Different formation mechanisms of PAH during wood and coal combustion under different temperatures

Journal Pre-proof Different formation mechanisms of PAH during wood and coal combustion under different temperatures Yong Han, Yingjun Chen, Yanli Feng, Wenhuai Song, Fang Cao, Yanlin Zhang, Qing Li, Xin Yang, Jianmin Chen PII:

S1352-2310(19)30723-X

DOI:

https://doi.org/10.1016/j.atmosenv.2019.117084

Reference:

AEA 117084

To appear in:

Atmospheric Environment

Received Date: 9 April 2019 Revised Date:

22 October 2019

Accepted Date: 26 October 2019

Please cite this article as: Han, Y., Chen, Y., Feng, Y., Song, W., Cao, F., Zhang, Y., Li, Q., Yang, X., Chen, J., Different formation mechanisms of PAH during wood and coal combustion under different temperatures, Atmospheric Environment (2019), doi: https://doi.org/10.1016/j.atmosenv.2019.117084. This is a PDF file of an article that has undergone enhancements after acceptance, such as the addition of a cover page and metadata, and formatting for readability, but it is not yet the definitive version of record. This version will undergo additional copyediting, typesetting and review before it is published in its final form, but we are providing this version to give early visibility of the article. Please note that, during the production process, errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain. © 2019 Published by Elsevier Ltd.

1

Different Formation Mechanisms of PAH during Wood and Coal Combustion

2

under Different Temperatures

3

Yong Hanξ,б, Yingjun Chenξ,υ,б*, Yanli Feng£, Wenhuai SongΦ, Fang CaoΦ, Yanlin

4

ZhangΦ,*, Qing Liξ, Xin Yangξ,υ, Jianmin Chenξ

5

ξ

Shanghai Key Laboratory of Atmospheric Particle Pollution and Prevention (LAP3), Department

6

of Environmental Science and Engineering, Fudan University, Shanghai 200438, P.R. China

7

б

8

υ

9

£

State Key Laboratory of Pollution Control and Resources Reuse, Shanghai 200092, China Shanghai Institute of Pollution Control and Ecological Security, Shanghai 200092, P.R. China Institute of Environmental Pollution and Health, School of Environmental and Chemical

10

Engineering, Shanghai University, Shanghai 200444, P.R. China

11

Φ

12

and Environment Change (ILCEC), Nanjing University of Information Science and Technology,

13

Nanjing 210044, China, Jiangsu Provincial Key Laboratory of Agricultural Meteorology, College

14

of Applied Meteorology, Nanjing University of Information Science and Technology, Nanjing

15

210044, China

16

*

17

Yanlin Zhang, [email protected], [email protected]

Yale–NUIST Center on Atmospheric Environment, International Joint Laboratory on Climate

Corresponding authors: Yingjun Chen, [email protected]

18 19 20 21 22 1

23

Abstract:

24 25

Residential solid fuel combustion (RSFC) is a major contributor to polycyclic

26

aromatic hydrocarbons (PAHs) in the atmosphere, which are strongly related to

27

negative health impacts. During RSFC, the variations of PAH emission factors (EFs)

28

and size-resolved profiles are known to be highly affected by fuel type and

29

combustion temperature. In this study, to investigate the behavior of emitted PAH,

30

combustion experiments were performed using three wood and three coal types under

31

different temperatures (500°C and 800°C) in a quartz tube furnace. The results show

32

that the average EFs of PAH (17-EPA-PAHs) from low temperature coal combustion

33

were nearly three times higher than those from low temperature wood combustion.

34

However, with high temperature, PAH emissions from wood combustion increased

35

two-fold and that from coal combustion decreased by two orders. Furthermore, And

36

the proportion of high-molecular-weight PAHs (HPAHs) increased with increasing

37

temperature in wood combustion, but decreased in coal combustion. This indicates

38

that PAH synthesis was the dominant process during wood combustion, while

39

pyrolysis of coal supramolecular structure was the main formation pathway of PAH

40

during coal combustion. In addition, more low-molecular-weight PAHs (LPAHs) were 2

41

emitted, with 0.006 µm – 0.050 µm and 0.223 µm – 1 µm particles in the early

42

burning stage, while more HPAHs were emitted in the later burning stage, with larger

43

particles in the size range of 0.050 µm – 0.223 µm. This means that the PAH

44

formations were different during each burning stage.

45

Key words: PAH, Size distribution, Residential solid fuel combustion, Fuel type,

46

Temperature, Formation mechanism.

47

1. Introduction

48

The emission of polycyclic aromatic hydrocarbons (PAHs) from residential solid

49

fuel combustion (RSFC) is the most significant contributor to air pollution and

50

impacts on human health; and has been reported to cause over 2 million deaths

51

annually (Lim et al. 2012). In particular, RSFC contributes to 53% of the PAHs in the

52

atmosphere in China (Lei et al. 2011, Xu et al. 2006). Thus, the emission factors (EFs)

53

of PAH from RSFC, including that of wood and coal combustion, have been well

54

studied in recent years (Shen et al. 2010, Shen et al. 2013, Chen et al. 2015, Saleh et

55

al. 2014, Czech et al. 2018). However, the PAH EFs values obtained by different

56

studies are highly variable, resulting in significant uncertainty with respect to the

57

estimation of emissions of PAH from RSFC. It therefore remains difficult to

58

understand the substantial variations between different studies.

