Emission factors of inorganic ions from road traffic: a case study from the city of Naples (Italy) Angelo Riccio1 , Elena Chianese, Giuseppina Tirimberio 5
Department of Science and Technology, Parthenope University, Centro Direzionale, Isola C4, 80143 Napoli (Italy)
Maria V. Prati Istituto Motori – National Research Council of Italy (IM-CNR), Via Marconi 4, 80125 – Napoli (Italy)
Abstract 10
PM10 samples were collected in the urban tunnel of Naples (southern Italy) during a monitoring campaign on March 2015. Two sets of samples were collected at both sides of the tunnel, each set representing the daily cycle at a 1 hour time resolution. Distance-based – mass per kilometer – emission factors (EFs) were calculated using mass concentrations, traffic flow rates
15
and wind speed as a function of fleet composition. Samples were analyzed for mass and water-soluble inorganic ions (Na+, NH4+ ,K+, Ca2+, Mg2+, Cl–, 2− NO− 3 and SO4 ) with the aim of investigating the influence of road traffic on
the contribution of these species to PM levels. Road traffic directly emits inorganic ions, both from the exhaust and non20
exhaust components. Analysis of ionic composition highlighted the increase in calcium concentration, which may derive from non-exhaust sources (road dust, wear of brake pads, clutches, tires) and calcium sulfonates, phenates or salicylates, often added to motor oils. Sulphate, added to lubricant oils, Email address:
[email protected] (Angelo Riccio) Preprint submitted to Transportation Research Part D: Transport and EnvironmentApril 15, 2017
is also directly emitted at a rate higher than the gaseous sulphur dioxide 25
emission. According to our analysis, nearly 10% of PM mass is composed by watersoluble inorganic ions, most of which directly emitted by automobiles. This suggests that an important contribution to PM emissions may derive from the inorganic component and more efforts should be devoted to constrain
30
these emissions if PM concentration had to effectively comply with air quality standards. Keywords: particulate matter, road traffic, inorganic ions 1. Introduction The effects of particulate matter (PM) on the atmosphere, climate and
35
health are among the central topics in current environmental research (P¨oschl, 2005; Lohmann and Feichter, 2005). Road traffic is one of the major contributors to atmospheric pollution in urban areas (Negral et al., 2008; Pey et al., 2009; Lonati et al., 2011; Riccio et al., 2016). Automobiles both directly emit particles (primary PM) and gaseous precursors, e.g. SO2 , NOx , organic
40
compounds and NH3 , leading to the formation of secondary particles. Direct emissions account for approximately 40-50% of the observed PM mass and gas to particle transformation processes account for the remaining 50–60% (Seinfeld and Pandis, 2006). Despite the constant developments on environmentally friendlier fuels and the reduction of pollutant emissions, 30 to 40%
45
of the European urban population was exposed at concentrations above allowable EU air quality standards for PM from 2000 to 2012 (Guerreiro et al., 2014).
2
To understand the health effects and formation/emission of PM, it is essential to unravel its chemical characteristics and mechanisms of formation. 50
PM characterization from road traffic is often based on the content of heavy metals and carcinogenic components (Adachi and Tainosho, 2004; Amato et al., 2011; Deng et al., 2006; Hjortenkrans et al., 2007; Huang et al., 2012; Johansson et al., 2009; Riccio et al., 2016), but less attention has been devoted to inorganic reactions. Most of the inorganic reactions are complex
55
and competing with each other: ammonia plays a primary role in neutralising atmospheric acids, such as nitric, sulphuric and hydrochloric acid to form ammonium sulfate (NH4)2SO4, ammonium bisulfate (NH4HSO4), ammonium nitrate (NH4NO3), and ammonium chloride (NH4Cl) aerosols (Seinfeld and Pandis, 2006). These reactions depend on ambient air temperature, relative
60
humidity, solar radiation and concentration of the precursor gases to a large extent; moreover, the long-range transport of these salts, joined to the reversible equilibrium with their gaseous precursors, facilitates the transport and release of inorganic salts to areas far from the emission sources. Automobiles greatly influence inorganic chemistry, emitting NOx, NH3
65
and SO2. Nowadays, the introduction of the Euro V fuel standard (< 20 ppmS) drastically reduced SO2 emissions in many European countries, although sulphates are added to lubricant oils, so that road vehicles must be considered as primary emitters for sulphate, too (Stepina and Vesely, 1992). Ammonia is mainly emitted from agriculture practices, which contributes for
70
more than 90% to the total European emission (http://www.eea.europa. eu/data-and-maps/indicators/eea-32-ammonia-nh3-emissions-1/assessment4), but road traffic is the most important non-agricultural source (Sutton
3
et al., 2000; Sapek, 2013). Specifically, the introduction of catalysts has much increased ammonia emissions. For example, it is estimated that gasoline pas75
senger cars may emit up to 40 mg km−1 vec−1 in Italy (http://www.sinanet. isprambiente.it/it/sia-ispra/fetransp/fattori-emissione-trasportostradale/view). Therefore, road traffic is now recognized as an important source of ammonia in the urban environment, and as a result, the production of ammonium salts from this source may have an important role in the
80
build up of particles. Recently, highway tunnel (Emmenegger et al., 2004; Cui et al., 2016; Kean et al., 2009) and chassis dynamometer studies (Durbin et al., 2002; Huai et al., 2003) indicate that ammonia emissions from vehicles may be larger than previously estimated. Livingston et al. (2009) estimated that mobile sources might be responsible for as much as 18% of total NH3
