The winter season (P1–P2) is characterized by the most considerable weather fluctuations due to the large-scale temperature and pressure gradients between midlatitudes and the Arctic (Sjöblom, 2010). During the sampling period the temperatures ranged from -18 to +4 ∘C, being -5 ∘C on average. In winter, the Arctic weather is strongly controlled by extratropical cyclones, commonly referred to as low-pressure systems, which are at their strongest and most frequent in December–February (Zhang et al., 2004) and have demonstrated an increasing tendency in the areas around Svalbard over the last few decades (Wickström et al., 2020b). These cyclones are the main atmospheric heat source to the Arctic in winter (Fig. S4), and they are also associated with high winds and precipitation (Wickström, 2020). Such inflows from southern latitudes interrupt the Arctic atmosphere thermal stability and cause the BLH fluctuations (Fig. 2). The southeasterly wind was observed during all the sampling days in the cold P1–P3 periods (Fig. S5a). Thus, the sampling site was upwind of the coal-burning emissions at the power plant in winter and spring (Fig. S1c).

The average P1 winter ∑22 PAH total (gaseous and particulate, G + P) concentrations were 1068±449 pg m-3. The level of ∑29 oxy-PAHs (5934±1480 pg m-3) was about 6-fold higher than PAHs, while ∑35 nitro-PAHs (84±19 pg m-3) were an order of magnitude lower. As discussed earlier, the LRAT of pollutants to the Arctic lower troposphere is maximum in winter, and the PAH multidecadal monitoring data from the Zeppelin station reflect this trend well (Fig. S3 and Yu et al., 2019). The measured PAH concentrations in Longyearbyen were similar (about 19 % higher) to those detected at the Zeppelin station (Table S7), confirming LRAT of anthropogenic pollutants as a dominant source of PAHs in the Arctic air during winter. According to the 10 d back-trajectory analysis, the transported air masses in winter mainly originated from the European sector, including the areas of Scandinavia, northern Europe, western Russia, and the West Siberian plain (Fig. 3a). Compared to the winter 2018 PAH levels in the background continental air in northern Finland (Pallas, Matorova station; 68∘00′ N, 24∘14′ E; 340 m a.s.l.), from where the air masses are frequently delivered, the Longyearbyen PAH concentrations are a factor of 2 lower (Table S7), which may give us a rough estimation of the PAHs' LRAT efficiency though it largely depends on the weather along the transport pathway, as discussed earlier.

At the end of January, the Longyearbyen area starts receiving sunlight. This promotes photochemical processes and notably PAC degradation. As a consequence, a factor of 5 lower PAH concentrations were measured at the Zeppelin station in spring compared to the winter levels (Table S7), also in agreement with earlier studies (Singh et al., 2017; Halsall et al., 1997; Fu et al., 2009). In contrast, a double concentration of ∑86 PACs (13277±1295 pg m-3) was detected in Longyearbyen air in February (P2), and even higher levels (19281±4876 pg m-3) were found during the March-to-April period P3. A pronounced increase in concentrations was revealed for most of the compounds. Compared to the winter levels, a 6-fold increase was detected for ∑22 PAHs (5902±2421 pg m-3), and 2-fold-higher concentrations of ∑29 oxy-PAHs (13229±2681 pg m-3) and ∑35 nitro-PAHs (150±34 pg m-3) were determined in spring P3. These levels are about 1 order of magnitude lower than the average European urban and suburban PAC concentrations measured during cold seasons (Tomaz et al., 2016; Albinet et al., 2008), though the difference is less pronounced with the annual levels in background air (Nežiková et al., 2020): the sum spring concentrations of PAHs (n=15) and high-molecular-weight (HMW) oxy-PAHs (n=4) were about a factor of 2 lower, while low-molecular-weight (LMW) oxy-PAHs (n=4) and the sum nitro-PAHs (n=15) were in a similar range to up to 1 order of magnitude higher. Several individual PAC concentrations increased by a factor of 2 to 525 (depending on the compound) compared to the winter levels, and several of them significantly exceeded (more than twice) the annual mean urban levels in central Europe (Tables S1 to S3; Tomaz et al., 2016; Albinet et al., 2008). As for the Arctic scale, the ∑20 PAH concentrations measured at UNIS were a factor of 30 higher than those at the Zeppelin background station (oxy- and nitro-PAHs are not monitored at the station; Table S7). Such a large difference implies a significant contribution of local emissions to the overall PAC levels in Longyearbyen as the UNIS and Zeppelin stations have similar air masses transport from lower latitudes. According to the backward trajectory probability analysis, in spring, the source regions shifted from east to west and significantly northward: from continental locations in Russia and Europe (P1–P2; Fig. 3a–b) to marine ones over the North Atlantic (Fig. 3c), mainly from the area north of the 75 ∘ N latitude with no significant PAC sources. Thus, the influence of local emissions might be prevailing in spring.

