Road traffic emission factors for heavy metals

ARTICLE IN PRESS Atmospheric Environment xxx (2008) 1–8

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Road traffic emission factors for heavy metals Christer Johansson a, b, *, Michael Norman a, Lars Burman a a b

Stockholm Environment and Health Administration, Box 8136, S-10420 Stockholm, Sweden Department of Applied Environmental Science, Stockholm University, S-106 91 Stockholm, Sweden

a r t i c l e i n f o

a b s t r a c t

Article history: Received 24 January 2008 Received in revised form 13 October 2008 Accepted 15 October 2008

Quantifying the emissions and concentrations of heavy metals in urban air is a prerequisite for assessing their health effects. In this paper a combination of measurements and modelling is used to assess the contribution from road traffic emissions. Concentrations of particulate heavy metals in air were measured simultaneously during 1 year at a densely trafficked street and at an urban background site in Stockholm, Sweden. Annual mean concentrations of cadmium were 50 times lower than the EU directive and for nickel and arsenic concentrations were 10 and six times lower, respectively. More than a factor of two higher concentrations was in general observed at the street in comparison to roof levels indicating the strong influence from local road traffic emissions. The only compound with a significantly decreasing trend in the urban background was Pb with 9.1 ng m3 in 1995/96 compared to 3.4 ng m3 2003/04. This is likely due to decreased emissions from wear of brake linings and reduced emissions due to oil and coal combustion in central Europe. Total road traffic emission factors for heavy metals were estimated using parallel measurements of NOx concentrations and knowledge of NOx emission factors. In general, the emission factors for the street were higher than reported in road tunnel measurements. This could partly be due to different driving conditions, since especially for metals which are mainly emitted from brake wear, more stop and go driving in the street compared to in road tunnels is likely to increase emissions. Total emissions were compared with exhaust emissions, obtained from the COPERT model and brake wear emissions based on an earlier study in Stockholm. For Cu, Ni and Zn the sum of brake wear and exhaust emissions agreed very well with estimated total emission factors in this study. More than 90% of the road traffic emissions of Cu were due to brake wear. For Ni more than 80% is estimated to be due to exhaust emissions and for Zn around 40% of road traffic emissions are estimated to be due to exhaust emissions. Pb is also mainly due to exhaust emissions (90%); a fuel Pb content of only 0.5 mg L1 would give similar emission factor as that based on the concentration increment at the street. This is the first study using simultaneous measurements of heavy metals at street and roof enabling calculations of emission factors using a tracer technique. Ó 2008 Elsevier Ltd. All rights reserved.

Keywords: Copper Lead Cadmium Nickel Chromium Antimony PM10 Urban aerosol

1. Introduction Many studies have documented the potential toxicity of heavy metals on airborne particulate matter (see e.g. recent review by Schlesinger et al., 2006). Even though there are several studies of the concentrations in urban air (e.g. Dongarra et al., 2007; Fang et al., 2005), quantitative knowledge of source contributions is lacking. More detailed investigations, both measurements and air quality dispersion modelling are needed in order to compare the levels to the new directive regulating heavy metals in air (2004/ 107/EG) and assess population exposure and health effects. The

* Corresponding author. Stockholm Environment and Health Administration, Box 8136, S-10420 Stockholm, Sweden. Tel.: þ46 709 383 086. E-mail address: [email protected] (C. Johansson).

sources include vehicle components, pavement material, road equipment, road maintenance activities (e.g. de Miguel et al., 1997; Moreno et al., 2006; Shah et al., 2006; Dongarra et al., 2007; Querol et al., 2007; Hjortenkrans et al., 2007). In some cities point sources such as steel plants and other industrial activities, foundries, waste incinerating plants and energy production plants can make a substantial contribution as well (e.g. Espinosa et al., 2004; Shah et al., 2006; Querol et al., 2007). But in the centre of urban areas road traffic is likely to make the most important contribution to people’s exposure. Emissions from road traffic occur at ground level along streets in the most densely populated parts of the cities. They are due to wear of brakes, tires, other vehicle components, road pavement and also due to exhaust emissions. Corrosion of metallic material alongside streets may also contribute. In this paper we estimate real world traffic emission factors for a number of metals using NOx as a tracer. We also calculate the

