The impact of domestic wood burning on personal, indoor and outdoor levels of 1,3-butadiene, benzene, formaldehyde and acetaldehyde

Pernilla Gustafson *, Lars Barregard , Bo Strandberg and Gerd Sällsten
Department of Occupational and Environmental Medicine, Sahlgrenska University Hospital and Academy at Göteborg University, P.O. Box 414, S-405 30, Goteborg, Sweden. E-mail:; Fax: +46 31 40 97 28; Tel: +46 31 786 62 82

Received 28th September 2006 , Accepted 17th November 2006

First published on 7th December 2006


The aim of this study was to quantify personal exposure and indoor levels of the suspected or known carcinogenic compounds 1,3-butadiene, benzene, formaldehyde and acetaldehyde in a small Swedish town where wood burning for space heating is common. Subjects (wood burners, n = 14), living in homes with daily use of wood-burning appliances were compared with referents (n = 10) living in the same residential area. Personal exposure and stationary measurements indoors and at an ambient site were performed with diffusive samplers for 24 h. In addition, 7 day measurements of 1,3-butadiene and benzene were performed inside and outside the homes. Wood burners had significantly higher median personal exposure to 1,3-butadiene (0.18 µg m–3) compared with referents (0.12 µg m–3), which was also reflected in the indoor levels. Significantly higher indoor levels of benzene were found in the wood-burning homes (3.0 µg m–3) compared with the reference homes (1.5 µg m–3). With regard to aldehydes, median levels obtained from personal and indoor measurements were similar although the four most extreme acetaldehyde levels were all found in wood burners. High correlations were found between personal and indoor levels for all substances (rs > 0.8). In a linear regression model, type of wood-burning appliance, burning time and number of wood replenishments were significant factors for indoor levels of 1,3-butadiene. Domestic wood burning seems to increase personal exposure to 1,3-butadiene as well as indoor levels of 1,3-butadiene and benzene and possibly also acetaldehyde. The cancer risk from these compounds at exposure to wood smoke is, however, estimated to be low in developed countries.


Wood is burned for heating and cooking or just for pleasure in many homes worldwide. Almost three billion people rely on solid fuel (biomass and coal) as their primary source of domestic energy. Most of them live in developing countries but there are also many living in countries with a cold climate, such as Sweden. In Sweden the number of people using wood boilers, wood stoves and fireplaces is increasing because of higher costs of both fossil fuels and electricity as well as government policy of promoting a shift to renewable fuel consumption. Moreover, domestic wood burning is facilitated by the easy availability of wood in Sweden.

Burning wood releases a wide range of potentially harmful air pollutants , including particulate matter (PM ), carbon monoxide, benzene, 1,3-butadiene, aldehydes, polycyclic aromatic hydrocarbons (PAHs) and many other toxic organic compounds.1,2 These compounds are emitted to the outdoor air and, to a varying extent, also to the indoor air, depending on the type of burning appliance. In Sweden, different kinds of wood-burning appliances are in use, including boilers for water-based heating and hot water production. Stoves and fireplaces are often used as secondary heating sources. By contrast, the burning appliances in developing countries are often simple, poorly designed and poorly maintained stoves with low combustion efficiency.3 These factors, in combination with extensive use of solid fuels, may lead to high indoor levels of incomplete combustion products causing adverse health effects including acute respiratory infections, chronic obstructive pulmonary disease, and lung cancer (from coal), and probably also cataracts, tuberculosis, asthma attacks and adverse pregnancy outcomes.3 Indoor smoke from solid fuels was estimated to be one of the top ten global risk factors for mortality and burden of diseases in 2001, mainly in developing countries.4 Cooking with solid fuels has been estimated to cause a significant proportion, about 3%, of the global burden of disease, which corresponds to over 1.6 million premature deaths annually.4 In modern society, domestic wood combustion contributing to ambient levels of PM has been associated with adverse health effects such as asthma.5 In a recent experimental study acute effects on inflammation and coagulation were shown after human exposure to wood smoke.6

Exposure to benzene, 1,3-butadiene, formaldehyde and acetaldehyde has the potential for adverse health effects, as these substances are suspected or known carcinogens.7–9 In addition, formaldehyde and acetaldehyde are irritants to the eyes and airways. The knowledge of personal exposure to these pollutants among people using domestic wood burning is very limited, which makes risk assessment problematic. In a recent study, domestic wood burning was shown to influence some trace elements in both personal PM exposure and indoor PM levels in Sweden.10

This study aims to quantify personal exposure and indoor levels of 1,3-butadiene, benzene, formaldehyde and acetaldehyde in a residential area where wood burning for space heating is common. The hypothesis was that people (wood burners) living in homes with daily use of wood-burning appliances have increased personal exposure and indoor levels. We compared the personal exposures and indoor levels between wood burners and referents who lived in the same residential area in Sweden.

Materials and methods

The study was carried out during winter, from 10 February to 12 March 2003, among subjects living in a residential area in the small Swedish town of Hagfors (5600 population). The weather was typical for the season, with full snow cover and mean daily temperatures of between –2 °C and –10 °C. The wind speed was low, <3 m s–1 for all days. During this season heating of the houses is necessary. Domestic wood burning is common in this residential area and no major industries are located within a distance of 6 km. Houses in the area are equally affected by the closest major road (3000 vehicles per day) passing the edge of the area. Local traffic within the area is limited. The area (400 × 1500 m) consists of 225 single-family homes, with about one-third using wood-burning appliances (boilers or fireplaces) continuously (11%), daily or weekly (10%) or less frequently (16%). The study was approved by the Ethics Committee at Göteborg University.

Subjects, study design and background information

Subjects were selected based on information obtained from the local chimney sweeping register as to the type of heating systems present in the inhabitants’ homes. To avoid contamination from other possible sources besides burning wood logs, houses heated by combustion of other fuels such as oil were not included. Households using wood pellets for burning were also excluded since their emissions are low compared with emissions from burning wood logs.2 The participants were adults who, during the measurement period, were unlikely to be exposed to the pollutants of interest at their work. People working at the local steel mill in Hagfors were therefore only investigated on non-working days. Smoking was not allowed by the subjects carrying the personal samplers or by anyone in their homes during the sampling period.