Regional high Arctic weather conditions established by spring favour pollution accumulation in the near-ground air layer. Continuous radiative cooling of snow-covered surfaces during several months of the polar night results in low temperatures at the ground, which creates thermally very stable air stratification. The median values of Ribulk for P2 and P3 periods were significantly higher than 5.25 (Fig. 2), indicating reduced vertical air mixing (Stull, 1988). Strong temperature inversions were frequent during this period with a median TIS of about 4.5 K. The inversion events were the most persistent in spring due to typical sustained cold ambient temperatures (-13.8±3.4 ∘C on average) in March and April (P3). The estimated BLHs were as low as 100±79 m with a minimum of 19 m under calm (1.9 m s-1 wind as the median), cold (-15.4 ∘C) conditions (Table S6). Thus, the spring near-ground PAC concentrations were greatly influenced by local emissions, which were frequently trapped beneath an inversion layer, and their dilution was limited to a very shallow volume of cold air under suppressed vertical mixing conditions. In addition, for secondary species such as nitro- and oxy-PAHs, secondary formation processes could also occur during these periods and might be promoted by such meteorological conditions. This will be further discussed in a forthcoming article.

Similarly to PACs, about 2-fold-higher EC and OC concentrations were measured in P2 and P3 periods (Fig. 2 and Table S1), and the peak values were found on weekends and during the Easter holidays. This highlights that a large part of OC was probably related to primary combustion emissions in these periods, although secondary processes (formation of secondary organic aerosols) may occur (Hallquist et al., 2009; Heald and Kroll, 2020; Jimenez et al., 2009; Kroll and Seinfeld, 2008). The coal-burning emissions from the power plant were upwind during all the sampling days. Thus, this source probably had no impact on the concentrations and patterns observed. Car traffic intensity is quite steady through winter to spring (P1–P3), and the same emission intensities are presumed as those of similarly cold ambient temperatures during these periods. However, significantly higher values of the ratio (BghiP + Cor) / ∑PAHp, representative of gasoline emissions (sum of filter concentrations of 12 PAHs mainly present in the particulate phase (PAHp), from benzo[a]anthracene to coronene according to our observations; Table S1) (Marchand et al., 2004; Albinet et al., 2007) were noticed during and after the polar sunrise (P2 and P3, respectively) compared to in the winter period P1 (Fig. 2). Human activities increase greatly with the sun's return. The town residents already start driving snowmobiles predominantly for recreational purposes during the twilight period P2. This is also the main attraction for tourists, the number of which significantly increases during the spring period P3 (about 16 000 hotel guest nights monthly are registered in March and April; Statistics Norway, 2016) when the greatest number of snowmobiles (2135 vehicles; Statistics Norway, 2018) are in use.

The Longyearbyen snowmobile park predominantly consists of four-stroke engine touring vehicles (60–150 hp, ca. 45–112 kW) and a small number (about 10 %) of two-stroke modern (sport models) and old snowmobiles (registered before 2010). All the snowmobiles are gasoline driven (10–20 L per 100 km; NS-EN 228 Norwegian standard winter gasoline 95 no. 2 with up to 15 % MTBE (methyl tert-butyl ether); Jøran Storø, LNS Spitsbergen, personal communication, 2020) with an addition of synthetic oil (about 2 L per 100 km) for two-stroke engines. Snowmobile emissions constitute a significant source of carbon monoxide, nitrogen oxides (NOx), particles, aldehydes, and a large number of aromatic hydrocarbons (Sive et al., 2003; Bishop et al., 2001; Shively et al., 2008; Zhou et al., 2010; GYC, 2011); further, several studies have confirmed significant contamination of snow along a snowmobile track by PAHs (McDaniel and Zielinska, 2014; Rhea et al., 2005; Oanh et al., 2019). Due to potentially hazardous consequences related to the emissions, the use of snowmobiles is regulated and controlled in the USA with the summed hydrocarbons and carbon monoxide emission limits of 15 and 90 g kW h-1, respectively (U.S. NPS, 2015). Significantly higher concentrations of numerous aromatic hydrocarbons were measured in Longyearbyen air in spring 2007 caused by snowmobile driving (Reimann et al., 2009). The last decade's shift from two- to four-stroke snowmobile engines, including in Svalbard, has reduced CO emissions. However, the recent inventory of 2017–2021 production year snowmobiles (499–1056 cm3 engine displacement; EPA, 2020) still shows emissions of hydrocarbons (and probably PACs) up to an order of magnitude higher from two-stroke modern sleds, the small number of which in Longyearbyen can cause severe air pollution (Reimann et al., 2009). In conclusion, the results obtained here highlighted a strong impact of the snowmobile emissions on the ambient air PAC concentration levels observed.