1352-2310/$ – see front matter Ó 2008 Elsevier Ltd. All rights reserved. doi:10.1016/j.atmosenv.2008.10.024

Please cite this article in press as: Johansson, C., et al., Road traffic emission factors for heavy metals, Atmospheric Environment (2008), doi:10.1016/j.atmosenv.2008.10.024

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C. Johansson et al. / Atmospheric Environment xxx (2008) 1–8

levels using a CFD model for this site (Gidhagen et al., 2004). For this study 1.4 g NOx veh1 km1 was used. NOx concentrations at the street and at urban background were measured during the same period as the sampling of heavy metals using commercial chemiluminescence analysers (Thermo Electron). The total road traffic emissions in Stockholm were estimated by multiplying the emission factors by the total road traffic transports (vehicle kilometres). Information on traffic flows, vehicle types, etc., was obtained from the emission inventory of the Regional Air Quality Management Association of Stockholm and Uppsala (Johansson et al., 1999; http://www.slb.nu/lvf/). The inventory includes some 20,000 road links and an annual traffic volume of 12,500 million vehicle km’s. Based on the emission factors road traffic contributions to the total urban background levels were estimated using a Gaussian air quality dispersion model (SMHI, Airviro; http://airviro.smhi.se/ iairviro) (see Johansson et al., 2007). Meteorological conditions were based on a climatology that was created from 10 years of meteorological measurements (15 min averages) in a 50 m high mast located in the southern part of Stockholm. The wind field for the whole model domain was calculated based on the concept first described by Danard, 1977. The dispersion calculations were performed on a 100 m resolution (122,500 receptor points). Effects of the dispersion of individual buildings are treated using a roughness parameter, i.e. they are not resolved.

local road traffic contribution and compare with all other source contributions based on the measured total concentrations. 2. Methodology Parallel filter (Zefluor) sampling of particulate heavy metals was made in a densely trafficked street canyon (Hornsgatan) and at an urban background site (at a roof) in central Stockholm, Sweden. The sites are described in Gidhagen et al. (2003). Once per month weekly samples were collected using Gent samplers (Hopke et al., 1997) with a 10 mm particle diameter cut-off during 1 year, 2003/ 2004. For each sample a total of between 130 and 190 m3 air was sampled. Air filters were prepared for analysis by microwave assisted digestion in a mixture of hydrochloric, nitric and hydrofluoric acids using a modified standard method (ASTM, 1983). Most analyses were performed by inductively coupled plasma (ICP) sector field mass spectrometry and ICP optical emission spectrometry using instrumental operating conditions given elsewhere (Engstro¨m et al., 2004). Hg was sampled and analysed as described by Zielonka et al. (2005). Road traffic emission factors (g/vehicle kilometres; g veh1 km1) were estimated using NOx as tracer for traffic emissions:

CMetal CMetal UB Ef Metal ¼ Ef NOx , Street NOx CNOx C Street UB where EfMetal and EfNOx are the emission factor for metal and NOx, respectively. Cstreet and CUB are the measured concentrations at street and urban background, respectively. This method has been successfully used in several earlier studies (Ketzel et al., 2003; Gidhagen et al., 2004, 2005; Omstedt et al., 2005). In our case we assume that the dispersion of the particles carrying the metals is similar to NOx. Since the timescale for deposition of metal containing micrometer or sub micrometer sized particles is several hours, which is much longer than the timescale for mixing and dilution, it is reasonable to assume that differences in deposition of NOx and heavy metals should have a minor influence. For our site the NOx emission factors is taken from the EVA emission model (Hammarstro¨m and Karlsson, 1994) and its factors have been found to yield NOx emission levels close to those measured in a car tunnel experiment (Kristensson et al., 2004). The estimated emission factor (1.465 g NOx veh1 km1 for the year 2000) was further supported by good agreement with calculated

