Effects of Atmospheric Processing on the Oxidative Potential of

Article Cite This: Environ. Sci. Technol. 2019, 53, 6747−6756

pubs.acs.org/est

Effects of Atmospheric Processing on the Oxidative Potential of Biomass Burning Organic Aerosols Jenny P.S. Wong,*,† Maria Tsagkaraki,‡ Irini Tsiodra,‡ Nikolaos Mihalopoulos,§ Kalliopi Violaki,∥ Maria Kanakidou,‡ Jean Sciare,# Athanasios Nenes,†,§,⊥,∥ and Rodney J. Weber† †

Earth and Atmospheric Sciences, Georgia Institute of Technology, Atlanta, 30331, United States Environmental Chemical Processes Laboratory, Department of Chemistry, University of Crete, 70013 Heraklion, Crete Greece § IERSD, National Observatory of Athens, Palea Penteli, 15236, Greece ∥ Laboratory of Atmospheric Processes and their Impacts, School of Architecture, Civil & Environmental Engineering, É cole Polytechnique Fédérale de Lausanne, Lausanne, 1015, Switzerland ⊥ School of Chemical and Biomolecular Engineering, Georgia Institute of Technology, Atlanta, 30331, United States # Energy Environment and Water Research, The Cyprus Institute, Nicosia 1645, Cyprus

Downloaded via UNIV AUTONOMA DE COAHUILA on July 25, 2019 at 04:13:17 (UTC). See https://pubs.acs.org/sharingguidelines for options on how to legitimately share published articles.

‡

S Supporting Information *

ABSTRACT: Oxidative potential (OP), which is the ability of certain components in atmospheric particles to generate reactive oxidative species (ROS) and deplete antioxidants in vivo, is a prevailing toxicological mechanism underlying the adverse health effects associated with exposure to ambient aerosols. While previous studies have identified the high OP of fresh biomass burning organic aerosols (BBOA), it remains unclear how it evolves throughout atmospheric transport. Using the dithiothreitol (DTT) assay as a measure of OP, a combination of field observations and laboratory experiments is used to determine how atmospheric aging transforms the intrinsic OP (OPDTT mass ) of BBOA. For ambient BBOA collected during the fire seasons in Greece, OPDTT mass was observed to increase by a factor of 2.1 ± 0.9 for samples of atmospheric ages up to 68 h. Laboratory experiments indicate that aqueous photochemical aging (aging by UVB and UVA photolysis; as well as OH oxidation), as well as aging by ozone and atmospheric dilution can transform the OPDTT mass of the water-soluble fraction of wood smoke within 2 days of atmospheric transport. The results from this work suggest that the air quality impacts of biomass burning emissions can extend beyond regions near fire sites and should be accounted for.

1. INTRODUCTION Correlation between increasing ambient particulate matter (PM) mass concentration and adverse human health is well documented.1−4 Ambient PM exposure is a leading environmental risk factor for global premature mortality,5 yet the mechanisms causing these effects are not well understood. Oxidative stress, an imbalance of cellular reactive oxygen species (ROS) and antioxidants, is a prevailing mechanism thought to be responsible for the adverse health effects associated with particle inhalation.6,7 Recent epidemiological studies demonstrated that the ability of particles to consume antioxidants and/or generate ROS, referred to as oxidative potential (OP), can be more strongly associated with the reduction in respiratory function than PM mass, suggesting that it is a health relevant metric for air quality.8−10 The dithiothreitol (DTT) assay is an acellular technique developed to quantify particle OP by monitoring the consumption of DTT, a surrogate for biological antioxidants, in the presence of particle components under physiologically relevant conditions.11,12 The biological relevance of this technique has also been demonstrated, as the consumption rate of DTT, also known as DTT activity (OPDTT), has been © 2019 American Chemical Society

linked with the expression of cellular markers of oxidative stress13 and inflammation,13−15 as well as cellular cytotoxic responses.16 Various chemical components in atmospheric aerosols have been demonstrated to be well-correlated with OPDTT, including water-soluble transition metal ions (e.g., water-soluble copper17−19) and organic compounds, such as water-soluble organic carbon,19,20 and more specifically, quinones.21 As summarized in Shiraiwa et al.,22 there have been extensive research efforts to characterize the intrinsic, or mass-normalized OPDTT (henceforth referred to as OPDTT mass ) of numerous particle sources, such as secondary organic aerosols (SOA) formed from biogenic volatile organic compounds and emissions from vehicle exhaust with respect to fuel types and emissions control technologies. Collectively, these research efforts suggest that the prevailing conceptual framework of aerosol toxicity is to link particle OP to its sources and emission characteristics. However, it is also critical to Received: Revised: Accepted: Published: 6747

February 18, 2019 May 7, 2019 May 15, 2019 May 15, 2019 DOI: 10.1021/acs.est.9b01034 Environ. Sci. Technol. 2019, 53, 6747−6756

Article

Environmental Science & Technology

filter samples with biomass burning influence, as described in detail in Supporting Information (SI) Section S1. This analysis also provided an estimation of the corresponding atmospheric transport time. An example of the analysis approach is shown in SI Figure S1. We note that there are uncertainties associated with the estimated atmospheric transport times for field samples, either due to unaccounted contributions from fires occurring more than 3 days prior to filter collection (i.e., HYSPLIT runs times of 72 h were used) or small fires (fire radiative power less than 100 MW), which were not included in the current analysis. Additionally, this analysis approach does not account for variability in the emission of BBOA from various fires. From the 65 samples collected, we focus on 24 that were impacted by wildfires. Separate filter punches from each filter were used to determine the water-soluble and total (i.e., water-soluble and insoluble components) DTT activity, as well as concentrations of water-soluble organic carbon (WSOC) and organic carbon (OC), described in Section 2.4. We note that for filter samples that were identified as influenced by biomass burning using the MODIS/HYSPLIT approach described above, had elevated concentrations of non sea salt potassium (ion chromatography methodology described in our previous work40), a commonly used chemical tracer for biomass burning, compared to samples that were identified as not being directly influenced by biomass burning emissions (P = 0.04). 2.2. Laboratory Experiments: BBOA Preparation and Bulk Photochemical Aging Experiments. The sample preparation and experimental procedures for the photochemical aging experiments are described in detail in our previous work.39,40 Briefly, primary BBOA was generated in the laboratory from the pyrolysis of dry hardwood (cherry) in an electronically heated pyrolyzation chamber in an oxygen-free atmosphere at 210 °C, mimicking particles formed under smoldering conditions. Particles formed were collected on PTFE filters (47 mm, 2 μm pore size, Pall Corporation) and extracted for water-soluble (WS) components using 15 mL of purified water via sonication for 60 min. The resulting extract was subsequently filtered (0.2 μm PTFE syringe filter, Fisher) to remove any insoluble materials and then divided into multiple glass vials (1 mL/vial) for photochemical aging. The photochemical aging experiments were conducted in a photoreactor, where multiple vials containing the WS-BBOA extract were continuously exposed to UVB or UVA radiation. For both types of UV lamps, most of the radiation fell in the range of 300−400 nm, with the maximum emission at 310 and 355 nm for the UVB and UVA lamps, respectively (measured wavelength dependent photon fluxes are shown in SI Figure S2). Note that while both UVB and UVA lamps have similar photon fluxes for wavelengths lower than 320 nm, the UVA lamps have much higher photon flux at higher wavelengths. These spectral differences allowed for assessing whether the effects of aging by UV radiation on OPDTT mass are wavelength dependent. For these experiments, the WS-BBOA was illuminated by either UVB or UVA radiation for up to 100 h. For aqueous OH oxidation experiments, H2O2 was added to each glass vial containing the WS-BBOA (final concentration of 1.5 mM) as a photolytic source of OH radicals upon irradiation by UVB lights (up to 18 h). The photolysis of the added 1.5 mM H2O2 resulted in a steady state OH concentration of 5.9 × 10−15 M, which was determined in separate experiments by using a commonly employed OH scavenger, benzoate.41,42 The procedure to determine the

