Assessment of the Contribution of Local and Regional Biomass Burning on PM2.5 in New York/New Jersey Metropolitan Area

Received: March 9, 2022
Revised: May 19, 2022
Accepted: June 7, 2022

 Copyright The Author(s). This is an open access article distributed under the terms of the Creative Commons Attribution License (CC BY 4.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original author and source are cited.

Cite this article:

Singh, S., Johnson, G., DuBois, D.W., Kavouras, I.G. (2022). Assessment of the Contribution of Local and Regional Biomass Burning on PM_2.5 in New York/New Jersey Metropolitan Area. Aerosol Air Qual. Res. 22, 220121. https://doi.org/10.4209/aaqr.220121

HIGHLIGHTS

• PM_2.5 mass concentrations have been declining due to decreasing secondary sulfate.
• Biomass burning was the predominant PM_2.5 source in metropolitan NY/NJ area.
• Primary and secondary fine particles associated with traffic emissions did change over time.

Keywords: Fine aerosol, Sources, Wildfires, Woodburning, Traffic

1 INTRODUCTION

Atmospheric particulate matter changes the earth’s energy budget directly through scattering or absorption of the solar and infrared radiation and indirectly through modification of the thermal diffusion parameters of cloud condensation nuclei (Shikwambana et al., 2020). Inhalation of atmospheric aerosol is linked to onset of cardiovascular and cerebrovascular diseases (Rajagopalan et al., 2018) including myocardial infarctions (Evans et al., 2017), ischemic stroke (Shah et al., 2015), and cardia arrythmia (Link et al., 2013), and heart failure (Shah et al., 2013). Due to implementation of a wide range of industrial emissions controls and fuel consumption policies, primary fine particle (PM_2.5, particles with aerodynamic diameter < 2.5 µm) releases and precursors of secondary inorganic aerosol species (sulfur dioxide and nitrogen oxides) were reduced up to 75% in the U.S. (U.S. EPA, 2017a). As a result, ambient PM_2.5 levels also declined, albeit at slower rate due to non-linear responses in atmospheric chemistry and biomass burning emissions (Chalbot et al., 2013; Squizzato et al., 2018a; Zhang et al., 2018; Blanchard et al., 2019, 2021).

Biomass burning includes wildland fires, agricultural fires, prescribed fires and anthropogenic bioenergy usage (wood burning, use of heating oil, and coal-fired power plant for producing electricity) and it is predominantly comprised of carbonaceous aerosol (i.e., organic carbon (OC) and elemental carbon (EC)) (Lee and Chan, 2015; Masiol et al., 2017a; Squizzato et al., 2018a). The seasonal pattern is frequently an indicator of biomass burning sources with domestic heating including wood burning in the cold months and fires in the warm period (Zhang et al., 2014). There has been a rapid increase in the frequency and intensity of large wildfires that has been linked to the longer and drier summer seasons (McClure and Jaffe, 2018). This increase was associated with higher temperatures during the warmer months, earlier snowmelt, and moisture deficits (Miller and Safford, 2012). Wildfires have been observed to steadily rise from 1984 to 2011 at a rate of 0.6–1.0 annually in the Northern Rockies and Pacific Northwest regions of the U.S. (Dennison et al., 2014). Climate change accounted for 50–60% of the larger wildfires from 1970 to 2016 in Northern Rockies (Westerling, 2016) and has doubled the cumulative wildfire areas burned from 1984 to 2015 in the western U.S. (Abatzoglou and Williams, 2016; Harvey, 2016). Emissions from wildfires have been consistently shown to contribute to ambient PM_2.5 levels in downwind urban metropolitan areas far away from the fires (Lall and Thurston, 2006; Jaffe et al., 2008; Chalbot et al., 2013; Blanchard et al., 2019; Masiol et al., 2019).

