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Organic Composition, Chemistry, and Photochemistry of Urban Film in Leipzig, Germany Sarah Anne Styler, Alyson M. Baergen, D. James Donaldson, and Hartmut Herrmann ACS Earth Space Chem., Just Accepted Manuscript • DOI: 10.1021/ acsearthspacechem.8b00087 • Publication Date (Web): 25 Jul 2018 Downloaded from http://pubs.acs.org on July 28, 2018

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ACS Earth and Space Chemistry

Organic Composition, Chemistry, and Photochemistry of Urban Film in Leipzig, Germany Sarah A. Styler,a,b Alyson M. Baergen,c D. J. Donaldson,c,d and Hartmut Herrmanna a

Leibniz-Institut für Troposphärenforschung (TROPOS), Atmospheric Chemistry Department (ACD), Permoserstraße 15, D-04318 Leipzig, Germany b Now at Department of Chemistry, University of Alberta, Edmonton, Alberta, Canada T6G 2G2 c Department of Chemistry, University of Toronto, Toronto, Ontario, Canada d Department of Physical and Environmental Sciences, University of Toronto at Scarborough, Toronto, Ontario, Canada Corresponding author email:

[email protected]

Keywords:

urban grime urban surface chemistry urban film growth heterogeneous photochemistry polycyclic aromatic hydrocarbons (PAH) n-alkanes particulate matter (PM10)

Abstract In polluted urban environments, windows and building surfaces are coated with a complex film of chemicals. Despite its high surface-to-volume ratio and direct exposure to sunlight, few studies have directly investigated the role that this “urban film” may play in promoting the chemistry and photochemistry of semi-volatile organic species contained within it. Here, we report results from a field investigation of the organic composition of urban film and particulate matter (PM10) samples collected at an urban site in Leipzig, Germany, in which we provide clear evidence for the influence of anthropogenic processes on film composition. In this study, we find that the ratio of watersoluble organic carbon (WSOC) to the total ionic content of film samples decreases with atmospheric exposure time, which suggests that urban film growth proceeds first via the condensation of semi-volatile species, and that the coating thus formed enhances the dry deposition of particles. Further, we find that the polycyclic aromatic hydrocarbon (PAH) abundance profiles in light-exposed films are different from those in films collected under light-shielded conditions, which represents the first direct evidence that urban films serve as a photochemical sink for semivolatile organic pollutants. Finally, we find that the PAH and alkane profiles of urban film samples differ substantially from co-located PM10 samples, which reflects both the contribution of settled coarse particulate matter to the overall film composition and the influence of in-film oxidative processes. Together, these results highlight the unique reactive environment afforded by urban film and underscore the need for further studies of urban surface chemistry.

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Introduction

In urban areas, windows and other impermeable surfaces are coated with a complex chemical mixture, which includes crustal material; trace metals; sulfate, nitrate, and other inorganic ions; alkanes, alkenes, and elemental carbon; and polar organics, including esters, alcohols, and carboxylic acids.1–7 This coating, often referred to as ‘urban film’ or ‘urban grime’, also contains trace quantities of a number of toxic species, including polychlorinated biphenyls (PCBs), polybrominated diphenyl ethers (PBDEs), perfluoroalkyl contaminants (PFCs), phthalate diesters (PAEs), and polycyclic aromatic hydrocarbons (PAHs).8–12 Field measurements suggest that urban film, once formed, enhances the particle capture efficiency of the surface and serves as a reservoir into which semivolatile species can partition.4,13–15 However, our understanding of the mechanism of film formation is currently limited by the fact that studies of film composition have typically employed single samples collected from surfaces with unknown histories: for example, although many studies have investigated PAH concentrations in window films,4,10,13,16–21 the vast majority of these films have been collected from windows where the date of last cleaning was either unknown or poorly defined. Field investigations of film development have largely employed time-resolved measurements of bulk physical parameters, including accumulated mass, particle coverage, and surface optical properties, to better understand the influence of pollutant deposition on the aesthetic and structural properties of urban surfaces.8,22–29 To date, only a handful of studies have measured the chemical composition of surface films as a function of atmospheric exposure time.5,7–9,17,30 In two of these studies, accumulation rates of elemental and organic carbon9 and trace organics9,17 in window films were used to draw conclusions regarding the relative contributions of particle deposition and condensation of semivolatile compounds from the gas phase to film formation and growth. These studies suggest that time-resolved measurements of film chemical composition have the potential to contribute significantly to our understanding of film formation mechanisms. As a result of its large surface area to volume ratio (~ 1.4 × 107) and direct exposure to sunlight, urban film has been suggested to serve as a medium for photochemical reactions.31 Currently, however, our knowledge regarding photochemistry occurring within real urban film samples is