understand whether particle OP evolves throughout its atmospheric lifespan, as it can be transported and affect human populations over great distances from emission sources. As particles are transported away from the source region, they are subjected to physical and chemical aging processes that transform their properties,23,24 such as gas-particle partitioning (e.g., evaporation or condensation of vapors), as well as interactions with solar radiation and atmospheric oxidants (e.g., ozone, OH and other radicals). Early work by Verma et al. observed that the OPDTT mass and fraction of watersoluble organic carbon of ambient particles collected in downtown Los Angeles increased during the afternoon, suggesting the effects of photochemical aging.25 Detailed mechanistic laboratory studies have also shown that exposures to ozone increased the OPDTT mass of diesel engine exhaust soot particles,26−28 likely from the formation of quinones, a class of DTT-active compounds.21 Other laboratory studies have demonstrated that photochemical aging by OH radicals can transform the toxicity of certain types of organic aerosols.29−33 A recent study demonstrated that reactions in fog droplets enhanced the toxicity of ambient organic aerosols in Italy.34 Other aging processes can modulate the OPDTT mass , such as the conversion of insoluble (or DTT-inaccessible) metal oxide forms of transition metals to soluble or DTT-accessible (i.e., surface-bound) forms, from acid dissolution35 and complexation with organic ligands.36 Given that biomass burning is a major contributor to global aerosols37 and OPDTT mass of biomass burning organic aerosols (BBOA) is found to be significantly higher than other identified components of organic aerosols in the southeast U.S.,38 BBOA may contribute substantially to the toxicity of ambient aerosols. The objective of this work is to investigate the transformation of OPDTT mass of BBOA by atmospheric aging, through field observations and detailed laboratory studies. Analysis of filter samples influenced by biomass burning emissions of different atmospheric ages, collected during the fire seasons in Greece, indicated that the OPDTT mass of BBOA increased throughout its atmospheric transport. To identify the potential aging mechanism that led to these ambient observations, the effects of multiple photochemical aging processes (aqueous OH oxidation, exposure to UVB and UVA radiation), along with aging by ozone and atmospheric dilution, were systematically investigated in the laboratory.

2. MATERIALS AND METHODS 2.1. Field Observations on Crete, Greece. Filter sampling was conducted on the island of Crete, Greece, during the fire seasons (July to October) of 2016 and 2017. A high volume total suspended particulate (TSP) sampler (TISCH) was used to collect PM10 particles on prebaked 8 × 10 in. quartz filters (Pall, 2500QAT-UP) for 22−24 h at a flow rate of 1.4 (2016 sampling period) and 2 m3 min−1 (2017 sampling period), respectively. Crete is an ideal sampling site to assess the effects of atmospheric aging of biomass burning as persistent prevailing winds blowing from the north during late summer across the Aegean Sea transport biomass burning emissions from different fires in the Mediterranean region and Eastern Continental Europe to the sampling site. Another advantage of the sampling site location is that no further biomass burning emissions can occur during the transport over the Aegean Sea. A combination of fire data (e.g., fire emissions, time and location) as detected by MODIS and air-mass back trajectories, computed using HYSPLIT, were used to identify 6748

DOI: 10.1021/acs.est.9b01034 Environ. Sci. Technol. 2019, 53, 6747−6756

Article

Environmental Science & Technology photon fluxes in the photoreactor and OH steady state concentration has also been previously described in our earlier work.39,40 For all photochemical aging experiments, at different illumination times, one vial from an identically prepared set was removed from the photoreactor to determine OPDTT mass as a function of photochemical age (described in Sections 2.4 and 2.5). Given that the photochemical aging experiments were conducted using BBOA dissolved in bulk aqueous solutions, these experimental conditions (i.e., solute concentrations and viscosity) most likely represent the bulk aqueous-phase aging of BBOA occurring in fog or cloud droplets and that not all processes occurring throughout the cloud cycling process, such as droplet evaporation, are examined. 2.3. Exploratory Laboratory Aging Experiments. Exploratory experiments were also conducted in the laboratory to assess the potential of aging by ozone and further atmospheric dilution that may occur sometime after emission, as processes that can transform the OPDTT mass of primary BBOA. The experimental setup to assess the potential effects of ozone is shown in SI Figure S3a. Varying mixing ratios of ozone (40− 600 ppb) were generated by an ozone calibrator (49 CPS, Thermo Science), using purified air. A portion of this ozone flow passed through a filter holder that contained six 1 cm2 PTFE filter punches of lab generated primary BBOA from cherry wood pyrolysis (described previously) for up to 6 h. At different times, one filter punch was removed and placed in a precleaned glass vial containing 1.1 mL of purified water (18.2 mΩ) and the WS-BBOA was extracted and filtered. The DTT resulting extract was analyzed to determine the OPmass WS‑DTT (Section 2.4). OPmass did not change in control experiments where ozone was not generated, suggesting that the exposure of filter punches to the air flow did not affect OPDTT mass (i.e., there was no volatilization of DTT active components from the air flow). The experimental setup used to investigate the potential effects of atmospheric dilution on the OPDTT mass of primary BBOA is shown in SI Figure S3b. For these experiments, prior to collection on filters, the smoke stream containing the laboratory generated BBOA from cherry wood pyrolysis was diluted by factors from 30 to 3000 using a two-step dilution system. Here, the generated smoke stream was mixed with the first dilution flow of purified air in a mixing volume (0.01 m3), then a portion of this smoke-containing flow was further diluted by a varying flow of purified air. This second dilution step was only employed to assess the effects of additional dilution (i.e., the BBOA generated for all other aging experiments experienced a dilution factor of ∼30 due to the mixing of the smoke plume with the first dilution flow). The resulting diluted smoke was then collected on PTFE filters, extracted, filtered and analyzed for OPWS‑DTT (Section 2.4). mass Note that activated carbon traps were not used to remove any of the volatile components of wood smoke and that a backup filter (i.e., behind that of the PTFE filter) was not employed in these experiments to assess positive sampling artifact arising from gas adsorption of semivolatile components onto BBOA collected on the filter.43−45 The experiment at each dilution factor was repeated at least three times to assess experimental variability. For all laboratory aging experiments, only the effects of aging on the OPDTT mass of the water-soluble fraction of BBOA were studied, as the photochemical aging experiments were conducted using BBOA dissolved in aqueous solutions. This is in contrast to filter samples collected in the field (Section

2.2) where the effects of aging on both the water-soluble and total (e.g., water-soluble and−insoluble) fractions of BBOA were examined. 2.4. DTT Activity. Using the dithiothreitol (DTT) assay, OP of the WS and total (i.e., water-soluble and -insoluble) fractions of BBOA were measured, using the procedure as outlined in Fang et al.46 and Gao et al.47 The detailed procedure for these measurements is also described in SI Section S2. Here, the WS fraction represents the BBOA extract that was filtered (0.2 μm PTFE syringe filter, Fisher) to remove insoluble materials and the total fraction represents the unfiltered extract, with the sample filter punch remaining in solution throughout the reaction with DTT. The decay rates of DTT due to its reaction with BBOA components at pH 7.4 and T = 37 °C were determined using absorption spectroscopy and were normalized by either the water-soluble organic carbon (WSOC) or organic carbon (OC) concentration of the aerosol component in the DTT reaction mixture, to represent the mass-normalized oxidative potential of the WS and total fractions of BBOA, henceforth denoted as OPWS‑DTT and mass OPTotal‑DTT , respectively. The measurement variability (i.e., mass precision) for both OPWS‑DTT and OPTotal‑DTT is approximately mass mass 20%, estimated from blanks, standards, and field sample duplicates. 2.5. Other Particle Measurements. Since the DTT assay is responsive to copper (Cu) in addition to organic compounds,17 the potential contribution of both total (i.e., water-insoluble plus water-soluble) and water-soluble (WS) Cu in ambient filter samples to the decay of DTT needs to be accounted for. Even though Cu is not expected to be a significant component of biomass burning smoke,48 the samples were collected using a TSP and so will include contributions of coarse mode species, such as mineral dust that can contain metals. This was assessed by determining the concentrations of total and WS-Cu in the ambient filter samples using inductively coupled plasma mass spectrometry (ICP−MS) (method described in described in SI Section S3) and the dependence of DTT decay rate on the Cu concentration (SI Figure S4; method described in SI Section S4).

