μg/ft2; vacuum 2: 23.5 μg/ft2 1 ) 

compared with dust wipes (5.6 μg/ft2 2 ).

3 Models may also be used in exposure assessment. For example, two dispersion models, 

4 the American Meteorological Society/Environmental Protection Agency Regulatory 

5 Model (AERMOD), and Industrial Source Complex-Plume Rise Model Enhancements 

6 (ISC-PRIME) were employed to model dispersion of Pb emissions from specific 

7 industrial facilities (Cimorelli et al., 2005; Perry et al., 2005; EPRI, 1997), and to

8 estimate ambient air Pb concentrations at some of the case studies included in the 2007 

9 Risk and Exposure Assessment (U.S. EPA, 2007g). These models assume plume 

10 dispersion follows a Gaussian distribution from a point source. For the two point source 

11 case studies included in the 2007 risk assessment, the plume models were used to track 

12 emissions to ambient air near homes located within a few miles of emitting facilities. 

13 However, dispersion models can also be used to track long distance transport of Pb 

14 emissions, as performed by Krell and Roeckner (1988) to model the dispersion and 

15 deposition of Pb and Cd from European nations into the North Sea.

16 Several models estimate blood Pb levels resulting from estimated exposure to Pb in 

17 environmental media. These models, which are described in detail in the 2006 Pb AQCD

18 (U.S. EPA, 2006b) include the IEUBK model, and the EPA All Ages Lead Model 

19 (AALM), which combines and expands the thorough exposure and absorption modules of 

20 the IEUBK model with the comprehensive biokinetic model of Leggett (1993).

21 The Stochastic Human Exposure and Dose (SHEDS) and NORMTOX models also are 

22 capable of modeling metals exposures through various routes including inhalation, 

23 ingestion, and dermal exposure (Loos et al., 2010; Burke et al., 2002). Pb exposure 

24 modeling can also be accomplished using the Modeling Environment for Total Risk 

25 (MENTOR) framework, in which airborne Pb levels could be modeled using AQS, 

26 dispersion modeling, or chemical transport modeling, while human exposure is modeled 

27 with SHEDS or a similar exposure model (Georgopoulos and Lioy, 2006). Additionally, 

28 housing data and time-activity data from the Consolidated Human Activity Database 

29 (CHAD) are incorporated into MENTOR to develop refined estimates of Pb exposure and 

30 tissue burden. However, a literature search did not produce any Pb exposure studies using 

31 the SHEDS, NORMTOX, or MENTOR modeling systems. In general, these models take 

32 input for several environmental Pb exposure media including soil, dust, food and water, 

33 outdoor air, and indoor air. The models are designed to evaluate different exposure 

34 scenarios based on specification of particular conditions. 
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4.1.3 Exposure Studies

4.1.3.1 Airborne Lead Exposure

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

2 Pb AQCD (U.S. EPA, 2006b). As described above, the NHEXAS study showed personal 

3 air Pb concentrations to be significantly higher than indoor or outdoor air Pb 

4 concentrations (Clayton et al., 1999). Indoor air Pb concentration was moderately 

5 correlated with floor dust and residential yard soil Pb concentration (Rabinowitz et al., 

6 1985). Egeghy et al. (2005) performed multivariate fixed effects analysis of the 

7 NHEXAS-Maryland data and found that Pb levels measured in indoor air were 

8 significantly associated with log-transformed outdoor air Pb levels, ambient temperature, 

9 number of hours in which windows were open, homes built before 1950, and frequency 

10 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: Reprinted with permission of Nature Publishing Group, Egeghy et al. (2005).

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

2 other trace metals. Adgate et al. (2007) measured the concentrations of several trace 

3 elements in personal, indoor, and outdoor air samples of PM2.5 and found that average 

4 personal Pb-PM2.5 concentration was roughly three times higher than outdoor air Pb5 PM2.5 concentration and two times higher than indoor Pb-PM2.5 concentration (Table 

6 4-3). Another study of indoor and outdoor air concentrations of Pb was carried out by 

7 Molnar et al. (2007). PM2.5 trace element concentrations were determined in homes, 

8 preschools and schools in Stockholm, Sweden. In all sampled locations, Pb-PM2.5

9 concentrations were higher in the outdoor environment than in the proximal indoor 

10 environment. The indoor/outdoor ratios for Pb-PM2.5 suggest an outdoor Pb-PM2.5 net 
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1 infiltration of ~0.6 for these buildings. Outdoor air Pb concentrations did not differ 

2 between the central and more rural locations. Indoor air Pb concentrations were higher in 

3 spring than in winter, which the authors attributed to greater resuspension of elements 

4 that had accumulated in road dust over the winter period and increased roadwear on days 

5 with dry surfaces. Pekey et al. (2010) measured indoor and outdoor trace element 

6 composition of PM2.5 and PM10 in Kocaeli, an industrial region of Turkey, and found that 

7 average airborne Pb concentrations were higher outdoors than indoors for both PM2.5 and 

8 PM10 during summer and for PM10 during winter, but that indoor Pb concentration was 

9 higher than outdoor Pb concentration for PM2.5 during winter. The indoor-to-outdoor 

10 ratio of airborne Pb varied by environment; it tended to be less than one, but the ratio 

11 varied from one microenvironment to another. In a pilot study in Windsor, Ontario, 

12 Rasmussen et al. (2007) observed that the concentration of Pb in PM2.5 from a personal 

13 exposure sample was roughly 40% higher than the concentration of Pb in outdoor PM2.5

14 and 150% higher than Pb in indoor PM2.5. The three studies that included personal 

15 samples recorded measurements that were consistently higher than indoor or outdoor 

16 levels. 

17 Domestic wood burning is a potential source of Pb compounds (Section 3.2.2.5). Molnar 

18 et al. (2005) measured trace element concentration in indoor and personal exposure PM2.5

19 samples for homes in which wood is burned and in a reference group where no wood 

20 burning occurs in the home. For both indoor and personal samples, Molnar et al. (2005)

21 observed that Pb concentrations were higher for the wood burning group, but that the 

differences were not significant (indoor concentration: 6.0 μg/m3 vs. 4.3 μg/m3 22 , p = 0.26; 

personal exposure: 4.6 μg/m3 vs. 3.0 μg/m3 23 , p = 0.06).

24 Indoor activity has been associated with resuspension of settled dust, which could cause 

25 airborne contact with particle-bound Pb. Qian et al. (2008) estimated a PM10 resuspension 

rate of 1.4x10-4 hr-1 26 for one person walking across a carpeted floor. Measurements of 

27 submicron particles illustrated a roughly two-fold increase of airborne particle 

28 concentration for particles smaller than 1.8 μm for activity vs. low activity periods, with 

29 maximum concentrations reaching 4-11 times the maximum value during low activity 

30 periods. For PM10, average concentration was 2.5 times higher than background levels 

31 during activity periods, while peak concentration was 4.5 times higher. Qian and Ferro

32 (2008) observed that resuspension rates depend on particle size, floor material, and 

33 ventilation position. Increases in walking speed and weight of the walker did not 

34 consistently produce increases in resuspension. 5-10 μm particles produced a higher 

35 resuspension rate compared with smaller particles. Newly carpeted areas produced 

36 significantly higher resuspension rates than vinyl floors. Zhang et al. (2008) modeled and 

37 conducted experiments of particle dispersion from walking and observed that human 

38 activity did affect resuspension. They found that larger particles were more readily 
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1 detached from the carpet by walking motion, but that smaller particles are more easily 

2 resuspended once detached. Hunt and Johnson (In Press) studied the duration and spatial 

3 extent of resuspension of 0.3-5.0 μm particles following walking by a soiled shoe. 

4 0.3-0.5 μm particle concentration remained increased over a time period of 23 min, while 

5 1-5 μm particles declined in concentration over the same time period. Experiments and 

6 computational fluid dynamics simulations by Eisner et al. (2010) for a mechanical foot 

7 moving on carpeting suggested that the rotating motion of the moving foot on the carpet 

8 induced rotating air movement beneath the foot that re-entrained the particles. If Pb 

9 adsorption onto particles varies with sources, size selectivity of resuspension processes 

10 could lead to enhanced Pb exposure from one source relative to another.

11 Several of the studies can be used to develop an understanding of how personal exposure 

12 to PM-bound Pb varies with other exposures. Molnar et al. (2007) reported Spearman 

13 correlations of Pb with PM2.5 and NO2 in three outdoor microenvironments (residence, 

14 school, and preschool) and found that Pb and other trace metals were generally well 

15 correlated with PM2.5 (r = 0.72-0.85), but Pb was not always well-correlated with NO2 (r 

16 = 0.24-0.75). In the case where Pb and NO2 were well-correlated, it is possible that the Pb 

17 was traffic related from resuspended pulverized wheel weights or impurities in unleaded 

18 on-road gasoline. For the other two sites where the correlation between Pb and NO2 was 

19 low, it is possible that they were less affected by traffic. Table 3-26 in the Appendix to 

20 Chapter 3 illustrates that Pb concentrations in the four studies summarized there are 

21 typically well below the level of the NAAQS. The higher personal air concentratoins 

22 occurred in a heavily industrialized area of Kocaeli, Turkey with an incinerator and 

23 several industrial facilities including metal processing, cement, petroleum refining, 

24 agriculture processing. Otherwise, concentrations were all between 0.002 and 

0.006 μg/m3 25 . The proportion of Pb compared with other trace metals varied with location 

26 and component. It was typically several times lower than S as well as crustal elements 

27 such as Ca and Fe. In the industrial area of Kocaeli, Pb comprised a greater proportion of 

28 the PM compared with other areas.
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Table 4-3 Comparison of personal, indoor, and outdoor Pb-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, 1997 13 6.6 8.5

Adgate et al. (2007) MinneapolisSt. 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

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

Rasmussen et al.

(2007)

Windsor, Ontario, 

Canada

Med. Pb-PM2.5

(mg/kg)

April, 2004 311 124 221

4.1.3.2 Exposure to Lead in Soil and Dust

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

2 potential source of exposure to Pb soil and dust. Thus, outdoor soil Pb may present an 

3 inhalation exposure if resuspended indoors or an ingestion exposure during hand-to4 mouth contact. A detailed description of studies of outdoor soil Pb concentration is 

5 provided in Section 3.6.1. Indoor measurements can reflect infiltrated Pb as well as Pb 

6 dust derived from debrided paint, consumer products, or soil that has been transported 

7 into the home via foot traffic. Table 4-4 presents indoor dust Pb concentrations for 

8 2006-2011 observational studies in which indoor dust Pb was measured.
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Table 4-4 Measurements of indoor dust Pb concentration from 2006-2011 

studies

Reference Study Location Metric (units) Sample Site

Indoor Pb 

Concentration

Caravanos et al.

(2006a) New York City, New York Weekly dust loading 

(μg/m2)

Glass plate next to open window 

of academic building Median: 52

Khoder et al. (2010) Giza, Egypt (extensive leaded gasoline use; 

industrial area)

Weekly dust loading 

(μg/m2

)

Glass plate in second-floor living 

room of apartments Median: 408

Brattin and Griffin

(2011)

Eureka, Utah near Eureka Mills Superfund 

Site

Dust concentration 

(mg/kg)

Indoor home site (not specified) 160-2000

Denver, CO, near VBI70 Superfund Site Indoor home site (not specified) 11-660

East Helena, MT, near East Helena 

Superfund Site Indoor home site (not specified) 68-1000

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

Wilson et al. (2007) Milwaukee, Wisconsin Dust concentration 

(μg/m2

)

Central perimeter Avg.: 107

Entry Avg.: 140

Window Avg.: 151

Zota et al. (2011) Ottawa County, Oklahoma (area 

surrounding the Tar Creek Superfund Site)

Dust concentration 

(mg/kg) Indoor (site not specified)

Avg.: 109

Median: 63

Max.: 881

Spalinger et al.

(2007)

Rural towns, Idaho Dust concentration 

(mg/kg)

Vacuum Median: 120

Max: 830

Floor Median: 95

Max: 1,300

Bunker Hill, Idaho Superfund site Dust concentration 

(mg/kg)

Vacuum Median: 470

Max: 2,000

Floor Median: 290

Max: 4,600

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

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

3 window, a 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 4 for the 

indoor site, 153 μg/m2 for the unsheltered outdoor site, and 347 μg/m2 5 for the sheltered 

6 outdoor site. This paper demonstrated the likely role of outdoor Pb in influencing indoor 

7 dust Pb loading and indicated that under quiescent conditions (e.g., no cleaning) near an 
February 2012 4-15 Draft – Do Not Cite or Quote

1 open second-story window, the indoor dust Pb level might exceed EPA’s hazard level for 

interior floor dust of 430 μg/m2 (40 μg/ft2 2 ). Khoder et al. (2010) used the same 

3 methodology to study Pb dust deposition in residential households in the town of Giza, 

4 Egypt, located between two industrial areas and where leaded gasoline is still in use; the 

investigators reported a median weekly deposition rate of 408 μg/m2 5 and an exterior 

median deposition rate of 2,600 μg/m2 6 . In the latter study, Pb deposition rate correlated 

7 with total dust deposition rate (R=0.92), Cd deposition rate (R=0.95), and Ni deposition 

8 rate (R=0.90). Statistically significant differences in Pb deposition rates were observed 

9 between old and new homes (p <0.01) in the Khoder et al. (2010) study, although the 

10 only quantitative information provided regarding home age stated that the oldest home 

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

12 statistically significant difference between Pb loadings when segregating the data by 

13 proximity to roadways. Recently, Brattin and Griffin (2011) performed linear regressions 

14 of dust Pb on soil Pb based on data collected previously for outdoor soil Pb and indoor 

15 dust Pb near mining and/or smelting Superfund sites in Utah, Colorado, and Montana

16 (U.S. EPA, 2005f, 2002a, 2001). They observed that the dust Pb concentration was 

17 4-35% of outdoor soil Pb. Excluding outliers on the regression, dust Pb concentration 

18 ranged from 160-2,000 mg/kg, 11-660 mg/kg, and 68-1,000 mg/kg at three sites.

19 Correlations between indoor and outdoor Pb content in dust can be partially explained 

20 with speciation. Beauchemin et al. (2011) used XANES to speciate in-home paint 

21 samples to assess the contributions of indoor paint and outdoor material to indoor dust Pb 

22 concentrations. In indoor dust samples of particles <150 μm in size, Pb oxide, Pb sulfate, 

23 and Pb carbonate were measured. These materials commonly were used in white paint. In 

24 the size fraction of particles <36 μm, half of the measured Pb was associated with Fe25 oxyhydroxides such as ferrihydrite and goethite and presumably adsorbed onto these 

26 species. This finding suggested that a mix of indoor and outdoor sources may affect the 

27 composition of dust in this size fraction.

28 Residual Pb dust contamination following cleaning activities has been documented. For 

29 instance, Hunt et al. (2008) performed tests where a test soil prepared by drying, 

30 grinding, and sieving Pb-contaminated yard soil samples from Herculaneum, MO was 

31 tracked onto a tile test surface and then repeatedly cleaned with a moistened wipe and/or 

32 vacuumed until visual inspection of the tiles uncovered no surface discoloration. The 

33 authors then used wet wipe samples to collect residual soil and estimate Pb deposition 

and concentration. After the first walk, tile Pb dust loading was 2,670 μg/m2 34 ; after the 

first vacuuming, it decreased to 398 μg/m2 35 . After multiple walks, tile Pb dust loading was 

7,100 μg/m2, and it decreased to 1,400 μg/m2 36 after multiple vacuuming. Scanning 

37 electron microscopy (SEM) of the wipe samples revealed that most of the residual dust 

38 particles were in the range of 1-3 μm in area equivalent diameter. This result indicates 
February 2012 4-16 Draft – Do Not Cite or Quote

1 that Pb-bearing fine particles are not completely captured by home cleaning. Yu et al.

2 (2006) analyzed dust samples from 50 homes in northern New Jersey (typically of older 

3 housing stock, although the study does not specify housing age). The investigators 

4 obtained dust from vacuuming carpet samples and found that total Pb concentration in 

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

6 Pb dust on floors, windowsills, and other accessible surfaces are potential exposure 

7 sources to small children who use touch to explore their environments. Gaitens et al.

8 (2009) used National Health and Nutrition Examination Survey (NHANES) data from 

9 1999 through 2004 to examine Pb in dust in homes of children ages 12-60 months. The 

median value of Pb dust loading on floors was reported to be 1.7 μg/m2 10 (mean: 

4.4 μg/m2 11 ), with floors that were not smooth and cleanable having a median Pb dust 

loading value of 5.6 μg/m2 (mean: 16 μg/m2 12 ). Floor Pb dust loading value was modeled 

13 against several survey covariates and was significantly associated (p <0.05) with floor 

14 surface condition, windowsill Pb dust loading, race and ethnicity, poverty-to-income 

15 ratio, year of home construction, presence of smokers in the home, and year of survey. It 

16 was nearly significantly associated (p = 0.056) with renovations made to pre-1950 homes. 

Median Pb dust loading on smooth windowsills was 25 μg/m2 (mean: 190 μg/m2 17 ). When 

windowsills were not smooth, the median Pb dust loading was 55 μg/m2 18 (mean: 

480 μg/m2 19 ). Windowsill Pb dust level was also significantly associated (p <0.05) with 

20 race and ethnicity, year of home construction, window surface condition, presence of 

21 smokers in the home, deterioration of indoor paint, and year of survey. Sill surface was 

22 nearly significantly associated (p = 0.076) with deterioration of outdoor paint when 

23 homes were built prior to 1950. Dust Pb loading was found by Egeghy et al. (2005) to be 

24 significantly associated with the log-transform of soil Pb concentration, cement content in 

25 the home entryway, indoor pesticide use, frequency of fireplace usage, number of hours 

26 in which windows were open, and homes built before 1950 (Table 4-2). Wilson et al.

27 (2007) studied Pb dust samples from homes in Milwaukee, WI children with and without 

28 elevated blood Pb ≥ 10 μg/dL. They found that Pb dust samples obtained from the floor 

29 were always significantly higher in residences of children with elevated blood Pb, with 

30 the exception of samples from the bathroom floor. Windowsill dust was not significantly 

31 higher in residences of children with elevated blood Pb.

32 Building demolition and renovation activities can create dust from interior and exterior 

33 paints with Pb content. Mielke and Gonzales (2008) measured Pb content in paint chips 

34 from paint applied prior to 1992 and found that median Pb levels were 420 mg/kg for 

35 interior paint and 77,000 mg/kg for exterior paint. Maximum levels were 63,000 mg/kg 

36 and 120,000 mg/kg for interior and exterior paint, respectively. Mielke et al. (2001)

37 compared dust samples from two New Orleans houses that were prepared for painting. 

38 One home was power sanded, while the other was hand-scraped. Immediately after 
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1 sanding, Pb dust samples ranged from <3 to 28,000 mg/kg at the sanded house. Pb dust 

2 samples from the scraped house ranged from 7 to 1,200 mg/kg. Pb in dust or paint 

3 samples was not quantified.

4 Dust Pb concentrations have also been reported for homes in the vicinity of historic and 

5 active metals mining and smelting sources. Near an active smelter in Port Pirie, Australia, 

6 median hand dust Pb loadings increased with age among a cohort of fourteen children 

followed over age 0-36 months (2-5 months: 54 μg/m2, 6-9 months: 173 μg/m2 7 , 10-15 

months: 424 μg/m2, > 15 months: 336 μg/m2 8 ) (Simon et al., 2007). Zota et al. (2011)

9 studied Pb dust and indoor Pb-PM2.5 concentration in Ottawa County, OK near the Tar 

10 Creek Superfund Site, in which a metals mine had closed. Statistically significant 

11 correlations among outdoor soil Pb concentration, indoor dust Pb concentration, indoor 

12 dust Pb loading, and indoor air Pb-PM2.5 concentrations were observed (r = 0.25-0.65), 

with an average dust Pb concentration of 109 mg/kg, dust Pb loading of 54 μg/m2 13 , soil Pb 

concentration of 201 mg/kg, and indoor Pb-PM2.5 concentration of 1 ng/m3 14 . Dust Pb 

15 concentrations were found to increase significantly with proximity to two chat (i.e., dry 

16 mining waste) sources and to decrease with distance to the street and presence of central 

air conditioning. Spalinger et al. (2007) measured Pb in dust in homes in a 34 km2 17 area 

18 surrounding a designated Superfund site where formerly a Pb and Zn smelter operated at 

19 Bunker Hill, ID. During spring of 1999, vacuum and floor mat samples were taken from 

homes in three towns within the 34 km2 20 area and five “background” towns further from 

21 the Superfund site. For the background towns, Pb concentration in vacuum dust had a 

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

23 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 24 per day. Among the background homes, median 

25 vacuum and floor mat Pb dust samples were 3 and 2.5 times higher, respectively, when 

26 comparing homes built before 1960 with those built after 1960. Deposition rate of Pb 

27 dust was 5 times higher in the older homes. In contrast, Pb in vacuum dust and floor mats 

28 for the towns contained within the Bunker Hill Superfund site had a median Pb 

29 concentration of 470 mg/kg with a maximum of 2,000 mg/kg and a median of 290 mg/kg 

30 with a maximum of 4,600 mg/kg, respectively. The median Pb loading rate for indoor 

dust in houses in these towns was 300 μg/m2 31 per day, and the maximum Pb dust loading 

rate was 51,000 μg/m2 32 per day. These results suggest that those living in close proximity 

33 to large Pb and Zn smelters or mines that are now Superfund sites are at much greater 

34 risk of exposure to Pb dust compared to the general population.

35 Pb exposure has been reported on children’s playgrounds. Mielke et al. (2011b) reported 

36 median soil Pb concentration of 558 mg/kg on playground soils at eleven New Orleans 

37 daycare or community centers. Following remediation efforts to cover playground soil 

38 with clean soil, median concentration dropped to 4.1 mg/kg. Duggan et al. (1985)
February 2012 4-18 Draft – Do Not Cite or Quote

1 reported on the concentration and size distribution of wipe samples on the hands of 368 

2 pre-school children from eleven schools in London, UK. Hand Pb residue (PbH) values 

3 were modeled as linear and power functions of Pb dust (PbD) from this study to obtain 

PbH = 16.1 + 0.0064PbD or PbH = 0.393PbD0.533 4 . Given that the Duggan et al. (1985)

5 study was performed when Pb additives were used in gasoline, dust Pb concentration 

6 values are not reported here. Measurements of the size distribution produced geometric 

7 mean diameters of 4.5 μm and 1.5 μm depending on the analysis method. The hand wipe 

8 samples were effectively limited to particles smaller than 10 μm, even when the soil itself 

9 exhibited a much larger particle size distribution. However, these studies focused on 

10 playground lead exposure. Que Hee et al. (1985) performed repeated wipe sampling of 

11 hands after rubbing in a reference dust sample and in reporting results for a “small adult 

12 hand” observed that loose particles up to 246 μm in diameter adhered, which is similar to 

13 the upper limit of 200-300 μm observed by Yamamoto et al. (2006). If a similar 

14 relationship holds for household dust, then exposure to Pb in dust by hand-to-mouth 

15 activity may be influenced by the dust particle size distribution and the relationship of Pb 

16 content to particle size fraction.

4.1.3.3 Dietary Lead Exposure

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

18 Included among those are drinking water, fish and meat, agriculture, urban gardening, 

19 dietary supplements, tobacco, cultural food sources, and breastfeeding. The breadth of 

20 dietary Pb exposures is illustrated in Figure 4-2, which illustrates the data obtained in the 

21 2008 FDA Total Diet Study market basket survey (FDA, 2008). Among the highest Pb 

22 concentrations were those for noodles, baby food carrots, baby food oatmeal, Swiss 

23 cheese, beef tacos from a Mexican restaurant, and fruit-flavored cereal. The source of the 

24 Pb in each case is unclear. Possible sources of Pb in food samples include introduction 

25 during processing or preparation with drinking water contaminated with Pb, deposition of 

26 Pb onto raw materials for each food, and Pb exposure in livestock that produce dairy or 

27 meat ingredients. Manton et al. (2005) used Pb isotope ratios to estimate sources of 

28 dietary Pb among a cohort of mothers and children from Omaha, NE using a combination 

29 of food samples, hand wipes, house dust wipes, and aerosol samples collected between 

30 1990 and 1997 and speciated for Pb isotopes. Drinking water Pb was not included in this 

31 study. The authors cited results from Egan et al. (2002) that imported vegetables 

32 contributed 55% of Pb dietary intake for infants, 30% for 2-6 y old children, and 20% for 

33 25-30 y old women. Imported candy contributed 10% of Pb dietary intake for 2-6 y old 

34 children and 9% for 25-30 y old women. Isotopic data from Manton et al. (2005)

35 suggested that, with the exception of children age 0-12 mos, house dust is a large 
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1 contributor to dietary Pb. The pattern of certain Pb-isotope ratios observed in the diet of 

2 children 0-12 mos are suggested to derive from Ca salts in limestone that may have been 

3 used in dietary supplements in baby formula. The contribution of ambient Pb aerosols to 

4 dietary Pb samples was not statistically significant for this urban exposure study.

Source Data: (FDA, 2008)

Note: from the 2008 FDA Total Diet Study. “BF” denotes baby food.

Figure 4-2 Market basket survey results for Pb concentration in foods.
February 2012 4-20 Draft – Do Not Cite or Quote

Drinking Water

1 Differences in sources and transport of drinking water may cause variation in water Pb 

2 levels. For example, Shotyk and Krachler (2009) measured the Pb concentration in tap 

3 water, commercially bottled tap water and bottled natural water. They found that, in 

4 many cases, tap water contained less Pb than bottled water. Excluding bottled water in 

5 glass containers because Pb can be leached from the glass, the median Pb concentration 

6 in the bottled water samples was 8.5 ng/L (range ≤ 1 to 761 ng/L). This level (i.e., 8.5 

7 ng/L) is significantly less than the Maximum Contaminant Level for Pb of 15 μg/L. Pb in 

8 drinking water supplies can derive from atmospheric deposition onto surface waters, as 

9 described in Section 3.3.1.2, or via corrosion of Pb in the distribution network 

10 exacerbated by contact with acidic disinfection byproducts, as described in the following 

11 paragraphs. 

12 It is now recognized that environmental nanoparticles (NPs) (~1-100 nm) can play a key 

13 role in determining the chemical characteristics of engineered as well as natural waters

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

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

16 the flocculation/coagulation stage of the treatment process, they may become effective 

17 transporters of contaminants such as Pb, particularly if these contaminants are leached 

18 from piping in the distribution system. Edwards and Dudi (2004) observed a red-brown 

19 particle-bound Pb in Washington, DC water that could be confused with innocuous Fe. 

20 The source of the particle-bound Pb was not known but was thought to originate from the 

21 source water.

22 Corrosion byproducts can influence Pb concentrations in drinking water. Schock et al.

23 (2008) characterized Pb pipe scales from 91 pipes made available from 26 different 

24 municipal water systems from across the northern U.S. They found a wide range of 

25 elements including Cu, Zn and V as well as Al, Fe and Mn. Interestingly, V was present 

26 at nearly one percent levels in pipes from many geographically diverse systems. In a 

27 separate study, Gerke et al. (2009) identified the corrosion product, vanadinite 

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

29 representing 8 different municipal drinking water distribution systems in the Northeastern 

30 and Midwest regions of the U.S. Vanadinite was most frequently found in the surface 

layers of the corrosion products. The vanadate ion, VO4

3- 31 , essentially replaces the 

32 phosphate ion in pyromorphite and hydroxyapatite structures. It is not known whether the 

33 application of orthophosphate as a corrosion inhibitor would destabilize vanadinite, but 

34 this substitution would have implications for V release into drinking water. The stability 

35 of vanadinite in the presence of monochloramine is also not known, and its stability 

36 might have implications for both Pb and V release into drinking water.
February 2012 4-21 Draft – Do Not Cite or Quote

1 In recent years, drinking water treatment plants in many municipalities have switched

2 from using chlorine to other disinfecting agents because their disinfection byproducts 

3 may be less carcinogenic. However, chloramines are more acidic than chlorine and can 

4 increase Pb solubility (Raab et al., 1991) and increase Pb concentrations in tap water. For 

5 example, after observing elevated Pb concentrations in drinking water samples, Kim and 

6 Herrera (2010) observed Pb oxide corrosion scales occurring after using acidic alum as a 

7 disinfection agent. High Pb concentrations in Washington, DC drinking water were 

8 attributed to leaching of Pb from Pb-bearing pipes promoted by breakdown products of 

9 disinfection agents (Edwards and Dudi, 2004). Maas et al. (2007) tested the effect of 

10 fluoridation and chlorine-based (chlorine and chloramines) disinfection agents on Pb 

11 leaching from plumbing soldered with Pb. When using chlorine disinfection agents alone, 

12 the Pb concentration in water samples doubled during the first week of application (from 

13 100 to 200 ppb) but then decreased over time. When adding fluorosilicic acid and 

14 ammonia, the Pb concentration spiked to 900 ppb and increased further over time. 

15 Similarly, Lasheen et al. (2008) observed leaching from pipes in Egypt. In this study, the 

16 authors tested polyvinyl chloride (PVC), polypropylene (PP), and galvanized iron pipes 

17 and observed leaching from both the PVC and PP pipes when exposed to an acid of pH = 

18 6, with PVC having greatest amount of leaching. Exposure to basic solutions actually 

19 resulted in reduction of Pb concentration in the drinking water. 

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

21 NC as a function of household age, use of chloramines and other covariates. Blood Pb 

22 levels were significantly associated with the year the home was built (p <0.001), use of 

23 chloramines (p <0.001), and the interaction between these two variables (p <0.001). 

24 When year in which the home was built was broken into categories for the independent 

25 variables and interaction terms, Miranda et al. (2007a) found that significance increased 

26 with the age of the home. However, the study did not control for the presence of Pb paint 

27 in the dwellings, so it is difficult to distinguish the influence of Pb pipes from Pb in paint 

28 on blood Pb levels.

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

30 disinfection agents. Edwards and Dudi (2004) hypothesized that Pb leaching occurs when 

31 chloramines cause the breakdown of brass alloys and solder containing Pb. After 

32 observing that nitrification also leads to increased Pb concentrations in water, they also 

33 proposed that chloramines may trigger nitrification and hence cause decreasing pH, 

34 alkalinity and dissolved oxygen that lead to corrosion after observing that nitrification 

35 also leads to increased Pb concentrations in water. However, Zhang et al. (2009b) found 

36 no evidence that nitrification brought about significant leaching of Pb from Pb pipes. 