59

Researchers have attributed this variation in PAH EFs to different combustion

60

conditions, and different forms of fuel (Wang et al. 2013, Xu et al. 2017). For

61

example, previous studies reported that the efficiency of combustion is the most

62

significant influencing factor for the emission of PAH from RSFC (Truesdale et al. 3

63

1982, Zhang et al. 2017, Wang et al. 2016, Shen et al. 2010, Li et al. 2016, Chen et al.

64

2015). Hence, many “improved” stoves have been designed to reduce PAHs emission.

65

Nevertheless, the increased quantities of nano/ultrafine particles, together with a

66

greater amount of toxic matter, have been observed in combustions conducted in some

67

“improved” cookstoves (Just et al. 2013, Zhi et al. 2009, Jetter et al. 2012). This is

68

despite them being more thermally efficient than traditional cookstoves and aiming to

69

reduce the PAH EFs. Similarly, other techniques have also been used to reduce the

70

emission of pollutants such as those that involve changing the form of fuel, for

71

example biomass briquetting and coal briquetting. These methods were successful in

72

reducing the black carbon (BC) and organic carbon (OC) emissions but failed to

73

reduce PAH emissions (Rapp et al. 2016, Sheng et al. 2012, Chen et al. 2015). It is

74

apparent that the influence mechanism of combustion conditions and fuel forms on

75

PAH emissions are unclear.

76

In general, the temperature of combustion and the composition of fuel are two

77

significant factors for PAH emissions in RSFC. Usually, PAH formation is positively

78

related to temperature because the chemical synthesis of PAHs may be enhanced by

79

increasing temperature (Bikau et al. 2010, Pergal et al. 2013). However, PAHs may

80

also be oxidized or transformed into elemental carbon (Thomas et al. 2008) which

81

conversely reduces PAH emissions. With regards to fuel type, various solid fuels have

82

been employed for combustion in Chinese households, with biomass (in the form of

83

wood and crop residues), and coal being two of the most common kinds of solid fuels

84

in China. Usually, emissions of PAH from wood are lower than those from bituminous 4

85

coal. For example, Lee et al. (2005) found the EFs for a total of 17-PAHs from a

86

controlled open fire to be up to 22 mg/kg and up to 8 mg/kg from bituminous coal and

87

wood, respectively. However, higher PAH emissions have been detected from wood

88

combustion than from coal combustion (Ross et al. 2002, Oanh et al. 1999). Their

89

results indicated that the emissions of PAH from coal and biomass combustion may

90

have different pathways during different studies, with various conditions.

91

To understand the variation of PAH emission between different studies, an

92

in-depth investigation of the origin of PAHs from RSFC is urgently needed. PAHs

93

emitted from RSFC may have two origins: 1) the release of inherent PAHs that are

94

originally trapped in the fuel and 2) the chemical formation pathway from PAH

95

precursors (Mastral et al. 2000). The first origin is mostly controlled by the properties

96

of the fuel. For fuels that contain many PAHs or PAH-like structures (such as coal),

97

the evaporation of raw PAHs or the pyrolysis of macromolecule structures were two

98

primary sources of combustion emitted PAHs (Dong et al. 2013). Recently, an article

99

reported that the PAH emissions from the combustion of carbonized fuels were 97 ± 2%

100

lower than from the combustion of raw fuels (Li et al. 2019). The explanation for this

101

finding is that most of the volatile constituents in raw fuels had been removed, which

102

can prove that the inherent PAHs are one of the significant sources. Furthermore, a

103

significant correlation between the particle-bound PAHs emitted during residential

104

coal combustion and the PAHs contained in raw coals (Chen et al. 2015) has also been

105

observed in a solvent extraction experiment. With the second origin, PAHs can be

106

formed from precursors via complex processes, for example, by synthesis reactions 5

107

through hydrogen abstraction/acetylene addition (HACA) or through phenyl

108

addition/cyclization (PAC) mechanisms (Dong et al. 2003, Liu et al. 2009, Böhm et al.

109

2001). For wood and coal combustion, the PAH formation mechanism can be very

110

dissimilar due to their different compositions. Wood mostly constitutes lignin,

111

cellulose, and hemicellulose (Sullivan et al. 2012), while coal contains a large amount

112

of condensed aromatic hydrocarbons which transform from the biomass contents

113

during coalification. Furthermore, the different compositions of wood and coal may

114

have different response characteristics to temperature changes.