The Arctic lower troposphere's summer composition is primarily controlled by local Arctic weather conditions and local anthropogenic and natural emission sources due to a less efficient LRAT mechanism and enhanced sink processes for PACs. The influence of the midlatitude emissions is more pronounced at higher altitudes because the summer atmospheric transport is directed to the higher troposphere (Klonecki, 2003; Stohl, 2006; Bozem et al., 2019). The Arctic age of summer air near the surface north of 75∘ N is 13–17 d (Stohl, 2006), and the estimated travel time for gaseous pollutants is about 3 weeks (Bozem et al., 2019) while particles are efficiently scavenged by wet precipitation (Willis et al., 2018, and Table S5). Thus, local emissions are of great importance in summer in the high Arctic.

Increased incoming solar radiation (24 h of sunlight) and warmer ambient temperatures are the most notable changes in the high-latitude summer. During the sampling period P4 the temperatures ranged from -0.1 to +9.3 ∘C with an average value of +3.6 ∘C on the sampling days. Loss of land-based snow cover and retreat of sea ice decrease surface albedo, which consequently creates the dominance of downward short-wave radiation in contrast to the spring period P3 (Fig. 1). Moreover, large areas of open ocean and increased solar insolation create relatively high atmospheric humidity. It allows the formation of low-level liquid-containing clouds (Browse et al., 2012), which are ubiquitous in the Arctic during summer (Lawson et al., 2001; Cesana et al., 2012). These low-level clouds and fogs produce frequent drizzle, which leads to efficient removal of particle-associated (Browse et al., 2012) and gaseous-phase water-soluble pollutants, including the PACs (Lei and Wania, 2004; Shahpoury et al., 2018). Thereby, the Arctic troposphere during summer is generally cleaner than in winter (AMAP, 2006; Stohl et al., 2007) and is often considered pristine (Browse et al., 2012; Garrett et al., 2010). The 2-decade atmospheric measurements at several high Arctic monitoring stations confirm that PAH concentrations throughout the year are minimal during summers (Yu et al., 2019). The summer PAH levels are usually 1–2 orders of magnitude lower than in winter and often close to detection limits (Table S7 and Yu et al., 2019; Prevedouros et al., 2004; Singh et al., 2017). Nonetheless, summer ∑22 PAHs concentrations (1779±1210 pg m-3) measured in Longyearbyen air were about 67 % higher than the winter P1 levels (Table S1). For individual PAHs, the largest differences (concentrations ≥2 times higher in summer) were observed for phenanthrene, 1- and 2-methylnaphthalene, pyrene, acenaphthene, and chrysene while other compounds showed comparable concentrations in both seasons. The ∑29 oxy-PAH (7219±1692 pg m-3) concentrations were 22 % higher, while the levels of ∑35 nitro-PAHs (66±9 pg m-3) were 22 % lower (Tables S2–S3) during P4 compared to winter P1. The PAH levels observed in Longyearbyen were a factor of 19 higher compared to levels measured at the Zeppelin background station during the same period of summer 2018 (Table S7), stressing the predominant contribution of the local anthropogenic emissions. Due to the town's lower demand for electricity and central heating during brighter and warmer midnight sun conditions, coal-burning emissions at the power plant decrease (Bøckman, 2019) as do car traffic emissions (Statistics Norway, 2016), also because of preferences for biking and walking within the town. Meanwhile, boat traffic increases dramatically, mainly due to tourist cruises (about 53 000 vessel passengers during the summer season; Statistics Norway, 2016) and cargo delivery, and can significantly contribute to the local air pollution (Eckhardt et al., 2013; Zhan et al., 2014; Law et al., 2017; Ferrero et al., 2016).

