Consecutive aerosol measurement campaigns were performed at existing field sites in a rural and an urban setting in southern Sweden, during the period of late July to early October 2018. The weather did not change dramatically between the two campaigns; hence seasonality is not expected to play a major role in the results. Figure S1 shows that the BC concentrations at the rural site during July–October did not change drastically between the times of the two campaigns. The urban campaign (September–October) took place at a curbside measurement site near a busy road junction surrounded by four- to seven-story buildings (about 30 000 vehicles per day) in Malmö, which has about 300 000 inhabitants, situated about 30 km east of Copenhagen. Air was sampled at 3 m above the pavement using 1/4” stainless steel tubing and a PM2.5 cyclone. Total inlet length was 4–5 m and the flow was 7–8 L min-1. The validity of the curbside site measurements as an indicator of the BC concentrations in the city was assessed by comparing simultaneously collected data at a rooftop urban background site (20 m a.g.l., PM2.5 inlet). The rural measurements (July–August) were performed at the Hyltemossa Research Station, about 65 km northeast of Malmö–Copenhagen. This station is part of both the ACTRIS (Aerosol, Clouds, and Trace gases European Research Infrastructure) and ICOS (Integrated Carbon Observation System) European environmental research infrastructure networks. The field site is surrounded by a managed spruce forest. Aerosols were sampled at a height of 30 m above ground through a PM10 inlet and 25 mm stainless steel tube, at a flow rate of 16.7 L min-1. The single-particle soot measurements at the rural site were conducted through a 1/4” tube at 20 L min-1 (turbulent flow), due to simultaneous eddy covariance flux measurements (not described here).

An overview of the BC measurement results is shown in Table 1. As expected, the curbside site has higher concentrations than the rural site, with campaign average factors of 2.5 for rBC mass, 2.2 for eBC mass and 3.2 for rBC number concentration. The differences in concentrations are due to local emissions, clearly shown in the diurnal pattern (Fig. 2), which follows traffic intensity. The comparison between the urban street level and urban background eBC levels are very similar in time series but with a lower daily maximum for the rooftop measurements (Fig. S5). This suggests that the curbside measurements are indicative of the city. The average eBC mass concentration at the rooftop site during the campaign was 15 % lower than at the curbside site. The largest difference between the urban and rural sites occur during morning rush hour when rBC mass concentrations are up to a factor 3.7 higher in the city. The mass concentrations are still higher in the city during nighttime, before the first rush hour peaks, but only by a factor of 1.2–1.3. The weekends have lower rBC mass concentration (MrBC of 0.11±0.06 µg m-3) compared to weekdays in the city, but it is still a factor of 1.8 higher than the average for the rural background station. The AAE is slightly higher at the urban site, which could be due to the differences in BC size distribution and coating (Virkkula, 2021).

The number fraction of particles containing a BC core was estimated by comparing the daily average of BC number concentration from the SP2 to the number concentration of particles larger than 64 nm as measured by the mobility particle sizers. At the rural site this fraction was 2.7±0.6 (1σ) %, while for the urban site the fraction was 13.4±5.5 (1σ) %. These estimates are biased low because the BC size distribution (as measured by a mobility particle sizer) is shifted to larger sizes if shape factors and coating are added. If the cutoff size for total particle number measured by the SMPS is increased to DP>130 nm (assuming a coating thickness of similar magnitude as the rBC core mass equivalent diameter), the fractions are increased to 6 % in the rural campaign and 45 % in the urban campaign. In a study outside Paris (Laborde et al., 2013), the number of BC particles counted by the SP2 (with no correction for particles outside the measurement range) was 0 %–15 % of the total particle number concentration (above 20 nm) measured by an SMPS. In airborne measurements over Europe (using an SP2 and a passive cavity aerosol spectrometer probe), Reddington et al. (2013) found that the number fraction of approximately 14 % of particles with dry diameters above 260 nm contained a BC core. In an older study, Rose et al. (2006), using a volatility tandem-DMA setup, measured summertime BC particle number fractions of 2 %–7 % at a rural site and 20 %–60 % in a street canyon, depending on the selected size (30–150 nm in mobility diameter).

