Particulate hydroxy-PAH emissions from a residential wood log stove using different fuels and burning conditions

Published Date
September 2016, Vol.140:1–9, doi:10.1016/j.atmosenv.2016.05.041
Open Access, Creative Commons license

Title 
Particulate hydroxy-PAH emissions from a residential wood log stove using different fuels and burning conditions

• Author 

• Rozanna Avagyan ^a
• Robin Nyström ^b
• Robert Lindgren ^b
• Christoffer Boman ^b
• Roger Westerholm^a,,

• aDepartment of Environmental Science and Analytical Chemistry, Arrhenius Laboratory, Stockholm University, SE-106 91 Stockholm, Sweden
• bDepartment of Applied Physics and Electronics, Thermochemical Energy Conversion Laboratory, Umeå University, SE-901 87 Umeå, Sweden

Received 18 November 2015. Revised 20 May 2016. Accepted 23 May 2016. Available online 24 May 2016.

Highlights

• •
  Particulate OH-PAHs from residential wood log stove combustion were determined.
• •
  Four different types of wood logs and two combustion conditions were examined.
• •
  Pine burning in high burn rate had largest impact on particulate OH-PAH emissions.
• •
  OH-PAH nominal burn rate emissions correspond to 15% of high burn rate emissions.
• •
  Emissions of OH-PAHs correspond on average to 32% of PAH emissions.

Abstract

Hydroxylated polycyclic aromatic hydrocarbons are oxidation products of polycyclic aromatic hydrocarbons, but have not been studied as extensively as polycyclic aromatic hydrocarbons. Several studies have however shown that hydroxylated polycyclic aromatic hydrocarbons have toxic and carcinogenic properties. They have been detected in air samples in semi urban areas and combustion is assumed to be the primary source of those compounds. To better understand the formation and occurrence of particulate hydroxylated polycyclic aromatic hydrocarbons from residential wood log stove combustion, 9 hydroxylated polycyclic aromatic hydrocarbons and 2 hydroxy biphenyls were quantified in particles generated from four different types of wood logs (birch, spruce, pine, aspen) and two different combustion conditions (nominal and high burn rate). A previously developed method utilizing liquid chromatography – photo ionization tandem mass spectrometry and pressurized liquid extraction was used. Polycyclic aromatic hydrocarbons were analyzed along with hydroxylated polycyclic aromatic hydrocarbons. The hydroxylated polycyclic aromatic hydrocarbon emissions varied significantly across different wood types and burning conditions; the highest emissions for nominal burn rate were from spruce and for high burn rate from pine burning. Emissions from nominal burn rate corresponded on average to 15% of the emissions from high burn rate, with average emissions of 218 μg/MJ_fuel and 32.5 μg/MJ_fuel for high burn rate and nominal burn rate, respectively. Emissions of the measured hydroxylated polycyclic aromatic hydrocarbons corresponded on average to 28% of polycyclic aromatic hydrocarbons emissions.

This study shows that wood combustion is a large emission source of hydroxylated polycyclic aromatic hydrocarbons and that not only combustion conditions, but also wood type influences the emissions of hydroxylated polycyclic aromatic hydrocarbons and polycyclic aromatic hydrocarbons. There are few studies that have determined hydroxylated polycyclic aromatic hydrocarbons in emissions from wood combustion, and it is therefore necessary to further investigate the formation, occurrence and distribution of these compounds as they are present in significant amounts in wood smoke particles.

Keywords

• OH-PAHs
• Hydroxy-PAHs
• PAHs
• Wood combustion
• Wood burning
• Wood log stove