85
emissions in the South Coast Air Basin of California. US EPA concluded that mobile sources contributed 3% of NH3 emissions nationwide in 2014 (EPA, 2015). Road tunnels are excellent locations to measure vehicle emissions as the influence of meteorological conditions is minimized and the emissions undergo
90
‘real world’ dilution effects with minimal interference from other sources, enabling the vehicle emission factors to be determined. More importantly, road tunnel studies provide emissions from a large number of vehicles under ‘real world conditions’, that is, vehicles operating at stable conditions when going through a road tunnel. Tunnel measurements have been frequently used in
95
the past to validate model EFs. Weingartner et al. (1997), Abu-Allaban et al. (2002) and Gertler and Pierson (1996) describe the methodology for calculation of emission factors using tunnel measurements. Pant and Harrison
4
(2013) reviewed all these methods and their applications. Unfortunately, the emission of inorganic ions in particles and their precur100
sor gases from vehicle exhaust are rarely measured and estimated. To better characterize the inorganic ion content from automobiles and their emission factors, a sampling campaign was performed in an urban tunnel in the city of Naples, with particulate matter samplers, meteorological instrumentation and traffic counters. This paper presents data on the inorganic content of
105
atmospheric aerosols sampled at the two sides of an urban tunnel in Naples (South Italy) and the corresponding emission factors (EFs). 2. Experimentals 2.1. Sampling The ‘4 giornate’ tunnel is a main thoroughfare connecting two densely
110
populated districts in the urban area of Naples and has a length of about 800 m. The number of vehicles passing through the tunnel is relatively high, with more than 25,000 vehicles crossing the tunnel per day. The speed limit is 50 km/h with a traffic light at the beginning of the tunnel. During the time in which measurements have taken place, the traffic in the tunnel was
115
unidirectional with two lanes in a single hole (a second separate hole operates in the opposite direction). The lane roadway shows different gradients, a rise of 2.6% in the first 400 m after the entrance, then a section with almost flat conditions up to the exit. The ventilation system was not active during the measuring campaign, so that the fresh air was mainly influenced by the
120
‘piston effect’ along the driving direction. PM10 (particulate matter with an aerodynamic diameter less than 10 µm) 5
samples were collected at both sides of the tunnel using two inter-calibrated low-volume gravimetric samplers (Tecora Bravo-H and Tecora Echo models) equipped with PM10 inlets. Filters (47 mm borosilicate glass Emfab filters 125
TX40HI20 by PALL Life Sciences) were collected during a two days campaign on 25-26 March, 2015, at a 1 hour time resolution, as already described in Riccio et al. (2016). The first day was devoted to instrument inter-calibration and the second day to the actual collection of samples. Before and after sampling, filters were pre-conditioned in a box with constant humidity (50%) and
130
temperature (20 °C) for 24 hours, electrostatically neutralized and weighted with a micro balance to determine the deposited aerosol mass, according to the EN 12341:2014 standard (CEN/TC 264). A mobile laboratory was equipped with a video camera for characterizing the traffic flow as the number of passenger cars, mopeds, motorcycles, light-
135
duty vehicles (LDVs), heavy-duty vehicles (HDVs) and buses/coaches. An ultrasonic anemometer (Delta Ohm HD2003.1) was used to measure wind speed along the tunnel axis. 2.2. Chemical analysis Each filter was analysed regarding its inorganic ions chemical composi-
140
tion. Ions were extracted in 15 ml ultra-pure water (≥ 18.2 MΩ resistivity) using a micro-wave assisted extraction technique (Sathrugnan and Balasubramanian, 2005). The water-soluble ionic composition (Na+, NH4+ , K+, Mg2+ , Ca2+, Cl–, Br–, SO42–, PO43–, NO2–, NO3– and several organic acids) was determined by ion chromatography (IC) with an ICS-1100 Dionex sys-
145
tem, using an injection volume of 25 µl in a Dionex IonPac CS12A column and a cation self-regenerating suppressor CSR-ULTRA II (2 mm) for cations; 6
methanesulfonic acid was used as eluent at a concentration of 20 mM and a flow rate of 0.25 ml/min. For anions, an injection volume of 25 µl in a Dionex IonPac AS14A column and an anion self-regenerating suppressor ASRS-300 150
(4 mm) was used. A 8:1 mM carbonate/bicarbonate buffer solution at a flow rate of 1 ml/min was used as eluent. Before the analysis, calibration plots were made for anions and cations, using single- and multi-component standard solutions for IC. In addition to the ionic chemical analysis, the sea-salt contribution to PM mass was estimated, according to standard methods. Typically Na+ is used to determine the sea-salt component in the aerosol phase since the water-soluble Na+ concentration is assumed to originate solely from seawater. Specifically, sea-salt aerosol concentration was calculated as: sea salt(µg/m3 ) = Cl– + 1.47 × Na+ (µg/m3 )
(1)
where 1.47 is the seawater ratio of (Na+ + K+ + Mg2+ + Ca2+ + SO42– + 155
HCO3–)/Na+ (Lewis and Schwartz, 2013; Millero, 2006). This approach prevents the inclusion of non-sea-salt (K+, Mg2+, Ca2+, SO42– and HCO3–) in the sea-salt mass and allows for the loss of Cl– mass through Cl– depletion processes. This approach is based on the assumption that all measured Na+ derives from seawater; results of Savoie and Prospero (1980) indicate that
160
soil dust has a minimal contribution to measured soluble sodium concentrations. Non sea-salt (nss) ionic contributions were calculated from Na+ concentrations and the ratio of corresponding ion to sodium in seawater.
