Data from the Total Exposure Assessment Methodology studies, conducted from 1980 to 1987 in New Jersey (NJ) and California (CA), and the 1990 California Indoor Exposure study were analyzed using positive matrix factorization, a receptor-oriented source apportionment model. Personal exposure and outdoor concentrations of 14 and 17 toxic volatile organic compounds (VOCs) were studied from the NJ and CA data, respectively. Analyzing both the personal exposure and outdoor concentrations made it possible to compare toxic VOCs in outdoor air and exposure resulting from personal activities. Regression analyses of the measured concentrations versus the factor scores were performed to determine the relative contribution of each factor to total exposure concentrations. Activity patterns of the NJ and CA participants were examined to determine whether reported exposures to specific sources correspond to higher estimated contributions from the factor identified with that source. For a subset of VOCs, a preliminary analysis to determine irritancy-based contributions of factors to exposures was carried out. Major source types of toxic VOCs in both NJ and CA appear to be aromatic sources resembling automobile exhaust, gasoline vapor, or environmental tobacco smoke for personal exposures, and automobile exhaust or gasoline vapors for outdoor concentrations.
Hazardous air pollutants (HAPs) are generally defined as pollutants that are known or suspected to cause cancer or other serious health effects, or to cause harm to the environment (OAQPS, 1998). The 1990 Clean Air Act Amendments (CAAA), Section 112, seek to reduce human exposure to HAPs by defining a statutory list of these compounds. More than 80% of the compounds on the federal HAPs list are toxic volatile organic compounds (VOCs). The main pathways for exposure to toxic VOCs include breathing contaminated air, ingesting contaminated water or soil, and dermal contact. This study focuses on identifying sources of inhalation exposure to toxic VOCs, most of which are on the current HAPs list.
The majority of toxic VOCs in ambient air originate from sources that emit to the outdoors, such as drycleaners, power plants, and vehicle emissions (UATW, 2000). It is unclear, however, whether these sources are the predominant contributors to human exposure. Although not major sources of emissions, personal activities and indoor sources may be the dominant sources of exposure for many compounds (Wallace, 1991). Some examples of sources of personal exposure to toxic VOCs include household cleaners, vehicle exhaust, gasoline vapors, drycleaned clothes, and environmental tobacco smoke (ETS).
With concentration data, receptor modeling may be used to identify source types that contribute significantly to measured concentrations. Although receptor modeling has been widely used for estimating source contributions for particulate matter (PM) air pollution, relatively few receptor modeling studies have been conducted for VOCs. The studies that have been done have focused primarily on estimating source contributions to VOCs in outdoor air rather than to personal exposure. Early examples of receptor modeling applied to VOCs include apportionment of ambient VOCs (Lin and Milford, 1994; Mukund et al., 1996), and the use of acetylene (Lonneman et al., 1974; Whitby and Altwicker, 1978) and aromatic organic compounds (Singh et al., 1985; Edgerton et al., 1989) as indicators of motor vehicle emissions.
The primary goals of this study were to determine sources contributing to personal exposure and outdoor concentrations, and the relative contribution of each source to total concentrations. Personal exposure and outdoor concentration data for residents in Elizabeth and Bayonne, NJ, and Los Angeles (L.A.), Pittsburg, and Antioch, CA from the United States Environmental Protection Agency's (USEPA) Total Exposure Assessment Methodology (TEAM) and California Air Resources Board (CARB) California Indoor Exposure studies were analyzed using positive matrix factorization (PMF; Paatero and Tapper, 1994).
The NJ TEAM data used were accessed from the Total Human Exposure Risk database and Advanced Simulation Environment (THERdbASE), version 1.2 (Pandian et al., 1989; NERL, 2000). This database includes data from the main TEAM studies, which measured 24-h exposures of 600 people in NJ, CA, North Dakota (ND) and North Carolina (NC) to various toxic chemicals in air and drinking water. Monitored compounds were selected based on their toxicity, carcinogenicity, and amenability to collection on Tenax sorbent (Wallace, 1987; Wallace et al., 1984). The CA TEAM and CARB data were accessed from the Californian Exposures Database (CED) (Clayton and Perritt, 1993). CARB studies were carried out using the same general procedures as the TEAM studies. A key feature of these studies was the use of a probability-based sampling design to represent the exposures of large populations in the various cities.