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Abbreviations

1-8-DHAQ

1,8-dihydroxy anthraquinone

1-HP

1-hydroxy pyrene

1-N

1-naphtol

2-HBP

2-hydroxy biphenyl

2-HPh

2-hydroxy phenanthrene

2-H-9-F

2-hydroxy-9-fluorenone

2-N

2-naphtol

3-HBaP

3-hydroxy benzo[a]pyrene

3-HPh

3-hydroxy phenanthrene

4-HBP

4-hydroxy biphenyl

6-HC

6-hydroxy chrysene

ACN

Acetonitrile

Ant

Anthracene

APPI

Atmospheric pressure photo ionization

BaP

Benzo[a]pyrene

Chr

Chrysene

R²

Coefficient of determination

Flu

Fluoranthene

GC

Gas chromatograph

HB

High burn rate

HPLC

High performance liquid chromatography

OH-PAHs

Hydroxylated polycyclic aromatic hydrocarbons

IS

Internal standards

LOD

Limit of detection

LOQ

Limit of quantification

LC

Liquid chromatography

MDL

Method detection limit

MQL

Method quantification limit

MTBE

Methyl tert-butyl ether

NB

Nominal burn rate

Phe

Phenanthrene

PAHs

Polycyclic aromatic hydrocarbons

PLE

Pressurized liquid extraction

PTV

Programed temperature vaporizer

Pyr

Pyrene

RSD

Relative standard deviation

SRM

Single reaction monitoring

CRM

Certified reference materials

MS/MS

Tandem mass spectrometry

UPLC

Ultra performance liquid chromatography

1 Introduction

Hydroxylated polycyclic aromatic hydrocarbons (OH-PAHs) are oxidation products of polycyclic aromatic hydrocarbons (PAHs), environmental contaminants with carcinogenic properties (el-Bayoumy, 2008, Boffetta et al., 1997, Mastrangelo et al., 1996 and Rantanen, 1983). OH-PAHs have not been studied as extensively as PAHs, however, several studies have shown that they are potentially more toxic and carcinogenic than native PAHs (Yu, 2002, Wang et al., 2009, Xu et al., 2010, Kazunga et al., 2001 and Traczewska, 2000). Furthermore, several OH-PAHs have been shown to have estrogenic and anti-estrogenic activities (Kamiya et al., 2005, Hayakawa et al., 2007 and Hirose et al., 2001). The sources of OH-PAHs in the urban environment are incomplete combustion processes, microbiological and/or photochemical degradation of PAHs (Vione et al., 2004), while in the body OH-PAHs are formed as metabolites of PAHs (Boström et al., 2002). OH-PAHs have low volatility and tend to be associated with particles (Vione et al., 2004). They have been detected in both PM₁₀ and PM_2.5 samples collected from semi urban areas. The concentrations vary between 0 and 200 pg/m³, with average concentrations of around 50 pg/m³ (Barrado et al., 2012, Barrado et al., 2013, Wang et al., 2007 and Kishikawa et al., 2004). Although particle emissions from combustion are assumed to be the primary source of OH-PAHs in urban air and the environment, there are few studies in the literature that have determined OH-PAHs in wood smoke particles (Cochran et al., 2013) and investigated the formation mechanisms during combustion (Simoneit et al., 2007 and Bi et al., 2008).

It has however been shown that the emissions of carbonaceous matter, i.e. soot and organic compounds (e.g. PAHs) from small scale biomass combustion are strongly linked to combustion conditions and also can vary with different wood types used (Orasche et al., 2012, Pettersson et al., 2011, Lamberg et al., 2011 and McDonald et al., 2000). In addition, it has been illustrated that the emission profile of PAHs during different burning phases under batch-wise combustion of wood are significantly influenced by the detailed combustion conditions in a stove (Eriksson et al., 2014). Consequently, the aim of this study was to gain a better understanding of the formation and occurrence of particulate OH-PAHs from residential wood log stove combustion using different fuels and burning conditions. A previously developed liquid chromatography – photo ionization tandem mass spectrometry (LC-APPI-MS/MS) method (Avagyan et al., 2015) utilizing pressurized liquid extraction (PLE) was used for simultaneous determination of 9 OH-PAHs and 2 hydroxy biphenyls in particles generated from four different types of wood logs (birch, spruce, pine, aspen) and two different combustion conditions. Furthermore, 45 PAHs were analyzed along with the OH-PAHs.

2 Materials and methods

2.1 Chemicals and solvents

Detailed information on OH-PAH and PAH standards used in this study are presented in Table S1 in the Electronic Supplementary Material (ESM). Methyl tert-butyl ether (MTBE), hexane, acetonitrile (ACN), acetone, toluene and methanol, all HPLC-grade, were purchased from Rathburn Chemicals Ltd (UK). Anhydrous dodecane (>99%) was purchased from Sigma-Aldrich (USA), anisole (99%) from Merck-Schuchardt (Germany) and ammonium acetate (analysis grade) from Merck (Germany). Water was purified using a Millipore Synergy 185 (Millipore Corp., USA) water purification system equipped with a Millipak 0.22 μm membrane filter (Millipore Corp., USA).

2.2 Combustion appliance, fuels and burning conditions

Four different wood types were combusted in a common Nordic wood stove (manufactured 1994), operated in a laboratory setup that enabled controlled combustion and sampling conditions. The stove was a natural draft wood stove with a nominal heat output of 9 kW and has been used and described in other studies (Pettersson et al., 2011 and Eriksson et al., 2014). The inside of the stove was lined with 5 tiles of soapstone, each tile having an approximate thickness of 25 mm and a height of 0.27 m. A sketch of the stove with dimensions can be found elsewhere (Pettersson et al., 2011). The wood fuels used were; aspen, birch, spruce and pine, all with a moisture content of approximately 12–13%. In the setup, a flue gas fan was used to regulate the under pressure and maintain the same chimney draft regardless of outside influences such as varying weather. This ensured stable combustion conditions with the same airflow throughout the whole combustion cycle, thus ensuring higher repeatability of the experiments. The stove was operated in two different combustion modes, i.e. nominal burn rate (NB) and high burn rate (HB). The recommendation for combustion from the manufacturer was used for NB. Each batch fired consisted of three wood logs with a combined weight of approximately 2.5 kg (moist wood), all logs roughly of the same size. The HB was achieved by following the same recommendations as NB, but slightly overloading the stove with more and smaller logs without exceeding the probable variations during real-life operation. The total weight of the batch was in this case approximately 3.5 kg (moist wood), divided over 5 smaller logs of roughly the same size (smallest approximately 0.5 kg each). The average burn rates were 2.1–2.7 and 4.0–4.2 kg/h for NB and HB, respectively. The most important difference between NB and HB was that during the HB mode episodes of hot and air-starved combustion occurred, that have been shown to cause elevated emissions of soot and PAHs (Eriksson et al., 2014, Fine et al., 2004 and Rau, 1989). In both cases (NB and HB), each batch of wood was added to glowing embers and at least one previous batch of the same fuel was fired beforehand to avoid “cold start conditions”. The sampling times, amount of particles collected and O₂, CO and NO levels during NB and HB conditions are summarized in Table S2 in ESM.