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

38 water following a change from free chlorine to chloramines disinfection is attributed to 
February 2012 4-22 Draft – Do Not Cite or Quote

1 the formation of the Pb(II) mineral hydroxypyromorphite (Pb5(PO4)3OH) instead of 

2 Pb(IV) oxide. Xie et al. (2010) further investigated the mechanisms by which Pb(II) 

3 release is affected by chloramines. Two opposing mechanisms were proposed: Pb(IV)O2

4 reduction by an intermediate species from decomposition of monochloramine; and 

5 increasing redox potential which decreases the thermodynamic driving force for 

6 reduction. They suggest that the contact time of monochloramine with PbO2 and the 

7 Cl2:N ratio in monochloramine formation will determine which mechanism is more 

8 important. Free chlorine can control Pb concentrations from dissolution under flowing 

9 conditions but for long stagnation periods, Pb concentrations can exceed the action level: 

10 4-10 days were required for Pb concentrations to exceed 15 μg/L (for relatively high 

11 loadings of PbO2 of 1 g/L). Thus, under less extreme conditions, it was concluded that 

12 chloramination was unlikely to have a major effect on the release of Pb into drinking 

13 water.

Agriculture

14 The 2006 Pb AQCD (2006b) states that surface deposition “represents a significant 

15 contribution to the total Pb in and on the plant”, while uptake through a plants roots can 

16 also contribute to a plant’s Pb concentration. Consequently, Pb content in plants may 

17 contribute to human dietary exposure. Uptake of Pb by plants growing in contaminated 

18 soil has been repeatedly demonstrated in some species during controlled potted plant 

19 experiments (Del Río-Celestino et al., 2006). In this study, most species retained Pb in 

20 the roots with little mobilization to the shoots of the plants. However, certain species 

21 Cichorium intybus [chicory], Cynodon dactylon [Bermuda grass], Amaranthus blitoides

22 [matweed or mat amaranth], and Silybum marianum [milk thistle]) were able to mobilize 

23 Pb from the roots to the shoots of the plant; these specific species could lead to human 

24 exposures through consumption of grazing animals. Lima et al. (2009) conducted similar 

25 greenhouse experiments with several vegetable crops grown in soil contaminated by Pb26 containing residue from battery recycling waste. In this study, carrots had high 

27 bioaccumulation, measured as the percent of Pb concentration measured in the plant 

28 compared with the Pb concentration in the soil, with little translocation of the Pb to the 

29 shoots, measured as the percent of Pb mass in the shoots compared to the Pb mass within 

30 the entire plant, of the Pb to the shoots. Conversely, beets, cabbages, sweet peppers, and 

31 collard greens had low bioaccumulation but moderate to high translocation. Okra, 

32 tomatoes, and eggplants had moderate bioaccumulation and moderate to high 

33 translocation. Sesli et al. (2008) also noted uptake of Pb within wild mushrooms. 

34 Vandenhove et al. (2009) compiled bioaccumulation data for plant groupings from 

35 various references; these data are reproduced in Table 4-5. Based on this review, grasses 

36 had the highest uptake, followed by leafy vegetables and root crops grown in sandy soils; 
February 2012 4-23 Draft – Do Not Cite or Quote

1 these references also suggested high transfer from roots to shoots among root crops, with 

2 shoots having roughly four times higher Pb bioaccumulation than roots. This is consistent 

3 with the 2006 Pb AQCD (2006b), which reported that Pb deposition onto leafy 

4 vegetables accounted for most Pb.

5 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 6 =0.96) in lettuce plants cultivated 

7 in the courtyard of a smelter. They reported that lettuce grown 250-400 m from the 

8 smelter had concentrations that were 10-20 times lower, which is consistent with findings 

9 described in Section 3.3 that deposition of Pb containing material drops off with distance 

10 from a source. Pb contamination of crops can occur through Pb emissions during aerial 

11 application of fertilizers and pesticides. In 2009, the U.S. Federal Aviation 

12 Administration (FAA) recorded 960,000 hours of flight time for aerial application. This 

13 term encompasses crop and timber production including seeding cropland and fertilizer 

14 and pesticide application. It is estimated that 42% of these flight-hours involved piston 

15 engine aircraft utilizing leaded fuel (FAA, 2009).

16 Fernandez et al. (2010; 2008; 2007) measured Pb from atmospheric deposition in two 

17 adjacent plots of land having the same soil composition but different uses: one was 

18 pasture land and one was agricultural. In the arable land, size distributions of soil 

19 particle-bound Pb, were uniformly distributed. In pasture land, size distributions of soil 

20 particle-bound Pb were bimodal with peaks around 2-20 μm and 50-100 μm (Fernandez 

21 et al., 2010). For the agricultural plot, Pb concentration was constant around 70 mg/kg in 

22 samples taken over the first 30 cm of soil, at which time it dropped below 10 mg/kg at 

23 soil depths between 35 and 100 cm. In contrast, Pb concentration in pasture land peaked 

24 at a depth of 10 cm at a concentration of roughly 70 mg/kg and then dropped off 

25 gradually to approach zero concentration at a depth of approximately 50 cm. The sharp 

26 change in concentration for the arable land was attributed to a combination of plowing 

27 the soil and use of fertilizers to change the acidity of the soil and hence the 

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

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

30 There is some evidence that Pb in crops can originate with treatment of crops. For 

31 example, compost produced from wastewater sludge has the potential to add Pb to crops. 

32 Cai et al. (2007) demonstrated that production of compost from sludge enriched the Pb 

33 content by 15-43% prior to its application. Chen et al. (2008a) observed that the median 

34 concentration of Pb in California crop soil samples was 16.2 mg/kg (range: 

35 6.0-62.2 mg/kg). Chen et al. (2008b) further observed that in three of the seven California 

36 agricultural regions sampled, concentrations of Pb increased following addition of 

37 fertilizer, but the increase was less than that for P and Zn indicators of fertilizer. In four 
February 2012 4-24 Draft – Do Not Cite or Quote

1 regions, there was no increase of Pb at all. Furthermore, Tu et al. (2000) observed a 

2 decrease in Pb fraction with increasing P application. Nziguheba and Smolders (2008)

3 also surveyed phosphate-based fertilizers sold in European markets to determine the 

4 contribution of these fertilizers to heavy metal concentrations in agricultural products. 

5 They reported a median fertilizer Pb concentration of 2.1 mg/kg based on total weight of 

6 the fertilizer, with a 95th percentile concentration of 7.5 mg/kg. Across Europe, 

7 Nziguheba and Smolders (2008) estimated that the amount of Pb applied via fertilizers to 

8 be only 2.6% of that resulting from atmospheric deposition. 

9 Although Pb in on-road vehicle gasoline has been phased out in the U.S., some imported 

10 crops are produced in countries that still use Pb antiknock agents in on-road gasoline. For 

11 example, high concentrations of Pb have been found in chocolate from beans grown in 

12 Nigeria, during the time when leaded gasoline was still legally sold. Rankin et al. (2005)

observed that the ratios of 207Pb to 206Pb and 208Pb to 207 13 Pb were similar to those of Pb in 

14 gasoline. Although this study showed that Pb concentration in the shelled cocoa beans 

15 was low (~1 ng/g), manufactured cocoa powder and baking chocolate had Pb 

16 concentrations similar to those of the cocoa bean shells, on the order of 200 ng/g, and Pb 

17 concentration in chocolate products was roughly 50 ng/g (Rankin et al., 2005). It is 

18 possible that the increases were attributed to contamination of the cocoa by the shells 

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

20 the source of contamination. Likewise, it is possible that resuspended Pb that originated 

21 from legacy mobile and industrial sources could deposit on crops. 

22 The extent to which soil Pb contributes Pb to agricultural crops varies with soil 

23 characteristics. For example, Jin et al. (2005) tested soil Pb, bioaccessibility of soil Pb 

24 (determined by CaCl2 extraction), and Pb in tea samples from tea gardens. They observed 

25 that the Pb concentration in tea leaves was proportional to the bioaccessible Pb in soil.
February 2012 4-25 Draft – Do Not Cite or Quote

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% 2,500%

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% 2,500%

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% 1,600%

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: Reprinted with permission of Elsevier Publishers, Vandenhove et al. (2009).

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

2 be contaminated with Pb, as described in Section 4.1.3.2. For instance, Clark et al. (2006)

3 tested the soil in 103 urban gardens in two Boston neighborhoods. They found that Pb4 based paint contributed 40-80% of Pb in the urban garden soil samples, with the rest 

5 coming from historical gasoline emissions. Furthermore, Clark et al. (2006) estimated 
February 2012 4-26 Draft – Do Not Cite or Quote

1 that Pb consumption from urban gardens can be equivalent to 10-25% of the exposure to 

2 Pb from drinking water for children living in the Boston neighborhoods studied. Because 

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

4 exposures in urban garden vegetables will vary.

Game

5 Game meat consumption also may pose a risk of Pb exposure. In Pb mining or smelting 

6 areas, several studies have documented Pb concentrations in game [e.g., (Nwude et al., 

7 2010; Reglero et al., 2009a)]. Jankovská et al. (2010) studied Pb accumulation in sheep 

8 infected with tapeworm and found that sheep harboring the parasites contained 

9 significantly less Pb compared with their uninfected counterparts (p ≤ 0.05). 

10 Potential Pb exposure through consumption of animals exposed to or killed with Pb shot 

11 has also been well documented (Hunt et al., 2009; Tsuji et al., 2009; Tsuji et al., 2008; 

12 Hunt et al., 2006). For example, Martínez-Haro et al. (2010) observed Pb in the feces of 

13 mallards that ingested gunshot of 34-13,930 mg/kg with a median of 1,104 mg/kg, while 

14 mallards that did not ingest gunshot had feces Pb levels < 12.5 mg/kg. Mateo et al. (2011)

15 studied Pb bioaccessibility as a function of cooking method for breast meat from 

16 partridges killed with gunshot. They observed that preparation in cold or hot vinegar 

17 increased bioaccessibility compared with total Pb in the samples. 

Fish

18 Accumulation in fish could also lead to human exposure to Pb (U.S. EPA, 2006b, 1986a).

19 Ghosh et al. (2007) demonstrated in laboratory experiments that exposure to Pb in water 

20 can lead to linearly increasing accumulation in fish. Several studies have documented the 

21 potential for human exposure through fish and seafood. Welt et al. (2003) conducted a 

22 survey of individuals who fished in Bayou St. John, Louisiana in conjunction with 

23 sampling Pb content in sediment. They found that median sediment Pb concentrations 

24 ranged from 43 to 330 mg/kg in different locations, while maximum sediment Pb 

25 concentrations ranged from 580 to 6,500 mg/kg. In total, 65% of the surveyed individuals 

26 fished for food from the Bayou, with 86% consuming fish from the Bayou each week. In 

27 a study of the effect of coal mining on levels of metals in fish (measured as blood Pb) in 

28 northeastern Oklahoma, Schmitt et al. (2005) found that Pb concentrations in blood

29 varied with respect to species of fish, but Pb concentrations were higher in fish in areas 

30 close to mining activities. Similarly, Besser et al. (2008) observed higher levels of blood 

31 Pb in fish close to mining activities in southeastern Missouri. In a related study of fish 

32 species in the same region of Missouri, blood Pb levels in fish were found to be 

33 significantly higher in sites within 10 km downstream of active Pb-Zn mines (p <0.01) 
February 2012 4-27 Draft – Do Not Cite or Quote

1 compared with fish located further from the mines (Schmitt et al., 2007a), and elevated 

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

3 noted that the Ozark streams where these studies were performed were often used for 

4 recreational fishing. 

Breast Milk

5 Studies breastfeeding women suggest that infants may be exposed to Pb in breast milk. 

6 Ettinger et al. (2004a) observed among Mexico City women studied in 1994-1995 that at 

7 1 month postpartum, 88 women breastfeeding exclusively (with mean blood Pb level of 

8 9.4 μg/dL) had breast milk Pb concentrations of 1.4 ± 1.1 μg/L, and 165 women 

9 breastfeeding partially (with mean blood Pb level of 9.5 μg/dL) had breast milk Pb 

10 concentrations of 1.5 ± 1.2 μg/L. During the same time period, Ettinger et al. (2006)

11 studied breastfeeding women in Mexico City over a child’s first year of life and sampled 

12 Pb concentration in breast milk at 1, 4, and 7 mo post-partum. They observed that mean 

13 breast milk concentrations dropped from 1.4 μg/L at 1 mo (mean maternal blood Pb = 

14 9.3 μg/dL) to a mean of 1.2 μg/L at 4 mo (mean maternal blood Pb = 9.0 μg/dL) to 

15 0.9 μg/L at 7 mo (mean maternal blood Pb = 8.1 μg/dL); this reduction was statistically 

16 significant (p < 0.00001). Among the 310 women included in the study, 181 had previous 

17 pregnancies. In one study of nursing mothers living in Pb contaminated city in Australia, 

18 10 of the 11 mothers had breast milk concentrations <5 μg/L (Simon et al., 2007). The 

19 authors hypothesized that breast milk concentration was too low to be a major contributor 

20 to blood Pb level in these infants relative to other factors such as hand loading of lead. 

21 However, one mother with a blood Pb level of 25 μg/dL had a breast milk Pb level of 

22 28 μg/L (Simon et al., 2007).

4.1.3.4 Occupational

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

24 day occupational exposures are briefly discussed below in the context of understanding 

25 potential exposures that are not attributed to ambient air. For example, Miller et al.

26 (2010) obtained personal and area samples of particle-borne Pb in a precious metals 

27 refinery; year of the study was not reported. It was not stated explicitly, but it is likely 

28 that Miller et al. (2010) measured the PM as TSP because the Occupational Safety and 

29 Health Administration (OSHA) permissible exposure limit (PEL) for Pb is based on TSP 

30 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 31 , and 

concentrations from area samples ranged from 4 to 14 μg/m3. The OSHA PEL is 5 μg/m3 32 . 
February 2012 4-28 Draft – Do Not Cite or Quote

1 In steel production, sintering was found to be the largest source of airborne Pb exposure 

2 in a survey of operations (Sammut et al., 2010), with Pb enrichment in PM reported to be 

3 20,000 mg/kg, although total PM concentration, reported to have 75% of particulate mass 

4 below 2.5 μm diameter, was not reported; year of the study was also not reported.

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

6 in addition to a residential exposure. Rodrigues et al. (2010) reported exposures to 

7 airborne Pb among New England painters, who regularly use electric grinders to prepare 

8 surfaces for painting. Two-week averaged airborne Pb concentrations, sampled with an 

9 Institute of Medicine inhalable PM sampler designed to capture PM smaller than 100 μm, 

were reported to be 59 μg/m3, with a maximum daily value of 210 μg/m3 10 . The Pb 

11 concentrations reported here were corrected by the National Institute for Occupational 

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

13 protection factors were not reported by Rodrigues et al. (2010). Information on the air Pb14 blood Pb relationship can be found in Section 4.5.1. Nwajei and Iwegbue (2007)

15 measured Pb contamination in sawdust; such contamination has been reported to occur 

16 when trees are grown in soil contaminated with Pb (Andrews et al., 1989). Sawdust 

17 samples from fifteen locations in Nigerian sawmills were reported to have Pb 

18 concentrations ranging from 2.0 to 250 mg/kg.

4.1.3.5 Exposure to Lead from Consumer Products

19 Pb is present in varying amounts in several consumer products including alternative 

20 medicines, candies, cosmetics, pottery, tobacco, toys, and vitamins (Table 4-6). Several 

21 of these categories suggest children may incur regular exposures. Pb concentrations were 

22 reported to range from non-detectable levels up to 77% by mass, for the case of one 

23 medicinal product. Exposure to these products, which originate in a range of different 

24 countries, can account for substantial influence on Pb body burden (Miodovnik and 

25 Landrigan, 2009; Levin et al., 2008).

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)
February 2012 4-29 Draft – Do Not Cite or Quote

Product 

Category Product

Location of 

Purchase Pb Content (units) Reference

Yugmijihwang-tang, Bojungigki-tang, Sibjeondaebo-tang, Kuibi-tang, 

Ojeogsan

Korea 7.9×10-6 to 2.5×10-5 mg/kg body 

weight/day

Kim et al. (2009a)

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, Chama Black Stone, Portugal Antiseptic Soap, 

Edysol Antiseptic Soap, H-Nal, M-Reg, Veins Flocher, Diabor, C-Candi, CCysta, 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 U.K. 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

Note that the country of origin is not provided because it was not published in the references cited.
February 2012 4-30 Draft – Do Not Cite or Quote

4.2 Kinetics

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

2 humans. The large amount of empirical information on Pb biokinetics in humans and 

3 animal models has been integrated into mechanistic biokinetics models (U.S. EPA, 

4 2006b). These models support predictions about the kinetics of Pb in blood and other 

5 selected tissues based on the empirically-based information about Pb biokinetics. In 

6 Section 4.3 (and Section 4.2.2.1), Pb biokinetics is described from the context of model 

7 predictions.

4.2.1 Absorption

8 The focus of the following sections within absorption is on inhalation and ingestion 

9 because these are the major exposure routes of Pb in humans. The 2006 Pb AQCD also 

10 presented dermal absorption of inorganic and organic Pb compounds, which is generally 

11 considered to be much less than by inhalation or ingestion. A study published subsequent 

12 to the 2006 Pb AQCD measured rates of absorption of Pb in skin patches harvested from 

13 nude mice (Pan et al., 2010). Following application of 12 mg Pb as Pb-acetate or Pb 

14 nitrate, the absorption rate (measured over a 10-hour observation period) was 

approximately 0.02 μg Pb/cm2 15 per hour. Absorbed Pb was detected in liver and kidney of 

16 nude mice following a 120-hr occluded dermal application of approximately 14 mg Pb as 

17 either Pb-acetate or Pb nitrate. Uptake of Pb into the skin at the site of application was 

18 greater when Pb-acetate was applied to the skin compared to lead nitrate; however, liver 

19 and kidney Pb concentrations observed at the conclusion of the study (120 hours 

20 following the application of Pb) were not different for the two Pb compound. No 

21 additional information provides evidence of dermal absorption being a major exposure 

22 route of environmental Pb.

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

24 absorbed from the respiratory or gastrointestinal tract. The term bioavailability, as it is 

25 used in this section, refers to the fraction of the amount of Pb ingested or inhaled that 

26 enters the systemic circulation. If properly measured (e.g., time-integrated blood Pb), 

27 under most conditions Pb bioavailability is equivalent (or nearly equivalent) to Pb 

28 absorption. Bioaccessibility is a measure of the physiological solubility of Pb in the 

29 respiratory or gastrointestinal tract. Pb must become bioaccessible in order for absorption 

30 to occur. 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+ 31 ). 

32 Bioaccessibility is typically assessed by measuring the fraction of Pb in a sample that can 
February 2012 4-31 Draft – Do Not Cite or Quote

1 be extracted into a physiological or physiological-like solution (e.g., gastric juice or 

2 solution similar to gastric juice).

4.2.1.1 Inhalation

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

4 and solubility, as well as by the pattern of regional deposition within the respiratory tract. 

5 Fine particles (<1 μm) deposited in the bronchiolar and alveolar region can be absorbed 

6 after extracellular dissolution or can be ingested by phagocytic cells and transported from 

7 the respiratory tract (Bailey and Roy, 1994). Larger particles (>2.5 μm) that are primarily 

8 deposited in the ciliated airways (nasopharyngeal and tracheobronchial regions) can be 

9 transferred by mucociliary transport into the esophagus and swallowed, thus being 

10 absorbed via the gut. 

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

12 regardless of chemical form (Morrow et al., 1980; Chamberlain et al., 1978; Rabinowitz 

13 et al., 1977). Absorption half-times (t1/2) have been estimated for radon decay progeny in 

adults who inhaled aerosols of Pb and bismuth isotopes generated from decay of 220 14 Rn or 

222 15 Rn. The absorption half-time for Pb from the respiratory tract to blood was estimated 

16 to be approximately 10 hours in subjects who inhaled aerosols having an activity median 

17 particle diameter of approximately 160 nm (range 50-500 nm) (Marsh and Birchall, 

18 1999), and approximately 68 min for aerosols having diameters of approximately 0.3–

19 3 nm (Butterweck et al., 2002). Given the submicron particle size of the exposure, these 

20 rates are thought to represent, primarily, absorption from the bronchiolar and alveolar 

21 regions of the respiratory tract.

22 Several studies have quantified the bioaccessibility of Pb in atmospheric PM, based on 

23 various in vitro extraction methods. In a study of PM10 and PM2.5 samples from 

24 downtown Vienna, Austria, Falta et al. (2008) used synthetic gastric juice to investigate 

25 the bioaccessibility of metals including Pb. The rationale was that inhaled particles in the 

26 2.5-10 μm size range are mostly deposited in the tracheal and bronchial regions of the 

27 lung from where they are transported within hours by mucociliary clearance, i.e., they are 

28 mainly swallowed. In contrast, the <2.5 μm particles are deposited in the pulmonary 

29 alveoli where they can stay for months to years. The study aimed to determine the 

30 bioaccessibility of the 2.5-10 μm PM. It is important to note that they do not isolate the 

31 2.5-10 μm size range; instead, they infer the characteristics from the difference between 

32 the PM2.5 and PM10 fractions. The Pb concentrations associated with the two fractions 

33 were almost identical, as was the percentage extracted by synthetic gastric juice (86% and 

34 83% Pb for PM2.5 and PM10 fractions, respectively). The mean daily bioavailable mass 
February 2012 4-32 Draft – Do Not Cite or Quote

1 was calculated to be 16 ng for the PM2.5-10 size range. Since the quantitative clearance of 

2 these particles to the stomach was assumed, this value represents an upper estimate for 

3 the amount of bioavailable Pb. Niu et al. (2010) determined the bioaccessibility of Pb in 

4 fine (100-1,000 nm) and ultrafine-sized (<100 nm) urban airborne PM from two sites 

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

6 particles smaller than 10 μm were 8,800 and 7,800 mg/kg for the two different locations. 

7 The bioaccessibility was based on ammonium acetate extractability and it was found that, 

8 within the fine and ultrafine-size ranges, 13-28% Pb was extracted. The Falta et al.

9 (2008) and Niu et al. (2010) results illustrate that different extraction techniques result in 

10 different bioavailable fractions. The main finding from Niu et al. (2010) was that the 

11 highest values (~28% and ~19% for the two different locations) were found for the 

12 <57 nm particles, with percent bioaccessibility decreasing with increasing particle size. 

13 This result indicated that Pb was potentially most bioavailable in the ultrafine-size range.

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

15 urban road dust in Manchester, UK, and considered the health implications of inhalation 

16 and ingestion of such material. Human exposure via inhalation is likely to involve only 

17 the finest grained fractions (up to 10 μm) and unfortunately this study characterized only 

18 the <38 μm fraction. Pb-goethite and PbCrO4 comprised the largest fractions, 45% and 

19 21% respectively, of Pb in the <38 μm fraction. These forms tend to be less 

20 bioaccessibility if ingested compared with PbO or Pb-acetate because they are less 

21 soluble.

22 The above considerations indicate that the relationship between air Pb exposure and 

23 blood Pb will depend on numerous exposure variables (e.g., particle size, solubility, 

24 exposure frequency and duration) and physiological variables (age, activity level, 

25 transport and absorption in the respiratory tract, blood Pb kinetics). Mechanistic models 

26 provide one means for integrating these variable into predictions of blood Pb – air Pb 

27 relationships; although, predictions are subject to simplifications and generalizations 

28 made in constructing the models. As an example, the ICRP (Pounds and Leggett, 1998; 

29 ICRP, 1994; Leggett, 1993) model (Section 4.3 for a brief description) can be used to 

30 predict blood Pb – air Pb slopes for specific direct Pb inhalation exposure scenarios. For a 

31 long-term continuous (24 hours/day) exposure of a typical adult male engaged in light 

exercise (ventilation rate 20-22 m3 32 /day) to Pb-bearing particles having a 1 μm uniform 

particle size, the predicted blood Pb – air Pb slopes range from 0.7 μg/dL per μg/m3 33 (for 

low solubility particles; e.g., Pb oxide) to 3 μg/dL per μg/m3 34 (for highly soluble Pb; 

35 e.g., Pb salts). Empirical estimates of blood Pb – air Pb slopes for various populations, 

36 derived from epidemiological studies, are summarized in Section 4.5.1.
February 2012 4-33 Draft – Do Not Cite or Quote

Organic Lead

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

2 nearly completely absorbed following deposition in the respiratory tract. As reported in 

the 2006 Pb AQCD, a single exposure to vapors of radioactive (203 3 Pb) tetraethyl Pb 

4 resulted in 37% initially deposited in the respiratory tract, of which ~20% was exhaled in 

5 the subsequent 48 hours (Heard et al., 1979). In a similar experiment conducted with 

203Pb tetramethyl Pb, 51% of the inhaled 203 6 Pb dose was initially deposited in the 

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

8 Estimation of bioavailability of organic Pb is relevant to some aviation fuel exposures 

9 (e.g., persons exposed to leaded gasoline used piston-engine aircraft). Mahaffey (1977)

10 estimated that 40% of inhaled Pb in urban air (largely attributed to combustion of 

11 gasoline containing tetraethyllead) 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 

14 detectable in the blood within 50 hours of inhalation, and the rest was found to deposit in 

15 bone or tissue. Chamberlain et al. (1975) estimated that continuous inhalation of Pb in 

engine exhaust from fuel containing tetraethyllead at a concentration of 0.001 μg/m3 16 for a 

17 period of months could produce a 1 μg/dL increment in blood Pb.

4.2.1.2 Ingestion

18 The extent and rate of GI absorption of ingested inorganic Pb are influenced by 

19 physiological states of the exposed individual (e.g., age, fasting, nutritional calcium and 

20 iron status, pregnancy) and physicochemical characteristics of the Pb-bearing material 

21 ingested (e.g., particle size, mineralogy, solubility). Pb absorption in humans may be a 

22 capacity-limited process, in which case the percentage of ingested Pb that is absorbed 

23 may decrease with increasing rate of Pb intake. Numerous observations of nonlinear 

24 relationships between blood Pb concentration and Pb intake in humans provide support 

25 for the likely existence of a saturable absorption mechanism or some other capacity26 limited process in the distribution of Pb in humans (Sherlock and Quinn, 1986; Sherlock 

27 et al., 1984; Pocock et al., 1983; Sherlock et al., 1982). While evidence for capacity28 limited processes at the level of the intestinal epithelium is compelling, the dose at which 

29 absorption becomes appreciably limited in humans is not known.

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

31 chloride, Pb nitrate, Pb-acetate) range from 3 to 10% in fed subjects (Maddaloni et al., 

32 1998; Watson et al., 1986; James et al., 1985; Heard and Chamberlain, 1982; Rabinowitz 

33 et al., 1980). The absence of food in the GI tract increases absorption of water-soluble Pb 
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1 in adults. Reported estimates of soluble Pb absorption range from 26 to 70% in fasted 

2 adults (Maddaloni et al., 1998; James et al., 1985; Blake et al., 1983; Heard and 

3 Chamberlain, 1982; Rabinowitz et al., 1980). Reported fed:fasted ratios for soluble Pb 

4 absorption in adults range from 0.04 to 0.2 (James et al., 1985; Blake et al., 1983; Heard 

5 and Chamberlain, 1982; Rabinowitz et al., 1980).

6 Limited evidence demonstrates that GI absorption of water-soluble Pb is higher in 

7 children than in adults. Estimates derived from dietary balance studies conducted in 

8 infants and children (ages 2 weeks to 8 years) indicate that ~ 40-50% of ingested Pb is 

9 absorbed (Ziegler et al., 1978; Alexander et al., 1974). Experimental studies provide 

10 further evidence for greater absorption of Pb from the gut in young animals compared to 

11 adult animals (Aungst et al., 1981; Kostial et al., 1978; Pounds et al., 1978; Forbes and 

12 Reina, 1972). The mechanisms for an apparent age difference in GI absorption of Pb have 

13 not been completely elucidated and may include both physiological and dietary factors

14 (Mushak, 1991). Eating breakfast was shown to be significant predictor of blood Pb 

15 concentrations in 1,344 children 3-5 year s old (Liu et al., 2011b). Blood Pb 

16 concentrations were lower in children who regularly ate breakfast compared to children 

17 who did not eat breakfast, and the difference persisted after controlling for nutritional 

18 variables (blood iron, calcium, copper, magnesium, zinc). This observation may be 

19 explained by lower GI absorption of Pb ingested with or in close temporal proximity to 

20 meals. Direct evidence for meals lowering GI absorption of Pb has been reported for 

21 adults (Maddaloni et al., 1998; James et al., 1985).

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

23 particularly in children. Children who are iron-deficient have higher blood Pb 

24 concentrations than similarly exposed iron-replete children, suggesting that iron 

25 deficiency may result in higher Pb absorption or, possibly, other changes in Pb 

26 biokinetics that contribute to altered blood Pb concentrations (Schell et al., 2004; Marcus 

27 and Schwartz, 1987; Mahaffey and Annest, 1986). Studies conducted in animal models 

28 have provided direct evidence for interactions between iron deficiency and increased Pb 

29 absorption, perhaps by enhancing binding of Pb to iron-binding proteins in the intestine

30 (Bannon et al., 2003; Morrison and Quarterman, 1987; Barton et al., 1978b). An analysis 

31 of data from a sample 448 woman (age 20-55 years) did not find a significant association 

32 between iron body stores (indicated from serum ferritin concentration) and blood Pb 

33 concentrations, although depleted irons stores (serum ferritin of <12 μg/L) was associated 

34 with higher blood concentrations of cadmium, cobalt and manganese higher (Meltzer et 

35 al., 2010).The effects of iron nutritional status on blood Pb include changes in blood Pb 

36 concentrations in association with genetic variation in genes involved in iron metabolism. 

37 For example, genetic variants in the hemochromatosis (HFE) and transferrin genes are 

38 associated with higher blood Pb concentrations in children (Hopkins et al., 2008). In 
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1 contrast, HFE gene variants are associated with lower bone and blood Pb levels in elderly 

2 men (Wright et al., 2004).