Exposures were measured using personal monitors, which collected two 12-h samples, representing overnight and daytime exposures. Concurrent outdoor samples were collected in two 12-h samples from the backyards of a subset of the study homes. Air samples were analyzed by gas chromatography/mass spectrometry. Both studies used time-weighted average exposure concentrations as the measure of exposure. Two questionnaires were administered to participants. At the beginning of each study, household questionnaires were administered regarding age, gender, occupation, smoking status, and customary activities of the participants. Immediately following each study, participants filled out a 24-h activity diary to establish an activity pattern for each subject. These questionnaires provided information needed to identify the likely sources and human activities contributing to exposures.
PMF, a multivariate technique, was used for source apportionment (Paatero and Tapper, 1994). Previous PMF applications include identifying sources of bulk wet deposition concentrations of strong acids in Finland (Antilla et al., 1995), and more recently, sources of PM from the USEPA Particle TEAM study (Yakovleva et al., 1999) and PM in Phoenix urban aerosol (Ramadan et al., 2000). PMF incorporates error estimates of the data to solve matrix factorization of a linear model as a constrained, weighted least-squares problem. These error estimates account for sampling errors, detection limits, missing data observations, and outliers. The input data must be finite, positive numbers. One portion of the model solution is a matrix of factors. These factors, which are roughly interpreted as source profiles, represent the relative amounts of each compound in each source. Because source profiles are spatially and temporally variable, the factor profiles presented are a qualitative guide to the types of sources to which people could be exposed and have been qualitatively related to the best available profiles in the literature. Each factor is constrained to be nonnegative. This requirement decreases the rotational freedom used to produce meaningful factors, and oftentimes, the result is fully unique with no rotational freedom (Paatero, 1998).
PMF was applied to 24-h personal and outdoor concentration data obtained by averaging overnight and daytime samples, as well as to separate daytime and overnight personal concentration data, to investigate potential sources of exposure to toxic VOCs. The compounds included in this study are listed in Table 1. Due to insufficient data collected for participants in ND, NC, and Woodland, CA, data from these locations were not used in this study.
Two-way PMF seeks to solve the matrix equation
where Xij is an element of an n by m matrix of observed data with n chemical species and m study participants. Gih is an element of an n by p unknown factor matrix, which is interpreted as a matrix of source profiles. Fhj is an element of a p by m matrix of unknown factor scores, whose relative magnitudes indicate how much the hth factor contributes to the exposure of the jth participant. Eij is the residual error. The number of factors, p, is chosen by the user. Given p, PMF solves this matrix problem by minimizing the sum of squares, Q:
where Sij is the standard deviation representing the uncertainty in the observation Xij. In this study, the Xij are assumed to be lognormally distributed and the values of Sij are iteratively refined so that the solution to Equation 1 approximates a maximum likelihood solution. Further details in the error model specification are given below and by Paatero (1997).
In this study, two-way PMF was applied separately to the personal exposure and outdoor concentration data from NJ and CA.
Three-way PMF seeks to solve a similar matrix equation for an extended set of observations that includes multiple sample modes, such as corresponding personal and outdoor samples. The three-way factor model can be written as
where i indexes the chemical species, j indexes the study participant, k indexes the sample mode, and h indexes the factors. Ajh is an element of the m by p matrix of unknown factor scores, Bih is an element of the n by p unknown factor matrix, and Ckh is an element of the q by p matrix that weights the factor contributions between sample modes. Q is calculated as shown in Equation 2.
Three-way PMF was applied here to the subset of TEAM and CARB study participants for which both personal and outdoor concentrations were measured, as well as to daytime and overnight concentrations for personal exposures.
The Q value (Equation 2), used to estimate the best number of factors, is an indicator of the goodness of fit and a measure of how close the modeled values are to the observed values. Because more of the variation in the data can be explained with more factors, the Q value will generally continue to decrease as the number of factors increases. The additional factors, however, may not be physically meaningful and may just be explaining noise in the data. Thus, other diagnostic measures, as well as judgment of the user, are required to select an appropriate number of factors. Additional diagnostics used include residuals for individual species and for total VOC concentrations.
In the presence of outliers in the data, it can be difficult to determine if the reported Q value is too high (Paatero, 1998). Therefore, the distribution of the scaled residuals for individual compounds may be used as a diagnostic measure as well. Ideally, the residuals should be centered around zero, forming a sharp peak at zero.