2.3 Particle sampling for OH-PAH and PAH analysis

Sampling was performed during the whole batch, starting slightly before the addition of logs to the embers proceeding until the last flame had extinguished. This resulted in varying sampling times for the different combustion situations tested, i.e. combustion modes and fuel used, as the burnout times differed (Table S2 in ESM). For each combustion case, triplicate particle samples were collected during three succeeding fuel batch additions.

The particulate matter was sampled in the flue gas channel approximately 2.5 m from the top of the stove at an approximate flue gas temperature of 250–300 °C. A porous tube diluter was used (Lyyranen et al., 2004) to collect material for subsequent analysis of OH-PAHs on 90 mm glass fibre filters (Munktell Filter AB, Sweden). The average dilution rate was around 4 times and although the dilution was rather low, care was taken to avoid sampling temperatures exceeding 40 °C. Furthermore, sampling on 47 mm tissue quartz filters (Pall Corporation, Port Washington, NY, USA) was performed for PAH analysis. Two Dekati ejector diluters were used in this sampling line and the dilution rate was around 100 times.

2.4 Sample extraction and clean-up

The analytical methods for OH-PAHs and PAHs are described in detail elsewhere (Avagyan et al., 2015 and Ahmed et al., 2015). Briefly, filters with collected particulate samples were placed in stainless steel extraction cells, the deuterated internal standards (IS) (phenanthrene-d₁₀, pyrene-d₁₀, benz[a]anthracene-d₁₂, benzo[a]pyrene-d₁₂, benzo[ghi]perylene-d₁₂, dibenzo[a,i]pyrene-d₁₄ for PAHs and 1-hydroxy pyrene-d₉ for OH-PAHs) were added and the samples were extracted using an ASE 200 accelerated solvent extraction system (Dionex Corporation, USA). OH-PAHs were extracted with methanol at 200 °C and 2000 psi in two 10 min extraction cycles, with a pre-heating time of 3 min, equilibration time of 9 min, purge with nitrogen in 1 min and 30% solvent flush. PAHs were extracted with toluene at 200 °C and 2000 psi in three 30 min extraction cycles, no pre-heating, an equilibration time of 9 min, purge with nitrogen in 1 min and 30% solvent flush. The extracts were then reduced in volume under a gentle stream of nitrogen in a TurboVap^® LV evaporator (Zymark Corporation, USA) at 65 °C to a final volume of approximately 0.5 mL.

Silica cartridges (100 mg Isolute, Biotage Sweden) were used for sample clean-up. They were conditioned with 3 mL hexane before 0.5 mL of the sample extract was added. The cartridges were washed with 0.5 mL hexane and the OH-PAHs were eluted with 3 mL methanol while PAHs were eluted with 3 mL hexane. The eluted fractions were evaporated under a gentle stream of nitrogen to 0.5 mL and filtered by 0.2 μm Nylon filters (SUN-SRi, USA) prior the analysis. Blank filters with added IS, were extracted, cleaned up, and analyzed same way as the real samples.

2.5 Instrumental analysis

The analyses of OH-PAHs were performed on an ACQUITY ultra performance liquid chromatography (UPLC) system coupled to a Xevo TQS tandem mass spectrometer (MS/MS) (Waters, Milford, MA, USA). The chromatographic separation was performed on an ACQUITY UPLC HSS T3 C₁₈ column (100 × 2.1 mm, 1.8 μm particle size) (Waters, Ireland). Mobile phase A was 20 mM ammonium acetate and mobile phase B was ACN. The flow was 0.5 mL/min, the column temperature was set to 60 °C, and the injection volume was 5 μL. The gradient program was as follows: 0.0 min 20% B, 2.5 min 60% B, 4 min 90% B, 5 min 90% B, 5.1 min 20% B. The system was equilibrated for 3 min with 20% B after each run.

The MS was equipped with an APPI source and was operated in negative ionization mode. The repeller voltage was set to 1.01 kV, the APPI probe temperature was 550 °C and the desolvation and cone gas flows (nitrogen) were set to 250 and 150 L/h, respectively. The collision gas (argon) pressure flow was set to 0.15 mL/min and the nebulizer gas flow to 4 bar. The dopant (flow: 20 μL/min), comprising 50:50 toluene/anisole was infused into the LC eluent flow using the fluidics system of the instrument.

Data was acquired using single reaction monitoring (SRM). The identification of the analytes was based on comparison of SRM transitions and retention times with pure standard solutions. The analyte specific parameters and SRM transitions are shown in Table S3 in ESM.