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

4 by decreasing absorption of Pb in the GI tract and by decreasing the mobilization of Pb 

5 from bone stores to blood. In experimental studies of adults, absorption of a single dose 

6 of Pb (100-300 μg Pb chloride) was lower when the Pb was ingested together with 

7 calcium carbonate (0.2 g calcium carbonate) than when the Pb was ingested without 

8 additional calcium (Blake and Mann, 1983; Heard and Chamberlain, 1982). A similar 

9 effect of calcium occurs in rats (Barton et al., 1978a). Similarly, an inverse relationship 

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

11 suggesting that children who are calcium-deficient may absorb more Pb than calcium12 replete children (Elias et al., 2007; Schell et al., 2004; Mahaffey et al., 1986; Ziegler et 

13 al., 1978). These observations suggest that calcium and Pb share and may compete for 

14 common binding and transport mechanisms in the small intestine which are regulated in 

15 response to dietary calcium and calcium body stores (Fullmer and Rosen, 1990; Bronner 

16 et al., 1986). However, animal studies have also shown that multiple aspects of Pb 

17 toxicokinetics are affected by calcium nutritional status. For example, feeding rats a 

18 calcium deficient diet is associated with increased Pb absorption, decreased whole body 

19 Pb clearance, and increased volume of distribution of Pb (Aungst and Fung, 1985). These 

20 studies suggest that associations between calcium nutrition and blood Pb that have been 

21 observed in human populations may not be solely attributable to effects of calcium 

22 nutrition on Pb absorption. Other potential mechanisms by which calcium nutrition may 

23 affect blood Pb and Pb biokinetics include effects on bone mineral metabolism and renal 

24 function.

25 Blood Pb concentrations in young children have also been shown to increase in 

26 association with lower dietary Zn levels (Schell et al., 2004). Mechanisms for how Zn 

27 affects blood Pb concentration, i.e., whether it involves changes in absorption or changes 

28 in distribution and/or elimination of Pb, have not been determined.

29 Dissolution of Pb from the soil/mineralogical matrix in the stomach appears to be the 

30 major process that renders soil Pb bioaccessible for absorption in the GI tract. Absorption 

31 of Pb has been shown to vary depending upon the Pb mineralogy and physical 

32 characteristics of the Pb in the soil (e.g., encapsulated or exposed) and size of the Pb33 bearing grains. GI absorption of larger Pb-containing particles (>100 μm) tends to be 

34 lower than smaller particles (Healy et al., 1992; Barltrop and Meek, 1979). Absorption of 

35 Pb in soils and dust has been most extensively studied in the in vivo swine model. Gastric 

36 function of swine is thought to be sufficiently similar to that of humans to justify use of 

37 swine as a model for assessing factors that may affect GI absorption of Pb from soils in 
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1 humans (Juhasz et al., 2009; U.S. EPA, 2007b; Casteel et al., 2006; Casteel et al., 1997; 

2 Weis and Lavelle, 1991). Other practical advantages of the swine model over rodent 

3 models have been described, and include: absence of coprophagia; ease with which Pb 

4 dosing can be administered and controlled; and higher absorption fraction of soluble Pb 

5 (e.g., Pb-acetate) in swine, which is more similar to humans than rats (Smith et al., 

6 2009a). The swine studies measure blood and/or tissue Pb (e.g., kidney, liver, bone) 

7 concentrations following oral dosing of swine with either soil or with a highly water 

8 soluble and fully bioaccessible form of Pb (e.g., Pb-acetate). A comparison of the internal 

9 concentrations of Pb under these two conditions provides a measure of the bioavailability 

10 (i.e., absorption) of Pb in soil relative to that of Pb-acetate, which is typically referred to 

11 as relative bioavailability (RBA). Relative bioavailability measured in the swine assay is 

12 equivalent to the ratio of the absorbed fraction (AF) of ingested dose of soil Pb to that of 

13 water-soluble Pb-acetate (e.g., RBA = AFSoil Pb/AFPb-acetate). 

14 Collectively, published studies conducted in swine have provided 39 estimates of Pb 

15 RBA for 38 different soil or “soil-like” test materials (Bannon et al., 2009; Smith et al., 

16 2009a; Casteel et al., 2006; Marschner et al., 2006). The mean of RBA estimates from 25 

17 soils is 49% (± 29[SD]), median is 51%, and 5th to 95th percentile range is 12 to -89%. 

18 RBA estimates for soils collected from 8 firing ranges were approximately 100%

19 (Bannon et al., 2009). The relatively high RBA for the firing range soils may reflect the 

20 high abundance of relatively un-encapsulated Pb carbonate (30-90% abundance) and Pb 

21 oxide (1-60%) in these soils. Similarly, a soil sample (low Pb concentration) mixed with 

22 a NIST paint standard (55% Pb carbonate, 44% Pb oxide) also had a relatively high 

23 bioavailability (72%) (Casteel et al., 2006). Samples of smelter slag, or soils in which the 

24 dominant source of Pb was smelter slag, had relatively low RBA (14-40%, n = 3), as did 

25 a sample from a mine tailings pile (RBA = 6%), and a sample of finely ground galena 

26 mixed with soil (Casteel et al., 2006).

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

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

29 (>0.75 [75%]) categories (Casteel et al., 2006). Figure 4-3 shows some of the materials 

30 that fall into these three categories. Mineral phases observed in mineralogical wastes can 

31 be expected to change over time (i.e., weathering), which could change the RBA over 

32 time. The above observations in swine are supported by various studies conducted in rats 

33 that have found RBA of Pb in soils to vary considerably and to be less that 100% (Smith 

34 et al., 2009a, 2008; Freeman et al., 1996; Freeman et al., 1994; Freeman et al., 1992).
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Source: Casteel et al. (2006).

Note: based on results from juvenile swine assays.

Figure 4-3 Estimated relative bioavailability (RBA, compared to Pb-acetate) 

of ingested Pb in mineral groups.

1 Drexler and Brattin (2007) developed an in vitro bioaccessibility (IVBA) assay for soil 

2 Pb that utilizes extraction fluid comprised of glycine, deionized water, and hydrochloric 

3 acid at a pH of 1.50 that is combined with sieved test material (<250 μm) for 1 hour. The 

4 assay was tested for predicting in vivo RBA of 18 soil-like test materials that were 

5 assayed in a juvenile swine assay (Casteel et al., 2006). A regression model relating 

6 IVBA and RBA was derived based on these data (Equation 4-1):

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 7

8 for the relationship (weighted for error in the IVBA and RBA estimates) was 0.924 

9 (p <0.001). The IVBA assay reported in Drexler and Brattin (2007) has been identified by 

10 the U.S. EPA as a validated method for predicting RBA of Pb in soils for use in risk 

11 assessment (U.S. EPA, 2007e). A review of soil Pb RBA estimates made using the IVBA 

12 assay described above and Equation 4-1 identified 270 estimates of Pb RBA in soils 

13 obtained from 11 hazardous waste sites. The mean for the site-wide RBA estimates (n = 
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1 11 sites) was 57% (SD 15), median was 63%, and 5th to 95th percentile range was 34 to 

2 71%.

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

4 developed for estimating Pb bioaccessibility without validation against in vivo RBA 

5 measurements made on the same test materials. Comparisons of outcomes among 

6 different in vitro assays applied to the same soil test materials have found considerable 

7 variability in IVBA estimates (Juhasz et al., 2011; Smith et al., 2011; Saikat et al., 2007; 

8 Van de Wiele et al., 2007). This variability has been attributed to differences in assay 

9 conditions, including pH, liquid:soil ratios, inclusion or absence of food material, and 

10 differences in methods used to separate dissolved and particle-bound Pb 

11 (e.g., centrifugation versus filtration). Smith et al. (2011) found that algorithms for 

12 predicting RBA based on two different IVBA assays did not yield similar predictions of 

13 RBA when applied to the same material. Given the dependence of IVBA outcomes on 

14 assay conditions, in vitro assays used to predict in vivo RBA should be evaluated against 

15 in vivo RBA estimates to quantitatively assess uncertainty in RBA predictions (U.S. 

16 EPA, 2007e).

17 Absorption of Pb in house dusts has not been rigorously evaluated quantitatively in 

18 humans or in experimental animal models. The RBA for paint Pb mixed with soil was 

19 reported to be approximately 72% (95% CI: 44, 98) in juvenile swine, suggesting that 

20 paint Pb dust reaching the gastrointestinal tract maybe highly bioavailable (Casteel et al., 

21 2006). The same material yielded a bioaccessibility value (based on IVBA assay) of 75%

22 (Drexler and 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 

24 Equation 4-1 identified 100 estimates of Pb RBA in dusts obtained from two hazardous 

25 waste sites. Mean Pb RBAs for the Herculaneum site were 47% (SD 7, 10 samples) for 

26 indoor dust and 69% (SD 3, 12 samples) for soil. At the Omaha site, mean Pb RBAs were 

27 73% (SD 10, 90 samples) for indoor dust and 70% (SD 10, 45 samples) for soil. Yu et al.

28 (2006) applied an IVBA method to estimate bioaccessibility of Pb in house dust samples 

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

30 reporting to the state department of health of at least on child with a blood Pb 

31 concentration >15 μg/dL and Pb paint dust may have contributed to indoor dust Pb. The 

32 mean IVBA was 64.8% (SD 8.2, age: 52.5 to 77.2 months). 

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

34 evaluated against in vivo RBA estimates for dust samples. Although, expectations are 

35 that a validated IVBA methodology for soil would perform well for predicting RBA of 

36 interior dust, this validation has not actually been experimentally confirmed. Factors that 

37 may affect in vitro predictions of RBA of interior dust Pb could include particle size 
February 2012 4-39 Draft – Do Not Cite or Quote

1 distribution of interior dust Pb and the composition of the dust matrix, which may be 

2 quite different from that of soil.

3 Other estimates of bioaccessibility of Pb in house dusts have been reported, based on 

4 results from in vitro extraction assays that have not been validated for predicting in vivo 

5 bioavailability. Bioaccessibility assays that sequentially extract soil at gastric pH 

6 followed by intestinal pH tend to show higher bioaccessibility of soil and dust Pb when 

7 incubated at gastric conditions (Juhasz et al., 2011; Lu et al., 2011; Smith et al., 2011; 

8 Roussel et al., 2010; Yu et al., 2006). Yu et al. (2006) dissolved Pb dust, obtained from 

9 vacuuming carpet samples into simulated gastric and intestinal acids (also 

10 Section 4.1.3.2). The carpet samples were obtained from homes located in northern New 

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

12 Pb dissolving in simulated gastric acid and 5-32% dissolving in simulated intestinal 

13 acids. In a similar test in the U.K., Turner and Simmonds (2006) observed median Pb 

14 dust concentrations of 178 mg/kg with approximately 80% bioaccessibility in simulated 

15 gastric acid. Jin et al. (2005) observed that bioaccessibility of Pb in soil was proportional 

16 to the soil acidity and organic matter content of the soil.

4.2.2 Distribution

17 A simple conceptual representation of Pb distribution is that it contains a fast turnover 

18 pool, comprising mainly soft tissue, and a slow pool, comprising mainly skeletal tissues

19 (Rabinowitz et al., 1976). The highest soft tissue concentrations in adults occur in liver 

20 and kidney cortex (Gerhardsson et al., 1995; Oldereid et al., 1993; Gerhardsson et al., 

21 1986; Barry, 1975; Gross et al., 1975). Pb in blood (i.e., plasma) exchanges with both of 

22 these compartments.

4.2.2.1 Blood 

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

24 the RBCs (Smith et al., 2002; Manton et al., 2001; Bergdahl et al., 1999; Bergdahl et al., 

25 1998; Hernandez-Avila et al., 1998; Bergdahl et al., 1997a; Schutz et al., 1996). δ26 aminolevulinic acid dehydratase (ALAD) is the primary binding ligand for Pb in 

27 erythrocytes (Bergdahl et al., 1998; Xie et al., 1998; Bergdahl et al., 1997a; Sakai et al., 

28 1982). Two other Pb-binding proteins have been identified in the RBC, a 45 kDa protein 

29 (Kmax 700 μg/dL; Kd 5.5 μg/L) and a smaller protein(s) having a molecular weight 

30 <10 kDa (Bergdahl et al., 1998; Bergdahl et al., 1997a; Bergdahl et al., 1996). Of the 

31 three principal Pb-binding proteins identified in RBCs, ALAD has the strongest affinity 
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1 for Pb (Bergdahl et al., 1998) and appears to dominate the ligand distribution of Pb (35 to 

2 84% of total erythrocyte Pb) at blood Pb levels below 40 μg/dL (Bergdahl et al., 1998; 

3 Bergdahl et al., 1996; Sakai et al., 1982). Pb binding to ALAD is saturable; the binding 

4 capacity was estimated to be ~850 μg/dL RBCs (or ~40 μg/dL whole blood) and the 

5 apparent dissociation constant has been estimated to be ~1.5 μg/L (Bergdahl et al., 1998).

6 Binding to fetal hemoglobin within the RBC has a higher affinity for Pb than adult 

7 hemoglobin. This suggests that erythrocytes of neonates should be able to store more Pb 

8 than infants (Simon et al., 2007). Hematocrit is somewhat higher in the neonate at birth 

9 (51%) than in later infancy (35% at 6 months), which may lead to a decrease in the total 

10 binding capacity of blood over the first 6 months of life that results in a redistribution of 

11 Pb among other tissues (Simon et al., 2007).

12 Saturable binding to RBC proteins contributes to an increase in the plasma/blood Pb ratio 

13 with increasing blood Pb concentration and curvature to the blood Pb–plasma Pb 

14 relationship (Kang et al., 2009; Jin et al., 2008; Barbosa et al., 2006b; Smith et al., 2002; 

15 Manton et al., 2001; Bergdahl et al., 1999; Bergdahl et al., 1998; Bergdahl et al., 1997b; 

16 DeSilva, 1981; Rentschler et al., In Press). An example of this is shown in Figure 4-4. 

17 Saturable binding of Pb to RBC proteins has several important consequences. As blood 

18 Pb increases and the higher affinity binding sites for Pb in RBCs become saturated, a 

19 larger fraction of the blood Pb is available in plasma to distribute to brain and other Pb20 responsive tissues. This change in distribution of Pb contributes to a curvature in the 

21 relationship between Pb intake (at constant absorption fraction) and blood Pb 

22 concentration. Plasma Pb also exhibits faster kinetics. Following exposures of 5 adults 

23 that resulted in relatively high blood Pb concentrations (56-110 μg/dL), the initial (fast24 phase) elimination half-time for plasma Pb (38 ± 20 [SD] days) was approximately half 

25 that of blood (81 ± 25 days) (Rentschler et al., In Press).

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

27 Pb is found in plasma (Marcus, 1985; Manton and Cook, 1984; DeSilva, 1981). However, 

28 as previously noted, plasma Pb may be the more biologically labile and toxicologically 

29 active fraction of the circulating Pb. Approximately 40-75% of Pb in the plasma is bound 

30 to proteins, of which albumin appears to be the dominant ligand (Al-Modhefer et al., 

31 1991; Ong and Lee, 1980a). Pb in serum that is not bound to protein exists largely as 

32 complexes with low molecular weight sulfhydryl compounds (e.g., cysteine, 

33 homocysteine) and other ligands (Al-Modhefer et al., 1991).
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Source: Adapted with permission of Elsevier Publishing and the Finland Institute of Occupational Health, Bergdahl et al. (1999; 

1997b).

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

and children.

1 As shown in Figure 4-4, the limited binding capacity of Pb binding proteins in RBCs 

2 produces a curvilinear relationship between blood and plasma Pb concentration. The 

3 limited binding capacity of RBC binding proteins also confers, or at least contributes, to a 

4 curvilinear relationship between Pb intake and blood Pb concentration. A curvilinear 

5 relationship between Pb intake and blood Pb concentration has been observed in children

6 (Sherlock and Quinn, 1986; Lacey et al., 1985; Ryu et al., 1983). As shown in Figure 4-5, 

7 the relationship becomes pseudo-linear at relatively low daily Pb intakes 

8 (i.e., <10 μg/day/kg) and at blood Pb concentrations <25 μg/dL. 

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
February 2012 4-42 Draft – Do Not Cite or Quote

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

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).

Figure 4-5 Relationship between Pb intake and blood Pb concentration in 

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

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

2 Pb concentrations in a 4-year old child using the ICRP model (Pounds and Leggett, 1998; 

3 ICRP, 1994; Leggett, 1993), see Section 4.3 for a brief description of the ICRP model]. 

4 The abrupt inflection point that occurs at approximately 25 μg/dL blood Pb is an artifact

5 of the numerical approach to simulate the saturation of binding using discontinuous first6 order rate constants for uptake and exit of Pb from the RBC. A continuous function of 

7 binding sites and affinity, using empirical estimates of both parameters, yield a similar 

8 but continuous curvature in the relationship (Bergdahl et al., 1998; O'Flaherty, 1995).

9 Nevertheless, either approach predicts a pseudo-linear relationship at blood Pb 

10 concentrations below approximately 25 μg/dL which, in this model, corresponds to an 

11 intake of approximately 100 μg/day (absorption rate ≈ 30 μg/day) (upper panel). An 

12 important consequence of the limited Pb binding capacity of RBC proteins is that the 

13 plasma Pb concentration will continue to grow at a linear rate above the saturation point 

14 for RBC protein binding. One implication of this is that a larger fraction of the Pb in 

15 blood will become available to distribute to brain and other Pb-responsive tissues as 

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

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

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

Figure 4-6 Simulation of quasi-steady state blood and plasma Pb 

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

ingestion rates.

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

2 kinetics influencing distribution of Pb to tissues. In these studies, the relationship 

3 between the ingested dose of Pb and tissue Pb concentrations (e.g., liver, kidney, bone) 

4 was linear, whereas, the relationship between dose and blood Pb was curvilinear with the 

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

0.6

0.8

1.0

0

10

20

30

40

50

60

0 100 200 300 400

Plasma Pb (μg/dL)

Blood Pb (

μg/dL)

Intake (μg/day)

Blood Plasma
February 2012 4-44 Draft – Do Not Cite or Quote

1 slope decreasing as the dose increased (Casteel et al., 2006). Saturable binding of Pb to 

2 RBC proteins also contributes to a curvilinear relationship between urinary Pb excretion 

3 and plasma Pb concentration (Section 4.2.3) (Besser et al., 2008; Bergdahl et al., 1997b).

4.2.2.2 Bone

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

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

6 body burden in children (Barry, 1975). Bone is comprised of two main types, cortical (or 

7 compact) and trabecular (or spongy or cancellous). The proportion of cortical to 

8 trabecular bone in the human body varies by age, but on average is about 80 to 20 percent

9 (O'Flaherty, 1998; Leggett, 1993; ICRP, 1973).

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

11 =0.19 and 0.23 hours for trabecular and cortical bone, respectively) (Leggett, 1993).

12 Some Pb diffuses from the bone surface to deeper bone regions (adult t1/2=150 days) 

13 where it is relatively inert (in adults) and part of a “nonexchangeable” pool of Pb in bone

14 (Leggett, 1993).

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

16 osteoclasts, osteocytes) and exchanges with proteins and minerals in the extracellular 

17 matrix (Pounds et al., 1991). Pb forms highly stable complexes with phosphate and can 

18 replace calcium in the calcium-phosphate salt, hydroxyapatite, which comprises the 

19 primary crystalline matrix of bone (Meirer et al., 2011; Brès et al., 1986; Miyake, 1986; 

20 Verbeeck et al., 1981). Several intracellular kinetic pools of Pb have been described in 

21 isolated cultures of osteoblasts and osteoclasts which appear to reflect physiological 

22 compartmentalization within the cell, including membranes, mitochondria, soluble 

23 intracellular binding proteins, mineralized Pb (i.e., hydroxyapatite) and inclusion bodies

24 (Long et al., 1990; Pounds and Rosen, 1986; Rosen, 1983). Approximately 70-80% of Pb 

25 taken up into isolated primary cultures of osteoblasts or osteocytes is associated with 

26 mitochondria and mineralized Pb (Pounds et al., 1991).

27 Pb accumulates in bone regions having the most active calcification at the time of 

28 exposure. Pb accumulation is thought to occur predominantly in trabecular bone during 

29 childhood and in both cortical and trabecular bone in adulthood (Aufderheide and 

30 Wittmers, 1992). Early Pb uptake in children is greater in trabecular bone due to its larger 

31 surface area and higher metabolic rate. With continued exposure, Pb concentrations in 

32 bone may increase with age throughout the lifetime beginning in childhood, indicative of 

33 a relatively slow turnover of Pb in adult bone (Park et al., 2009a; Barry and Connolly, 

34 1981; Barry, 1975; Gross et al., 1975; Schroeder and Tipton, 1968). The cortical and 
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1 trabecular bones have distinct rates of turnover and Pb release. For example, tibia has a 

2 turnover rate of about 2% per year whereas trabecular bone has a turnover rate of more 

3 than 8% per year in adults (Rabinowitz, 1991).

4 A high bone formation rate in early childhood results in the rapid uptake of circulating Pb 

5 into mineralizing bone; however, bone Pb is also recycled to other tissue compartments 

6 or excreted in accordance with a high bone resorption rate (O'Flaherty, 1995). Thus, most 

7 of the Pb acquired early in life is not permanently fixed in the bone (60-65%)

8 (O'Flaherty, 1995; Leggett, 1993; ICRP, 1973). However, some Pb accumulated in bone 

9 does persist into later life. McNeill et al. (2000) compared tibia Pb levels and cumulative 

10 blood Pb indices in a population of 19- to 29-year-olds who had been highly exposed to 

11 Pb in childhood from the Bunker Hill, Idaho smelter; they concluded that Pb from 

12 exposure in early childhood had persisted in the bone matrix until adulthood.

13 A key factor affecting Pb uptake into bone is the fraction of bone surface in trabecular 

14 and cortical bone adjacent to active bone marrow. Of the total bone surface against red 

15 marrow, 76% is trabecular and 24% is cortical endosteal (Salmon et al., 1999). The 

16 fraction of total bone marrow that is red and active decreases from 100% at birth to about 

17 32% in adulthood (Cristy, 1981). However, bone marrow has much lower Pb 

18 concentrations than bone matrix (Skerfving et al., 1983).

4.2.2.3 Soft Tissues

19 Most of the Pb in soft tissue is in liver and kidney (Gerhardsson et al., 1995; Oldereid et 

20 al., 1993; Gerhardsson et al., 1986; Barry, 1975; Gross et al., 1975). Presumably, the Pb 

21 in these soft tissues (i.e., kidney, liver, and brain) exists predominantly bound to protein. 

22 High affinity cytosolic Pb-binding proteins have been identified in rat kidney and brain

23 (DuVal and Fowler, 1989; Fowler, 1989). The Pb-binding proteins in rat are cleavage 

24 products of α2μ globulin, a member of the protein superfamily known as retinol-binding 

25 proteins that are generally observed only in male rats (Fowler and DuVal, 1991). Other 

26 high-affinity Pb-binding proteins (Kd ~14 nM) have been isolated in human kidney, two 

27 of which have been identified as a 5 kDa peptide, thymosin 4 and a 9 kDa peptide, acyl28 CoA binding protein (Smith et al., 1998). Pb also binds to metallothionein, but does not 

29 appear to be a significant inducer of the protein in comparison with the inducers Cd and 

30 Zn (Waalkes and Klaassen, 1985; Eaton et al., 1980).

31 The liver and kidneys rapidly accumulate systemic Pb (t1/2=0.21 and 0.41 hours, 

32 respectively), which amounts to 10-15% and 15-20% of intravenously injected Pb, 

33 respectively (Leggett, 1993). A linear relationship in dose-tissue Pb concentrations for 

34 kidney and liver has been demonstrated in swine, dogs, and rats (Smith et al., 2008; 
February 2012 4-46 Draft – Do Not Cite or Quote

1 Casteel et al., 2006; Casteel et al., 1997; Azar et al., 1973). In contrast to Pb in bone, 

2 which accumulates Pb with continued exposure in adulthood, concentrations in soft 

3 tissues (e.g., liver and kidney) are relatively constant in adults (Treble and Thompson, 

4 1997; Barry, 1975), reflecting a faster turnover of Pb in soft tissue relative to bone.

4.2.2.4 Fetus

5 Evidence for maternal-to-fetal transfer of Pb in humans is derived from cord blood to 

6 maternal blood Pb ratios. Group mean ratios range from about 0.7 to 1.0 at the time of 

7 delivery for mean maternal blood Pb levels ranging from 1.7 to 8.6 μg/dL (Amaral et al., 

8 2010; Kordas et al., 2009; Patel and Prabhu, 2009; Carbone et al., 1998; Goyer, 1990; 

9 Graziano et al., 1990). In a study of mothers having blood Pb levels of <14 μg/dL, the 

10 ratio of cord blood Pb to maternal blood Pb decreased with decreasing maternal blood Pb 

11 (r=0.82) (Carbone et al., 1998). In addition, the similarity of isotopic ratios in maternal 

12 blood and in blood and urine of newly-born infants provide further evidence for placental 

13 transfer of Pb to the fetus (Gulson et al., 1999).

14 Transplacental transfer of Pb may be facilitated by an increase in the plasma/blood Pb 

15 concentration ratio during pregnancy (Montenegro et al., 2008; Lamadrid-Figueroa et al., 

16 2006). Maternal-to-fetal transfer of Pb appears to be related partly to the mobilization of 

17 Pb from the maternal skeleton. Evidence for transfer of maternal bone Pb to the fetus has 

18 been provided by stable Pb isotope studies in cynomolgus monkeys exposed during 

19 pregnancy. Approximately 7-39% of the maternal Pb burden transferred to the fetus was 

20 derived from the maternal skeleton, with the remainder derived from contemporaneous 

21 exposure (O'Flaherty, 1998; Franklin et al., 1997).

4.2.2.5 Organic Lead 

22 Information on the distribution of Pb in humans following exposures to organic Pb is 

23 extremely limited. However, as reported in the 2006 Pb AQCD, the available evidence 

24 demonstrates near complete absorption following inhalation of tetraalkyl Pb vapor and 

25 subsequent transformation to trialkyl Pb metabolites. One hour following brief inhalation 

exposures to 203Pb tetraethyl or tetramethyl Pb (1 mg/m3), ~50% of the 203 26 Pb body burden 

was associated with liver and 5% with kidney; the remaining 203 27 Pb was widely distributed 

throughout the body (Heard et al., 1979). The kinetics of 203 28 Pb in blood showed an initial 

29 declining phase during the first 4 hours (tetramethyl Pb) or 10 hours (tetraethyl Pb) after 

30 the exposure, followed by a reappearance of radioactivity back into the blood after ~20 

31 hours. The high level of radioactivity initially in the plasma indicates the presence of 
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1 tetraalkyl/trialkyl Pb. The subsequent rise in blood radioactivity, however, probably 

2 represents water-soluble inorganic Pb and trialkyl and dialkyl Pb compounds that were 

3 formed from the metabolic conversion of the volatile parent compounds (Heard et al., 

4 1979).

5 Alkyl Pb compounds undergo oxidative dealkylation catalyzed by cytochrome P450 in 

6 liver and, possibly, in other tissues. Trialkyl Pb metabolites have been found in the liver, 

7 kidney, and brain following exposure to the tetraalkyl compounds in workers

8 (Bolanowska et al., 1967); these metabolites have also been detected in brain tissue of 

9 nonoccupational subjects (Nielsen et al., 1978).

4.2.3 Elimination

10 The rapid-phase (30-40 days) of Pb excretion amounts to 50-60% of the absorbed 

11 fraction (Chamberlain et al., 1978; Rabinowitz et al., 1976; Kehoe, 1961a, b, c).

12 Absorbed Pb is excreted primarily in urine and feces, with sweat, saliva, hair, nails, and 

13 breast milk being minor routes of excretion (Kehoe, 1987; Chamberlain et al., 1978; 

14 Rabinowitz et al., 1976; Griffin et al., 1975; Hursh et al., 1969; Hursh and Suomela, 

15 1968).

16 Approximately 30% of intravenously injected Pb in humans (40-50% in beagles and 

17 baboons) is excreted via urine and feces during the first 20 days following administration

18 (Leggett, 1993). The kinetics of urinary excretion following a single dose of Pb is similar 

19 to that of blood (Chamberlain et al., 1978), likely due to the fact that Pb in urine derives 

20 largely from Pb in plasma. Evidence for this is the observation that urinary Pb excretion 

21 is strongly correlated with the rate of glomerular filtration of Pb (Araki et al., 1986) and 

22 plasma Pb concentration (Bergdahl et al., 1997b; Rentschler et al., In Press)

23 (i.e., glomerular filtration rate × plasma Pb concentration), and both relationships are 

24 linear. While the relationship between urinary Pb excretion and plasma Pb concentration 

25 is linear, the plasma Pb relationship to blood Pb concentration is curvilinear (as described 

26 in Section 4.2.2.1 and demonstrated in Figure 4-6). This relationship contributes to an 

27 increase in the renal clearance of Pb from blood with increasing blood Pb concentrations

28 (Chamberlain, 1983). Similarly, a linear relationship between plasma Pb concentration 

29 and urinary excretion rate predicts a linear relationship between Pb intake (at constant 

30 absorption fraction) and urinary Pb excretion rate, whereas the relationship with blood Pb 

31 concentration would be expected to be curvilinear (Section 4.3.7). 

32 Estimates of urinary filtration of Pb from serum (or plasma) range from 13-22 L/day, 

33 with a mean of 18 L/day (Araki et al., 1986; Manton and Cook, 1984; Manton and 

34 Malloy, 1983; Chamberlain et al., 1978), which corresponds to half-time for transfer of 
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1 Pb from plasma to urine of 0.1-0.16 days for a 70-kg adult who has a plasma volume of 

2 ~3 L. The rate of urinary excretion of Pb was less than the rate of glomerular filtration of 

3 ultrafilterable Pb, suggesting that urinary Pb is the result of incomplete renal tubular 

4 re-absorption of Pb in the glomerular filtrate (Araki et al., 1986); although, net tubular 

5 secretion of Pb has been demonstrated in animals (Victery et al., 1979; Vander et al., 

6 1977). On the other hand, estimates of blood-to-urine clearance range from 

7 0.03-0.3 L/day with a mean of 0.18 L/day (Diamond, 1992; Araki et al., 1990; Berger et 

8 al., 1990; Koster et al., 1989; Manton and Malloy, 1983; Ryu et al., 1983; Chamberlain et 

9 al., 1978; Rabinowitz et al., 1973), consistent with a plasma Pb to blood Pb concentration 

10 ratio of ~0.005–0.01 L/day (Klotzback et al., 2003). Based on the above differences, 

11 urinary excretion of Pb can be expected to reflect the concentration of Pb in plasma and 

12 variables that affect delivery of Pb from plasma to urine (e.g., glomerular filtration and 

13 other transfer processes in the kidney).