A linear regression analysis of the factor scores versus the sum of the measured VOC concentrations was used to help ascertain the optimal number of factors. In addition, the regression was used to determine the relative contribution of each factor to exposure. Multiplying the factor scores (unitless) by the linear regression coefficients (concentration units) gives a total modeled VOC concentration from each factor. Summing these modeled concentrations over all factors gives a total modeled concentration that allows the user to compare the model-predicted total VOC concentration to the measured total VOC concentration for each observation. In addition, the percent contribution of each factor to total VOCs can be determined by dividing the modeled concentration for each factor by the total modeled concentration. The number of factors was selected, in part, to avoid negative regression coefficients. R-squared values and the ratios of modeled to measured concentrations were examined to determine how well the regression model fit the measured data for total VOCs.
Treatment of Missing Data and Uncertainties
Compounds were included in the PMF analysis only if more than 60% of all study participants had values for that compound. Values reported as below the detection limit were included in the analysis to create the largest concentration matrix possible, but were assigned large errors and thus given low weight. Similarly, participants were included only if they had data for at least 60% of the compounds. In this case, missing values for a given compound were filled in using the median value for that compound across all participants. Tables Table 2 and Table 3 show that relatively few data points were missing after compounds and participants not meeting the specified criteria were excluded.
One limitation of PMF is its inability to extract factors that fit widely varying exposure concentrations, which is apparent when comparing modeled concentrations from the regression analysis to measured concentrations. When the measured concentrations analyzed include both very high and very low values, the modeled concentrations are generally much higher or lower than the true values. When the measured concentrations are restricted to a smaller range, the modeled concentrations reflect measured concentrations much more closely. Because of this limitation, only participants with total exposure concentrations of less than 2000 μg/m3 were included in the NJ personal exposure data analysis, excluding 4% of the participants. This cutoff concentration was chosen because it was an obvious divider between the high and low concentrations. No outdoor concentrations were greater than 2000 μg/m3. In the CA studies, only one participant had a total exposure concentration greater than 2000 μg/m3. Therefore, all participants were included in the analysis for CA.
All data were treated as coming from lognormal distributions, as indicated by examination of histogram plots, means, and medians for the concentration data. The error model used by PMF for lognormally distributed data is:
where Yij is the fitted concentration value, represents normally distributed measurement error and Vij is associated with inherent randomness in the data. For this study, Tij was determined for each observation using the following formula:
where DLi is the detection limit and Mi is the median value for the compound of interest across all observations, as given by Pandian et al. (1989) and Clayton and Perritt (1993). A constant value of 0.1 was used for Vij. For the three-way analysis, Sij is replaced by Sijk, Tij by Tijk, etc. It was found that changing the error estimates had a negligible effect on the factors obtained by the model.
Geometric means (GM) and geometric standard deviations (GSD) of the personal exposure and outdoor concentration data used in the PMF modeling are presented in Table 2 for the NJ data and Table 3 for the CA data. NJ and CA average personal exposure concentrations for all compounds are higher than outdoor concentrations.
Standard deviations of the factors are plotted as error bars on all source apportionment results. These values are based on a global least squares fit, in which all three matrices, A, B and C, are determined simultaneously (Paatero, 1998). In general, the standard deviation matrices represent both individual random uncertainty in the factor elements and uncertainty due to factor rotation. If there is negligible rotational freedom, however, the error estimates in A, B, and C reflect only the random uncertainty in the factor elements, and the values in the standard deviation matrices will be small (Paatero, 1998). Although error bars are plotted for each compound in the factor profiles, some errors are small enough such that they are not visible, suggesting negligible rotational freedom of the factors.
Factorization was performed using from four to nine factors for both the two-way and three-way PMF data sets. Six factors and five factors were chosen as the optimal number of factors for the personal and outdoor two-way PMF data sets, respectively. Eight factors were chosen for the three-way PMF model. Adding more factors slightly improved the Q values and R-squared values, but investigation of the resulting factor profiles revealed that the additional factors were not easily interpreted and were likely just explaining noise in the data. Factor profiles and modal associations for the three-way model are shown in Figure 1. The right-hand side of each bar graph shows the relative contribution of each factor to personal exposure (PER) or outdoor (OUT) concentrations, and also to personal daytime (PER DT) versus personal overnight (PER ON) exposures. In all factor profiles, the height of the bars can only be compared within a given factor, and not across all factors, because the sum of chemical contributions in a factor is normalized to one.
