Fine particle and gaseous emissions from normal and smouldering wood combustion in a conventional masonry heater

Atmospheric Environment 42 (2008) 7862–7873

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Fine particle and gaseous emissions from normal and smouldering wood combustion in a conventional masonry heater J. Tissari a, *, J. Lyyra¨nen b, K. Hyto¨nen a, O. Sippula a, U. Tapper b, A. Frey c, K. Saarnio c, A.S. Pennanen d, R. Hillamo c, R.O. Salonen d, M.-R. Hirvonen d, J. Jokiniemi a, b a

Fine Particle and Aerosol Technology Laboratory, University of Kuopio, P.O. Box 1627, FIN-70211 Kuopio, Finland VTT, Technical Research Centre of Finland, Fine Particles, P.O. Box 1000, FIN-02044 VTT, Espoo, Finland c Air Quality Research, Finnish Meteorological Institute, Erik Palme´nin Aukio 1, P.O. Box 503, FIN-00101 Helsinki, Finland d National Public Health Institute, Department of Environmental Health, P.O. Box 95, FIN-70701 Kuopio, Finland b

a r t i c l e i n f o

a b s t r a c t

Article history: Received 30 January 2008 Received in revised form 7 July 2008 Accepted 8 July 2008

The fine particle and gas emissions from the residential wood combustion (RWC) appear to be a major contributor to winter-time pollution in Europe. In this study, we characterised the effect of two different combustion conditions on particulate and gaseous emissions from a conventional masonry heater. Normal combustion (NC) is the best available operational practice for the heater, whereas smouldering combustion (SC) mimicked slow heating combustion. It was found that the operational practice in RWC can significantly influence the quantity and quality of particle and gaseous emissions into the atmosphere. In SC, the emissions of carbon monoxide were 3.5-fold, total volatile organics 14-fold and PM1 6-fold to those of NC, whereas the mass of the inorganic compounds (‘‘fine ash’’) and particle number emissions were lower from SC than from NC. According to electron microscopy analyses, the observed fine ash particles seemed to occur mainly as separate spherical or irregularly shaped particles but not as agglomerates. Ultrafine (<100 nm) fine ash particles were composed mainly of K, S and Zn, but also in a lesser extent of C, Ca, Fe, Mg, Cl, P and Na. Large agglomerates were found to contain mainly carbon and are considered to be primarily soot particles. The larger spherical and irregularly shaped particles were composed of same alkali metal compounds as ultrafine particles, but they were probably covered with heavy organic compounds. From SC, particles were composed mainly of carbon compounds and they had a more closed structure than the particles from NC, due to organic matter on the particles. In the present experiments, the ultrafine mode in the particle number distributions seemed to be determined mainly by the amount of released ash forming material in combustion, and the shifting of particle size during different combustion conditions seemed to be determined by the amount of condensed organic vapour in the flue gas. Ó 2008 Elsevier Ltd. All rights reserved.

Keywords: Wood smoke Fine particles Size distribution Smouldering combustion Single particle composition Gaseous emissions

1. Introduction Residential wood combustion (RWC) for heat production is an important source of both gaseous and particulate pollutants (McDonald et al., 2000; Johansson et al., 2004;

* Corresponding author. Tel.: þ358 40 355 3237; fax: þ358 17 163 229. E-mail address: jarkko.tissari@uku.fi (J. Tissari). 1352-2310/$ – see front matter Ó 2008 Elsevier Ltd. All rights reserved. doi:10.1016/j.atmosenv.2008.07.019

Oanh et al., 2005; Tissari et al., 2007; Sippula et al., 2007). According to the latest studies, the health effects of inhaled fine particles (PM2.5: particulate mass, i.e. 2.5 mm or smaller in aerodynamic size) from wood combustion may be more harmful than has previously been thought (Boman et al., 2003; Naeher et al., 2007). The emissions from RWC have been demonstrated to be highly variable and dependent on many practical factors (Johansson et al., 2004; Nussbaumer, 2003; Johansson