14 The value for fecal:urinary excretion ratio (~0.5) was observed during days 2-14 

15 following intravenous injection of Pb in humans (Chamberlain et al., 1978; Booker et al., 

16 1969; Hursh et al., 1969). This ratio is slightly higher (0.7-0.8) with inhalation of 

17 submicron Pb-bearing PM due to ciliary clearance and subsequent ingestion. The transfer 

18 of Pb from blood plasma to the small intestine by biliary secretion in the liver is rapid 

19 (adult t1/2 = 10 days), and accounts for 70% of the total plasma clearance (O'Flaherty, 

20 1995).

Organic Lead 

21 Pb absorbed after inhalation of tetraethyl and tetramethyl Pb is excreted in exhaled air, 

22 urine, and feces (Heard et al., 1979). Fecal:urinary excretion ratios were 1.8 following 

23 exposure to tetraethyl Pb and 1.0 following exposure to tetramethyl Pb (Heard et al., 

24 1979). Occupational monitoring studies of workers exposed to tetraethyl Pb showed that 

25 tetraethyl Pb is excreted in the urine as diethyl Pb, ethyl Pb, and inorganic Pb (Vural and 

26 Duydu, 1995; Zhang et al., 1994; Turlakiewicz and Chmielnicka, 1985).

4.3 Lead Biomarkers

27 This section describes the biological measurements of Pb and their interpretation as 

28 indicators of exposure or body burden. 

29 For any health endpoint of interest, the most useful biomarker of exposure is one that 

30 provides information about the Pb dose at the critical target organ and, moreover, reflects 

31 the exposure averaging time that is appropriate to the underlying pathogenetic processes 
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1 (e.g., instantaneous, cumulative over lifetime, or cumulative over a circumscribed age 

2 range). In recent studies of Pb and health, the exposure biomarkers most frequently used 

3 are Pb in blood and bone. For outcomes other than those relating to hematopoiesis and 

4 bone health, these biomarkers provide information about Pb dose that is some distance 

5 from the target organ. For example, given that the central nervous system is considered 

6 the critical target organ for childhood Pb toxicity, it would be most helpful to be able to 

7 measure, in vivo, the Pb concentrations at the cellular site(s) of action in the brain. 

8 However, because such measurements are not currently feasible, investigators must rely 

9 on measurements of Pb in the more readily accessible but peripheral tissues. The 

10 relationship between brain Pb and Pb in each of these surrogate tissues is still poorly 

11 understood, although the pharmacokinetics clearly differs among these compartments. 

12 As an exposure biomarker, blood Pb concentration has other limitations. Only about 5% 

13 of an individual’s total body Pb burden resides in blood. Furthermore, blood consists of 

14 several subcompartments. More than 90% of Pb in whole blood is bound to red cell 

15 proteins such as hemoglobin, with the balance in plasma. From a toxicological 

16 perspective, the unbound fraction is likely to be the most important subcompartment of 

17 blood Pb because it distributes into soft tissues. The concentration of Pb in plasma is 

18 much lower than in whole blood (<1%). The greater relative abundance of Pb in whole 

19 blood makes its measurement much easier (and more affordable) than measurement of Pb 

20 in plasma. The use of whole blood Pb as a surrogate for plasma Pb could be justified if 

21 the ratio of whole blood Pb to plasma Pb were well characterized, but this is not so. At 

22 least some studies suggest that it varies several-fold among individuals with the same 

23 blood Pb level. Moreover, binding Pb in red blood cells is limited, so the ratio of blood 

24 Pb to plasma Pb would be expected to be nonlinear. Thus, interpreting whole blood Pb 

25 level as a proxy for plasma Pb level, which, itself, is a proxy for brain Pb level, will result 

26 in some exposure misclassification. 

27 Another limitation of blood Pb as an exposure biomarker is that the kinetics of Pb in 

28 blood is relatively fast compared to the kinetics of Pb in bone, and therefore, of the whole 

29 body burden. Thus, a high blood Pb concentration measured at any given time does not 

30 necessarily indicate a high body Pb burden. Similarly, individuals who have the same 

31 blood Pb level will not necessarily have similar body burdens or exposure histories. The 

32 rate at which blood Pb changes with time/age depends on exposure history due to re33 equilibration of Pb stored in the various body pools.

34 The development of X-ray-fluorescence (XRF) methods for measuring Pb in mineralized 

35 tissues offers another approach for characterization and reconstruction of exposure 

36 history. Such tissues are long-term Pb storage sites, with a half-life measured in decades 
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1 and contain ~90% of the total body Pb burden in adults and 70% in children. Thus, bone 

2 Pb reflects a long exposure averaging time.

3 Mechanistic models are used throughout the section as a means to describe basic 

4 concepts that derive from the wealth of information on Pb toxicokinetics. Although 

5 predictions from models are inherently uncertain, models can serve to illustrate expected 

6 interrelationships between Pb intake and tissue distribution that are important in 

7 interpreting human clinical and epidemiologic studies. Thus, models serve as the only 

8 means we have for synthesizing our extensive, but incomplete, knowledge of Pb 

9 biokinetics into a holistic representation of Pb biokinetics. Furthermore, models can also 

10 be used to make predictions about biokinetics relationships that have not been thoroughly 

11 evaluated in experiments or epidemiologic studies. In this way, models can serve as 

12 heuristic tools for shaping data collection to improve our understanding of Pb biokinetics.

13 Mechanistic toxicokinetics models can make predictions about hypothetical populations 

14 and exposure scenarios. When a model is run as a single simulation, the output represents 

15 average outcomes from what is in reality a distribution of possible outcomes that would 

16 be expected in the population (or in any single individual) where intra-individual and 

17 inter-individual variability in exposure and toxicokinetics exist. More realistic predictions 

18 for the population can be developed by running a series of model simulations in which 

19 ranges (i.e., distributions) of parameter values are considered that may better represent 

20 the population of interest. In this section, only single simulations are used to demonstrate 

21 relationships between various biomarkers (e.g., blood Pb and bone Pb) that would apply 

22 to a population having “typical” or “average” exposure and toxicokinetics. These single 

23 simulations are used for illustrative purposes to describe general concepts and patterns. 

24 Variability would be expected in real populations. 

25 Numerous mechanistic models of Pb biokinetics in humans have been proposed, and 

26 these are described in the 2006 Pb AQCD (U.S. EPA, 2006b) and in the supporting 

27 literature cited in that report. In this section, for simplicity and for internal consistency, 

28 we have limited the discussion to predictions from a single model, the ICRP Pb 

29 biokinetics model (Pounds and Leggett, 1998; ICRP, 1994; Leggett, 1993). The ICRP 

30 model consists of a systemic biokinetics model (Leggett, 1993) and a human respiratory 

31 tract model (ICRP, 1994). The Leggett model simulates age-dependent kinetics of tissue 

32 distribution and excretion of lead ingestion and inhalation intakes. This model was 

33 originally developed for the purpose of supporting radiation dosimetry predictions and it 

34 has been used to develop cancer risk coefficients for internal radiation exposures to lead 

35 and other alkaline earth elements that have biokinetics similar to those of calcium (ICRP, 

36 1993). Although the ICRP model has not been validated by U.S. EPA as a regulatory 

37 model for lead risk assessment, it has been applied in Pb risk assessment (Abrahams et 
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1 al., 2006; Lorenzana et al., 2005; Khoury and Diamond, 2003). Portions of the model 

2 have been incorporated into an AALM that is being developed by EPA (2005a). In 

3 addition to the above considerations regarding previous applications of the ICRP model, 

4 the model was selected for use in the ISA because it has several useful features for 

5 predicting exposure-body burden relationships. The model simulates blood Pb and tissue 

6 Pb concentration dynamics associated with the uptake and elimination phases of 

7 exposures of ≥ 1 day in duration; and it simulates age-dependent and particle size8 dependent deposition and clearance of inhaled lead in the respiratory tract. These types of 

9 simulations can only be approximated with the U.S. EPA IEUBK Model for lead in 

10 children because it simulates exposures in time steps of 1 year (i.e., age-year average 

11 exposures); lumps the simulation of deposition, mechanical clearance, and absorption of 

12 inhaled lead into a single absorption term representing the combined processes of 

13 gastrointestinal and respiratory tract absorption of inhaled lead; simulates steady state 

14 blood Pb concentrations and was does not allow access to the underlying simulations of 

15 tissue Pb concentrations which serve as intermediate variables in the model for predicting 

16 steady state blood Pb concentrations. Other models have been developed that allow 

17 simulations of tissue Pb concentrations (e.g., O'Flaherty, 1995; Leggett, 1993) and 

18 comparisons of these models have been previously described (Maddaloni et al., 2005).

19 Pb biokinetics in adolescents is poorly characterized by all existing Pb biokinetics 

20 models. Individuals undergo rapid changes in sexual development, growth, food and 

21 water intake, bone growth and turnover, behavior, etc. during adolescence. There is a 

22 paucity of experimental measurements of Pb biomarkers during this time developmental 

23 window. The individual biological and kinetic parameters for adolescents are largely 

24 interpolated rather than based on solid experimental and toxicological measurements. 

25 These deficiencies limit the validity of model predictions in this age group.

4.3.1 Bone Lead Measurements

26 For Pb measurements in bone, the most commonly examined bones are the tibia, 

27 calcaneus, patella, and finger bone. For cortical bone, the midpoint of the tibia is 

28 measured. For trabecular bone, both the patella and calcaneus are measured. The tibia 

29 consists of more than 95% cortical bone, the calcaneus and patella comprise more than 

30 95% trabecular bone, and finger bone is a mixed cortical and trabecular bone although 

31 the second phalanx is dominantly cortical. Recent studies favor measurement of the 

32 patella for estimating trabecular bone Pb, because it has more bone mass and may afford 

33 better measurement precision than the calcaneus. 
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1 Bone Pb measurements are typically expressed in units of Pb/g bone mineral. This 

2 convention may potentially introduce variability into the bone Pb measurements related 

3 to variation in bone density. Typically, potential associations between bone density and 

4 bone Pb concentration are not evaluated in epidemiologic studies (Theppeang et al., 

5 2008a; Hu et al., 2007a). An important consequence of expressing bone Pb measures 

6 relative to bone mineral content is that lower bone mineral density is associated with 

7 greater measurement uncertainty in bone Pb. This can have important implications for 

8 studies in older women for whom low bone mineral density is more common than in 

9 other populations including men and younger adults.

10 Methods of direct analysis of bone tissue samples include flame atomic absorption 

11 spectrometry (AAS), anode stripping voltammetry (ASV), inductively coupled plasma 

12 atomic emission spectroscopy (ICP-AES), inductively coupled plasma mass spectrometry 

13 (ICP-MS), laser ablation inductively coupled plasma mass spectrometry (LA-ICP-MS), 

14 thermal ionization mass spectrometry (TIMS), synchrotron radiation induced X-ray 

15 emission (SRIXE), particle induced X-ray emission (PIXE), and X-ray fluorescence 

16 (XRF). Non-invasive, in vivo measurements of bone Pb is achieved with XRF. The 

17 upsurge in popularity of the XRF method has paralleled a decline in the use of the other 

18 methods. More information on the precision, accuracy, and variability in bone Pb 

19 measurements can be found in the 2006 Pb AQCD (U.S. EPA, 2006b).

20 Two main approaches for XRF measurements have been used to measure Pb 

21 concentrations in bone, the K-shell and L-shell methods. The K-shell method is the most 

22 widely used, as there have been relatively few developments in L-shell devices since the 

23 early 1990s. However, Nie et al. (2011) recently reported on the use of a new portable L24 shell device for human in vivo Pb measurements. Advances in L-shell device technology 

25 resulted in much higher sensitivity than previous L-shell devices. The new L-shell device 

26 showed sensitivity similar to that of K-shell methods (detection limit was approximately 

27 8 μg/g bone mineral with 2 mm of soft tissue overlay targeted bone) and a high 

28 correlation with results obtained from K-shell methods (intraclass correlation = 0.65). 

29 Behinaen et al. (2011) described application of a 4-detector system (“clover leaf array”) 

30 for the K-shell method that provided higher precision and lower minimum detection 

31 limits (MDL) for tibia and calcaneus Pb measurements (3.25 and 4.78 μg/g bone mineral, 

32 respectively) compared to measurements made with single detectors (8-12 μg/g and 

33 14-15 μg/g, respectively). 

34 Since 1986, several investigators have reported refinements to hardware and software to 

35 improve the precision and accuracy of XRF measurements and there have been a number 

36 of investigations into the precision, accuracy and variability in XRF measurements 

37 [e.g., (Todd et al., 2002; Todd et al., 2001; Aro et al., 2000; Todd et al., 2000)]. Todd 
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1 et al. (2000) provided a detailed discussion of factors that influence the variability and 

2 measurement uncertainty, including repositioning, sample measurement duration, 

3 overlying tissue, operator expertise, detector resolution, and changes to measurement 

4 process over time. Some of these aspects were also discussed by Hu et al. (1995). From 

5 their cadaver and in vivo measurements, Todd et al. (2000) concluded that the uncertainty 

6 in an individual measurement was an underestimate of the standard deviation of replicate 

7 measurements, suggesting a methodological deficiency probably shared by most current 

109 8 Cd-based K-shell XRF Pb measurement systems. In examining the reproducibility of 

9 the bone Pb measurements over a 41⁄2 month period, Todd et al. found the average 

10 difference between the XRF results from short term and longer term measurements was 

11 1.2 μg/g, indicating only a small amount of variability in the XRF results over a sustained 

12 period of time.

13 In the epidemiologic literature, XRF bone Pb data have typically been reported in two 

14 ways: one that involves a methodological approach to assessing the minimum detection 

15 limit and the other termed an epidemiologic approach by Rosen and Pounds (1998). In 

16 the former approach, a minimum detection limit is defined using various methods, 

17 including two or three times the square root of the background counts; one, two, or three 

18 times the SD of the background; or two times the observed median error. This approach 

19 relies upon the minimum detection limit to define a quantitative estimate that is of 

20 sufficient precision to be included in the statistical analysis, as demonstrated by Bellinger 

21 et al. (1994a), Gerhardsson et al. (1993), and Christoffersson et al. (1986).

22 With the epidemiologic approach, all values are used (including negative values) to 

23 determine the minimum detection limit of an instrument that results in extremely low 

24 detection limits. Rosen and Pounds (1998) noted that this approach yields population 

25 bone Pb averages that were artificially low. However, not including values that are 

26 negative or below the detection limit, or assigning these values a fixed number is also of 

27 concern. Using the epidemiologic approach of retaining all point estimates of measured 

28 bone Pb concentrations provided the least amount of bias and the greatest efficiency in

29 comparing the mean or median levels of bone Pb of different populations (Kim et al., 

30 1995).

4.3.2 Blood Lead Measurements

31 Analytical methods for measuring Pb in blood include AAS, graphite furnace atomic 

32 absorption spectrometry (GFAAS), ASV, ICP-AES, and ICP-MS. GFAAS and ASV are 

33 generally considered to be the methods of choice (Flegal and Smith, 1995). Limits of 

34 detection for Pb using AAS are on the order of 5-10 μg/dL for flame AAS measurements 
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1 and approximately 0.1 μg/dL for flameless AAS measurements (Flegal and Smith, 1995; 

2 NIOSH, 1994). A detection limit of 0.005 μg/dL has been achieved for Pb in blood 

3 samples analyzed by GFAAS.

4 For measurement of Pb in plasma, ICP-MS provides sufficient sensitivity (Schutz et al., 

5 1996). While the technique has been applied to assessing Pb exposures in adults, ICP-MS 

6 has not received widespread use in epidemiologic studies.

7 The primary binding ligand for Pb in RBC, ALAD, is encoded by a single gene in 

8 humans that is polymorphic in two alleles (ALAD1 and ALAD2) (Scinicariello et al., 

9 2007). Since the ALAD1 and ALAD 2 alleles can be co-dominantly expressed, 3 

10 different genotypes (ALAD 1-1, ALAD 1-2, and ALAD 2-2) are possible. The ALAD 

11 1-1 genotype is the most common. Scinicariello et al. (2010) tested genotypes in civilian, 

12 non-institutionalized U.S. individuals that participated as part of NHANES III from 

13 1988–1994 and found that 15.6% of non-Hispanic whites, 2.6% non-Hispanic blacks, and 

14 8.8% Mexican Americans carried the ALAD2 allele. 

15 The 2006 Pb AQCD document reports that many studies have shown that, with similar 

16 exposures to Pb, individuals with the ALAD-2 allele have higher blood Pb levels than 

17 those without (Kim et al., 2004; Pérez-Bravo et al., 2004; Bergdahl et al., 1997b; Smith et 

18 al., 1995a; Wetmur, 1994; Wetmur et al., 1991a; Astrin et al., 1987). More recent meta 

19 analyses provide further support for ALAD2 carriers having higher blood Pb levels than 

20 ALAD1-1 homozygotes (Scinicariello et al., 2007; Zhao et al., 2007). The mechanism for 

21 this association may be higher Pb binding affinity of ALAD2. Although, this 

22 interpretation would be consistent with the structural differences that result in greater 

23 electronegativity of ALAD1 compared to ALAD2 (Wetmur, 1994; Wetmur et al., 

1991b), measurements of Pb binding affinity to ALAD1 and ALAD2 (i.e., Pb2+ 24

displacement of Zn2+ 25 binding to recombinant ALAD1 and ALAD2) have not revealed 

26 differences in Pb binding affinity (Jaffe et al., 2000). In a meta-analysis of 24 studies, 

27 Scinicariello et al. (2007), observed the greatest differences for ALAD2 compared to 

28 ALAD1 in highly exposed adults with little difference among environmentally-exposed 

29 adults; large differences were also observed for children at low exposures. However, 

30 there are few studies that evaluated children and the largest study contributing to the meta 

31 analysis may have been influenced by selection bias (Scinicariello et al., 2007).

32 Individual studies find similar results in occupationally-exposed adults, with blood Pb 

33 levels being higher in individuals with ALAD2 alleles (Miyaki et al., 2009; Shaik and 

34 Jamil, 2009). A subsequent meta analysis of adult data from NHANES III did not find 

35 any differences in blood Pb level between all carriers of either the ALAD 1-1 or ALAD 

36 1-2/2-2 allele (Scinicariello et al., 2010). Other studies provide further support for no 

37 blood Pb differences among ALAD1 and ALAD2 carriers (Sobin et al., 2009; Rabstein et 
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1 al., 2008; Montenegro et al., 2006; Wananukul et al., 2006) or lower blood Pb levels for 

2 individuals with ALAD 1-2/2-2 (Krieg et al., 2009; Chia et al., 2006).

3 Genetic polymorphism in the gene that encodes for peptide transporter 2 (PEPT2) has 

4 been associated with variability in blood Pb concentrations in children (Sobin et al., 

5 2009). PEPT2 expression in the brain and renal proximal tubule has been associated with 

6 transport of di- and tri-peptides and may function in the transport of δ-ALA into brain 

7 and renal tubular re-absorption of peptides. The PRPT2*2 polymorphism was associated 

8 with increased blood Pb concentrations in a sample of 116 children of Mexican9 American/Hispanic (age 4-12 years, mean blood Pb concentration 3-6 μg/dL).

10 Analyses of serial blood Pb concentrations measured in longitudinal epidemiologic 

11 studies found relatively strong correlations (e.g., r = 0.5-0.8) between individual blood 

12 Pb concentrations measured after 6-12 months of age (Schnaas et al., 2000; Dietrich et 

13 al., 1993b; McMichael et al., 1988; Otto et al., 1985a; Rabinowitz et al., 1984). These 

14 observations suggest that, in general, exposure characteristics of an individual child 

15 (e.g., exposure levels and/or exposure behaviors) tend to be relatively constant across 

16 age. However, a single blood Pb measurement may not distinguish between a history of 

17 long-term lower-level Pb exposure from a history that includes higher acute exposures

18 (Mushak, 1998). This concept is illustrated in Figure 4-7. Two hypothetical children are 

19 simulated. Child A has a relatively constant Pb intake from birth, whereas Child B has 

20 the same Pb intake as Child A for the first two years of life, then a 1-year elevated intake 

21 beginning at age 24 months (Figure 4-7, upper panel) that returns to the same intake as 

22 Child A at 36 months. The absorption fraction is assumed to be the same for both 

23 children. Blood Pb samples 1 and 5 for Child A and B, or 2 and 4 for Child B, will yield 

24 similar blood Pb concentrations (~3 or 10 μg/dL, respectively), yet the exposure contexts 

25 for these samples are very different. Two samples (e.g., 1 and 2, or 4 and 5), at a 

26 minimum, are needed to ascertain if the blood Pb concentration is changing over time. 

27 The rate of change can provide information about the magnitude of change in exposure, 

28 but not necessarily about the time history of the change (Figure 4-7, lower panel). Time29 integrated measurements of Pb concentration may provide a means for accounting for 

30 some of these factors and, thereby, provide a better measure of long-term Pb exposure.
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Note: 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).

Figure 4-7 Simulation of temporal relationships between Pb exposure and 

blood Pb concentration in children.
February 2012 4-57 Draft – Do Not Cite or Quote

4.3.3 Urine Lead Measurements

1 Standard methods that have been reported for urine Pb analysis are, in general, the same 

2 as those analyses noted for determination of Pb in blood. Reported detection limits are 

3 ~50 μg/L for AAS, 5-10 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) 

5 and the urine flow rate. Urine flow rate requires collection of a timed urine sample, which 

6 is often problematic in epidemiologic studies. Collection of untimed (“spot”) urine 

7 samples, a common alternative to timed samples, requires adjustment of the Pb 

8 measurement in urine to account for variation in urine flow (Diamond, 1988). Several 

9 approaches to this adjustment have been explored, including adjusting the measured urine 

10 Pb concentration by the urine creatinine concentration, urine osmolality, or specific 

11 gravity (Fukui et al., 1999; Araki et al., 1990). Urine flow rate can vary by a factor or 

12 more than 10, depending on the state of hydration and other factors that affect glomerular 

13 filtration rate and renal tubular reabsorption of the glomerular filtrate. All of these factors 

14 can be affected by Pb exposure at levels that produce nephrotoxicity (i.e., decreased 

15 glomerular filtration rate, impaired renal tubular transport function). Therefore, urine Pb 

16 concentration measurements provide little reliable information about exposure (or Pb 

17 body burden), unless they can be adjusted to account for unmeasured variability in urine 

18 flow rate (Araki et al., 1990).

19 Urinary Pb concentration reflects, mainly, the concentration of Pb in the blood. As such, 

20 urinary concentrations by reflect both recent and past exposures to Pb (see Section 4.3.5). 

21 A single urinary Pb measurement cannot distinguish between a long-term low level of 

22 exposure or a higher acute exposure. Urinary Pb measurements would be expected to 

23 correlate with concurrent blood Pb (see Section 4.3.6 for additional discussion of the 

24 relationship between blood and urine Pb). Chiang et al. (2008) reported a significant, but 

25 relatively weak correlation between urinary Pb levels (μg/dg creatinine) and individual 

26 Pb intakes (μg/day) estimated in a group of 10- to 12-year-old children (β: 0.053, R = 

27 0.320, p = 0.02, n = 57). A contributing factor to the relatively weak correlation may have 

28 been the temporal displacement between the urine sampling and measurements used to 

29 estimate intake, which may have been as long as 6 months for some children. 

30 Thus, a single urine Pb measurement, or a series of measurements taken over short-time 

31 span, is likely a relatively poor index of Pb body burden for the same reasons that blood 

32 Pb is not a good indicator of body burden. On the other hand, long-term average 

33 measurements of urinary Pb can be expected to be a better index of body burden (Figure 

34 4-8).
February 2012 4-58 Draft – Do Not Cite or Quote

Note: 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).

Figure 4-8 Simulation of relationship between urinary Pb excretion and body 

burden in adults.

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)

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Urine

Body

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20

0

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Time-Averaged Urine Pb

(μg/day)

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Body
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4.3.4 Lead in Other Biomarkers

1 There was extensive discussion in the 2006 Pb AQCD regarding the utility of other Pb 

2 biomarkers as indicators of exposure or body burden. Due to the fact that most 

3 epidemiologic studies continue to use blood Pb or bone Pb, and other potential 

4 biomarkers (i.e., teeth, hair, and saliva) have not been established to the same extent as 

5 blood or bone Pb, only summaries are 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, 

7 tooth Pb levels are generally considered an estimate of cumulative Pb exposure. The 

8 tooth Pb-blood Pb relationship is more complex than the bone Pb-blood Pb relationship 

9 because of differences in tooth type, location, and analytical method. Although 

10 mobilization of Pb from bone appears well established, this is not the case for Pb in teeth. 

11 Conventional wisdom has Pb fixed once it enters the tooth. Although that may be the case 

12 for the bulk of enamel, it is not true for the surface of the enamel and dentine (Gulson et 

13 al., 1997; Rabinowitz et al., 1993). Limited studies have demonstrated moderate-to-high 

14 correlations between tooth Pb levels and blood Pb levels (Rabinowitz, 1995; Rabinowitz 

15 et al., 1989).

16 Teeth are composed of several tissues formed pre- and postnatal. Therefore, if a child’s 

17 Pb exposure during the years of tooth formation varied widely, different amounts of Pb 

18 would be deposited at different rates (Rabinowitz et al., 1993). This difference may allow 

19 investigators to elucidate the history of Pb exposure in a child. Robbins et al. (2010)

20 found a significant association between environmental Pb measures that correlated with 

21 leaded gasoline use and tooth enamel Pb in permanent teeth. Costa de Almeida et al.

22 (2007) discerned differences between tooth enamel Pb concentration in biopsy samples 

23 from children who lived in areas having higher or lower levels of Pb contamination. 

24 Gulson and Wilson (1994) advocated the use of sections of enamel and dentine to obtain 

25 additional information compared with analysis of the whole tooth (e.g., (Tvinnereim et 

26 al., 1997; Fosse et al., 1995). For example, deciduous tooth Pb in the enamel provides 

27 information about in utero exposure whereas that in dentine from the same tooth provides 

28 information about postnatal exposure until the tooth exfoliates at about 6-7 years of age.
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4.3.4.2 Hair

1 The 2006 Pb AQCD discussed applications of hair Pb measurements for assessing Pb 

2 body burden or exposure and noted methodological limitations (e.g., external 

3 contamination) and lack of a strong empirical basis for relating hair Pb levels to body 

4 burden or exposure. No new methodological or conceptual advances regarding hair Pb 

5 measurements have occurred since 2006, and widespread application of hair Pb 

6 measurements in epidemiologic studies has not occurred.

7 Pb is incorporated into human hair and hair roots (Bos et al., 1985; Rabinowitz et al., 

8 1976) and has been explored as a noninvasive approach for estimating Pb body burden

9 (Wilhelm et al., 2002; Gerhardsson et al., 1995; Wilhelm et al., 1989). Hair Pb 

10 measurements are subject to error from contamination of the surface with environmental 

11 Pb and contaminants in artificial hair treatments (i.e., dyeing, bleaching, permanents) and 

12 are a relatively poor predictor of blood Pb concentrations, particularly at blood Pb levels 

13 less than 10-12 μg/dL (Rodrigues et al., 2008; Campbell and Toribara, 2001; Esteban et 

14 al., 1999; Drasch et al., 1997). Temporal relationships between Pb exposure and hair Pb 

15 levels, and kinetics of deposition and retention of Pb in hair have not been evaluated. 

16 Although hair Pb measurements have been used in some epidemiologic studies (Shah et 

17 al., 2011; U.S. EPA, 2006b), an empirical basis for interpreting hair Pb measurements in 

18 terms of body burden or exposure has not been firmly established.

4.3.4.3 Saliva

19 A growing body of literature on the utility of measurements of salivary Pb has developed 

20 since the completion of the 2006 Pb AQCD (U.S. EPA, 2006b). Earlier reports suggested 

21 a relatively strong correlation between salivary Pb concentration and blood Pb 

22 concentration (Omokhodion and Crockford, 1991; Brodeur et al., 1983; P'an, 1981);

23 however, more recent assessments have shown relatively weak or inconsistent 

24 associations (2011; 2010; Costa de Almeida et al., 2009; Barbosa et al., 2006c; Nriagu et 

25 al., 2006). The differences in these outcomes may reflect differences in blood Pb 

26 concentrations, exposure history and/or dental health (i.e., transfer of Pb between dentin 

27 and saliva) and possibly methods for determining Pb in saliva. Barbosa et al. (2006c)

28 found a significant but relatively weak correlation (log[blood PB] versus log[saliva Pb], r 

29 = 0.277, p = 0.008) in a sample of adults, ages 18-60 years (n = 88). The correlation was 

30 similar for salivary and plasma Pb. Nriagu et al. (2006) found also found a relatively 

weak association (R2 31 = 0.026) between blood Pb (μg/dL) and salivary Pb (μg/L) in a 

32 sample of adults who resided in Detroit, MI (n = 904). Costa de Almeida et al. (2009)

33 found a significant correlation between salivary and blood Pb concentrations in children 
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1 in a Pb-contaminated region in Sao Paulo State, Brazil (r = 0.76. p = 0.04, n = 7) prior to 

2 site remediation; however, the correlation degenerated (r = 0.03, p = 0.94, n = 9) 

3 following remediation. Nevertheless, salivary Pb concentrations in the group of children 

4 who lived in the contaminated area were significantly elevated compared to a reference 

5 population. It is possible, that salivary Pb measurements may be a useful non-invasive 

6 biomarker for detecting elevated Pb exposure; however, it is not clear based on currently 

7 available data, if salivary Pb measurements would be a more reliable measure of 

8 exposure than blood Pb measurements.