For the PAH analysis, an online liquid chromatograph coupled to gas chromatograph mass spectrometer (LC–GC/MS) system was used consisting of a CMA/200 microsampler (CMA Microdialysis AB, Sweden), an HPLC pump (Varian Inc, Palo Alto, CA, USA), a UV detector (SPD-6A, Shimadzu, Japan), with a normal phase LC column (Nucleosil 100-5NO2 124 × 4.6 mm, 5 μm) and a GC/MS system consisting of an Agilent 6890N GC (Agilent Technologies, Palo Alto, CA, USA) with an Agilent 5973N MSD (Agilent Technologies). After the HPLC clean-up step that removed alkanes and monoaromatic compounds utilizing back flush technique, the PAH fraction was transferred online through a transfer line to the programed temperature vaporizer (PTV) GC injector. A detailed description of the system setup, operation and parameters can be found elsewhere (Christensen et al., 2005, Bergvall and Westerholm, 2006 and Sadiktsis et al., 2014).

2.6 Quality assurance/quality control (QA/QC)

The analytical method for OH-PAHs was validated as described in detail elsewhere (Avagyan et al., 2015). The method detection limits (MDL) and method quantification limits (MQL) ranged from 1.6 to 9.6 pg injected and 5.2–32 pg/g, respectively, and the coefficients of determination (R²) of all calibration curves were ≥0.99. The extraction recoveries and the relative standard deviations (RSDs) of repeated experiments were determined in a previous study (Avagyan et al., 2015) and ranged from 70 to 102% and 2–12%, respectively. MDL, MQL, as well as recoveries and repeatability are shown in Table S4 in ESM.

Validation procedure and data of PAH analysis have been reported in detail in other studies (Schantz et al., 2012, Sadiktsis et al., 2014 and Ahmed et al., 2015). Briefly, limits of detection (LOD) and limits of quantification (LOQ) ranged from 2.1 to 14 pg and 7.1–49 pg, respectively, and the R² of all calibration curves were ≥0.98. The extraction efficiency has been determined using certified reference materials (CRM); 1649a, 1650b and 2975 (Ahmed et al., 2015), from the National Institute of Standards and Technology (NIST, Gaithersburg, MD, USA).

3 Results and discussion

3.1 Sampling artefacts and gas – particle partition

Filter sampling is one of the most common ways of collecting particle associated PAHs. The technique is however suffering from a number of sampling artefacts, e.g. the concentrations of PAHs associated with particles can be underestimated due to volatilization of PAHs from the particles to the gaseous phase. The gas-phase PAHs can also adsorb on the filter substrates and particles collected on the filter, causing overestimation of the concentrations (Gundel et al., 1995). Furthermore, the different sampling methods used in this study for the collection of particles for PAH and OH-PAH analysis might also have introduced biases into the sampling process. However, larger amount of particles needed for OH-PAH analysis required 90 mm filters (compared to 47 mm for PAH analysis) and higher sampling flows, as the sampling was performed during one whole wood burning batch (from fuel addition until the last flame had extinguished) and each sampling period was limited in time.

Most of the two- and three- ring PAHs have high vapor pressures, are volatile and/or semi volatile compounds and can be present in the gaseous phase, while the OH-PAHs have lower volatility and vapor pressures and are usually associated with particles (Vione et al., 2004). One of the smallest OH-PAHs quantified in this study, 2-N, has a vapor pressure of 0.00032 mm Hg at ambient temperature (Bidleman, 1988), which can be compared with the corresponding PAH, naphthalene with vapor pressure of 0.085 mm Hg (Ambrose et al., 1975) and usually found in the gaseous phase (thus not quantified in this study). However, the gas – particle partitioning and vapor pressures are only valid at ambient temperature and can change during combustion at very high temperatures. When collecting combustion particles, the partition of the semi volatile PAHs also depends on the dilution of the combustion exhaust. In hot and undiluted exhaust, the semi volatile PAHs are present in the gaseous phase. When the combustion exhaust is diluted with clean air at ambient temperature, the temperature of the sample decreases, affecting the equilibrium towards partition on particles. This might also be the case with OH-PAHs, although they have low vapor pressures at ambient temperatures, they can be vaporized to the gaseous phase at high temperatures. However, the relative low dilution (∼4 times) in this study, the sharp temperature decrease at the sampling point most probably enhanced partitioning of the semi volatile OH-PAHs to the particles (condensation).