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limited to the results of three studies, all of which focused on nitrate photochemistry and its influence on urban NOx (NO + NO2) budgets.30,32,33 Although some indirect field evidence exists for the photochemical loss of organic contaminants in urban film,9,15,17 a comprehensive assessment of the significance of film-phase processing of photoactive organics is still missing. Since, as discussed above, urban film is formed not only by particle deposition but also by condensation of semivolatile compounds from the gas phase,13–15 and since species contained within it are potentially exposed to light and atmospheric oxidants for considerable periods of time, its composition might be expected to differ from that of urban aerosol. Although a number of field studies have indeed identified differences in the inorganic and organic composition of urban film as compared to typical urban aerosol,1,5,10,13 few studies to date have directly compared the chemical composition of simultaneously collected film and aerosol samples.6,30 In the present study, which serves as a companion to Baergen et al.,30 urban film samples were collected under controlled conditions in a high-traffic urban location in Leipzig, Germany. In order to improve our understanding of the processes that govern film formation and growth, concentrations of both bulk and trace organic film components were monitored as a function of atmospheric exposure time. The influence of photochemistry on the lifetime of pollutants present within the film was explored by comparing the relative abundances of PAHs in samples collected under dark and light conditions. Finally, the organic composition of film samples was compared to that of PM10 and size-resolved particle samples collected simultaneously at the same location. Together, these results provide strong evidence that urban film is a dynamic, photochemically active, and compositionally unique environment, and highlight the need for further studies of urban surface chemistry.

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Materials and Methods

2.1

Sampling methodology

Urban film and particle sampling was conducted from September 16 to October 25, 2014 at the “Leipzig-Mitte” (51.33°N, 12.38°E) air quality monitoring station, which is located in a high-traffic location opposite the main train station and operated by the Saxon State Agency for Environment,

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Agriculture and Geology (LfULG).34 A summary of meteorological and air quality data for the sampling period is presented in Figure S1. 2.1.1 Film collection Urban film samples were collected using a newly developed custom-built three-stage sampler, a photograph of which is shown in Figure S2. The roof of the sampler consists of a polymethylmethacrylate (PMMA) plate (PLEXIGLAS® GS, UV transmitting, Clear 2458; EBLAGmbH) which shields samples from precipitation while still allowing for illumination of the uppermost stage of the sampler. The two lower stages are shielded from light by a stainless-steel sheet placed beneath the first stage. Stainless-steel slats around the outside of each stage further shield samples from precipitation while still allowing for airflow through the sampler. Each stage of the sampler consists of a 16-compartment stainless steel mesh tray. In these experiments, only the top two stages of the sampler were used for sample collection. Soda-lime glass beads (d = 3 mm; Sigma Aldrich) were used as surrogate window surfaces.8,24,30,35 Before deployment, beads were immersed in a base bath for 1 h to remove residual organics, rinsed 10 times with tap water, soaked overnight in deionized (Milli-Q; 18 MΩ) water, rinsed 8 times with 18 MΩ water, and dried overnight at 373K. On the first day of the campaign, bead samples (~ 80 g) were placed in a single layer in 15 of the 16 compartments on each of the two top stages of the sampler. Laboratory blanks (n = 3), consisting of clean bead samples that were not transported to the field site, and field blanks (n = 3), consisting of clean bead samples that were placed onto the sampler on the first day of the campaign and immediately collected, were also analyzed. Beads were transported to and from the sampling site in amber bottles and were refrigerated until extraction and analysis. All reported analyte concentrations are blank-corrected by subtraction; since field and laboratory blanks were similar for all analytes, the average of the six blank values was used in these corrections. Bead samples were collected from a single compartment on the light and dark stages every 3 days (72 h) at 11:00 local time (GMT+2). The choice of sample compartments was determined using a random number generator. In order to examine the extent to which our results were influenced by variable airflow within the sampler, samples were taken from three different compartments on