115

We therefore, compare the particle bound PAH emissions from wood and coal

116

combustion under different combustion temperatures, and investigate the influence of

117

temperature on particle bound PAH emission and the different PAH formation

118

pathways between wood and coal combustion. In this study, a dilution sampling

119

system (FPS-4000, Dekati Inc, Finland), together with an electrical low-pressure

120

impactor (ELPI+, Dekati Inc, Finland), were used to obtain highly time-resolved

121

particle mass size distributions (PMSDs) and highly size-segregated particle samples.

122

The PMSD was used in the segmentation of the burning stages and then the particle

123

samples were analyzed for elemental carbon and organic carbon (EC and OC) and

124

PAHs offline. To investigate the differences in EC and OC emissions and

125

compositions, we examined EC/OC and PAH contents of different sized particles.

126

This work continues our previous research on RSFC emission pollutants (Han et al.

127

2018) and provides further understanding of PAH formation mechanisms during

128

different fuel combustion. This is meaningful for understanding the variation of PAH 6

129

EFs through different studies.

130

2. Material and Methods

131

2.1. Combustion, dilution sampling, and measurement systems

132

The combustion, dilution sampling, and measurement systems were similar to

133

those in our previous work (Han et al. 2018) and are briefly described in the following:

134

We performed the combustion experiments in a quartz tube furnace that can be heated

135

to 1200°C. This furnace, a vertical quartz tube (Diameter= 60 mm, High=1200 mm),

136

was modified to fit this study with a matched hoist connected to a sample table

137

(Figure S1 in supporting information). Samplers and particle monitors were set

138

downstream of the dilution sampling system (FPS-4000, DEKATI Inc, Finland) and

139

connected by electric rubber tubing to prevent electrostatic losses. To monitor the

140

particle

141

electrical-low-pressure impactor plus (ELPI+, DEKATI Inc, Finland), which can

142

collect the size-resolved particles at the same time as the particle mass and size

143

distributions were obtained. During the whole combustion, a flue gas analyzer

144

(Photon-Ⅱ, Madhur, Italy) was used to monitor the content of CO2, CO, NO, NO2 and

145

SO2 in the flue gas. Clean air was fed as air supply for all combustion tests.

146

2.2. Fuel sampling and experimental design

mass

and

size

distribution,

the

monitor

system

included

an

147

In this study, we tested three types of wood (Pinus Sekiya, Cystinosis indica,

148

Melia azedarach), denoted as PKW, CIW, MAW, and three types of coal from

149

different mines (Shen Hua in Shan Xi Province, Xu Zhou in Jiang Su Province and Yi

150

Ma in He Nan Province, respectively), denoted as SHC, XZC, and YMC, with the 7

151

volatile matter content on a dry and ash-free basis ranging from 30% to 40% (Table

152

S1).

153

The experimental design was similar to that in our previous work (Han et al.

154

2018) and is briefly described in the following: Before each combustion experiment,

155

the furnace was heated to the setting temperature (500°C or 800°C). A 5-g fuel sample

156

was accurately weighed and placed onto the sample table. The sample table was hung

157

in the low temperature zone. When the system passed the air tightness check, the

158

sample table was lifted to the middle of the high temperature zone at a constant speed

159

(0.2 m/s). The clean compressed air was let into the quartz tube from the bottom in a

160

constant velocity (5 L/min) controlled by a mass flow controller (DF-200C, KOFLOC,

161

Japan). The flue gas rose through the quartz tube and was divided into two streams.

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One stream entered the dilution sampling system which was connected to the quartz

163

tube on the top side. The other was sent into the flue gas analyzer (Photon-Ⅱ, Madur,

164

Italy) (1 L/min) for the measurement of CO2, CO, NO, NO2 and SO2. The dilution air

165

was first decontaminated to remove moisture and particles and later heated to 40 °C.

166

The dilution ratio was approximately 10 during the tests. The diluted flue gas was

167

subsequently divided in a flow divider to form three different streams for the monitor

168

and sampling systems. One stream (10 L/min) led to an ELPI+, which could measure

169

the particle number concentration in the diameter range from 6 nm to 9.8 µm for

170

every 6 sec. At the same time, the size segregated particulate samples were obtained

171

by ELPI+. The particles with diameters ranging from 0.015 µm to 9.8 µm was

172

separated into 13 stages according to their diameters. The details are as follows: 8

173

0.0156 µm, 0.0303 µm, 0.0549 µm, 0.0943 µm, 0.153 µm, 0.256 µm, 0.382 µm,

174

0.602 µm, 0.947 µm, 1.63 µm, 2.47 µm, 3.65 µm, 5.36 µm, and 9.88 µm. The second

175

stream (30 L/min) was introduced to a filter sampler to collect the PM onto a quartz

176

fiber filter (QFF) (90 mm in diameter) for PM EFs analysis. The third stream was

177

vented to ambient air to maintain balanced system pressure. All the stream flows were

178

controlled by using a mass flow controller and were checked by a flow meter (Model

179

4043; TSI Incorporated; Shoreview, MN).