3.2 The effect of burning conditions on OH-PAH and PAH emissions

The OH-PAH and hydroxy biphenyl (for simplicity, presented together with OH-PAHs) emissions, varying significantly across different wood types and burning conditions, are summarized in Table 1 and Fig. 1 and Fig. 2 (in μg/MJ_fuel). The emissions from NB samples were generally lower, corresponding on average to 15% of the emissions from HB samples, with average emissions of 218 μg/MJ_fuel and 32.5 μg/MJ_fuel for HB and NB samples, respectively. The most abundant compound for NB samples was 2-naphthol (2-N) with an average emission of 7.66 μg/MJ_fuel, followed by 1-naphthol (1-N) and 1-hydroxy pyrene (1-HP) with average emissions of 4.49 and 4.42 μg/MJ_fuel, respectively. Conversely, 1-HP was the most abundant compound in the HB samples (64.7 μg/MJ_fuel), followed by 2-N (34.8 μg/MJ_fuel). 3-hydroxy benzo[a]pyrene (3-HBaP) was the sixth and third most abundant compound in NB and HB samples, (3.18 and 28.9 μg/MJ_fuel, respectively). While 1-N was the second most abundant compound in NB samples, it was only the fourth most abundant compound (25.6 μg/MJ_fuel) in HB samples. 3-hydroxy phenanthrene (3-HPh) was most abundant of the hydroxylated phenanthrenes, the fifth and third most abundant compound in NB and HB samples, respectively (3.56 and 25.9 μg/MJ_fuel, respectively). While 2-hydroxy biphenyl (2-HBP) was one of the less abundant compounds in both NB and HB samples (0.80 and 1.10 μg/MJ_fuel, respectively), the average emissions of 4-hydroxy biphenyl (4-HBP) were 3.80 μg/MJ_fuel in NB samples and 13.9 μg/MJ_fuel in HB samples. Regarding hydroxy-quinones, 2-hydroxy-9-fluorenone (2-H-9-F) was one of the more abundant compounds with average emissions of 2.64 and 10.0 μg/MJ_fuel, while 1,8-dihydroxy anthraquinone (1-8-DHAQ) was one of the less abundant compounds (0.26 and 0.88 μg/MJ_fuel) in NB and HB samples, respectively. The least abundant OH-PAH in NB and second least abundant in HB samples was 6-hydroxy chrysene (6-HC) (0.18 and 1.40 μg/MJ_fuel), respectively.

Table 1. Average emissions and standard deviations in brackets of OH-PAHs (μg/MJ_fuel) for the different wood types and burning conditions (n = 3).

	Nominal burn rate				High burn rate
	Birch	Aspen	Spruce	Pine	Birch	Aspen	Spruce	Pine
1-8-DHAQ	0.17 (<0.01)	0.14 (0.05)	0.58 (0.24)	0.16 (0.03)	0.45 (0.09)	1.08 (0.14)	1.03 (0.14)	0.96 (0.13)
3-HBaP	2.14 (0.29)	1.42 (0.03)	8.28 (0.85)	0.90 (0.17)	4.02 (1.16)	4.51 (1.31)	8.11 (2.32)	98.8 (17.2)
2-HBP	0.54 (0.03)	0.42 (0.16)	0.85 (0.08)	1.39 (0.27)	0.66 (0.17)	0.00 (0.00)	1.77 (0.40)	1.97 (0.15)
4-HBP	2.56 (0.08)	2.32 (0.36)	6.69 (0.70)	3.63 (0.12)	4.68 (0.63)	8.80 (1.89)	10.5 (2.40)	31.7 (3.86)
6-HC	0.09 (0.01)	0.06 (0.01)	0.44 (0.09)	0.10 (0.04)	0.21 (0.04)	0.55 (0.16)	0.29 (0.10)	4.54 (1.11)
2-H-9-F	1.29 (0.12)	1.53 (0.17)	6.07 (0.86)	1.68 (0.41)	5.34 (0.60)	11.1 (2.26)	10.4 (1.73)	13.3 (0.22)
1-HP	3.10 (0.21)	2.87 (0.50)	9.11 (1.09)	2.61 (0.14)	13.6 (3.38)	43.2 (10.8)	29.2 (6.25)	173 (31.1)
2-HPh	0.73 (0.11)	0.94 (0.09)	3.70 (0.60)	0.73 (0.08)	4.23 (0.39)	8.86 (2.74)	4.91 (0.96)	26.4 (1.42)
3-HPh	1.70 (0.25)	2.15 (0.22)	8.63 (1.41)	1.77 (0.18)	9.92 (0.90)	20.7 (6.40)	11.5 (2.23)	61.5 (3.33)
1-N	0.80 (0.31)	1.30 (0.05)	14.7 (2.99)	1.11 (0.47)	1.13 (0.31)	2.75 (0.42)	2.06 (0.17)	96.4 (4.62)
2-N	1.73 (0.26)	3.63 (0.40)	23.8 (5.08)	1.52 (1.12)	3.60 (0.64)	7.86 (2.29)	11.0 (4.49)	117 (8.62)
Tot. OH-PAH	14.8	16.8	82.8	15.6	47.8	109	90.8	625
Phe/Tot.OH-Phe	0.30	0.59	0.39	0.87	0.24	0.17	1.09	0.71
Phe/2-HPh	1.00	1.95	1.30	2.97	0.79	0.56	1.56	2.36
Pyr/1-HP	1.20	3.10	1.58	2.77	1.01	0.59	1.92	0.25
Chr/6-HC	63.0	111	25.5	70.1	78.5	100	137	58.6
BaP/3-HBaP	1.27	2.52	0.70	5.90	2.39	4.34	3.42	1.81

Fig. 1. Relative (%) OH-PAH and PAH emissions from different wood types and burning conditions (n = 3).

Fig. 2. OH-PAH emissions (μg/MJ_fuel) and total OH-PAH emissions in logarithmic scale from different wood types and burning conditions (n = 3).