J. Tissari et al. / Atmospheric Environment 42 (2008) 7862–7873

et al., 2003; Tissari et al., 2007; Sippula et al., 2007). For example, numerous types and models of wood combustion appliances are being used and the wood fuel can originate from several tree species. The operational practices of RWC also vary widely (e.g., fuel seasoning, combustion patterns, combustion rates, kindling approaches, etc.) and often these practises are not well established from the emission point of view. Generally, the highest emissions have been found in the smouldering combustion conditions that are mainly due to poor operational practices (e.g. restricted combustion air). These occur for example in old wood boilers without heat-storing tanks (Johansson et al., 2004), and in appliances where the storage of heat is not possible, e.g. in light metal stoves (Jordan and Seen, 2005) because their operation is often performed with restricted air to slow down the heating. The poor combustion conditions can also be associated to natural fires and uncontrolled combustion which are a large source of organic matter to the atmosphere (e.g. Robinson et al., 2007). Masonry heaters are enclosed combustion appliances made of masonry products, a combination of masonry products and ceramic materials, or soapstone (Stehler, 2000; Tissari et al., 2007). Others are covered with decorative tiles. They have a very high mass, typically from 800 to 3000 kg, and can be up to 6000 kg. In these heaters wood is combusted in a relatively short period of time and at high power. Typically, the heaters have an upright firebox with a glass door. In the contraflow system, the exhaust gas flows from the firebox to an upper combustion chamber, and goes down through the ducts into the chimney from the bottom or top of the heater. The energy released is efficiently stored in the large mass surrounding the firebox and the ducts. Masonry heaters produce both primary and supplemental heat, when the heat stored in the stone mass slowly releases (at an average rate of 1–3 kW) into the indoor air during the next 1–2 days (Tissari et al., 2007). In Finland, there are approximately 3.7 million wood combustion appliances; those are mainly masonry heaters and sauna stoves. In addition, almost all new detached houses are equipped with masonry heater. The most important parameters for complete combustion conditions are a high combustion temperature, a sufficient amount of combustion air supply, and adequate mixing of combustion air and fuel gas (e.g. Nussbaumer, 2003; van Loo and Koppejan, 2008). The combustion temperature affects primarily the burn out of combustion compounds. The oxidation reactions are faster and more complete, and the combustion time shorter in high temperatures than at low ones. In RWC appliances, there must be an overall excess of oxygen, and the overall air-to-fuel ratio is usually above 1.5, to ensure a sufficient local oxygen concentrations for combustion reactions. However, the combustion temperature decreases as a function of the excess air, mainly due to the heating of inert nitrogen in the air. Particularly in the closed combustion appliance, where the sizes of air intakes are restricted, the gasification rate of fuel is a very important parameter. This is because too fast fuel pyrolysis leads to insufficiency of air supply, and high emissions of incomplete combustion compounds. The gasification rate is controlled mainly by the primary air supply, but the log

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and batch sizes (i.e. total surface area of wood logs) also strongly affect the gasification rate of wood. Combustion conditions have a great influence on the fine particle and gas emissions particularly from RWC appliances. In good combustion conditions, e.g. in largescale power plants or pellet combustion in small units, the air supply is sufficient and the combustion temperature is high. The hot secondary air in the top of the primary combustion chamber and a good mixing of secondary air and combustion gases enhance the ignition of hydrocarbons, leading to low gaseous emissions. In addition, the soot particles, which are formed in the flame from pyrolysis gases, burn away in the flame or in the post-flame oxygenrich zone (e.g. Amann and Siegla, 1982; Wiinikka et al., 2007). Thus, in good combustion conditions the fine particle emission is formed mainly by the vaporisation of ash forming elements from the wood fuel (i.e. ‘‘fine ash’’ that means here inorganic compounds present in fine particle fraction). The behaviour of ash forming elements in biomass combustion has been studied intensively more than 10 years and it is generally well-known (e.g. Baxter, 1993; Christensen, 1995; Dayton et al., 1995; Valmari et al., 1998; Davidsson et al., 2002; Sippula et al., 2007). The vaporisation is dependent on the chemical composition of the wood and the reactions of inorganic species. The combustion temperature also has an important influence on the vaporisation, so that greater amount of fine ash particles is released at high than at low temperatures (e.g. Davidsson et al., 2002; Knudsen et al., 2004). The volatile ash forming elements in wood fuels are mainly alkali metals, chlorine and sulphur, and the ash compounds measured in the fine particles from RWC are primarily potassium sulphates, chlorides, carbonates and hydroxides (e.g. Valmari et al., 1998; Sippula et al., 2007). In incomplete combustion conditions, e.g. due to a low combustion temperature and/or local or overall lack of available oxygen for combustion, the gaseous CO and volatile hydrocarbon emissions increase remarkably. Besides the increase in soot formation, heavy and complex organic compounds condense onto the surfaces of the primary particles, or may form very small organic particles by nucleation (Tissari et al., 2007). Thus, in incomplete combustion conditions, the fine particles are mainly composed of organic matter, elemental carbon (EC, i.e. soot) and fine ash and in addition, physical properties of fine particles are very complex. Fine et al. (2001) reported that from 40% to almost 100% of fine particle mass was organic material, in case of a conventional fireplace. The fractions of EC in fine particles have been reported to vary between 1 and 31% from fireplace and wood stove (Schauer et al., 2001; Fine et al., 2001). In a study of Tissari et al. (2007) the fine particle emissions from different type of masonry heaters contained 31–78% organic material, 15–41% EC and 6–32% fine ash. Because RWC produces fine particles and hazardous organic compounds causing air quality problems (Glasius et al., 2006), there is a need to decrease the particle and gaseous emissions from wood combustion in small scale appliances. In many cases, flue gas filtering systems are still not economically feasible and on the other hand there is a large potential to decrease emissions by developing the