3. RESULTS AND DISCUSSION 3.1. Field Observations. The OC and WSOC concentrations, along with OPTotal‑DTT and OPWS‑DTT of ambient mass mass filters with biomass burning influence were binned according to the transport times estimated from back trajectories, to illustrate the effects of atmospheric aging (Figure 1a,b). Here, the bins “fresh,” “intermediate,” and “more aged” represent ambient filters with atmospheric transport times of 0−5 h, 5.1−26 h, and 26.1−68 h, respectively. Higher time resolution binning was not used due to limited number of filters collected. Total‑DTT Field observations indicated that both OPmass and WS‑DTT OPmass increased with atmospheric ages, with the majority of the increase occurring within the first 26 h of atmospheric transport. For fresh and intermediate ambient BBOA, the changes in OPTotal‑DTT (ANOVA, P = 0.077) and OPWS‑DTT (P mass mass = 0.064) are statistically significant, whereas the corresponding change for intermediate to more aged BBOA are not (P > 0.5). DTT The trends of increasing OPTotal and OPWS‑DTT follow mass mass closely to those of decreasing OC and WSOC, suggesting that the enhancement in the OPDTT could be driven by the loss of less or non-DTT active organic components in BBOA (recall that values of OPTotal‑DTT and OPWS‑DTT represent the OC/ mass mass 6749


Article


the more aged ambient samples may have enhanced contributions from non biomass burning sources of insoluble copper compared to WS-Cu. We note that these non biomass burning sources of Cu appear not to be associated with the sources of the organic component in these ambient sample as poor correlations (Pearson’s r) between the concentrations of Total-Cu and WS-Cu to WSOC (r = −0.17 to −0.14) and OC (r = 0.19 to 0.31) were observed. Nevertheless, in order to assess whether the observed increases in OPDTT mass were due to increases in copper concentrations, we assumed the approach by Charrier et al.17 where the effects of copper and organic compounds to DTT consumption were considered to be additive. Accounting for the contributions of WS-Cu to DTT consumption, the trend of increasing OPDTT mass as a function of atmospheric transport time remains (SI Figure S5a), suggesting that the majority of the changes in OPTotal‑DTT mass and OPWS‑DTT were due to the changes in the OP of organics mass in BB aerosols from atmospheric aging. In addition, the fractional contribution of Cu to DTT consumption rates did not change significantly with atmospheric age (SI Figure S5b), where on average, Cu contributed 38 ± 12% of the DTT consumption rates for all ambient samples (analysis approach is described in SI Section S3). This approach likely result in an lower estimate on intrinsic DTT activity of BBOA, since Yu et al.,49 have shown that WS-Cu mixed with humic like substances (HULIS) result in a small reduction (∼25%) in DTT compared to the sum of the DTT consumption by WSCu and HULIS individually. However, it remains unclear whether the interactions of compounds in BBOA with Cu have a net antagonistic effect on DTT activity, and if so, to what extent. Moreover, it remains unclear whether these interactions are affected when the organics are aged. In general, the effects of organic-transition metal interaction on OPDTT mass are warrant future studies. 3.2. Laboratory Experiments. To identify potential atmospheric aging processes that could have resulted in the ambient observations, we systematically investigated how various atmospheric aging processes transformed the OPWS‑DTT mass of laboratory generated BBOA. 3.2.1. Aqueous-Phase Photochemical Aging. Relative WS‑DTT changes in the WSOC concentration and OPmass with respect to initial (i.e., preaging) values for laboratory generated BBOA from wood smoke due to different bulk photochemical aging processes (exposure to UV radiation and aqueous OH oxidation) are shown in Figure 2. Note that in these was not determined and that field experiments OPTotal‑DTT mass observations (Section 3.1) indicated that the WS fraction of ambient BBOA contributed to the majority of OPTotal‑DTT. For of fresh BBOA was 48 ± 6 pmol min−1 reference, OPWS‑DTT mass −1 ug of WSOC, which is comparable to other studies of for fresh BBOA emitted from laboratory burns of OPWS‑DTT mass hickory wood under smoldering and flaming conditions ((9.0− 15) pmol min−1 ug−1 of organic aerosol),31 as well as for various types of grain crop residues ((9.0−30) pmol min−1 ug−1 of organic aerosol) under flaming conditions,50 assuming the conversion of OC to organic aerosol concentrations using a factor of 2.51 Considering the effects of UVB and UVA exposure first, as shown in Figure 2, initial increases in OPWS‑DTT were observed mass in the first 5 h of illumination. Given that there are no significant changes in WSOC during this time period, this suggests that the reactions involving WS-BBOA components initiated by the exposure to UV radiation form products that

Figure 1. Box and whisker plots of (a) OPTotal‑DTT (black) and mass (blue); (b) organic carbon (OC, black) and water-soluble OPWS‑DTT mass organic carbon (WSOC, blue) concentrations; (c) total copper (Total-Cu; black) and water-soluble copper (WS-Cu; blue) concentrations for fresh (≤5 h atmospheric transport time), intermediate (5.1−26 h) and more aged (26.1−68 h) BBOA ambient filters collected at Heraklion on Crete Island, Greece. The lines in the boxes indicate the median values, the upper and lower box boundaries indicate the 75th and 25th percentiles, the whiskers indicate the 90th and 10th percentiles, and the markers indicate the mean values for each corresponding age bin. The bracketed numbers reflect the number of ambient filters in each atmospheric age bin.

WSOC mass-normalized OPDTT). The increase in OPDTT mass may also be due to the formation of more DTT-active components, especially during the initial period of atmospheric transport (i.e., from “fresh” to “intermediate” aged BBOA), where on average, OPTotal‑DTT and OPWS‑DTT increased by a factor of 1.8, mass mass whereas OC and WSOC decreased by only a factor of 1.3. These field observations suggest that as the biomass burning plume is transported away from the fire site, atmospheric aging processes overall increase the toxicity (as measured by the DTT assay) of ambient BBOA. Figure 1 also shows that for BBOA of all atmospheric age bins, OPTotal‑DTT is lower than mass OPWS‑DTT , indicating that a larger portion of the watermass insoluble fraction of BBOA is less or non-DTT active compared to the water-soluble fraction. This is consistent with a previous study where the mass-normalized OP of watersoluble organic compounds in fog droplets were found to be higher compared to interstitial organic aerosols, potentially due to either the scavenging of water-soluble DTT active components by fog droplets or their formation in the aqueous phase.34 For all the filter samples considered, the WS fraction of BBOA contributed 82 ± 16% of the OPTotal‑DTT, suggesting that majority of the DTT active components in BBOA are water-soluble. Total‑DTT We note that the increases in both OPmass and WS‑DTT OP mass may be partially driven by the increasing concentrations of WS-Cu for more aged ambient samples (Figure 1c), in addition to aging of the organic component. The increase in WS-Cu concentrations may be due to the conversion of insoluble forms of copper to water-soluble forms, due to aging processes.35 The increase in Total-Cu is more significant with age compared to WS-Cu, which suggests that 6750