3. Results and discussion 3.1. Concentrations in street canyon and at urban background The concentrations of heavy metals at the two sites are presented in Tables 1 and 2. Compared to the target values (year 2013) according to the EU directive, the levels at the street site were more than six and 50 times lower for As and Cd, respectively. For Ni concentrations were 10 times lower. Note that for As and Cd most values were below the detection limit. The concentrations observed at the roof site may be compared with measurements made 8 years earlier (1995–1996) (Table 1). The only compound with a significant decrease was Pb; 9.1 ng m3 in 1995/96 compared to 3.4 ng m3 2003/04. According to Hjortenkrans et al. (2007), emissions of Pb and Cd due to wear of brake linings have decreased by a factor of 10 during the period 1998–2005. Also for Cd and Co a slight decrease

Table 1 Metal concentrations measured at a roof-top site in central Stockholm (urban background). Metal

Arsenic (As) Cadmium (Cd) Cobalt (Co) Chromium (Cr) Copper (Cu) Manganese (Mn) Nickel (Ni) Lead (Pb) Vanadium (V) Zinc (Zn) Molybdenum (Mo) Wolfram (W) Tin (Sn) Antimony (Sb) Mercury (Hg, particulate) Mercury (Hg, gaseous)

Mean# (standarddeviation)

September 2003–September 2004 No. of values

Minimum

Maximum

No. of values below detection limit

Mean (standard-deviation)

No. of values

0.88 (0.92) 0.11 (0.09) 0.15 (0.11) 2.3 (2.1) 7.7 (6.9) 5.5 (4.0) 2.3 (1.9) 3.4 (3.1) 1.6 (1.6) 17 (15) 1.6 (2.5) 0.47 (0.66) 14 (9.8) 2.2 (1.7) 11.7 (6.2)* 1.5 (0.9)

12 12 12 12 12 12 12 12 12 12 12 12 12 12 12 12

0.03 0.014 0.014 0.30 1.2 1.1 0.37 0.45 0.06 3.2 0.12 0.03 1.9 0.22 2.9* 0.79

3.17 0.26 0.36 7.1 22.5 14 6.1 8.5 24 52 9.3 2.4 34 5.1 23* 4.3

11 11 6 2 2 1 3 1 2 0 3 1 0 1 0 0

0.88 0.31 0.37 3.0 8.7 6.1 2.7 9.1 2.9 19

(0.36) (0.36) (0.38) (1.1) (29) (5.8) (1.8) (13) (2.1) (14)

17 14 26 5 24 26 13 26 23 25

8.5 (6.5)* 1.7 (0.35)

26 24

August 1995–September 1996

Unit: ng m3. # Mean is calculated assuming concentration is equal to detection level if concentration is below detection level. 3 * pg m .


ARTICLE IN PRESS C. Johansson et al. / Atmospheric Environment xxx (2008) 1–8

Co, Mn, Zn, Mo and Pb street levels were found to be 4.0, 3.1, 2.8, 2.4, 2.2 and 2.1 times higher compared to the roof site. Fig. 1 shows that the monthly variations of the levels at street and roof were quite similar for most metals. In general, relatively low levels were found during September, October, February and March and relatively high levels were found during November–January, April and August. In Table 3 the observed levels at Hornsgatan are compared with measurements at kerb-side sites heavily influenced by traffic and urban background sites in other European cities where the same particle fraction as at Hornsgatan (PM10) were sampled and analysed. The range of the concentrations of Pb and V at the kerb-side sites is a factor 10. Metal concentrations at Hornsgatan (Stockholm) show lower levels for As, Cd, Cr, Ni, Pb, V and Zn compared to the other cities, while the levels of Cu, Co, Mn, Mo and Sb where about the same as in Athens, and Palermo. The levels at urban background in Stockholm can be compared with other urban background measurements Sevilla, Spain (Fernandez Alvarez et al., 2004; only Pb, As, Ni and Cd), Alcobendas, Tarragona and the Canaries in Spain (Moreno et al., 2006) and Palermo (Dongarra et al., 2007). For Cu, Co, Mn, Mo, Ni and Sb the lowest urban background concentrations are found in Stockholm. The concentrations of As, Cr and Zn were also lower than at other sites (except for Canaries). Differences in observed concentrations at different sites may have many explanations, but clearly the location of the measurement stations in relation to traffic and other sources is important. Different meteorological conditions at the sites may also play a role. For all metals the ranges of the concentrations at traffic sites overlap with the ranges at urban background sites. There are also data from other traffic influenced sites on the metal content of total suspended particulate matter (TSP) (Zereini et al., 2005; Bem et al., 2003), but they are not included since TSP includes particles larger than 10 mm.