The factors obtained from the three-way PMF analysis generally matched the factors obtained from the two-way PMF analysis. Both analyses showed similar contributions of factors to total concentrations. This occurs despite the fact that the three-way model was applied to only a subset of the participants included in the two-way models. Factor 6p, however, did not have an equivalent three-way factor, nor was there an equivalent factor to the three-way Factor 3 in the two-way outdoor model. This is likely due to the use of different data sets in the two-way and three-way models.
In addition to 24-h personal and outdoor data, 12-h daytime and overnight personal observations were modeled using three-way PMF. Daytime sources were generally the same as overnight sources, though the sources contributed differently to exposure. Modal associations for daytime and overnight personal factors are shown in Figure 1.
Figure 2 shows examples of the residuals for benzene and TCE for the personal and outdoor modes of the three-way PMF model. If the majority of the scaled residuals (Eij/Sij) are between ±2 standard deviations (σ), the fit of the model is acceptable (Paatero, 1998). The percentages of residuals within ±2 σ are shown in Table 4.
To estimate the contributions of factors to total personal exposure and outdoor concentrations (summed over all VOCs), the observations of these values were regressed against the factor scores produced by the PMF model. Figure 3 shows an example of the factor scores for each NJ participant for Factor 1 in the three-way model. The regression analysis results, showing the average source contributions to personal exposure and outdoor concentrations for both the two-way and three-way NJ PMF data sets, are presented in Table 5. For example, Factor 3 contributed, on average, 22% to participants' total 24-h personal VOC exposure, whereas Factor 4 contributed approximately half as much (11%) to total exposure. Averaged across all of the study participants, Factors 1 and 2 contribute most to total personal exposures, whereas Factors 5 and 6 contribute most to total outdoor concentrations of the 14 VOCs studied.
The measured concentrations were well predicted by the regression. The personal modeled concentrations were, on average, slightly overpredicted at 110±40% and 105±12% of measured concentrations for the three-way and two-way models, respectively. The predicted outdoor concentrations were approximately 100±32% and 97±22% of measured concentrations for the three-way and two-way models, respectively.
Factorization was attempted for the CA data using from 4 to 11 factors for both the two-way and three-way PMF data sets. Eight factors and three factors were chosen as the optimal number of factors for the personal and outdoor two-way PMF data sets, respectively. Nine factors were chosen for the three-way PMF model. Factor profiles and modal associations for the three-way model are shown in Figure 4. As with the NJ data, the factors obtained from three-way PMF generally matched the factors obtained from the two-way PMF analysis. Factor 1 from the three-way model, however, did not appear in the personal two-way model. The percentages of residuals within ±2 σ are shown in Table 4.
Daytime and overnight personal observations were modeled using three-way PMF for the CA data as well. Modal associations for the daytime and overnight personal factors are shown in Figure 4. Factor 6 from the three-way model did not have an equivalent daytime/overnight factor.
The regression analysis results for the two-way and three-way CA PMF models are presented in Table 6. Averaged across all of the study participants, Factors 1, 5, and 9 contribute most to personal exposure and Factor 7 contributes most to outdoor concentrations of the 17 VOCs studied.
The measured concentrations were well predicted by the regression. The personal modeled concentrations were, on average, slightly overpredicted at 120±41% and 100±11% of measured concentrations for the three-way and two-way models, respectively. The predicted outdoor concentrations were approximately 104±18% and 102±13% of measured concentrations for the three-way and two-way models, respectively.
Interpretation of Factor Profiles
Factors identified in three-way PMF were interpreted as 10 different source types contributing to NJ and CA VOC concentrations that were reflected in either one or both of the data sets. The interpretation was made on the basis of qualitative comparison to source or exposure profiles reported in the literature, as detailed below. Note that precise matches were not expected due to variability in the reported profiles and the fact that VOCs undergo chemical degradation from the time they are emitted. The sources are: contaminated water (TCE); solvents (TCA); gasoline vapors, automobile exhaust, or ETS (aromatics, including benzene, ethylbenzene, and xylenes); wastewater treatment plant emissions (chloroform, TCA, TCE, benzene, xylenes); drycleaning chemicals (PERC); deodorizers or mothballs (p-DCB); consumer products (PERC, styrene, benzene, n-octane); building materials (alkanes and xylenes); background ambient concentrations (TCA, benzene, ethylbenzene, PERC, xylenes); and cleaners (α-pinene). These sources were, in general, confirmed by two-way PMF results. A summary of the factorization results for the NJ and CA PMF modeling is presented in Table 7.