In accordance with OH-PAHs, also the PAH emissions (OH-PAHs not included) varied between different burning conditions and fuel types, as summarized in Table S5 in ESM and Fig. 1 and Fig. 3. In general, the PAH emissions were clearly higher in HB samples compared to NB samples for all the fuels. The average emissions of all 45 PAHs analyzed were 79.4 and 966 μg/MJ_fuel for NB and HB samples, respectively, and the emissions from NB samples corresponded on average to 18% of the emissions from HB samples. This clear general influence of burning conditions seen in this study, are in line with previous studies (Pettersson et al., 2011 and Eriksson et al., 2014) using the same stove model. During incomplete combustion conditions, there is an oxygen-deficiency resulting in elevated levels of PAHs and particles. For NB conditions, the batch average O₂ ranged from 10.3 to 13.2% (for all wood types) and for HB conditions from 6.1 to 8.1% (Table S2 in ESM), indicating that there was a higher overall oxygen deficiency during the HB conditions. In like manner, the NO levels were lower in HB (46–70 ppm) compared to NB (69–85 ppm), except for birch burning having similar NO levels, another indication of overall oxygen deficiency in HB. The burning (=sampling) times for NB were also much longer compared to HB. Consequently, a fast and air starved combustion in HB resulted in higher emissions of carbonaceous particulate matter and PAHs. The OH-PAH emissions (hydroxyl biphenyls not included) corresponded on average to 35 and 21% of PAH emissions from NB and HB samples, respectively. The emissions of 3-HBaP were even higher, corresponding on average to 70 and 50% of BaP emissions in NB and HB, respectively. It is worth highlighting that approximately 50% of BaP emissions in urban air in Stockholm have been estimated to be from wood combustion (Westerholm et al., 2012).

Fig. 3. PAH emissions (μg/MJ_fuel) and total PAH emissions in logarithmic scale from different wood types and burning conditions (n = 3).

Furthermore, it was observed that OH-PAH emissions from birch burning corresponded to 27% of PAH emissions for NB and 26% for HB, respectively. Similarly, OH-PAH emissions from aspen burning in NB corresponded to 20% and in HB to 26%, respectively of PAH emissions. For spruce burning in NB the OH-PAH emissions corresponded to 60% of PAH emissions, while it decreased to 16% in HB. The opposite trend was seen for pine burning, the OH-PAH emissions in NB corresponded to 14% and in HB to 21% of PAH emissions. Those relatively high emissions of OH-PAHs, particularly in spruce NB particles, can be explained by the high fraction of 2-ring aromatics, e.g. 1-N and 2-N present in the samples. The emissions of those compounds were then decreased when the burning condition was changed to HB. With some exceptions, the emissions of smaller OH-PAHs were higher for NB combustion compared to HB combustion. An overall increase of high molecular weight PAHs and OH-PAHs was observed when the combustion conditions were changed from NB to HB, e.g. the total emissions of dibenzopyrene isomers increased from 1.87 to 56.1 μg/MJ_fuel and emissions from HB pine accounted for the largest increase. That low molecular PAHs are predominant at NB conditions, while more high weight PAHs are more abundant during HB conditions have been reported in other studies as well (Pettersson et al., 2011 and Ramdahl et al., 1982). Ramdahl and co-workers observed that the speciation was shifted toward PAHs with molecular weights in the range 192–228 when the conditions were changed to HB (Ramdahl et al., 1982). This indicates that the formation of OH-PAHs and PAHs depends on the amount of oxygen and the combustion temperatures; the higher temperatures in absence of oxygen, the larger PAHs are formed. Contrariwise, smaller PAHs and phenols are formed at lower temperatures (Orasche et al., 2013).

For comparison with OH-PAH the corresponding PAHs, i.e. anthracene (Ant), phenanthrene (Phe), fluoranthene (Flu), pyrene (Pyr), chrysene (Chr) and benzo[a]pyrene (BaP) were used. In NB samples, Pyr was the most abundant compound (11.3 μg/MJ_fuel), followed by Flu (10.7 μg/MJ_fuel), Chr (8.25 μg/MJ_fuel), BaP (4.49 μg/MJ_fuel), Phe (3.04 μg/MJ_fuel) and Ant (1.18 μg/MJ_fuel). While in HB samples, Chr was the most abundant compound (145 μg/MJ_fuel), followed by BaP (58.8 μg/MJ_fuel), Flu (44.2 μg/MJ_fuel), Pyr (34.7 μg/MJ_fuel), Phe (22.1 μg/MJ_fuel) and Ant (5.31 μg/MJ_fuel). The relative content of Pyr, Phe and BaP were similar to those of corresponding OH-PAHs, except from Chr, which was one of the most abundant compounds, while 6-HC was one of the least abundant.