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Flue gas fan

Filtration CO2 Analyzer

Hood Air valve Gas analyzing rack

Stack

Dilution tunnel

F Condenser

OGC

T

CO CO2

P

Constant volume pump

T

Precyclone

MFC

Heater

O2

F

Zero gas N2

FMPS

Ball valve

F DLPI T SEM/ TEM

P F

T

Masonry Heater

FTIR

PM1 Impactor TF

T

Balance Needle valve = Heated sampling line or thermal insulation

Pump

ELPI

Flow meter Gas meter

Fig. 1. The experimental set-up: FTIR, gas analyzer; ELPI, Electrical Low Pressure Impactor; DLPI, Dekati Low Pressure Impactor; FMPS, Fast Mobility Particle Sizer; MFC, mass flow controller; TF, Teflon filter; F, filter; T, thermocouple; P, pressure sensor; TEM/SEM, TEM/SEM sampling unit.

combustion technology itself. Thus, there is actual need to get detailed information from the particle and gas emissions in different kind of RWC appliances to be able to develop low emission combustion techniques. However, extensive information on emissions from RWC appliances are still scarce. The aim of this study was to investigate the fine particle and gas emissions from the conventional masonry heaters in two clearly different combustion conditions. The main focus of the work was in the extensive and detailed characterisation of fine particles, but also particle and gas emission factors were defined as we also wanted to provide uncertainty ranges of the emission factors used in emission inventories.

A conventional masonry heater (CMH) was used in the experiments. It is a commercially available relatively small soapstone masonry heater (Fig. 2) with dimensions of 1.23  0.72  0.45 m3, mass of 800 kg and firebox volume of 27.7 dm3. The heater uses a conventional combustion

2. Methods 2.1. Combustion facilities and appliance The experiments were conducted in the emission test laboratory of the University of Kuopio. The experimental set-up is shown in Fig. 1. During the combustion experiments the masonry heater was situated on a balance to enable the measurement of fuel mass flow. To mimic a natural draught, the combustion gases were led through an externally insulated steel stack of 180/230 mm diameter placed below a hood. The draught in the stack can be adjusted by using a flue gas fan, changing the location of the hood, and with a damper (Fig 1).

Fig. 2. Sketch of the conventional masonry heater used in this study.

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Table 1 Parameters of operational practice in normal combustion (NC) and smouldering combustion (SC)

NC SC a

Na

Batch size (kg)

Log size (kg)

Total loading (kg)

Batch time (min)

Arrangement of logs

Combustion rate (kg hr1)

8 3

2.4  0.1 3.5  0.1

0.48 0.23

9.1  0.3 8.9  0.2

20 30

Compact Crosswise

7.8 7.1

Number of experiments.

technique with a conventional grate and an upright enclosed firebox with a glass window in the front door. The main air supply is through the grate (Fig. 2). The main air damper in the ash box door controls the combustion rate during the operation. The auxiliary air dampers just below and above the glass door control a minor air flow to flush the window. The exhaust gas flows from the firebox to an upper combustion chamber, down through the two side ducts and then into the chimney from the bottom of the heater. The energy released during combustion is stored very efficiently (typically more than 75%) in the large stone mass around the firebox and the ducts. 2.2. Combustion procedures and conditions The CMH was operated with smouldering (SC, smouldering combustion) and with good operational practise (NC, normal combustion). SC was produced by restricting the combustion air supply and slightly overloading the firebox with log-wood to achieve a low heat output. Respectively, NC describes the best applicable operational practice of the heater from the emission point of view, and was defined in careful pre-tests. In this study, eight NC and three SC combustion tests were performed. Both the combustion procedures were started with a similar ignition/kindling batch, 1.7 kg of wood logs and 0.15 kg of wood sticks with a minor amount of birch bark. Fifteen minutes after ignition the kindling batch was followed by two (SC) or three (NC) subsequent batch additions (Table 1). All of the burn cycles from start up to burn-out of all batches were included to the emission analysis (however, without the combustion of burn out residues from certain point, Fig. 3). The experiments lasted from 75 to 85 min.