Article


aging processes to transform the OP of BBOA in the atmosphere, where three main effects were observed (Figure 2): (1) the increase in OPWS‑DTT during the initial stages of mass aging by UVB radiation; as well as the decrease in OPWS‑DTT mass due to (2) later stages of UVB aging; and (3) aqueous OH oxidation. 3.2.2. Comparison of Photochemical Aging Lab Experiments to Field Observations. To compare these laboratory results to field observations, which will provide insight into whether the aging mechanisms studied in the laboratory experiments could have contributed to the changes in WS‑DTT OPmass for the ambient samples, the rate constants corresponding to the three main effects observed due to aging by UVB radiation and aqueous OH oxidation were determined using the approach outlined in Wong et al. (SI Table S1),40 to estimate the rates of OPWS‑DTT transformation mass under atmospherically relevant conditions. Note that a previous ambient study of biomass burning aerosols on Crete Island, Greece during the 2012 fire season observed a drastic decrease in the heterogeneity in mixing state and chemical composition for particles above the size threshold for cloud droplet formation compared to those below the threshold, suggesting that the larger aerosols were cloud processed.52 Assuming typical conditions in atmospheric cloud droplets (i.e., average steady state OH concentration of 1 × 10−14 M),53 the laboratory results suggest that aqueous OH oxidation of BBOA in cloud droplets will lead to a 50% reduction of OPWS‑DTT within 3.6 h in-cloud time. Given that an increase in mass was observed for ambient samples of atmospheric OPWS‑DTT mass transport time up to 68 h, this suggests that the ambient BBOA may not have been subjected to aqueous OH oxidation in cloud droplets. Assuming that reaction kinetics for the OH oxidation of BBOA components in aqueous particles is comparable to that in cloud droplets, the corresponding halflife for OPWS‑DTT due to OH oxidation of BBOA in aqueous mass particles is estimated to be much longer, at 36 h, since the steady state concentration of OH in aqueous particles is estimated to be an order of magnitude lower than that in a cloud or fog droplet.53 Given that OPWS‑DTT for ambient mass BBOA was observed to increase for samples with longer atmospheric transport times, it is unlikely that OH oxidation of BBOA components in aqueous particles had a significant . impact on the observed transformation of OPWS‑DTT mass Figure 2 shows that exposure to UV radiation can lead to both an increase and decrease in OPWS‑DTT for BBOA. Given mass that the integrated UV photon fluxes in these laboratory experiments are roughly ∼90% of the sun at solar noon (i.e., 1 h of UV exposure in the laboratory is equivalent to ∼0.9 h of ambient UV exposure at solar noon), and assuming that the compounds that participate in the initial enhancement in OPWS‑DTT do not have additional atmospheric sources (e.g., mass partitioning from the gas-phase), then the OPWS‑DTT for BBOA mass is expected to reach its maximum due to aging by UV in about 6 h of atmospheric transport. OPWS‑DTT is also estimated to mass reduce to 50% of preaging values after ∼45 h of atmospheric transport. Given that the actual actinic flux over the course of a day is lower than that of solar noon, assuming that the average daily actinic flux is 0.25 times of that at solar noon, OPWS‑DTT is mass expected to reach its maximum after ∼20 h of atmospheric transport. This approximate analysis suggests that aging by UVB radiation may have contributed to the observed increase for ambient BBOA, at least in the first few hours in OPWS‑DTT mass

Figure 2. Changes (relative to preaging values) in (a) OPWS‑DTT and mass (b) WSOC for laboratory generated BBOA due to aqueous OH oxidation (green triangles), exposure to UVB (blue circles) and UVA (black squares) radiation. The shaded regions represent the variability (±1σ) of multiple experiments (n = 3).

are more DTT active and of higher intrinsic DTT activity (i.e., higher OPWS‑DTT ). This rate of increase slowed down after mass about 1 h of illumination, which may be due to the depletion of compounds participating in the reactions leading to this enhancement, or that additional aging by UV exposure, at least up to 5 h of illumination, do not result in further net increases in the DTT activity of the WS-BBOA (e.g., competition between reactions that leads to an increase and decrease in DTT activity). After this initial period (∼5 h) of increased OPWS‑DTT , further exposure to UV radiation led to decreases in mass WSOC and OPWS‑DTT , suggesting that later stages of aging by mass UV radiation depletes DTT active WS-BBOA components. Despite the enhanced photon flux at wavelengths larger than 320 nm for UVA (by a factor of 10−27) compared to UVB lights, no significant differences were observed for the changes in WSOC and OPWS‑DTT , suggesting that reactions resulting in mass the changes in OPWS‑DTT were mostly induced by UVB mass radiation. Due to the presence of UVB lights to generate OH radicals (from the photolysis of H2O2), the effects of aqueous OH oxidation were taken to be the difference between the UVB + H2O2 and UVB experiments. Shown in Figure 2, up to 1 h of illumination, OH oxidation suppressed the enhancement in OPWS‑DTT due to UVB exposure (i.e., the green trace is lower mass than the blue trace), following which additional reactions with OH lead to rapid decrease in OPWS‑DTT . After 18 h of aqueous mass OH aging, only 10% of the initial OPWS‑DTT remained. Given mass that aqueous−OH oxidation did not significantly affect concentrations of WSOC, the decrease in OPWS‑DTT is likely mass driven by the formation of non- or comparatively less-DTT active components by the OH oxidation of WS BBOA. These laboratory results suggest that aqueous OH oxidation of WSBBOA components generally leads to a reduction in OPWS‑DTT . mass This was also reported recently by Jiang and Jang31 where OH photooxidation of laboratory generated BBOA (from the pyrolysis of hickory wood) for 7 h decreased the OPWS‑DTT up mass to 75%. This decrease in OPWS‑DTT correlated with a decrease mass in quinone concentrations in the BBOA,31 suggesting that the OH photooxidation products of quinones in BBOA are less DTT activSe compared to their precursors. Collectively, the laboratory results from these aging experiments demonstrate the potential of photochemical 6751


Article

Environmental Science & Technology (for actinic fluxes corresponding to solar noon) and up to approximately 1 day (for actinic fluxes 0.25 times of solar noon) of atmospheric transport. However, the observed increase in OPWS‑DTT for ambient BBOA persisted for a mass sufficiently longer period of time (up to 68 h), suggesting that other aging processes contributed to the observed increase in OPWS‑DTT at longer atmospheric transport times. In addition, mass most of the observed increase in OPWS‑DTT for ambient BBOA mass is likely driven by the decrease in WSOC, as noted above and seen in Figure 1, opposed to the UV aging experiments for where WSOC did not change significantly. These contrasting behaviors further suggest that at longer atmospheric transport times, the contribution of other aging processes resulting in the increase in OPWS‑DTT is likely to be more significant. mass 3.3. Other Potential Aging Mechanisms. We also explored the potential of other atmospheric aging processes to transform OPWS‑DTT of BBOA in the laboratory and mass compared these results to field observations; of particular interest is aging by ozone and the evaporation of volatile components in BBOA due to atmospheric dilution. The effects of ozone were investigated by exposing labgenerated BBOA on filters to varying levels of ozone (Figure 3). Despite the considerable scatter, initial increasing trends in

OPWS‑DTT were observed for ozone exposures higher than 4 × mass 103 molecules cm−3 hr, indicating that the transformation of OPWS‑DTT by ozone occurs during early stages of fire plume mass aging and leads to increased aerosol toxicity. Given that the typical ozone level in the Mediterranean region is approximately 50 ppbv,59 these laboratory results suggest that for ambient BBOA that are aged by ozone, its OPWS‑DTT would peak within approximately 8 h of atmospheric mass transport, indicating that aging by ozone could have contributed to the initial increase in the OPWS‑DTT observed mass for ambient BBOA. Beyond 8 h of aging by ozone, OPWS‑DTT mass of laboratory generated BBOA by ozone decreased, whereas OPWS‑DTT of ambient BBOA persisted for a much longer mass period of time. This suggests that at longer atmospheric transport times, other aging processes, such as enhancement in OPWS‑DTT due to aging by UV (up to 1 day of atmospheric mass transport), which were discussed previously, are likely WS‑DTT important processes affecting the OPmass for ambient BBOA. To simulate the effects of dilution, which occurs during the mixing of the biomass burning plume with background airmasses, the emitted smoke plume from wood pyrolysis in the laboratory experiments was mixed with purified air at different dilution factors (DF) prior to filter collection. Shown in Figure 4, significant changes in WSOC and OPWS‑DTT were observed mass

Figure 3. Changes (relative to preaging values) in WSOC (black (red squares) for laboratory generated BBOA circles) and OPWS‑DTT mass vs varying ozone exposure. The equivalent hours of ambient ozone exposures at 50 ppbv are provided in the top axis to aid in the comparison to field observations. The error bars represent the variability (±1σ) of three experimental replicates, respectively.

Figure 4. Changes (relative to preaging values) in WSOC (black (red squares) for laboratory generated BBOA circles) and OPWS‑DTT mass vs varying dilution factors. The error bars represent the variability (±1σ) of three experimental replicates, respectively. Note that for during the generation of BBOA, the emissions experienced a dilution factor of 30 (i.e., the initial condition), which is indicated by the arrow.