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October 7, 2014 (Day 21). As shown in Figure 1, the variability in measured concentrations varied with analyte, and ranged from 7% for WSOC to a maximum of 25% for Σalkanes. Figure 1 also shows that there was no systematic difference between the accumulation of unreactive organics (i.e. alkanes) on the middle (dark) and upper (light-exposed) stages, which provides further evidence for the relative uniformity of airflow within the sampler. As a result of the 72 h sampling collection protocol, some urban film growth periods included weekend days, whereas others did not. However, we did not see conclusive evidence for a relationship between film growth over each sampling period and the number of weekday/weekend sampling days in the same period. Indeed, air quality at the sampling site during the study, as represented by NOx (Figure S1) and PM10 levels (data not shown), did not display a weekday/weekend pattern. In addition, a lack of weekday/weekend pattern has previously been observed for traffic PM10 at the same sampling site.36 Since the melting point of the glass beads employed in this study is below the temperatures employed in organic and elemental carbon (OC–EC) analysis, film samples for bulk organic analysis were collected on pairs of quartz–fibre filters (n = 6) placed on the top two sampler stages. The surface presented by this substrate is different from that presented by the glass beads; a discussion of these differences, and their implications, is presented in Section 3.3. 2.1.2 Particle collection For the first 33 days of the sampling campaign (September 16–October 19), 72 h particle samples (n = 11) were collected using a low-volume particle sampler (Partisol 2000, Rupprecht and Patashnick; 1 m3 h-1 sampling rate) equipped with a PM10 inlet. Samples were collected on quartzfibre filters (d = 45 mm; Munktell MK 360) that had been previously heated for 24 h at 378 K to minimize blank organic carbon values. Each sampling period began and ended at 11:00 (GMT+2). Filter samples were transported to the laboratory in clean, covered Petri dishes, divided (see Section 2.2), and stored at 253 K until analysis. All reported values are corrected by subtraction for analyte concentrations in laboratory blanks (n = 3), which consisted of clean quartz-fibre filters that were not transported to the sampling site.

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During the last five days of the campaign (October 21–25), size-resolved particle samples were collected using a five-stage (0.05–0.14 μm, 0.14–0.42 μm, 0.42–1.2 μm, 1.2–3.5 μm, 3.5–10 μm aerodynamic particle diameter) stainless-steel Berner cascade impactor (Hauke; 4.5 m3 h-1 sampling rate). Preheated (2 h at 623 K) aluminum foils were used as sampling substrates. After sampling, the entire Berner impactor was returned to the laboratory, where the aluminum foils were divided (see Section 2.2) and stored at 253 K until analysis. All reported values are corrected by subtraction for analyte concentrations in a field blank (n = 1), which was obtained by placing the Berner impactor, equipped with aluminum sampling substrates, in the sampler without turning on the sampling pump. 2.2

Sample analysis

PM10 filter samples were divided as follows: a 1 cm × 1 cm square was used for analysis of organic and elemental carbon (OC–EC), two 6 mm diameter filter punches were reserved for Curie-point pyrolysis (CPP-GC-MS) analysis of PAH and alkanes, and the remainder was used for analysis of water-soluble ions30 and water-soluble organic carbon (WSOC). Quartz filters used as substrates for urban film collection were divided as follows: two 1 cm × 1 cm squares were used for OC–EC analysis, and the remainder was used for analysis of water-soluble ions and WSOC. Glass bead samples were divided as follows: 4 g of beads were used for each of analysis of water-soluble ions30 and WSOC, and the remainder was used for CPP-GC-MS analysis of PAH and alkanes. 2.2.1 Organic and elemental carbon (OC–EC) The OC–EC content of quartz fiber filters used for active PM10 and passive urban film sampling was determined via thermal-optical analysis (Dual-Optical Carbonaceous Analyzer, Sunset Laboratory) using the EUSAAR 2 thermal and charring correction (transmission mode) protocol.37 Filters used for passive film collection were analyzed in duplicate or triplicate. Since the Al foils employed as Berner impactor substrates are not suitable for thermal-optical analysis,38 the OC–EC content of size-resolved particle samples was determined using a commercial carbon analyzer (C/S-Max, Seifert Instruments). A two-step thermographic method was employed: first, OC was vaporized in a N2 atmosphere at 923 K and detected as CO2 after catalytic conversion;