The difference between rBC and eBC is significant at both sites (Fig. S6). For the full campaign averages, eBC is more than 3 times higher than rBC at both sites, with median ratios (eBC / rBC) of 3.04 and 3.65 at the urban and rural sites, respectively. Several studies have found similar discrepancies between rBC and eBC (Holder et al., 2014; Raatikainen et al., 2015; Sharma et al., 2017; Tasoglou et al., 2018; Li et al., 2019). The rBC mass in this study is biased low because of the limited range of the SP2, but given the size distributions measured (example shown in Fig. S7), this mass bias should be on the order of 10 %–20 %, which would still only lower the ratio between eBC and rBC to 2.4–2.7. One reason for biased high eBC values is that the measured absorption can be amplified by non-absorbing coatings of BC cores (Kalbermatter et al., 2022). It is well known that site-specific MAC can differ from the default values from the manufacturer of the AE33 (e.g., Cui et al., 2016). At a site nearby the rural site of this study, a MAC of 12.64 m2 g-1 at 880 nm has previously been derived using Aethalometer (AE33) absorption coefficients and elemental carbon (EC) measurements (Martinsson et al., 2017). This value is 1.6 times higher than the default, and if it is used, the ratio between eBC and rBC of the rural campaign would be ∼2.

The average geometric mean diameter (GMD) of the rBC mass size distribution, from lognormal fits of 24 h data, was highest in the rural setting (Table 1). The average mass size distributions for the full campaigns are shown in Fig. 3. Mass equivalent diameters of 168 and 148 nm for the rural and urban sites, respectively, correspond to BC particle masses of 4.5 and 3.1 fg. The number size distribution generally peaked close to the SP2 detection limit (64 nm, corresponding to 0.25 fg, for both campaigns), and the GMD of this distribution is therefore more uncertain and not presented here. An example of 24 h number and mass size distributions are shown in Fig. S7. The difference in mass GMD is statistically significant (P<0.01) and can be explained by different BC particle sources, air masses and coagulation. Figure S1 shows that while BC concentrations at the rural site were similar during the two campaigns, the source sectors changed slightly. Notably, more industrial emissions and less shipping emissions affected the concentrations during the urban campaign. Figure 4 shows the diurnal variation in GMD from mass size distributions averaged over 2 h. The urban site has a higher fraction of particles coming from local traffic, which have been shown to be smaller than particles from, e.g., biomass burning (Schwarz et al., 2008; Laborde et al., 2013; Saarikoski et al., 2021). This shifts the size distribution during daytime to slightly lower sizes, although the measurement uncertainties, including calibrations, of the SP2 are close to the difference in average GMD between the sites.

The BC particles at the rural site were characterized by a thicker coating than at the urban location, as measured by the SP2 delay time method (a typical example is shown in Fig. S8). The fraction of “thickly coated” BC particles with an incandescence peak height corresponding to a BC core diameter of mass 2.4 fg (∼136 nm mass equivalent diameter) was on average 71±8 % (1σ) at the rural site and 38±7 % (1σ) at the urban site, with no big difference during weekends. The short-term APM measurements verified the bimodal structure in effective density of the urban aerosol. The mass distribution of 150 nm particles is shown in Fig. 5. The two modes suggest that the aerosol is externally mixed, with more and less dense particles present. The two lognormal curve fittings have a mean mass of 0.86±0.02 fg and 2.49±0.02 fg, respectively. These masses correspond to effective densities of 0.49±0.01 and 1.41±0.01 g cm-3, respectively, which is very similar to the effective densities presented by Rissler et al. (2014), who reported two effective densities for 150 nm particles sampled in downtown Copenhagen of 0.53±0.05 and 1.36±0.15 g cm-3. For 50 and 75 nm particles, the average mass distribution was unimodal with a mean mass of 0.062±0.001 and 0.26±0.01 fg, respectively, corresponding to effective densities of 0.95±0.01 and 1.19±0.02 g cm-3. The average mass distribution of 100 nm particles was bimodal with mean masses of 0.33±0.01 and 0.68±0.01 fg, corresponding to effective densities of 0.63±0.02 and 1.29±0.01 g cm-3. These values are also similar to the densities in Rissler et al. (2014), but we did not observe an additional mode at higher masses for 50 and 75 nm particles but only the less dense, fresh soot mode. However, in single APM spectra, the mean mass mode was sometimes different from the average, and on a couple of occasions, a bimodal distribution was observed also for the smaller particles. No mass mode could be extracted for APM measurements of larger particle sizes (250 and 350 nm) due to low particle concentrations and consequently poor counting statistics.