3rd Batch (NC)

NC SC

2nd Batch (NC)

Ignition

300 Batch 1st Batch (NC)

200

Burn Out

100 1st Batch (SC)

2nd Batch (SC)

0 0

20

40

60

80

2.3. Fuel The fuel used was birch wood, which is widely used in Finland. The logs in this study were 230 mm in length. Normally, moisture content of wood logs is between 15 and 25%, but even 10% moisture contents have been found in the operation of masonry heaters in winter (Tissari et al., 2007). However, in this study the wood was stored in the laboratory and the moisture content was as low as 7%. The ash content and the net heating value of the fuel were not determined. Typically, the ash content of Finnish birch logwood is below 0.5% of dry fuel, and the net heating value is 18–19 MJ kg1 (e.g. Sippula et al., 2007).

Oxygen concentration (%, dry)

Flue gas temperature (°C)

400

Different batch sizes, log sizes and different arrangements of logs were used in NC and SC in order to get similar pyrolysis rates of wood for both combustion cases. In SC, the restricted combustion air decreases the pyrolysis of wood, but the use of small log size, large batch size and crosswised arrangements of logs (i.e. higher total surface area of wood logs) increases the pyrolysis of wood. In NC, the free combustion air supply increases the pyrolysis, but the large log size and compact arrangement of logs decreases the pyrolysis of wood. Thus, the fuel consumption rates were almost similar in both cases, on average 7.8 kg wood/h in NC, and 7.1 kg wood/h in SC. The flue gas temperature and oxygen concentration are shown in Figs. 3 and 4 and CO concentration in Fig. 5. The cyclic nature of the combustion process is clearly seen in the figures. The firing and burn out phases were similar in both cases studied. The flue gas temperature was clearly lower in SC (159  28  C (average  standard deviation) than in NC (253  3  C).

100

Combustion time (min) Fig. 3. The flue gas temperature during normal combustion (NC) and smouldering combustion (SC).

20

15

10

5

NC SC

0 0

20

40

60

80

100

Combustion time (min) Fig. 4. The flue gas oxygen concentration during normal combustion (NC) and smouldering combustion (SC).

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in the undiluted flue gas and CO2,BG is the CO2 concentration in the background dilution air.

60000 NC SC

4000

40000

20000

2000

0 0

20

40

60

80

CO in SC (ppm, dry)

CO in NC (ppm, dry)

6000

0 100

Combustion time (min) Fig. 5. The flue gas carbon monoxide (CO) concentration (raw data) during normal combustion (NC) and smouldering combustion (SC).

2.4. Gas analyses The sample for the gas analysers was taken straight from the stack through an insulated and externally heated (180  C) sample line. Particles were removed from the sample air by a ceramic filter unit. The exhaust gas was measured continuously with a combination of gas analysers (ABB Cemas Gas Analysing Rack) for carbon monoxide (CO), carbon dioxide (CO2) and oxygen (O2). The organic gaseous carbon (OGC) was analysed with a flame ionisation detector that was calibrated against propane. Nitrogen oxides (NOx) and volatile organic compounds (VOCs) were measured with a Fourier Transform Infrared analyser (FTIR, Gasmet Technologies Ltd.) through an insulated and externally heated (180  C) sample line (Fig. 1).

2.6. Particle measurements and analyses Particle number emissions and number size distributions were measured in a parallel real-time set-up using an Electrical Low Pressure Impactor (ELPI, Dekati Ltd.), with a flow rate of 10 lpm and a cut-off size from 24 to 9.38 mm with sintered impactor plates, and a Fast Mobility Particle Sizer (FMPS, TSI 3091) with a flow rate of 10 lpm and a cut-off size from 6 to 523 nm. It must be noted that FMPS measures particle electrical mobility diameter, whereas ELPI measures particle aerodynamic diameter. Particle mass size distributions were measured using a Dekati Low Pressure Impactor (DLPI, Dekati Ltd.) with a flow rate of 10 lpm and a cut-off size from 28 to 9.84 mm with greased Al-foils as collection substrates. From one to three PM1 samples per combustion experiment were collected on 47-mm Teflon membrane filters (Gelman Scientific, Teflo) with a flow rate of 10 lpm. A cascade impactor (Dekati Ltd.) was used to remove particles larger than 1 mm (Fig. 1). The filters for gravimetric analysis were kept for 24 h at a constant temperature of 20  C and a relative humidity of 40% before weighing, and were weighed using a microbalance of 1 mg sensitivity. The weighing procedure has been presented in detail in Tiitta et al. (2002). Particle samples for electron microscopy were collected in the stack on holey carbon copper grids using thermophoretic sampling. The samples were analysed by a scanning electron microscope (SEM, Leo DSM 982 Gemini) and a transmission electron microscope (TEM, Philips CM-200 FEG/STEM operated at 200 kV), including elemental analyses of the single particles from TEM samples by Energy Dispersive Spectroscopy (EDS).