The contaminated water factor, comprised of mainly TCE, may be a reflection of exposure to contaminated drinking and showering water or solvents (McKone, 1987; Wallace et al., 1989). The main exposure route to TCE appears to be from personal activities, as evidenced by Factors 1 and 6 from the NJ and CA three-way models, and the corresponding Factors 1p and 6p from the NJ and CA personal two-way models.
The TCA source likely reflects exposure to solvents, household cleaners, and household pesticides (Wallace et al., 1989). The main exposure route to TCA, as evidenced by NJ Factors 3 and 3p, and CA Factors 5 and 5p, is due to personal activities as well.
The aromatics factor, resembling ETS (Daisey et al., 1994), automobile exhaust or gasoline vapors (Rappaport et al., 1987; Scheff et al., 1991; MacIntosh et al., 1995; Wallace, 1996), is present as both a personal exposure and outdoor source in NJ and CA. Exposure to ETS could result from smoking or being proximal to a smoker. Personal exposure to gasoline or exhaust could result from activities such as sitting in traffic, pumping gasoline, or from being exposed to outdoor air with automobile exhaust or gasoline vapors that has penetrated indoors. NJ Factors 2 and 2p, and 6 and 4o consist mainly of ethylbenzene and xylenes. Benzene, a major component of gasoline, automobile exhaust and ETS, is not included in these factors. It is, however, the dominant component in other NJ personal and outdoor factors (Factors 8 and 6p, and Factors 5 and 2o, respectively). This suggests that benzene is present in widely varied concentrations in several sources. PMF, to account for the variability in the data, might make benzene a separate source. As shown with CA Factors 9 and 8p, and Factors 7 and 3o, an aromatics factor consisting of benzene, ethylbenzene, and xylenes was a personal and outdoor source in the CA data as well.
The outdoor source comprised of chloroform, TCA, xylenes, and TCE resembles wastewater treatment plant emissions (Scheff et al., 1989). This source is evident with NJ Factor 5, which appears to be a combination of Factors 2o and 3o in the two-way model, and CA Factor 1, which affects both personal exposure and outdoor concentrations. This factor did not appear in the CA two-way personal model, but was present in the two-way outdoor results, as shown by Factor 1o.
The PERC source could result from exposure to drycleaned clothes, or being inside or living near a drycleaning shop (Wallace, 1989; Wallace et al., 1989). This source was only extracted from the NJ data, although PERC was measured in CA too. As shown in NJ Factors 4, 4p and 1o, the PERC source affects both personal exposure and outdoor concentrations.
The p-DCB source is most likely a reflection of exposure to mothballs and room deodorizers (Wallace, 1989). Because this compound was not included in the CA modeling due to a limited number of measurements, the p-DCB source appeared only in the NJ data. NJ Factors 7 and 5p show that p-DCB primarily affects personal exposures. A p-DCB source was present in the NJ two-way PMF outdoor results, although it was not significant to outdoor concentrations.
The source resembling consumer products (Samfield, 1992; Sheldon et al., 1992), such as adhesives and aerosols, is comprised of PERC, styrene, benzene, and n-octane. This source, appearing only in the CA results, was reflected in CA Factor 2, a personal exposure factor that appears to be a combination of Factors 1p, 2p and 3p from the two-way model.
The building material and carpeting source (Mølhave, 1982; Wallace, 1987), composed of higher alkanes (n-decane, n-undecane, n-dodecane) and xylenes, was extracted from only the CA data as well, as shown in CA Factors 3 and 4p. These alkanes were not measured in the NJ study.
As shown in Factors 4 and 2o, the outdoor source comprised of TCA, benzene, ethylbenzene, PERC and xylenes is present only in the CA results. This factor resembles background outdoor concentrations of these compounds (Rosenbaum et al., 1999).
The last source, extracted only from the CA data, is comprised of α-pinene, which is most likely a reflection of exposure to deodorizers and cleaners (Apte and Daisey, 1999). Factors 8 and 7p show that α-pinene primarily affects personal exposures. This compound was not measured in the NJ study.
Activity pattern data from questionnaires administered in the CARB and TEAM studies were analyzed along with the personal factor scores to determine if factors with high scores were reflected in the reported activities. For example, we expected that participants who reported having smoked or pumped gasoline should have higher scores for factors dominated by aromatics than those who did not report being exposed to smoke or gasoline. Although many activities were analyzed, only a subset is discussed here.