Table 2 Metal concentrations measured at a street canyon site (Hornsgatan) in central Stockholm September 2003–September 2004. Metal

Mean# (standard deviation)

No. of values

Minimum

Maximum

No. of values below detection limit

Arsenic (As) Cadmium (Cd) Cobalt (Co) Chromium (Cr) Copper (Cu) Manganese (Mn) Nickel (Ni) Lead (Pb) Vanadium (V) Zinc (Zn) Molybdenum (Mo) Wolfram (W) Tin (Sn) Antimony (Sb) Mercury (Hg, particulate) Mercury (Hg, gaseous)

1.04 0.12 0.47 6.1 57.6 15.6 2.9 7.2 2.4 41 3.6

(1.0) (0.11) (0.40) (5.1) (48.9) (11.2) (2.7) (5.5) (2.2) (39) (3.6)

12 12 12 12 12 12 12 12 12 12 12

0.14 0.015 0.067 1.13 12.8 3.5 0.26 1.5 0.012 7.8 0.26

3.0 0.32 1.3 18.3 171 34.6 7.5 15.4 8.4 118 10.5

10 10 1 0 0 0 2 0 1 0 1

1.89 25.6 15.5 19.5

(1.97) (18.3) (12.8) (8.1)*

12 12 12 11

0.067 3.4 2.9 10*

6.8 57 39 34*

0 0 0 0

12

0.81

5.8

0

2.2 (1.4)

Unit: ng m3. # Mean is calculated assuming concentration is equal to detection level if concentration is below detection level. 3 * pg m .

in the concentrations is observed, while other metals shows no or only minor changes in the concentrations compared to 1995/96. For most metals substantially higher levels were observed in the street canyon as compared to the roof site. The largest difference was found for Cu and Sb, with on average more than seven times higher levels in the street compared to roof. For W,

180 Street

Roof

160

40 Street

Roof

35

140

ng/m3

100 80 60

10 5 0 sep okt nov dec jan feb mar apr maj jun aug sep

Nickel, Ni

8

Roof

Roof

16

2 0

4 3

40

4

1

2

0

Antimony, Sb Street

0 sep okt nov dec jan

sep okt nov dec jan feb mar apr maj jun aug sep

feb mar apr maj jun aug sep

Tin, Sn

60

Roof

Street

Wolfram, W

8

Roof

7

50

35

Street

Roof

6

30

40

ng/m3

ng/m3

8 6

25 20 15

ng/m3

45

10

2


Roof

12

5

ng/m3

ng/m3

ng/m3

4

Street

14

6 8

Lead, Pb

18 Street

7

10

6

15


Molybdenum, Mo Street

1

20

20 0

12

Roof

25

40


Street

30

120

ng/m3

ng/m3

Manganese, Mn

Cupper, Cu

Chromium, Cr 20 18 16 14 12 10 8 6 4 2 0

3

30 20

5 4 3 2

10

10

5 0

1

0 sep okt nov dec jan feb mar apr maj jun aug sep

0 sep okt nov dec jan feb mar apr maj jun aug sep

sep okt nov dec jan

feb mar apr maj jun aug sep

Fig. 1. Example of measured heavy metal concentrations at street and roof level during 2003/2004 in central Stockholm.