7
3. Methods Given the concentrations measured at the two sides of the tunnel, from a straightforward mass balance it is possible to estimate the emission factors (EFs). EFs are functional relations that predict the quantity of an emitted pollutant as a function of the activity causing the emission. In this work the calculation is based on the following well known equation (Abu-Allaban et al., 2002; Amato et al., 2012): ∆ci =
L X EFj Nj vA j
(2)
∆ci , the increase in concentration between the tunnel exit and entrance for 165
the ith measurement and Nj the number of vehicles for the jth category; v the velocity of air, A the tunnel cross section. EFj can be interpreted as the quantity of a pollutant emitted per distance driven by the jth vehicle category. The sum is over all vehicle categories included in the estimation. 4. Results
170
4.1. Diurnal variation of traffic density profiles The vehicle classification (based on video observation) allowed to determine the exact vehicle fleet composition during the monitoring campaign. Automatic radar detectors allowed to measure vehicle speeds, which were approximately constant all along the day, at a value of about 50 km/h. Fig-
175
ure 1 shows the traffic density profiles for two-wheels vehicles, passenger cars, LDVs and the sum of HDVs, urban buses and coaches. During the day, 4928 mopeds/motorcycles, 13195 passenger cars, 674 delivery vans, 275 heavy duty
8
vehicles and 50 coaches/urban buses passed through the tunnel. Diurnal patterns of vehicles exhibited a rather steady level, with a barely visible peak in 180
the early afternoon, corresponding to schools and offices closures. According to the official statistics (http://www.aci.it/laci/studi-e-ricerche/ dati-e-statistiche/autoritratto/autoritratto-2015.html, data of 2015), in the Naples province about 77% of the vehicle fleet is composed by passenger cars (66% gasoline- and 34% diesel-powered), 14% by motorcycles
185
and about 6% by delivery vans. Seventy percent of the gasoline vehicles are equipped with catalytic converters (3-way catalysts); 34% of these have engine capacities between 1400 and 2000 cm3 and only 4% higher than 2000 cm3 . The commercial vehicles (diesel and gasoline powered) represent 7% of the fleet, among which 5.4% are Light Duty and 1.6% Heavy Duty Vehi-
190
cles. It may be noted that, compared to the average fleet composition in the Naples province, a larger proportion of vehicles passing through the tunnel is represented by mopeds and motorcycles, and a lower percentage by passenger cars. 4.2. PM10 and ionic components
195
Table 1 lists the average ± standard deviation (SD), median, minima and maxima concentration of each ionic species determined at the tunnel entrance and exit. Concentrations at the entrance are in general agreement with other aerosol studies in Naples (Riccio et al., 2014) and in the same range of levels reported for other European urban (Van Dingenen et al., 2004) and Mediter-
200
ranean coastal sites (Perrone et al., 2011; Nicolas et al., 2009; Lazaridis et al., 2008) during high pollution events; PM10 concentration remained approximately constant throughout the day at the tunnel entrance, at a level of 9
45
two-wheels pass. cars
40
LDV HDV+CO+UB
35
vehicles/min
30 25 20 15 10 5 9
10
11
12
13
14 Time
15
16
17
18
19
Figure 1: Diurnal variation of traffic density profiles, at a 1 hour time resolution, for twowheels mopeds and motorcycles, passenger cars, light-duty vehicles (LDVs), and the sum of light-duty vehicles (LDVs), coaches (CO) and urban buses (UB)
about 85 µg/m3 , a relatively high concentration, since the reference level is 50 µg/m3 , not to be exceeded for more than 35 days per year (Air Quality 205
Directive 1999/30/EC). Particle mass concentrations at the tunnel exit were significantly elevated relative to entrance concentrations during the whole experimental campaign; depending on the hour of the day, PM10 concentration ranged between 300 and 600 µg/m3 . Inorganic anions included NO2–, NO3–, Br–, PO3− 4 and oxalate; however
210
these constituents were not detected or with values very close to the detection limit and were therefore excluded from further analysis. Of the organic anions analysed, acetate and formate concentrations appeared to be impacted 10