WSOC and OPWS‑DTT due to exposures of BBOA on filters to mass ozone up to 1 × 1013 molecules cm−3 hr were observed, where both WSOC and OPWS‑DTT increased up to a factor of 1.4 (tmass test, P = 0.026). Given that there are no additional sources of organics in the gas phase in these experiments, we speculate that this increase in both WSOC and OPWS‑DTT is likely from mass the reaction of ozone with BBOA on the filters, of which the oxidation products are more water-soluble and DTT active. For example, ozonolysis of polycyclic aromatic hydrocarbons (PAHs), which have been identified in ambient BBOA54,55 and are not highly water-soluble nor DTT active,17,56 are known to form quinones,57,58 which are highly DTT active12,17 and potentially more water-soluble due to increased functionalization. It is possible that other oxidation products could have also contributed to this increase in OPWS‑DTT as it has been mass previously demonstrated that for soot aged by ozone, only 45% of the increased OPWS‑DTT can be attributed to the formation mass of quinones from the reaction of ozone with PAHs (or functionalized PAHs) present on soot particles.21 Following this period of increased OPWS‑DTT , additional exposures to mass ozone from 2 to 4 × 103 molecules cm−3 hr appears to reduce the OPWS‑DTT to the initial preaging value (t-test, P = 0.037), mass suggesting that upon further aging by ozone, some of the DTT active compounds are depleted. No changes in WSOC and

at DF of 200 and greater; on average, WSOC decreased by a increased by a factor of 1.8. Note factor of 2.3 and OPWS‑DTT mass that the changes in WSOC for DF 70 and 130 are not statistically significant (ANOVA, P > 0.40). The loss of WSOC due to increasing dilution indicates that some water-soluble components of BBOA are semivolatile. Given that the coincident changes in WSOC and OPWS‑DTT are also similar mass in magnitude, we speculate that this increase in OPWS‑DTT is mass likely driven by the loss of the semivolatile organic compounds that are non- or comparatively less-DTT active, as they partition from the particle to the gas phase during dilution. Increasing DFs from 600 to 3000 did not lead to additional WS‑DTT changes in WSOC and OPmass , suggesting that the remaining components in BBOA (DTT active and not) are of low-volatility. We note that these results are qualitative, as these experiments were likely conducted at much higher particle concentration compared to ambient conditions, which will shift the gas-particle partitioning of semivolatile components toward the particle phase. Additionally, the loss of WSOC does not scale with the DF (e.g., for a completely nonvolatile OA, the loss in WSOC should scale linearly with 6752


Article


focused on the aging of primary BBOA (i.e., aerosols emitted directly from biomass burning emissions); the semivolatile nature of BBOA means that significant concentrations of organic compounds are also present in the gas phase where aging of these gas-phase components is known to form new organic aerosols (e.g., SOA).62 For example, SOA generated from the photooxidation of naphthalene, one of the dominant gas phase PAHs emitted during biomass burning,63 are very DTT active, with OPWS‑DTT of 153 ± 50 pmol min−1 μg−1.64 mass DTT The OPmass of SOA generated from other gaseous biomass burning emissions remains largely uncharacterized and should be explored to assess whether this aging process affects the WS‑DTT OPDTT for mass of BBOA. Results also indicated that OPmass ambient BBOA is transformed throughout the whole day, and not limited to processes occurring only during the daylight hours, such as photochemical aging, which has generally been the focus of previous studies examining the effects of aging on OPDTT of atmospheric aerosols.29−33 While aqueous OH oxidation in cloud droplets likely did not contribute to the observed changes in OPDTT mass for ambient BBOA, laboratory results suggest that for BBOA that are subjected to cloud processing, their OPDTT mass will rapidly diminish within ∼4 h. These results indicate that the evolution of OPDTT mass for BBOA is strongly dependent on the ability of particles to be cloud processed, which is dependent on their hygroscopicity and meteorological conditions (i.e., clear-sky vs cloudy days). We also note that the effects of individual atmospheric aging processes on the water insoluble components of BBOA were not investigated in the laboratory and should be explored in future studies. Collectively, the results from this work highlight the fact that toxicity of BBOA does not only depend on their emissions but also on atmospheric aging, and that their air quality impacts can extend beyond regions near fire sites. In particular, while concentrations of BBOA decrease as smoke plumes are diluted during atmospheric transport, the results from this work suggest that the degree of potential toxicity due to exposure to BBOA may remain high, as the intrinsic toxicity of BBOA was observed to increase with atmospheric age. In addition, the use of OPDTT mass determined from emissions to estimate DTT activity per volume of air (OPDTT vol ), which is the metric relevant for human exposure and any resulting adverse health effects, would significantly underestimate OPDTT since OPDTT mass was observed to increase with age. Accounting for aging effects is critical for accurate predictions of the health impacts of BBOA.

the increase in DF), which is consistent with a positive artifact from the adsorption of semivolatile organic gases emitted during wood pyrolysis onto the particles collected on filters.42,44 Thus, the observations from these experiments likely represent a lower limit of the effects of dilution on the OPWS‑DTT of BBOA (i.e., the increase in OPWS‑DTT may be mass mass more significant and may occur at DFs lower than 200 in the atmosphere). The DFs experienced by the ambient BBOA are required to establish a connection between the field observations and laboratory results. While this is unknown, results from a microphysics dispersion model suggest that the DFs for fire emissions is related to the fire size and atmospheric stability, as the emitted plumes from smaller fires experience larger DFs more rapidly during atmospheric transport under turbulent meteorological conditions, compared to larger plumes emitted from larger fires during stable conditions.60 For example, simulations conducted by Bian et al.60 indicated that for fire sizes between 1 × 10−4 and 100 km2 (0.01 to 1 × 104 hectares), which represent a small agricultural prescribed burn to large regional wildfire, the corresponding fire plumes experienced DFs of approximately over 1.8 × 105 for small fires versus 1.3 for large fires after 2 h of atmospheric transport under stable meteorological conditions. Under turbulent meteorological conditions, where the fire plume will mix with background air masses more rapidly compared to stable meteorological conditions, the corresponding fire plumes would experience higher DFs after 2 h of atmospheric transport. Considering that the median size of fires in the Mediterranean and continental Eastern Europe regions are typically ∼1 km2 (100 ha),61 for the fires that contributed to the ambient BBOA considered in this study, we estimate the emissions likely experienced DFs (e.g., DFs > 200) that would have led to significant changes in WSOC and OPWS‑DTT within mass several hours of atmospheric transport. Note, DFs increase significantly early in the plume aging and stabilize with time and that the time period during which this rapid increase in DF occurs strongly depends on the stability of the atmosphere.60 This analysis suggests that atmospheric dilution could potentially be the most important aging process that resulted in the observed decrease in WSOC and increase in OPWS‑DTT mass in the ambient samples (Figure 1a,b).

4. ATMOSPHERIC IMPLICATIONS Both laboratory and field results from this work demonstrated that atmospheric aging of BBOA leads to significant changes in OPDTT mass . While the field observations demonstrate for the first time that OPDTT mass for ambient BBOA evolves throughout atmospheric transport, this trend should be explored further. Given that the average atmospheric lifetime of particles in the atmosphere is approximately 1 week with respect to deposition, additional work investigating whether the OPTotal‑DTT and OPWS‑DTT of BBOA continue to evolve beyond mass mass atmospheric transport times of 68 h is necessary. Comparison of field observations to laboratory results indicated that exposures to UV radiation and ozone, along with dilution represent atmospheric aging processes that could have resulted in the observed trend of increasing OPWS‑DTT mass with atmospheric age for ambient BBOA. While this work suggests that the evaporative loss of non- or less-DTT active semivolatile organic compounds likely resulted in the increase in OPWS‑DTT , the relationship between volatility and OPDTT mass mass should be characterized for BBOA. Our laboratory study

■

ASSOCIATED CONTENT

S Supporting Information *

The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acs.est.9b01034. Additional experimental details, Figures S1−S5, and Table S1 (PDF)