180

The average PAHs, EC, and OC emission values from wood combustion, under

181

high/low ignition temperature was denoted as wood-HIT/wood-LIT, and coal

182

combustion under high/low ignition temperature was denoted as coal-HIT/coal-LIT.

183

2.3 Measurements of PM, OC, and EC

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Particle number size distributions were obtained by ELPI+ during combustion.

185

The particle mass distributions were calculated based on the particle number size

186

distributions, and used the aerodynamic diameter as the particle equivalent diameter

187

that was defined as a diameter of a unit density (ρp) spherical particle having the same

188

settling velocity (VTS) as the actual particle. The assumed particle density was 1.2

189

g/cm3, the same as the density of high OC and EC content particles (Cha et al. 2018).

190

Carbon concentrations of quartz fiber filter membrane (QFF) samples were obtained

191

using an improved thermal/optical reflectance (IMPROVEA-TOR) protocol

192

(Atmoslytic Inc. Model 2001A) (Chow et al. 2001). The quartz filter was stepwise

193

heated to 120, 250, 450, and 550 °C in the pure He atmosphere. Next, 2% O2/98% He

194

atmosphere was employed, and the filter was continuously heated stepwise to 550, 9

195

700, and 850 °C for the quantification. The pyrolyzed organic carbon (POC) was

196

determined by measuring the combusted carbon before the laser reflectance signal

197

returned to the initial value. The EC was obtained by subtracting the POC from

198

EC1+EC2+EC3 and OC was obtained by adding POC to OC1+OC2+OC3+OC4. The

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EC1 in the subsequent analysis was defined as EC1-POC, and OC4 was defined as

200

OC4+POC. The emission factors of EC and OC (EFEC and EFOC) for each

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fuel/temperature combination were calculated according to the EC and OC masses of

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the QFF sample combined with sampled fractional ratio and fuel weight with the

203

following formula: EF (g/kg) = c ∗ A ∗ f ∗ 10 /W

204

where c (µg/cm2) is the OC or EC concentration of multiple QFF portions; A is the

205

whole area of QFF (4.9 cm2); f is the fractional ratio of sampled to total emissions;

206

and W (g) is tested coal weight.

207

2.4 Analysis of PAH and quality control

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In this study, we used an Optical-4 thermal desorption (TD) sample injection port

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coupled with an Agilent GC7890B/MS5977A (Agilent Technologies; Santa Clara, CA)

210

system. The TD method was similar to the study reported by Ding et al. (2009) and

211

briefly described here. Quartz filter samples of 8 mm diameter were cut from each 25

212

mm quartz filter and loaded into the TD tube. The tube was then heated to 310°C at a

213

rate of 12°C/min and desorbed at 310°C for 3 min. The desorbed organic compounds

214

were trapped on the head of the GC-column (DB-5MS, 0.25 *0.25*30). The initial

215

GC oven temperature was 40°C, rose to 120°C at a rate of 10°C/min, then rose again 10

216

to 310°C at a rate of 5°C/min and remained at 310°C for 14 min.

217

We identified 17 target PAHs based on retention time and qualified ions of

218

standards, namely 16 US-Environmental Protection Agency (EPA) priority parent

219

PAHs (p-PAHs) and (BeP). The method detection limits (MDLs) ranged from 0.2

220

pg/mm2 for Acenaphthene (Ace) to 0.6 pg/mm2 for Indeno (1,2,3-cdpyrene) (Incdp).

221

Naphthalene-D8,

222

Perylene-D12 were used to check desorption efficiency every 10 detections. The

223

desorption efficiency for all compounds was higher than 90%.

Acenaphthene-D10,

Phenanthrene-D10,

Chrysene-D12,

and

224

All the air used in this study was purified by a nanoparticle filter. Hence, the

225

background particle concentration in the system was relatively low compared with the

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coal combustion-emitted particle (approximately 2-3 orders of magnitudes lower).

227

Furthermore, the background was subtracted from the detected particle number

228

concentration, EC, OC, and PAH results.

229

2.5 Quality control and assurance

230

To evaluate the reproducibility of the combustion and sampling system built in

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our laboratory, three coal combustion experiments and one wood combustion

232

experiment were repeated three times. The reproducibility of coal combustion has

233

been presented in our previous work (Han et al. 2018) and is briefly described,

234

together with a description of that of wood combustion as follows: the coefficient of

235

variation (CV, n=3) of EFPM, for the repeated experiments ranged from 1.9% for SHC

236

at 800°C combustion to 8.9% for SH coal at 500°C combustion and 4.9% for PKW at

237

500°C combustion. All of the CV values were lower than 10%, indicating that the 11

238

combustion technology is acceptable.

239

3. Results and Discussion

240

3.1 Influence of combustion temperature and fuel type on particle bound PAH

241

emissions

242

The EFs of total PAHs from wood and coal combustions were mainly influenced

243

by temperature, but variations of EFs of total PAHs (EFsPAH) were different for wood

244

and coal combustions. The EFsPAH ranged from 6.2 mg/kg in CIW to 31.3 mg/kg in

245

XZC at 500ºC and from 0.1 mg/kg in SHC to 20.4 mg/kg in CIW at 800ºC. During

246

low temperature combustion, coal property is a significant factor for PAH emission.