by occasional contamination and close to their limit of detection; thus also acetate and formate were not considered further. 215
About 30% of the total PM10 mass is explained by ionic elements at the tunnel entrance, predominantly sodium chloride, calcium and sulphate. Total Na+ concentrations varied from 1.0 to 6.5 µg/m3 and averaged 4.2 ± 1.4 µg/m3 , with no significant variation at the tunnel exit. Under the assumption that sodium originates solely from seawater, the resulting sea-salt concentra-
220
tion varies from 7.0 to 16.0 µg/m3 and a mean level of 10.7 ± 3.0 µg/m3 , explaining about 10% of the total PM10 mass. Table 1 shows that the mean Cl–/Na+ mass ratio is close to 1.0, i.e. lower than that expected for fresh sea-salt particles (1.8). Cl– depletion effects have already been observed to occur at Mediterranean sites; Ko¸cak et al. (2007)
225
found at Erdemli in the northeastern Mediterranean that mean Cl–/Na+ mass ratio was 0.7 and 1.7 in the fine and coarse fraction, respectively, while Nicolas et al. (2009) found at Elche, in southwestern Mediterranean, a mean Cl–/Na+ mass ratio of 0.7 and 0.6 in the fine and coarse fraction, respectively. The lack of Cl– may be attributable to the desorption of hydrochloric acid,
230
during the aerosol aging process, due to sulfate (or nitrate) substitution (Song and Carmichael, 1999): H2SO4 + 2 NaCl HNO3 + NaCl
Na2SO4(s,l)↓ + 2 HCl↑ NaNO3(s,l)↓ + HCl↑
These reactions might be responsible for the Cl− disappearance, enriching particles by sodium. Whereas HNO3 displacement has been well demonstrated to dominate chloride depletion for coarse mode sea salt particles, the 11
Table 1: Mean, median, minimum, maximum and standard deviations (SD) of atmospheric PM10 and ion concentrations collected at the tunnel entrance and exit. Concentrations are in µg/m3 . ‘lod’ stands for ‘limit of detection’. Charge balance stands for the ratio of positive to negative equivalent charges. The ‘ss’ prefix stands for sea-salt and ‘nss’ for non sea-salt. Components for which the one-tailed t-test for the null hypothesis (concentration at the tunnel exit greater than that at the entrance) cannot be rejected at the 5% significance level are highlighted in bold.
tunnel entrance mean ± SD median min
tunnel exit max
mean ± SD median min
max
PM10
86.3 ± 16.6 84.9
61.7 115.9
411 ± 113
407
279
567
Na+ NH+ 4 K+ Mg2+ Ca2+
4.2 ± 1.4 < lod 2.4 ± 0.9 0.4 ± 0.1 9.0 ± 1.3
4.0 < lod 2.3 0.4 8.2
1.0 < lod 1.1 0.3 8.0
6.5 < lod 3.9 0.6 11.4
3.3 ± 1.3 2.2 ± 1.3 1.8 ± 0.6 0.9 ± 0.2 27.4 ± 4.8
3.3 2.5 1.9 0.9 27.6
1.9 < lod 0.9 0.5 20.6
5.5 4.2 3.0 1.1 34.7
F− CH3 COO− HCOO− Cl− SO2− 4
0.4 ± 0.3 1.0 ± 0.5 2.4 ± 2.3 4.5 ± 1.5 5.6 ± 1.9
0.4 0.8 1.1 4.6 5.9
< lod < lod < lod 3.0 2.7
0.7 1.8 5.0 7.4 9.0
0.5 ± 0.3 1.7 ± 2.2 1.0 ± 1.3 5.7 ± 2.8 8.0 ± 1.6
0.3 < lod 0.3 5.0 8.2
< lod < lod < lod 3.5 5.6
1.1 4.2 2.5 12.8 10.8
Charge balance 2.9 ± 0.7
2.8
1.9
4.4
5.3 ± 1.2
5.0
3.8
6.9
Mg2+ /Na+
0.3 ± 0.1
0.3
0.2
0.7
0.6 ± 0.2
0.5
0.3
1.1
sea salt ssSO2− 4 nssSO2− 4 nssCa2+ nssMg2+ nssK+
10.7 ± 3.0 1.1 ± 0.4 4.2 ± 1.6 8.9 ± 1.3 0.8 ± 0.2 2.2 ± 0.9
10.0 1.0 4.9 8.1 0.8 2.1
7.0 0.5 1.7 7.8 0.6 1.0
16.0 0.6 5.8 11.7 1.3 3.7
10.6 ± 4.4 0.8 ± 0.3 7.1 ± 1.4 27.3 ± 4.8 1.2 ± 0.2 1.7 ± 0.6
9.0 0.8 7.4 27.4 1.2 1.8
5.2 0.3 5.3 20.5 1.0 0.8
20.6 1.4 9.4 34.5 1.5 2.8
12
235
predominant mechanism of fine mode chloride depletion has been mostly attributed to sulfate substitution (Hara et al., 2002; Johansen and Hoffmann, 2004; Li et al., 2011). Magnesium is slightly enriched with respect to the bulk seawater, with a Mg2+/Na+ mass ratio of 0.3 at the entrance and 0.6 at the tunnel exit,
240
greater than the bulk seawater one (0.12). Ca2+ and K+ are much more enriched with respect to their expected seawater ratio, because well known additional sources, e.g. biomass burning for potassium and crustal erosion for calcium, contribute to their concentrations (Pio et al., 2008). Ammonium is below its detection limit at the tunnel entrance but achieves
245
an average concentration of 2.2 ± 1.3 µg/m3 at the tunnel exit. This is because of vehicles equipped with catalytic converters, generating ammonia through the reaction taking place in the converter between NO and H, particularly when the air to fuel ratio in the combustion chamber is lower than the stoichiometric value (Franco et al., 2013).