Due to the town's coastal location and the summer-specific land–sea thermal gradient, the northwesterly winds become frequent during summer. These winds are characterized by lower velocity (4.0 m s-1 as the median) compared to winds that are southeasterly (7.4 m s-1 as the median), the annual prevailing wind direction (Table S6). About half of the P4 period air samples were collected under northwesterly wind conditions (Fig. S5c) when the sampling site was downwind of the coal power plant and the harbour (Fig. S1c and d). The Flu / (Flu + Pyr) coal combustion diagnostic ratio value of 0.81±0.09 was found to be similar to previously reported values (Drotikova et al., 2020) with low variability over the entire summer sampling period P4, indicating steady contamination from the power plant. Also, the power plant plumes are emitted at high elevation (95 m a.s.l.) and a short distance from UNIS (about 800 m) and thus might not be easily detectable at the UNIS sampling site due to higher BLH, less frequent thermal inversion events, increased air layers' convection, and generally windy conditions during summer (4.9 m s-1 during P4 sampling days). However, about 29 %–88 % higher ∑86 PAC, EC, and OC concentrations were detected on the days with a northwesterly wind, and the harbour emissions seemed likely responsible for it as substantial air pollution can be caused by ship emissions when ships are docked and during port manoeuvering (Huang et al., 2018a). Moreover, these summer PAC concentrations (measured during the P4 period of May to June 2018) were found to be about 3-fold higher than those reported previously for the autumn (measured during the period of late August to September 2018; Drotikova et al., 2020), when ship traffic in the fjord was already notably lower as the sailing season was close to its end. The average (n=5) concentrations when the wind blew from the southeast direction (from Adventdalen; n=5) were for ∑22 PAHs 1116±278, for ∑29 oxy-PAHs 6557±1821, and for ∑35 nitro-PAHs 65±10 pg m-3 and likely mainly influenced by local car traffic. On average, a factor of 2 and 1.2 higher concentrations of ∑22 PAHs (2253±1416 pg m-3) and ∑29 oxy-PAHs (7691±1553 pg m-3), respectively, were measured on the days with northwesterly wind (from the harbour; n=7). The ∑35 nitro-PAH (66±10 pg m-3) level did not change markedly. The maximum summer concentrations (∑22 PAHs = 4633, ∑29 oxy-PAHs = 10 256, and ∑35 nitro-PAHs = 75 pg m-3) were found downwind of the harbour when the source strength (summed total tonnage of ships registered in the harbour during the sampling hours) was maximal and hourly median wind speed varied between 1.8 and 5.7 m s-1. Dispersal of ship plumes and their visibility at a fixed measurement point are dependent not only on the source strength but also on the atmospheric stability, wind velocity, and turbulence intensity (Contini et al., 2011). Higher local air pollution by ship emissions and its accumulation can be expected at lower wind speeds, similarly to in the summer 2009 case study in Ny-Ålesund, Svalbard (Eckhardt et al., 2013).

To further highlight ship emissions' influence, only samples collected when the moderate wind (4.0 m s-1 as the median) blew from the northwesterly sector (292–313∘; n=7) were considered (Fig. 4). The total (arrival + departure) gross tonnage of vessels (Kystdatahuset, 2018) was used as an indicator of local ship traffic. A correspondence was found between the summed total tonnage of ships registered in the harbour during the sampling hours and the Flt / (Flt + Pyr) and BaAnt / (BaAnt + Chry) diagnostic ratios, providing clear evidence of the harbour emissions' contribution (Zhang et al., 2019; Drotikova et al., 2020). A factor of 4 higher concentrations of the PAHp were measured on the days with northwesterly wind compared to the days when wind blew from a southeast direction, which is also in agreement with a recent study on ship emissions in the harbour of Longyearbyen (Dekhtyareva, 2019). Such a large increase in particle-associated PAHs also suggests the harbour being the predominant influence as only about 6 % of the power plant total PAC emissions (gaseous + particulate phases) were associated with the particulate phase (Drotikova et al., 2020). Finally, a good correlation of the PAHp / OC ratio values with the summed vessels' total tonnage was also found (Fig. 4), highlighting a higher content of PAHs in OC under such a harbour influence.