ARTICLE IN PRESS 4


Table 3 Comparison of atmospheric heavy metal concentrations in particle samples (PM10) at different sites in some European cities. Description

1) TR

As Cd Co Cr Cu Hg Mn Mo Ni Pb Sb Sn V W Zn

1.04 0.12 0.47 6.1 57.6 0.020 15.6 3.6 2.9 7.2 15.5 25.6 2.4 1.89 41

Ni/V Pb/Ni Cu/Zn Mn/Cr Cu/Sb

1.2 2.5 1.4 2.6 3.7

4) TR

7) TR

8) TR

Range TR sites

2) UB

3) UB

1.3

1.4 0.3 9.3 83

8.1 2.5 4.3 18 7.4

18 7 8 17 19

17

22

45

60

0.88 0.11 0.15 2.3 7.7 0.012 5.5 1.6 2.3 3.4 2.2 14 1.6 0.47 17

3.4 0.49

0.2 4.9 34

1.0–1.4 0.12–0.8 0.2–0.47 6.1–18.3 34–83 0.02 8.1–35.1 2.5–7 2.9–14.2 7.2–79.2 7.4–19 25.6 2.4–27.5 1.89 41–60

0.25 4.2 0.76 1.6 4.6

0.36 2.1 1.4 1.9 4.4

0.25–1.2 2.1–5.6 0.76–1.4 1.6–1.9 3.7–4.6

1.4 1.5 0.45 2.4 3.5

0.8 18.3 40.6 35.1 14.2 79.2

27.5

0.52 5.6 2.4

5) UB

6) UB

1.8

1.8

9) UB

10) UB

11) UB

12) UB

0.6 0.3

0.8 0.3

0.3 0.2

1.5 0.7

0.2 3.1 9.9

0.3 3.9 24

2.5 28.1

2.9 32.9

1.9 22.9

8.2 48.5

6.6 0.5 4.6 9.8 3.3

9.3 1.8 3.7 20 6.7 22

9.7 3.9 2.3 22.2 8.2 1.5 3.6

9.2 2.2 4.2 25.5 6.9 1.7 7.7

11 2.2 3.8 15.3 7.3 0.5 7.5

22.8 4 7.3 57.1 10.8 4.4 15.3

10 17

35

95.7

35

14.5

97.3

0.46 2.1 0.58 2.1 3.0

0.17 5.4 0.69 2.4 3.6

0.070

2.9 14.0

4.9

0.64 9.6 0.29 3.9 3.4

0.55 6.1 0.94 3.2 4.8

0.51 4.0 1.6 5.8 3.1

0.48 7.8 0.50 2.8 4.5

Range UB sites 0.3–3.4 0.2–0.7 0.15–0.3 1.9–8.2 7.7–48.5 0.07 5.5–22.8 0.5–4 2.3–7.3 3.4–57.1 2.2–10.8 1.5–14 1.6–22 0.47 17–97.3 0.17–1.4 1.5–9.6 0.29–1.6 2.1–5.8 3.0–4.5

TR ¼ kerb-side site influenced by road traffic. UB ¼ urban background site. Unit ng m3. 1) Stockholm, Sweden, this study; 2) Stockholm, Sweden, this study; 3) Sevilla, Spain, Fernandez Alvarez et al., 2004; 4) Athens, Greece, particles 0.8 are shown as bold.



5

Table 5 Emission factors for different metals for Hornsgatan in Stockholm based on measurements 2003/2004 and comparison with other studies (mg veh1 km1). Metal

Concentration increment (Cstreet–CUB) ng m3 (this study)

Emission factor (this study)

Kaisermu¨hlen-tunnel (Austria, 2002)a

Tingstad and Lundby tunnels (Gothenburg, Sweden, 1999/2000)b

Gelezinis Vilkas tunnel (Vilnius, Estonia, 2000)c

So¨derleden road tunnel (Stockholm, Sweden, 1999)d

Cobalt (Co) Chromium (Cr) Copper (Cu)

0.32 3.8 49.9

3.5 41 542

30

Tingstad: 172 Lundby: 147

159

214

Manganese (Mn) Nickel (Ni) Lead (Pb)

10.1 0.6 3.8

110 6.5 41

1.8 9.5

54

0.14 4.8

Zinc (Zn)

24

260

34

Tingstad: 36.9 Lundby: 35.1 Tingstad: 205 Lundby: 239

206

24

Molybdenum (Mo) Wolfram (W) Tin (Sn) Antimony (Sb) Mercury (Hg, part.) Mercury (Hg, gaseous)

2 1.42 11.6 13.3 7.8 103 0.7 103

22 15 126 144 85 103 7.6 103

a b c d

Laschober et al. (2004). Sternbeck et al. (2002). Valiulis et al. (2002). Kristensson et al. (2004).