4.3.4.4 Serum δ-ALA and ALAD

9 The association between blood Pb and blood ALAD activity and serum δ-aminolevulinic 

10 acid (δ-ALA) levels was recognized decades ago as having potential use as a biomarker 

11 of Pb exposure (Mitchell et al., 1977; Hernberg et al., 1970). More recently reference 

12 values for blood ALAD activity ratio (the ratio of ALAD activity in the blood sample to 

13 that measured after fully activating the enzyme in the sample) have been reported

14 (Gultepe et al., 2009). Inhibition of erythrocyte ALAD by Pb results in a rise in the 

15 plasma concentration of the ALAD substrate δ-ALA. The δ-ALA biomarker can be 

16 measured in serum and has been used as a surrogate for Pb measurements in studies in 

17 which whole blood samples or adequately prepared plasma or serum samples were not 

18 available for Pb measurements (Opler et al., 2008; Opler et al., 2004).

4.3.5 Relationship between Lead in Blood and Lead in Bone

19 The kinetics of elimination of Pb from the body reflects the existence of fast and slow 

20 pools of Pb in the body. The dominant phase of Pb kinetics in the blood, exhibited shortly 

21 after a change in exposure occurs, has a half-life of ~20-30 days (Leggett, 1993; 

22 Rabinowitz et al., 1976). Studies of a limited number of adults (four individuals with hip 

23 or knee replacement, a married couple, and 10 female Australian immigrants) in which 

24 the Pb exposure was from historical environmental sources have found that bone Pb 

25 stores can contribute 40-70% to blood Pb (Smith et al., 1996; Gulson et al., 1995; 

26 Manton, 1985). Bone Pb burdens in adults are slowly lost by diffusion (heteroionic 

27 exchange) as well as by resorption (O'Flaherty, 1995). Half-times for the release of Pb in 

28 bone are dependent on age and intensity of exposure. Bone compartments are much more 

29 labile in infants and children than in adults as reflected by half-times for movement of Pb 

30 from bone into the plasma (e.g., cortical t1/2 = 0.23 years at birth, 1.2 years at 5 years of 

31 age, 3.7 years at 15 years of age, and 23 years in adults; trabecular t1/2 = 0.23 years at 
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1 birth, 1.0 years at 5 years of age, 2.0 years at 15 years of age, and 3.9 years in adults)

2 (Leggett, 1993). Slow transfer rates for the movement of Pb from nonexchangeable bone 

3 pools to the plasma are the dominant transfer process determining long-term 

4 accumulation and elimination of bone Pb burden.

5 The apparent slow clearance of Pb from the blood over months and years following the 

6 cessation or reduction in exposures reflects elimination of Pb stores in bone. Longer fast7 phase elimination half-times (60-120 days) were reported for five adults with lead 

8 poisoning following roughly a month to 12 years of exposure and relatively high blood 

9 Pb concentrations of 70-110 μg/dL (Rentschler et al., In Press). A slower phase becomes 

10 evident with longer observation periods following a decrease in exposure. Nilsson et al.

11 (1991) reported in tri-exponential decay in the blood Pb concentrations of 14 individuals 

12 having a median occupational exposure period of 26 years. Representing 22% of blood 

13 Pb, the fast compartment had a clearance half time of 34 day. The intermediate 

14 compartment, 27% of blood Pb, had a clearance half time of 1.12 year. And, the slow 

15 compartment, 50% of blood Pb, had a clearance half time of 13 years. The authors 

16 attributed the fast, intermediate, and slow compartment clearance to elimination of Pb 

17 from blood and some soft tissues, from trabecular bone, and cortical bone, respectively. 

18 Children who have been removed from a relatively brief exposure to elevated 

19 environmental Pb also exhibit faster slow-phase kinetics than children removed from 

20 exposures that lasted several years, with half-times of 10 and 20-38 months, respectively

21 (Manton et al., 2000). Rothenberg et al. (1998) also showed that exposures in the first 6 

22 months of life could contribute to elevated blood lead through at least 3 years relative to 

23 children with lower early life exposures, despite similar environmental exposures at later 

24 time points. In both adults and children, the longer half-times measured under the latter 

25 conditions reflect the contribution of bone Pb stores to blood Pb following a change in 

26 exposure.

27 The longer half-life of Pb in bone compared to blood Pb, allows a more cumulative 

28 measure of Pb dose. Pb in adult bone can serve to maintain blood Pb levels long after 

29 external exposure has ceased (Fleming et al., 1997; Inskip et al., 1996; Smith et al., 1996; 

30 Kehoe, 1987; O'Flaherty et al., 1982), even for exposures that occurred during childhood

31 (McNeill et al., 2000). The more widespread use of in vivo XRF Pb measurements in 

32 bone and indirect measurements of bone processes with stable Pb isotopes have enhanced 

33 the use of bone Pb as a biomarker of Pb body burden.

34 Several studies have found a stronger relationship between patella Pb and blood Pb than 

35 tibia Pb and blood Pb (Park et al., 2009a; Hu et al., 1998; Hernandez-Avila et al., 1996; 

36 Hu et al., 1996b). Hu et al. (1998) suggest that trabecular bone is the predominant bone 

37 type providing Pb back into circulation under steady-state and pathologic conditions. The 
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1 stronger relationship between blood Pb and trabecular Pb compared with cortical bone is 

2 probably associated with the larger surface area of trabecular bone allowing for more Pb 

3 to bind via ion exchange mechanisms and more rapid turnover making it more sensitive 

4 to changing patterns of exposure.

4.3.5.1 Children

5 As mentioned in Section 4.2.2.2, bone growth in children will contribute to accumulation 

6 of Pb in bone, which will comprise most of the Pb body burden. As a result, Pb in bone 

7 will more closely reflect Pb body burden than blood Pb. However, changes in blood Pb 

8 concentration in children (i.e., associated with changing exposures) are thought to more 

9 closely parallel changes in total body burden. Figure 4-9 shows a simulation of the 

10 temporal profile of Pb in blood and bone in a child who experiences a period of constant 

11 Pb intake (from age 2-5) via ingestion (μg Pb/day) followed by an abrupt decline in 

12 intake. The figure illustrates several important general concepts about the relationship 

13 between Pb in blood and bone. While blood Pb approaches a quasi-steady state after a 

14 period of a few months with a constant rate of Pb intake (as demonstrated by the vertical 

15 dashed line), Pb continues to accumulate in bone with continued Pb intake after the 

16 quasi-steady state is achieved in blood. The model also predicts that the rate of release of 

17 Pb from bone after cessation of exposure is faster than in adults. This difference has been 

18 attributed to accelerated growth-related bone mineral turnover in children, which is the 

19 primary mechanism for release of Pb that has been incorporated into the bone mineral 

20 matrix.

21 Empirical evidence in support of this conclusion comes from longitudinal studies in 

22 which relatively high correlations were found between concurrent (r = 0.75) or average 

23 lifetime (obtained at 6-month intervals from birth to age 10 or 12) blood Pb 

24 concentrations (r = 0.85) and tibia bone Pb concentrations (measured by XRF) in a 

25 sample of children in which the group mean concurrent blood Pb concentration exceeded 

26 20 μg/dL; the correlations was much weaker (r <0.15) among the group of children with 

27 a mean concurrent blood Pb concentration <10 μg/dL (Wasserman et al., 2003).

28 Two alternative blood Pb metrics depicted in Figure 4-9 include the time-averaged and 

29 time-integrated blood Pb concentrations. Both the time-averaged and time-integrated 

30 blood Pb metrics display rates of change in response to the exposure event that more 

31 closely approximate the slower kinetics of bone Pb and body burden, than the kinetics of 

32 blood Pb concentration, with notable differences. The time-averaged blood Pb 

33 concentration increases during the exposure event and decays following the event, 

34 consistent with the changing body burden. The time-integrated blood Pb concentration 
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1 (conceptually identical to cumulative blood lead index [CBLI] used in epidemiologic 

2 studies) is a cumulative function and increases throughout childhood; however, the slope 

3 of the increase is higher during the exposure event than prior to or following the event. 

4 Following cessation of exposure, the time-integrated blood Pb and body burden diverge. 

5 This result is expected, as the time-integrated blood Pb curve is a cumulative function 

6 which cannot decrease over time and bone Pb levels will decrease with cessation of 

7 exposure.

8 The time-integrated blood Pb concentration will be a better reflection of the total amount 

9 of Pb that has been absorbed, than the body burden at any given time. The time10 integrated blood Pb concentration will also reflect cumulative Pb absorption, and 

11 cumulative exposure if the absorption fraction is constant. This is illustrated in the 

12 hypothetical simulations of an exposure event experienced by a child (Figure 4-10). This 

13 pattern is similar for adults.
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Note: 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).

Figure 4-9 Simulation of relationship between blood Pb concentration and 

body burden in children, with a constant Pb intake from age 

2 to 5.
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Note: 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).

Figure 4-10 Simulation of relationship between time-integrated blood Pb 

concentration and cumulative Pb absorption in children.

4.3.5.2 Adults

1 In adults, where a relatively large fraction of the body burden residing in bone has a 

2 slower turnover compared to blood, a constant Pb uptake (or constant intake and 

3 fractional absorption) gives rise to a quasi-steady state blood Pb concentration, while the 

4 body burden continues to increase over a much longer period, largely as a consequence of 

5 continued accumulation of Pb in bone. This pattern is illustrated in Figure 4-11 which 

6 depicts a hypothetical simulation of an exposure event consisting of a 20-year period of 

7 daily ingestion of Pb in an adult. The exposure event shown in the simulations gives rise 

8 to a relatively rapid increase in blood Pb concentration, to a new quasi-steady state, 

9 achieved in ~75-100 days (i.e., approximately 3-4 times the blood elimination half-life). 

10 In contrast, the body burden continues to increase during this period. Following cessation 

11 of the exposure, blood Pb concentration declines relatively rapidly compared to the 

12 slower decline in body burden. Careful examination of the simulation shown in Figure 

13 4-11 reveals that the accumulation and elimination phases of blood Pb kinetics are not 

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
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1 symmetrical; elimination is slower than accumulation as a result of the gradual release of 

2 bone Pb stores to blood. This response, known as the prolonged terminal elimination 

3 phase of Pb from blood, has been observed in retired Pb workers and in workers who 

4 continued to work after improved industrial hygiene standards reduced their exposures. In 

5 the adult simulation shown in Figure 4-11, the initial phase of elimination (the first 

6 5 years following cessation of exposure at 50 years) has a half-time of approximately 

7 14 years; however, the half-time increases to approximately 60 years during the period 

8 5-30 years after cessation of exposure. These model predictions are consistent with the 

9 slow elimination of Pb from blood and elimination half-times of several decades for bone 

10 Pb (e.g., 16-98 years) that have been estimated from observations made on Pb workers

11 (Wilker et al., 2011; Fleming et al., 1997; Gerhardsson et al., 1995). Based on this 

12 hypothetical simulation, a blood Pb concentration measured 1 year following cessation of 

13 a period of increased Pb uptake would show little or no appreciable change from prior to 

14 the exposure event, whereas, the body burden would remain elevated. These simulations 

15 in Figure 4-11 illustrate how a single blood Pb concentration measurement, or a series of 

16 measurements taken over a short-time span, could be a relatively poor index of Pb body 

17 burden. 

18 One important potential implication of the profoundly different kinetics of Pb in blood 

19 and bone is that, for a constant Pb exposure, Pb in bone will increase with increasing 

20 duration of exposure and, therefore, with age. In contrast, blood Pb concentration will 

21 achieve a quasi-steady state. As a result, the relationship between blood Pb and bone Pb 

22 will diverge with increasing exposure duration and age. This divergence can impart 

23 different degrees of age-confounding when either blood Pb or bone Pb is used as an 

24 internal dose metric in dose-response models. In a review of epidemiologic studies that 

25 evaluated the associations between blood Pb, bone Pb and cognitive function, the effects 

26 of bone Pb were more pronounced than blood Pb (particularly for longitudinal studies) 

27 for older individuals with environmental Pb exposures and low blood Pb levels (Shih et 

28 al., 2007). In contrast, occupational workers with high current Pb exposures had the 

29 strongest associations for blood Pb levels with cognitive function, thus providing 

30 evidence for this divergence (Shih et al., 2007).

31 The aforementioned expectation for an increase in bone Pb and body burden with age 

32 applies to scenarios of constant exposure but not necessarily to real world populations in 

33 which individual and population exposures have changed over time. Longitudinal studies 

34 of blood and bone Pb trends have not always found strong dependence on age (Nie et al., 

35 2009; Kim et al., 1997). Kim et al. (1997) found that bone Pb levels increased with 

36 increasing age in elderly adults (age 52-83) years), only when the data were analyzed 

37 cross-sectionally. When analyzed longitudinally, the trend for individual patella Pb was a 

38 23% decrease over a 3-year period (approximate t1/2 of 8 years), whereas tibia Pb levels 
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1 did not change with over the same period. Therefore, although older individuals tended to 

2 have higher bone Pb levels, the 3-year temporal trend for individuals was a loss of Pb 

3 from the more labile Pb stores in trabecular bone. Nie et al. (2011) observed that 

4 longitudinal observations of blood and bone Pb in elderly adults did not show a 

5 significant age effect on the association between blood Pb and bone Pb (patella and tibia), 

6 when the sample population (n=776) was stratified into age tertiles (mean age 62, 69 or 

7 77 years). Nie et al. (2009) did find that regressed function bone Pb and appeared to level 

8 off at bone Pb levels >20 μg/g bone mineral.

Note: A constant baseline intake results in 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-steady state, and a slower change in body burden. The 

long-term time-averaged blood Pb concentration more closely tracks the 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 

uptake (lower panel). Simulation based on ICRP Pb biokinetics model (Leggett, 1993).

Figure 4-11 Simulation of relationship between blood Pb concentration, bone 

Pb and body burden in adults.
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1 Tibia bone Pb is correlated with time-integrated blood Pb concentration (i.e., CBLI). 

2 McNeill et al. (2000) compared tibia Pb levels and cumulative blood Pb indices in a 

3 population of 19- to 29-year-olds who had been highly exposed to Pb in childhood from 

4 the Bunker Hill, Idaho smelter. They concluded that Pb from exposure in early childhood 

5 had persisted in the bone matrix until adulthood. The bone Pb/CBLI slopes from various 

6 studies range from 0.022 to 0.067 μg/g bone mineral per μg-year/dL (Healey et al., 2008; 

7 Hu et al., 2007a). Because the CBLI is a cumulative function which cannot decrease over 

8 time, CBLI and bone Pb would be expected to diverge following cessation of exposure, 

9 as bone Pb levels decrease. This divergence was observed as a lower bone Pb/CBLI slope 

10 in retired Pb workers compared to active workers and in worker populations whose 

11 exposures declined over time as a result of improved industrial hygiene (Fleming et al., 

12 1997; Gerhardsson et al., 1993).

13 Although differences in kinetics of blood and bone Pb degrade the predictive value of 

14 blood Pb as a metric of Pb body burden, within a population that has similar exposure 

15 histories and age demographics, blood and bone Pb may show relatively strong 

16 associations. A recent analysis of a subset of data from the Normative Aging Study (an 

17 all male cohort) showed that cross-sectional measurements of blood Pb concentration 

18 accounted for approximately 9% (tibia) to 13% (patella) of the variability in bone Pb 

19 levels. Inclusion of age in the regression model accounted for an additional 7-10% of the 

20 variability in bone Pb (Park et al., 2009a).

Mobilization of Lead from Bone in Adulthood

21 In addition to changes in exposure (e.g., declines in exposure discussed in prior section), 

22 there are physiological processes during different life circumstances that can increase the 

23 contribution of bone Pb to blood Pb. These life circumstances include times of 

24 physiological stress associated with enhanced bone remodeling such as during pregnancy 

25 and lactation (Hertz-Picciotto et al., 2000; Silbergeld, 1991; Manton, 1985), menopause 

26 or in the elderly (Silbergeld et al., 1988), extended bed rest (Markowitz and Weinberger, 

27 1990), hyperparathyroidism (Kessler et al., 1999) and severe weight loss (Riedt et al., 

28 2009).

29 During pregnancy, bone Pb can serve as a Pb source as maternal bone is resorbed for the 

30 production of the fetal skeleton (Gulson et al., 2003; Gulson et al., 1999; Franklin et al., 

31 1997; Gulson et al., 1997). Increased blood Pb during pregnancy has been demonstrated 

32 in numerous studies and these changes have been characterized as a “U-shaped” pattern 

33 of lower blood Pb concentrations during the second trimester compared to the first and 

34 third trimesters (Lamadrid-Figueroa et al., 2006; Gulson et al., 2004a; Hertz-Picciotto et 

35 al., 2000; Gulson et al., 1997; Lagerkvist et al., 1996; Schuhmacher et al., 1996; 
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1 Rothenberg et al., 1994a). The U-shaped relationship reflects the relatively higher impact 

2 of hemodilution in the second trimester versus the rate of bone Pb resorption 

3 accompanying calcium releases for establishing the fetal skeleton. In the third trimester, 

4 fetal skeletal growth on calcium demand is greater, and Pb released from maternal 

5 skeleton offsets hemodilution. Gulson et al. (1998a) reported that, during pregnancy, 

6 blood Pb concentrations in the first immigrant Australian cohort (n = 15) increased by an 

7 average of about 20% compared to non-pregnant migrant controls (n = 7). Skeletal 

8 contribution to blood Pb, based on the isotopic composition for the immigrant subjects, 

9 increased in an approximately linear manner during pregnancy. The mean increases for 

10 each individual during pregnancy varied from 26% to 99%. Interestingly, the percent 

11 change in blood Pb concentration was significantly greater during the post-pregnancy 

12 period than during the second and third trimesters. This is consistent with Hansen et al.

13 (2011b) that demonstrated the greatest blood Pb levels at 6 weeks postpartum compared 

14 to the second trimester in 211 Norwegian women. Increased calcium demands of 

15 lactation (relative to pregnancy) may contribute to the greater change in blood Pb 

16 observed post pregnancy compared to the second and third trimesters. The contribution of 

17 skeletal Pb to blood Pb during the post-pregnancy period remained essentially constant at 

18 the increased level of Pb mobilization.

19 Gulson et al. (2004a) observed that calcium supplementation was found to delay 

20 increased mobilization of Pb from bone during pregnancy and halved the flux of Pb 

21 release from bone during late pregnancy and postpartum. In another study, women whose 

22 daily calcium intake was 850 mg per day showed lower amounts of bone resorption 

23 during late pregnancy and postpartum than those whose intake was 560 mg calcium per 

24 day (Manton et al., 2003). Similarly, calcium supplementation (1,200 mg/day) in 

25 pregnant Mexican women resulted in an 11% reduction in blood Pb level compared to 

26 placebo and a 24% average reduction for the most compliant women (Ettinger et al., 

27 2009). When considering baseline blood Pb levels in women who were more compliant 

28 in taking calcium supplementation, the reductions were similar for those <5 μg/dL and 

29 those ≥ 5 μg/dL (14% and 17%, respectively). This result is in contrast to a study of 

30 women who had blood Pb concentrations <5 μg/dL, where calcium supplementation had 

31 no effect on blood Pb concentrations (Gulson et al., 2006b). These investigators 

32 attributed their results to changes in bone resorption with decoupling of trabecular and 

33 cortical bone sites.

34 Miranda et al. (2010) studied blood Pb level among pregnant women aged 18-44 years 

35 old. The older age segments in the study presumably had greater historic Pb exposures 

36 and associated stored Pb than the younger age segments. Compared with the blood Pb 

37 levels of a reference group in the 25-29 years old age category, pregnant women ≥ 

38 30 years old had significant odds of having higher blood Pb levels (aged 30-34: OR = 
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1 2.39, p <0.001; aged 35-39: OR = 2.98, p <0.001; aged 40-44: OR = 7.69, p <0.001). 

2 Similarly, younger women had less chance of having higher blood Pb levels compared 

3 with the reference group (aged 18-19: OR = 0.60, p = 0.179; aged 20-24: OR = 0.54, p = 

4 0.015). These findings indicate that maternal blood Pb levels are more likely the result of 

5 Pb mobilization of bone stores from historic exposures as opposed to contemporaneous 

6 exposures.

7 Blood Pb levels increase during lactation due to alterations in the endogenous bone Pb 

8 release rate. After adjusting for patella Pb concentration, an increase in blood Pb levels of 

9 12.7% (95% CI: 6.2, 19.6) was observed for women who practiced partial lactation and 

10 an increase of 18.6% (95% CI: 7.1, 31.4) for women who practiced exclusive lactation 

11 compared to those who stopped lactation (Tellez-Rojo et al., 2002). In another Mexico 

12 City study, Ettinger et al. (2006; 2004b) concluded that an interquartile increase in patella 

13 Pb was associated with a 14% increase in breast milk Pb, whereas for tibia Pb the 

14 increase was ~5%. Breast milk:maternal blood Pb concentration ratios are generally <0.1, 

15 although values of 0.9 have been reported (Koyashiki et al., 2010; Ettinger et al., 2006; 

16 Gulson et al., 1998b). Dietary intake of polyunsaturated fatty acids (PUFA) has been 

17 shown to weaken the association between Pb levels in patella and breast milk, perhaps 

18 indicating decreased transfer of Pb from bone to breast milk with PUFA consumption

19 (Arora et al., 2008).

20 The Pb content in some bones (i.e., mid femur and pelvic bone) plateau at middle age and 

21 then decreases at older ages (Drasch et al., 1987). This decrease is most pronounced in 

22 females and may be due to osteoporosis and release of Pb from resorbed bone to blood

23 (Gulson et al., 2002). Two studies indicate that the endogenous release rate in 

24 postmenopausal women ranges from 0.13-0.14 μg/dL in blood per μg/g bone and is 

25 nearly double the rate found in premenopausal women (0.07-0.08 μg/dL per μg/g bone)

26 (Popovic et al., 2005; Garrido Latorre et al., 2003). An analysis of data on blood lead 

27 concentrations and markers of bone formation (serum alkaline phosphatase) and 

28 resorption (urinary cross-linked N-telopeptides, NTx) in a sample of U.S. found that 

29 blood Pb concentrations were higher in women (pre- or post-menopausal) who exhibited 

30 the highest bone formation or resorption activities (Jackson et al., 2010). Calcium or 

31 vitamin D supplementation decreased the blood lead concentrations in the highest bone 

32 formation and resorption tertiles of the population of post-menopausal women. 

33 Significant associations between increasing NTx and increasing blood Pb levels 

34 (i.e., increased intercept of regression model relating the change in blood Pb per change 

35 in bone Pb) has also been observed in elderly males (Nie et al., 2009).

36 Studies of the effect of hormone replacement therapy on bone Pb mobilization have 

37 yielded conflicting results (Popovic et al., 2005; Berkowitz et al., 2004; Garrido Latorre 
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1 et al., 2003; Korrick et al., 2002; Webber et al., 1995). In women with severe weight loss 

2 (28% of BMI in 6 months) sufficient to increase bone turnover, increased blood Pb levels 

3 of approximately 2.1 μg/dL (250%) were reported, and these blood Pb increases were 

4 associated with biomarkers of increased bone turnover (e.g., urinary pyridinoline cross5 links) (Riedt et al., 2009).

4.3.6 Relationship between Lead in Blood and Lead in Soft Tissues

6 Figure 4-12 shows simulations of blood and soft tissues Pb (including brain) for the same 

7 exposure scenarios previously displayed. Pb uptake and elimination in soft tissues is 

8 much faster than bone. As a result, following cessation of a period of elevated exposure, 

9 Pb in soft tissues is more quickly returned to blood. The terminal elimination phase from 

10 soft tissue mimics that of blood, and it is similarly influenced by the contribution of bone 

11 Pb returned to blood and being redistributed to soft tissue. 
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Note: Simulation based on ICRP Pb biokinetics model (Leggett, 1993).

Figure 4-12 Simulation of blood and soft tissue (including brain) Pb in 

children and adults who experience a period of increased Pb 

intake.

1 Information on Pb levels in human brain are limited to autopsy data and the simulation of 

2 brain Pb shown in Figure 4-13 reflects general concepts derived from observations made 

3 in non-human primates, dogs and rodents. These observations suggest that peak Pb levels 

4 in the brain are reached 6 months following a bolus exposure and within two months 

5 approximately 80% of steady state brain Pb levels are reached (Leggett, 1993). There is a 

6 relatively slow elimination of Pb from brain (t1/2 ≈ 2 years) compared to other soft tissues

0.0

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Soft Tissue

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Blood Pb (μg/dL)

Age (year)

Blood

Soft Tissue
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1 (Leggett, 1993). This slow elimination rate is reflected in the slower elimination phase 

2 kinetics in shown Figure 4-13. Although in this model, brain Pb to blood Pb transfer half3 times are assumed to be the same in children and adults, uptake kinetics are assumed to 

4 be faster during infancy and childhood, which achieves a higher fraction of the soft tissue 

5 burden in brain, consistent with higher brain/body mass relationships. The uptake half 

6 times predicted by Leggett (1993) vary from 0.9 to 3.7 days, depending on age. Brain Pb 

7 kinetics represented in the simulations are simple outcomes of modeling assumptions and 

8 cannot currently be verified with available observations in humans. 
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Note: Simulation based on ICRP Pb biokinetics model (Leggett, 1993).

Figure 4-13 Simulation of blood and brain Pb in children and adults who 

experience a period of increased Pb intake.

1 Urinary filtering and excretion of Pb is associated with plasma Pb concentrations. Given 

2 the curvilinear relationship between blood Pb and plasma Pb, a secondary expectation is 

3 for a curvilinear relationship between blood Pb and urinary Pb excretion that may 

4 become evident only at relatively high blood Pb concentrations (e.g., >25 μg/dL). Figure 

5 4-14 shows these relationships predicted from the model. In this case, the exposure 

0

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24

32

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48

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Total Brain Pb (

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Blood Pb (

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Age (year)

Blood

Brain
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1 scenario shown is for an adult (age 40 years) at a quasi-steady state blood Pb 

2 concentration; the same relationships hold for children. At lower blood Pb concentrations 

3 (<25 μg/dL), urinary Pb excretion is predicted to closely parallel plasma Pb concentration 

4 for any given blood Pb level (Figure 4-14, top panel). It follows from this that, similar to 

5 blood Pb, urinary Pb will respond much more rapidly to an abrupt change in Pb exposure 

6 than will bone Pb. One important implication of this relationship is that, as described 

7 previously for blood Pb, the relationships between urinary Pb and bone Pb will diverge 

8 with increasing exposure duration and age, even if exposure remains constant. 

9 Furthermore, following an abrupt cessation of exposure, urine Pb will quickly decrease 

10 while bone Pb will remain elevated (Figure 4-14, lower panel).
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Note: 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).

Figure 4-14 Relationship between Pb in urine and Pb in blood.

0.00

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Urine Pb (μg/day)

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Bone Urine Blood
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4.4 Observational Studies of Lead Biomarker Levels

4.4.1 Lead in Blood

1 Overall, trends in blood Pb levels have been decreasing among U.S. residents over the 

2 past 20 years. Blood Pb concentrations in the U.S. general population have been 

3 monitored in the NHANES. Analyses of these data show a progressive downward trend 

4 in blood Pb concentrations during the period 1976-2008, with the most dramatic declines 

5 coincident with the phase out of leaded gasoline (Pirkle et al., 1998; Brody et al., 1994; 

6 Pirkle et al., 1994; Schwartz and Pitcher, 1989). The temporal trend for the period 

7 1988-2008 is shown in Figure 4-15. Summary statistics from the most recent publically 

8 available data (1999-2008) are presented in Table 4-7 (CDC, 2011). The geometric mean 

9 Pb concentration among children 1-5 years of age, based on the sample collected during 

10 the period 2007-2008, was 1.51 μg/dL (95% CI: 1.37, 1.66), which was a slight increase 

11 from the previous year (1.46 μg/dL, 95% CI: 1.36, 1.57). Figure 4-16 uses NHANES data 

12 to illustrate temporal trends in the distribution of blood Pb levels among U.S. children 

13 aged 12-60 months. The median blood Pb in this age group was 1.4 μg/dL with a 95th

14 percentile value of 4.1 μg/dL in 2007-2008 (NCHS, 2010). For 2005-2008, 95% of 

15 childhood blood Pb levels were less than 5 μg/dL. The geometric mean blood Pb 

16 concentration among adults ≥ 20 years of age was 1.38 μg/dL (95% CI: 1.31, 1.46) for 

17 the sample collected during the period 2007-2008 (CDC, 2011). Based on these same 

18 data, the geometric mean for all males (aged ≥ 1 y) was 1.47 μg/dL (95% CI: 1.39, 1.56), 

19 and for females (aged ≥ 1 y) was 1.11 μg/dL (95% CI: 1.06, 1.16).

20 There has been a steep decline in mean blood Pb levels from 1975 through 2010 among 

21 all birth cohorts from 1975 to 2010; Figure 4-17. For all cohorts, blood Pb generally 

22 decreases with age during childhood until adolescence; following adolescence (in the 

23 early 20s), blood Pb generally levels off or even increases with age. It is possible that 

24 bone growth in young people and occupational exposure for adults influences the shape 

25 of these curves. For the 1960 to 1970 birth cohort, the mean blood Pb is the highest of the 

26 cohorts in the 1970s, but beginning in 1993 the mean blood Pb is one of the lowest of the 

27 cohorts. This interaction between time and cohort may be due to the faster release of Pb 

28 from bone in younger people (Rabinowitz, 1991). This interaction is also apparent for 

29 some of the other more recently born cohorts. In comparison, the slopes of blood Pb over 

30 time are nearly parallel among the cohorts born before 1930. This suggests that the time31 cohort-interaction diminishes among older people. Also, the leveling of the blood Pb in 

32 the 2000s could be due to aging of the birth cohort and consequent slowing of their Pb 

33 release from bone.
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1 When race/ethnicity groups were compared for years 1999-2004, blood Pb levels in 

2 children were highest in the ethnicity category non-Hispanic black (GM 2.8, 95% CI: 2.5, 

3 3.0) compared to the categories Mexican-American (GM 1.9, 95% CI: 1.7, 2.0) and 

4 non-Hispanic white (GM 1.7, 95% CI: 1.6, 1.8) (Jones et al., 2009a). Figure 4-18

5 demonstrates the change in percent of children (aged 1-5 years) with various blood Pb 

6 levels by race/ethnicity between the survey during 1988-1991 and that during 1999-2004. 