For the New York/New Jersey metropolitan statistical area (MSA), the most populous in the U.S., the 2020 PM_2.5 weighted annual average and 24-hr concentrations were 8.5 µg m^–3 and 21 µg m^–3, both being below the federal national ambient air quality standards (NAAQS). The 2020 8-hr max ozone (O₃) concentration was 73 ppbv (above the federal NAAQS). Secondary sulfate and nitrate, motor vehicle emissions, road dust, sea salt, and oil combustion were previously identified as the predominant PM_2.5 sources (Ito et al., 2004; Li et al., 2004). More recently, biomass burning has also been recognized in the region (Blanchard et al., 2019; Masiol et al., 2017a, 2017b, 2019). The latter was associated with increasing OC levels from upwind areas (Blanchard et al., 2019, 2021; Chen et al., 2022).

The objectives for this study were: (i) to identify and quanitify the contributions of PM_2.5 sources across the NY/NJ MSA; (ii) to estimate the annual trends of PM_2.5 sources and (iii) to assess the role of anthropogenic and biomass burning emissions. The PM_2.5 chemical speciation data in four sites during the 2007–2017 period were analyzed using factor analysis to apportion PM_2.5 sources. Understanding the temporal trends of PM_2.5 sources is essential to evaluate the efficacy of air pollution controls on industrial and transportation sources over the pase decades. Moreover, the possible contribution of regional and continental wildfires on PM_2.5 mass and chemical content was assessed.

 
2 METHODS

 
2.1 Sampling Sites

The concentrations of PM_2.5 mass and chemical species were retrieved from the U.S. Environmental Protection Agency’s Air Data system (U.S. EPA, 2017b) for the NCore site in Queens, New York (Site #1, EPA AIRS ID: 36-081-0124) and three PM_2.5 chemical speciation network (CSN) sites, Lower Manhattan (Site #2, Division Street, EPA AIR ID: 36-061-0134), Elizabeth New Jersey (NJ) (Site #3, Elizabeth Lab, EPA AIR ID 34-039-0004) and Chester NJ (Site 4, EPA AIR ID: 34-027-3001) for the 2007-2017 period. Fig. 1 shows the locations of the four sites within the NY/NJ MSA. Site #1 is at Queens College, near LaGuardia Airport in New York City. Site #2 is on the roof of Public School 124 in Lower Manhattan. Site #3 is at the intersection of interstate highways I-95 and I-278, near many industrial facilities including oil refineries. Site #4 is on the Department of Public Works building at Chester town in Morris County, NJ. The sites in Queens, Division Street and Elizabeth Lab are in heavily populated areas (more than 3,000,000 residents within 8 km). Chester, a smaller town, is in western NJ, upwind of the New York City metropolitan area.

Fig. 1. The PM_2.5 chemical speciation sites, population, and major road network in NY/NJ MSA.

Daily measurements of PM_2.5 mass, organic carbon (OC) and elemental carbon (EC), ions (sulfate (SO₄^2–), nitrate (NO₃^–), ammonium (NH₄⁺), potassium (K⁺) and sodium (Na⁺) were measured in 1 every 3 days frequency in Sites #1, 2, 3 and 1 every 6 days in Site #4. PM_2.5 particles collected on polytetrafluoroethylene membrane filters (PTFE) using the MetOne SASS/SuperSASS sampler were analysed gravimetrically for PM_2.5 mass and by X-ray fluorescence spectrophotometer (PANalytical Epsilon 5 analyzer). Elemental and organic carbon were measured by the thermal optical reflectance (TOR) (Sunset analyzer) for PM_2.5 collected on quartz fiber filters using the URG 3000N sampler. Water-soluble PM_2.5 ions were measured by ion chromatography (Dionex ICS-2000, ICS-3000 and Aquion systems) of nylon filters using MetOne SASS/SuperSASS sampler. The U.S. EPA laboratory standard operating protocols and QA/QC protocols are described elsewhere (U.S. EPA, 2019).

 
2.2 Aerosol Types and Source Apportionment

The major aerosol species (inorganic secondary SO₄^2–, NO₃^– and NH₄⁺), organic mass (OM), elemental carbon (EC) and soil dust concentrations were analyzed using the Interagency Monitoring of Protected Visual Environments (IMPROVE) PM_2.5 mass reconstruction scheme (Malm et al., 2004) according to Eqs. (1)–(4).

where [EC], [OC], [NO₃^–], [NH₄⁺], [SO₄^2–], [Al], [Si], [Ca], [Fe] and [Ti] were the elemental carbon, organic carbon, nitrate, ammonium, sulfate, aluminum, silica, calcium, iron and titanium concentrations (in µg m^–3), respectively. The OC/OM factor of 1.6 was used for urban PM_2.5 aerosol (Turpin and Lim, 2001). Soil dust was estimated as the sum of the crystal elements as oxides. Discrepancies between measured and reconstructed particle mass may be associated with organic aerosol with a higher OC/OM factor, that is typical of oxygenated and polyfunctional organic compounds in biomass burning aerosol (Kavouras et al., 2012). Sea salt is also not considered in the IMPROVE PM_2.5 mass reconstruction scheme, which may be important in remote marine environments.