250
Sulphate is the most represented anion in PM10 samples. Average SO2− 4 concentrations observed at the tunnel entrance, 5.6 ± 1.9 µg/m3 , are of the same order of magnitude as reported in other investigations performed in Southern Italy coastal sites during high pollution episodes (Amodio et al., 2008) and in other tunnel studies (Allen et al., 2001; Gillies et al., 2001;
255
Laschober et al., 2004; Cheung et al., 2009). Sulphate can be decomposed into its sea-salt (ss) and non sea-salt (nss) contributions. ssSO2− 4 , a natural component of seawater salts, explains about one fifth of total sulphate, and, as expected, remains approximately constant at the tunnel exit; vice versa, nssSO2− 4 almost doubles its concentration at the
13
260
tunnel exit. nssSO2− 4 at the tunnel entrance, deriving from the oxidation of anthropogenically emitted SO2 and of biogenically emitted dimethylsulfide, represents the predominant fraction of total sulphate. The mean molar ratio of NH4+ to nssSO2− 4 at the tunnel exit average 2.0± 0.7, indicating complete neutralization between sulphate and ammonium. More formally, to assess the availability of ammonia for neutralization of acidic components (e.g. H2SO4 and HNO3) in the tunnel atmosphere, the ammonium availability index was estimated on the basis of hourly data. The ammonium availability index, J, was defined as the molar ratio of the observed ammonium cation concentration to the amount needed to neutralize the SO2− 4 and nitrate anion concentration, and is expressed as: J=
[NH4+] – 2 × [nssSO2− 4 ] + [NO3 ]
(3)
For J < 1, there is an ammonium deficit; for J > 1, there is an ammonium surplus. A similar index was also defined by Adams et al. (1999) and Chu 265
(2004) in their modeling studies and it was referred to as the degree of neutralization (DON). Nitrate concentration was negligible during the sampling period, so that J was determined solely by ammonium and sulphate concentration. As already noted by Chu (2004), J is only an indicator for the acidity of particles, since it does not exactly correspond to the pH value of
270
the observed ambient aerosols as a whole. Other species such as calcium, sodium, and various organic acids could also contribute to the aerosol pH value. The mean value of J was 1.0 ± 0.3, and even if total sulphate in taken into consideration the mean value of J is 0.9 ± 0.3, indicating an almost
275
complete neutralization between excess sulphate produced by automobile 14
emissions and ammonium. This is more clearly examined in Figure 2, where the ammonium concentration in PM10 vs. stoichiometric nss-sulphate has been reported. Figure 2 shows that samples are close to the 1:1 line in several cases and that average deviations from the 1:1 line are ±15%, indicating that 280
most of the time there is sufficient NH3 for the complete neutralization of sulphate, i.e. to form NH4(SO4)2 as much as possible.
300
3 2 × nss-[SO2− 4 ] (nmol/m )
250
200
150
100
50
0 0
50
100
150 200 3 [NH+ 4 ] (nmol/m )
250
300
Figure 2: nss-sulphate against equivalent ammonium at the tunnel exit. The blue dashed line is the linear best fit. The solid black line is the 1:1 line.