■

AUTHOR INFORMATION

Corresponding Author

*E-mail: [email protected]. ORCID

Jenny P.S. Wong: 0000-0002-8729-8166 Notes

The authors declare no competing financial interest. 6753


Article


■

Particulate Matter Supersites Program. Aerosol Sci. Technol. 2004, 38 (sup1), 68−81. (13) Teenhof, M.; Gosens, I.; Strak, M.; Godri, K. J.; Hoek, G.; Cassee, F. R.; Mudway, I. S.; Kelly, F. J.; Harrison, R. M.; Lebret, E.; Brunekreef, B.; Janssen, N. A.; Pieters, R. H. In Vitro Toxicity of Particulate Matter (PM) Collected at Different Sites in the Netherlands Is Associated with PM Composition, Size Fraction and Oxidative Potential–the RAPTES Project. Part. Fibre Toxicol. 2011, 8, 26. (14) Delfino, R. J.; Staimer, N.; Tjoa, T.; Gillen, D. L.; Schauer, J. J.; Shafer, M. M. Airway Inflammation and Oxidative Potential of Air Pollutant Particles in a Pediatric Asthma Panel. J. Exposure Sci. Environ. Epidemiol. 2013, 23 (5), 466−473. (15) Janssen, N. A. H.; Strak, M.; Yang, A.; Hellack, B.; Kelly, F. J.; Kuhlbusch, T. A. J.; Harrison, R. M.; Brunekreef, B.; Cassee, F. R.; Steenhof, M.; Hoek, G. Associations between Three Specific ACellular Measures of the Oxidative Potential of Particulate Matter and Markers of Acute Airway and Nasal Inflammation in Healthy Volunteers. Occup. Environ. Med. 2015, 72 (1), 49−56. (16) Velali, E.; Papachristou, E.; Pantazaki, A.; Choli-Papadopoulou, T.; Planou, S.; Kouras, A.; Manoli, E.; Besis, A.; Voutsa, D.; Samara, C. Redox Activity and in Vitro Bioactivity of the Water-Soluble Fraction of Urban Particulate Matter in Relation to Particle Size and Chemical Composition. Environ. Pollut. 2016, 208 (Pt B), 774−786. (17) Charrier, J. G.; Anastasio, C. On Dithiothreitol (DTT) as a Measure of Oxidative Potential for Ambient Particles: Evidence for the Importance of Soluble Transition Metals. Atmos. Chem. Phys. 2012, 12 (19), 9321−9333. (18) Saffari, A.; Daher, N.; Shafer, M. M.; Schauer, J. J.; Sioutas, C. Seasonal and Spatial Variation in Dithiothreitol (DTT) Activity of Quasi-Ultrafine Particles in the Los Angeles Basin and Its Association with Chemical Species. J. Environ. Sci. Health, Part A: Toxic/Hazard. Subst. Environ. Eng. 2014, 49 (4), 441−451. (19) FFang, T.; Verma, V.; Bates, J. T.; Abrams, J.; Klein, M.; Strickland, M. J.; Sarnat, S. E.; Chang, H. H.; Mulholland, J. A.; Tolbert, P. E.; Russell, A. G.; Weber, R. J. Oxidative Potential of Ambient Water-Soluble PM2.5 in the Southeastern United States: Contrasts in Sources and Health Associations between Ascorbic Acid (AA) and Dithiothreitol (DTT) Assays. Atmos. Chem. Phys. 2016, 16 (6), 3865−3879. (20) Verma, V.; Fang, T.; Guo, H.; King, L.; Bates, J. T.; Peltier, R. E.; Edgerton, E.; Russell, A. G.; Weber, R. J. Reactive Oxygen Species Associated with Water-Soluble PM2.5 in the Southeastern United States: Spatiotemporal Trends and Source Apportionment. Atmos. Chem. Phys. 2014, 14 (23), 12915−12930. (21) Antiñolo, M.; Willis, M. D.; Zhou, S.; Abbatt, J. P. D. Connecting the Oxidation of Soot to Its Redox Cycling Abilities. Nat. Commun. 2015, 6, 6812. (22) SShiraiwa, M.; Ueda, K.; Pozzer, A.; Lammel, G.; Kampf, C. J.; Fushimi, A.; Enami, S.; Arangio, A. M.; Fröhlich-Nowoisky, J.; Fujitani, Y.; Furuyama, A.; Lakey, P. S. J.; Lelieveld, J.; Lucas, K.; Morino, Y.; Poschl, U.; Takahama, S.; Takami, A.; Tong, H.; Weber, B.; Yoshino, A.; Sato, K. Aerosol Health Effects from Molecular to Global Scales. Environ. Sci. Technol. 2017, 51 (23), 13545−13567. (23) Robinson, A. L.; Donahue, N. M.; Shrivastava, M. K.; Weitkamp, E. A.; Sage, A. M.; Grieshop, A. P.; Lane, T. E.; Pierce, J. R.; Pandis, S. N. Rethinking Organic Aerosols: Semivolatile Emissions and Photochemical Aging. Science 2007, 315 (5816), 1259−1262. (24) Jimenez, J. L.; Canagaratna, M. R.; Donahue, N. M.; Prevot, A. S. H.; Zhang, Q.; Kroll, J. H.; DeCarlo, P. F.; Allan, J. D.; Coe, H.; Ng, N. L.; Aiken, A. C.; Docherty, K. S.; Ulbrich, I. M.; Grieshop, A. P.; Robinson, A. L.; Duplissy, J.; Smith, J. D.; Wilson, K. R.; Lanz, V. A.; Hueglin, C.; Sun, Y. L.; Tian, J.; Laaksonen, A.; Raatikainen, T.; Rautiainen, J.; Vaattovaara, P.; Ehn, M.; Kulmala, M.; Tomlinson, J. M.; Collins, D. R.; Cubison, M. J.; Dunlea, J.; Huffman, J. A.; Onasch, T. B.; Alfarra, M. R.; Williams, P. I.; Bower, K.; Kondo, Y.; Schneider, J.; Drewnick, F.; Borrmann, S.; Weimer, S.; Demerjian, K.; Salcedo, D.; Cottrell, L.; Griffin, R.; Takami, A.; Miyoshi, T.; Hatakeyama, S.;

ACKNOWLEDGMENTS Funding for this work was provided by the Electric Power Research Institute (EPRI) through contract no. 00-10003806, and from the project PyroTRACH (ERC-2016-COG) funded from H2020-EU.1.1. - Excellent Science - European Research Council (ERC), project ID 726165. Additional support was also provided by NASA through contracts NNX14A974G and NNX16AE19G.