2.5. Aerosol dilution

2.7. Temperature measurements

Before particle measurement, the sample air flow was diluted in a dilution tunnel (diameter 300 mm). The tunnel is constructed according to ISO 8178-1 standard which is intended for emission tests of internal combustion engines (Fig. 1). A partial flow from the stack was led through an externally insulated and heated 12 mm steel tube to the dilution tunnel by the low pressure in the tunnel. The dilution air was filtered in three stages, where a pre-filter removed coarse particles, a chemical filter removed hydrocarbons and nitrogen oxides, and a post-filter removed fine particles. The total air flow in the tunnel was adjusted with a constant volume pump (flow rate 0–1200 m3 h1) and the low pressure of the tunnel was controlled with an air valve situated after the filters, giving dilution ratios (DR) of 180–330. The DR was calculated on the basis of the concentrations of CO2 (dry) in raw and diluted exhaust gas with the equation

Temperatures were monitored continuously from the exhaust gas, the outlet of the partial flow tube, the dilution tunnel and the laboratory room air with K-type thermocouples.

DR ¼

CO2;FG  CO2;BG ; CO2;D  CO2;BG

where CO2,D is the CO2 concentration in the diluted gas (CO2-sensor: Sensorex Ltd.), CO2,FG is the CO2 concentration

2.8. Data collection and calculations The real-time particle and gas concentration data was collected at 10 s intervals from the firing to burn out phase. One particle mass size distribution sample was collected from each test. One PM1 sample was collected from the NC, whereas due to the filling of the filter, two to three sequential samples were collected during SC (one per batch). Particle samples for electron microscopy were collected from different combustion phases at couple of second sampling time. The emissions from both real-time data and filters were calculated also as an average of each test. The uncertainty in the emission values was defined as standard deviations which were calculated from average values of eight NC and three SC emission experiments. The nominal emission values were calculated in relation to energy input to the combustion process (SFS 5624, 1990)

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Table 2 Average particulate emissions from a conventional masonry heater in normal (NC) and smouldering combustion (SC) conditions

NC  SD SC  SD

Particle number (FMPS) (1015 kg1)

GMD (FMPS) (nm)

Particle number (ELPI) (1015 kg1)

GMD (ELPI) (nm)

PM1 (Filter) (g kg1)

PM10 (DLPI) (g kg1)

3.9  0.5 1.4  0.4

56  3 118  21

3.9  0.7 1.9  0.7

65  4 160  30

1.8  0.5 11.1  3.9

2.1  0.3 11.3  3.8

and then in relation to the amount of fuel used in units of g fuel kg1 (dry) by multiplying the value by the net heating value of fuel (18.3 MJ kg1). The particle size distributions were calculated as mass concentrations (mg m3) and number concentrations (# cm3) normalised to 13% oxygen in the dry flue gas. The particle emission and concentration values were dilution-corrected. 3. Results and discussion 3.1. Particle emissions

3.1.2. Particle mass emissions and mass size distributions Most of the particles were below 1 mm in aerodynamic diameter (Fig. 10). The ratio of PM1-to-PM10 measured by DLPI was 0.8 in both combustion conditions. Filter-based PM1 from NC was 1.8 g kg1 and about 6-fold from SC (Table 2). The mass size distributions peaked in the size range of 150–200 nm, 400–500 nm, and 2–3 mm in NC and at 500–700 nm in SC (one mode) (Fig. 10). The mean GMD measured by DLPI was 243  24 nm in NC and 534  36 nm in SC. The PM1 emission from NC was well comparable with previous studies. In masonry heaters, baking ovens and stoves, emissions of 0.6–1.6 g kg1 for PM1 have been reported (Tissari et al., 2007). The PM1 emissions from NC was also on the same level as particle mass emission from stoves (2.3–7.2 g kg1, McDonald et al., 2000) and cookstoves (0.9–2.8 g kg1 (Venkataraman and Uma Maheswara Rao, 2001). In SC, PM1 emission was on average 11.1 g kg1 which is a relatively high emission. However, Johansson

0.18 6.E+08

NC SC

NC SC

GMD (µm)

Particle number concentration, N (# cm−3, in flue gas conditions)