The charge balance, i.e. the sum of Na+, K+, Mg2+, Ca2+ and NH4+ against the sum of Cl–, Br– and SO42– at the tunnel exit is shown in Figure 3 (left panel). The anion and cation microequivalents are calculated from the corre285
sponding concentrations, taking into account for the number of ion charges. The discrete correlation (R2 = 0.84) between cation and anion equivalents for all samples indicates that the five cations and three anions were the 15
2.5
2.0
calcium microequivalents/m3
cation microequivalents/m3
2.0
1.5
1.0
0.5
0.0 0.0
0.5
1.0 1.5 anion microequivalents/m3
2.0
1.5
1.0
0.5
0.0 0.0
2.5
0.5 1.0 1.5 excess charge microequivalents/m3
2.0
Figure 3: Charge balance between positive and negative ions at the tunnel exit (left) and calcium equivalent concentration against excess charges at the tunnel exit (right). The blue dashed line is the linear best fit. The solid black line is the 1:1 line.
major ions extracted from the filters. As found in other studies (Oliveira et al., 2010; Shen et al., 2008), the charge balance is far from the 1:1 line, 290
indicating an excess of cations in all conditions. This unbalance also holds at the tunnel entrance, albeit to a lesser extent. The explanation for the deficit of negative charges is due to the presence of undetected carbonates and other organic anions (although the contribution of water-soluble organic acids have been shown to be negligible in this study). Carbonates were not
295
determined from PM samples, but an evidence for the presence of carbonates is shown on the right panel in Figure 3, where the equivalent calcium concentration is reported against excess charges; all points are close to the 1:1 line (R2 = 0.97) , indicating that calcium is almost completely balanced by excess charges, i.e. carbonates. Therefore, the most likely explanation for
300
− the anion deficit is that a significant amount of CO2− 3 and HCO3 exists in − the aerosol phase. CO2− 3 and HCO3 are important atmospheric constituents
16
because atmospheric chemistry and aerosol characteristics are affected by their presence, favoring the uptake and conversion of SO2 to SO2− 4 and NOx to NO− 3 , as well as removal of HNO3 and H2SO4 from the gas phase (Den305
tener et al., 1996). The buffering capacity of CO2− 3 has been demonstrated for precipitation in China and more generally for the northern hemisphere (Dianwu et al., 1988; Sequeira, 1993). More to the point, based on results in Figure 3, assuming that the predominant carbonate form is CO2− 3 , it accounts for about 3.5% (≈ 15 µg/m3 ) of PM10 mass at the tunnel exit and
310
about 6% (≈ 6 µg/m3 ) at the tunnel entrance. Similar results has also been obtained worldwide (Cao et al., 2005; Davis and Jixiang, 2000; Querol et al., 2008). Calcium is the most abundant ion at the tunnel entrance, with a mean value of 9.0 ± 1.3 µg/m3 and a negligible contribution from sea salts. Cal-
315
cium may derive from several non-exhaust sources; primarily, from friction processes by tires on the road surface, brakes and re-suspension of particles due to traffic-induced turbulence (Abu-Allaban et al., 2003; Bukowiecki et al., 2010; Thorpe and Harrison, 2008; Amato et al., 2012), since it is a major component of earth crust (Mason and Moore, 1982), and secondly
320
because it is added in motor oils as calcium sulfonates, phenates and salicylates for their detergent properties and as a filler in brake pads (Willermet, 1998; Yamaguchi et al., 2002). One striking feature of Table 1 is the enrichment in calcium at the tunnel exit, achieving a level of 27.4 ± 4.8 µg/m3 , i.e. three times the entrance level, pointing to the importance of road traffic in
325
determining its concentration into particulate mass.
17
4.3. Emission factors Given the vehicle fleet composition, wind speed and tunnel geometry, it is possible to apply equation (2) for the estimation of emission factors for major ions (calcium, sulphate and ammonium, whose concentration at the tunnel exit is significantly higher than that at the entrance), re-written as: ∆c =
NL EF vA
(4)
where the increase in concentration has been expressed as a function of a fleetaveraged emission factor. Given the observed wind speed (v), the number of vehicles passing through the tunnel (N) and the length (L) and cross-section 330
area (A) of the tunnel, equation (4) regresses NL/vA onto ∆c and the regression coefficient is the fleet-averaged emission factor. We used the DREAM software (Vrugt et al., 2009) to obtain a posterior estimation of emission factors in the light of experimental data, in a manner similar to that described in Riccio et al. (2016). The estimated emission factors, together with their
335
inter-quartile ranges, are shown in Table 2; for the sake of comparison, the emission factor for PM10 is also shown. The most abundantly emitted ion is calcium, explaining about 4% of total PM10 emission, highlighting the importance of non-exhaust sources, probably related to road abrasion since calcium-containing minerals (e.g. plagioclase
340
feldspars and carbonates) are major constituents of the Earth’s crust (Mason and Moore, 1982). Sulphate and ammonium are also largely emitted; the average mass emission rate of particulate ammonium is 1.37 mg/km, which is a factor three times smaller than found in the Caldecott tunnel (Allen et al., 2001) but 18
Table 2: Fleet-averaged emission factors and corresponding IQR (inter-quartile range) for Ca2+, SO42– and NH4+
Ion
345
emission factor
IQR
units
Ca2+
8.3
[7.9, 8.7] mg/km
SO42–
3.0
[2.8, 3.2] mg/km
NH4+
1.37