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then, the remaining carbon (i.e. EC) was combusted in an O2 atmosphere and detected as CO2. Quantitative determination of CO2 was performed using a non-dispersive infrared detector. 2.2.2 Water-soluble organic carbon (WSOC) PM10 filter samples and quartz filters used for urban film collection were placed in 2-mL vials with 1.5 mL deionized water; Al impactor substrates were placed in 2.0 mL deionized water. The vials were placed in a test-tube rack in an ultrasonic bath for 10 minutes, shaken, and returned to the bath for an additional 10 minutes. The resultant extracts were filtered (0.45 µm Acrodisc, Pall Corporation). After removal of aliquots for ion chromatographic analysis (400 µL)30 and analysis of dicarboxylic acids (100 µL), the remainder was diluted to 5 mL for analysis of water-soluble organic carbon (WSOC). Glass bead samples were subjected to sonication (15 min) in 7 mL deionized water and filtered as above prior to analysis. In all cases, WSOC was determined using a TOC-VCPH analyzer (Shimadzu, Japan) in the non-purgeable organic carbon (NPOC) mode.39 2.2.3 Polycyclic aromatic hydrocarbons (PAHs) and alkanes The PAH and alkane content of PM10 filter samples and Berner impactor substrates were determined directly (i.e. without extraction). In brief, filter punches and Al foils were spiked with a standard solution containing 2.5 ng of three deuterated PAH and two deuterated alkanes, wrapped in a Pyrofoil with a Curie-point temperature of 783 K, and analyzed using Curie-point pyrolysis GCMS (CPP-GC-MS; 6890N GC/5973 MSD, Agilent Technologies, Waldbronn, Germany).40 Samples were quantified against single-point calibration standards containing 2.5 ng of the deuterated standards described above. In this study, we quantified the C21–C32 n-alkanes; the PAH/PAH pairs phenanthrene (PHE), anthracene (ANT), fluoranthene (FLU), pyrene (PYR), benzo[a]anthracene (BaA), chrysene + triphenylene (CHR + TRI), 2,2’-binaphthalene (BIN), benzo[b + k]fluoranthene (BbF + BkF), benzo[a]pyrene (BaP), and benzo[ghi]perylene (BghiP); and benzo[b]naphtho[1,2d]thiophene (BNT). In urban film samples, we also quantified indeno[1,2,3-cd]pyrene (IND). The PAH and alkane content of urban film samples was also determined using CPP-GC-MS. Here, glass bead samples (65–70 g) were twice extracted by shaking in 24 mL dichloromethane (Pestinorm, VWR) for 3 min.8 The combined extracts were filtered (0.2 µm PTFE Acrodisc, Pall Corporation), 10 µL of dodecane (analytical standard grade, Sigma Aldrich) was added as keeper

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solvent,41 and the dichloromethane was evaporated under a nitrogen flow. The resultant dodecane extract (0.5 µL) was placed on a Pyrofoil, which was subsequently spiked with 1 ng of the deuterated standard mix described above. PAH were quantified against a five-point external calibration curve prepared from standards containing 0.02–0.5 ng PAH, 1 ng of the deuterated standard mix described above, and 16 ng of the alkane standard described above in 0.5 µL dodecane. Alkane concentrations were determined using single-point calibration (i.e. against the 16 ng alkane calibration standard).

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Results and discussion

3.1

Urban film growth and organic composition

A number of previous studies have reported linear growth of both bulk urban film parameters (e.g. film mass, EC, OC, and inorganic ions)5,8,12,30 and trace organic constituents of film8 as a function of atmospheric exposure time. Other studies, by contrast, have reported saturation effects in the accumulation of trace and bulk organics,9,12 surface particulate,42 and overall film mass.35 In the present study, we did not observe any obvious saturation/plateau in any of the measured components of urban film: as shown in Figure 1, the concentrations of ΣPAH, Σalkanes, and WSOC in both dark and light-exposed films increased as a function of time over the duration of the campaign. Although we note that the data presented in Figure 1 seems to suggest that the accumulation of PAH/alkanes is slowing during the last week of the sampling campaign, assessment of whether a plateau in the concentration of these species would ultimately be reached would require longer-term sampling data. Finally, the bulk organic content (OC–EC) of co-located quartzfiber filters also increased with sample exposure time, although the scatter in these data was more significant (Figure S3). As illustrated in Figure 1, the WSOC content of films collected under both light and dark conditions exhibited an initial, rapid increase followed by a slower, linear increase with sample exposure time. In addition, as shown in Figure 2, the ratio of WSOC to the total ionic mass of the film (the latter reported in Baergen et al.30) decreased substantially over the first 9 days of film development, after which time it remained relatively constant. Together, these results provide support for previous suggestions that urban film first develops via the condensation of semivolatile organic species and that the resultant organic coating enhances the particle capture efficiency of the surface.4,9,9,13