The United States Environmental Protection Agency’s Positive Matrix Factorization (PMF) model (Version 5.0) was employed (Norris et al., 2014; Hopke, 2015, 2016). Chemical species with more than 50% of measurements above the limit of detection were included. Missing concentrations and uncertainties were substituted by the geometric mean of the measured concentrations and, four times the geometric mean of measured uncertainties, respectively. The concentrations of chemical species were analyzed by a least-squares method imposing a non-negative restriction on factor (i.e., source) contributions (G(nxp)) and profiles (F(pxm)), during minimization of the objective function (Paatero, 1997). The retained sources were rotated using the F_peak variable to reduce ambiguity of the unrotated solution. The best possible number of sources and the rotation was evaluated by a set of statistical tools (Paatero et al., 2005), and by comparison of previously published source profiles. There were up to 990 samples for Sites #1–3 and 660 samples for Site #4. In total, 25 chemical species were used. The S/N ratio varied from 2.27 to 8.49. We ran the model using the robust method for factors varying from 3 to 20 with a random seed and 20 runs per configuration during screening and 100 runs for the final solution. The Q_robust/Q_expected was 1.19. The correlation coefficient varied from 0.25 (for soluble Na⁺) to 0.95. An eight-factor model with a rotation with F_peak = 1.0 was selected. Base and F_peak bootstrapping included 200 runs using a minimum R of 0.75 and block size of 6 with random seeding. More than 80% of base and 100% of F_peak bootstrapped factors were mapped into the original factors, with no clear pattern for the unmapped factors.

 
2.3 Statistical Analysis

The non-parametric Mann-Whitney U-test (for two groups) and Kruskal-Wallis (for more than 2 groups) at α = 0.05 was used to test the significant of the difference. The annual trend was computed using the monthly PM_2.5 source contributions by applying the non-parametric sequential Mann-Kendall test at a confidence level of 95% (Kganyago and Shikwambana, 2020; Shikwambana et al., 2020). Analyses were done using IBM SPSS (Version 27) (IBM Analytics, Armonk, NY).

 
3 RESULTS AND DISCUSSION

 
3.1 PM_2.5 Types and Characteristics

Table 1 shows the PM_2.5 diagnostic ratios and reconstructed aerosol type concentrations at the four sites. The K/Fe enrichment factor (EF) was computed as the ratio of measured K/Fe to the crystal Soil K/Fe (0.56). The S/SO₄^2– mass ratio (from 3.03 to 3.67) and NH₄⁺/SO₄^2– molar ratio (from 2.50 to 2.77) demonstrated the presence of ammonium sulfate ((NH₄)₂SO₄) and ammonium bisulfate ((NH₄)HSO₄) particles (Malm et al., 2002).The OC/EC mass ratio suggested a mixture of combustion sources with a strong fossil fuel signature in urban sites (#1, 2 and 3) (from 2.34–3.32) and biomass in Chester (Site #4, 5.83 ± 0.16). The significant contribution of biomass burning was further corroborated by the increased abundance of soluble potassium (K⁺, a tracer of biomass burning) (from 58 to 64% of total K) and the EF(K/Fe) values (from 0.78 to 2.56).