As a last piece of evidence indicating a significant influence from the harbour, the OC / EC ratio values ranged between 2.5 and 7.4 (3.6 as the median) on the selected days. Marine fuel quality and engine operating conditions cause large variation in OC and EC emissions. OC / EC ratio values between 2 and 5 are often detected for low- and high-power diesel-operated vessel plumes, while the higher values are typical of heavy-fuel-oil engine emissions (Zhang et al., 2019, 2020; Sippula et al., 2014), as well as of the ship emissions at low-speed manoeuvering during the harbour departure and arrival (Zhang et al., 2016; Sippula et al., 2014; Huang et al., 2018b). Furthermore, a plot of the daily OC / EC ratio values versus total tonnage demonstrates the ratio increasing with the tonnage (Fig. S6).

Phenanthrene, fluorene, naphthalene, benzo[a]anthracene, and 1-methylnaphthalene accounted for 76±8 % of total PAHs measured in Longyearbyen. Phenanthrene was the predominant compound in all the periods (P1–P4) and accounted for about 30 % on average, being higher in summer. The contribution of naphthalene varied from 5 % to 24 %, with a higher proportion in the spring period P3. This compound is not included in further discussion as it is not source-specific.

The resulting winter profile P1 (Fig. 5a) was dominated by phenanthrene and fluorene (60 %) with a smaller contribution of about 8 % of anthracene, which was the highest compared to the other periods. Anthracene is one of the most reactive PAHs (Keyte et al., 2013). Its higher proportion in winter may indicate accumulation of anthracene due to a lower level of atmospheric oxidants under dark conditions during the P1 period.

The twilight P2 and spring P3 profiles were strongly influenced by benzo[a]anthracene and 1-methylnaphthalene, accounting for 40 % of the total PAH relative compositions with a higher proportion on weekends and public holidays, when snowmobile traffic is increased. Comparing the PAH spring weekend profiles (n=7) with the rest of the cold-period day profiles (P1–P3; n=12) confirmed a significant increase in the benzo[a]anthracene proportion caused by snowmobiling, as well as the higher contribution of 1-methylnaphthalene, fluoranthene, acenaphthene, and coronene to the total PAH profile (Fig. S7). Overall, benzo[a]anthracene, phenanthrene, and 1-methylnaphthalene accounted for 68 % of the average PAH profile of these days with known high snowmobile traffic. The predominance of phenanthrene and methylated PAHs in snow samples contaminated by snowmobile emissions has previously being reported (McDaniel and Zielinska, 2014), although the contribution of 2-methylnaphthalene and 2-methylfluoranthene in the present study was found to be negligible. Different engine technologies and fuel quality may have been responsible for this.

Like in winter, phenanthrene and fluorene were the predominant compounds of the summer PAH profile P4 and accounted for 59 % of it. The proportion of 2-methylnaphthalene was found to be higher in summer. The summer profile represents an average picture of two sample groups collected under different wind conditions (direction) and thus influenced by different sources, though car traffic is similar (Fig. S1c). To investigate the source-specific emissions, profiles of the days with northwesterly wind (n=7) were compared with those samples collected when wind was from Adventdalen (southeasterly wind; n=5). The difference between the average profiles was mainly driven by the significantly higher contribution of fluoranthene, 2-methylfluoranthene, and chrysene on the days with northwesterly wind from the harbour and the coal power plant (Fig. S8). As specified above, increased air convection in summer, the sampling site's proximity to the power plant, and its almost exclusively gaseous emissions mean that the ground-level emissions from the harbour are more detectable at UNIS than the high-elevation power plant plumes. Our assumption is supported by frequent reports of fluoranthene and chrysene as major PAHs in ship emissions (Huang et al., 2018b; Zhang et al., 2019; Zhao et al., 2020, 2019) and an absence of contribution changes of phenanthrene, fluorene, and pyrene, the predominant compounds emitted from the power plant (Drotikova et al., 2020).