Participants in NJ who reported visiting the drycleaners (4 out of 321 participants) had factor scores for the personal PERC factor that were approximately twice as high as those who did not report visiting the drycleaners. It should be noted, however, that the sample size for NJ is likely not large enough to produce conclusive evidence as to exposure from being at the drycleaners. In addition, significant exposure can arise from wearing or being around drycleaned clothes. Therefore, the highest exposures may have occurred outside of the drycleaners (Wallace, 1989).
Three hundred five out of 321 NJ participants and 282 out of 311 CA participants reported drinking tap water, as opposed to bottled water. These participants had approximately four times higher factor scores for the TCE factor (Factor 1p in NJ and Factor 6p in CA), which is thought to reflect exposure to contaminated water. CA participants who reported having showered or bathed (204 out of 234 participants) had two times higher factor scores for Factor 6p. This activity was not reported for the NJ participants.
CA participants who reported having used paint or solvents (16 out of 94 participants) had approximately five times higher factor scores for the 1,1,1-TCA factor (Factor 5p) and seven times higher factor scores for the alkanes factor (Factor 4p). This activity was not reported for the NJ participants.
The NJ participants who reported having smoked or lived with smokers in NJ (224 out of 321 participants) had approximately 85% higher factor scores for the aromatics factor (Factor 2p) and 40% higher factor scores for the benzene factor (Factor 6p) than people who neither smoked nor lived with smokers. Those who reported having pumped gasoline during the study (22 out of 321 participants) showed 80% higher factor scores than those who did not pump gasoline for Factor 2p and 20% higher factor scores for Factor 6p. The fact that higher factor scores for the aromatics factors are linked to both ETS and pumping gasoline may be evidence of PMF having difficulty separating sources with similar profiles.
The CA participants who reported having smoked or lived with smokers (145 out of 305 participants) did not have elevated factor scores for the aromatics factor compared to those who did not smoke or live with smokers. This suggests that either ETS was not strongly reflected as a source of exposure in the PMF factors or that most participants received similar levels of exposure to ETS, whether they reported exposure to ETS or not. A study published in 1996 by Pirkle et al. (1996) indicates widespread exposure to ETS in the U.S. Of American nonsmokers tested, 88% showed evidence of exposure to ETS over the previous 1 to 2 days.
Forty-eight out of 305 CA participants reported having pumped gasoline and 50 out of 93 reported having an attached garage, but neither group showed an increase in the factor scores, on average, for the aromatics factor (Factor 8p). Exposure to gasoline or automobile exhaust, therefore, may not have been a strong personal source of exposure for CA participants or, as with ETS, it may have been a common source of exposure that most participants received.
Tables Table 5 and Table 6 show mass-based contributions of various sources to exposures. Because this research focuses on exposure to toxic VOCs, a preliminary analysis using irritancy-based contributions of sources to exposures was carried out. Mass fractions of each factor were scaled by unitless irritancy factors (IFs) to determine contributions of each source to total irritancy for a given participant. IFs are calculated using RD50 values (Ten Brinke, 1995; Kasanen et al., 1998), or the concentration of airborne irritant causing a 50% decrease in the respiratory rate of an animal. The RD50 for toluene is divided by the RD 50 for the compound of interest (Ten Brinke, 1995). Compounds for which IFs were available (Ten Brinke, 1995; Kasanen et al., 1998) and that were included in this calculation are indicated in Table 1. Using these IFs, the greatest irritancy-weighted contribution to exposure in NJ is from Factor 5p (p-DCB), which contributes an average of 48% of irritancy. In CA, the personal aromatics factor resembling either ETS, gasoline vapors or automobile exhaust (Factor 8p) contributes the greatest amount to irritancy, on average, with 66% of the total contribution. A factor with a similar profile (Factor 2p) is the second highest contributor to irritancy in NJ, averaging 37% of the total contribution, followed by the benzene source (Factor 6p), with 12% of irritancy. The second highest contributor to irritancy in CA is α-pinene (Factor 7p), with 12% of irritancy, followed by the consumer products factor (Factor 2p), with 11%. All other source contributions to total irritancy were small (0.5–7%) compared to these factors.