Inorganic species were the predominant PM_2.5 aerosol components (from 2.76 to 4.99 µg m^–3), followed by OM (from 2.71 to 4.98 µg m^–3). Inorganic species and OM concentrations accounted for more than 80% of PM_2.5 mass with a west-to-east spatial gradient from Chester (#4) to Division Street (#2) and slightly decline for Queens College (#1). For EC and soil dust, the highest concentrations were measured at the Elizabeth Lab site (#3), at the intersection of two busy interstate highways. The ratio of the aerosol type concentration measured in urban sites (Queens, Division Str and Elizabeth Lab) as compared to that measured in Chester varied from 1.28–1.38 for inorganic species and 1.30–1.84 for OM, indicating that regional upwind sources may account for most of inorganic species and OM measured at the urban sites. On the other hand, the EC concentration ratio (2.18–3.76) and soil dust (2.36–3.16) indicated the significant contribution of local sources.

 
3.2 Source Apportionment

The reconstructed PM_2.5 mass concentrations accounted from 90% at Chester to 106% at Division Street, of measured PM_2.5 mass (Table 2). Figs. 2 and 3 illustrate the profiles and seasonal contributions of the eight PM_2.5 sources. The mean contribution of each source of PM_2.5 mass by site is presented in Table 2. The first factor was assigned to biomass burning with high concentrations of OC, EC, S and SO₄^2– (Fig. 2). The OC/EC ratio (4.92 ± 0.10) was indicative of contemporary biomass burning (Turpin and Lim, 2001). The SO₄^2–/S ratio (2.57 ± 0.13) was characteristic of a mixture of fresh and aged aerosol (Malm et al., 2002). Biomass burning contributed from 2.9 ± 0.1 µg m^–3 at Queens College to 3.5 ± 0.1 µg m^–3 at Elizabeth Lab (Table 2) with slightly higher contributions in summer and winter as compared to fall and spring (Fig. 3). Biomass burning accounted for about 37% of PM_2.5 in urban sites and 55% of PM_2.5 mass in Chester. The seasonal trend suggested the influence of local residential wood burning and regional wildland fires (Chalbot et al., 2013).

Fig. 2. Profiles of PM_2.5 sources in the NY/NJ MSA during 2007–2017.

Fig. 3. Seasonal contributions of PM_2.5 sources in the NY/NJ MSA during 2007–2017.

The second factor was assigned to secondary NO₃^– (in the form of NH₄NO₃) with high concentrations of NO₃^–, NH₄⁺, OC, EC, K⁺, K and SO₄^2– (Fig. 2). The contribution of secondary NO₃^– particles varied from 0.4 ± 0.1 µg m^–3 in Chester to 0.7 ± 0.1 µg m^–3 in Queens (from 6% to 8% of PM_2.5 mass) with the highest contributions during winter and the lowest during summer (Fig. 3). This was consistent with the favorable conditions for gas-to-particles conversion of HNO₃ in low ambient temperatures. The third factor was attributed to diesel combustions because of the high Ni, V, EC, NO₃^–, NH₄⁺, SO₄^2– and to a lesser extent to Fe and Mn concentrations (Chalbot et al., 2013). The presence of NO₃^– and SO₄^2– indicated the contribution from both industrial activities and transportation diesel engines. It accounted for less than 0.1 µg m^–3 of PM_2.5 mass in Chester, up to 0.7 ± 0.1 µg m^–3 in Queens (Table 2) (from 5% to 8% of PM_2.5 mass in Queens and Division Str. and less than 2% in Elizabeth Lab and Chester) with higher concentrations in the winter (Fig. 3).

The high concentrations of crystal Al, Si, Ca, Fe and Ti on the fourth factor indicated the presence of soil particles. The components OC, EC, SO₄^2–, S and Mg hinted at the mechanical resuspension of contaminated road dust (Fig. 2). Mineral and road dust accounted from 0.1 ± 0.1 µg m^–3 in Chester to 0.4 ± 0.1 µg m^–3 (from 2% to 4% of PM_2.5 mass). The highest contributions were computed in spring and summer due to accumulation during the winter of dust and debris deposited in curbs and road shoulders available for resuspension by traffic and patterns of regional dust transport (Fig. 3) (Chalbot et al., 2013). Sea salt particles are correlated with Na, Na⁺ and Cl in the sixth factor. It also exhibited high contributions from S, SO₄^2– and, to a lesser extent, NO₃^– and OC, indicating the possible contribution of harbor and shipping emissions. This was further supported by the SO₄^2–/S ratio (3.00 ± 0.36) and lack of NH₄⁺, indicating free H₂SO₄. The New York port is the destination of container ships (37%), oil/chemical tankers (9%), passenger cruise ships (7%), vehicles carriers (6%) and crude oil tanker (5%). It contributed, on average, less than 0.1 µg m^–3 on PM_2.5 mass concentration (less and 1%) across all four sites. Slightly higher contributions were computed in spring than those measured in winter and summer (Fig. 3).