7 When these data for children aged 1-5 years were aggregated for all survey years from 

8 1988 to 2004, residence in older housing, poverty, age, and being non-Hispanic black 

9 were significant predictors of higher Pb levels (Jones et al., 2009a).

Note: 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.

Figure 4-15 Temporal trend in blood Pb concentration.

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
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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 8,945

2003-2004 1.43 1.36, 1.50 8,373

2005-2006 1.29 1.23, 1.36 8,407

2007-2008 1.27 1.21, 1.34 8,266

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 1,044

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 1,011

12-19 yr

1999-2000 1.10 1.04, 1.17 2,135

2001-2002 0.94 0.90, 0.99 2,231

2003-2004 0.95 0.88, 1.02 2,081

2005-2006 0.80 0.75, 0.85 1,996

2007-2008 0.80 0.74, 0.86 1,074

≥ 20 yr

1999-2000 1.75 1.68, 1.81 4,207

2001-2002 1.56 1.49, 1.62 4,772

2003-2004 1.52 1.45, 1.60 4,525

2005-2006 1.41 1.34, 1.48 4,509

2007-2008 1.38 1.31, 1.46 5,364

Males

1999-2000 2.01 1.93, 2.09 3,913

2001-2002 1.78 1.71, 1.86 4,339

2003-2004 1.69 1.62, 1.75 4,132

2005-2006 1.52 1.42, 1.62 4,092

2007-2008 1.47 1.39, 1.56 4,147

Females

1999-2000 1.37 1.32, 1.43 4,057

2001-2002 1.19 1.14, 1.25 4,606

2003-2004 1.22 1.14, 1.31 4,241

2005-2006 1.11 1.05, 1.17 4,315

2007-2008 1.11 1.06, 1.16 4,119

Mexican - Americans

1999-2000 1.83 1.75, 1.91 2,742

2001-2002 1.46 1.34, 1.60 2,268

2003-2004 1.55 1.43, 1.69 2,085

2005-2006 1.29 1.21, 1.38 2,236

2007-2008 1.25 1.15, 1.36 1,712

Non-Hispanic blacks

1999-2000 1.87 1.75, 2.00 1,842

2001-2002 1.65 1.52, 1.80 2,219

2003-2004 1.69 1.52, 1.89 2,293

2005-2006 1.39 1.26, 1.53 2,193

2007-2008 1.39 1.30, 1.48 1,746
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Survey Stratum Period Geometric Mean (μg/dL) 95% Confidence Interval Number of Subjects

Non-Hispanic whites

1999-2000 1.62 1.55, 1.69 2,716

2001-2002 1.43 1.37, 1.48 3,806

2003-2004 1.37 1.32, 1.43 3,478

2005-2006 1.28 1.19, 1.37 3,310

2007-2008 1.24 1.16, 1.33 3,461

Source: Adapted from data from the NHANES (CDC, 2011).

Age strata correspond to the NHANES study design.

Source: Adapted from data from the NHANES (NCHS, 2010)
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Note: Top: all data. Bottom: data for subjects having blood Pb levels less than 15 μg/dL.

Figure 4-16 Box plots of blood Pb levels among U.S. children (1-5 years old) 

from the NHANES survey, 1988-2008.

Source: Adapted from data from the NHANES (NCHS, 2010)

Note: The means of logged blood Pb were weighted to represent national averages. Data were from the publically NHANES II, 

NHANES III for 1988-1991 and 1992-1994, and the continuous NHANES in 1999-2000, 2003-2004, 2005-2006, 2007-2008. 

Continuous NHANES data from 2001-2002 and 2009-2010 are not included because there were only 551 blood Pb samples in each 

of those data sets. The year plotted for exam year was the reported exam year for NHANES II, the middle year of each of the 

phases of NHANES III, and the second year of each of the continuous NHANES.

Figure 4-17 Blood Pb cohort means versus year of exam.
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Source: Data used with permission of the American Academy of Pediatrics, Jones et al. (2009a)

Figure 4-18 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 

2 [e.g., (Gulson et al., 2008; Laidlaw et al., 2005; Haley and Talbot, 2004; Johnson et al., 

3 1996)]. Seasonal variation in blood Pb concentration was evident in Australian children 

4 (n=107) with a group mean blood Pb level of 2.57 μg/dL when repeated blood Pb 

5 measurements were made over a 5-year period, with lower levels in summer compared 

0

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70

< 1 1 - < 2.5 2.5 - < 5 5 - < 7.5 7.5 - <10 ≥ 10

Percent of Children

1988-1991

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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
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1 with winter (Gulson et al., 2008). A cross-sectional study conducted in New York State 

2 from 1995-1998 demonstrated seasonality, with the greatest percent of one and two year 

3 old children with blood Pb ≥ 10 μg/dL (n=262,687) occurring in August and the lowest 

4 percent in March and April (Haley and Talbot, 2004). Meteorological factors appear to 

5 contribute to blood Pb seasonality. Laidlaw et al. (2005) analyzed the temporal 

6 relationships between child blood Pb concentrations and various atmospheric variables in 

7 three cites (Indianapolis, IN: 1999-2002; Syracuse, NY: 1994-1998; New Orleans, LA: 

8 1998-2003). Blood Pb data was obtained from public health screening programs 

9 conducted in the three cities. Blood Pb samples were dominated by children <5 years of 

10 age and age distribution varied across the three cities. The number of blood Pb 

11 measurements included in the analyses were as follows: Indianapolis, 15,969; Syracuse, 

12 14,457 (Johnson and Bretsch, 2002; Johnson et al., 1996); New Orleans, 2,295 (Mielke et 

13 al., 2007a). The temporal variation in blood Pb concentrations in each city were predicted 

14 by multivariate regression models that included the following significant variables: PM10, 

15 wind speed, air temperature, and soil moisture; as well as dummy variables accounting 

16 for temporal displacement of the effects of each independent variable on blood Pb. 

Laidlaw et al. (2005) reported R2 17 values for the regression models, but did not report the 

actual regression coefficients. The R2 18 values were as follows: Indianapolis 0.87 (p = 

19 0.004); Syracuse 0.61 (p = 0.0012); New Orleans 0.59 (p <0.00001). 

20 Studies have examined the change in blood Pb with changes in potential Pb sources. 

21 Gulson et al. (2004b) observed that children living near a Zn-Pb smelter in Australia had 

22 blood Pb levels ranging from 10 to 42 μg/dL, with 55-100% of Pb attributed to the 

23 smelter based on isotope ratio analysis. Rubio-Andrade et al. (2011) followed a cohort of 

24 6-8 y old children living within 3.5 km of a Mexican smelter at 0, 6, 12, and 60 months 

25 after environmental intervention took place. Soil Pb was concurrently obtained but not 

26 reported at 6, 12, or 60 months. Median blood Pb level at initiation of the study was 

27 10.1 μg/dL for the 598 initial participants, and median soil Pb was 3,300 mg/kg at the 

28 start of the study. After 60 months, median blood Pb level was 4.4 μg/dL for the 

29 remaining 232 participants, and median soil Pb concentration was 370 mg/kg at that time. 

30 Bonnard and McKone (2009) modeled blood Pb of French children ages 21-74 months 

31 living within a village containing a Pb smelter and estimated blood Pb levels of 

32 3.2-10.9 μg/dL. It should be noted that these studies are suggestive but not conclusive in 

33 showing that exposure to sources elevates blood Pb because these studies do not control 

34 for factors such as non-ambient in-home exposures. For this reason, Newhook et al.

35 (2003) point out that they utilize World Health Organization guidelines on fenceline 

36 concentrations in lieu of blood Pb levels in the vicinity of industrial sources to quantify 

37 exposures related to the sources. Lanphear et al. (1998) noted that the probability of 

38 children having blood Pb ≥ 10 μg/dL increases both with exterior soil Pb content and 

39 interior Pb dust loading. Mielke et al. (2011a) noted significant increases in percentages 
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1 of children younger than 7 y old with blood Pb level ≥ 10 μg/dL for those living in inner 

2 city New Orleans housing developments (22.9%) compared with children living in 

3 communities located on the city outskirts (9.1%). At the same time, median soil Pb was 

4 significantly higher in the inner city (438 mg/kg) compared with the city outskirts 

5 (117 mg/kg). 

6 For infants <1 year old, very little data are available on blood Pb levels. Simon et al.

7 (2007) followed a cohort of 13 children living near an Australian smelter from birth 

8 through 36 months. In general, immediately after birth blood Pb levels fell for 1-2 months 

9 to approximately 47% of birth blood Pb level. After this initial fall, all infants' blood Pb 

10 levels rose with age until approximately 12 months old (Simon et al., 2007). Median 

11 blood Pb level among the children was 1.9 μg/dL at 2 months and increased to 

12 13.6 μg/dL at 16 months. Geometric mean hand-Pb loading of the child and the mother

13 were significant contributors to the area under the curve for infant blood Pb, with 46% 

14 and 60% of the variance being explained by these variables, respectively; geometric 

15 mean mothers' blood Pb explained 46% of the variance (Simon et al., 2007). Across all 

16 the data, there was a good correlation between child blood Pb level and child hand Pb 

loading (R2 17 = 0.70). In another study, blood Pb levels of 15 infants aged 6-12 months 

18 were compared with cord blood Pb levels, resulting in significantly lower blood Pb levels 

19 later in life (2.24 μg/dL vs. 4.87 μg/dL) (Carbone et al., 1998)

20 Pb body burden has been reported among individuals known to consume wild game 

21 hunted with Pb shot. For example, fifty men from Nuuk, Greenland participated in a 

22 study in which they recorded their diet and produced blood samples (Johansen et al., 

23 2006). Men who regularly ate hunted game had an average blood Pb concentration of 

24 12.8 μg/dL, in contrast with those who did not and had an average blood Pb 

25 concentration of 1.5 μg/dL. Umbilical cord blood was collected from a cohort of Inuit 

26 newborns from northern Quebec, where the Inuit population consumes game killed with 

27 Pb shot (Lévesque et al., 2003). The geometric mean cord blood Pb level was

28 0.19 μmol/L, with a range of 0.01-1.31 μmol/L; the Canadian level of concern for cord 

29 blood Pb is 0.48 μmol/L. The authors contrasted the finding that 7% of Inuit newborns 

30 had cord blood Pb concentration ≥ 0.48 μmol/L in contrast with 0.16% of the Caucasian 

31 population in southern Quebec.

32 Recent studies have sought to characterize human exposure to Pb from piston-engine 

33 aircraft emissions. Section 3.2.2.1 describes a study by Carr et al. (2011) in which Pb 

34 concentrations, both modeled and monitored, extended beyond airport property. Miranda 

35 et al. (2011) used GIS to study the association between blood Pb level and distance from 

36 airports in six North Carolina Counties. They observed that the trend in blood Pb level 

37 decreases monotonically with distance class from the airports, with subjects within 500 m 
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1 of the airports having significantly increased blood Pb levels with residential proximity to 

2 airports (β = 0.043, 95% CI (0.006,0.080), p < 0.05) compared with the general 

3 population for a given county after controlling for proportion of black, Hispanic, percent 

4 receiving public assistance, and household median income at the census block group 

5 level and including dummy variables for season during which the children were screened 

6 for blood Pb.

7 Trends in blood Pb levels have been accompanied by changes in Pb isotope ratios for 

8 blood Pb. Isotopic ratios, described in Sections 3.2 and 3.3 as a tool for source 

9 apportionment, have been used to associate blood Pb measurements with anthropogenic 

10 sources of Pb in the environment. Changes in Pb isotopic ratios in blood samples reflect 

11 the changing influence of sources of Pb following the phase-out of tetraethyl Pb 

12 antiknock agents in automotive gasoline and changes in Pb usage in paints and other 

13 industrial and consumer products (Gulson et al., 2008; Ranft et al., 2008; Gulson et al., 

14 2006a; Ranft et al., 2006). Gulson et al. (2006a) illustrated how a linear increase in the 

isotopic ratio 206Pb/204 15 Pb occurred in concert with a decrease in blood Pb levels among 

16 selected study populations in Australia during the period 1990-2000 (Figure 4-19). 

Gulson et al. (2006a) point out that the isotopic signature of 206Pb/204 17 Pb derived from 

18 Australian mines (median ~16.8) differs from that of European and Asian mines, where 

206Pb/204 19 Pb varies between ~17.4 and ~18.1. Liang et al. (2010) also examined the trends 

20 in blood Pb level over the 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 21 Pb for 

22 blood and environmental samples showed overlap between the isotopic signature for coal 

23 combustion ash and that measured in blood. This result suggests a growing influence of 

24 Pb from coal ash in Shanghai in the absence of Pb in automobile emissions. Oulhote et al.

25 (2011) examined Pb isotope ratios in blood Pb samples of 125 French children aged 6 m26 6 y. The study found that Pb isotope ratios could be used to attribute Pb exposure to one 

27 source for 32% of children and to eliminate an unlikely source of Pb exposure in 30% of 

28 children.
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Source: Reprinted with permission of Academic Press, Gulson et al. (2006a)

Figure 4-19 Trends in 206Pb/204Pb isotope ratio in blood Pb (a) and trends in 

blood Pb levels (b) 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 

2 in children and adults, as described in Section 4.4.1. Bone Pb concentrations are less well 

3 characterized. Table 4-8 and Table 4-9 are compilations of data from epidemiologic 

4 studies that provided bone Pb concentrations by K-XRF and/or variability in 

5 concentrations among individuals without reported occupational exposure and those with 

6 occupational exposures, respectively. In non-occupationally exposed individuals, typical 

7 group mean tibia bone Pb concentrations ranged from 10 to 30 μg/g. Patella bone Pb 

8 levels are typically higher than tibia bone Pb levels in the studies considered (Table 4-8). 

9 For example, in the Normative Aging Study, patella bone Pb concentrations were 

10 approximately 32 μg/g, whereas tibia bone Pb concentrations were about 22 μg/g. 

11 Occupationally exposed individuals generally had greater bone Pb concentrations than 

12 seen in control groups (i.e., unexposed). Bone Pb data in Table 4-9 for occupationally 

13 exposed individuals were also generally higher compared to non-occupationally exposed 

14 individuals (Table 4-8).
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Table 4-8 Epidemiologic studies that provide bone Pb measurements for nonoccupationally exposed populations

Reference Study Methods Prior Pb Exposure

Bone Pb 

biomarker Bone Pb Conc. (μg/g)

Distribution of 

Bone Pb (μg/g)

BandeenRoche et al.

(2009)

Cohort: Baltimore Memory Study cohort

Age (yrs):50-70

N:1140

Location: Baltimore, MD

Study Period:2001-2005

Cumulative Tibia Mean±SD

Tibia: 18.8 ± 11.6

Not reported

Bellinger et al.

(1994a)

Cohort: Not reported

Age (yrs): 5-8 (recruited); 19-20 (follow-up)

N: 79

Location: Boston, MA

Study Period: 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)

Cohort: Normative Aging Study cohort

Age (yrs): Mean±SD: 

Normotensive: 65.49 ± 7.17 

Borderline hypertension: 68.3 ± 7.79 

Definite hypertension: 67.93 ± 6.79

N: 833 males

Location: Boston, MA

Study Period: 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)

Cohort: Participants from Henry Ford 

Health System (HFHS)

Age (yrs): ≥ 50; Mean: 69.9

N: 121 cases; 414 controls

Location: Southeastern Michigan

Study Period: 1995-1999 (participants 

received primary health care services)

Cumulative Tibia

Calcaneus

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

Elmarsafawy 

et al. (2006)

Cohort:Normative Aging Study

Age (yrs): Not reported

N: 471 elderly males

Location: Greater Boston area, MA

Study Period: 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)

Cohort:Baltimore Memory Study

Age (yrs): Mean: 59.4; Range: 50-70

N: 1,001

Location: Baltimore, MD

Study Period: 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

Hsieh et al.

(2009b)

Cohort:Not reported

Age (yrs): Mean: Control: 46.06

N: 18 controls

Location: Not reported

Study Period: 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.

(1996b)

[As reported in 

Navas-Acien 

et al., (2008)]

Cohort:Normative Aging Study

Age (yrs): 48-92; Mean ± SD: 66.6 ± 7.2

N: 590 males

Location: Boston, MA

Study Period: 8/1991-12/1994

Cumulative 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
February 2012 4-89 Draft – Do Not Cite or Quote

Reference Study Methods Prior Pb Exposure

Bone Pb 

biomarker Bone Pb Conc. (μg/g)

Distribution of 

Bone Pb (μg/g)

Jain et al.

(2007)

Cohort:VA-Normative Aging Study

Age (yrs): Not reported

N: 837 males

Location: Greater Boston, MA

Study Period: 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)

Cohort:Not reported

Age (yrs): 30-80

N: 256 controls (Bone samples collected from 

41 controls)

Location: New England (Boston, MA)

Study Period: 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.

(2009b)

Cohort:1982 Pb Occupational Study

Age (yrs): Control mean: 55

N: 51 controls

Location: Eastern Pennsylvania

Study Period: 1982-2004

Control group for 

occupational exposure 

group

Tibia Median (IQR)

Tibia Control: 12 (-8-32)

Not reported

Korrick et al.

(1999) (As 

reported in 

from Navas- Acien et al.,

(2008)

Cohort: Nurses’ Health Study

Age (yrs): 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

N: 284 females; (89 cases; 195 controls)

Location: Boston, MA

Study Period: 7/1993-7/1995

Nonoccupationally 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.

(2001a) [As 

reported in 

Navas-Acien 

et al., (2008)]

Cohort: Not reported

Age (yrs): 22.0-60.2 Mean ± SD: Controls: 

34.5 ± 9.1

N: 135 controls

Location: South Korea

Study Period: 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
February 2012 4-90 Draft – Do Not Cite or Quote

Reference Study Methods Prior Pb Exposure

Bone Pb 

biomarker Bone Pb Conc. (μg/g)

Distribution of 

Bone Pb (μg/g)

Martin et al.

(2006)

Cohort: Baltimore Memory Study

Age (yrs): 50-70; Mean: 59.4

N: 964

Location: Baltimore, MD

Study Period: 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)

Cohort: Not reported

Age (yrs): 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

N: 194 male youth cases; 146 male youth 

controls

Location: Allegheny County, PA (cases); 

Pittsburgh, PA (controls)

Study Period: 4/1996-8/1998

Not 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 of paper 

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)]

Cohort: Not reported

Age (yrs): Median: 41.5

N: 19 male controls

Location: Not reported

Study Period: Not reported

Control group for 

occupational exposure 

group

Finger bone Median (range)

Finger Bone Controls: 

4 (-19-18)

Not reported

Park et al.

(2006)

Cohort: Normative Aging Study

Age (yrs): Mean: 72.9 ± 6.5

N: 413 males

Location: Greater Boston, MA

Study Period: 11/14/2000-12/22/2004; (HRV 

measurements taken); 1991-2002 (bone Pb 

measurements 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.

(2009b)

Cohort: Normative Aging Study

Age (yrs): Mean: 67.3 ± 7.2

N: 613 males

Location: Greater Boston, MA

Study Period: 8/1991 - 12/1995

Not reported Tibia

Patella

Median (IQR)

Tibia: 19 (14-27)

Patella: 26 (18-37)

Table 1 of paper 

distributes tibia and 

patella Pb by 

genotype; Table 2 of 

paper distributes tibia 

and patella Pb by 

number of gene 

variants

Park et al.

(2010)

Cohort: VA Normative Aging Study cohort

Age (yrs): Mean: 64.9 (at bone 

Pb measurement)

N: 448 males

Location: Eastern Massachusetts

Study Period: 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 of paper 

provides ageadjusted mean bone 

Pb levels (age, race, 

education, smoking 

[pack-yr], 

occupational noise, 

noise notch, BMI, 

hypertension, 

diabetes)

Payton et al.

(1998)

Cohort: VA Normative Aging Study cohort

Age (yrs): Mean: 66.8

N: 141 males

Location: Boston, MA

Study Period: 4/1993-3/1994

Not reported Tibia

Patella

Mean ± SD

Tibia: 22.5 ± 12.2

Patella: 31.7 ± 19.2

Not reported
February 2012 4-91 Draft – Do Not Cite or Quote

Reference Study Methods Prior Pb Exposure

Bone Pb 

biomarker Bone Pb Conc. (μg/g)

Distribution of 

Bone Pb (μg/g)

Peters et al.

(2007)

Cohort: Normative Aging Study cohort

Age (yrs): Mean: 66.9

N: 513 male cases

Location: Boston, MA

Study Period: 1991-1996

Cumulative Tibia

Patella

Mean ± SD

Tibia: 21.5 ± 13.4

Patella: 31.5 ± 19.3

Not reported

Rajan et al.

(2007)

Cohort: VA Normative Aging Study Cohort

Age (yrs): Mean: 67.5 (at bone scan)

N: 1075 males

Location: Boston, MA

Study Period: 1991-2002

Not reported Tibia

Patella

Mean ± SD

Tibia: 22.1 ± 13.8

Patella: 31.4 ± 19.6

Not reported

Rajan et al.

(2008)

Cohort: VA Normative Aging Study Cohort

Age (yrs): ≥ 45

N: 720 males

Location: Boston, MA

Study Period: 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)

Cohort: VA Normative Aging Study Cohort

Age (yrs): Mean: 67.1

N: 526 males

Location: Boston, MA

Study Period: 1/1/1991-12/31/1995

Not 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)

Cohort: Not reported

Age (yrs): 30-60

N: 68 males

Location: Belgium

Study Period: 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. (2002b), as 

reported in 

Navas-Acien 

et al. (2008)

Cohort: Not reported

Age (yrs): 15-44; Mean ± SD: 31.0 ± 7.7

N: 720 females

Location: Los Angeles, CA

Study Period: 6/1995-5/2001

Not reported Tibia

Calcaneus

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)

Cohort: Baltimore Memory Study cohort

Age (yrs): Mean: 59.39

N: 985

Location: Baltimore, MD

Study Period: Not reported

Not reported Tibia Mean ± SD:

Tibia: 18.7 ± 11.2

Not reported
February 2012 4-92 Draft – Do Not Cite or Quote

Reference Study Methods 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)

Cohort: Not reported

Age (yrs): 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)

N: 257 cases; 276 controls

Location: Silver Valley, ID; Spokane, WA

Study Period: 7/10/1994-8/7/1994

Cumulative (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 

Wijngaarden 

et al. (2009)

Cohort: Not reported

Age (yrs): Mean: 61.5

N: 47

Location: Rochester, NY

Study Period: Not reported

Cumulative Tibia

Calcaneus

Mean ± SD

Tibia: 2.0 ± 5.2

Calcaneus: 6.1 ± 8.5

Not reported

Wasserman et 

al. (2003)

Cohort: Yugoslavia Prospective Study of 

Environmental Pb Exposure

Age (yrs): 10-12

N: 167 children

Location: Kosovska, Mitrovica, Kosovo, 

Yugoslavia; Pristina, Kosovo, Yugoslavia

Study Period: 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 of paper 

distributes tibia Pb by 

sex, ethnicity, 

address at birth 

relative to factory, and 

maternal education

Weisskopf et 

al. (2004), as 

reported in 

Shih et al.

(2007)

Cohort: Normative Aging Study

Age (yrs): Mean ± SD: 67.4 ± 6.6

N: 466 males

Location: Boston, MA

Study Period: 1991-2002

Environmental Tibia

Patella

Median (IQR)

Tibia: 19 (12,26)

Patella: 23 (15, 35)

Tibia IQR: 14

Patella IQR: 20

Table 3 of paper 

shows mean Pb 

levels across categorical variables (yr of 

education, smoking 

status, computer 

experience, first 

language English)

Weisskopf et 

al. (2007b)

Cohort: VA Normative Aging Study cohort

Age (yrs): Mean: 

Lowest Patella quintile: 73.2; 

Highest Patella quintile: 80.7

N: 31 males

Location: Boston, MA

Study Period: Bone Pb measured: 

1994-1999 Scans performed: 2002-2004

Not reported Tibia

Patella

Median (IQR)

Tibia

Lowest quintile: 13 (9-17)

Highest quintile: 41 (38-59)

Patella

Lowest quintile: 9 (5-15)

Highest quintile: 63 (43-86)

Not reported

Weisskopf et 

al. (2007a)

Cohort: VA Normative Aging Study cohort

Age (yrs): Mean: 68.7

N: 1,089 males

Location: Boston, MA

Study Period: 1993-2001

Concurrent and cumulative Tibia

Patella

Median (IQR)

Tibia: 20 (13-28)

Patella: 25 (17-37)

Table 1 of paper 

shows distribution of 

Pb biomarkers by 

categories of covariates (age, education, 

smoking status, alco- hol intake, physical 

activity, computer 

experience, first 

language English)
February 2012 4-93 Draft – Do Not Cite or Quote

Reference Study Methods Prior Pb Exposure

Bone Pb 

biomarker Bone Pb Conc. (μg/g)

Distribution of 

Bone Pb (μg/g)

Weisskopf et 

al. (2009)

Cohort: Normative Aging Study; (95% white)

Age (yrs): Mean ± SD (at Patella baseline); 

Tertile 1: 65.2 ± 7.1; 

Tertile 2: 66.5 ± 6.5

Tertile 3: 70.2 ± 7.2

N: 868 males

Location: Greater Boston area, MA

Study Period: 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)

Cohort: BUMC, BWH, BIDMC, HVMA, 

Normative Aging Study (NAS), Harvard 

Cooperative Program on Aging (HCPOA)

Age (yrs): Mean:

Cases: 66.5; Controls: 69.4

N: 330 cases; 308 controls

Location: Boston, MA

Study Period: 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

Weuve et al.

(2006)

Cohort: VA Normative Aging Study cohort

Age (yrs): ≥ 45

N: 720 males

Location: Boston, MA

Study Period: 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 of paper 

shows distribution of 

mean Pb biomarker 

levels by 

characteristics of 

participants (age, 

education, computer 

experience, smoking 

status, alcohol 

consumption, tertile of 

calcium intake, tertile 

of physical activity, 

diabetes)

Weuve et al.

(2009)

Cohort: Nurses’ Health Study cohort

Age (yrs): 47-74

N: 587 females

Location: Boston, MA

Study Period: 1995-2005

Recent and cumulative Tibia

Patella

Mean ± SD:

Tibia: 10.5 ± 9.7

Patella: 12.6 ± 11.6

Not reported

Wright et al.

(2003b), as 

reported in 

Shih et al.

(2007)

Cohort: Normative Aging Study

Age (yrs): Mean ± SD: 68.2 ± 6.9

N: 736 males

Location: Boston, MA

Study Period: 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 lowest-highest 

quartile: 47
February 2012 4-94 Draft – Do Not Cite or Quote

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

occupationally exposed populations

Reference Study Methods 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)

Cohort: Canada Lead Study

Age (yrs): 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

N: 80 males

Location: Canada

Study Period: 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.

(2007a)

Cohort: Not reported

Age (yrs): Mean: 39.7

N: 61

Location: Northern Canada

Study Period: Not Reported

Occupational (primary Pb 

smelter workers)

Tibia Mean:

Tibia: 38.6

Not reported

Caffo et al.

(2008)

Cohort: Not reported 

Age (yrs): Mean: 60.39

N: 513 males

Location: Delaware and New Jersey, U.S.

Study Period: 1994-1997 

(Phase 1 recruitment); 

2001-2003 (Phase 2 recruitment)

Cumulative

Occupational

(Former organolead 

manufacturing workers)

Tibia Mean ± SD:

Peak Tibia: 23.99 ± 18.46

Not reported

Dorsey et al.

(2006)

Cohort: Not reported

Age (yrs): Mean: 43.4

N: 652

Location: Korea

Study Period: 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)

Cohort: Not reported

Age (yrs): 40-70; Mean: 55.8 (baseline)

N: 496 males

Location: Eastern U.S.

Study Period: 6/1994-6/1996 (enrolled); 

6/1998 (follow-up period ended)

Occupational

(Chemical manufacturing 

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)

Cohort: Not reported

Age (yrs): 0-36.2 (baseline); 

Mean ± SD: 41.4 ± 9.5 (baseline)

N: 575; (76% male; 24% female)

Location: South Korea

Study Period: 10/1997-6/2001

Cumulative 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

Hanninen et 

al. (1998), as 

reported in 

Shih et al.

(2007)

Cohort: Not reported

Age (yrs): 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

N: 54; (43 males, 11 females)

Location: Helsinki, Finland

Study Period: 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
February 2012 4-95 Draft – Do Not Cite or Quote

Reference Study Methods Prior Pb Exposure

Bone Pb 

biomarker

Bone Pb Concentration 

(μg/g)

Distribution of Bone 

Pb (μg/g)

Hsieh et al. 

2009 (2009b)

Cohort: Not reported

Age (yrs): Mean:

Cases: 45.71

Controls: 46.06

N: 22 cases; 18 controls

Location: Not Reported

Study Period: 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)

Cohort: Not reported

Age (yrs): 30-80

N: 109 cases; 256 controls; 

(Bone samples collected from 

104 cases and 41 controls)

Location: New England (Boston, MA)

Study Period: 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.

(2009b)

Cohort: 1982 Pb Occupational 

Study cohort

Age (yrs): Mean:

Cases: 54

Controls: 55

N: 83 cases; 51 controls

Location: Eastern Pennsylvania

Study Period: 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)

Cohort: Not reported

Age (yrs): Median: 41.5

N: 38 male cases; 19 male controls

Location: Not reported

Study Period: 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

Roels et al.

(1994)

Cohort: Not reported

Age (yrs): 30-60

N: 76 male cases; 68 male controls

Location: Belgium

Study Period: 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 
February 2012 4-96 Draft – Do Not Cite or Quote

Reference Study Methods Prior Pb Exposure

Bone Pb 

biomarker

Bone Pb Concentration 

(μg/g)

Distribution of Bone 

Pb (μg/g)

Schwartz

(2000c) et al., 

as reported in 

Shih et al.,

(2007)

Cohort: U.S. Organolead Study

Age (yrs): Mean ± SD:

Cases: 55.6 ± 7.4

Controls: 58.6 ± 7.0

N: 535 male cases

118 male controls

Location: Eastern U.S.