IFs were available for only 8 out of 14 and 8 out of 17 compounds studied in the NJ and CA analyses, respectively. Therefore, this analysis serves primarily to illustrate alternate methods of analyzing source apportionment results. IFs are only one of several methods of scaling source contributions to analyze different aspects or consequences of exposure to chemicals. For example, cancer slope factors or reference doses could be used to analyze chronic cancer risk or acute risk, respectively.
The main sources of personal exposure to toxic VOCs in NJ and CA in the mid to late 1980s appear to be ETS and/or automobile exhaust or gasoline vapors, followed by TCA sources, such as solvents and cleaning agents. Good agreement was found between reported activity patterns and factor scores for pumping gasoline, being exposed to ETS, drinking tap water, and visiting the drycleaners in NJ, and drinking tap water, showering, and using paints and solvents in CA.
We are continuing this study by analyzing a synthetic data set with known sources and contributions to test the robustness of PMF as a source-apportionment technique. We will also be analyzing the NJ, CA and synthetic data sets with the principal components analysis (PCA), chemical mass balance (CMB), and GRACE/SAFER (Henry et al., 1994) models. The results from all models will be compared.
Antilla P, Paatero P, Tapper U, and Järvinen O, Source identification of bulk wet deposition in Finland by positive matrix factorization, Atmos Environ (1995) 29: 1705–1718
Apte MG and Daisey JM, VOCs and “Sick Building Syndrome”: application of a new statistical approach for SBS research to U.S. EPA base study data, Indoor Air (1999 177–122
Clayton CA and Perritt RL, Data base development and data analysis for California Indoor Exposure Studies. California Air Resources Board 1993 pp. A133–187
Daisey JM Mahanama KRR and Hodgson AT, Toxic volatile organic compounds in environmental tobacco smoke: emissions factors for modeling exposures of California populations. California Air Resources Board 1994 pp. A133–186
Edgerton SA Holdren MW Smith DL and Shah JJ, Inter-urban comparison of ambient volatile organic compounds concentrations in U.S. cities, J. Air Waste Manage Assoc (1989) 39: 729–732
Henry RC Lewis CW and Collins JF, Vehicle-related hydrocarbon source compositions from ambient data: the GRACE/SAFER method, Environ Sci Technol (1994) 28: 823–832
Kasanen JP Pasanen AL Pasanen P Liesivuori J Kosma VM and Alarie Y, Stereospecificity of the sensory irritation receptor for nonreactive chemicals illustrated by pinene enantiomers, Arch Toxicol (1998) 72: 514–523
Lin C, and Milford JB, Decay-adjusted chemical mass balance receptor modeling for volatile organic compounds, Atmos Environ (1994) 28: 3261–3276
Lonneman WA Kopczynski SL Darley PE and Sutterfield FD, Hydrocarbon composition of urban air pollution, Environ Sci Technol (1974) 8: 229–236
MacIntosh DL Xue J Ozkaynak H Spengler JD and Ryan PB, A population-based exposure model for benzene, J Exposure Anal Environ Epidemiol (1995) 5: 375–403
McKone T, Human exposure to volatile organic compounds in household tap water: the indoor inhalation pathway, Environ Sci Technol (1987) 21: 1194–1201
Mølhave L, Indoor air pollution due to organic gases and vapours of solvents in building materials, Environ Int (1982) 8: 117–127
Mukund R Kelly TJ and Spicer CW, Source attribution of ambient air toxic and other VOCs in Columbus, Ohio, Atmos Environ (1996) 30: 3457–3470
National Exposure Research Laboratory (NERL), USEPA, THERdbASE Exposure Assessment Software http://www.epa.gov/nerl/heasd/therd-home.htm(accessed July 5, 2000)
Office of Air Quality Planning and Standards (OAQPS), USEPA, Taking toxics out of the air: progress in setting “maximum achievable control technology” standards under the Clean Air Act. EPA/451/K-98-001 (1998
Paatero P, Least squares formulation of robust non-negative factor analysis, Chemom Intell Lab Syst (1997) 37: 15–35
Paatero P, User's guide for positive matrix factorization programs PMF2 and PMF3 (1998) ftp://rock.helsinki.fi/pub/misc/pmf/
Paatero P, and Tapper U, Positive matrix factorization: a non-negative factor model with optimal utilization of error estimates of data values, Environmetrics (1994) 5: 111–126