Benzophenone, phthalic anhydride, 9-fluorenone, 1-naphthaldehyde, and 1,2-naphthalic anhydride were the predominant substances and accounted for 89±4 % of the total oxy-PAHs in all periods (Fig. 5b). The main changes in the profiles were caused by a 16 % increased contribution of benzophenone in spring and a 17 % higher proportion of phthalic anhydride in summer. Among the minor compounds (ranging from 25 to 400 pg m-3) accounting for about 10 % of the total oxy-PAH profile, higher proportions of 2,3-naphthalenedicarboxylic anhydride and anthrone were found in summer. Detailed analysis of the average summer profiles based on different prevailing wind directions (Fig. S9) did not reveal any primary local source influence. This may indicate the secondary formation of phthalic anhydride, 2,3-naphthalenedicarboxylic anhydride, and anthrone from PAH photochemical reactions (Keyte et al., 2013; Chan et al., 2009; Perraudin et al., 2007; Bunce et al., 1997; Lee and Lane, 2009). A higher contribution of 9-fluorenone, 1-naphthaldehyde, 1-acenaphthenone, 6H-dibenzo[b,d]pyran-6-one, 9,10-anthraquinone, 2-methylanthraquinone, and benzanthrone to the total oxy-PAH summer profile was discovered on days with northwesterly wind (Fig. S9). Several of these compounds have previously been detected in the ship plumes (Zhao et al., 2020; Czech et al., 2017; Sippula et al., 2014). Note that 9-fluorenone and 9,10-anthraquinone are the main oxy-PAHs emitted after the local coal burning at the plant (Drotikova et al., 2020).

For the twilight P2 and spring P3 periods, a higher contribution of several oxy-PAHs was found on weekends and public holidays, namely benzophenone, 9,10-phenanthrenequinone, 2-formyl-trans-cinnamaldehyde, anthrone, 9,10-anthraquinone, and benzanthrone (Fig. S10a), and the greatest increases in proportion were found among the minor compounds 1,2-naphthoquinone, phthaldialdehyde, 2-methylanthraquinone, and 1,8-naphthalic anhydride (ranging from 33 to 102 pg m-3; Fig. S10b). These changes seemed likely to be caused by higher human activities and mostly snowmobile driving, as mentioned above. The use of oxidation catalysts as part of four-stroke snowmobile exhaust after-treatment (Meldrum, 2017) may lead to efficient oxidation of parent PAHs and consequent formation of their oxygenated derivatives.

1-Nitronaphthalene and dinitropyrenes accounted for 40±5 % of total nitro-PAHs in all periods (Fig. 5c). The contribution of 1-nitronaphthalene decreased in summer while it increased for 1-nitrobenzo[a]pyrene. Different wind directions did not cause notable changes to the P4 summer profile.

The twilight P2 and spring P3 period profiles differ greatly from the winter P1 and summer P4 total nitro-PAH relative compositions. The P2 and P3 profiles were enhanced by higher proportions of 9-nitroanthracene, 6-nitrochrysene, 3-nitrophenanthrene, 5-nitroacenaphthene, and 3-nitrofluoranthene. A study of daily profiles of the entire cold period (P1–P3) showed evident pattern changes with a significantly greater contribution of the listed nitro-PAHs during non-work days (Fig. S11). No scientific studies on PAC emissions from snowmobiles have been reported yet. Nonetheless, snowmobile driving is a well-documented major source of a list of aromatic hydrocarbons (Zhou et al., 2010; Reimann et al., 2009; Eriksson et al., 2003; Shively et al., 2008), and PAHs were detected in snow along snowmobile tracks (McDaniel and Zielinska, 2014; Rhea et al., 2005; Oanh et al., 2019). Moreover, a factor of 3 higher NOx emissions from snowmobiles compared to gasoline-driven cars in Svalbard are reported by the Norwegian Climate and Pollution Agency (Vestreng et al., 2009) with the main contribution from four-stroke snowmobiles. Modern-technology snowmobiles have a significant side effect of high production of NOx (Ray et al., 2012) due to oxidation reactions in exhaust after-treatment (Aubin et al., 2017). A recent field survey by Dekhtyareva (2019) in Svalbard also confirmed enhanced NOx emissions from snowmobiles. Excess of NOx during combustion drives nitration of PAHs producing nitrated PAH derivatives via electrophilic substitution (Ringuet et al., 2012; Heeb et al., 2008; Carrara and Niessner, 2011; Carrara et al., 2010; Hu et al., 2013; Keyte et al., 2013). Note that pattern changes of the daily total nitro-PAH profiles were the most easily applicable to recognize snowmobile emissions compared to PAHs and oxy-PAHs, and 9-nitroanthracene, 6-nitrochrysene, 3-nitrophenanthrene, 5-nitroacenaphthene, and 3-nitrofluoranthene seem to have strong specificity to the source.