It has been shown that different aromatic compounds released from wood combustion are thermal decomposition products of lignin (Steiber, 1993). Lignin is an aromatic macromolecule present in wood and constitutes 25–30% of the wood weight and the amount of lignin in the wood increases the formation rate of PAHs during combustion. Lignin is also a thermally stable molecule and high temperatures are needed for the pyrolysis of this compound (Hueglin et al., 1997). The formation of PAHs starts by a reaction between small radicals and monocyclic and dicyclic aromatic hydrocarbons (e.g. naphthalene, cyclopentadiene, indene) generated during the early stages of decomposition and thermal transformations, to yield larger and larger PAHs as the temperatures increase (Orasche et al., 2013). The PAH formation processes from lignin structures were demonstrated in several studies where catechol (i.e. a representative of lignin) was pyrolyzed at temperatures between 500 and 1000 °C and the presence of varying amounts of oxygen (Wornat et al., 2001, Thomas et al., 2007 and Thomas and Wornat, 2008). Below 800 °C, PAH products obtained during both pyrolysis and oxygen-rich conditions were two-ring compounds, especially indene and naphthalene, while above 800 °C and in the absence of oxygen, much larger PAHs were generated. Furthermore, OH-PAHs are oxidation products of PAHs, a reaction that occurs through rapid radical reaction between hydroxyl radicals and PAHs. Hence, the emissions of OH-PAHs and PAHs are higher in HB compared to NB combustion.

3.3 OH-PAH and PAH emissions in particles from different wood types

The OH-PAH emissions from different wood types are shown as box plots in Fig. 4. Generally, the emissions from birch and aspen in NB were slightly lower compared to HB, emissions from spruce having same levels in both NB and HB, while emissions from pine were significantly higher in HB, whereby a clear increase of OH-PAH emissions could be observed for spruce in NB and pine in HB compared to the other fuels. For NB samples, birch had the lowest emissions of OH-PAHs with an average emission of 1.35 μg/MJ_fuel, followed by pine (1.42 μg/MJ_fuel), aspen (1.53 μg/MJ_fuel) and spruce with significantly higher emissions of OH-PAHs (7.53 μg/MJ_fuel) than the other wood types. For HB, another order was seen with birch having the lowest emissions of OH-PAHs (4.35 μg/MJ_fuel) followed by spruce (8.25 μg/MJ_fuel), aspen (9.94 μg/MJ_fuel) and finally pine (56.8 μg/MJfuel) with significantly higher OH-PAH emissions. The formation mechanisms of OH-PAHs in combustion and gasification are not well elucidated and have only been studied in coal combustion (Simoneit et al., 2007), thus, the considerably elevated levels of OH-PAHs for spruce in NB and pine in HB are difficult to further discuss based on the present study. The increased emissions of OH-PAHs for those fuel types might depend on the amount of lignin, combustion temperatures and the absence of oxygen during the combustion. However, for pine in HB, the low O₂ and significantly higher CO levels compared to NB indicate that oxygen-starvation have occurred during the combustion. Periods of O₂ below 2% were also observed (not observed for the other wood types), which is well known to cause higher PAH emissions (Eriksson et al., 2014). It can therefore be assumed that the high emission levels of OH-PAHs might have been caused by a higher degree of oxygen-deficiency compared to the other wood species.

Fig. 4. Box and whisker plots showing the total OH-PAH emissions (μg/MJ_fuel) from NB and HB samples for different wood types (the bottom and the top of the box plot are the first and the third quartiles, the band is the second quartile (the median), the ends of the whisker are the minimum and maximum of all the data and the black square represent the average).

1-HP was the OH-PAH with the highest measured emissions (maximum value in the box plot) in all wood types for HB samples, while for NB samples it had the highest measured emissions only from birch burning. Instead, 2-N had the highest emissions from aspen and spruce burning and 4-HBP from pine burning in NB. 6-HC had the lowest measured emissions (minimum value in the box plot) in all wood types and burning conditions, except from HB pine burning, where 1-8-DHAQ had the lowest measured emissions.

Generally the PAH emissions in NB were rather similar for the different fuel types, while a clear increasing trend was seen from birch to aspen to spruce and finally pine, during HB conditions. The substantially elevated emissions of both PAHs and OH-PAHs for pine during HB is most probably due to the even higher burn rates observed during initial combustion phase compared to the other wood fuels. All tests were performed using the same firing procedures, thus the explanation in this specific set-up is fuel related. As stated above, even if the manufacturer recommendations are followed, different wood types burn differently because of their chemical and physical properties, e.g. lignin content, density, etc. and those parameters might result in local oxygen-deficiency and lower burning temperatures and consequently elevated levels of PAHs/OH-PAHs. Some similarities when comparing to OH-PAH emissions for different wood types were however observed (Fig. 5). Similar to OH-PAHs, HB birch burning had the lowest emissions of PAHs (3.87 μg/MJ_fuel), while pine burning (65.8 μg/MJ_fuel) had the highest, followed by spruce (11.3 μg/MJ_fuel) and aspen (8.86 μg/MJ_fuel). The PAH emissions in NB samples were also similar to OH-PAH emissions, birch having the lowest emissions (0.50 μg/MJ_fuel) and spruce the highest (2.94 μg/MJ_fuel), followed by pine (1.79 μg/MJ_fuel) and aspen (1.63 μg/MJ_fuel). The most abundant PAH in NB birch burning samples (maximum value in the box plot) was Chr, while Flu was the most abundant in aspen and spruce burning samples and Pyr in pine samples. Chr had the highest measured emissions in HB birch and aspen burning samples, while Flu had the highest measured emission in spruce burning samples and benzo(ghi)fluoranthene in pine burning samples. PAHs with lowest emissions (minimum value in the box plot) in all types of NB and HB samples were dibenzopyrene isomers.