Pandian MD Bradford J and Behar JV, THERDBASE: total human exposure relational database. Proceedings of the EPA/AWMA Specialty Conference: Total Exposure Assessment Methodology, A New Horizon. Air & Waste Management Association, Pittsburgh, PA 1989 pp. 204–209
Pirkle JL Flegal KM Bernett JT Brody DJ Etzel RA and Maurer KR, Exposure of the U.S. population to environmental tobacco smoke, J Am Med Assoc (1996) 275: 1233–1240
Ramadan Z, Song X-H, and Hopke PK, Identification of source of Phoenix aerosol by positive matrix factorization, J Air Waste Manage Assoc (2000) 50: 1308–1320
Rappaport SM Selvin S and Waters MA, Exposures to hydrocarbon components of gasoline in the petroleum industry, Appl Ind Hyg (1987) 2: 148–154
Rosenbaum AS Axelrad DA Woodruff TJ Wei Y Ligocki MP and Cohen JP, National estimates of outdoor air toxics concentrations, J Air Waste Manage Assoc (1999) 49: 1138–1152
Samfield MM, Indoor air quality data base for organic compounds. EPA-600-R-92-025 (1992)
Scheff PA Porter JA and Doskey PV, Improvement of VOC source fingerprints for vehicles and refineries, Paper No. 91–79.5, Air & Waste Management Association (1991)
Scheff PA Wadden RA Bates BA and Aronian PF, Source fingerprints for receptor modeling of volatile organics, J Air Pollut Control Assoc (1989) 39: 469–478
Sheldon L, Clayton A, Jones B, Keever J, Perritt R, Smith D, Whitaker D, and Whitmore R, Indoor pollutant concentrations and exposures. California Air Resources Board (CARB) A833-156 1992 pp. 9.39–9.47
Singh HB Salas LJ Cantrell BK and Redmond RM, Distribution of aromatic hydrocarbons in the ambient air, Atmos Environ (1985) 19: 1911–1919
Ten Brinke J, Development of new VOC exposure metrics and their relationship to “sick building syndrome” symptoms. Lawrence Berkeley Laboratory. LBL-37652. 1995
Unified Air Toxics Website (UATW), USEPA, Unified air toxics website: the pollutants. http://www.epa.gov/ttn/uatw/(accessed June 30, 2000)
Wallace LA, The Total Exposure Assessment Methodology (TEAM) study: summary and analysis: vol. I. EPA/600/6-87/002a (1987)
Wallace LA, The Total Exposure Assessment Methodology (TEAM) study: an analysis of exposures, sources and risks associated with four volatile organic chemicals, J Am Coll Toxicol (1989) 8: 883–895
Wallace LA, Comparison of risk from outdoor and indoor exposure to toxic chemicals, Environ Health Perspect (1991) 95: 7–13
Wallace LA, Environmental exposure to benzene: an update, Environ Health Perspect (1996) 104: 1129–1135
Wallace LA Pellizari ED Hartwell T Rosenzweig M Erickson M Sparacino C and Zelon H, Personal exposure to volatile organic compounds: I. Direct measurement in breathing-zone air, drinking water, food, and exhaled breath, Environ Res (1984) 35: 293–319
Wallace LA Pellizari ED Hartwell TD Davis V Michael LC and Whitmore RW, The influence of personal activities on exposure to volatile organic compounds, Environ Res (1989) 50: 37–55
Whitby RA and Altwicker ER, Acetylene in the atmosphere: sources, representative ambient concentrations, and ratios to other hydrocarbons, Atmos Environ (1978) 12: 1289–1296
Yakovleva E Hopke PK and Wallace L, Receptor modeling assessment of Particle Total Exposure Assessment Methodology data, Environ Sci Technol (1999) 33: 3645–3652
This research was supported by U.S. EPA grant number R82-6788-010. The PMF model was used under a licensing agreement with Pentti Paatero of the University of Helsinki. We are grateful to Eileen Daly and Bill Oliver for their assistance with the statistical analysis of the data.
Although the research described in this article has been funded by the U.S. EPA, it has not been subjected to any EPA review and therefore does not necessarily reflect the views of the Agency, and no official endorsement should be inferred.
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ANDERSON, M., MILLER, S. & MILFORD, J. Source apportionment of exposure to toxic volatile organic compounds using positive matrix factorization. J Expo Sci Environ Epidemiol 11, 295–307 (2001). https://doi.org/10.1038/sj.jea.7500168
- factor analysis
- personal exposure
- positive matrix factorization
- receptor modeling
- source apportionment
- volatile organic compounds
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