247

Different coal types in this study have different volatile contents and this has been

248

reported in previous studies to cause a high range of variation of EFsPAH (Chen et al.

249

2005, Chen et al. 2004). Coal-LIT emitted approximately two orders of magnitude

250

more PAHs (19.5±8.8 mg/kg) than coal-HIT (0.2±0.05 mg/kg). For wood combustion,

251

the average PAH emission of wood-HIT was 14.1 mg/kg, two-fold that of wood-LIT.

252

EFsPAH found in this study were comparable to those in the literature, and the

253

variation range of EFsPAH for different ignition temperatures could cover the values

254

obtained in many studies, as shown (Table S2). This suggests that the combustion

255

experiment with different ignition temperatures may be typical of residential solid

256

combustions.

12

257 258

Figure 1. (a) Emission factors of PAH emitted from wood and coal combustions

259

at 500°C and 800°C; (b) ring-resolved PAH groups from wood and coal combustions

260

at 500Ⅱ and 800Ⅱ.

261

PAH emissions in wood combustions were promoted by high temperature, while

262

those in coal combustions were significantly reduced (Fig. 1a). This was comparable

263

with EFsPAH reported in previous laboratory studies (Keshtkar et al. 2007, Peng et al.

264

2016, Chen et al. 2005, Chen et al. 2015, Liu et al. 2009) and it was clear that PAH

265

emission is highly affected by temperature. Other studies report that PAH formation is

266

promoted by high temperatures (Dong et al. 2013, Wen et al. 2016, Liu et al. 2012),

267

which is in agreement with our observations for wood combustions. The majority of

268

PAHs emitted from the combustion of wood barely contained protogenetic PAHs

269

(Hajaigol et al. 2001), suggesting that they were generated through synthetic reactions.

270

The generation rate of PAHs during wood combustion was significantly improved

271

when the combustion temperature was high, which may be due to the abundance of

272

PAH precursors in wood, which can enhance the formation rate. However, the PAHs

273

emitted from coal combustion decreased at high temperatures, perhaps because 13

274

pyrolysis of the coal-condensed aromatic structure, rather than synthetic reactions,

275

was the main formation pathway of PAH (Dong et al. 2013). In addition, the synthesis

276

and cracking were balanced dynamically for PAH emission. Hence, coal combustions

277

emitted more PAHs than wood combustions at low temperatures because most PAHs

278

in raw coal are evaporated during LIT combustions. Nevertheless, during HIT

279

combustion, the crack reaction and transformation of the coal-condensed aromatic

280

structure and synthetic PAHs dominates the emission process. Furthermore, PAHs

281

could also transform into derived-PAHs during combustion. In this study, these

282

conversions were not included, but it would be useful to investigate PAH formation

283

mechanisms in future studies.

284

Ring-resolved PAHs profiles support our hypothesis that wood and coal

285

combustions have different PAH emission mechanisms. Five-ring PAHs were

286

dominant in all the fuel and combustion temperature combinations and six-ring PAHs

287

were the second most abundant compound (Fig. 1b). The difference was that the

288

proportion of high-molecular-weight PAHs (HPAHs) in wood combustions increased

289

with high temperature but decreased in coal combustion. This is in concordance with

290

our hypothesis that the synthesis reaction of PAHs in wood combustion was promoted

291

by high temperature and resulted in the increase of more stable HPAHs. In contrast,

292

the HPAH emissions from low temperature coal combustions were higher because the

293

emission of native PAHs was the largest contributor for coal combustion PAH

294

emission. This is supported by a previous study which reported that HPAHs were the

295

main contributors to the total PAHs in raw coal (Liu et al. 2012). The yield of 14

296

low-molecular-weight PAHs (LPAHs) increased at 800°C. The reaction rate of

297

cracking of the coal-condensed aromatic structure exceeded that of PAH formation at

298

this temperature.

299 300 301

Figure 2. PAH profiles in size segregated particles under different ignition temperature conditions.

302

It is notable that the increase in temperature positively influenced the formation

303

of the examined PAHs molecules. For wood and coal combustion at 500°C, the

304

compound with the highest EFs were benzo(a)pyrene/benzo(e)pyrene (BAP/BEP)

305

(Fig. 2a, b), which concords with the results of previous studies (Hays et al. 2003,

306

Chen et al. 2005). This five-ring molecule was not contained in wood and was a result

307

of the processes occurring during combustion (Hajaligol et al. 2001). It can be

308

synthesized through the hydrogen abstraction/acetylene addition (HACA) or phenyl

309

addition/cyclization (PAC) mechanism. The formation of BAP and BEP is expected to

310

be promoted by high temperatures. The observed data show that BAP and BEP

311

emissions increased when the combustion temperature was 800Ⅱ (Fig. 2a, b). It

312

validates that PAH formation is positively correlated with temperature. Although BAP

313

and BEP emission decreased with high temperature in this study (Fig. 2b), other 15

314

research suggests that it could be also formed during combustion (Homann 1998,

315

Frenklach 2002, Cain et al. 2014, Wang 2011). Hence, emissions of small PAHs such

316

as naphthalene, acenaphthylene, acenaphthene, and phenanthrene increased when the

317

combustion temperature was 800Ⅱ (Fig. 2b). This indicates that PAH formation could

318

also be promoted by high temperature during coal combustion.