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In the only study to date of PAH accumulation in urban film, Pan et al. reported that whereas threering PAH compounds exhibited fast initial growth followed by a plateau, five-ring PAH compound concentrations increased linearly during the early stages of film growth.17 In this study, by contrast, we found that the concentrations of both light and heavy PAH increased with sample exposure time for the duration of the sampling campaign (Figure S4). In addition, as shown in Figure 2, the relative abundances of both phenanthrene (PHE), a representative semivolatile PAH, and benzo[a]pyrene (BaP), a representative particle-associated PAH, did not display strong trends over the duration of the sampling campaign. Since the concentrations of several PAH were below our limit of quantitation for the first two timepoints (i.e. 3 and 6 days; these quantification limitations are also the reason for the apparent hysteresis in PAH/alkane accumulation during the early stages of the field campaign), we cannot exclude preferential accumulation of semivolatile PAH via condensation during the first days of film development. However, these results suggest that the first, condensational, stage of film development is effectively complete within ~ 1 week of atmospheric exposure. At the end of the sampling period, ΣPAH and Σalkane concentrations within the film were ~ 1800 ng m-2 and ~ 100000 ng m-2, respectively. As shown in Table S1, whereas ΣPAH levels measured in the present study are comparable to those reported in previous investigations of urban film composition, Σalkane concentrations are significantly higher. Insight into this result is provided by the size-segregated particle samples collected in the final week of the campaign: as illustrated in Figure S6, whereas ΣPAH levels are highest in submicron particles (Berner impactor stages 1–3), a substantial portion of Σalkanes is present in the largest impactor size fraction (3.5–10 μm). In this context, the elevated Σalkane levels observed in the film likely reflect contributions from settled coarse particulate matter, which would be expected to be promoted under our sampling conditions (i.e. rain-sheltered beads with both horizontal and vertical collection surfaces22 as opposed to vertical window surfaces13), and which we visually observed during the sample extraction and filtration process. The influence of coarse particulate matter on the composition of the urban film samples collected in this study is also discussed in Section 3.3. 3.2

Evidence for film-phase organic photochemistry

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Since PAH have been shown to undergo photolysis on a variety of laboratory and environmental surfaces, including silica and alumina,43 soot,44 fly ash,45,46 and vegetation,47,48 we hypothesized that the PAH abundance profile in light-exposed urban film samples would differ from that in samples grown under dark conditions. In order to explore the influence of light on the PAH profile of urban film, we first compared the concentration ratios of three isomeric pairs of PAH—anthracene (ANT)/phenanthrene (PHE), pyrene (PYR)/fluoranthene (FLU), and benzo[a]anthracene (BaA)/(chrysene (CHR) + triphenylene (TRI))—in light-exposed films with those in films collected under dark conditions. These PAH pairs were chosen because they have similar physical properties (i.e. volatility) but different photochemical reactivities: at the surface of silica gel, for example, the photochemical half-lives of ANT (1.6 h), PYR (35.5 h), and BaA (2.4 h) are significantly shorter than those of PHE (268 h), FLU (99.1 h), and CHR (101 h).45 As shown in Figure 3, the ANT/PHE, PYR/FLU, and BaA/(CHR+TRI) ratios in light-exposed films were consistently lower than those in films collected under dark conditions, which provides convincing evidence that PAH molecules undergo photooxidation in urban film. The average ANT/PHE, PYR/FLU, and BaA/(CHR+TRI) ratios under light conditions were 85 ± 6%, 94 ± 6%, and 85 ± 7% of those observed under dark conditions for samples collected on the same day (here, the uncertainties indicate the standard deviations of the calculated values). The modest reduction in the PYR/FLU ratio in illuminated films is unsurprising, since pyrene is by far the least photoreactive of the three ‘reactive’ PAH, and since the difference between the photochemical reactivities of pyrene and fluoranthene is significantly smaller than those of the other two PAH pairs studied.45 The photochemical oxidation of surface-sorbed PAH displays a strong substrate dependence: whereas PAH photolysis rates at the surface of silica gel and alumina are largely determined by the inherent photoreactivity of each PAH,43,45 those at the surface of carbonaceous material, such as soot, fly ash, and carbon black, are also influenced by light absorption by the substrate (i.e. an “inner-filter” effect),43,45 and in some cases are limited by the diffusion of PAH from underlying, light-shielded layers to near-surface, illuminated areas.44,46,49 For these reasons, PAH associated with carbonaceous material are often resistant to photodegradation; further, their photooxidation rates are often less different than one would expect from reactivity considerations alone. The light-