Fig. 5. Box and whisker plots showing the total PAH emissions (μg/MJ_fuel) from NB and HB samples for different wood types (same type of box and whisker plots as for Fig. 4).

3.4 Trends and correlations

Statistically significant differences (p ≤ 0.05) in the average emissions of OH-PAHs were found using t-test for unequal variances for birch and pine burning when the same wood type was compared for different burning conditions, while no statistically significant differences (p ≤ 0.05) were observed for aspen and spruce. Statistically significant differences (p ≤ 0.05) in the average PAH emissions were found in all cases for the same wood type but different burning conditions. Two factor ANOVA and Bonferroni correction were also performed to determine the differences within the two burning conditions (Table S6 in ESM). In NB samples, the average OH-PAH emissions from spruce were statistically different (significance level of a = 0.05) from the three other wood types and emissions from pine were statistically different (a = 0.05) from the others in HB samples. In HB samples, the average PAH emissions from pine burning were statistically different (a = 0.05) from the other wood types. In NB samples, the average PAH emissions from birch were statistically different (a = 0.05) from spruce, while no statistically significant difference (a = 0.05) was observed for the other wood types. Two factor ANOVA was also performed within the same batch of replicates (n = 3) for each combustion experiment and no statistically significant differences were observed for any of the wood types (Table S6 in ESM). There is always a possibility that the variance between the replicates for the same wood type and firing procedure is bigger than between different combustion experiments. The variance in this study (spread of the whiskers in Fig. 4 and Fig. 5) depends however on the emission differences of different analytes rather than the variance of the same analyte emission in the replicates of the same combustion experiment.

Correlation calculations were made for Phe/HPh, Pyr/1-HP, Chr/6-HC, BaP/3-HBaP, and positive correlations were observed in all cases; emissions of OH-PAH increased with increased PAH emissions. However, R² were between 0.10 and 0.45 and were not further discussed in the manuscript. The poor correlation depends probably on that only 9 OH-PAHs (one isomer for each parent PAH except for Phe) were determined in this study, while there might be many possible oxidation product isomers for each parent PAH.

Ratios between the parent PAH and OH-PAH were also calculated for Phe/2-HPh, Pyr/1-HP, Chr/6-HC, BaP/3-HBaP (Table 1). To the best of our knowledge there are no studies in the literature that have determined PAH to OH-PAH ratios in air particulate in regions dominated by wood combustion, e.g. in the Nordic countries. There are some studies that have determined those compounds in air particulate in Spain, China and Japan (Barrado et al., 2012, Barrado et al., 2013, Wang et al., 2007 and Kishikawa et al., 2004). OH-PAHs were shown to be more abundant during winter and it was partially explained by the increased coal burning for house heating in the cold season. However, the environmental concentrations of OH-PAHs can also be affected by meteorological parameters, e.g. temperature, solar radiation, UV light intensity and ozone concentration. The ratios for Phe/2-HPh and Pyr/1-HP were compared with ratios found in air particulate collected in Spain (Barrado et al., 2012). The ratio for Phe/2-HPh (average ratio 1.56) was in good agreement with the literature values (average for winter 1.86 and for summer 2.97), while the ratio for Pyr/1-HP in this study was much lower (average ratio 1.55) than the literature values (average ratio for winter 13.8 and for summer 19.7).

4 Conclusions

This study shows that not only combustion conditions, but also wood type influences the emissions of OH-PAHs and PAHs. In general, higher emissions of both OH-PAHs and PAHs are observed during HB conditions compared to NB conditions, with some exceptions. The highest emissions of OH-PAHs and PAHs for NB samples are from spruce and for HB samples from pine burning. Similarly, both 3-HBaP and BaP emissions in HB samples are highest from pine burning, while for NB samples the highest emissions are from spruce burning. Pine burning in HB has been shown to have largest impact on OH-PAH and PAH emissions. Emissions of OH-PAHs correspond on average to 28% of PAH emissions, while emissions of 3-HBaP, a compound shown to be both toxic and have estrogenic activity, correspond to 60% of BaP emissions. This study shows that wood burning is a large emission source of OH-PAHs and it is therefore necessary to further investigate the formation, occurrence and distribution of these compounds as they are present in significant amounts in wood smoke particles. Furthermore, other influences e.g. moisture content of the fuel should be examined in more detail. It is also important to emphasize that only 9 OH-PAHs have been quantified in this present study, (compared to 45 PAHs), and it can therefore be assumed that several other OH-PAHs and also other isomers of the OH-PAHs quantified have been omitted. Consequently, further investigations on occurrence of other OH-PAHs are needed both in wood smoke and air particles and also toxicological investigations on health effects of exposure to emissions containing OH-PAHs.

Acknowledgements

Tomas Alsberg at Department of Environmental Science and Analytical Chemistry(Stockholm University) is acknowledged for his assistance with the instrumentation. This study has been funded by Stockholm University and the Swedish Research Council, contract no. 621-2012-3802. The anonymous reviewers are acknowledged for helping us to improve this article.

Appendix A Supplementary data

The following is the supplementary data related to this article:

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