319

However, EFsPAH from coal combustion shows that the emission of larger PAHs

320

decreased significantly in high temperature combustion (Fig. 1b). This suggests that

321

the generation of PAH was not the only influencing factor for PAH emissions.

322

Considering the sharp decrease in PAHs during coal combustion, the elimination of

323

PAHs could also influence PAH emissions in solid fuel combustion. There were

324

usually two pathways for PAH elimination: 1) high temperature could promote the

325

oxidation of PAHs, although it can promote the synthesis reaction rate, but result in a

326

decline in PAHs emission, and 2) the high content of aromatics in bituminous coal

327

produces more soot (Shen et al. 2017), which also makes coal-HIT combustion emit

328

less PAHs.

329

3.2 Combustion temperature and fuel type influence on EC and OC emissions

330

and profiles

16

331 332

Figure 3. a) Average emission factors of OC and EC from wood and coal combustions

333

at 500Ⅱ and 800Ⅱ. b) Profiles of EC from wood and coal combustions at 500Ⅱ and

334

800Ⅱ.

335

The EC/OC analysis also supports the observation that the PAHs in bituminous

336

coal were easier to turn into EC, which made high temperature coal combustion emit

337

less PAHs. The total carbonaceous matter EFs decreased one order of magnitude

338

when temperature increased from 500Ⅱ to 800Ⅱ (Fig. 3a). For wood combustion, the

339

average OC and EC EFs decreased from 4621 mg/kg and 1242 mg/kg to 427 mg/kg

340

and 242 mg/kg respectively. For coal combustion, the average OC EFs decreased

341

from 4701 mg/kg to 51 mg/kg, while the EC EFs increased from 303 mg/kg to 343

342

mg/kg.

343

The oxidation rate of carbonaceous matter could have been enhanced when the

344

ignition temperature increased. Thus, the EFs of the total carbonaceous matter

345

decreased but nevertheless EC, which was more stable than OC, showed a significant

346

decline during wood combustion. This indicated that the EC fraction was also 17

347

oxidized when the ignition temperature was 800Ⅱ. However, the EC EFs increased

348

from 303 mg/kg to 343 mg/kg with increasing temperature during coal combustion.

349

This means that not only the elimination but also the generation rate was promoted at

350

high temperatures and supports the observation that PAHs from coal combustion tend

351

to turn into EC under the HIT condition.

352

The different trends for EC emissions from wood and coal combustions indicate

353

that the ECs fraction in the two fuels were not the same. Based on the oxidation

354

temperature, EC is usually separated into three fractions in the IMPOROVA-TOR

355

method: EC1 (550Ⅱ), EC2 (700Ⅱ), and EC3 (850Ⅱ). Figure 3b shows the EC profiles

356

for wood and coal combustion under different ignition temperatures. Wood

357

combustion emitted EC containing more of the EC1 fraction, which can be defined as

358

char-EC according to our previous work (Han et al. 2018). Coal combustion produced

359

more of the EC2 fraction, which can be defined as soot-EC. Hence, the reduction in

360

EC during wood combustion may be due to the oxidation of EC1, which was more

361

active than EC2 and EC3. According to our previous study (Han et al. 2018), the

362

different EC fractions may be formed through different mechanisms during the

363

different stages of fuel combustion. As an EC precursor, PAH emission should also be

364

different during the different burning stages.

365

3.3 Burning stages by time-resolved PMSD of wood and coal combustion

18

366 367

Figure 4. Time resolved particle mass size distributions from wood and coal

368

combustions at 500°C and 800°C.

369

Time-resolved particle mass size distribution reveals that the combustion had

370

different stages, with distinct characters of particle emission and size. For low

371

temperature combustion, particle sizes are close to 0.022 µm in the beginning of the

372

combustion, then they increased rapidly to approximately 1 µm after 150 sec (for

373

wood) and 450 sec (for coal). Thereafter, the particle sizes decreased to approximately

374

0.129 – 0.205 µm to the end of combustion (Figure 4). For the high temperature

375

combustion, the size variation of particles is similar to that from low temperature

376

combustion although the particle emissions decreased one order of magnitude when

377

the temperature increased from 500Ⅱ to 800Ⅱ (Table S2).