[1.31, 1.44] mg/km
PM10
190
[182, 198] mg/km
one-third higher than the emission rates determined by Gillies et al. (2001). Deviations in the emission rate of ammonium are expected since ammonia emissions depend on the car fleet composition. In the present study only particulate phase ammonia is determined, whereas total ammonia emissions (in the gas phase) from dynamometer studies (Kean et al., 2009) indicate
350
emission rates as high as 30 mg/km. The average sulphate emission rate was 3.0 mg/km, a value that accords well with the emission rates found in the Caldecott (Allen et al., 2001) and Sepulveda (Gillies et al., 2001) tunnel. In atmosphere, sulphate and nitrate both compete for reacting with ammonia to form ammonium sulphate and ammonium nitrate, but the reaction
355
rate between NH3 and H2SO4 (reaction rate constant equal to 1.5×10−4 sec−1 , (Seinfeld and Pandis, 2006)) is much faster. From Figure 2 and Table 2, it is shown that excess ammonia quickly combines with sulphate, resulting in an equivalent ammonium emission approximately twice as that of sulphate (76 µmol/km and 31 µmol/km, respectively). Due to the high number of
360
vehicles, we expect that the excess of ammonia were available to combine 19
with nitrate to form NH4 NO3 , but, unlike sulphate, nitrate was found at a negligible concentration in the tunnel atmosphere (see Table 1); the lack of nitrate con be explained by the fact that it forms from the oxidation of NO2 with ozone or hydroxyl radical, but both reactions are negligible in the 365
tunnel atmosphere, rich in ozone-depleting nitric oxide. For all other elutable ions no significant difference between tunnel exit and entrance concentrations was found, preventing the calculation of emission factors. 5. Discussion
370
Measurements in traffic tunnels have been used successfully for quantifying emissions of a large number of both gaseous and particulate pollutants (see El-Fadel and Hashisho (2001) for a review). In this paper, real-world emission factors for ionic components in PM10 from road traffic have been estimated. This was achieved by a mass balance exercise from measurements
375
at the two sides of an urban tunnel in the city of Naples, sampling particulate mass from both exhaust and non-exhaust emissions. It is important to bear in mind that factors such as variations in speed, aerodynamic conditions in the tunnel (the so-called ‘piston effect’), fleet characteristics, i.e. proportion of heavy- (HDVs) and light-duty vehicles (LDVs),
380
vehicle load and the unknown proportion of vehicles in cold-start mode, cause uncertainties in measurements (He et al., 2008). In addition, vehicles in a tunnel often drive at a steady speed which does not happen under other urban conditions where traffic follows a stop-and-go pattern (El-Fadel and Hashisho, 2001; Gertler et al., 2002), and resuspension of road dust, an im20
385
portant contributor to PM emissions, depends on parameters like humidity or cleanness of the roadway (Sternbeck et al., 2002; Vardoulakis et al., 2003), factors which are usually highly variable. Despite all these complications, tunnel studies yield valuable data regarding the actual emission behavior of real-world road vehicles. From the
390
analysis of ions in PM10 samples, we found significant emissions for ammonium, calcium, and sulphate. The t-test (Table 1) confirms that the average ammonium concentration is significantly higher at the the tunnel exit, due to ammonia emissions from vehicles, especially gasoline passenger cars, which quickly combines with available sulphate. An increase of ammonia emissions
395
has been reported both in USA (Fraser and Cass, 1998; Kean et al., 2000) and European (Perrino et al., 2002) experimental studies. Recently, Wu et al. (2016) argued that ammonia emission in China plays a determinant role in the formation of secondary inorganic aerosols. As far as Italy is concerned, estimates from the Italian Agency for the Protection of the Environment in-
400
dicate that ammonia emission from road transports increased from 1618 tons (less than 0.4%) in 1993, when catalytic converters were introduced in Italy, to 19796 tons in 2000 (about 4.3% of total NH3 emissions) (ISPRA, 2016). Agriculture practices and natural sources emit almost all of ammonia (about 95% in 2000). According to the latest European Environment Agency
405
report (EEA, 2016), NH3 is the only main air pollutant whose emissions increased in the EU-28 (by 0.9%) between 2013 and 2014. Also, ammonia from road transport has fallen since 2000, and is projected to fall in the future as the second generation of catalysts, which emit lower levels of NH3 than the first generation catalysts, penetrate the vehicle fleet. In 2013 ammonia
21
410
emissions from road transport has been estimated at 7097 tons (less than 2% of total emissions) in Italy (ISPRA, 2016), still remaining road transport the most important emitting sector after agricultural practices (and the most important in urban areas). The higher nss-sulphate concentration at the end the tunnel may be ex-
415
plained by sulphate in lubricant oils or sulphur content in fuels which is converted to sulphate by oxidation treatments (Stepina and Vesely, 1992). It is interesting to compare our results with official SOx emissions. Sulphur oxides emissions from road traffic steadily decreased from 1990 to 2014 (from ≈ 370 mg/km to ≈ 1 mg/km, respectively), largely due to the more stringent
420