3.1.1. Particle number emissions and size distributions From NC, the particle number emission was relatively high, 3.9  1015 particles per kilogram of dry wood combusted, measured by ELPI and FMPS. In contrast in SC, the particle number emission was less than half of those from NC and it was lower during the whole combustion period (Table 2; Fig. 6). The particle number emissions were well comparable or slightly higher than in other studies. In masonry heaters, baking ovens and stoves, emission factor of 0.7–2.6  1015 particles per kilogram (measured by FMPS) have been reported (Tissari et al., 2007). Hedberg et al. (2002) measured emission factors of 0.13–9.7  1014 particles per kilogram measured by DMPS. In this study, the geometric mean diameters (GMDs) of emitted particles from NC were on average 56 nm by FMPS and 65 nm by ELPI (Table 2). The mean GMD from SC was 2–2.5-fold to that of NC. In NC after adding a new batch, the GMD of the particles peaked shortly up (around 5 min period) and at the same time the flue gas oxygen concentration dropped down (Figs. 4 and 7). For example in the case of Fig. 7, the GMD by ELPI peaked shortly up to 110 nm

after adding the batch, but stabilised rapidly to 40–60 nm. Respectively from SC experiments, the GMD was large, about 140 nm, during the main pyrolysis period and decreased only at the end of the batch to 60–70 nm (Fig. 7). From NC, the average particle number size distributions seemed to be unimodal and the repeatability was good (Fig. 8). However, in addition to ultrafine mode of particles, also accumulation mode particles were observed in the temporary distributions (Fig. 9a,b). In SC, the number size distribution was composed of two particle size classes with peaks at 70 and 250 nm, measured with ELPI (Fig. 9c). The nucleation mode was not observed in either of the two combustion conditions in the present study.

4.E+08

2.E+08

0.E+00

0.12

0.06

0.00 10

20

30

40

50

Combustion time (min) Fig. 6. A time series of particle number concentrations during combustion of a wood-log batch in normal (NC) and smouldering (SC) combustion as measured by an Electrical Low Pressure Impactor (ELPI).

10

20

30

40

50

Combustion time (min) Fig. 7. A time series of particle geometric mean diameters (GMDs) during combustion of a wood-log batch in normal (NC) and smouldering (SC) combustion as measured by an Electrical Low Pressure Impactor (ELPI).

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dN/dlogDp [#/cm3, red. 13% O2]

8.0E+08 NC, ELPI

NC, FMPS

4.0E+08

NC, Test 1 NC, Test 2 NC, Test 3

NC, Test 1 NC, Test 2 NC, Test 3

NC, Test 4 NC, Test 5 NC, Test 6

NC, Test 4 NC, Test 5 NC, Test 6

NC, Test 7 NC, Test 8

NC, Test 7 NC, Test 8

0.0E+00

dN/dlogDp [#/cm3, red. 13% O2]

8.0E+08 SC, FMPS

SC, ELPI

SC, Test 1 SC, Test 2

SC, Test 1

SC, Test 3

SC, Test 3

SC, Test 2

4.0E+08

0.0E+00 0.01

0.1

1

10

Dem (µm)

100

0.01

0.1

1

10

100

Da (µm)

Fig. 8. The average particle number size distributions during normal (NC, eight experiments) and smouldering (SC, three experiments) combustion as measured by a Fast Mobility Particle Sizer (FMPS) and an Electrical Low Pressure Impactor (ELPI).

et al. (2004) found even 4-fold higher particle mass emission (42 g kg1) from the combustion of an oldtype wood log boiler with a big batch size. In addition, Jordan and Seen (2005) observed as high as 40 g kg1 particle emissions from a wood stove with restricted combustion air. 3.1.3. Particle morphology and composition of single particles In SEM and TEM analyses, both large agglomerates (Figs. 9d–f and 11) and separate spherical and irregularly shaped particles were observed (Figs. 9d–f, 12 and 13). Separate particles were to some extent connected with agglomerates. According to elemental analyses of the single particles, agglomerates were composed mainly of carbon and were considered to be soot. According to literature, they are formed in the flame and most of these particles burn in the oxygen-rich zone in the flame (e.g. Wiinikka et al., 2007). Minor part of the soot particles is released, mainly as agglomerates composed of about 30–50 nm solid carbon spherules (e.g. Kocbach et al., 2005). Similar spherule sizes were also observed in this study (Fig. 11). Further, ultrafine (<100 nm) separate particles composed mainly of K, S and Zn were observed in the samples (Figs. 9e and 12). Also O, Ca, Fe, Mg, Cl, P and Na were observed, but only in some of the analysed particles. In addition, carbon, Cu and Si were identified from the