3.2. Emission factors and total emissions in Stockholm Table 5 presents emission factors and total road traffic emissions of some of the metals. Compared to our estimates the emission factors in two different road tunnels in Gothenburg (Sweden) (Sternbeck et al., 2002) were significantly lower for Cu but quite similar for Pb and Zn. Kristensson et al. (2004) also found much lower values for Cr, Ni, Pb and Zn in the So¨derleden road tunnel in Stockholm. Laschober et al. (2004) estimated emission factors for a number of heavy metals based on measured concentrations in the Kaisermu¨hlen-tunnel in Austria. For Zn, Cu, Pb and Ni they found 34, 30, 9.5 and 1.8 mg veh1 km1, respectively, which all are much lower than the values obtained for the street in our study. Valiulis et al. (2002) found lower emission factors for Cu, somewhat higher emission factor for Pb and similar emission factor for Zn. The higher value for Pb obtained by Valiulis et al. (2002) in Estonia, compared to both our study and Sternbeck et al. (2002) from Sweden and also the Austrian study by Laschober et al. (2004) may be due to differences in Pb content in the vehicle fuels or lubricating oils. It is to be expected that metal emission factors for brake linings will vary depending on the traffic conditions. Our measurement site is quite close to a road traffic crossing and a slightly sloping road may lead to increased emissions due to brake wear. Several studies have found higher emissions of Zn, Pb, and Cu at sites with stop and go conditions as compared to sites with lower traffic and presumably less braking manoeuvres (Laschober et al., 2004; Allen et al., 2001; Sternbeck et al., 2002; Furusjo¨ et al., 2007; Valiulis et al., 2002). In Table 6 the estimated total road traffic emission factors for Cu, Cr, Ni, Pb and Zn are compared with emission factors for wear of brake linings based on the composition of the linings, estimated turnover time of linings and total vehicle transport in Stockholm. Based on analysis of metals in linings made in 1998 (Westerlund and Johansson, 2002) and 2005 (Hjortenkrans et al., 2007), Hjortenkrans et al. (2007) concluded that brake wear emissions have decreased drastically for Pb. The emission factor has dropped from 65 to 4.1 mg veh1 km1. The values for Cu and Zn are about the same 1998 as 2005. The small changes for Cu and Zn are not considered significant due to the large variation in metal contents of different types of brake linings (Hjortenkrans et al., 2006). The emission factor of Cu from brake wear is 84% of the total emission factor as estimated for Hornsgatan in our study (Table 6). Also for Zn a large part of the emission may be due to brake wear (50%). For

Cr, Ni and Pb, only 2, 16 and 10%, respectively, may be due to brake wear (Table 6). The brake wear emission factor estimated for all vehicles in Stockholm based on the analysis of brake linings is expected to be lower than for Hornsgatan due to the dense traffic at this street compared the average traffic of Stockholm. Then the brake wear contribution to the total emission factor is underestimated. Table 6 gives exhaust emission factors (due to fuel and lubricant oil content and engine wear) based on the COPERT model. In this model no losses in the engine are considered, except for Pb of which 25% is assumed to be lost. For Pb only fuel content is considered. According to COPERT the emission factor for Cu is 130 mg veh1 km1, almost a factor 4 lower compared to the value for wear of brake linings. The sum of exhaust and brake wear is 590 mg veh1 km1, which is quite close to the total emission factor for Hornsgatan calculated in this study (542 mg veh1 km1). Also for Ni and Zn the sum of brake wear and exhaust emissions agree rather well with the total emission factor estimated in study. For Ni most

Table 6 Comparison of emission factors for wear of brake linings and vehicle exhaust with the total emission factor based on measured concentrations at street and roof level in central Stockholm. Metal