The sixth factor was attributed to secondary SO₄^2– because of the high levels of S, SO₄^2–, NH₄⁺ and to a lesser extent of EC and OC. This was further supported by the SO₄^2–/S ratio (3.02 ± 0.09). This source accounted for 1.5 ± 0.1 µg m^–3 of PM_2.5 in Chester to 2.4 ± 0.1 µg m^–3 of PM_2.5 in Elizabeth Lab (from 24% to 28%), with the highest contributions being measured in the summer (Fig. 3). Traffic exhausts were identified due to OC, EC, S, SO₄^2–, NO₃^–, Zn and Fe high concentrations in the seventh factor. The low OC/EC ratio (1.84 ± 0.07) was comparable to those computed for urban aerosol and traffic exhausts (Turpin and Lim, 2001). Traffic added 1.5 ± 0.1 µg m^–3 of PM_2.5 in Chester up to 2.2 ± 0.1 µg m^–3 of PM_2.5 in Elizabeth Lab (Table 2) (17–23%) with slightly higher contribution in the fall at Elizabeth Lab and Division Street and no seasonal variation at Queens and Chester (Fig. 3). Lastly, ferrous, and chrome-related industrial emissions were identified because of the concentrations of heavy metals (Cr, Fe, Ni, Cu). This source contributed minimally (less than 0.1 µg m^–3 of PM_2.5; less than 1%) throughout the year (Table 2).

 
3.3 Annual Trends

The annual trends of PM_2.5 mass and source contributions are presented in Table 3. The p-value of the trend analysis is shown in parentheses if p > 0.001). PM_2.5 mass declined by 0.23 ± 0.02 µg m^–3 yr^–1 in Chester to 0.54 ± 0.05 µg m^–3 yr^–1 in Division Str (p < 0.001) This was mostly due to the significant decrease of secondary SO₄^2– (from –0.19 ± 0.02 µg m^–3 yr^–1 in Chester to –0.36 ± 0.03 µg m^–3 yr^–1 in Division Str (p < 0.001). The declining trends among all four sites were indicative of the regional SO₄^2– origins from power plants (Emami et al., 2018). The annual SO₄^2– concentration in the NY/NJ MSA dropped by 77.5%, from 3.8 µg m^–3 in 2007 to 0.9 µg m^–3 in 2017. This was comparable to national SO₂ reductions (78.2%, from 11.7 × 10⁶ tons in 2007 to 1.9 × 10⁶ tons in 2017) (U.S. EPA, 2017a).

For locally important sources such as secondary nitrate, traffic exhausts and diesel emissions, site-specific annual trends were observed. More specifically, secondary NO₃^– and primary traffic-related gasoline exhausts declined in Queens (–0.19 ± 0.02 µg m^–3 yr^–1 for secondary NO₃^– and –0.04 ± 0.01 µg m^–3 yr^–1 for traffic exhausts, p < 0.001) but slightly increased in the remaining sites, albeit without statistically insignificance (Table 3) The NO₃^– annual concentrations declined, from 1.7 µg m^–3 in 2007 to 1.0 µg m^–3 in 2017 was consistent with the national NO_x emission reduction (45.6%) (U.S. EPA, 2017a). The good agreement between reductions of secondary SO₄^2– and NO₃^– and their precursors emissions (SO₂ and NO_x), despite that NH₃ emissions remained relatively unchanged nationally, further confirmed strong regional contributions and the negligible influence of local NH₃ emissions. For diesel exhausts, contributions decreased in the three urban sites, from –0.02 ± 0.01 µg m^–3 yr^–1 in Elizabeth Lab to –0.06 ± 0.01 µg m^–3 yr^–1 in Queens and Division Str. The annual trend of mineral and road dust was comparable to that of traffic exhausts, with an increasing trend in Division Str (0.01 ± 0.01 µg m^–3 yr^–1) and Elizabeth Lab (0.04 ± 0.01 µg m^–3 yr^–1) and minimal changes in Queens and Chester. No significant trends were computed for contaminated sea salt and industrial emissions.