Study Period: 6/1994-10/1997 (enrolled); 

Completed 2-4 annual follow-up visits; 

Tibia Pb taken in 3rd year

Occupational

(tetraethyl and tetramethyl 

Pb manufacturing facility)

Tibia Mean ± SD

Current Tibia: Cases: 14.4 ± 

9.3

Peak Tibia:

Cases: 22.6 ± 16.5

Not reported

Schwartz et 

al. (2000b), as 

reported in 

Navas-Acien 

et al. (2008)

Cohort: Not reported

Age (yrs): 41.7-73.7 (Combined)

Mean ± SD: 

Combined: 57.6 ± 7.6

Hypertensive: 60.2 ± 6.9

Nonhypertensive: 56.6 ± 7.5

N: 543 males

Location: Eastern U.S.

Study Period: 1995 (recruited); 

1996-1997 (Tibia Pb 

taken during the 3rd yr)

Occupational

(former organolead 

manufacturing 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.

(2001a)

Cohort: Not reported

Age (yrs): Mean:

Exposed: 40.4

Control: 34.5

N: 803 cases; 135 controls

Location: South Korea

Study Period: 10/24/1997-8/19/1999

Occupational (battery 

manufacturing, secondary 

smelting, Pb oxide 

manufacturing, car radiator 

manufacturing)

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)

Cohort: Not reported

Age (yrs): Mean at 1st visit: 41.4

N: 576

Location: South Korea

Study Period: 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

Stewart et al.

(1999), as 

reported in 

Shih et al.,

(2007)

Cohort: U.S. Organolead Study

Age (yrs): 40-70 (in 1995)

38% ≥ 60 yrs

Mean: 58

N: 534 males

Location: Eastern U.S.

Study Period: Not Reported

Occupational

(tetraethyl and tetramethyl 

Pb manufacturing 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)

Cohort: Not reported

Age (yrs): Mean: 56.1

N: 532 males

Location: Eastern U.S.

Study Period: 1994-1997; 2001-2003

Cumulative

Occupational

(Organolead workers - not 

occupationally 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)

Cohort: Not reported

Age (yrs): Mean ± SD: 43.3 ± 9.8

N: 652

Location: South Korea

Study Period: 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
February 2012 4-97 Draft – Do Not Cite or Quote

1 Urine Pb concentrations in the U.S. general population have been monitored in the 

2 NHANES. Data from the most recent survey (CDC, 2011) are shown in Table 4-10. The 

3 geometric mean for the entire sample for the period 2007-2008 (n = 2,627) was 0.52 μg/g 

4 creatinine (95% CI: 0.48, 0.55). The geometric means for males (n = 1,327) and females 

5 (n = 1,300) were 0.50 μg/g creatinine (95% CI: 0.47, 0.53) and 0.53 μg/g creatinine (95% 

6 CI: 0.49, 0.57), respectively.

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

Survey Stratum Period Geometric Mean (μg/g CR)a 95% Confidence Interval Number of Subjects

All

1999-2000 0.721 0.700, 0.742 2,465

2001-2002 0.639 0.603, 0.677 2,689

2003-2004 0.632 0.603, 0.662 2,558

2005-2006 0.546 0.502, 0.573 2,576

2007-2008 0.515 0.483, 0.549 2,627

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 1,406

2001-2002 0.658 0.617, 0.703 1,559

2003-2004 0.641 0.606, 0.679 1,543

2005-2006 0.573 0.548, 0.600 1,520

2007-2008 0.546 0.513, 0.580 1,857

Males

1999-2000 0.720 0.679, 0.763 1,227

2001-2002 0.639 0.607, 0.673 1,334

2003-2004 0.615 0.588, 0.644 1,281

2005-2006 0.551 0.522, 0.582 1,271

2007-2008 0.502 0.471, 0.534 1,327

Females

1999-2000 0.722 0.681, 0.765 1,238

2001-2002 0.639 0.594, 0.688 1,355

2003-2004 0.648 0.601, 0.698 1,277

2005-2006 0.541 0.507, 0.577 1,305

2007-2008 0.527 0.489, 0.568 1,300

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
February 2012 4-98 Draft – Do Not Cite or Quote

Survey Stratum Period Geometric Mean (μg/g CR)a 95% Confidence Interval Number of Subjects

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 1,132

2003-2004 0.623 0.592, 0.655 1,074

2005-2006 0.541 0.500, 0.585 1,041

2007-2008 0.506 0.466, 0.550 1,095

a

Values are μg Pb/g creatinine (CR)

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).

2 Tooth enamel samples from 127 subjects born between 1936 and 1993 were analyzed for 

3 Pb concentration and Pb isotope ratios of the tooth enamel and compared with those 

4 parameters for sediment cores and estimates of Pb emissions from gasoline during the 

5 years when 50% enamel formation was estimated to occur. They found that the log6 transform of tooth enamel concentration was significantly predicted by the log-transform 

7 of Lake Erie sediment core data obtained by Graney et al. (1995) (p <0.00001) and by the 

8 log-transform of U.S. consumption of Pb in gasoline (p <0.00001); Figure 4-20. 

Additionally, Robbins et al. (2010) found that 207Pb/206 9 Pb was significantly predicted by 

the 207Pb/206 10 Pb observed in the Lake Erie sediment cores obtained by Graney et al. (1995)

11 (p <0.0001) and for this study (p <0.0002).
February 2012 4-99 Draft – Do Not Cite or Quote

Source: Reprinted with permission of Elsevier Publishing, Robbins et al. (2010).

Note: (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.

Figure 4-20 Comparison of relative temporal changes in tooth enamel Pb 

concentration.

1 Several Brazillian studies have found increased levels of Pb in teeth in areas where Pb 

2 sources are present. For example, Costa de Almeida et al. (2007) reported Pb 

3 concentration in tooth enamel among 4-6 y old kindergarteners in São Paulo, Brazil to be 

4 significantly higher for children living near a Pb-acid battery processing plant in the 

5 Baruru neighborhood compared with 4-6 y old children in other parts of the city 

6 (non-exposed median: 206 mg/kg, n = 247; exposed median: 786 mg/kg, n = 26; p < 

7 0.0001). Subsequent analysis revealed that 55% of 4-6 y old children from Baruru had 

8 tooth enamel Pb concentrations greater than 600 mg/kg, forming a significant comparison 

9 with other neighborhoods having 0-33% of 4-6 y old children with tooth enamel Pb 

10 greater than 600 mg/kg (p < 0.0001) (de Almeida et al., 2008). The authors did not 

11 describe controlling for additional factors, such as socioeconomics or housing conditions. 

12 Arruda-Neto et al. (2009) studied Pb in tooth samples among São Paulo children to 

13 compare exposures of children age 4-12 living near a dam with heavy metal sediments 

14 with those of children ages 4-13 living in a control area thought to have few exposures. 

15 They observed a significant comparison (near dam: avg 1.28 ± 0.11 mg/kg, n = 50; 

16 control region: avg 0.91 mg/kg, n = 24). In a related study of Pb measures in teeth among 

17 the general population ages 7-60 y, Arruda-Neto et al. (2009) observed peak Pb 

18 concentrations of 115% of age 7 y concentrations at age 10 y followed by a steep drop to 

19 50% among 20 y old subjects. Tooth Pb concentrations stayed fairly constant throughout 

20 adulthood but then dropped to just above 30% among 65 y old subjects. Note that the 
February 2012 4-100 Draft – Do Not Cite or Quote

1 authors did not clarify if average or median values were presented, nor did they adjust for 

2 potentially confounding factors.

4.5 Empirical Models of Lead Exposure-Blood Lead Relationships

3 Multivariate regression models, commonly used in epidemiology, provide estimates of 

4 the contribution of variance in the internal dose metric to various determinants or control 

5 variables (e.g., air Pb concentration, surface dust Pb concentration). Structural equation 

6 modeling links several regression models together to estimate the influence of 

7 determinants on the internal dose metric. Regression models can provide estimates of the 

8 rate of change of blood or bone Pb concentration in response to an incremental change in 

9 exposure level (i.e., slope factor). One strength of regression models is that they are 

10 empirically verified within the domain of observation and have quantitative estimates of 

11 uncertainty imbedded in the model structure. However, regression models are based on 

12 (and require) paired predictor-outcome data, and, therefore, the resulting predictions are 

13 confined to the domain of observations and are typically not generalizable to other 

14 populations. Regression models also frequently exclude numerous parameters that are 

15 known to influence human Pb exposures (e.g., soil and dust ingestion rates) and the 

16 relationship between human exposure and tissue Pb levels, parameters which are 

17 expected to vary spatially and temporally. Thus, extrapolation of regression models to 

18 other spatial or temporal contexts, which is often necessary for regulatory applications of 

19 the models, can be problematic.

4.5.1 Air Lead-Blood Lead Relationships

20 The 1986 Pb AQCD (U.S. EPA, 1986a) described epidemiological studies of 

21 relationships between air Pb and blood Pb. Of the studies examined, the total blood Pb-air 

22 Pb slope (when considering both direct and indirect exposures derived from air) was 

23 estimated to be approximately double the slope estimated from the direct contribution due 

24 to inhaled air alone (U.S. EPA, 1986a). Much of the pertinent earlier literature (e.g., prior 

25 to 1984) was summarized by Brunekreef (1984). Based on the studies available at that 

26 time that considered multiple air-related Pb exposure pathways in the aggregate, the 1986 

27 Pb AQCD concluded that “the blood Pb versus air Pb slope β is much smaller at high 

28 blood and air levels.” This is to say that the slope β was much smaller for occupational 

29 exposures where high blood Pb levels (>40 μg/dL) and high air Pb levels (much greater 

than 10 μg/m3 30 ) prevailed relative to lower environmental exposures which showed lower 

blood Pb and air Pb concentrations (<30 μg/dL and <3 μg/m3 31 ). For those environmental 

32 exposures, it was concluded that the relationship between blood Pb and air Pb “for direct 
February 2012 4-101 Draft – Do Not Cite or Quote

1 inhalation appears to be approximately linear in the range of normal ambient exposures 

(0.1-2.0 μg/m3 2 )” (pp 1–98 of the 1986 Pb AQCD). Based on meta-analysis of 18 studies of 

3 urban or industrial-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 4 = 0.69; Figure 4-21). 

This slope corresponds to an increase of 4.6 μg/dL blood Pb per μg/m3 5 air Pb at an air Pb 

concentration of 1.5 μg/m3 for all groups included in the study (n=96). The 1.5 μg/m3 6 value 

7 is the median of the air Pb concentrations that match the 96 blood Pb concentrations in 

8 Figure 3 of Brunekreef et al. (1984), taken from the Appendix to the same paper. When the 

9 analysis was limited to children whose blood Pb concentrations were <20 μg/dL, the slope 

was 0.2159 (R2=0.33), which corresponds to an increase of 4.8 μg/dL blood Pb per μg/m3 10 air 

Pb at the median air concentration (0.54 μg/m3, n=43)

12 Newer studies that provide estimates for the blood Pb-air Pb slope are described below. In 

13 some studies, the blood Pb-air Pb relationship was described with a non-linear regression 

14 function, in which the blood Pb-air Pb slope varied with air Pb concentration. In Table 4-11, 

15 slopes corresponding to the central estimate of the air Pb concentrations are provided, which 

16 are considered to be the best estimate of the slope from each study. These were calculated by 

evaluating each regression function at ± 0.01 μg/m

11 . 

3 17 from the central estimate of the air Pb 

18 concentration. Air Pb concentration ranges and central estimates varied across studies, 

19 making it difficult to interpret comparisons based solely on the central estimates of the 

20 slopes; therefore, Figure 4-21 depicts the relationship between the blood Pb-air Pb slope as a 

21 function of air Pb concentration for the range of air Pb concentrations evaluated in each 

22 study (the central estimate is also shown). Figure 4-21 provides a more informative picture 

23 of the extent to which slope estimates vary (and overlap) within and between studies. In the 

24 Schnaas et al. (2004) analysis, the effect of air Pb on blood Pb may have been 

25 underestimated due to inclusion of location and SES terms in their regression model. It was 

26 specifically noted by the authors that air Pb differed significantly between the locations and 

27 the poorer residential areas were usually the more industrialized areas with higher pollution. 

28 Hence, the inclusion of these terms may have accounted for some of the variance in blood Pb 

29 attributable to air Pb. With the exception of Ranft et al. (2008), all studies included in Table 

30 4-11 and Figure 4-22 include a blood Pb-air Pb relationship that reflects all air-related 

31 pathways of exposure. The Ranft et al. (2008)study includes a separate term for soil Pb, so 

32 the blood Pb-air Pb slope presented for that study underestimates the slope that would reflect 

33 all air-related pathways, since soil Pb encompasses deposited ambient air Pb. The Ranft et al.

34 (2008) model is log-linear, with the natural logarithm of blood Pb being a function of linear 

35 increase in air Pb. This results an upward curvature of the blood Pb-air Pb relationship. By 

36 comparison, log-log models predict an increase in the blood Pb-air Pb slope with decreasing 

37 air Pb concentration, whereas linear models predict a constant blood Pb-air Pb slope across 

38 all air Pb concentrations. 
February 2012 4-102 Draft – Do Not Cite or Quote

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 Slopea

Children Populations

Brunekreef et al. (1984)

Location: Various countries

Years: 1974-1983

Subjects: Children (varying age ranges; n>190,000)

Analysis: Meta analysis of 18 studies

Model: Log-Log

Blood Pb: 5-41 μg/dL 

(mean range for studies)

Air Pb: 0.1-24 μg/m3

(mean range for studies)

All children:

4.6 (1.5)c

Children <20 

μg/dL: 4.8 (0.54)d

Hayes et al. (1994)

Location: Chicago, IL

Years: 1974-1988

Subjects: 0.5-6 yr (n = 9,604)

Analysis: Regression of quarterly median blood Pb 

and quarterly mean air Pb 

Model: Log-Log

Blood Pb: 12-30 μg/dL 

(annual median range)

Air Pb: 0.05-1.2 μg/m3

(annual mean range)

8.2 (0.62)e

Hilts et al. (2003)

Location: Trail, BC

Years: 1989-2001

Subjects: 0.5-6 yr (Estimated n = 220-460, based on 292-536 blood 

Pb measurements/yr with 75-85% participation)

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 geometric mean range)

Air Pb: 0.03-1.1 μg/m3

(annual geometric mean range)

6.5 (0.48)f

Ranft et al. (2008)

Location: Germany

Years: 1983-2000

Subjects: 6-11 yr (n = 843) 

Analysis: Pooled regression 5 cross-sectional studies

Model: multivariate Log-Linear 

Blood Pb: 2.2-13.6 μg/dL 

(5th-95th percentile)

Air Pb: 0.03-0.47 μg/m3

(5th-95th percentile)

3.2 (0.1)g

Schnaas et al. (2004)

Location: Mexico City

Years: 1987-2002

Subjects: 0.5-10 yr (n = 321)

Analysis: Regression of lifetime blood Pb from longitudinal 

blood Pb measurements and annual average air Pb data

Model: Log-Log

Blood Pb: 5-12 μg/dL 

(annual GM range)

Air Pb: 0.07-2.8 μg/m3

(annual mean range in yr of birth)

2.2 (0.4)h

Schwartz and Pitcher

(1989), U.S. EPA (1986a)

Location: Chicago, IL

Years: 1976-1980

Subjects: Black children, 0-5 yr (n = 5,476)

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)b

8.6 (0.75)i

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 (0.45)j

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)

0.01 (58)k
February 2012 4-103 Draft – Do Not Cite or Quote

Mixed Child-Adult Populations

Schwartz and Pitcher

(1989), U.S. EPA (1986a)

Location: U.S.

Years: 1976-1980

Subjects: 0.5-74 yr (n = 9,987)

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 (0.75)l

a

Slope is predicted change in blood Pb (μg/dL per μg/m3) evaluated at ± 0.01 μg/m3 from central estimate of air Pb for the study (shown in parentheses) b

Based on data for U.S. (1986 Pb AQCD) c ln(PbB) = ln(PbA) x 0.3485 +2.853

d ln(PbB) = ln(PbA) x 0.2159 + 2.620 e ln(PbB) = ln(PbA) x 0.24 +3.17 f PbB = PbA x 6.5

g PbB =1.5 x EXP(0.9361 x (PbA-0.1)/0.44), where 1.5 μg/dL is the background PbB, and 0.1 μg/m3 is the median PbA for the study; model also adjusted for 

soil Pb concentration, which may reduce estimated slope

h ln(PbB) = Ln(PbA) x 0.213 +1.615 for the 1987 cohort, see text for more study details. i PbB = PbA x 8.6

j PbB =Pb A x 3.6

k ln(PbB) = ln(PbA) x 0.05 +2.12

l PbB = PbA x 9.63

GM, geometric mean; GSD, geometric standard deviation; PbB, blood Pb concentration (μg/dL); PbA, air lead concentration (μg/m3)

Data provided from Brunekreef (1984).

Note: The regression model is: (ln[μg/dL blood Pb] = 0.3485⋅ln[μg/m3 air Pb] + 2.85) for all children (n=96 subject groups) 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 (n=44 subject groups).

Figure 4-21 Predicted relationship between air Pb and blood Pb based on a 

meta analysis of 18 studies.

0

10

20

30

40

50

60

0 5 10 15 20 25

Blood Pb (

μg/dL)

Air Pb (μg/m3)

All children <20 μg/dL
February 2012 4-104 Draft – Do Not Cite or Quote

Note: Slopes are calculated for a change in air Pb (±0.01 μg/m3) over ranges of air Pb concentrations reported in each study (lines). 

The air Pb axis is truncated at 2 μg/m3

; the actual range for the Brunkreefe et al. (1984) study was 0.1-6.4 μg/dL per μg/m3 and for 

the Schnaas et al. (2004) study was 0.08-2.8 μg/dL per μg/m3

. The slope axis has been truncated at 40; the actual range for the 

Hayes et al. (1994) study was 5-56 μg/dL per μg/m3 (the high end of the range was estimated for the minimum annual average air 

Pb of 0.05 μg/m3

). The two estimates for Schwartz and Pitcher (1989) represent data for U.S. and Chicago. Models are log-log (solid 

lines), log-linear (dashed line), and linear (dotted lines). Symbols show the slope at the central estimate of air Pb (e.g., mean or 

median reported for each study).

Figure 4-22 Blood Pb – air Pb slope (μg/dL per μg/m3) predicted from various 

epidemiologic studies (links available in Table 4-11).

4.5.1.1 Children

1 Hilts et al. (2003) reported child blood Pb and air Pb trends for the city of Trail, British 

2 Columbia, over a period preceding and following installation of a new smelter process in 

3 1997 which resulted in lower air Pb concentrations. Blood Pb data were obtained from 

4 annual (1989-2001) surveys of children 6-60 months of age who lived within 4 km from 

5 the smelter (n: 292-536 eligible per year, 75-85% participation). Air Pb concentrations 

6 were obtained from high volume suspended particulate samplers placed within 2 km of 

7 the smelter that operated 24 hours every 6th day. Data on Pb levels in air, residential soil, 

8 interior dust, and blood for three sampling periods are summarized in Table 4-12. Based 

on these data, blood Pb decreased 6.5 μg/dL per 1 μg/m3 9 air Pb and by 0.068 μg/dL per 

10 mg/kg soil Pb (based on linear regression with air or soil Pb as the sole independent 

11 variable). Several uncertainties apply to these estimates. Potential mismatching of air Pb 

12 concentrations (often termed misclassification) with individual blood Pb levels may have 

13 occurred as a result of air Pb being measured within 2 km of the smelter, whereas, the 

0

10

20

30

40

0.0 0.5 1.0 1.5 2.0

Blood Pb Slope (μg/dLper μg/m3)

Air Pb (μg/m3)

Brunekreef (1984) 

<20 μg/dL

Hayes (1994)

Hilts (2003)

Ranft(2008)

Schnaas (2004)

Schwartz (1989) 

Chicago

Schwartz (1989) 

U.S.

Tripathi (2001)
February 2012 4-105 Draft – Do Not Cite or Quote

1 blood Pb data included children who resided >2 km from the smelter. The regression 

2 estimates were based on group mean estimates for three sampling dates, rather than on 

3 the individual blood Pb estimates, which included repeated measures on an unreported 

4 fraction of the sample. The limited number of data pairs (three) constrained parameter 

5 estimates to simple regression coefficients. Other important factors probably contributed 

6 to blood Pb declines in this population that may have been correlated with air, soil and 

7 dust Pb levels. These factors include aggressive public education and exposure 

8 intervention programs (Hilts et al., 1998; Hilts, 1996). Therefore, the coefficients shown 

9 in Table 4-12 are likely to overestimate the influence of air, dust, or soil Pb on blood Pb 

10 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 0.069 ± 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 geometric means; 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).

11 Ranft et al. (2008) reported a meta-analysis of five cross-sectional surveys of air and soil 

12 Pb levels and blood Pb concentrations in children living in Duisburg, Germany. The 

13 analysis included observations on 843 children (6-11 years of age) made during the 

14 period 1983-2000. Pb was measured in PM10 samples collected in a 200 meter by 200 

15 meter grid that encompassed the city. Pb in surface soil (0-10 cm) was measured at 145 

16 locations in the city. Air and soil Pb concentrations were assigned to each participant by 

17 spatial interpolation from the sampling grid data to each home residence. The 5th-95th 

percentile ranges were 0.025-0.465 μg Pb/m3 18 for air and 72-877 mg Pb/kg for soil. The 

19 results of multivariate regression analyses were reported in terms of the relative increase 

20 (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 21 = 0.586) that 

22 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 23 =0.484, p <0.001) 

and 1.30 per 800 mg/kg soil Pb (95% CI: 1.19, 1.43, R2 24 = 0.017, p <0.001). Based on the 

values for R2 25 , the regression model accounted for approximately 59% of the total 

26 variance in blood Pb and, of this, 83% was attributed to air Pb. Values for GMR for soil 
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1 Pb varied depending on the sampling data and ranged from 1.41 to 2.89, with most recent 

2 data (from the year 2000) yielding a value of 1.63 per 800 mg/kg increase in soil Pb. The 

3 GMR values can be converted to regression slopes (slope = [starting blood 

4 Pb×ln(GMR)]/[95th − 5th percentile air 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 5 increase in 

air Pb at the median air Pb concentration for the study (0.1 μg/m3 6 ) and assuming a 

7 background blood Pb concentration of 1.5 μg/dL. Based on the GMR estimate of 1.63 for 

8 soil Pb, a 1,000 mg/kg increase in soil Pb would be associated with an increase in blood 

9 Pb of 0.9 μg/dL per mg/kg soil at the median soil Pb concentration of 206 mg/kg and 

10 assuming a background blood Pb concentration of 1.5 μg/dL. The degree of confounding 

11 of the GMR and estimates resulting from the air and soil Pb correlation was not reported, 

12 although the correlation coefficient for the two variables was 0.136 for the whole data set 

13 and 0.703 when data collected in 1983 was omitted. Because the model also included Pb 

14 levels in soil, the blood Pb-air Pb ratio may be underestimated since some of the Pb in 

15 soil was likely derived from air. The blood Pb-air Pb slope does not include the portion of 

16 the soil/dust Pb ingestion pathway that derives from air Pb, such as recently airborne Pb 

17 deposited to soil and dust which remains available for inhalation and ingestion. To 

18 estimate the blood Pb-air Pb ratio that included all air-related pathways, data for 

19 geometric mean of blood Pb and air Pb among the cohort of children studied were 

20 extracted from Figure 1 in Ranft et al. (2008) for each of the five study years. The 

21 extracted values of the geometric mean of blood Pb and air Pb were used in regressions 

22 employing linear and log-log fits. The linear model obtained was: PbB = (13.65×PbA) + 

2.96 (R2 23 = 0.92); i.e., the linear regression produced a constant slope of 14 μg/dL per 

μg/m3. The log-log model was: ln(PbB) = (0.48×ln(PbA) + 2.61 (R2 24 = 0.91), resulting in 

an inverse curve for dPbB/dPbA vs. PbA with a slope of 22 μg/dL per μg/m3 25 at PbA = 

0.1 μg/m3 26 .

27 Schnaas et al. (2004) analyzed data on blood Pb and air Pb concentrations during and 

28 after the phase out of leaded gasoline use in Mexico (1986-1997) in children as part of a 

29 prospective study conducted in Mexico City. The sample included 321 children born 

30 during the period 1987 through 1992. Repeated blood Pb measurements were made on 

31 each child at 6-month intervals up to age 10 years. Air Pb measurements in PM10 (annual 

32 average of quarterly means) were derived from three area monitors which represented 

33 distinct study zones. Children were assigned to study zones based on their current address 

34 and were assigned the corresponding annual average air Pb concentrations for their year 

35 of birth and appropriate air monitoring zones. Associations between lifetime (across the 

36 first 10 years of life) blood Pb concentration, air Pb concentration for year of birth and 

37 other variables (e.g., age, year of birth, family use of glazed pottery) were evaluated 

38 using multivariate regression models. 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 39 ); the 
February 2012 4-107 Draft – Do Not Cite or Quote

1 predicted slope for this group of children was 0.213 (95% CI: 0.114-0.312) ln [μg/dL 

blood] per ln[μg/m3 2 air]. This slope corresponds to an increase of 2.2 μg/dL blood Pb per 

μg/m3 at the median annual air Pb concentration of 0.4 μg/m3 3 estimated over the years of 

4 the study. Slopes for other birth cohorts ranged from -0.003 (1992) to 0.166 (1988) with a 

5 median of 0.153 for all six cohorts. This median slope corresponds to an increase of 

1.7 μg/dL blood Pb per μg/m3 at the median annual air Pb concentration of 0.4 μg/m3 6

7 estimated over the years of the study. Considering all cohorts simultaneously, data for 

8 annual geometric mean of blood Pb and air Pb were extracted from Figure 1 in Schnaas et 

9 al. (2004). However, in employing this approach, blood Pb is confounded by age and year 

10 because in the early years of the study, only younger children were available and in the 

11 later years of the study, only older children contributed data. The extracted values of the 

12 geometric mean of blood Pb and mean air Pb were used in regressions employing linear 

13 and log-log models for comparison to other studies and a log-linear model as employeed 

by the authors. The linear model obtained was: PbB = (2.50×PbA) + 5.61 (R2 14 = 0.84), 

i.e., the linear model produced a constant slope of 2.50 μg/dL per μg/m3 15 . However, 

16 inspection of the graph (not shown here) suggested a bi-linear fit. Regression of the data 

over the interval 0.1-0.4 μg/m3 produced a slope of 9.0 μg/dL per μg/m3 (R2 17 = 0.83), and 

regression of the data over the interval 0.4-2.8 μg/m3 18 produced a slope of 1.52 μg/dL per 

μg/m3 (R2 = 0.83). The log-log model was: ln(PbB) = (0.26×ln(PbA)) + 2.20 (R2 19 = 0.94), 

20 resulting in an inverse curve for dPbB/dPbA vs. PbA, with a slope of 4.5 μg/dL per 

μg/m3 at PbA = 0.4 μg/m3 21 . The log-linear model as employed by the authors was: 

ln(PbB) = (0.32×PbA) + 1.73 (R2 22 = 0.77); and described the data least well. The log23 linear model produced an exponential curve of dPbB/dPbA vs. PbA, with a slope of 

2.04 μg/dL per μg/m3 at PbA = 0.4 μg/m3 24 .

25 Schwartz and Pitcher (1989) reported a multivariate regression analysis of associations 

26 between U.S. gasoline Pb consumption (i.e., sales) and blood Pb concentrations in the 

27 U.S. population during the period 1976-1980 when use of Pb in gasoline was being 

28 phased out. Although this analysis did not directly derive a slope for the air Pb-blood Pb 

29 relationships, other analyses have shown a strong correlation between U.S. gasoline Pb 

30 consumption and ambient air Pb levels during this same period (U.S. EPA, 1986a).

31 Therefore, it is possible to infer an air Pb-blood Pb relationship from these data. Two 

32 sources of blood Pb data were used in Schwartz and Pitcher (1989): NHANES II 

33 provided measurements for U.S. children 6 months to 74 years of age (n = 9,996) during 

34 1976-1980, and the City of Chicago blood Pb screening program provided approximately 

35 7,000 blood Pb measurements in black children during 1976-1980. Gasoline Pb 

36 consumption was estimated as the product of monthly gasoline sales in the U.S. and 

37 quarterly estimates of Pb concentrations in gasoline reported to U.S. EPA. Based on the 

38 NHANES blood Pb data for white children, the regression coefficient was 2.14 μg/dL

39 blood per 100 metric tons of gasoline Pb/day (SE=0.19, p=0.0000); results for black 
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1 children were essentially identical. Based on the Chicago blood Pb data the regression 

2 coefficient was 16.12 (μg/dL per 1,000 metric tons gasoline Pb/quarter (SE=1.37, 

3 p=0.0001), which is roughly equivalent to 1.79 μg/dL blood per 100 metric tons of 

4 gasoline Pb/day (the value cited in Schwartz and Pitcher (1989) is 1.97 μg/dL blood per 

5 100 metric tons of gasoline Pb/day). U.S. EPA (1986a) reported data on gasoline Pb 

6 consumption (sales) and ambient Pb levels in the U.S. during the period 1976-1984 

7 (Table 4-13). Based on these data, air Pb concentrations decreased in association with 

8 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 9 = 0.95, p <0.0001). If this 

10 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 8.6 μg/dL per μg/m3 11 , 

12 based on the NHANES and Chicago data, respectively (e.g., 2.14/0.23 = 9.3 and 

13 1.97/0.23=8.6).

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 (1986a).

14 Tripathi et al. (2001) reported child blood Pb and air Pb trends for the city and suburbs of 

15 Mumbai, India over the period 1984-1996. Blood Pb data were obtained from children 

16 6-10 years of age (n = 544) who lived in 13 locations within the Mumbai area. Air Pb 

17 concentrations were measured from high volume PM samplers (with the majority of Pb in 

18 the respirable size range) placed at a height of 1.6 meters that operated 24 hours. Data on 

19 Pb concentrations in air, residential soil, interior dust, and blood for three sampling 

20 periods are summarized in Table 4-14. Based on these data, blood Pb increased 3.6 μg/dL 

per 1 μg/m3 21 air Pb (based on linear regression with air or soil Pb as the sole independent 

22 variable). Several uncertainties apply to these estimates, including potential exposure 

23 misclassification since the mean air Pb concentration was used for each suburb over the 

24 entire study period. The regression estimates were based on group mean blood Pb 
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1 estimates for the 13 sampling locations, rather than on the individual blood Pb estimates, 

2 which included repeated measures on an unreported fraction of the sample.