Eighteen households in the area, with daily use of wood-burning appliances, were approached and in 14 of them (78%) one householder each (ten men and four women) agreed to participate in the study. Of these subjects eleven had boilers and three had fireplaces. One of the boilers was located in a shelter outside the house. As referents, individuals living in homes with electrical heating or a heat pump were selected and the participation rate was 83% (five men and five women from ten out of twelve households).

The features of the study design together with details of the sampling techniques are summarized in Table 1. Personal exposure and indoor measurements of 1,3-butadiene, benzene, formaldehyde and acetaldehyde were carried out for 24 h together with simultaneous measurements in ambient air taken on the roof of a single car garage attached to the home of a referent with electrical heating. In addition, measurements of 1,3-butadiene and benzene were performed for 7 days inside and outside the participants’ homes. The 1 day samples were collected on the first day of the 7 day sampling periods. The sampling was performed for wood burners and the reference group in parallel except on one occasion.

Table 1 Features of the study design including type of sampler, sampling time, uptake rate used, limit of detection (LoD ), and number of samples of the pollutants (N), taken at the different locations
Pollutant Samplera Sampling time/days Uptake rate/ml min–1 LoD /µg m–3 Personal (N) Living room (N) Outside home (N) Ambient site (N)
a PE = Perkin Elmer (PE) sampler (Perkin Elmer, Wellesley, MA, USA). SKC = SKC Ultra Passive Sampler (SKC Inc., Eighty Four, PA, USA). Umex 100 Passive Sampler (SKC Inc., Eighty Four, PA, USA).b The results of the 1,3-butadiene measurements outside the homes were not reliable and have not been presented.
1,3-Butadiene SKC 1 14.9 0.015 24 24 9
1,3-Butadiene PE 7 0.56 0.03 24 24b
Benzene SKC 1 16 0.03 24 24 9
Benzene PE 7 0.59 0.04 24 24
Formaldehyde Umex 1 25 1.9 24 24 9
Acetaldehyde Umex 1 21 2.3 24 24 9

Personal samplers were attached within the breathing zone and placed close to the bed at night. Indoor measurements were performed in the living room, with the sampler hanging 1.5 m above the floor and at least 0.5 m away from lamps, walls and windows. At the outdoor sites the samplers were placed 1.5 m above the ground and on top of the garage (4 m above the ground). Temperature and relative humidity were registered with Tinytag data loggers (Gemini Data Loggers UK Ltd, Chichester, UK) indoors in the homes as well as outdoors on top of the garage.

During the 24 h sampling period each participant completed a diary, noting where they had spent the day and listing all activities at home and work. In this way we collected information about possible exposure to the pollutants of interest (e.g. filling the car with petrol, and environmental tobacco smoke (ETS), gas cooking and wood burning). The wood burners also noted when they or someone else in the household had made a fire (frequency of wood replenishment), and how much and what type of wood they had burned. For the 7 day sampling, participants completed a limited questionnaire regarding activities in the home, which could possibly have affected the concentrations. Information on the age and type of boiler, presence of an accumulator tank (i.e. a water tank used as a heat reservoir), firewood storage, and type of firewood was collected from the participants. The mean age of the eleven boilers was 18 years (range 3–51 years). Two had environmental certification from the Swedish National Testing and Research Institute, and another three were equipped with accumulator tanks. The wood had been stored for between 4 months and 3 years, and in all cases it had been stored outdoors under a roof or in a shed. Three of the homes (one wood-burning home and two reference homes) had an attached garage.

Sampling methods

1,3-Butadiene and benzene were sampled for 24 h using the SKC Ultra Passive Sampler (SKC Inc., Eighty Four, PA, USA), which has been validated for measurements in ambient air.11,12 It is a plastic, badge-type sampler (diameter: 30 mm; thickness: 15 mm) containing about 600 mg adsorbent , in this case Carbopack X 60–80 mesh (Supelco, Bellefonte, PA, USA). For 1 week measurements, the Perkin Elmer (PE) sampler (Perkin Elmer, Wellesley, MA, USA) was used. This sampler consists of a steel tube (90 mm × 6.3 mm outside diameter (o.d.) × 5.0 mm inside diameter (i.d.)) filled with about 300 mg Carbopack X 60–80 mesh. This sampler has been validated for 1 week ambient air monitoring.13 Before and after sampling, the samplers were stored at room temperature and wrapped in aluminium foil. The uptake rates used are presented in Table 1. Field duplicate measurements were performed indoors and outdoors. The coefficient of variation (CV) of the duplicate pairs of 1,3-butadiene and benzene levels was 4% and 9% measured with the SKC samplers (n = 8), and 12% (only indoors, n = 5) and 14% (n = 11) with the PE sampler, respectively.

The Umex 100 Passive Sampler (SKC Inc., Eighty Four, PA, USA), used to measure formaldehyde and acetaldehyde, is a modification of the GMD sampler.14 It consists of polypropylene housing (20 × 30 × 5 mm) containing two reagent (2,4-dinitrophenylhydrazine (DNPH))-coated filters, one working as a blank filter. Before and after sampling the sampler was placed in a sealed pouch and stored in a deep freeze. The GMD and the Umex sampler have been validated for formaldehyde for 8 h and 1 week sampling.14,15 Low concentrations of formaldehyde in air have been measured for 24 h with the GMD sampler16,17 and with the Umex sampler.18 The uptake rates used are presented in Table 1. The CVs of field duplicate measurements performed both indoors and outdoors (n = 5) were 11% for formaldehyde and 31% for acetaldehyde. The lack of validation and the high CV for the acetaldehyde measurements with the Umex 100 sampler must be considered when interpreting the results.