Biomass burning contributions declined substantially in Elizabeth Lab (–0.12 ± 0.01 µg m^–3 yr^–1, p < 0.001) and Division Str (–0.04 ± 0.01 µg m^–3 yr^–1, p = 0.22) but increased in Chester (slightly, 0.01 ± 0.01 µg m^–3 yr^–1, p = 0.1) and Queens (0.07 ± 0.01 µg m^–3 yr^–1, p < 0.001). The inconsistent annual trends may be influenced by climatology and local domestic emissions (Rattigan et al., 2016; Squizzato et al., 2018b; Pitiranggon et al., 2021). Wood burning for heating and recreational reasons is allowed in both New York and New Jersey; however due to the type of residential units, its prevalence may be higher in communities with single house units (e.g., Chester and Queens) as compared to building apartments (e.g., Division Str). The site in Elizabeth Lab is in an industrial area.

 
3.4 Biomass Burning Sources

Fig. 4 illustrates the annual trends of biomass burning in summer (May–September) and winter (November to March), the number and area burnt (in acres) by lightning ignited wildland fires in the U.S. (data obtained from the National Interagency Fire Center) and the monthly minimum temperature and number of days with temperature less than 0°C in New York City (New York City Central Park; NOAA NCDC site: USW00094728) during the 2007–2017 period. Since 2010, there appears to be a temporal correlation of the number of wildfires naturally induced by lightning and, to a lesser extent, the area burnt and summertime biomass burning contributions. Wildfires burned more than 15,000 km² per year in the U.S. in 2007–2009, as compared to 17,000 km² during the 2011–2013 period and over 20,000 km² during the 2015–2017 period. We have previously shown that changes in the El-Nino southern oscillation were correlated with wildfires in Eastern United States and episodes of smoke-related high ozone pollution (Singh and Kavouras, 2022). Higher wintertime biomass burning contributions prior to 2010 and in 2014 were associated with colder winters with an average minimum temperature of less than 9°C and prolonged periods of extreme cold more than 60 days. On the other hand, local biomass burning contributions declined when the average minimum temperature was above 0°C and the number of cold days declined.

Fig. 4. Annual trend of biomass burning contributions in summer (May–September) and winter (November–March), number and area burnt by wildfires and, minimum winter temperatures and number of days with temperature less than 0°C.

4 CONCLUSIONS

The sources of fine particles in the NY/NJ MSA for the 2007–2017 period were biomass burning (37–55%), secondary sulfate (24–28%), primary traffic emissions (17–23%), secondary nitrate (6–8%), diesel emissions (1–8%), road dust (2–4%), sea salt (< 1%), and industrial emissions (< 1%). The seasonality of secondary NO₃^–, secondary SO₄^2– and dust was consistent with those previously observed in urban environments and the effect of local meteorology and emissions. PM_2.5 mass concentrations declined by 0.23–0.54 µg m^–3 yr^–1. Secondary NO₃^– and secondary SO₄^2– declines were consistent with the reductions on NO_x and SO₂ emission from mobile and point sources, respectively. The seasonal variability of biomass burning to PM_2.5 mass was indicative of local wood burning for domestic heating and recreational activities and regional smoke from wildfires contributed to PM_2.5 in winter and summer, respectively. Wintertime biomass burning appeared to be related to the number of cold days and average minimum temperature in the region. The number and area burnt by wildfires was also associated with the interannual variability of biomass burning contribution in the summer. The findings of this study show that changes in PM_2.5 mass concentrations in the NY/NJ MSA are responding to reductions of secondary sulfate and nitrate precursors from anthropogenic sources. As a result, biomass burning is the predominant PM_2.5 source. Its contributions appear to be related to local and regional climatology affecting the frequency and intensity of cold weather in the winter and wildfires in the summer.

 
ACKNOWLEDGMENTS

The data analysis by Subraham Singh was supported in part by the Dean’s Dissertation Award.

 
DISCLAIMER

The authors declare no potential conflicts of interest.