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dependent PAH ratios observed in the present study, therefore, suggest that these six PAH were not entirely associated with elemental carbon contained within the film but rather were at least in part associated with less strongly absorbing particulate and/or contained within the film’s bulk organic fraction. This interpretation is supported by comparison of the ratio of the 5-ring PAH benzo[a]pyrene (BaP) to the 6-ring PAHs indeno[1,2,3-cd]pyrene (IND) and benzo[ghi]perylene (BghiP) under light and dark conditions. Although the photochemical half-life of BaP on silica gel (2.7 h) is significantly lower than that of IND (62 h) and BghiP (10.5 h), their photoreactivity is relatively similar on lightabsorbing carbonaceous substrates.45 Here, the average BaP/IND and BaP/BghiP ratios in lightexposed films were 94 ± 8% and 95 ± 11% of those in films collected under dark conditions for samples collected on the same day (again, the uncertainties indicate the standard deviations of the calculated values). Since these species would be expected to be exclusively particle associated,50 these results suggest that the photooxidation of particle-associated PAH in film is less significant than that of semi-volatile PAH—such as ANT, PYR, and BaA, discussed above— that are more likely to partition into the film’s organic portion. In the companion paper to this study, Baergen et al.30 provided the first direct field evidence that nitrate anion undergoes photochemical loss in urban film; here, we provide evidence that urban film also serves as a photochemical sink for semivolatile organic pollutants. Because the campaign took place during the autumn period, the actinic flux measured over its duration was significantly lower than that measured during the summer months at the same location (see Figure S5). Given the substrate effects discussed above for PAH photooxidation, we suggest that the results obtained in the present study therefore do not represent the maximum possible magnitude of film-phase processing of photoactive organics. Further study of urban film composition under summer illumination conditions and/or in equatorial regions (i.e. regions with lower solar zenith angles51) would help to clarify this issue. Although some indirect evidence exists for photochemical loss of organic contaminants in urban films,9,15,17 only one other study to date has explicitly compared the organic composition of urban film collected under dark and light-exposed conditions. In this study, Wu et al. reported that the congener profile of polychlorinated biphenyls (PCBs) in films collected on glass beads exposed to

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the elements was similar to those collected on sheltered glass beads.8 In this case, the authors attributed this observation to a lack of PCB photodegradation in the film, which is understandable given that the environmental lifetimes of PCBs are lengthy.52 However, it is possible that sample-tosample scatter in film composition could obscure subtle effects of photochemical processing even for more photochemically active species. The internal ratioing techniques discussed in this section represent a useful strategy for extracting from complex and variable samples information regarding the influence of photochemistry on urban film composition. 3.3

Film composition differs from that of colocated PM10

As shown in Figure 4, the OC–EC ratios of film samples collected on quartz-fiber filters are substantially higher than those of simultaneously collected PM10 samples. Insight into this observation is provided by the size-segregated particle samples collected in the final week of the campaign: as illustrated in Figure S7, whereas EC is primarily associated with submicron particles (Berner impactor stages 1–3), a substantial fraction of OC is present in supermicron particles. The lower relative abundance of EC in the film, therefore, may reflect contributions from particles > 10 µm in diameter, which would be expected to be low in EC. The OC–EC ratios of film samples in this study are also higher than those measured in previous studies of urban film. For example, Lam et al. found an OC–EC ratio of ~ 7 in a composite urban film sample obtained from 15 windows in downtown Toronto,1 and Favez et al. reported a range of ~ 0.4–4.6 for vertical glass samples exposed for several months to several years in five European locations.5 The higher OC–EC ratios associated with film samples in the present study may also reflect the lack of rain wash-off as a loss mechanism for OC31 and the partitioning of gas-phase organics to the film surface. We note that the calculated OC–EC ratios likely represent an upper limit for those in films collected on glass beads, since the quartz filters themselves provide a substantial internal surface for adsorption of gas-phase organics.53,54 Indeed, as shown in Figure S8, the area-normalized WSOC concentrations measured in film samples collected on quartz-fiber filters are significantly higher than those measured in film samples collected on glass beads. The alkane profiles of both light-shielded urban film samples and simultaneously collected PM10 samples are shown in Figure 5. As illustrated there, both sample types reflect significant contributions from the odd carbon-number alkanes C27–C33, which are emitted into the atmosphere 13