■

REFERENCES

(1) Dockery, D. W.; Pope, I.; Xu, X.; Spengler, J. D.; Ware, J. H.; Fay, M. E.; Ferris, J.; Speizer, F. E. An Association between Air Pollution and Mortality in Six U.S. Cities. N. Engl. J. Med. 1993, 329 (24), 1753−1759. (2) Samet, J. M.; Dominici, F.; Curriero, F. C.; Coursac, I.; Zeger, S. L. Fine Particulate Air Pollution and Mortality in 20 U.S. Cities, 1987−1994. N. Engl. J. Med. 2000, 343 (24), 1742−1749. (3) Pope III, C.; Burnett, R.; Thun, M. J.; Calle, E. E.; Aljawhary, D.; Ito, K.; Thurston, G. D. Lung Cancer, Cardiopulmonary Mortality, and Long-Term Exposure to Fine Particulate Air Pollution. JAMA 2002, 287 (9), 1132−1141. (4) Brunekreef, B.; Holgate, S. T. Air Pollution and Health. Lancet 2002, 360 (9341), 1233−1242. (5) Gakidou, E.; Afshin, A.; Abajobir, A. A.; Abate, K. H.; Abbafati, C.; Abbas, K. M.; Abd-Allah, F.; Abdulle, A. M.; Abera, S. F.; Aboyans, V.; Abu-Raddad, L. J.; Abu-Rmeileh, N. M. E.; Abyu, G. Y.; Adedeji, I. A.; Adetokunboh, O.; Afarideh, M.; Agrawal, A.; Agrawal, S.; Ahmadieh, H.; Ahmed, M. B.; Aichour, M. T. E.; Aichour, A. N.; Aichour, I.; Akinyemi, R. O.; Akseer, N.; Alahdab, F.; Al-Aly, Z. Global, Regional, and National Comparative Risk Assessment of 84 Behavioural, Environmental and Occupational, and Metabolic Risks or Clusters of Risks, 1990−2016: A Systematic Analysis for the Global Burden of Disease Study 2016. Lancet 2017, 390 (10100), 1345− 1422. (6) Tao, F.; Gonzalez-Flecha, B.; Kobzik, L. Reactive Oxygen Species in Pulmonary Inflammation by Ambient Particulates. Free Radical Biol. Med. 2003, 35 (4), 327−340. (7) Kuenzli, N.; Shi, T.; Goetschi, T.; Kelly, F.; Mudway, I.; Burney, P.; Forsberg, B.; Heinrich, J.; Jarvis, D.; Soon, A.; Lucynska, C.; PayoLosa, F.; Poli, A.; Weyler, J.; Hazenkamp, M.; Norback, D.; Borm, P. Beyond the Mass: Oxidative Properties of PM2.5 in the European Community Respiratory Health Survery (ECRHS). Epidemiology 2004, 15 (4), S43. (8) Bates, J. T.; Weber, R. J.; Abrams, J.; Verma, V.; Fang, T.; Klein, M.; Strickland, M. J.; Sarnat, S. E.; Chang, H. H.; Mulholland, J. A.; Rolbert, P. E.; Russell, A. G. Reactive Oxygen Species Generation Linked to Sources of Atmospheric Particulate Matter and Cardiorespiratory Effects. Environ. Sci. Technol. 2015, 49 (22), 13605−13612. (9) Yang, A.; Janssen, N. A. H.; Brunekreef, B.; Cassee, F. R.; Hoek, G.; Gehring, U. Children’s Respiratory Health and Oxidative Potential of PM2.5: The PIAMA Birth Cohort Study. Occup. Environ. Med. 2016, 73 (3), 154−160. (10) Abrams, J. Y.; Weber, R. J.; Klein, M.; Samat, S. E.; Chang, H. H.; Strickland, M. J.; Verma, V.; Fang, T.; Bates, J. T.; Mulholland, J. A.; Russell, A. G.; Tolbert, P. E. Associations between Ambient Fine Particulate Oxidative Potential and Cardiorespiratory Emergency Department Visits. Environ. Health Perspect. 2017, 125 (10), 107008. (11) Kumagai, Y.; Koide, S.; Taguchi, K.; Endo, A.; Nakai, Y.; Yoshikawa, T.; Shimojo, N. Oxidation of Proximal Protein Sulfhydryls by Phenanthraquinone, a Component of Diesel Exhaust Particles. Chem. Res. Toxicol. 2002, 15 (4), 483−489. (12) Cho, A. K.; Stefano, E. D.; You, Y.; Rodriguez, C. E.; Schmitz, D. A.; Kumagai, Y.; Miguel, A. H.; Eiguren-Fernandez, A.; Kobayashi, T.; Avol, E.; Froines, J. R. Determination of Four Quinones in Diesel Exhaust Particles, SRM 1649a, and Atmospheric PM2.5 Special Issue of Aerosol Science and Technology on Findings from the Fine 6754


Article

Environmental Science & Technology Shimono, A.; Sun, J. Y.; Zhang, Y. M.; Dzepina, K.; Kimmel, J. R.; Sueper, D.; Jayne, J. T.; Herndon, S. C.; Trimborn, A. M.; Williams, L. R.; Wood, E. C.; Middlebrook, A. M.; Kolb, C. E.; Baltensperger, U.; Worsnop, D. R. Evolution of Organic Aerosols in the Atmosphere. Science 2009, 326 (5959), 1525−1529. (25) Verma, V.; Ning, Z.; Cho, A. K.; Schauer, J. J.; Shafer, M. M.; Sioutas, C. Redox Activity of Urban Quasi-Ultrafine Particles from Primary and Secondary Sources. Atmos. Environ. 2009, 43 (40), 6360−6368. (26) Li, Q.; Wyatt, A.; Kamens, R. M. Oxidant Generation and Toxicity Enhancement of Aged-Diesel Exhaust. Atmos. Environ. 2009, 43 (5), 1037−1042. (27) McWhinney, R. D.; Gao, S. S.; Zhou, S.; Abbatt, J. P. D. Evaluation of the Effects of Ozone Oxidation on Redox-Cycling Activity of Two-Stroke Engine Exhaust Particles. Environ. Sci. Technol. 2011, 45 (6), 2131−2136. (28) Rattanavaraha, W.; Rosen, E.; Zhang, H.; Li, Q.; Pantong, K.; Kamens, R. M. The Reactive Oxidant Potential of Different Types of Aged Atmospheric Particles: An Outdoor Chamber Study. Atmos. Environ. 2011, 45 (23), 3848−3855. (29) Tuet, W. Y.; Chen, Y.; Fok, S.; Gao, D.; Weber, R. J.; Champion, J. A.; Ng, N. L. Chemical and Cellular Oxidant Production Induced by Naphthalene Secondary Organic Aerosol (SOA): Effect of Redox-Active Metals and Photochemical Aging. Sci. Rep. 2017, 7 (1), 15157. (30) Chowdhury, P. H.; He, Q.; Lasitza Male, T.; Brune, W. H.; Rudich, Y.; Pardo, M. Exposure of lung epithelial cells to photochemically aged secondary organic aerosol shows increased toxic effects. Environ. Sci. Technol. Lett. 2018. 5424. (31) Jiang, H.; Jang, M. Dynamic Oxidative Potential of Atmospheric Organic Aerosol under Ambient Sunlight. Environ. Sci. Technol. 2018, 52 (13), 7496−7504. (32) Wang, S.; Ye, J.; Soong, R.; Wu, B.; Yu, L.; Simpson, A. J.; Chan, A. W. H. Relationship between Chemical Composition and Oxidative Potential of Secondary Organic Aerosol from Polycyclic Aromatic Hydrocarbons. Atmos. Chem. Phys. 2018, 18 (6), 3987− 4003. (33) Zhou, J.; Zotter, P.; Bruns, E. A.; Stefenelli, G.; Bhattu, D.; Brown, S.; Bertrand, A.; Marchand, N.; Lamkaddam, H.; Slowik, J. G.; Prevot, A. S. H.; Baltensperger, U.; Nussbaumer, T.; El-Haddad, I.; Dommen, J. Particle-Bound Reactive Oxygen Species (PB-ROS) Emissions and Formation Pathways in Residential Wood Smoke under Different Combustion and Aging Conditions. Atmos. Chem. Phys. 2018, 18 (10), 6985−7000. (34) Decesari, S.; Sowlat, M. H.; Hasheminassab, S.; Sandrini, S.; Gilardoni, S.; Facchini, M. C.; Fuzzi, S.; Sioutas, C. Enhanced Toxicity of Aerosol in Fog Conditions in the Po Valley, Italy. Atmos. Chem. Phys. 2017, 17 (12), 7721−7731. (35) Fang, T.; Guo, H.; Zeng, L.; Verma, V.; Nenes, A.; Weber, R. J. Highly Acidic Ambient Particles, Soluble Metals, and Oxidative Potential: A Link between Sulfate and Aerosol Toxicity. Environ. Sci. Technol. 2017, 51 (5), 2611−2620. (36) Paris, R.; Desboeufs, K. V. Effect of Atmospheric Organic Complexation on Iron-Bearing Dust Solubility. Atmos. Chem. Phys. 2013, 13 (9), 4895−4905. (37) Tsimpidi, A. P.; Karydis, V. A.; Pandis, S. N.; Lelieveld, J. Global Combustion Sources of Organic Aerosols: Model Comparison with 84 AMS Factor-Analysis Data Sets. Atmos. Chem. Phys. 2016, 16 (14), 8939−8962. (38) Verma, V.; Fang, T.; Xu, L.; Peltier, R. E.; Russell, A. G.; Ng, N. L.; Weber, R. J. Organic Aerosols Associated with the Generation of Reactive Oxygen Species (ROS) by Water-Soluble PM2.5. Environ. Sci. Technol. 2015, 49 (7), 4646−4656. (39) Wong, J. P. S.; Nenes, A.; Weber, R. J. Changes in Light Absorptivity of Molecular Weight Separated Brown Carbon Due to Photolytic Aging. Environ. Sci. Technol. 2017, 51 (15), 8414−8421. (40) Wong, J. P. S.; Tsagaraki, M.; Tsiodra, I.; Mihalopoulos, N.; Violaki, K.; Kanakidou, M.; Nenes, A.; Weber, R. J. Atmospheric