378

The variation in particle mass size distributions followed the same trend for

379

every fuel and temperature combination. Based on PMSD (Fig. 4), the burning

380

process can be divided into four stages: 1) ignition, 2) first flaming, 3) second flaming, 19

381

and 4) burnout. The ignition stage produces a very high average number concentration

382

of particles with a wide size range from 0.006 µm – 1 µm. Small particles within the

383

range of 0.04 µm – 0.223 µm were the most abundant particles during this stage. In

384

the second stage, large particles with diameters within 0.223 µm – 1 µm were

385

dominant. Particles emitted at the third stage had a similar size distribution to that of

386

the first stage, but with lower number concentration. At the fourth stage, particles

387

were barely generated.

388

Generally, particle size is positively correlated with volatile matter content. Thus,

389

a higher gasification rate of fuel in stage 2 may result in a higher concentration of

390

pyrolysis gas in the flaming zone, which may lead to an increase in particle size. This

391

result is consistent with those of previous studies that reported particles considerably

392

forming during intense flaming combustion due to the less efficient transport of

393

oxygen into the interior flame zone.Growth in the size of particles was fast because

394

the coagulation rate of particles is roughly proportional to the square of their number

395

concentration (Pan et al. 2017). Thus, later burning stages generated particles that

396

were small due to the lack of volatiles which were consumed in stage 2.

397

3.4 PAH emissions in size–resolved particles under different temperatures

398 20

399

Figure 5. Percentage of total PAH emissions in size resolved particles from wood and

400

coal combustions at 500°C and 800°C temperatures.

401

Based on particle size, the combustion process can be divided into different

402

stages which may also have different characteristics of PAH emission (Fig. 5). For

403

wood combustion, the size distributions of total particulate phase PAHs were

404

unimodal, with peak values in 0.354 µm – 0.585 µm and 0.121 µm – 0.223 µm size

405

bins for wood-LIT and wood-HIT, respectively. For coal combustion, the peak value

406

for EFs of total particulate phase PAHs occurred within 0.585 µm – 0.946 µm and

407

0.354 µm – 0.585 µm size bins for coal-LIT and coal-HIT, respectively. The highest

408

total particulate phase PAH emission occurred at the second burning stage with larger

409

particles.

410

3.5. Ring-resolved PAH profiles in size resolved particles: implications for PAH

411

formation during different burning stages

412 413

Figure 6. Mass percent of particulate bound PAHs in size resolved particles for a) 21

414

wood HIT combustion, b) wood LIT combustion, c) coal HIT combustion, and d) coal

415

LIT combustion.

416

Similarly, profiles of particulate bound PAHs are distinct characteristics for

417

different burning stages, suggesting that the emission mechanism of PAH may be

418

different at each burning stage. The size distribution of ring resolved particulate

419

bound PAH profiles is presented in Figure 6. For wood combustion, bimodal

420

distributions for LPAHs with 2 – 4 rings were observed in the particles with peak

421

values at 0.010 µm and 3 µm. In addition, unimodal distributions for HPAHs with 5 –

422

6 rings were observed in the particles with peak value at 0.100 µm. Similarly, bimodal

423

distributions for LPAHs were observed in coal combustion, but with peak values

424

occurring at the particle diameter approximately 0.010 µm and 0.300 µm. The peak

425

value for HPAHs occurred in particle sizes at 0.068 µm – 0.121 µm for coal HIT

426

combustion and 0.121 µm – 0.223 µm for coal LIT combustion.

427

The proportion of HPAHs did not always increase with decreasing size but did

428

with peak values within 0.100 µm – 0.200 µm. The size distribution of individual

429

particulate bound PAH compound in particles between fine (<2.1 µm) and coarse (2.1

430

µm –10 µm) particles was such that the smaller the particle was, the higher the

431

content of HPAHs from wood and crop residue burning. This phenomenon may be

432

explained by the following processes (Keshtkar et al. 2007, Kawanaka et al. 2009,

433

Hays et al. 2003):

434

(1) the difference in diffusivity that is correlated with molecular weight;

435

(2) enhanced vapor pressures of lower molecular weight compounds that have 22

436

higher volatilization rates from small particles, and

437

(3) more organic matter in larger particles may enhance absorption.

438

In general, the less volatile five and six-ring PAHs were present predominantly

439

on smaller particles where they condensed in the early stages of combustion, while the

440

more volatile three- and four-ring PAHs formed on larger particles as the smoke

441

cooled. Condensation is more likely to occur with fine particles because they can

442

provide a bigger specific surface area. Nevertheless, LPAHs were inhibited on small

443

PM according to the Kelvin effect.

444

Based on our observations, the burning period was another important factor for

445

PAH formation. Stage 2 emitted particles (0.223 – 1 µm) contained more LPAHs

446

while stage 3 particles (0.040 – 0.223 µm) contained more HPAHs, indicating that the

447

second burning stage emitted more LPAHs, while the third stage emitted more

448

HPAHs. As discussed above, large particles (>0.223 µm) were mainly generated in the

449

flaming stage through gas-phase reaction which is usually promoted by fuel rich

450

flame condition with many LPAHs. This agrees with EC/OC analysis for size

451

segregate particles (Figure S2 in supporting information) which demonstrated that

452

large particles contain significantly more soot-EC formed through the gas-phase

453

HACA mechanism (Wang 2011, Frenklach et al. 1986, Xu et al. 2017). However, for

454

the third burning stage, most of the particles were in the range of 0.05 µm – 0.223 µm,

455

which corresponds to more HPAHs. This can be due to the lack of volatile matter

456

content during this stage. However, fixed carbon was mostly burned in this stage,

457

which makes the fuel temperature higher than that in the former burning stage, 23

458

resulting in a higher generation rate of HPAHs (Liu et al. 2000, Pergal et al. 2013).