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from abrasion of epicuticular plant waxes.55 Figure 5 also shows significant contributions from even-numbered alkanes, which are of petrogenic origin, for both sample types. The distribution profile of alkanes in urban particulate matter has been shown to display a strong temperature dependence, with reduced contributions from more volatile lower alkanes at elevated temperatures.56 Here, the PM10 petrogenic signature peaks at C28. This profile is similar to that previously observed at an urban site in summertime Vienna, Austria,57 and is intermediate between typical winter and summer alkane profiles reported in Nagoya, Japan.56 Interestingly, Figure 5 also shows that the relative contributions of C30 and C32 alkanes in this study are significantly higher in urban film than in PM10, which may reflect contributions from settled paved road dust and tire tread particulate, both of which are enriched in higher alkanes.58 The carbon preference index (CPI), here defined as the ratio of odd (ΣC21–31) to even (ΣC22–32) alkanes, is commonly used to provide an indication of the relative contribution of biogenic and anthropogenic sources to alkane profiles in particulate matter59 and urban film.3,10 Since biogenic alkanes display a strong carbon number preference, CPI values significantly greater than 1 indicate a significant contribution from biogenic emissions; by contrast, since petrogenic alkane sources do not display a carbon number preference, a CPI of ~ 1 indicates that anthropogenic emissions dominate. The average CPI value for urban film in this study (1.5 ± 0.1), is lower than that for PM10 (1.9 ± 0.2), which most likely reflects the contributions from C30 and C32 alkanes discussed above. The urban film CPI value is also slightly lower than those obtained in one study of urban window films in Toronto, Canada,3 and significantly lower than those obtained in two other studies conducted in the same city (the Diamond et al.13 paper does not calculate a CPI, but does provide alkane profiles for urban film samples collected from seven locations),10,13 which found the alkane composition in several cases to be almost completely dominated by compounds from plant waxes (C29 and C31 alkanes). Together, these results highlight that the anthropogenic influence on urban film composition can vary even within urban centres. As illustrated in Figure 5, the PAH profile in light-shielded urban film samples obtained in this study differs from that of co-located PM10 samples. In one of the few studies to date that have compared the organic composition of surface films with that of aerosol samples collected simultaneously at the same location, Chabas and coworkers reported that the organic carbon

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content of films was lower than that of aerosols, and attributed this reduction to the oxidative degradation of organics at the film surface.6 In order to assess the extent to which differences in the PAH profile between urban film and PM10 reflect film-phase reactive processing of PAH, we again used a PAH concentration ratio approach. As shown in Figure 3, the BaA/(CHR + TRI) and BaP/BghiP ratios in the film were consistently lower than those in PM10—over the duration of the sampling campaign, the average BaA/(CHR + TRI) and BaP/BghiP ratios in the light-shielded film were 83 ± 13% and 66 ± 15% of those in PM10, where the reported errors correspond to the standard deviations of the calculated values. Since laboratory studies have shown that the heterogeneous reactivities of BaA and BaP with the gasphase oxidants NO2,60 O3,61,62 and NO363 are higher than those of CHR and BghiP, these results provide strong evidence that PAH undergo reactive processing in urban film, even under dark conditions. In the following paragraphs, we consider the potential role of each of these three oxidants in turn. As shown in Figure S1, the NO2 concentrations are elevated at the measurement site: the median hourly NO2 concentration was 88 µg m-3 over the duration of the campaign, and ~ 8% of the hourly measurements were above 200 µg m-3. However, studies have shown that the reactivity of PAH with NO2 at the surface of soot,64 graphite65 and particulate diesel exhaust66 is slow, and further does not depend strongly on PAH identity. Since BaP and BghiP would both expected to be exclusively particle-associated,50 the reduction in the BaP/BghiP ratio in urban film as compared to PM10 therefore most likely does not arise via film-phase processing by surface-sorbed NO2. The heterogeneous reaction of PAH with O3—unlike with NO2—displays a strong dependence on PAH identity on a variety of surfaces, including graphite61 and model urban film,67 which suggests a role for O3 in the film-phase processing of PAH. Unfortunately, however, ozone measurements were not available at the sampling site. Finally, since the reaction of PAH with NO3 is rapid68,69 and, in ambient PM2.5, also displays a dependence on PAH identity,63 we cannot exclude the influence of nighttime NO3 chemistry on the PAH abundance profile in urban film. This latter process is especially important, since the reaction of NO3/N2O5 with PAH leads to the formation of mutagenic nitro-PAH.63,70