evolution of molecular weight separated brown carbon from biomass burning. Atmos. Chem. Phys. Discuss. 2018.1, accepted, (41) Zhou, X.; Mopper, K. Determination of Photochemically Produced Hydroxyl Radicals in Seawater and Freshwater. Mar. Chem. 1990, 30, 71−88. (42) Anastasio, C.; McGregor, K. G. Chemistry of Fog Waters in California’s Central Valley: 1. In Situ Photoformation of Hydroxyl Radical and Singlet Molecular Oxygen. Atmos. Environ. 2001, 35 (6), 1079−1089. (43) Fine, P. M.; Cass, G. R.; Simoneit, B. R. T. Chemical Characterization of Fine Particle Emissions from Fireplace Combustion of Woods Grown in the Northeastern United States. Environ. Sci. Technol. 2001, 35 (13), 2665−2675. (44) Fine, P. M.; Cass, G. R.; Simoneit, B. R. T. Chemical Characterization of Fine Particle Emissions from the Fireplace Combustion of Woods Grown in the Southern United States. Environ. Sci. Technol. 2002, 36 (7), 1442−1451. (45) Lipsky, E. M.; Robinson, A. L. Effects of Dilution on Fine Particle Mass and Partitioning of Semivolatile Organics in Diesel Exhaust and Wood Smoke. Environ. Sci. Technol. 2006, 40 (1), 155− 162. (46) Fang, T.; Verma, V.; Guo, H.; King, L. E.; Edgerton, E. S.; Weber, R. J. A Semi-Automated System for Quantifying the Oxidative Potential of Ambient Particles in Aqueous Extracts Using the Dithiothreitol (DTT) Assay: Results from the Southeastern Center for Air Pollution and Epidemiology (SCAPE). Atmos. Meas. Tech. 2015, 8 (1), 471−482. (47) Gao, D.; Fang, T.; Verma, V.; Zeng, L.; Weber, R. J. A Method for Measuring Total Aerosol Oxidative Potential (OP) with the Dithiothreitol (DTT) Assay and Comparisons between an Urban and Roadside Site of Water-Soluble and Total OP. Atmos. Meas. Tech. 2017, 10 (8), 2821−2835. (48) Fang, T.; Guo, H.; Verma, V.; Peltier, R. E.; Weber, R. J. PM2.5 Water-Soluble Elements in the Southeastern United States: Automated Analytical Method Development, Spatiotemporal Distributions, Source Apportionment, and Implications for Heath Studies. Atmos. Chem. Phys. 2015, 15 (20), 11667−11682. (49) Yu, H.; Wei, J.; Cheng, Y.; Subedi, K.; Verma, V. Synergistic and Antagonistic Interactions among the Particulate Matter Components in Generating Reactive Oxygen Species Based on the Dithiothreitol Assay. Environ. Sci. Technol. 2018, 52 (4), 2261−2270. (50) Fushimi, A.; Saitoh, K.; Hayashi, K.; Ono, K.; Fujitani, Y.; Villalobos, A. M.; Shelton, B. R.; Takami, A.; Tanabe, K.; Schauer, J. J. Chemical Characterization and Oxidative Potential of Particles Emitted from Open Burning of Cereal Straws and Rice Husk under Flaming and Smoldering Conditions. Atmos. Environ. 2017, 163, 118−127. (51) Turpin, B. J.; Lim, H.-J. Species Contributions to PM2.5 Mass Concentrations: Revisiting Common Assumptions for Estimating Organic Mass. Aerosol Sci. Technol. 2001, 35 (1), 602−610. (52) Bougiatioti, A.; Bezantakos, S.; Stavroulas, I.; Kalivitis, N.; Kokkalis, P.; Biskos, G.; Mihalopoulos, N.; Papayannis, A.; Nenes, A. Biomass-Burning Impact on CCN Number, Hygroscopicity and Cloud Formation during Summertime in the Eastern Mediterranean. Atmos. Chem. Phys. 2016, 16 (11), 7389−7409. (53) Arakaki, T.; Anastasio, C.; Kuroki, Y.; Nakajima, H.; Okada, K.; Kotani, Y.; Handa, D.; Azechi, S.; Kimura, T.; Tsuhako, A.; Miyagi, Y. A General Scavenging Rate Constant for Reaction of Hydroxyl Radical with Organic Carbon in Atmospheric Waters. Environ. Sci. Technol. 2013, 47 (15), 8196−8203. (54) Poulain, L.; Iinuma, Y.; Müller, K.; Birmili, W.; Weinhold, K.; Brü g gemann, E.; Gnauk, T.; Hausmann, A.; Lö s chau, G.; Wiedensohler, A.; Herrmann, H. Diurnal Variations of Ambient Particulate Wood Burning Emissions and Their Contribution to the Concentration of Polycyclic Aromatic Hydrocarbons (PAHs) in Seiffen, Germany. Atmos. Chem. Phys. 2011, 11 (24), 12697−12713. (55) Lin, Y.; Ma, Y.; Qiu, X.; Li, R.; Fang, Y.; Wang, J.; Zhu, Y.; Hu, D. Sources, Transformation, and Health Implications of PAHs and 6755


Article

Environmental Science & Technology Their Nitrated, Hydroxylated, and Oxygenated Derivatives in PM2.5 in Beijing. J. Geophys. Res., Atmos. 2015120 (14), 7219−7228. . (56) Cho, A. K.; Sioutas, C.; Miguel, A. H.; Kumagai, Y.; Schmitz, D. A.; Singh, M.; Eiguren-Fernandez, A.; Froines, J. R. Redox Activity of Airborne Particulate Matter at Different Sites in the Los Angeles Basin. Environ. Res. 2005, 99 (1), 40−47. (57) Mmereki, B. T.; Donaldson, D. J.; Gilman, J. B.; Eliason, T. L.; Vaida, V. Kinetics and Products of the Reaction of Gas-Phase Ozone with Anthracene Adsorbed at the Air-aqueous Interface. Atmos. Environ. 2004, 38 (36), 6091−6103. (58) Kwamena, N.-O. A.; Earp, M. E.; Young, C. J.; Abbatt, J. P. D. Kinetic and Product Yield Study of the Heterogeneous Gas-Surface Reaction of Anthracene and Ozone. J. Phys. Chem. A 2006, 110 (10), 3638−3646. (59) Kopanakis, I.; Glytsos, T.; Kouvarakis, G.; Gerasopoulos, E.; Mihalopoulos, N.; Lazaridis, M. Variability of Ozone in the Eastern Mediterranean during a 7-Year Study. Air Qual., Atmos. Health 2016, 9 (5), 461−470. (60) Bian, Q.; Jathar, S. H.; Kodros, J. K.; Barsanti, K. C.; Hatch, L. E.; May, A. A.; Kreidenweis, S. M.; Pierce, J. R. Secondary Organic Aerosol Formation in Biomass-Burning Plumes: Theoretical Analysis of Lab Studies and Ambient Plumes. Atmos. Chem. Phys. 2017, 17 (8), 5459−5475. (61) Hernandez, C.; Drobinski, P.; Turquety, S.; Dupuy, J.-L. Size of Wildfires in the Euro-Mediterranean Region: Observations and Theoretical Analysis. Nat. Hazards Earth Syst. Sci. 2015, 15 (6), 1331−1341. (62) Ortega, A. M.; Day, D. A.; Cubison, M. J.; Brune, W. H.; Bon, D.; Gouw, D.; A, J.; Jimenez, J. L. Secondary Organic Aerosol Formation and Primary Organic Aerosol Oxidation from BiomassBurning Smoke in a Flow Reactor during FLAME-3. Atmos. Chem. Phys. 2013, 13 (22), 11551−11571. (63) Samburova, V.; Connolly, J.; Gyawali, M.; Yatavelli, R. L. N.; Watts, A. C.; Chakrabarty, R. K.; Zielinska, B.; Moosmüller, H.; Khlystov, A. Polycyclic Aromatic Hydrocarbons in Biomass-Burning Emissions and Their Contribution to Light Absorption and Aerosol Toxicity. Sci. Total Environ. 2016, 568, 391−401. (64) Tuet, W. Y.; Chen, Y.; Xu, L.; Fok, S.; Gao, D.; Weber, R. J.; Ng, N. L. Chemical Oxidative Potential of Secondary Organic Aerosol (SOA) Generated from the Photooxidation of Biogenic and Anthropogenic Volatile Organic Compounds. Atmos. Chem. Phys. 2017, 17 (2), 839−853.

6756


Effects of Atmospheric Processing on the Oxidative Potential of

Recommend Documents