Brake wear 1998a

Brake wear 2005b

Exhaust emission (due to fuel or lubricating oil and motor wear)c

Sum of brake wear and exhaust emission factors

Total emission factor (this study)

Copper, Cu Chromium, Cr Nickel, Ni Lead, Pb Zinc, Zn Wolfram, W

470 0.84 1.0 65 118

460 – – 4.1 131

130 3.8 5.3 71d 76

590 4.64 6.3 75.1 207

542 41 6.5 41 261 15

Unit mg veh1 km1. a Westerlund and Johansson (2002), estimated based on analysis of brake linings and estimates of the total traffic transport in Stockholm assuming that 35% of the wear gets airborne according to Garg et al. (2000). b Estimate based on brake lining analysis 2005 by Hjortenkrans et al. (2006). c COPERT 3 (from year 2000, Ntziachristos and Samaras, 2000) and COPERT 4 (from 2007, available at http://lat.eng.auth.gr/copert/) give the same emission factors for these substances. The emission factors in COPERT are given on a fuel-basis and were converted using gasoline fuel consumption of 0.095 L km1 for vehicles on Hornsgatan (2005). d Assuming a Pb content of 1 mg L1 (see text) and that 25% is lost in the exhaust system of the vehicle.


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Table 7 Calculated contribution from local road traffic in Stockholm to the total urban background concentrations in Stockholm. Metal

Calculated urban background concentration due to local road traffic (ng m3)

Rural backgrounda (ng m3)

Measured total concentration in urban background (ng m3)

Fraction due to local road traffic

Fraction due to other local sourcesb

Chromium (Cr) Copper (Cu) Nickel (Ni) Lead (Pb) Zinc (Zn) Wolfram (W) Mercury (Hg, part phase)* IVL

0.53 7.0 0.084 0.53 3.4 0.19 1.1c

0.28 0.83 0.57 1.6 3.9 No data 5.2c

2.3 7.7 2.3 3.4 17.0 1.9 11.1c

23% 91% 4% 16% 20% 10% 10%

65% 2% 72% 37% 57% – 43%

a b c

(59%) (7%) (59%) (13%) (46%) (20%)

Measurements at Birkenes in southern Norway (Aas and Breivik, 2005). Values in paranthesis refer to the contribution from other local sources if rural concentrations were 50% higher in the Stockholm area. pg m3.

emissions are due to the exhaust, whereas brake wear is more important for emissions of Zn. The sum of the exhaust and brake wear emission factors for Pb is much larger than the total emission factor estimated for Hornsgatan; 75.1 mg veh1 km1 compared to 41 mg veh1 km1 (Table 6). Most of the Pb is due to exhaust emissions, but the actual Pb content in fuels is not well known. According to reports from the fuel companies the Pb content is less than 5 mg L1. In Table 6 we assume 1 mg L1, but it could be even lower. For Cr the exhaust emission factor is 3.8 mg veh1 km1 and for brake wear the estimated emission factor in 1998 was 0.84 mg veh1 km1. Assuming similar emission factor of Cr for brake wear in 2003/ 2004 the total emission factor (4.64 mg veh1 km1) is only 10% of the total emission factor as determined in this study for Hornsgatan, indicating that there are other road traffic sources for Cr. It should be noted that the emission factor for brake wear according to Westerlund and Johansson (2002) and Hjortenkrans

et al. (2006), rely on the assumption that 35% of the wear particles become airborne, the rest being deposited onto wheel covers and other surfaces. However, Sanders et al. (2003) report that 50–70% might become airborne. This would mean that the emission factors for brake linings would be a factor 2 higher than shown in Table 6. Obviously, the fraction airborne will depend on vehicle construction and the driving conditions. 3.3. Air quality model calculations Based on the emission factors obtained from measurements at street and roof level, the contribution from road traffic was calculated for the Greater Stockholm area (35 km 35 km) using the Gaussian air quality model (Table 7 and Fig. 2). Calculated road traffic contributions to the total measured concentrations of Cu, Cr, Zn, Pb, W, Hg(p) and Ni at roof level was 91, 23, 20, 16, 10, 10 and 4%,