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Interestingly, as shown in Figure 3, the ANT/PHE and PYR/FLU ratios in film collected under dark conditions are nearly identical to those in PM10. Since these species are semivolatile, we would expect them to enter the film not only by particle deposition but also by condensation. Indeed, Figure 5 shows that the film is enriched in fluoranthene and pyrene relative to PM10, which provides further evidence that the film acts as a reservoir into which semivolatile organics can partition.13,14,18 In this context, these results do not necessarily imply that these PAH do not undergo differential oxidative processing in the film. A complete understanding of the behaviour of semivolatile PAH in the film environment, however, would require knowledge of the gas-phase PAH profile at the sampling site.

4

Acknowledgments

We thank Sylvia Haferkorn and Yoshi Iinuma for assistance with CPP-GC-MS measurements; Anett Dietze, Susanne Fuchs, and Anke Rö dger for performing sample analysis; Konrad Mü ller, Gerald Spindler, and Achim Grü ner for assistance with sampler deployment and helpful discussions; and Olaf Bö ge, Ricarda Grä fe, and Cornelia Kurze for assistance with sampler construction. We appreciate the support of the Saxon State Agency for Environment, Agriculture and Geology (LfULG) in conducting the sampling at Leipzig-Mitte as well as their provision of additional data. We thank NSERC for ongoing research support. A.M.B. acknowledges NSERC for the award of a CGS-D graduate fellowship and the Department of Chemistry at the University of Toronto for travel support.

5

Supporting information

Figures and tables showing selected meteorological and air quality parameters at the Leipzig-Mitte air quality monitoring station for the duration of the sampling campaign; images of the urban film sampler employed in the study; organic carbon (OC) and elemental carbon (EC) content of quartz filters exposed at the sampling site; concentrations of selected polycyclic aromatic hydrocarbons (PAH) within urban film; irradiance at the sampling location; average mass concentrations of ΣPAH, Σalkanes, OC, EC, and water-soluble organic carbon (WSOC) in 24-h size-segregated particle samples; a comparison of WSOC concentrations in glass bead samples and co-located quartz filter samples; and a summary of previous measurements of PAH and alkanes in urban film. 16


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5


Figures

Figure 1 Area-normalized concentrations of a) water-soluble organic carbon (WSOC), b) polycyclic aromatic hydrocarbons (PAH), and c) alkanes as a function of urban film exposure time. The black and yellow circles denote samples collected under light-shielded and light-exposed conditions, respectively. The hollow red circle represents a sample for which the concentrations of all alkanes were below the limit of quantification (here, defined as the average blank concentration + 10x the standard deviation of the blank concentration). In order to assess the influence of variable airflow within the sampler, three samples were taken on Day 21.

17



Figure 2 The mass ratio of a) water-soluble organic carbon (WSOC) to the total urban film ion content, b) phenanthrene (PHE) to ΣPAH and c) benzo[a]pyrene (BaP) to ΣPAH as a function of atmospheric exposure time. The black and yellow circles denote samples collected under light-shielded and light-exposed conditions, respectively.

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Figure 3 PAH abundance ratios in urban film as a function of atmospheric exposure time: a) anthracene (ANT)/phenanthrene (PHE), b) pyrene (PYR)/fluoranthene (FLU), c) benzo[a]anthracene (BaA)/(chrysene (CHR) + triphenylene (TRI)), and d) benzo[a]pyrene (BaP)/benzo[ghi]perylene (BghiP). The black and yellow circles denote samples collected under light-shielded and lightexposed conditions, respectively; the red circles denote PAH abundance ratios in co-located PM10 samples.

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Figure 4 Organic and elemental carbon (OC–EC) ratios of urban film as a function of atmospheric exposure time (n = 2 or n = 3 filter punches for each exposure time). The black and yellow circles denote samples collected under light-shielded and light-exposed conditions, respectively; the red circles denote OC–EC ratios in co-located PM10 samples.

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Figure 5 PAH and n-alkane distributions in urban film and co-located PM10 samples. Carbon preference index (CPI) reflects the ratio of odd (ΣC21–31) to even (ΣC22–32) alkanes. Error bars denote the standard deviation of the samples for which all reported analytes were above the limit of quantification (here, defined as the average blank concentration + 10x the standard deviation of the blank concentration).

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