Analytical methods

1,3-Butadiene and benzene were analysed with automatic thermic desorption using an ATD 400 (Perkin Elmer) coupled to a gas chromatograph (Autosystem GLX; Perkin Elmer) equipped with a flame ionization detector (FID ).11 The adsorbent tube (the adsorbent in the SKC Ultra sampler was transferred to Perkin Elmer tubes) was desorbed by heating and concentrated on a cold trap. The sample was injected into the separation column (PLOT fused-silica) by rapid heating of the cold trap to 250 °C.11 The detection limits are presented in Table 1.

Formaldehyde and acetaldehyde were analysed using a high-performance liquid chromatography (HPLC) system (consisting of a Varian 2510 pump and an HP autosampler 1100 injector) with ultraviolet (UV) detection.14 The aldehyde–dinitrophenylhydrazone was eluted from the filter with acetonitrile and separated on a Nucleosil C-18 column using a mobile phase of 64% methanol in water. The hydrazone was detected with an adsorbance detector (Varian 2550) at 365 nm. All samples were corrected with the individual blanks. The detection limits are presented in Table 1.

Statistical methods

The differences in air pollution levels between wood burners and the reference group were assessed using the Wilcoxon rank-sum test for both personal exposure and indoor levels, since the levels were, in general, not normally distributed. The tests we performed were one-sided, since the hypothesis was that the wood-burning group would display higher concentrations. Wilcoxon’s signed-rank tests (two-tailed) were performed on the differences in concentrations within the pairs of personal and indoor samples for the entire dataset, as well as for wood burners and the reference group separately. Spearman’s rank correlation coefficient (rs) was used to express correlations between the different sampling locations (personal, indoors and outdoors) and between levels of pollutants . Associations between the levels of the pollutants (log-transformed data) and the characteristics of wood-burning appliances (presence of water storage tank and type of appliance: boiler or fireplace) and wood-burning behaviour (frequency of wood replenishments, amount and type of wood: deciduous trees or a mixture of deciduous and coniferous trees, as well as wood-burning time) were investigated with stepwise backward multiple regression analysis . P-values < 0.05 were considered statistically significant. Statistical calculations were made using SAS for Windows, version 9.1.19 For values below the limit of detection (LoD ), the calculated value of the LoD divided by the square root of 2 was used in the statistical calculations.20


Background data on wood-burning behaviour and activities

Most of the wood-burning households used a mix of deciduous and coniferous trees. Four used only deciduous trees. The subjects spent 91% (mean, range 71–100%) of the 24 h sampling time indoors (77% at home, 9.6% at work and 4.7% in other indoor places), about 6% of the sampling time outdoors, and 3% in cars or buses. No differences in activities were found between the wood-burning group and the reference group. There were no significant differences in temperature or relative humidity between the two groups.

Personal exposure, and indoor and outdoor levels

Tables 2 and 3 summarize the results of personal, indoor and outdoor measurements of 1,3-butadiene, benzene, formaldehyde and acetaldehyde among the wood-burning group and the reference group. Note that the indoor result from one wood burner was omitted because the boiler involved was located in a separate shed and not in the main building. In addition, the 7 day indoor results for two wood burners were omitted because of some smoking indoors during the sampling period.
Table 2 Results of measurements of personal exposure, indoor levels and levels outside the homes, expressed in µg m–3. Means, medians and 95% confidence intervals (CIs) of medians are presented for the wood-burning group and the reference group for 1 day and 7 day sampling. Significant differences between wood burners and the reference group are indicated in bold, and the p-value given
Pollutant /location Sampling time/days Wood-burning group (n = 14) Reference group (n = 10)
Mean Median 95% CI Mean Median 95% CI
a Wood burners, n = 13; b Wood burners, n = 11. c p = 0.015. d p = 0.03.e p = 0.03.f p = 0.05.
Personal 1 0.33 0.18 c 0.12–0.54 0.14 0.12 0.06–0.23
Living room 1a 0.38 0.20 d 0.08–0.48 0.11 0.10 0.05–0.14
Living room 7b 0.31 0.23 e 0.10–0.72 0.11 0.11 0.06–0.17
Personal 1 4.3 2.2 1.4–7.9 2.4 1.7 1.0–4.0
Living room 1a 3.9 2.6 1.3–7.3 2.0 1.4 0.95–2.9
Living room 7b 5.7 3.0 f 1.4–14 2.5 1.5 1.0–2.6
Outside home 7a 1.3 1.2 1.0–1.7 1.0 1.1 0.80–1.2
Personal 1 24 19 11–36 26 28 20–33
Living room 1a 28 26 11–36 27 28 14–42
Personal 1 17 14 11–23 13 12 9.4–16
Living room 1a 17 13 9.6–25 12 11 8.1–16

Table 3 Ambient air concentrations (µg m–3) measured with the samplers placed in the middle of the study area. Means, medians and 95% confidence intervals (CIs) are presented for 1 day sampling
Pollutant n Mean Median 95% CI
1,3-Butadiene 9 0.12 0.11 0.04–0.16
Benzene 9 1.2 1.2 1.1–1.4
Formaldehyde 9 3.7 3.1 2.5–5.8
Acetaldehyde 9 3.2 3.1 1.6–4.0

Wood burners had significantly higher personal exposure to 1,3-butadiene (median 0.18 µg m–3) than did the reference group (median 0.12 µg m–3), as shown in Table 2 and Fig. 1a. Likewise, significantly higher indoor levels were found for wood burners, both for 1 day and for 7 day sampling, than for the reference group. The personal exposure and indoor levels showed a large variation among wood burners, the personal exposure of, and indoor levels for, some subjects being considerably higher than the medians (Fig. 1a and 2). The 1 day ambient levels measured on the roof of the garage in the middle of the study area were similar to the personal exposure and indoor levels for the reference group (Table 3).