samples, but according to background beam, they were mainly originated from the TEM collection grids. According to previous studies, fine ash particles are formed after the soot formation process, when the temperature decreases. First, the alkali sulphate vapours begin to condense (Sippula et al., 2007), which leads to new particle formation by homogeneous nucleation or to condensation on existing seed particles (Jokiniemi et al., 1994). In the following steps, the alkali hydroxide vapours convert partly to alkali carbonates and the alkali chloride vapours condense (Sippula et al., 2007). Particularly in the firing phase of combustion, samples contained also a fraction of large spherical and irregularly shaped particles, composed of alkali metal compounds and carbon, similarly to observed ultrafine particles (Fig. 13). However, the behaviour of the particle in the analysis under the beam differed from typical soot or fine ash particles. The particle material volatilised more easily under the beam. We suggest that this material was mainly composed of heavy organic compounds, as presented later in the SC particles. From SC, particles were composed mainly of carbon compounds. The particles were more irregular and had a more closed (sintered-like) structure than the particles from NC (Figs. 9f and 14). The particle effective density can

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8

a

d

GMD 63 nm N 4.6×108 # cm-3

6

4 120 nm

Particle number concentration, dN dlogDp-1 (×108 # cm-3 red. 13% O2)

2 2 µm 0 8

b

e

GMD 56 nm N 4.2×108 # cm-3

6

80 nm

4

2

2 µm 0

1.6

c

f 250 nm

1.2 GMD 230 nm N 7.6×107 # cm-3 0.8

110 nm 0.4 2 µm 0.0 0.01

0.1

1

Particle aerodynamic size (µm) Fig. 9. Temporary particle number size distributions during firing phase (a), combustion phase in NC (b) and combustion phase in SC (c) and comparable SEM figures from the equivalent combustion conditions (d–f, respectively).

be estimated by comparing particle aerodynamic and electric mobility size (e.g. Ristima¨ki et al., 2002; Zelenyuk et al., 2008). On the basis of simultaneous ELPI (aerodynamic size) and FMPS (electric mobility size) measurements, it seems that particles had higher effective density

in SC than in NC. In NC the aerodynamic and mobility GMDs were almost equal while in SC the aerodynamic GMD was clearly higher than mobility GMD. In SC, both larger particle size and higher effective density of the particles may indicate organic matter condensation on the agglomerate

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dM/dlogDp [mg/m3, red. 13 % O2]

200

NC, DLPI NC, Test 1 NC, Test 2 NC, Test 3 NC, Test 4 NC, Test 5 NC, Test 6 NC, Test 7 NC, Test 8

150

100

50

0

dM/dlogDp [mg/m3, red.13 % O2]

1600

SC, DLPI

1200

Fig. 11. SEM micrograph of typical agglomerate particles from NC.

SC, Test 1 SC, Test 2 SC, Test 3

800

particles. This can be also seen as closed structure of particles in SC (Fig. 9f). Similar differences between chemical compositions of NC and SC were observed in the parallel study of Frey et al. (in press). From filter samples taken in the same experiments, the particles were mainly of POM (33%) and EC (32%) in NC. In SC, the portions were 67–69% and 22–27%, respectively. Furthermore, the emission of fine ash compounds was clearly higher from NC (emission factor about 0.54 g kg1) than from SC (0.16–0.27 g kg1). In summary, the results of this study suggest that condensation on the existing fine particles is the primary physical mechanism in the formation of POM. Based on

400

0 0.01

0.1

1

10

100

Da (µm) Fig. 10. The average particle mass size distributions during normal (NC, eight experiments) and smouldering (SC, three experiments) combustion as measured by a Dekati Low Pressure Impactor (DLPI).

200

a

a

Counts/Channel

C 160 Cu

O

120

Zn/Cu

Zn

80 Si 40

Cl

S

K

0

b

b

Cu

C

Counts/Channel

50 nm

500 400 300 O

200 100

P Cu

Cu

Ca

Fe

K 0

0

1

2

3

4

5

6

7

8

9 10 11 12

Energy [keV] Fig. 12. Compilation of four TEM micrographs of typical ultrafine (<100 nm) particles from NC and the elemental compositions of the particles (a) and (b).

J. Tissari et al. / Atmospheric Environment 42 (2008) 7862–7873

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35 Cu 30

Counts/Channel

C S

25 20 15 10

K

EDS

Cu 5

50 nm

0

0

1

2

3

4

5

6

7

8

9

10

11

12

Energy [keV] Fig. 13. TEM micrograph of typical large spherical particles from the firing phase and the elemental composition of the core of particle.

the amount of condensed organic vapour available in the flue gas.