sulphur abatement requirements (http://www.sinanet.isprambiente.it/ it/sia-ispra/fetransp/fattori-emissione-trasporto-stradale/view and http://www.sinanet.isprambiente.it/it/sia-ispra/serie-storicheemissioni/dati-trasporto-strada/view). Nowadays, sulphur oxides are mainly emitted from combustion processes in energy production, manufac-
425
turing industries, production processes and harbor activities: ≈ 45, 33, 31 and 22 ktons in 2014 in Italy, respectively. As shown in Tables 1 and 2, sulphate concentration almost doubles at the tunnel exit, corresponding to an emission factor of about 3.0 mg/km, i.e. a value three times higher than the direct gaseous sulphur emissions. Even if gaseous emissions were further
430
reduced, this direct sulphate emission should represent an unabated lower limit. As long as sulphate is emitted, ammonia combines with sulphate to produce particulate mass. From emission factors in Table 2, it can be estimated a direct production of about 4.4 mg/km of (NH4)2SO4. Results in Tables 1 and 2 also highlights the importance of non-exhaust
22
435
sources. This is made very clear by the high calcium emission rate and by the charge imbalance, due to undetected carbonates mainly emitted by road abrasion. From the sea-salt budget (equation 1), it was possible to demonstrate that the predominant component of water soluble calcium concentration does not come from seawater; even if all soluble sodium were assumed
440
to derive from seawater, nearly all calcium concentration can be ascribed to nssCa2+ (see Table 1). Calcium may derive from different sources: first it is a major natural component of crustal earth (the upper crustal composition comprises about 3.6% of calcium oxide, Mason and Moore (1982) and Rudnick and Gao (2003)), but it is also added in motor oils as calcium
445
sulfonates, phenates and salicylates for their detergent properties (Willermet, 1998; Yamaguchi et al., 2002); moreover, calcite (CaCO3 ) is used as a filler in brake pads because it imparts heat stability to the friction material, thereby improving the brake pads’ properties (Eriksson et al., 2002; Chan and Stachowiak, 2004). Modern brake lining materials are composites of many
450
different ingredients, which are responsible for differing wear characteristics, physical, and chemical properties of both the bulk lining material and the emitted particles produced during the braking process (Eriksson et al., 1999; Chan and Stachowiak, 2004). Unfortunately, brakes and tyres vary considerably in their formulation and each manufacturer uses a different composition
455
which makes it difficult to predict brake dust composition from the entire vehicle fleet. According to literature data, calcium concentration from brake dust ranges in the interval 103 − 104 mg/kg (Garg et al., 2000; Thorpe and Harrison, 2008). It is important to stress that dust from brake’s wear introduces many
23
460
other hazardous elements (As, Cd, Sb, among others) (Garg et al., 2000; Sternbeck et al., 2002), so that large calcium emissions are presumably related to harmful conditions. Rexeis and Hausberger (2009) predicted, using a detailed emission model for the Austrian fleet, that the percentage of nonexhaust emissions would increase from about 50% between 2005 and 2010
465
up to some 80-90% by 2020. Presently, it has been estimated that nonexhaust emissions may contribute up to 50% of local road traffic emissions (Bukowiecki et al., 2010). In Scandinavian countries, the contribution in early spring may be even up to 90% of PM10 along roads, due to the use of studded tyres (Hussein et al., 2008).
470
6. Conclusions In the past many efforts have been made to reduce PM emissions: more stringent regulations, technological and non-technological upgrades, e.g. congestion charge schemes, park-and-ride systems, restricted-traffic zones and car sharing initiatives. Despite these efforts, road traffic still provides a sig-
475
nificant contribution to airborne PM concentrations. A large number of studies have been devoted to the characterization of trace elements contained in PM particle, e.g. PAHs and heavy metals, but little attention has been paid to the direct emission of inorganic ions. In this work we show that, according to measured values, PM10 emissions are
480
surely relevant, both from exhaust and non exhaust emissions and that a large proportion of PM mass is due to inorganic compounds (mainly sulphate, nitrate, ammonium and calcium), part of which directly emitted from exhaust components. In this study we are not able to distinguish between the 24
contribution of exhaust and non-exhaust emissions, but we conjecture that a 485
relevant proportion of calcium (and consequentially of PM) emissions come from the unregulated non-exhaust fraction, e.g. from road dust abrasion and resuspension. Also, cars directly emit sulphate, contributing to the formation of ammonium sulphate, probably at a rate higher than the emission of gaseous sulphur oxides.
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In the near future, non-exhaust emissions (road dust resuspension, road surface abrasion, brake, clutch and tyre wear) are expected to dominate the remaining traffic-related PM emissions, mostly due to the steep decrease in exhaust emissions, which are heavily regulated, while non-exhaust emissions are not yet affected by regulation standards (Denier van der Gon et al., 2013).
495
Since PM is still heavily emitted by road traffic, it is important continuing to monitor if its concentration effectively comply with air quality standards.
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