Personal exposure and indoor levels (µg m–3) of (a) 1,3-butadiene, and (b) benzene for the wood-burning group (grey boxes) and the reference group (unfilled boxes). Indoor levels are shown for both 1 day and 7 day sampling times. The box plots show the 10th, 25th, 50th (i.e. the median), 75th and 90th percentiles.
Fig. 1 Personal exposure and indoor levels (µg m–3) of (a) 1,3-butadiene, and (b) benzene for the wood-burning group (grey boxes) and the reference group (unfilled boxes). Indoor levels are shown for both 1 day and 7 day sampling times. The box plots show the 10th, 25th, 50th (i.e. the median), 75th and 90th percentiles.

The association between (1 day) personal exposure and indoor levels for 1,3-butadiene, benzene, formaldehyde and acetaldehyde. Spearman’s rank correlation coefficient is presented. One high value of 1,3-butadiene is not shown (personal 1.3 µg m–3; indoors 2.1 µg m–3).
Fig. 2 The association between (1 day) personal exposure and indoor levels for 1,3-butadiene, benzene, formaldehyde and acetaldehyde. Spearman’s rank correlation coefficient is presented. One high value of 1,3-butadiene is not shown (personal 1.3 µg m–3; indoors 2.1 µg m–3).

The 1,3-butadiene levels measured personally and indoors showed very high correlations both for the combined group (rs = 0.94; p < 0.0001) and for the two groups separately, as illustrated in Fig. 2. Personal exposure was significantly higher than indoor levels for the combined group (p = 0.02) and for the reference group (p = 0.04), but not for wood burners (p = 0.27).

For benzene, we found significantly higher indoor levels for the 7 day sampling for wood burners than for referents, but there were no significant differences in personal exposure or 1 day indoor levels (Table 2). However, the 1 day mean levels of benzene measured personally and indoors in the wood-burning group were about twice as high as those in the reference group (Table 2). The personal exposure was in the same range as the indoor levels (Table 2 and Fig. 1b), while the outdoor levels in the middle of the residential area and outside the homes were clearly lower in both groups (Tables 2 and 3). One sample collected outside a home with a boiler was contaminated and has been excluded. The personal exposures and indoor levels of some of the wood burners were considerable higher than the medians, as shown in Fig. 1b and 2.

The correlations between 1 day personal exposure and indoor levels were high for benzene in the combined group (rs = 0.94; p < 0.0001) as well as for the two groups separately (Fig. 2), but no correlation was observed between 7 day levels measured indoors and outside the homes. Statistically significant differences between personal exposure and indoor levels, and between levels indoors and outside the homes, were observed for the two groups combined and within the groups (personal exposure > indoor levels: p = 0.022 for wood burners, p = 0.004 for referents and p = 0.0003 for the combined group; indoor levels > outdoor levels: p = 0.006 for the wood-burning group, p = 0.004 for the reference group and p < 0.0001 for the combined group).

No significant differences between wood burners and the reference group were observed for formaldehyde and acetaldehyde (Table 2). The median personal exposure to formaldehyde was 23 µg m–3 (combined group) and the median indoor level was 28 µg m–3. For acetaldehyde, the median levels obtained from personal and indoor measurements were similar (13 µg m–3 and 12 µg m–3, respectively), although the four extreme values were all measured in wood burners. Personal exposure and indoor levels of formaldehyde and acetaldehyde were clearly higher than ambient levels (Tables 2 and 3).

High correlation coefficients between personal exposure and indoor levels were found for formaldehyde and acetaldehyde, rs = 0.87 and rs = 0.86 (p < 0.0001), respectively, in the combined group and in the two groups separately (Fig. 2). There was no significant difference between indoor levels and personal exposure levels of formaldehyde and acetaldehyde, either in the combined group or in the separate groups.

Associations between exposure and wood-burning behaviour, and between the pollutants

In the multiple regression analysis , a model with log-transformed 1,3-butadiene indoor levels (24 h samples) showed a significant impact of number of wood-burning hours (p = 0.012), type of appliance (higher for wood boilers, p = 0.012) and number of wood replenishments (higher with few replenishments, p = 0.004) among the wood burners. This model yielded an adjusted R2 of 0.64. Similar associations were found for acetaldehyde and to some extent for formaldehyde, but not for benzene.

In Table 4 the correlation coefficients for the associations between the different pollutants measured personally and indoors are presented for the combined group and the wood-burning group. Personal exposure and indoor levels of acetaldehyde were significantly correlated with 1,3-butadiene and formaldehyde in both these groups. Correlations between 1,3-butadiene and benzene measured personally or indoors were statistically significant in the combined group but not for the wood-burning group alone.

Table 4 Correlation coefficients (rs) for 1 day personal exposure and 1 day indoor concentrations of the pollutants of interest (Bd = 1,3-butadiene; Bz = benzene; Fo = formaldehyde; Ac = acetaldehyde) in the combined group and in the wood-burning group. Statistically significant values are indicated in bold
  Personal   Indoors
  Combined group (n = 24) Wood-burning group (n = 14)   Combined group (n = 23) Wood-burning group (n = 13)
  Bd Bz Fo Ac Bd Bz Fo Ac   Bd Bz Fo Ac Bd Bz Fo Ac
Bd 1 0.49 0.21 0.51 1 0.38 0.25 0.56 Bd 1 0.55 0.42 0.58 1 0.41 0.45 0.63
Bz   1 0.25 0.41   1 0.33 0.45 Bz   1 0.31 0.38   1 0.39 0.38
Fo     1 0.47     1 0.72 Fo     1 0.52     1 0.64
Ac       1       1 Ac       1       1


In this study, significantly higher indoor levels of 1,3-butadiene and benzene were found in homes with daily use of wood-burning appliances compared with homes with other types of heating though located in the same residential area. For 1,3-butadiene, the difference was also reflected in personal exposure. For formaldehyde and acetaldehyde, no impact of domestic wood burning was observed, although a few wood burners had high levels of acetaldehyde. To our knowledge, this is the first study investigating the influence of wood burning on personal exposure to these compounds.

Concentrations and comparisons with other studies

For 1,3-butadiene, there was no statistically significant difference between personal exposures and indoor levels for wood burners, while the referents had higher personal exposure compared with indoor levels. These facts suggest that the personal exposure for the wood-burning group had a dominant indoor source. Exposure to 1,3-butadiene has only been measured in a few recently published studies, due to lack of validated measurement methods. Although wood burners in the town of Hagfors had higher levels than their referents, their median personal exposure was half the median found in a non-smoking general population from Umeå, a middle sized Swedish city.21 Zhu et al.22 reported indoor and outdoor levels in Canada comparable to those in our study. Kim et al.23 reported a median personal exposure of 0.4 µg m–3 among twelve non-smoking persons in the UK and a similar median indoor level in non-smoking homes. The level was, however, almost twice as high in homes with a smoker.24 Kinney et al.25 reported mean personal exposure and indoor levels among students in the USA that were three times higher than for the wood burners in our study. The outdoor levels in the US study were similar to those in our study, indicating the presence of stronger indoor sources in the US homes. Moreover, ten to twenty times higher median personal exposure and indoor levels, and also higher outdoor levels, were found in Mexico City.26 In a study of different environments in Australia, the highest levels of 1,3-butadiene and benzene were recorded in locations affected by traffic pollution and cigarette smoke (e.g. inside cars, basement car parks and night clubs) and in two homes heated with wood.27 The higher traffic intensity in more densely populated cities and the impact of tobacco smoke are factors which may have contributed to the generally higher levels found in other studies compared with ours.

For benzene, the personal exposure and indoor levels were higher in the wood-burning group but the differences were not statistically significant, except for the indoor 7 day measurements. The median personal exposure in the combined group (1.9 µg m–3) was somewhat higher than the exposure (1.4 µg m–3) in non-smokers in another Swedish study.21 Because of the high personal exposure observed for some of the wood burners, the mean personal exposure of this group was twice that of the referents. Benzene in petrol has been reduced in many countries during the 1990’s, and therefore a comparison with published results should focus on the last decade (1996 to 2006). In a European study (EXPOLIS) performed in 1996–1997, the personal exposure levels ranged from 2 to 3 µg m–3 in Helsinki, Basel and Oxford, while higher levels were observed in Prague and Athens (8 and 12 µg m–3, respectively).28 Relatively high personal exposure and indoor levels have also been found in the UK, with levels being twice as high in homes with smokers compared with non-smoking homes.23,24 The levels in our study were in agreement with some recently published studies from the USA25,29–31 and other countries.22,32–34 Higher personal exposure and indoor and outdoor levels were found in Mexico City.26 Son et al.35 reported very high personal exposure, as well as indoor and outdoor levels, among a population in Korea (about 25–40 µg m–3), similar to results reported by Sinha et al.36 for people in India who cook indoors with wood fuels (23–50 µg m–3). Our outdoor levels were somewhat lower than our personal and indoor levels, which is consistent with results found in many other studies.21,22,25,29 The benzene content in petrol, as well as traffic intensity, presence of wood burning, and smoking habits are important factors determining benzene levels.

For formaldehyde and acetaldehyde, no significant difference could be seen between the two groups regarding either personal exposure or indoor concentrations. Formaldehyde is a widely measured indoor air pollutant , while few studies have reported acetaldehyde levels. The median personal exposure to formaldehyde (23 µg m–3) and acetaldehyde (13 µg m–3), as well as the indoor levels, in our study are in agreement with previous studies from Sweden, Mexico, Great Britain, France, and Finland.17,26,34,37,38 Kinney et al.25 reported lower personal exposure and indoor levels of formaldehyde (about 12 µg m–3) in the USA compared with our results, while their acetaldehyde levels were similar to ours. Outdoor levels of formaldehyde and acetaldehyde were considerably lower than personal exposure and indoor levels in this study, which is consistent with the previously mentioned studies. The only three studies we could find on indoor aldehyde levels and domestic wood burning in the developed world are from Canada, where Lévesque et al.39 found no difference between homes with and homes without wood-burning appliances regarding indoor formaldehyde levels. The same result was reported by Gilbert et al.18,40 for formaldehyde and for acetaldehyde,40 consistent with findings in our study. In developing countries, however, where simple and often poorly ventilated cook stoves are used, levels are much higher. A median level of formaldehyde as high as 652 µg m–3 was found inside 20 Indian homes using wood stoves for cooking.41 Increased levels of formaldehyde and acetaldehyde were also found in an experimental wood smoke study.42 This shows that wood burning can indeed increase levels of aldehydes if the level of wood smoke is high.

Our findings of a few wood burners having high acetaldehyde levels, together with the positive association between 1,3-butadiene and acetaldehyde in wood burners (Table 4), could possibly have been caused by acetaldehyde in wood smoke.

Personal exposure versus indoor concentrations

Personal exposure reflects the time spent in different microenvironments and the concentration within the microenvironments. In this study a relatively large proportion of time was spent indoors at home (77%) compared with other Swedish studies (65–70%).17,21 This explains the high correlations between personal exposure and indoor levels (rs > 0.8 and p < 0.0001 for all compounds). Similar results have been shown in other studies for formaldehyde and acetaldehyde,17,43 and also to some extent for benzene.44 Indoor residential levels of benzene, 1,3-butadiene, formaldehyde and acetaldehyde compounds are generally dependent on both outdoor sources (such as engine exhaust and emissions from industries) and indoor sources (such as tobacco smoking, cooking, and emissions from consumer products, construction materials, and furnishings). In the residential area studied, wood burning constitutes an additional source for these compounds. In addition, indoor concentrations are also influenced by home characteristics such as ventilation rates and emission rates.45 Most Swedish people rarely open their windows and doors in wintertime (in our study, windows and doors were only open for 0.1 h (mean) during the 24 h sampling period), resulting in reduced air flow in the homes during this time of the year. Air pollution levels emitted by indoor sources therefore build up inside the home.

Wood burning

Exposure to wood smoke can occur both outdoors from ambient air and indoors as a result of direct release from cooking and heating devices and fireplaces, leakage from boilers and stoves, and the infiltration of outdoor air pollution . For wood burners, lighting the fire may have contributed to the wide range of personal exposure levels. Wood combustion emissions have been shown to be highly dependent on many factors related to the type of wood, appliance and burning behaviour.1,2 In this study, information on these factors was provided by the wood burners, and was included in a multiple regression model investigating the impact on 24 h indoor levels, for which we consider the information to be more precise compared with the 1 week period. We found that a model including type of wood-burning appliance, burning time and number of wood replenishments explained 64% of the variation in 24 h indoor levels of 1,3-butadiene. McDonald et al.1 quantified emission rates of 350 elements, inorganic compounds, and organic compounds, among them 1,3-butadiene, benzene, formaldehyde and acetaldehyde, from residential wood combustion. They found higher emission rates for wood stoves than for fireplaces and higher emissions among fireplaces burned with hardwood (deciduous trees) compared with softwood (coniferous trees).1 In our study, boilers were associated with higher indoor levels of 1,3-butadiene. In Sweden, birch and spruce are the main types of wood burnt, and most of the subjects in our study used a mixture of these. Moreover, a long burning time was associated with higher indoor levels of 1,3-butadiene, while a large number of wood replenishments was associated with lower levels. Lower emissions can be achieved with frequent replenishments because of more complete combustion obtained with a smaller amount of wood.2 Also, connecting the wood boiler to a heat storage tank can reduce the emissions.2 This factor was not significant in our model for 1,3-butadiene, but it was significantly associated with lower indoor levels of acetaldehyde. This is a small study including a limited number of persons compared with the number of variables entered into the model, and further investigations are needed.

Other sources

1,3-Butadiene, benzene, formaldehyde and acetaldehyde are emitted from all types of incomplete combustion of organic material, not only wood burning, and are ubiquitous in traffic environments. Elevated personal benzene exposure has been associated with time spent commuting in a car, and also with refuelling owing to the presence of benzene in petrol.33,46 The variation in time spent in cars and buses or outdoors may have contributed to the variation in personal exposure observed in both of our groups. The impact of traffic on the indoor concentration in our study is considered to have been low and evenly distributed for all the homes since they are located in a small town and in the same residential area.

Another source of 1,3-butadiene, benzene, formaldehyde and acetaldehyde is cigarette smoking,38,47 and high levels of benzene and 1,3-butadiene have been reported in homes of smokers and in pubs and restaurants.24,46 The interference of smoking was avoided in this study by selecting non-smokers and not allowing smoking inside the homes during the sampling period, although we did not verify non-exposure status with any biological or indoor measurements of cotinine or nicotine. Seven of the households included a smoker, but they generally smoked outdoors, and if smoking was reported during the sampling period the measurement was excluded. Environmental tobacco smoke is not always avoidable and was noted in the diaries of two wood burners (20 min and 1 h, respectively). Burning incense has been shown to produce high levels of benzene, formaldehyde and acetaldehyde indoors.48,49 However, in Sweden burning of incense is very uncommon and it constituted no indoor source in our study.

Both formaldehyde and acetaldehyde are dominated by indoor sources.45 For formaldehyde, indoor sources such as furnishings, construction materials (e.g. particle board and insulation) and consumer products may overshadow the impact of wood burning on the indoor levels. Elevated indoor formaldehyde levels have been reported to be dependent on many factors, such as increased temperature and relative humidity , type of home (single-family house versus apartment; newly built or refurbished), presence of particle board flooring, and, for both formaldehyde and acetaldehyde, smoking and low air-exchange rate.17,34,40,45,50 The influence of house type on indoor levels of formaldehyde was not an issue in our study since all subjects lived in single-family homes.

Correlations between the four compounds

We found an association between benzene and 1,3-butadiene both for personal exposure and for indoor levels (Table 4). These results are in agreement with two other studies.21,24 A positive correlation is to be expected, since the two compounds are both constituents of wood smoke as well as of motor vehicle exhaust and tobacco smoke, though the source strength may be different. High correlations between formaldehyde and acetaldehyde were found, both for personal and for indoor measurements, which is in agreement with a high correlation found inside Canadian homes.40 By contrast, Jurvelin et al.43 did not find any correlation between these two aldehydes, either for personal exposure or for indoor levels. A few wood burners had fairly high acetaldehyde levels in their homes, and the subject with the highest level also had the highest level of formaldehyde in addition to relatively high 1,3-butadiene and benzene levels.

Risk assessment

The personal exposure to 1,3-butadiene for wood burners and referents in this study is well below the low-risk value of 2.5 µg m–3, which represents a lifetime cancer risk of 1 × 10–5 in the Swedish population.51 The unit risk estimate calculated by the US Environmental Protection Agency (EPA)52 results in a much lower low-risk value, of 0.3 µg m–3, and is within the range of health-based guideline values (0.2–1.0 µg m–3) discussed in Sweden when applying an uncertainty factor to the low-risk value.51 Only one wood burner exceeded this range. For benzene, both the wood burners and the referents had a median personal exposure higher than the low-risk guideline of 1.3 µg m–3 used in Sweden.53 The US EPA52 and the WHO47 have similar risk estimates. Considering the risk estimates, and also the practical possibility of limiting the benzene exposure, the European Commission has adopted an ambient air quality limit value for benzene of 5 µg m–3 to be met by the member countries by 1 January 2010 (Directive 2000/69/EC). The median formaldehyde exposure was within the guideline value range of 12–60 µg m–3 recommended in Sweden, which is based on the irritative effects and not on the carcinogenic potency.53 The WHO has likewise based their air quality guideline value of 100 µg m–3 (30 min average) on irritation, and this is the lowest concentration that has been associated with nose and throat irritation in humans after short-term exposure.47 By contrast, the US EPA52 recommends a low-risk value of 0.8 µg m–3. In Sweden, a guideline value for acetaldehyde has not been recommended, but the personal exposure levels in our study are much lower than the levels known to cause irritation.54 The US EPA52 has estimated a cancer risk of 1 × 10–5 at a lifetime exposure of 5.0 µg m–3.

The risk assessment indicates that for the average person living in a house heated with wood, the risk of developing cancer is low as a consequence of the additional contribution to the personal exposure from 1,3-butadiene, benzene and aldehydes. Using the point estimates for the additional contribution in the present study, it would mean less than one case per year per one million people using wood-burning appliances in Sweden. A wood burner exposed to benzene in wood smoke over a lifetime has an estimated cancer risk comparable with an individual living with a smoker exposed to benzene in ETS for a lifetime, while for 1,3-butadiene, the associated cancer risk is more than a factor of 10 lower for wood burners.55 The cancer risk associated with domestic wood burning may be significant in developing countries where a larger number of people are exposed and often at much higher levels. Wood smoke also contains a significant amount of PAHs, many of which are carcinogenic, and this must be taken into consideration when estimating the total cancer risk.


Domestic wood burning seems to increase indoor levels of 1,3-butadiene and benzene, which in the present study was reflected in personal exposure levels of 1,3-butadiene. However, the 1,3-butadiene levels measured personally, indoors and outdoors were low compared with those reported in other studies. By contrast, benzene levels were in the same range as reported for other European countries and the USA, although much lower than in Asia, Mexico and some developing countries. For formaldehyde and acetaldehyde, the generally reported pattern of personal and indoor levels being several times higher than outdoor levels was also seen in this study and the levels were in the same range as in previous studies. No clear effect of wood burning on the aldehyde levels was found, although in some wood burners there were also increased acetaldehyde levels. The cancer risk from these compounds at exposure to wood smoke is probably low in Sweden, but may be substantial in developing countries where frequent use of unvented stoves causes higher indoor levels. For risk management, the replacement of old wood-burning appliances with new, environmentally approved ones would dramatically reduce emissions of air pollutants from domestic wood burning.


This project was funded by the Swedish Energy Agency and is a part of the Swedish National Air Pollution and Health Effects Program (SNAP). The authors also acknowledge the subjects in the study for participating.


  1. J. D. McDonald, B. Zielinska, E. M. Fujita, J. C. Sagebiel, J. C. Chow and J. G. Watson, Environ. Sci. Technol., 2000, 34, 2080 CrossRef.
  2. L. Johansson, B. Leckner, L. Gustavsson, D. Cooper, C. Tullin and A. Potter, Atmos. Environ., 2004, 38, 4183 CrossRef CAS.
  3. J. J. Zhang and K. R. Smith, Br. Med. Bull., 2003, 68, 209 Search PubMed.
  4. A. D. Lopez, C. D. Mathers, M. Ezzati, D. T. Jamison and C. J. L. Murray, Lancet, 2006, 367, 1747 CrossRef.
  5. B. C. Boman, B. A. Forsberg and B. G. Järvholm, Scand. J. Work, Environ. Health, 2003, 29, 251 Search PubMed.
  6. L. Barregard, G. Sällsten, P. Gustafson, L. Andersson, L. Johansson, S. Basu and L. Stigendal, Inhalation Toxicol., 2006, 18, 845 Search PubMed.
  7. IARC Monographs on the Evaluation of Carcinogenic Risks to Humans, 2004, vol. 88, accessed at Search PubMed.
  8. IARC Monographs on the Evaluation of Carcinogenic Risks to Humans, 1999, vol. 71, accessed at Search PubMed.
  9. IARC Monographs on the Evaluation of Carcinogenic Risks to Humans, 1982, vol. 29, accessed at Search PubMed.
  10. P. Molnár, P. Gustafson, S. Johannesson, J. Boman, L. Barregård and G. Sällsten, Atmos. Environ., 2005, 39, 2643 CrossRef CAS.
  11. B. Strandberg, A.-L. Sunesson, K. Olsson, J.-O. Levin, G. Ljungqvist, M. Sundgren, G. Sällsten and L. Barregard, Atmos. Environ., 2005, 39, 4101 CrossRef CAS.
  12. B. Strandberg, A.-L. Sunesson, M. Sundgren, J.-O. Levin, G. Sällsten and L. Barregard, Atmos. Environ., 2006, 40, 7686–7695 CrossRef CAS.
  13. J.-O. Levin, A.-L. Sunesson, M. Sundgren and B. Strandberg, Proceedings of the Fifth International Symposium on Modern Principles of Air Monitoring, Loen, Norway, 2005 Search PubMed.
  14. J.-O. Levin, R. Lindahl and K. Andersson, Environ. Technol. Lett., 1988, 9, 1423 CAS.
  15. R. Lindahl and M. Rhen, Proceedings of the Conference Indoor Air, Beijing, China, 2005 Search PubMed.
  16. R. Lindahl, J.-O. Levin and B. Järvholm, Proceedings of the Conference Indoor Air, Edinburgh, Scotland, 1999 Search PubMed.
  17. P. Gustafson, L. Barregård, R. Lindahl and G. Sällsten, J. Exposure Anal. Environ. Epidemiol., 2005, 15, 252 Search PubMed.
  18. N. L. Gilbert, D. Gauvin, M. Guay, M.-E. Héroux, G. Dupuis, M. Legris, C. C. Chan, R. N. Dietz and B. Lévesque, Environ. Res., 2006, 102, 1 CrossRef CAS.
  19. SAS Statistical software version 9.1, SAS Institute Inc., Cary, NC, 2003 Search PubMed.
  20. R. W. Hornung and L. D. Reed, Appl. Occup. Environ. Hyg., 1990, 5, 46 CAS.
  21. L. Modig, A.-L. Sunesson, J.-O. Levin, M. Sundgren, A. Hagenbjörk-Gustafsson and B. Forsberg, J. Environ. Monit., 2004, 6, 1 RSC.
  22. J. Zhu, R. Newhook, L. Marro and C. Chan, Environ. Sci. Technol., 2005, 39, 3964 CrossRef CAS.
  23. Y. M. Kim, S. Harrad and R. M. Harrison, Environ. Sci. Technol., 2002, 36, 5405 CrossRef CAS.
  24. Y. M. Kim, S. Harrad and R. M. Harrison, Environ. Sci. Technol., 2001, 35, 997 CrossRef CAS.
  25. P. L. Kinney, S. N. Chillrud, S. Ramstrom, J. Ross and J. D. Spengler, Environ. Health Perspect., 2002, 110, 539 CAS.
  26. P. I. Serrano-Trespalacios, L. Ryan and J. D. Spengler, J. Exposure Anal. Environ. Epidemiol., 2004, 14, 118 Search PubMed.
  27. A. L. Hinwood, H. N. Berko, D. Farrar, I. E. Galballt and I. A. Weeks, Chemosphere, 2006, 63, 421 CrossRef CAS.
  28. K. Saarela, T. Tirkkonen, J. Laine-Ylijoki, J. Jurvelin, M. J. Nieuwenhuijsen and M. Jantunen, Atmos. Environ., 2003, 37, 5563 CrossRef CAS.
  29. K. Sexton, J. L. Adgate, G. Ramachandran, G. C. Pratt, S. J. Mongin, T. H. Stock and M. T. Morandi, Environ. Sci. Technol., 2004, 38, 423 CrossRef CAS.
  30. J. L. Adgate, L. E. Eberly, C. Stroebel, E. D. Pellizzari and K. Sexton, J. Exposure Anal. Environ. Epidemiol., 2004, 14, 4 Search PubMed.
  31. J. L. Adgate, T. R. Church, D. R Ryan, G. Ramachandran, A. L. Fredrickson, T. H. Stock, M. T. Morandi and K. Sexton, Environ. Health Perspect., 2004, 112, 1386 CAS.
  32. P. Schneider, I. Gebefügi, K. Richter, G. Wölke, J. Schnelle, H. E. Wichmann, J. Heinrich and INGA Study Group, Sci. Total Environ., 2001, 267, 41 CrossRef CAS.
  33. A. Horton, F. Murray, M. Bulsara, A. Hinwood and D. Farrar, Atmos. Environ., 2006, 40, 2596 CrossRef CAS.
  34. G. J. Raw, S. K. D. Coward, V. M. Brown and D. R. Grump, J. Exposure Anal. Environ. Epidemiol., 2004, 14, 85 Search PubMed.
  35. B. Son, P. Breysse and W. Yang, Environ. Int., 2003, 29, 79 CrossRef CAS.
  36. S. N. Sinha, P. K. Kulkarni, S. H. Shah, N. M. Desai, G. M. Patel, H. N. Mansuri and H. N. Saiyed, Sci. Total Environ., 2006, 357, 280 CrossRef CAS.
  37. N. Gonzales-Flesca, A. Cicolella, M. Bates and E. Bastin, Environ. Sci. Pollut. Res., 1999, 6, 95 Search PubMed.
  38. J. A. Jurvelin, M. Vartiainen, M. J. Jantunen and P. Pasanen, J. Air Waste Manage. Assoc., 2001, 51, 17 CAS.
  39. B. Lévesque, S. Allaire, D. Gauvin, P. Koutrakis, S. Gingras, M. Rhainds, H. Prud’Homme and J.-F. Duchesne, Sci. Total Environ., 2001, 281, 47 CrossRef CAS.
  40. N. L. Gilbert, M. Guay, J. D. Miller, S. Judek, C. C. Chan and R. E. Dales, Environ. Res., 2005, 99, 11 CrossRef CAS.
  41. C. V. Raiyani, S. H. Shah, N. M. Desai, K. Venkahia, J. S. Patel, D. J. Parikh and S. K. Kashyap, Atmos. Environ., 1993, 27A, 1643 CrossRef CAS.
  42. G. Sällsten, P. Gustafson, L. Johansson, P. Molnar, B. Stranberg, C. Tullin and L. Barregard, Inhalation Toxicol., 2006, 18, 855 Search PubMed.
  43. J. A. Jurvelin, R. D. Edwards, M. Vartiainen, P. Pasanen and M. J. Jantunen, J. Air Waste Manage. Assoc., 2003, 53, 560 CAS.
  44. M. L. Phillips, N. A. Esmen, T. A. Hall and R. Lynch, J. Exposure Anal. Environ. Epidemiol., 2005, 15, 35 Search PubMed.
  45. S. S. Sax, D. H. Bennett, S. N. Chillrud, P. L. Kinney and J. D. Spengler, J. Exposure Anal. Environ. Epidemiol., 2004, 14, 95 Search PubMed.
  46. R. D. Edwards and M. J. Jantunen, Atmos. Environ., 2001, 35, 1411 CrossRef CAS.
  47. WHO, Air Quality Guidelines for Europe, WHO, Copenhagen, 2nd edn, 2000, accessed at Search PubMed.
  48. S. S. H. HO and J. Z. Yu, J. Environ. Monit., 2002, 4, 728 RSC.
  49. S.-C. Lee and B. Wang, Atmos. Environ., 2004, 38, 941 CrossRef CAS.
  50. B. Clarisse, A. M. Laurent, N. Seta, Y. Le Moullec, A. El Hasnaoui and I. Momas, Environ. Res., 92, 245 Search PubMed.
  51. N. Finnberg, P. Gustavsson, J. Högberg, G. Johansson, G. Sällsten, M. Warholm and K. Victorin, Summary Risk Assessment of 1,3-butadiene, IMM-report 1/2004 (in Swedish), Institute of Environmental Medicine, Karolinska Institute, Stockholm, 2004, Search PubMed.
  52. US Environmental Protection Agency. Integrated Risk Information System, 2006, accessed at Search PubMed.
  53. K. Victorin, Risk assessment of carcinogenic air pollutants IMM Report 1/98 (in Swedish), Institute of Environmental Medicine, Karolinska Institute, Stockholm, 1998, accessed at Search PubMed.
  54. Environment Canada and Health Canada, Priority substances list assessment report: acetaldehyde, Canadian Environmental Protection Act, 1999, 2000 Search PubMed.
  55. W. W. Nazaroff and B. C. Singer, J. Exposure Anal. Environ. Epidemiol., 2004, 14, 7 Search PubMed.

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