3.2. Gas emissions

Fig. 14. SEM micrograph of a particle formed in smouldering combustion. The particle contains organic matter on the surface.

high number of ultrafine ash particles and large size of soot agglomerates, we suggest that the ultrafine mode in particle number size distribution is mainly formed from fine ash particles. Thus, it also seems that the primary particle size and number emission are determined by the amount of volatilised ash forming material from wood fuel. The shifting of particle size during different combustion conditions seems to be mainly determined by

The average gas emission factors during the tests are shown in Table 3. CO emission from SC was 3.5-fold to that of NC. NOx emissions (1.3–1.6 g kg1) were almost the same in all of the samples. The largest differences in gas emissions between the NC and SC were observed for the gaseous organic emissions. The OGC emission measured with the FID method from the SC was 14-fold to that of NC. Emissions of single hydrocarbons were also clearly higher from SC than from NC. In SCs, there were problems in the measurements of single hydrocarbons with FTIR (too high concentrations), so the results from only one measurement are presented. In this successful experiment, the sample gas for FTIR was diluted with heated nitrogen and the dilution was regulated accurately with a mass flow controller. The emissions of six common VOCs are presented in Table 3. From SC, the emission of methane was 11-fold, acetylene 14-fold, ethene 15-fold, propene 6-fold, 1.3-butadiene 7-fold, and benzene 12-fold those from NC (Table 3). The CO emission values from the NC were well comparable with those reported in other studies of stoves and masonry heaters (e.g. Venkataraman and Uma Maheswara Rao, 2001; Koyuncu and Pinar, 2007;Tissari et al., 2007). Extremely high emissions, CO up to 300 g kg1 and OGC up to 90 g kg1, have been measured from oldtype wood log boilers with large batch size (Johansson et al., 2004), since the SC produced emission factors of

Table 3 Average gaseous emissions from a conventional masonry heater in normal (NC) and smouldering combustion (SC) conditions

NC  SD SC  SD a

CO (g kg1)

NOx (g kg1)

OGC (gC kg1)

CH4 (g kg1)

C2H2 (g kg1)

C2H4 (g kg1)

C3H6 (g kg1)

C4H6 (g kg1)

C6H6 (g kg1)

42  4 148  43

1.4  0.1 1.3a

2.2  0.4 31  14

0.58  0.13 6.3a

0.18  0.06 2.6a

0.27  0.07 4.0a

0.21  0.08 1.2a

0.11  0.03 0.8a

0.33  0.09 3.9a

Only one experiment.

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150 g kg1 CO and 31 g kg1 OGC. The OGC emission from SC was well comparable with that reported in an earlier study: according to Hu¨bner et al. (2005), the OGC emissions from stoves in the field studies have been 1–3 mg MJ1, which corresponds to about 17–56 gC kg1. 4. Conclusions It was found that the operational practice in RWC can significantly influence the quantity and quality of gaseous and particulate emissions into the atmosphere. In masonry heaters, the storage of heat to stone mass makes a low average heat output possible without necessarily leading to smouldering combustion (i.e. a low combustion rate). However, if the gasification process of wood is not controlled, smouldering combustion conditions may also occur in these appliances. In this study, smouldering combustion conditions were found to cause large increases in particle and gaseous emissions, with the largest influence on organic emission. The OGC emission from SC was 14-fold to that of NC. In addition, PM1 from SC was 6-fold and CO 3.5-fold to those of NC. According to SEM and TEM analyses, the observed fine ash particles seemed to occur mainly as separate spherical or irregularly shaped particles but not as agglomerates. Large agglomerates were found to contain mainly carbon and are considered to be primarily soot particles. The separate ultrafine particles were composed mainly of K, S and Zn, but also in a lesser extent of C, Ca, Fe, Mg, Cl, P and Na. The larger spherical and irregularly shaped particles were composed of same alkali metal compounds, but they were probably covered with heavy organic compounds. From SC, particles were composed mainly of carbon compounds. The particles were more irregular and had a more closed (sintered-like) structure than the particles from NC, due to organic matter on the particles. The fine ash mass emissions and particle number emissions from SC were less than half of those from NC. Thus, it seems that the released ash particles may play an important role in the formation of the particle number emission. In the present experiments, the ultrafine mode in the particle number distributions seemed to be determined mainly by the amount of released ash forming material in combustion. Respectively, the shifting of particle size during different combustion conditions seemed to be determined by the amount of condensed organic vapour in the flue gas. In summary, a restriction of air supply and too large fuel batches, which are the common operational errors in log-wood heating, causes significant changes in the particle and gaseous emissions, especially on the organic emissions.

Acknowledgements The authors are grateful for the financial support of the Finnish Funding Agency for Technology and Innovation (Grant no. 40229/05) in the ‘‘small-scale production and use of wood fuels’’ technology programme. We thank Pentti Willman of the University of Kuopio for technical support during this work.

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