Fig. 2. Geographic distribution of Cu concentrations due to road traffic emissions in the greater Stockholm area. Calculated values based on air quality dispersion modeling and emission factors obtained in this study for Hornsgatan.



respectively. For these calculations it is assumed that the emission factors estimated for Hornsgatan are representative for all roads in the area. Thus, if for example, the emissions due to brake wear at Hornsgatan are larger than other streets, the calculated contributions are overestimated. Due to lack of background (rural) measurements it is difficult to determine the total local (urban) contribution to the levels. Based on background data from southern Norway (Birkenes) most of the Cu in the urban background air in Stockholm is due to local road traffic. Our calculation then indicates that brake wear emission is the main source of Cu. This is in accordance with Hulskotte et al. (2006), who found that brake wear is the dominating source of atmospheric Cu in Western Europe and the main contributor to atmospheric Cu concentrations and deposition in this region. The calculations also indicate that more than 50% of the measured Ni, Cr and Zn are due to other local sources. 4. Conclusions Heavy metal emissions from road traffic has been estimated based on parallel filter sampling in a densely traffic street and at a roof-top site in central Stockholm. Even on one of the most densely trafficked streets in Stockholm the levels of As, Cd and Ni are several times lower than the EU directive. Concentrations at a street canyon site are two to four times higher than at an urban background site, except for Cu and Sb, for which the levels are seven times higher. Using NOx as tracer, emission factors have been estimated and compared with estimated emissions from brakes and vehicle exhaust. For Cu and Zn brake wear is likely to be very important for the emissions as indicated by comparing with estimates based on brake wear analysis and verified by high correlations between these metals and with Sb. Metals such as Cr, Ni, Mo, Co, V and Mn are part of a different steel alloys and are therefore also correlated. The emission factors estimated for the street canyon site are higher than estimates based on road tunnel sites, likely due to different driving conditions (more stop and go at the street canyon site). The total emission factors for Cu, Cr and Zn at the street site are much higher than emission factors for vehicle exhaust. For Pb and Ni vehicle exhaust emission is the dominating source. Acknowledgements This work was financed by Stockholm city as part of the project ‘‘New poisons – new tools’’ (Nya gifter – nya verktyg). References Aas, W., Breivik, K., 2005. Heavy metals and POP measurements 2003. EMEP Cooperative Programme for Monitoring and Evaluation of the Long-range Transmission of Air Pollutants in Europe. EMEP/CCC report 9/2005, Norwegian Institute for Air Research, P.O. Box 100, N-2027 Kjeller, Norway (http://www. nilu.no/projects/CCC/reports.html, accessed Jan., 2008). Allen, J.O., Mayo, P.R., Hughes, L.S., Salomon, L.G., Cass, G.R., 2001. Emissions of sizesegregated aerosols from on-road vehicles in the Caldecott tunnel. Environmental Science and Technology 35, 4189–4197. ASTM, 1983. Standard test method for trace elements in coal and coke by atomic absorption. Annual Book of American Society for Testing and Materials, Standards, Vol. 05.05; Gaseous Fuels, Coal and Coke; Atmospheric Analysis. ASTM, Philadelphia, ANSI/ASTM D3683-78. Bem, H., Gallorini, M., Rizzio, E., Krzeminska, M., 2003. Comparative studies on the concentrations of some elements in the urban air particulate matter in Lodz City of Poland and in Milan, Italy. Environmental International 29, 423–428. Chan, D., Stachowiak, G.W., 2004. Review of automotive brake friction materials. Journal of Automobile Engineering 218, 953–966. Danard, M., 1977. A simple model for mesoscale effects of topography on surface winds. Monthly Weather Review 99, 831–839. de Miguel, E., Llamas, J.F., Chacon, E., Berg, T., Larssen, S., Royset, O., Vadset, M., 1997. Origin and patterns of distribution of trace elements in street dust: unleaded petrol and urban lead. Atmospheric Environment 31, 2733–2740.

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