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).

3 Hayes et al. (1994) analyzed data collected as part of the Chicago, IL blood Pb screening 

4 program for the period 1974-1988, following the phase-out of leaded gasoline. The data 

5 included 9,604 blood Pb measurements in children (age: 6 months to 6 years) and 

6 quarterly average air Pb concentrations measured at 12 monitoring stations in Cook 

7 County, IL. Quarterly median blood Pb levels declined in association with quarterly mean 

8 air Pb concentrations. The regression model predicted a slope of 0.24 ln [μg/dL blood] 

per ln[μg/m3 9 air], as illustrated in Figure 4-23. This slope corresponds to an increase of 

8.2 μg/dL blood Pb per μg/m3 10 at the average annual mean air Pb concentration of 

0.62 μg/m3 11 .
February 2012 4-110 Draft – Do Not Cite or Quote

Modified from Hayes et al. (1994).

Note: The regression model is: (ln[μg/dL blood Pb] = 0.24⋅ln[μg/m3 air Pb] + 3.17).

Figure 4-23 Predicted relationship between air Pb and blood Pb based on data 

from Chicago, IL in children age 0-5 y (1974–1988).

4.5.1.2 Adults

1 Rodrigues et al. (2010) examined factors contributing to variability in blood Pb 

2 concentration in New England bridge painters, who regularly use electric grinders to 

3 prepare surfaces for painting. The study included 84 adults (1 female) who were observed 

4 during a 2-week period in 1994 or 1995. Subjects wore personal inhalable PM samplers 

5 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 6 (GSD 2.8), 

with a maximum daily value of 210 μg/m3 7 . These Pb concentrations were reported to 

8 have been corrected by the National Institute for Occupational Safety and Health 

9 (NIOSH) respirator protection factors, which were not reported by the authors. Hand 

10 wipe samples were collected at the mid-shift break and at the end of the shift (after the 

11 subjects had reportedly cleaned up for the day; GM = 793 μg, GSD 3.7). Blood Pb 

12 samples were collected at the beginning of the 2-week period (GM = 16.1 μg/dL, GSD 

13 1.7). Associations between exposure variables and blood Pb concentrations were 

14 explored with multivariate regression models (Table 4-15). When the model excluded 

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 hand-wipe data (not all participants who wore the personal air samplers agreed to provide 

2 hand-wipes), the regression coefficient for the relationship between ln[blood Pb 

concentration (μg/dL)] and ln[air Pb (μg/m3 3 )] was 0.11 (SE = 0.05, p = 0.03). This slope 

4 corresponds to a 1.3-fold increase in blood Pb concentration for a 10-fold increase in air 

Pb concentration and a slope of 0.009 μg/dL per μg/m3 5 , at the average occupational air 

Pb concentration for the study (58 μg/m3 6 ); non-occupational exposures were not included 

7 in the slope calculation. A second regression model included hand wipe Pb (n = 54) and 

8 yielded a regression coefficient of 0.05 (SE = 0.07, p = 0.45), which corresponds to a 

9 1.12-fold increase in blood Pb concentration per 10-fold increase in air Pb concentration 

and a slope of 0.02 μg/dL per μg/m3 10 , at the average occupational air Pb concentration for 

the study (58 μg/m3 11 ).

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).
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4.5.2 Environmental Lead-Blood Lead Relationships

1 Empirically-based relationships between blood Pb levels and Pb intakes and/or Pb 

2 concentrations in environmental media have provided the basis for what has become

3 known as slope factor models. Slope factor models are highly simplified representations 

4 of empirically based regression models in which the slope parameter represents the 

5 change in blood Pb concentration projected to occur in association with a change in Pb 

6 intake or uptake. The slope parameter is factored by exposure parameters (e.g., exposure 

7 concentrations, environmental media intake rates) that relate exposure to blood Pb 

8 concentration (Maddaloni et al., 2005; U.S. EPA, 2003c; Abadin and Wheeler, 1997; 

9 Stern, 1996; Bowers et al., 1994; Stern, 1994; Carlisle and Wade, 1992). In slope factor 

10 models, Pb biokinetics are represented as a linear function between the blood Pb 

11 concentration and either Pb uptake (uptake slope factor, USF) or Pb intake (intake slope 

12 factor, ISF). The models take the general mathematical forms:

PbB = E × ISF

Equation 4-2

PbB = E × AF × USF

Equation 4-3

13 where PbB is the blood Pb concentration, E is an expression for exposure (e.g., soil 

14 intake × soil Pb concentration) and AF is the absorption fraction for Pb in the specific 

15 exposure medium of interest. Intake slope factors are based on ingested rather than 

16 absorbed Pb and, therefore, integrate both absorption and biokinetics into a single slope 

17 factor, whereas models that utilize an uptake slope factor include a separate absorption 

18 parameter. In contrast to mechanistic models, slope factor models predict quasi-steady 

19 state blood Pb concentrations that correspond to time-averaged daily Pb intakes (or 

20 uptakes) that occur over sufficiently long periods to produce a quasi-steady state 

21 (i.e., >75 days, ~3 times the t1/2 for elimination of Pb in blood).

22 The U.S. EPA Adult Lead Methodology (ALM) is an example of a slope factor model 

23 that has had extensive regulatory use in the EPA Superfund program for assessing health 

24 risks to adults associated with non-residential exposures to Pb in contaminated soils

25 (Maddaloni et al., 2005; U.S. EPA, 1996a). The model was developed to predict maternal 

26 and fetal blood Pb concentrations that might occur in relation to maternal exposures to 

27 contaminated soils. The model assumes an uptake slope factor of 0.4 μg/dL blood per 

28 μg/day Pb uptake. Additional discussion of slope factor models that have been used or 

29 proposed for regulatory use can be found in the 2006 Pb AQCD (U.S. EPA, 2006b).
February 2012 4-113 Draft – Do Not Cite or Quote

1 Previous studies included in the 2006 Pb AQCD (U.S. EPA, 2006b) explored the 

2 relationship between blood Pb in children and environmental Pb concentrations. In a 

3 pooled analysis of 12 epidemiologic studies, interior dust Pb loading, exterior soil/dust 

4 Pb, age, mouthing behavior, and race were all statistically significant variables included 

5 in the regression model for blood Pb concentration (Lanphear et al., 1998). Significant 

6 interactions were found for age and dust Pb loading, mouthing behavior and exterior 

7 soil/dust level, and SES and water Pb level. In a meta-analysis of 11 epidemiologic 

8 studies, among children the most common exposure pathway influencing blood Pb 

9 concentration in structural equation modeling was exterior soil, operating through its 

10 effect on interior dust Pb and hand Pb (Succop et al., 1998). Similar to Lanphear et al.

11 (1998), in the linear regression model, interior dust Pb loading had the strongest 

12 relationships with blood Pb concentration. Individual studies conducted in Rochester, 

13 NY, Cincinnati, OH, and Baltimore, MD report similar relationships between children’s 

14 blood Pb and interior dust concentrations (Lanphear and Roghmann, 1997; U.S. EPA, 

15 1996b; Bornschein et al., 1985).

16 Dixon et al. (2009) reported a multivariate analysis of associations between 

17 environmental Pb concentrations and blood Pb concentrations, based on data collected in 

18 the NHANES (1999-2004). The analyses included 2,155 children, age 12-60 months. The 

19 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 20 of 40% (Table 4-16). The 

21 regression coefficient for the relationship between ln[blood Pb concentration (μg/dL)] 

and ln[floor dust Pb concentration (μg/ft2 22 )] was 0.386 (SE 0.089) for “not smooth and 

23 cleanable” surfaces (e.g., high-pile carpets) and 0.205 (SE 0.032) for “smooth and 

24 cleanable” surfaces (e.g., uncarpeted or low-pile carpets). These coefficients correspond 

25 to a 2.4-fold or 1.6-fold increase in blood Pb concentration, respectively, for a 10-fold 

26 increase in floor dust Pb concentration. 
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Table 4-16 Linear model relating environmental Pb exposure and blood Pb 

concentration in children

Variables Overall p-value Levelsa Estimate (SE) p-Value

Intercept 0.172 - 0.517 (0.373) 0.172

Age (in yr) < 0.001

Age

Age 2

Age 3

Age 4

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 missing

Slope

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

U.S.b

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 missing

Slope

–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

Children: n = 2,155 (age 10-60 months); R2 = 40% b

Includes the 50 states and the District of Columbia

Source: Dixon et al. (2009).

1 Mielke et al. (2007a) analyzed blood Pb and soil Pb concentration data collected as part 

2 of a blood Pb screening program in New Orleans (2000-2005). The data set included 

3 55,551 blood Pb measurements for children 0-6 years of age and 5,467 soil Pb 

4 measurements. Blood Pb and soil Pb concentrations were matched at the level of census 

5 tracts. The association between blood Pb concentration and soil Pb concentration was 

6 evaluated using non-parametric permutation methods. The resulting best-fit model 

(R2 7 =0.528) was:
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PbB = 2. 038 + (0. 172 × PbS0.5)

Equation 4-4

1 where PbB is the median blood Pb concentration and PbS is the median soil Pb 

2 concentration. The resulting curvilinear relationship predicts a twofold increase in blood 

3 Pb concentration for an increase in soil Pb concentration from 100 to 1,000 ppm (Figure 

4 4-24).

Note: 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 (Mielke et al., 2007a).

Figure 4-24 Predicted relationship between soil Pb concentration and blood 

Pb concentration in children based on data collected in the New 

5 In a subsequent re-analysis of the New Orleans (2000-2005) data, individual child blood 

6 Pb observations were matched to census tract soil concentrations (Zahran et al., 2011).

7 This analysis confirmed the association between blood Pb and both soil Pb and age 

8 reported in Mielke et al. (2007a). Regression coefficients for soil Pb (random effects 

generalized least squares regression) ranged from 0.217 to 0.214 (per soil Pb0.5 9 ), which is 

10 equivalent to approximately a 2-fold increase in blood Pb concentration for an increase in 

11 soil Pb concentration from 100 to 1,000 ppm.

12 Several studies have linked elevated blood Pb levels to residential soil exposures for 

13 populations living nearby industrial or mining facilities. Gulson et al. (2009) studied the 

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 blood Pb and isotopic Pb ratios of children younger than 5 years old and adults older than 

2 18 years old living in the vicinity of a mine producing Magellan Pb ore in western 

3 Australia. They observed a median blood Pb level of 6.6 μg/dL for the children, with 

4 isotopic ratios indicating contributions from the mine ranging from 27 to 93%. A weak 

5 but significant linear association between blood Pb level and percent Magellan Pb was 

observed (R2 6 = 0.12, p = 0.018). Among children with blood Pb levels over 9 μg/dL and 

7 among adults, the isotopic ratios revealed Pb exposures from a variety of sources. 

8 Garavan et al. (2008) measured soil Pb and blood Pb levels among children aged 1 month 

9 to 17.7 years old in an Irish town near a coal mine. The blood Pb measurements were 

10 instituted as part of a screening and community education program given that the 

11 presence of Pb had been documented in the environment. Garavan et al. (2008) found that 

12 over 3 years of the screening period, median blood Pb levels reduced by roughly 22% 

13 from 2.7 to 2.1 μg/dL. 

14 An extensive discussion of the relationships between environmental Pb levels and blood 

15 Pb concentrations in children at the Bunker Hill Superfund Site, a former Pb mining and 

16 smelting site, was provided in the 2006 Pb AQCD. In the most recent analysis

17 (TerraGraphics Environmental Engineering, 2004) of the data on environmental Pb levels 

18 and child blood Pb concentrations (1988-2002), blood Pb concentrations (annual GM) 

19 ranged from 2.6 to 9.9 μg/dL. Environmental Pb levels (e.g., dust, soil, paint Pb levels) 

20 data were collected at ~3,000 residences, with interior dust Pb concentrations (annual 

21 GM) ranging from ~400 to 4,200 mg/kg and yard soil Pb concentration (annual GM) 

22 ranging from ~150 to 2,300 mg/kg. Several multivariate regression models relating 

23 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 predicts significant associations 

25 between blood Pb concentration, age, interior dust, yard soil, neighborhood soil 

26 (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 

28 coefficients, the community soil Pb concentration had the largest effect on blood Pb 

29 concentration, followed by neighborhood soil Pb concentration, interior dust Pb 

30 concentration, and yard soil Pb concentration (Table 4-17). The model predicted a 

31 1.8 μg/dL decrease in blood Pb concentration in association with a decrease in 

32 community soil Pb concentration from 2,000 to 1,000 mg/kg. The same decrease in 

33 neighborhood soil Pb concentration, interior dust Pb concentration, or yard soil Pb 

34 concentration was predicted to result in a 0.8, 0.5, or 0.2 μg/dL decrease in blood Pb 

35 concentration, respectively. Note that the soil Pb component of the model was similar to 

36 that derived by Lewin et al. (1999), in which a model of blood Pb as a function of soil Pb 

37 among 0-6 y old children living near one of four industrial sites was given as PbB = 

38 0.2438ln(PbS} + 0.2758.
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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 

2 concentrations in a representative sample of Native American and white children (n = 

3 224, age 1-6 years) who resided in a former Pb mining region in Ottawa County, OK. 

4 The data set included measurements of blood Pb, yard soil Pb, residential interior dust Pb 

5 loading, first-draw water Pb, paint Pb assessment and other behavioral (i.e., hand-to6 mouth activity, hygiene rating) and demographic variables (i.e., hand-to-mouth activity, 

7 hygiene rating, poverty level, caregiver education). A multivariate regression model 

8 accounted for 34% of the observed variability in blood Pb. Yard soil Pb and interior dust 

9 Pb loading accounted for 10% and 3% of the blood Pb variability, respectfully. The 

10 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 11 ] dust Pb loading.

4.6 Biokinetic Models of Lead Exposure-Blood Lead 

Relationships

12 An alternative to regression models are mechanistic models, which attempt to specify all 

13 parameters needed to describe the mechanisms (or processes) of transfer of Pb from the 

14 environment to human tissues. Such mechanistic models are more complex than 

15 regression models; this added complexity introduces challenges in terms of their 

16 mathematical solution and empirical verification. However, by incorporating parameters 

17 that can be expected to vary spatially or temporally, or across individuals or populations, 

18 mechanistic models can be extrapolated to a wide range of exposure scenarios, including 

19 those that may be outside of the domain of paired predictor-outcome data used to develop 

20 the model. Exposure-intake models, a type of mechanistic models, are highly simplified 
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1 mathematical representations of relationships between levels of Pb in environmental 

2 media and human Pb intakes (e.g., μg Pb ingested per day). These models include 

3 parameters representing processes of Pb transfer between environmental media (e.g., air 

4 to surface dust) and to humans, including rates of human contact with the media and 

5 intakes of the media (e.g., g soil ingested per day). Intake-biokinetic models provide the 

6 analogous mathematical representation of relationships between Pb intakes and Pb levels 

7 in body tissues (e.g., blood Pb concentration). Biokinetic models include parameters that 

8 represent processes of Pb transfer (a) from portals of entry into the body and (b) from 

9 blood to tissues and excreta. Linked together, exposure-intake and intake-biokinetics 

10 models (i.e., integrated exposure-intake-biokinetics models) provide an approach for 

11 predicting blood Pb concentrations (or Pb concentrations in other tissues) that 

12 corresponds to a specified exposure (medium, concentration, and duration). Detailed 

13 information on exposure and internal dose can be obtained from controlled experiments, 

14 but almost never from epidemiological observations or from public health monitoring 

15 programs. Exposure intake-biokinetics models can provide these predictions in the 

16 absence of complete information on the exposure history and blood Pb concentrations for 

17 an individual (or population) of interest. Therefore, these models are critical to applying 

18 epidemiologic-based information on blood Pb-response relationships to the quantification 

19 and characterization of human health risk. They are also critical for assessing the 

20 potential impacts of public health programs directed at mitigation of Pb exposure or of 

21 remediation of contaminated sites. 

22 However, they are not without their limitations. Human exposure-biokinetics models 

23 include large numbers of parameters, which are required to describe the many processes 

24 that contribute to Pb intake, absorption, distribution, and elimination. The large number 

25 of parameters complicates the assessment of confidence in parameter values, many of 

26 which cannot be directly measured. Statistical procedures can be used to evaluate the 

27 degree to which model outputs conform to “real-world” observations and values of 

28 influential parameters can be statistically estimated to achieve good agreement with 

29 observations. Still, large uncertainty can be expected to remain about many, or even 

30 most, parameters in complex exposure-biokinetic models. Such uncertainties need to be 

31 identified and their impacts on model predictions quantified (i.e., sensitivity analysis or 

32 probabilistic methods).

33 Modeling of human Pb exposures and biokinetics has advanced considerably during the 

34 past several decades, although there have been relatively few developments since the 

35 2006 Pb AQCD was published. Still in use is the Integrated Exposure Uptake Biokinetic 

36 (IEUBK) Model for Lead in Children (U.S. EPA, 1994) and models that simulate Pb 

37 biokinetics in humans from birth through adulthood (O'Flaherty, 1995; Leggett, 1993; 
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1 O'Flaherty, 1993). The EPA AALM is still in development. A complete and extensive 

2 discussion of these models can be found in the 2006 Pb AQCD (U.S. EPA, 2006b).

4.7 Summary and Conclusions

4.7.1 Exposure

3 Exposure data considered in this assessment build upon the conclusions of the 2006 Pb 

4 AQCD (2006b), which found air Pb concentrations in the U.S. and associated biomarkers 

5 of exposure to Pb have decreased substantially following the ban on Pb in gasoline, 

6 house-hold paints, and solder. Pb exposure is difficult to assess because Pb has multiple 

7 sources in the environment and passes through various media. The atmosphere is the 

8 main environmental transport pathway for Pb, and, on a global scale, atmospheric Pb is 

9 primarily associated with fine particulate matter, which can deposit to soil and water. In 

10 addition to primary emission of particle-bearing or gaseous Pb to the atmosphere, Pb can 

11 be suspended to the air from soil or dust. Air-related pathways of Pb exposure are the 

12 focus of this assessment. In addition to inhalation of Pb from ambient air, air-related Pb 

13 exposure pathways include inhalation and ingestion of Pb from indoor dust and/or 

14 outdoor soil that originated from recent or historic ambient air (e.g., air Pb that has 

15 penetrated into the residence either via the air or tracking of soil). Non-air-related Pb 

16 exposures may include occupational exposures, hand-to-mouth contact with Pb17 containing consumer goods, hand-to-mouth contact with dust or chips of peeling Pb18 containing paint, or ingestion of Pb in drinking water conveyed through Pb pipes. Pb can 

19 cycle through multiple media prior to human exposure. Given the multitude of possible 

20 air-related exposure scenarios and the related difficulty of constructing Pb exposure 

21 histories, most studies of Pb exposure through air, water, and soil can be informative to 

22 this review. Other exposures, such as occupational exposures, contact with consumer 

23 goods in which Pb has been used, or ingestion of Pb in drinking water conveyed through 

24 Pb pipes may also contribute to Pb body burden.

25 A number of monitoring and modeling techniques have been employed for ambient Pb 

26 exposure assessment. Environmental Pb concentration data can be collected from 

27 ambient air Pb monitors, soil Pb samples, dust Pb samples, and dietary Pb samples to 

28 estimate human exposure. Exposure estimation error depends in part on the collection 

29 efficiency of these methods; collection efficiency for ambient air Pb FRM samplers is 

30 described in Section 3.4. Additionally, high spatial variability of the Pb concentrations in 

31 various media also can contribute to exposure error, as described in the 2009 PM ISA

32 (U.S. EPA, 2009). Models, such as the Integrated Exposure Uptake Biokinetic (IEUBK) 
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1 model, simulate human exposure to Pb from multiple sources and through various routes 

2 including inhalation, ingestion, and dermal exposure. IEUBK model inputs include soil 

3 Pb concentration, air Pb concentration, dietary Pb intake including drinking water, Pb 

4 dust ingestion, human activity, and biokinetic factors. Measurements and/or assumptions 

5 can be utilized when formulating the model inputs; errors in measurements and 

6 assumptions thus have the potential to propagate through the exposure models.

7 Section 4.1 presents data illustrating potential exposure pathways. Soil can act as a 

8 reservoir for deposited Pb emissions, and exposure to soil contaminated with deposited 

9 Pb can occur through resuspended PM as well as shoe tracking and hand-to-mouth 

10 contact, which is the main pathway of childhood exposure to Pb. In general, soil Pb 

11 concentrations tended to be higher within inner-city communities compared with 

12 neighborhoods surrounding the city. Recent data by Yamamoto et al. (2006) have shown 

13 that the size distribution of particles collected on children’s hands have a mode around 

14 40 μm with the upper tail of the distribution extending to 200-300 μm. Infiltration of Pb 

15 dust into indoor environments has been demonstrated, and Pb dust has been shown to 

16 persist in indoor environments even after repeated cleanings. Measurements of particle17 bound Pb exposures reported in this assessment have shown that personal exposure 

18 measurements for Pb concentration are typically higher than indoor or outdoor ambient 

19 Pb concentrations. These findings may be related to local resuspension with body 

20 movement.

4.7.2 Toxicokinetics

21 The majority of Pb in the body is found in bone (roughly 90% in adults, 70% in children); 

22 only about 1% of Pb is found in the blood. Pb in blood is primarily (~99%) bound to red 

23 blood cells (RBCs). It has been suggested that the small fraction of Pb in plasma (<1%) 

24 may be the more biologically labile and toxicologically active fraction of the circulating 

25 Pb. The relationship between Pb in blood and plasma is pseudo-linear at relatively low 

26 daily Pb intakes (i.e., <10 μg/day/kg) and at blood Pb concentrations <25 μg/dL, and 

27 becomes curvilinear at higher blood Pb concentrations due to saturable binding to RBC 

28 proteins. As blood Pb level increases and the higher affinity binding sites for Pb in RBCs 

29 become saturated, a larger fraction of the blood Pb is available in plasma to distribute to 

30 brain and other Pb-responsive tissues.

31 The burden of Pb in the body may be viewed as divided between a dominant slow 

32 (i.e., uptake and elimination) compartment (bone) and smaller fast compartment(s) (soft 

33 tissues). Pb uptake and elimination in soft tissues is much faster than in bone. Pb 

34 accumulates in bone regions undergoing the most active calcification at the time of 
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1 exposure. During infancy and childhood, bone calcification is most active in trabecular 

2 bone (e.g., patella); whereas, in adulthood, calcification occurs at sites of remodeling in 

3 cortical (e.g., tibia) and trabecular bone (Aufderheide and Wittmers, 1992). A high bone 

4 formation rate in early childhood results in the rapid uptake of circulating Pb into 

5 mineralizing bone; however, in early childhood bone Pb is also recycled to other tissue 

6 compartments or excreted in accordance with a high bone resorption rate (O'Flaherty, 

7 1995). Thus, much of the Pb acquired early in life is not permanently fixed in the bone.

8 The exchange of Pb from plasma to the bone surface is a relatively rapid process. Pb in 

9 bone becomes distributed in trabecular and the more dense cortical bone. The proportion 

10 of cortical to trabecular bone in the human body varies by age, but on average is about 

11 80% cortical to 20% trabecular. Of the bone types, trabecular bone is more reflective of 

12 recent exposures than is cortical bone due to the slow turnover rate and lower blood 

13 perfusion of cortical bone. Some Pb diffuses to deeper bone regions where it is relatively 

14 inert, particularly in adults. These bone compartments are much more labile in infants 

15 and children than in adults as reflected by half-times for movement of Pb from bone into 

16 to the plasma (e.g., cortical half-time = 0.23 years at birth, 3.7 years at 15 years of age, 

17 and 23 years in adults; trabecular half-time = 0.23 years at birth, 2.0 years at 15 years of 

18 age, and 3.8 years in adults) (Leggett, 1993).

19 Evidence for maternal-to-fetal transfer of Pb in humans is derived from cord blood to 

20 maternal blood Pb ratios. Group mean ratios range from about 0.7 to 1.0 at the time of 

21 delivery for mean maternal blood Pb levels ranging from 1.7 to 8.6 μg/dL. Transplacental 

22 transfer of Pb may be facilitated by an increase in the plasma/blood Pb concentration 

23 ratio during pregnancy. Maternal-to-fetal transfer of Pb appears to be related partly to the 

24 mobilization of Pb from the maternal skeleton.

25 The dominant elimination phase of Pb kinetics in the blood, exhibited shortly after a 

26 change in exposure occurs, has a half-life of ~20-30 days. An abrupt change in Pb uptake 

27 gives rise to a relatively rapid change in blood Pb, to a new quasi-steady state, achieved 

28 in ~75-100 days (i.e., 3-4 times the blood elimination half-life). A slower phase of Pb 

29 clearance from the blood may become evident with longer observation periods following 

30 a decrease in exposure due to the gradual redistribution of Pb among bone and other 

31 compartments. 

4.7.3 Lead Biomarkers

32 Overall, trends in blood Pb levels have been decreasing among U.S. children and adults 

33 over the past 20 years (Section 4.4). The median blood Pb level for the entire U.S. 

34 population is 1.2 μg/dL and the 95th percentile blood Pb level was 3.7 μg/dL, based on 
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1 the 2007-2008 NHANES data (NCHS, 2010). Among children aged 1-5 years, the 

2 median and 95th percentiles were slightly higher at 1.4 μg/dL and 4.1 μg/dL, 

3 respectively. 

4 Blood Pb is dependent on both the recent exposure history of the individual, as well as 

5 the long-term exposure history that determines body burden and Pb in bone. The 

6 contribution of bone Pb to blood Pb changes depending on the duration and intensity of 

7 the exposure, age, and various other physiological stressors that may affect bone 

8 remodeling (e.g., nutritional status, pregnancy, menopause, extended bed rest, 

9 hyperparathyroidism) beyond that which normally and continuously occurs. In children, 

10 largely due to faster exchange of Pb to and from bone, blood Pb is both an index of recent 

11 exposure and potentially an index of body burden. In adults and children, where exposure 

12 to Pb has effectively ceased or greatly decreased, a slow decline in blood Pb 

13 concentrations over the period of years is most likely due to the gradual release of Pb 

14 from bone. Bone Pb is an index of cumulative exposure and body burden. Even bone 

15 compartments should be recognized as reflective of differing exposure periods with Pb in 

16 trabecular bone exchanging more rapidly than Pb in cortical bone with the blood. This 

17 difference in the compartments makes Pb in cortical bone a better marker of cumulative 

18 exposure and Pb in trabecular bone more likely to be correlated with blood Pb, even in 

19 adults. 

20 Sampling frequency is an important consideration when evaluating blood Pb and bone Pb 

21 levels in epidemiologic studies, particularly when the exposure is not well characterized. 

22 It is difficult to determine what blood Pb is reflecting in cross-sectional studies that 

23 sample blood Pb once, whether recent exposure or movement of Pb from bone into blood 

24 from historical exposures. In contrast, cross-sectional studies of bone Pb and longitudinal 

25 samples of blood Pb concentrations over time provide more of an index of cumulative 

26 exposure and are more reflective of average Pb body burdens over time. The degree to 

27 which repeated sampling will reflect the actual long-term time-weighted average blood 

28 Pb concentration depends on the sampling frequency in relation to variability in 

29 exposure. High variability in Pb exposures can produce episodic (or periodic) oscillations 

30 in blood Pb concentration that may not be captured with low sampling frequencies.

31 Furthermore, similar blood Pb concentrations in two individuals (or populations), 

32 regardless of their age, do not necessarily translate to similar body burdens or similar 

33 exposure histories.

34 The concentration of Pb in urine follows blood Pb concentration, in that it mainly reflects 

35 the exposure history of the previous few months and therefore, is likely a relatively poor 

36 index of Pb body burden. There is added complexity with Pb in urine because 

37 concentration is also dependent upon urine flow rate, which requires timed urine samples 
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1 that is often not feasible in epidemiologic studies. Other biomarkers have been utilized to 

2 a lesser extent (e.g., Pb in teeth).

4.7.4 Air Lead-Blood Lead Relationships

3 The 1986 Pb AQCD described epidemiological studies of relationships between air Pb 

4 and blood Pb. Much of the pertinent earlier literature described in the 1986 Pb AQCD 

5 was drawn from a meta-analysis by Brunekreef (1984). Based on the studies available at 

6 that time that considered multiple air-related Pb exposure pathways in the aggregate, the 

7 1986 Pb AQCD concluded that “the blood Pb versus air Pb slope β is much smaller at 

8 high blood and air levels.” This is to say that the slope β was much smaller for 

9 occupational exposures where high blood Pb levels (>40 μg/dL) and high air Pb levels 

(much greater than 10 μg/m3 10 ) prevailed relative to lower environmental exposures which 

showed lower blood Pb and air Pb concentrations (<30 μg/dL and <3 μg/m3 11 ). For those 

12 environmental exposures, it was concluded that the relationship between blood Pb and air 

13 Pb “...for direct inhalation appears to be approximately linear in the range of normal 

ambient exposures (0.1-2.0 μg/m3 14 )” (pp 1-98 of the 1986 Pb AQCD). In addition to the 

15 meta-analysis of Brunekreef (1984), more recent studies have provided data from which 

16 estimates of the blood Pb-air Pb slope can be derived for children (Table 2-7). The range 

of estimates from these studies is 2-9 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 ). Most studies have 

19 described the blood Pb-air Pb relationship as either log-log (Schnaas et al., 2004; Hayes 

20 et al., 1994; Brunekreef, 1984), which predicts an increase in the blood Pb-air Pb slope 

21 with decreasing air Pb concentration or linear (Hilts, 2003; Tripathi et al., 2001; Schwartz 

22 and Pitcher, 1989), which predicts a constant blood Pb-air Pb slope across all air Pb 

23 concentrations. These differences may simply reflect model selection by the 

24 investigators; alternative models are not reported in these studies. The blood Pb-air Pb 

25 slope may also be affected in some studies by the inclusion of parameters (e.g., soil Pb) 

26 that may account for some of the variance in blood Pb attributable to air Pb. Other factors 

27 that likely contribute to the derived blood Pb-air Pb slope include differences in the 

28 populations examined and Pb sources, which varied among individual studies.

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