Figure 6 shows the wind direction probability together with eBC from the 5-week rural campaign. The Malmö–Copenhagen area is in the direction 200–230∘ as seen from the rural site. Based on the HYSPLIT trajectory analysis (Sect. 2.2.8), some peaks in eBC, non-refractory PM1 and total particle number concentrations at the rural site coincide with air masses originating over Malmö–Copenhagen. These urban plumes were discernible when the air mass was relatively clean (typically of western or northwestern origin). However, no conclusive results on the influence of the urban site on rural levels could be derived from analysis of the limited field campaign data. Also, when analyzing the complete eBC dataset from 2018, there is no significant difference in the median eBC concentration for air masses with and without influence from Copenhagen and/or Malmö, according to the trajectory analysis. However, if we only consider air masses with insignificant precipitation within 48 h upwind of the rural site (i.e., less than <1 mm precipitation according the HYSPLIT model), the measured eBC median concentration is significantly higher in the air masses that have moved over Copenhagen–Malmö, 354.7 ng m-3 compared to 226.9 ng m-3 without influence from Copenhagen–Malmö. Similar results are also found when only the air masses originating from the southwest are considered (see Fig. 7 and Tables S2–S4). Most likely the contribution from the BC emissions from Copenhagen–Malmö is not apparent when analyzing the whole eBC dataset from 2018 because the SW air masses are generally influenced by more precipitation (more BC wet scavenging) compared to other air masses. The HYSPLIT trajectory simulations for the complete year show that 66 % of all air masses that passed over Malmö–Copenhagen were influenced by >1 mm precipitation within 48 h upwind of the rural site, compared to 39 % for all other air masses. The lowest median eBC concentration is found in air masses from the northwest (101 ng m-3) followed by NE air masses (149.9 ng m-3) and SW air masses (192.0 ng m-3). SE air masses clearly stand out from any other air masses, with a median eBC concentration of 563 ng m-3. This can be partly explained by the low probability of precipitation in SE air masses. Only 23 % of these air masses are influenced by >1 mm precipitation within 48 h upwind of the rural site. However, when also considering the effect of precipitation upwind of Hyltemossa, the SE air masses have a median eBC concentration that is ∼2 times larger than the SW air masses.

A comparison of the 15 d of simultaneous urban and rural measurements is shown in Fig. 8. It is clear that non-refractory PM1 concentration and composition are driven by long-range pollution at both sites, while black carbon is more influenced by local emissions. The organic aerosol at the urban site was clearly dominated by oxygenated organic aerosol (OOA, 78 %) with only a minor contribution of HOA (9 %), suggesting that aged long-range sources dominate even the organic aerosol, for which one could expect a larger difference between the sites owing to the high urban traffic density (Glasius et al., 2011). However, in the original time resolution of 20 s (not shown), the urban AMS data feature transient plumes of traffic exhaust. Yet, as shown in Fig. 8, these plumes have little impact on average concentration. The OA mass spectrum from traffic plumes at the curbside site is shown in Fig. S9. The plume spectra are similar in mass spectral profiles to hydrocarbon-like organic aerosol (HOA), commonly associated with traffic emissions. This OA is likely present in thin coatings on the freshly emitted traffic BC and possibly also externally mixed with the BC and hence is not detected by the SP2 time delay method.

With regards to eBC, Fig. 8 shows that approximately half of the mass at the urban site is due to long-range transport, despite the high traffic intensity (about 30 000 vehicles per day). This can be deduced from the temporal variability and absolute concentrations at both sites. The result is corroborated by the rBC results shown in Fig. 2. Firstly, rBC concentrations are roughly doubled at the urban site compared to the rural site. Secondly, the diurnal pattern observed at the urban site (see Fig. 2) shows that levels between 1:00 and 5:00 LT, during which local traffic density is low, are about half of the average concentration.

Consistent with the high abundance of BC from long-range transport at the urban site, on average 40 % of urban BC particles with a diameter close to the mass GMD were found to be thickly coated based on the SP2 data. We did not directly measure the chemical composition of the thick coatings. However, as shown in Fig. 9, non-refractory PM1 concentration is highly correlated with thickly coated BC fraction (R2=0.87). Indeed, all major chemical species in non-refractory PM1 correlate well with the thickly coated BC fraction (R2= 0.82–0.88). This suggests that the BC coatings are similar in composition to non-refractory PM1 (dominated by secondary material), despite the fact that only a minor fraction of the particles in PM1 contains rBC as discussed above. As non-refractory PM1 is more widely and accurately measured than BC coating composition, such homogeneity simplifies the assessment of BC properties and impacts, for example concerning cloud formation and lung deposition. However, direct measurements of BC coating composition, ideally on a single-particle basis, are needed to support this conjecture.