459

The higher fuel temperature can be confirmed by the emission of SO2 which is usually

460

generated by the oxidation of pyrite above 727°C (Zhang et al. 2016). The SO2

461

concentration exhibited a peak in the middle of this stage, indicating that the

462

combustion temperature increased to its peak, and large amounts of inorganic

463

compounds such as sulfate were produced to form particles smaller than 0.03 µm

464

(Kulmala et al. 2007, Andreae 2013) (Figure S3 in supporting information). With the

465

exception of temperature, the modified combustion efficiency (MCE) is another

466

important factor for PAH emissions. As expected, the MCE value for stage 2 could be

467

as low as 0.6, 30% lower than stage 3. This indicates that combustion of stage 2 is

468

relative insufficient, resulting in the large amount of EC and PAH generation. During

469

stage 3, the MCE value substantially increased, which means most of the fuel was

470

burned into CO2, resulting in the lower EC and PAH emissions.

471

4. Conclusions and Implications

472

Although this study was performed under controlled combustion conditions with

473

fixed air supply and set temperatures that may not be identical to real-world solid fuel

474

combustion, it benefits from the simplified combustion scenario and some actual solid

475

fuel combustion processes are revealed.

476

We found that fuel type was an important influencing factor for PAH emissions

477

from residential combustion even when the temperatures are lower than those in

478

industrial combustion. We observed significantly different emission characteristics for

479

wood and coal due to their different compositions. The PAH emissions from 24

480

low-temperature coal combustion were four times higher than that from wood

481

combustion. Nevertheless, wood combustion generated PAHs that contain

482

significantly more HPAHs than coal combustion.

483

In addition, PAH emissions depend mainly on combustion temperature. High

484

temperature can reduce the PAH emissions significantly during coal combustion,

485

while during wood combustion, PAH emissions may be significantly promoted. This

486

indicates a possibility of different mechanisms for PAH emissions from wood and

487

coal combustion. Synthesis reactions may be the main source of PAH emitted from

488

wood combustion, while pyrolysis of the coal supramolecular structure may be the

489

main source of PAH from coal combustion. Hence, the synthesis rate can be enhanced

490

at high temperatures, increasing the PAH emission from wood combustion. However,

491

the formation of PAH is not the only influencing factor for PAH emission. Elimination

492

is also an important factor which can significantly reduce PAH emission. Usually, this

493

process includes the oxidation and transformation of PAHs. During elimination, PAHs

494

may be oxidized into COx and oxide PAHs or be converted into elemental carbon. The

495

conversion of PAHs to elemental carbon was potentiated because of the protogenetic

496

PAHs contained in coal, resulting in the decrease in PAH emission along with an

497

increase in elemental carbon during coal combustion.

498

Furthermore, PAH emissions changed during different combustion processes.

499

During the fast combustion stage with flame, a large amount of LPAHs were

500

synthesized through the HACA pathway because of the relatively low oxygen level,

501

which can be proven by the increasing of EC emission observed during this stage. 25

502

Large quantities of HPAHs were synthesized during the later burning stage as the

503

combustion temperature increased.

504

Although this study is insufficient to understand PAH formation mechanisms

505

which include many chemical reactions, we have found that the formation of PAHs

506

for different fuel combustion may have different pathways, and temperature is an

507

important factor for PAH formation. In future research, we should focus on the PAH

508

emissions from residential solid fuel combustion under more temperature conditions,

509

to determine at which temperature the PAH emission is the lowest for different fuel,

510

which is important for emissions reductions. Solid surface temperature should also be

511

included in future studies because it can help us to understand PAH formation during

512

different burning stages. Furthermore, numerous organic compounds such as alkanes,

513

methyl-PAHs, oxygenated PAHs, and nitrated PAHs, which can help understand the

514

formation process from that of small hydrocarbons to PAHs and elemental carbon,

515

should be considered.

516

Acknowledgement

517

This study was financially supported by the National Natural Science Foundation

518

of China (91744203, 41761134083, 41877371, 41473091).

519

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30

Quantification of PAHs in highly size resolved particles from 0.01 µm to 10 µm. PAHs from wood increase with temperature increasing while it is opposite from coal. PAHs emissions have different characteristics during each combustion processes. The formation pathway of PAHs during wood and coal combustions are different.

Declaration of interests ☒ The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. ☐The authors declare the following financial interests/personal relationships which may be considered as potential competing interests: