How Harmful Is Particulate Matter Emitted from Biomass Burning? A Thailand Perspective

  • Helinor J. JohnstonEmail author
  • William Mueller
  • Susanne Steinle
  • Sotiris Vardoulakis
  • Kraichat Tantrakarnapa
  • Miranda Loh
  • John W. Cherrie
Open Access
Human Health Effects of Environmental Pollution (KC Makris, Section Editor)
Part of the following topical collections:
  1. Topical Collection on Human Health Effects of Environmental Pollution


Purpose of Review

A large body of epidemiological evidence demonstrates that exposure to particulate matter (PM) is associated with increased morbidity and mortality. Many epidemiology studies have investigated the health effects of PM in Europe and North America and focussed on traffic derived PM. However, elevated levels of PM are a global problem and the impacts of other sources of PM on health should be assessed. Biomass burning can increase PM levels in urban and rural indoor and outdoor environments in developed and developing countries. We aim to identify whether the health effects of traffic and biomass burning derived PM are similar by performing a narrative literature review. We focus on Thailand as haze episodes from agricultural biomass burning can substantially increase PM levels.

Recent Findings

Existing epidemiology, in vitro and in vivo studies suggest that biomass burning derived PM elicits toxicity via stimulation of oxidative stress, inflammation and genotoxicity. Thus, it is likely to cause similar adverse health outcomes to traffic PM, which causes toxicity via similar mechanisms. However, there is conflicting evidence regarding whether traffic or biomass burning derived PM is most hazardous. Also, there is evidence that PM released from different biomass sources varies in its toxic potency.


We recommend that epidemiology studies are performed in Thailand to better understand the impacts of PM emitted from specific biomass sources (e.g. agricultural burning). Further, experimental studies should assess the toxicity of PM emitted from more diverse biomass sources. This will fill knowledge gaps and inform evidence-based interventions that protect human health.


Particulate Matter PM10 PM2.5 Biomass burning Toxicity Air pollution Hazard In vitro In vivo Epidemiology Thailand  


A large body of epidemiological evidence has clearly demonstrated that short- and long-term exposure to particulate matter (PM) is associated with increased morbidity and mortality (e.g. [1, 2, 3, 4, 5, 6, 7], reviewed in [8, 9, 10, 11, 12, 13, 14, 15]). A broad range of health effects are associated with PM exposure, which are summarised in Fig. 1. There is compelling evidence that PM has detrimental impacts on the respiratory and cardiovascular systems. For example, poor air quality is associated with an increased incidence of stroke and myocardial infarctions, an increase in emergency hospital admissions related to asthma, chronic obstructive pulmonary disease (COPD), respiratory infections and an increased incidence of lung cancer (Fig. 1). There is also emerging evidence that exposure to PM increases risk for neurological (e.g. dementia, cognitive impairment) and metabolic (e.g. diabetes) diseases and that PM can cause reproductive and developmental toxicity [12, 16, 17, 18, 19, 20, 21, 22] (Fig. 1). Susceptibility to the health impacts of PM can vary with respect to health, age and socioeconomic status [10]. For example, it is established that individuals with pre-existing disease (e.g. asthma, COPD, cardiovascular disease), young children and the elderly are more vulnerable to the toxicity of PM [10]. Socioeconomic status is also important, with adverse health impacts occurring more frequently in economically disadvantaged groups [10]. Due to concerns regarding PM toxicity, air quality standards have been adopted globally to protect the general public [10].
Fig. 1

Examples of the adverse health impacts associated with exposure to particulate air pollution. Short- and long-term exposure to PM can cause a spectrum of adverse health outcomes, which vary in their severity. Adapted from [15]. COPD = chronic obstructive pulmonary disease, CVD = cardiovascular disease, MI = myocardial infarction

The bulk of evidence available on the health impacts of PM is mainly associated with exposure to urban (e.g. traffic derived) PM, with relatively fewer studies addressing the toxicity of PM specifically emitted from other emission sources, such as biomass burning. In Southeast (SE) Asia, haze episodes from agricultural biomass burning commonly occur each year [23]. More specifically, fire is used to clear agricultural land in SE Asia, and the resulting smoke (termed haze) can spread to other regions and countries causing transboundary pollution. During haze events, pollutant levels can remain elevated for long periods, thereby having a detrimental impact on air quality [24]. Biomass burning therefore has the potential to substantially contribute to and elevate ambient PM levels, so there is a need to understand whether the implications for health following exposure to biomass-derived PM are the same as those observed for PM emitted from other sources (e.g. traffic). In this review, as part of the Thailand Air Pollution and Health Impact Assessment (TAPHIA) study, we have performed a narrative review of the available literature to identify whether PM emitted from biomass burning is toxic in order to (1) identify whether the health effects of urban (e.g. traffic) and biomass burning derived PM are similar by examining the findings from epidemiology, in vivo and in vitro studies, and (2) recommend future research priorities that will address current knowledge gaps relating to the toxicity of biomass burning–derived PM in order to strengthen the evidence base, and inform interventions to protect human health. We will focus our recommendations on Southeast Asia, and in particular on Thailand where biomass burning is an important source of PM emissions and overall exposures, and where currently the impacts of PM emitted from burning biomass on human health are uncertain.

Contribution of Emission Source to PM Toxicity: a Global Perspective

To demonstrate adherence to regulatory limits and to protect human health, ambient air pollutant levels are monitored in many large urban and rural locations across the world, with World Health Organisation (WHO) guideline levels set at a 24 h mean of 25 μg/m3/50 μg/m3 PM2.5/PM10 (i.e. PM with an aerodynamic diameter smaller than 2.5 and 10 μm, respectively) and an annual mean of 10 μg/m3/20 μg/m3 for PM2.5/PM10 [10]. These guidelines assume that all PM is equally toxic. However, PM emitted from different sources vary with respect to their physico-chemical properties (e.g. composition, shape, size and surface area), which potentially influences PM toxicity.

The impacts of PM on human health are relevant globally, in both urban and rural settings. The majority of epidemiology studies conducted to date have been performed in Europe and North America, but poor air quality associated with elevated levels of PM is a global problem. Importantly, the sources of pollution varies in different countries, influencing the level of PM exposure, and the physico-chemical properties of the emitted PM [25]. Regardless of source, PM is likely to elicit detrimental health effects; however, the specific role of PM sources in inducing adverse health impacts is still very much under investigation [26]. Untangling whether the source of PM influences its toxicity is essential as, whilst there is agreement that elevated levels of PM are associated with adverse health effects, harmful effects can be observed when PM levels are within permitted levels (e.g. [27, 28]). Though it is not possible to have an absence of PM (as PM also comes from natural, as well as anthropogenic sources), it is important to understand whether PM from specific anthropogenic sources varies in its toxicity, since this can help prioritise source control measures and ultimately protect public health [29, 30, 31, 32].

It has been hypothesised that the ultrafine particle component of PM (i.e. PM with an aerodynamic diameter smaller than 100 nm) is primarily responsible for its toxicity [33]. Experimental (laboratory) studies extensively investigated the cell and molecular mechanism of toxicity of air pollution particles to understand why adverse health effects were observed in epidemiology studies. The studies found that the ultrafine particles in PM stimulated oxidative stress and inflammation in the lung following inhalation, which led to detrimental health outcomes (for reviews, see [34, 35, 36, 37]). Existing laboratory studies that investigated the toxicity of PM primarily assessed the toxicity of PM10 samples obtained from traffic sources, or used ultrafine carbon black (ufCB; i.e. carbon particle with a diameter of 14 nm) as a surrogate for traffic-derived PM10. However, PM is heterogeneous, and variations in the characteristics of particles (e.g. particle size and surface area, and composition [e.g. polyaromatic hydrocarbon (PAH), metal, and endotoxin content]) released from different emission sources can influence the biological response (reviewed by [10, 12, 38]). Variance in the hazard of PM emitted from different sources (e.g. biomass burning) is therefore expected as they are likely to have different physico-chemical properties to that emitted from traffic.

Contribution of Biomass Burning to Ambient PM Levels

Burning of fossil fuels is the main source of ambient PM in the developing world (reviewed in [11, 13]). In urban areas, road traffic is typically the primary source of PM [39]. However, biomass burning can also increase the burden of PM in environments that span urban and rural locations, indoor and outdoor settings, as well as developed and developing countries [40, 41]. Activities that generate PM emissions from biomass combustion include electricity generation, commercial and domestic emissions (e.g. cooking, heating), agriculture, waste and natural emissions [40, 42]. Examples of open biomass burning include burning of agricultural waste, forest fires, wild fires, bushfires and land clearing, whereas examples of enclosed biomass burning includes domestic biofuel burning (e.g. wood, animal dung, crop residues) for cooking and heating and domestic/industrial burning for electricity [43]. As the use of renewable energy is appealing for many reasons (e.g. rising cost of using fossil fuels) in developed and developing countries, this has led to an increase in burning of biomass fuels globally [11].

To date, biomass burning–derived PM has primarily been a health concern for developing countries. However, exposure to biomass PM is increasing in developed countries mainly when used for domestic heating purposes, but also increasingly from wild fires, and can substantially contribute to ambient PM concentrations, particularly in winter months (reviewed in [41, 44]). For example, Glasius et al. assessed the influence of residential wood burning on local air quality in a rural town in Denmark (with limited traffic) and found that the levels of PM were similar to those recorded in urban areas (a busy street in Copenhagen) [45]. Furthermore, Fuller et al. demonstrated that wood burning (for domestic heating purposes) contributed to PM10 levels in London [46]. In addition, Snider et al. measured indoor and outdoor levels of PM in rural China to identify the level of PM associated with wood burning or gas/electric stoves and found that the use of wood burning stoves was associated with up to 55% higher PM2.5 levels indoors than the use of gas/electric stoves [47]. For some European countries (e.g. Sweden, Finland, Germany and Austria), it has been estimated that 15–25% of PM2.5 particles originate from residential biomass combustion, and it is expected that by 2020, biomass burning will contribute to almost 40% of fine PM emissions in Europe (reviewed by [41]).

Over recent years, there has also been a greater frequency, size and intensity of wildfires worldwide (e.g. bush, forest, vegetation, peat and grass fires). [48, 49]. Wildfires can be a significant source of air pollution, with emissions including PM, carbon monoxide (CO), PAHs and volatile organic compounds (VOCs) [50]. Air quality is severely affected during wildfire episodes, with the levels of PM emitted resulting in concentrations greatly exceeding target (regulated) levels. For example, He et al. investigated the impact of open biomass burning (forest and farmland burning) on air quality in Brisbane, Australia, and observed that PM levels were 10 times higher than the annual average, and exceeded WHO 24 h guidelines during the wildfire events [43]. Artaxo and Yamasoe quantified the levels of PM10 that were emitted during biomass burning season in the Amazon Basin and found that concentrations reached 150–700 μg/m3 (depending on the location), which is well over background PM10 concentrations of 10 μg/m3 [51]. In 2008, PM values in Sacramento River Valley (California, USA) during wildfires reached 262 μg/m3 (reviewed by [52]). Furthermore, wildfires in Southern California in 2003 saw PM10 levels increase by three- to fourfold, with one region reaching 215 μg/m3 [53]. Of additional concern is that, like other sources, PM emitted from wildfires may travel over long distances to affect regions far from the original source. Indeed, Sapkota et al. investigated the long range transport of pollutant emissions from forest fires in Quebec (Canada), and in particular, PM, and demonstrated that air quality was affected in Baltimore (USA), over a thousand kilometres from the source [50].

Sources of PM in Thailand

The population of Thailand is exposed to PM that is emitted from a range of different sources, with different sources dominating PM emissions in different regions of the country. For example, in the Bangkok Metropolitan Area, with a population of 12.6 million (in 2015), the number of cars on the road was 10 million and vehicle emissions were the main source of PM [54]. Emissions from power plants and industry are also large contributors of PM emissions in Thailand, particularly in areas surrounding Bangkok, including the central and eastern regions [55, 56]. As discussed previously, there is a wealth of information available on the health impacts of urban (e.g. traffic and industry) derived PM, and thus, it is possible to infer what the potential risks posed by these sources of PM to the Thailand population are from existing epidemiology and toxicology studies.

PM emissions from biomass burning can also influence ambient PM concentrations in Thailand. In particular, the burning of forests and crop residues for agricultural purposes can impact on air quality. More specifically, anthropogenic forest fires are used to clear the land so that it can be used for agricultural purposes, as the burnt land is often considered to be nutrient rich. Agricultural land is also intentionally burnt before and after harvesting (e.g. to clear crop residues and weeds) and to prepare the land for the next crop [57]. The frequency of agriculture-related biomass burning in Thailand (as well as other SE Asian countries) varies depending on the region and crop, but the greatest activity is typically observed during the dry season (January to April) and in northern regions of the country. Agricultural burning is often performed for financial reasons due to the high cost of the machinery required for land/forest clearing, as it is often quicker and easier to perform burning than using other approaches to clear land/forests, and as in some areas, there is no easy access for the required machinery.

Phairuang et al. assessed the influence of agricultural burning (e.g. crop residues, forest fires) on air quality in the upper northern, lower northern and northeast regions in Thailand and found that these activities correlated with elevations in ambient PM concentrations [58]. Indeed, it is established that agricultural (crop) burning in Thailand leads to concentrations of PM that are well over WHO guideline levels, which is likely to increase the incidence of adverse health effects [23, 59]. Furthermore, agricultural burning increased PM levels above national guideline levels in northern parts of the country, with the greatest effect observed during the dry season (January–April), where a tripling of the daily PM concentration was recorded [60]. In addition, markers of biomass burning in ambient PM samples, such as levoglucosan—an anhydrohexose that is produced from the pyrolysis of carbohydrates—have been shown to be up to 20× higher during the burning season in Thailand [61, 62]. These studies therefore indicate that PM emission sources can vary by season.

It has also been demonstrated that air quality in Thailand can be affected both in locations where biomass burning is intensive, but also in more distant regions. For example, it has been observed that the clearing of tropical forest in SE Asia for commercial agriculture causes the release of significant quantities of gases and aerosols, which disrupts air quality and produces haze in neighbouring countries [63, 64]. These transboundary effects are heavily dependent on meteorological conditions, including prevailing wind and precipitation levels. For example, Kliengchuay et al. demonstrated that air quality in the Mae Hong Son Province (in the north of Thailand) was influenced by several sources depending on the season, including local biomass burning (e.g. natural or anthropogenic forest fires and agricultural burning) and transboundary pollution, and that PM levels were strongly influenced by meteorological factors (e.g. wind direction, wind speed, rainfall, temperature) [65].

Whilst the impacts of agricultural burning on air quality in Thailand have been evaluated in several studies, it is also essential to assess the potential risks posed by the emitted PM on human health.

Review of Epidemiology and Experimental Toxicology Studies that Have Assessed the Harmful Effects of Biomass-Derived PM

Adverse health effects from biomass-derived PM can be assessed using epidemiology and experimental (laboratory) studies. Epidemiological studies are essential in demonstrating a link between PM exposure and detrimental impacts on human health but can be costly and time consuming to perform. Furthermore, it is more difficult to assess the toxicological mechanism of action of PM in epidemiology studies, as these are generally not controlled exposures. Therefore, in vitro and in vivo studies are often commonly used to assess the toxicity of PM. In vivo studies allow the toxicity of PM to be assessed in a whole organism, but there is a drive to move away from animal studies for ethical, financial and time reasons and also due to the fact that there are issues with the reproducibility of in vivo studies and that they may not be able to predict human responses (reviewed in [66]). In vitro studies can be criticised for their lack of physiological relevance, but these tests allow the toxicity of PM to be screened relatively cheaply and quickly. Investigation of the mechanism of toxicity of PM can also be probed using in vitro studies, which is critical in informing potential impacts on health and identifying markers to screen PM toxicity in the future. Furthermore, the use of alternatives to animal testing also has ethical benefits, and using human cells can help overcome issues with species differences which are apparent when using rodents [66].

Epidemiology, in vivo and in vitro studies that have investigated the toxicity of biomass burning-derived PM have been identified and their results discussed in order to identify if the emission source influences the toxicity of PM. This information will be used to inform whether biomass burning derived PM is likely to elicit similar adverse health outcomes to that observed for urban (e.g. traffic-related) PM. The experimental design of existing studies will also be critically appraised and knowledge gaps identified in order to inform the direction of future research.


We performed a narrative review of the literature to synthesise current epidemiological and toxicological evidence relating to the health impacts of PM emitted from biomass burning. In order to identify relevant scientific papers, a search of PubMed was performed using keywords such as ‘Biomass burning’ AND particulate matter OR ‘PM10/2.5’ OR ultrafine particles OR air pollution AND ‘toxicity’ OR ‘hazard’ AND ‘in vitro’ OR ‘in vivo’ OR ‘epidemiology’ OR ‘human health’. The abstract of each paper that was identified from the search was read to identify the relevance of the publication to our review. In addition, research publications that were referenced in these articles or in relevant review articles but were not identified in the PubMed search were also considered if their content was deemed relevant to the review. The quality of each study was assessed by appraising the suitability of the experimental design, presentation of results and appropriateness of conclusions.

Epidemiology Studies

Large-scale epidemiological studies, beginning with the Harvard Six Cities study [1] and the seminal American Cancer Study [7], have consistently documented that elevated levels of PM are linked to detrimental health outcomes associated with short- and long-term exposure (reviewed in, e.g. [8, 9, 10, 11, 12, 13, 14]). There is also a growing body of evidence from epidemiology studies that suggests exposure to non-traffic–derived PM, including industrial emissions and natural dust and salts, can cause adverse health effects (e.g. [67, 68]).

Epidemiology studies to assess whether exposure to biomass burning–derived PM is linked to a similar set of health outcomes as traffic PM were identified and their results are summarised in Table 1. Available studies have primarily focussed on investigation of negative impacts of biomass burning–derived PM on the respiratory and cardiovascular system (e.g. increased exacerbations of asthma and COPD, hospital admissions, use of medication, mortality, respiratory infections, stroke and myocardial infarction incidence) (reviewed in [11, 13, 78]), with a more limited number of studies assessing impacts on other adverse health effects (e.g. adverse birth outcomes).
Table 1

A summary of the empirical epidemiological studies which have investigated health impacts from exposure to PM derived from outdoor biomass burning. Studies presented in order of publication date



Study type

Population/sample size

Time period

Exposure source

Concentration levels

Endpoints investigated

Summary of findings

Tan et al. [69]


Panel study

National service men (n = 30)

June–December 1997

Wildfire smoke

Mean PM10 during haze = 125.4 μg/m3

Spirometry, white blood cell counts

Elevated neutrophil counts during haze period, no change in lung function

Swiston et al. [70]

British Columbia, Canada

Panel study

Seasonal forest firefighters (n = 52)

May–August 2004–2005

Wildfire smoke

Estimated respirable peak PM > 2 mg/m3

Symptoms, spirometry and sputum induction

65% reported respiratory symptoms

Increases following fire-fighting shifts:

% sputum granulocytes 6.5 to 10.9%; circulating white blood cells 5.55 × 109 to 7.06 × 109 cells/L; and band cells 0.11 × 109 to 0.16 × 109 cells/L

Jayachandran [71]



All (census data)


Wildfire smoke

Mean daily PM10 > 150 μg/m3 for long periods

Infant mortality

Estimated ~ 15,600 child, infant and fetal deaths from exposure to wildfires

Morgan et al. [72]

Sydney, Australia


All (population-based registry)


Wildfire smoke

Median PM10 on bushfire days = 62 μg/m3

Daily mortality and hospital admissions

No bushfire association with cardiovascular admissions or mortality

A 10-μg/m3 increase in bushfire PM10 with the following hospital admissions:

Respiratory disease = 1.24% (95% CI 0.22% to 2.27%)

COPD = 3.80% (95% CI 1.40 to 6.26)

Adult asthma = 5.02% (95% CI 1.77 to 8.37)

Rappold et al. [73]

North Carolina, USA


All (population-based registry)

June 2008

Peat bog wildfire smoke

Max 1-h PM2.5 levels > 200 μg/m3

Cardiopulmonary emergency department visits

Increased RRs during wildfire period:

Asthma = 1.65 (95% CI 1.25 to 2.10)

COPD = 1.73 (95% CI)

Pneumonia and acute bronchitis = 1.59 (95% CI 1.07 to 2.34)

Cardiopulmonary symptoms = 1.23 (95% CI 1.06 to 1.43)

Heart failure = 1.37 (95% CI 1.01 to 1.85)

Holstius et al. [74]

California, USA


All (population-based registry)


Wildfire smoke

PM10 during fire period > 40 μg/m3

Birth weight

Compared to before and after, mean birth weight during fire period:

7.0 g lower [95% CI − 11.8, − 2.2] when fire occurred during third trimester

9.7 g lower [95% CI − 14.5, − 4.8] when fire occurred during second trimester

No association in first trimester

Cândido da Silva et al. [75]



All (population-based registry)

July 2004–December 2005

Biomass burning in forest and grassland

Peak seasonal mean PM2.5 (July–August 2005) = 60.5 μg/m3

Low birth weight (< 2500 g)

Q4 compared to Q1 PM2.5 exposure:

2nd trimester: 1.51 (95% CI 1.04 to 2.17)

3rd trimester: 1.50 (95% CI 1.06 to 2.15)

Dennekamp et al. [76]

Melbourne, Australia


All (population-based registry)

July 2006–June 2007

Wildfire smoke

Mean PM10 during risk period = 55.2 μg/m3

Out-of-hospital cardiac arrests (OHCAs)

Significant OHCA increased rates in men during fire period:

11.1% (95% CI 1.55, 21.48); IQR = 13.7 μg/m3

Haikerwal et al. [77]

Victoria, Australia


All (population-based registry)

December 2006–January 2007

Wildfire smoke

Mean PM2.5 = 15.4 μg/m3

Max PM2.5 = 163.4 μg/m3

OHCAs, IHD, acute myocardial infarction and angina (admissions/emergency department visits)

Increase in IQR (9.04 μg/m3) in PM2.5 associated with:

OHCAs 6.98% (95% CI 1.03 to 13.29); IHD (emergency department) 2.07% (95% CI 0.09 to 4.09)

IHD (hospital admissions) 1.86% (95% CI 0.35 to 3.4)

CI = confidence interval, IHD = ischaemic heart disease, IQR = interquartile range, OHCA = out-of-hospital cardiac arrests, RR = relative risk

The majority of epidemiology studies investigating the toxicity of biomass burning–derived PM have focussed on health impacts associated with exposure to PM during wildfire events. For example, Rappold et al. reported an increase in hospital admissions for cardiopulmonary disease (e.g. asthma, COPD, pneumonia, heart failure) during a peat wildfire event in North Carolina, USA [73]. Similarly, Morgan et al. identified an association between the levels of bushfire PM and respiratory hospital admissions in Sydney, with a 3.8% increase in admissions for asthma and 5% increase for COPD observed with a 10-μg/m3 increase in bushfire PM [72]. A systematic review of 61 epidemiology studies investigating non-occupational exposure to wildfire smoke demonstrated that > 90% of the studies found significant positive associations with respiratory morbidity [79]. In occupational studies, wildland firefighters have shown a consistent trend of declines in forced expiratory volume in 1 s (FEV1) following the forest fire season, with some indication of a return to baseline FEV1 levels [80]. Furthermore, Swiston et al. demonstrated that acute exposure to woodsmoke (from forest fires in Canada) was associated with pulmonary and systemic inflammation in firefighters [70]. Wildfire smoke-haze episodes were found to raise levels of circulating neutrophils in the blood of exposed national service men in Singapore [69]. The effect of smoke from wild and prescribed biomass burning on human health is a complex framework of interrelations and trade-offs, as described by Williamson et al. [81]. For example, exposure to smoke from biomass burning (e.g. wildfires) can occur at low frequency and can potentially affect large areas and populations, whereas prescribed agricultural burning affects smaller areas but occurs more often. Furthermore, prescribed fires often avoid hot and windy conditions, which can result in smoke getting trapped under atmospheric inversions or transported over urban areas and thereby increasing ambient concentrations and population exposure.

Cardiovascular morbidity in relation to wildfire smoke has been reviewed by Reid et al. [82], with the conclusion that results, at present, are inconsistent. Many epidemiology studies have not found any associations between wildfire smoke exposure and grouped cardiovascular disease (CVD) outcomes. Specific endpoints, such as out-of-hospital cardiac arrests and hospitalisations for acute myocardial infarction, were associated with biomass-related PM in selected studies [73, 76, 77]. Studies investigating ischemic heart disease (IHD) morbidity (e.g. hospitalisations/emergency department visits related to IHD) associated with wildfire smoke exposure also showed inconsistent results, and there were only a limited number of studies for other CVD outcomes [82].

Adverse birth outcomes due to wildfire smoke exposure during pregnancy are plausible [82], with studies finding associations between wildfire or agricultural burning smoke and early life death [71], lower birth weight [74, 75], and small for gestational age [83]. However, currently available evidence for causation from biomass burning emissions is limited.

Whilst much of the epidemiology on biomass burning has focussed on high exposure scenarios (e.g. associated with wildfires), another body of literature focusses on indoor biomass burning for cooking and heating. Evidence is accumulating linking household air pollution due to biomass cooking or heating fuels and adverse birth and pregnancy outcomes [84]. Smith et al. performed a critical review of the literature and observed that there was an increased incidence of acute respiratory infections in children in developing countries due to indoor biomass burning [85]. Indoor biomass burning for cooking or fuel can increase PM levels to 20–80 mg/m3, presenting significant risks for women at home for COPD and lung cancer [86].

Taken together, the findings from epidemiology studies suggest that exposure to biomass burning–derived PM can cause adverse health effects. Existing studies have primarily focussed on investigation of impacts on the pulmonary and cardiovascular systems, and the findings have revealed that biomass burning derived PM causes adverse health outcomes which are similar to those observed for traffic-derived PM (outlined in Fig. 1). However, it would be beneficial to better characterise exposure–response relationships with these health outcomes, to refine comparability with urban PM. Also, it would be helpful to consider the impact of biomass burning derived PM on a wider range of adverse health outcomes in the future (e.g. neurological and metabolic diseases). Furthermore, to date, the majority of epidemiology studies have considered the impacts of PM derived from wildfire events on health, with relatively fewer studies investigating the toxicity of other biomass sources. Therefore, future studies should consider how other sources of biomass burning can impact on health.

In Vivo Studies

In vivo (rodent) studies have been performed to identify whether the emission source influences the toxicity of PM. A summary of the experimental design and main findings of the in vivo studies that have been conducted to date to assess the toxicity of biomass burning–derived PM is presented in Table 2.
Table 2

A summary of in vivo studies which have investigated the toxicity of PM derived from biomass burning. The experimental design employed in the study and a brief overview of the findings are presented. Studies are presented in order of publication date




Particle dose(s) and route of exposure

Time point(s)

Endpoints investigated

Summary of findings


Dubick et al. [87]

Male Sprague-Dawley rats

Wood smoke derived from burning fir and pine in the laboratory

16.25 min exposure to woodsmoke via inhalation

1, 12, 24, 48 and 96 h

BAL: differential cell count (inflammation), and protein levels

Lungs, liver, heart and kidney: antioxidant enzyme activities (GP, GR, SOD), lipid peroxidation (TBARS)

SEM of lung tissue

No inflammatory response stimulated

Woodsmoke increased protein levels in BAL

Woodsmoke increased lipid peroxidation in the lungs and liver

No changes in SOD activity in the lung, but GP activity was higher following woodsmoke exposure

Damage to lung epithelium by woodsmoke

Woodsmoke can stimulate oxidative stress in the lungs

Seagrave et al. [88]

Male F344/Crl BR rats

PM sampling from 4 southeastern US sites which represented different contributing PM sources (e.g. wood burning, traffic, industry) during winter and summer 2004

Woodsmoke PM samples were collected during prescribed forest burns

0.75, 1.5 and 3 mg/rat

Intratracheal instillation

24 h

BAL: differential cell count (inflammation), LDH, protein, and alkaline phosphatase (lung injury)

Histopathology of lung tissue

Urban (traffic) sites stimulated the greatest inflammatory and cytotoxic response

Winter samples were typically more toxic than PM collected in summer

PM toxicity differed depending on the site and season, implying that specific constituents and/or sources influenced its toxicity

Wood smoke–derived PM was less toxic than other sources (e.g. traffic and industry PM emissions)

Wegesser et al. [52]

Male BALB/c mice

Peat wildfire PM (coarse and fine) collected from North California in June 2008

Ambient PM samples collected 1 year earlier from the same area (not during a wildfire event)

10, 25 or 50 μg/mouse

Intratracheal instillation

24 h

BAL: differential cell count (inflammation), protein levels

Histology of lung tissue

An inflammatory response (neutrophil accumulation) was activated by wildfire PM in the lungs

Protein levels increased by wildfire PM

Damage to lung tissue was observed using histology

Wildfire PM caused a greater (10-fold) toxic response than ambient PM from the same area

Wegesser et al. [89]

Male BALB/c mice

Peat wildfire PM (coarse and fine) collected from North California in June 2008

Ambient PM samples collected 1 year earlier from the same area (not during a wildfire event)

10, 25, 50 or 100 μg

Intratracheal instillation

6 and 24 h

BAL: differential cell count and cytokine production (TNF, MIP-1a, KC, IP-10) (inflammation), antioxidant content (total antioxidant capacity)

Increase in cytokine production and depletion of antioxidants stimulated by wildfire PM

Wildfire PM caused a depletion in antioxidants that was greater than ambient PM samples

Wildfire PM stimulated an inflammatory response in the lungs and caused a depletion of antioxidants

The toxic response is greater than that stimulated by ambient PM from the same area when there was no wildfire activity

Danielsen et al. [90]

Male Fischer 344 rats

Ambient PM samples were collected from 2 rural sites of Denmark with and without wood stove emissions (in January–February 2007)

Woodsmoke PM samples were also generated by woodstoves under high or low oxygen combustion conditions

0.64 mg/kg

Intratracheal instillation and oral gavage

24 h

BAL: differential cell count (inflammation)

Lung (following i.t. administration) and liver (following oral gavage) tissue: DNA damage and gene expression (cytokines (e.g. MCP-1, MIP-2, and antioxidants (e.g. HO-1))

PM caused an increase in expression of inflammatory and antioxidant genes. The greatest effects were stimulated by ambient air–derived woodsmoke PM and PM generated from wood burning during low oxygen conditions

DNA damage by PM is only observed in the liver following gavage

Woodsmoke PM can cause oxidative stress, inflammation and genotoxicity

Happo et al. [91]

Male C57BL/6J mice

Urban PM samples (coarse and fine) were collected in a series of sampling campaigns conducted in 6 European cities (Duisburg, Prague, Amsterdam, Helsinki, Barcelona, Athens) in winter 2002–summer 2003

*Tracers of biomass burning used to identify contribution of this emission source to toxicity

10 mg/kg

Intratracheal instillation: single and 3 repeated doses equivalent to 0.27 mg/animal or cumulative dose of 0.82 mg/animal for repeated exposure (on days 1, 3 and 6)

24 h

BAL: differential cell counts and cytokine levels (TNF, IL-6, KC) (inflammation), total protein and LDH levels (lung injury)

Histopathology of lung tissue

No inflammatory response observed for PM samples after a single exposure

After repeated exposure samples from Duisburg, Amsterdam, Helsinki, Barcelona, and Athens stimulated an inflammatory response

All PM samples stimulated an increase in protein levels

No changes in LDH levels observed

Small changes in cytokine production observed

PM from different cities varies in its toxic potency—this is likely to derive from differences in the emission source of PM

Toxicity after repeat dosing was greater than that observed after single exposure

Kim et al. [92]

Female CD-1 mice

Peat wildfire PM (coarse and fine) was collected from North California in June and July 2008 during smouldering and nearly extinguished (glowing) phases of combustion

100 μg/mouse

Oropharyngeal aspiration

4 and 24 h

BAL: differential cell count and cytokine production (TNF, IL-6, MIP-2), (inflammation), lung injury (LDH, protein and GGT levels, activity of NAG)

Wildfire PM stimulated an inflammatory response (neutrophil accumulation and cytokine production) and an increase in protein

ROS production by BAL cells was increased by wildfire PM

Wildfire PM stimulates a pulmonary inflammatory and oxidative response

Coarse PM stimulated a greater response than fine PM

Mirowsky et al. [93]

Male and female FVB/N mice

Ambient PM samples (coarse and fine) were collected from 5 urban and rural sites in California (which represented different emission sources including traffic, coal, industrial, biomass burning and soil dust)

*Sources/tracers used to identify contribution of different emission sources to toxicity

50 μg/mouse

Oropharyngeal aspiration

*In vivo response was compared with an in vitro model (lung HPMEC-ST1.6R endothelial cells and BEAS-2B lung bronchial epithelial cells), which assessed cytotoxicity and ROS production

24 h

BAL: differential cell count (inflammation), and protein levels

Inflammatory response (neutrophil infiltration) activated by many PM samples. Particle size influences the toxic potency of the PM samples (coarse samples were more potent than fine)

Protein levels in BAL were greater for urban and coarse PM samples than rural and fine PM

Toxicity of PM was influenced by site (emission source) and particle size. Urban PM samples were more toxic than rural PM, and coarse samples more toxic than fine. PM derived from traffic and soil was identified as being the most toxic

A lack of correlation between the in vitro and in vivo study was observed

Sussan et al. [94]

Male C57BL/6 mice

Ambient PM collected from homes (in India) during cooking with biomass (wood or wood and cow dung)

Acute 20, 50, 250, 500 μg/mouse

*Most endpoints assessed at a dose of 250 μg

Sub-chronic 50 μg/mouse

Intranasal instillation

Acute: 6, 24, 72 h

Sub-chronic: 8 weeks

BAL: differential cell count (inflammation), cytokine production (panel of 32 assessed)

Histology of lung tissue

Pulmonary mechanics—airway resistance and hyper-responsiveness

Both PM samples elicited an inflammatory response (neutrophil accumulation)

Sub-chronic studies revealed that the neutrophil response was followed by eosinophil recruitment

Both PM samples elicited similar cytokine profiles: neutrophil chemokines (G-CSF, KC and MIP-1α), macrophage chemokines (MIP-1β, IP-10, MCP-1) and cytokines (IL-6, TNF, IL-12 and MIG)

PM caused damage to the airway epithelium

Inflammatory responses and lung injury was greater in mice exposed to cow dung PM compared with wood PM

Biomass cooking emission source of PM can influence its toxicity. Cow dung PM was more toxic than wood PM

Plummer et al. [95]

Male BALB/C mice

Source-oriented sampling was performed. Ambient urban PM was collected from California (USA) during summer 2008 and winter 2009 which represented different sources (e.g. vehicular emissions, residential heating (woodstoves and fireplaces) and residential (e.g. BBQing) and commercial cooking

50 μg/mouse

Oropharyngeal aspiration

24 h

BAL: differential cell count (inflammation), LDH and protein levels

Many PM samples increased neutrophil infiltration into the lungs and increased LDH and protein levels

PM samples derived from vehicle emissions were most potent

Biomass-derived PM samples stimulated an inflammatory response

Seasonal effects were observed—summer PM samples were typically more toxic

Seasonal differences in PM toxicity were observed which is likely to be influenced by the emission source and particle characteristics (e.g. size)

PM derived from vehicular emissions was particularly toxic, but biomass PM could also elicit toxic responses

Kim et al. [49]

Female CD-1 mice

Biomass smoke during flaming or smouldering phases of combustion from 5 different fuel types (red oak, peat, pine needles, pine and eucalyptus) were collected

100 μg/mouse

Oropharyngeal aspiration

4 and 24 h

BAL: differential cell count and cytokine (TNF, IL-6, MIP-2) production (inflammation, LDH, protein, GGT, NAG levels (lung injury)


Mutagenicity assay (Ames test)

Inflammatory response activated (neutrophil accumulation). Flaming peat and eucalyptus samples stimulated the greatest response

IL-6, TNF and MIP-2 were elevated by the flaming peat sample only

Increase in markers of lung injury for flaming peat sample

Flaming peat and pine samples caused mutagenicity

The combustion phase and fuel source influenced the toxicity of biomass PM to the lung

Peat and eucalyptus samples were the most potent. Flaming samples were more toxic than smouldering samples

Kim et al. [96]

Female Balb/cJ mice

Three biomass fuels were combusted (eucalyptus, peat, red oak) and samples obtained during flaming and smouldering phases were exposed to mice via inhalation

4 and 40 mg/m3


4 and 24 h

Ventilatory function (plethysmography)

BAL: differential cell count and cytokine levels (TNF, IL-6, MIP-2) (inflammation), LDH, protein, GGT, NAG levels (lung injury)


Inflammatory response activated (neutrophil accumulation) by flaming peat and eucalyptus samples

No change in inflammatory cytokines

Lung injury markers increased and changes in respiratory parameters observed for peat and eucalyptus samples

No changes in haematological parameters

The combustion phase and fuel source influenced the toxicity of biomass PM following inhalation. Flaming samples were more toxic than smouldering samples

Peat and eucalyptus PM samples were most toxic

Findings obtained aligned with those observed following oropharyngeal aspiration [49]

BAL = bronchoalveolar lavage, GGT = glutamyl transferase, GP = glutathione peroxidase, GR = glutathione reductase, IL = interleukin, IP = interferon-gamma–induced protein, i.t. = intratracheal instillation, KC = keratinocyte-derived chemokine, LDH = lactate dehydrogenase, MIP-2 = macrophage inflammatory protein, NAG = N-acetyl-β-d-glucoaminidase, ROS = reactive oxygen species, SEM = scanning electron microscopy, SOD = superoxide dismutase, TBARS = thiobarbituric acid reactive substances, TNF = tumour necrosis factor

The PM samples tested in existing in vivo studies are diverse. Several studies have compared the pulmonary toxicity of ambient PM samples obtained from different sites and seasons in order to identify whether different PM emission sources cause differential toxicity (e.g. [88, 91, 95]). Seasonal differences in the toxicity of ambient PM are investigated as the emission source is likely to vary throughout the year (e.g. more biomass burning is expected in homes in winter months). Of interest is that when ambient PM samples are tested, these are mainly derived from EU and US locations, and thus, testing samples from a wider array of locations is recommended to understand the global relevance of the findings obtained to date. For example, the samples tested to date include wood burning from wildfires or heating, but in SE Asian countries, such as Thailand, biomass burning also includes agricultural burning of crop residues and fires from different types of vegetation. As the type of fuel burned may be a determinant of biological effects, and atmospheric conditions may impact the composition and toxicity of the samples, we may not be able to extrapolate effects from European and American samples to Asia. PM from the burning of different types of biomass has also been generated in the laboratory and its toxicity assessed in vivo (e.g. [49, 87, 92, 96]). In addition, the toxicity of PM samples emitted when cooking with biomass as a fuel source has been tested in one study [94]. Finally, PM has been collected during wildfire events to better understand the potential detrimental impacts of wildfires on health (e.g. [52, 89, 92]).

Taken together, the results of existing in vivo studies suggest that biomass burning derived PM can elicit toxicity to the lung, but that PM samples from different biomass sources vary in their toxic potency. In addition, PM samples derived from traffic sources are typically more toxic than biomass-derived PM samples. However, not all in vivo studies conducted to date have included a traffic PM sample and instead are more focussed on comparing the toxic potency of different types of biomass sources. We therefore suggest that in the future, when performing in vivo studies that investigate the toxicity of biomass burning–derived PM a sample which represents urban/traffic-derived PM should be included as this helps to identify whether biomass-derived PM has a similar toxic potency and will elicit similar adverse health outcomes.

Interestingly, despite differences in the experimental design employed across the different in vivo studies performed to date (e.g. species/strain of animal selected, PM source, PM sampling procedure, administration route, PM dose, time point and endpoints assessed), there is evidence that similar biological responses are activated in the lung in vivo following exposure to biomass derived PM. More specifically, activation of inflammation is commonly observed following pulmonary exposure to biomass burning derived PM samples of diverse sources, including biomass burning (Table 2). Investigation of pulmonary inflammation is prioritised for investigation in vivo as it is established that inhaled PM can activate a pulmonary inflammatory response which is likely to drive its adverse health impacts, as previously discussed. However, assessment of the involvement of oxidative stress in the toxicity of inhaled biomass burning derived PM has not been routinely assessed in existing in vivo studies. This is surprising as existing knowledge suggests that oxidative stress also plays a key role in PM toxicity (e.g. [97]). Therefore, whilst reactive oxygen species (ROS) production and antioxidant depletion have been assessed in some studies, a more comprehensive investigation of whether biomass PM stimulates oxidative stress is required in the future. More specifically, employing a battery of tests which assess ROS levels, antioxidant levels/activity, oxidant-mediated damage, and the ability of antioxidants to protect against PM-mediated toxicity can aid in unravelling the role of oxidative stress in PM toxicity [66]. This will help identify whether similar adverse health impacts are expected for biomass burning–derived PM as those known for traffic-derived PM.

Investigation of impacts of biomass burning–derived PM on the lung has been prioritised in existing studies, as the main route of exposure to PM is inhalation. However, only a limted number of inhalation studies have been performed to date (e.g. [87, 96]), with intratracheal instillation/oropharyngeal aspiration being most commonly employed as the route of administration (Table 2). This is not surprising given the expense of performing inhalation studies and requirement for specialised equipment and larger quantities of PM. Of interest is that Kim et al. [96] exposed mice to biomass PM from different fuel sources obtained during different phases of combustion via inhalation, and their findings aligned to those observed following oropharyngeal aspiration [49, 92]. However, ideally, it would be beneficial to conduct more inhalation studies as administration via this method is more physiologically relevant than other approaches. Furthermore, short- and long-term exposure of humans to biomass PM is expected. To date, the majority of studies have acute responses in rodents following a single administration of biomass PM at one time point, and few studies have investigated the toxicity of biomass burning–derived PM following repeated exposures (e.g. [91]). Therefore, there is a need to perform chronic in vivo studies which encompass repeated and longer-term exposure to biomass burning–derived PM to address this knowledge gap. It would also be beneficial to consider other target sites in future studies, as inhaled particles can stimulate systemic effects to promote the initiation and progression of disease in extra-pulmonary organs (e.g. reviewed by Stone et al. [36]).

To summarise, the findings from in vivo studies conducted to date suggest that biomass burning derived PM can stimulate pulmonary toxicity. The majority of studies have used inflammation as an indicator of toxicity and demonstrated that biomass burning–derived PM can stimulate an acute inflammatory response. There is also evidence that biomass derived PM can stimulate oxidant-driven responses; however, this endpoint is less routinely assessed. Several PM samples that represent different biomass sources are often tested in each study, with the findings suggesting that the biomass type can influence the toxicity of the emitted PM. Furthermore, when comparative studies are performed, traffic derived PM is often identified as being more toxic than biomass burning derived PM. There are gaps in knowledge that need to be addressed in order to gain a better understanding of the toxicity of PM in vivo. In particular, it is advised that longer-term, repeat dose studies are prioritised as to date, and studies have typically investigated acute responses following a single administration of PM at one dose.

In Vitro Studies

There is a drive to reduce the use of animals when performing toxicology testing, and thus, it is prudent to consider if in vitro models can be used as an alternative to rodent testing to predict PM toxicity. Table 3 summarises the experimental design and findings from in vitro studies that have investigated the toxicity of PM derived from biomass burning.
Table 3

A summary of in vitro studies which have investigated the toxicity of PM derived from biomass burning. A brief description of the experimental design employed in the study and an overview of the findings are presented. Studies are presented in order of publication date


Cell model


Particle concentration(s)

Time point(s)

Endpoints investigated

Summary of findings


Leonard et al. [98]

Mouse RAW264.7 macrophage cells

Woodsmoke particles collected from wood (pine and fir) burning in a furnace

10–100 μL/mL

5 min–24 h

Acellular ROS production (ESR)

DNA damage (DNA strand break assay—acellular, and the SCG assay)

Lipid peroxidation (TBA assay)

NF-κB activation (electophoresis mobility shift assay)

Cytokine (TNF) production (ELISA)

Woodsmoke PM can generate ROS

Woodsmoke PM caused DNA damage and lipid peroxidation

TNF production and NF-κB activation stimulated by woodsmoke PM

Woodsmoke PM can cause pro-inflammatory and oxidative effects in vitro

Jalava et al. [99]

Mouse RAW264.7 macrophage cells

Long-range–transported wildfire-derived ambient PM samples were collected in August–September 2002 in Finland. For comparison, seasonal average and mixed episode PM samples were collected

15, 50, 150 and 300 μg/mL

24 h

Cytotoxicity (MTT assay), apoptosis (propidium iodide staining), cytokine production (TNF, IL-6, MIP-2)

NO production

All PM samples stimulated a decrease in cell viability

Concentration-dependent increases in NO and cytokine production were observed for all PM samples

PM toxicity was size dependent, and the toxicity of long-range–transported particles was lower than that observed for the other sample types investigated

Karlsson et al. [100]

Human A549 alveolar epithelial cells

Human macrophages

PM samples were collected from biomass (wood—3 samples obtained from chimneys of one family house with different types of boiler and wood sources), urban (e.g. street, subway) and tire sources

Epithelial cells 40 μg/cm2

(equivalent to 70 μg/mL)

Macrophages 100 μg/mL (equivalent to 50 μg/cm2)

Epithelial cells 4 h

Macrophages 18 h

Epithelial cells: genotoxicity (Comet assay)


Cytokine production (IL-6, IL-8 and TNF)

All PM samples caused genotoxicity. Subway-derived PM elicited the greatest response. All wood-derived PM samples stimulated a similar genotoxic potency

Only 1 wood sample elicited an increase in cytokine production. Street-derived PM stimulated the greatest response

PM collected from different sources caused different biological responses. For genotoxic and inflammatory endpoints, the biomass-derived PM samples were less toxic than the urban PM samples. Subway-derived PM was the most toxic

Leonard et al. [101]

Mouse RAW264.7 macrophage cells

PM derived from wildfires in Alaska (USA) in March 2004

100 μg/mL

30 min

ROS production (hydrogen peroxide concentration)

Lipid peroxidation (MDA assay)

DNA damage (DNA strand break assay—acellular)

ROS production, lipid peroxidation and DNA damage increased by PM

Wildfire PM caused toxicity that was size dependent, with smaller particles stimulating a greater effect

Karlsson et al. [102]

Human A549 alveolar epithelial cells

PM samples were collected from biomass (wood), ambient urban (e.g. street, subway), tire and diesel sources

20 μg/cm2 (equivalent to 35 μg/mL), 40 μg/cm2 (equivalent to 70 μg/mL)

*Concentration-dependent on assay

2, 8 h

*Time point dependent on assay

Mitochondrial depolarisation (TMRE probe)

ROS production (DCFH assay)

Genotoxicity (Comet assay)

All PM samples caused mitochondrial depolarisation (apart from tire PM). Wood-derived PM caused the greatest effect

Only subway-derived PM stimulated ROS production

Subway PM caused genotoxicity via an oxidant mechanism (the other PM samples were not tested in this study)

PM collected from different sources caused different biological responses. Subway-derived PM was the most toxic and caused toxicity via an oxidant mechanism. Biomass (wood)-derived PM caused mitochondrial depolarisation but no ROS production

Guastadisegni et al. [103]

Mouse RAW264.7 macrophage cells

PM samples collected from 7 sites in the Netherlands and Germany (which varied with respect to their proximity to major roads and the level of traffic.

Sites in Italy and Sweden with PM largely derived from domestic wood burning were also included

20 and 60 μg/cm2

5 h

Cytotoxicity (LDH assay)

IL-6 and TNF production (ELISA)

Arachadonic acid (AA) release

No cytotoxicity observed

Inflammatory response (cytokine and AA production) not related to traffic—the strongest responses were observed in low traffic sites

Wood burning site induced a relatively large response

Hypothesised that samples from high traffic sources would be more toxic

PM collected from sites in close proximity to traffic was not more toxic than other sites (e.g. with low traffic, wood burning)

Danielsen et al. [90]

Human A549 alveolar epithelial cells and human THP macrophages

PM samples were collected from 2 rural sites of Denmark with and without wood stove emissions (in January–February 2007), and the exhaust of a wood stove

2.5–100 μg/mL (equivalent to 1.65–66 μg/cm2)

3–24 h

Cytotoxicity (LDH and WST-1 assays)

Acellular and cellular ROS production (DCFH assay)

Genotoxicity (Comet assay)

DNA damage (8-oxodG assay, DNA adduct formation)

Cytokine and oxidative stress response gene expression (TNF, IL-8, IL-6, LFA-1, MCP-1, HO-1 and OGG1)

No cytotoxicity observed for all samples

ROS generation was observed for all samples

All samples caused DNA damage, but the greatest response was observed for woodsmoke-derived PM

Upregulation of the expression of pro-inflammatory and oxidative stress response genes was observed for all samples

Wood stove–derived particles produced ROS, DNA damage and had higher levels of gene expression

PM collected from rural locations can have a large contribution of particulates from biomass burning

The toxicity of ambient PM samples was related to biomass emissions

Franzi et al. [104]

Mouse RAW264.7 macrophage cells

PM derived from a rural area in North Californian (Escalon) during wildfires (June 2008) and 1 year later (i.e. when there were no wildfires, July 2009)

2.5–50 μg

30 min, 1 h, 2 h, 4 h, 6 h and/or 24 h

Cytotoxicity (trypan blue)

NF-κB activation (reporter gene)

Concentration- and time-dependent reduction in cell viability for both PM samples. Wildfire-derived PM was more toxic than ambient PM

NF-κB activation was greater for wildfire PM than ambient PM

Wildfire PM exhibited a greater toxic effect to macrophages than ambient PM

Wong et al. [105]

Human bronchial epithelial cells (HBE)

Ambient PM collected from an urban area of the San Joaquin Valley (USA) in June 2006

Wildfire PM collected from a rural area in North Californian (Escalon) during wildfires (June 2008)

10 μg/mL

3 h

Gene expression (microarray and PCR) in the presence and absence of ROS inhibitors

Genes involved in detoxification, inflammation and oxidative stress were affected by PM exposure. Wildfire PM induced different responses to ambient PM. ROS inhibitors decreased expression of some inflammatory genes by PM

Cells respond differently to urban PM and wildfire PM

Tapanainen et al. [106]

Mouse RAW264.7 macrophage cells

Human BEAS-2B bronchial epithelial cells

PM generated from burning wood (dry birch) in 4 different combustion appliances

15–300 μg/mL

24 h

Cytotoxicity (PI staining, DNA content analysis, MTT and trypan blue assays)

Cytokine production (RAW cells: TNF and MIP-2, BEAS-2B cells: IL-1β, IL-6, IL-8, IL-10, IL-12p70, INF-γ and TNF) (ELISA)

Genotoxicity (Comet assay)

Cytotoxicity stimulated by all PM samples

Cytokine production increased for macrophages only

PM samples caused DNA damage in both cell types

Combustion appliance used to generate PM samples influenced their toxicity

The combustion appliance used to burn wood influenced the toxicity of the emitted PM

Corsini et al. [107]

Human A549 alveolar epithelial cells and human THP macrophages

PM samples from burning of biomass (fir or beech pellets) under laboratory conditions

Diesel exhaust particles (DEP) included for comparative purposes

12.5–100 μg/mL

3–48 h

Cytotoxicity (LDH assay)

Gene expression (IL-8) in the presence and absence of an uptake inhibitor (cytochalasin D)

Cytokine production (IL-8)

Signal transduction pathway activation (e.g. NF-κB, PKC, p38 MAPK)

Genotoxicity (micronucleus and Comet assays)

ROS production (DCFH assay)

Biomass-derived PM and DEP stimulated cytokine production. Beech-derived PM elicited a greater response than fir, but DEP exhibited the greatest response. Cytokine production from macrophages was greater than that observed for epithelial cells. Particle uptake and p38 MAPK activation played a key role in cytokine production

All PM samples caused an increase in DNA damage (Comet assay). DEP stimulated the greatest response. No genotoxic response was observed with the micronucleus test

No increase in ROS production was observed for biomass-derived PM

Wood-derived PM can stimulate toxic effects, although the biomass source may influence the toxic potency of the PM. In general, DEP induced a stronger response for the endpoints investigated

Perrone et al. [108]

Human A549 alveolar epithelial cells

PM was collected from 3 sites (urban, rural and remote (high altitude)) in Italy, from different seasons (spring, summer, autumn, winter) in 2008–2009

*Tracers for biomass burning measured in samples

6 μg/cm2

3–24 h

Cytotoxicity (LDH and MTT assays)

Cytokine production (IL-8)

ROS production (DCFH assay)

DNA damage (Comet assay)

All PM samples reduced cell viability, but samples from spring and summer had the greatest effect

Limited changes in IL-8 production were observed and were site specific

ROS production was dependent on the season—greater effects observed in summer and autumn. ROS production was driven by particles derived from biomass burning

Genotoxic potential of particles was related to location and season and traffic-derived particles were associated with effects observed

Seasonal differences in PM toxicity were observed

Biomass burning was identified as being a significant contributor to PM toxicity in the samples collected in this study

Miousse et al. [109]

Mouse RAW264.7 macrophage cells

PM samples from a variety of sites were collected including an underground parking building (i.e. traffic exhausts), biomass PM (from a fireplace burning wood), soil dust, agriculture dust, and pollen, in Arkansas (USA) from January–April 2014

5 or 50 μg/mL, equivalent to 2.94 and 29.9 μg/cm2

24–72 h

Cytotoxicity (LDH assay)

Gene expression (oxidative stress responsive genes)

Mitochondrial function (oxygen consumption rate)

Epigenetic effects (e.g. DNA methylation)

DNA damage (chromosome aberrations)

Soil and agricultural PM had the greatest detrimental impact on cell viability. Road dust, biomass burning and traffic exhaust PM samples had no effect on cell viability

All PM samples (with the exception of pollen) stimulated an upregulation of the expression of genes related to oxidative stress

Soil and agricultural PM had detrimental impacts on cellular respiration

Soil PM had the greatest detrimental effect on epigenetic changes and DNA damage

PM from different emission sources exhibits differential biological effects

Longhin et al. [110]

Human HBEC3-KT bronchial epithelial cell line

Diesel exhaust and biomass burning PM emissions were collected in laboratory conditions

2.5 μg/cm2

24 h, 72 h, 2 weeks (2 exposures, 1/week)

Acellular ROS production (cytochrome c assay)

Gene expression (related to cellular metabolism, inflammation and oxidative stress)

Cytokine production (IL-1β, IL-6, IL-8)

Cell morphology (light microscopy) and migration (scratch wound assay)

Diesel-derived PM were the most reactive (greatest ROS production)

Increase in detoxification enzyme expression for biomass- and diesel-derived PM. Only diesel PM stimulated a change in the other genes

No cytokine production observed for biomass exposed cells. Time-dependent cytokine production for diesel PM

Only diesel PM caused a changed in cell morphology

The emission source can influence the toxic potency of PM

Diesel PM induced a more toxic response than biomass-derived PM

Jin et al. [111]

Mouse RAW264.7 macrophage cells

PM10 samples were collected from five direct emission sources in China: wood (pine), crop residue (cornstalk), coal, cigarette smoke and diesel from a diesel engine

Two urban ambient PM10 samples were collected from Beijing and Wuwei

50, 100, 200 or 400 mg/mL

16 h

ROS generation (DHE assay)

Cytokine (TNF) production (ELISA)

Biomass-derived PM samples stimulated ROS production in cells. No significant change in ROS production by PM originating from coal, diesel or urban ambient air

Urban PM, coal and diesel samples stimulated the greatest level of TNF production. Lower levels of cytokine production observed for biomass-derived PM

PM from different sources varies in its toxic potency; biomass-derived PM could induce strong oxidative stress responses

Park et al. [112]

Human airway cell lines (A549, H292, BEAS-2B and SAEC)

Chinese hamster ovary (CHO-K1) cell line


PM from combustion (diesel engine, gasoline engine, biomass (rice straw and pine stem), coal burning) and non-combustion (e.g. road dust, sea spray) sources were tested.

Combustion-derived PM was generated in the laboratory

1–150 μg/mL

5 min–24 h

Acellular ROS production (DTT assay and ESR)

Cell viability (neutral red and WST-1 assays)

Genotoxicity (Comet assay and Ames test)

ROS production (DCFH assay)

Cytokine production (IL-6 and IL-8)

Biomass PM caused cytotoxicity, cytokine production and ROS production (acellular and cellular)

Diesel engine PM were the most toxic across all endpoints

The emission source of PM influenced whether it stimulated mutagenic, oxidative and inflammatory responses

Diesel particles were the most toxic particle, but biomass-derived PM was able to elicit toxicity

Van Den Heuvel et al. [113]

Human Beas-2B bronchial epithelial cells

PM10 samples were collected from an urban and rural background site in Belgium

Seasonal effects were investigated as samples were taken during the period April 2013–May 2014 (assumed greater biomass burning in winter)

*Tracers for biomass burning measured in samples

12.5, 25, 50, 100 μg/mL equivalent to 62.5, 31.3, 15.6, 7.8 μg/cm2

24 h

Acellular ROS production (EPR)

Cell viability (neutral red assay)

IL-8 production (ELISA)

Mutagenicity (Ames test)

Acellular ROS production greatest for the urban site

Concentration-dependent decrease in cell viability—no difference between potency of urban and rural site. Toxicity greatest in winter

IL-8 production was greatest for the urban site and during spring/summer

Mutagenicity was greatest for urban sites and in winter

Regional and seasonal differences in PM10 toxicity

Cytotoxicity and mutagenicity correlated with tracers of biomass burning

Cytokine production not correlated to biomass burning

Marchetti et al. [114]

Human A549 alveolar epithelial cells

Biomass fuels (pellet, charcoal or wood) was burnt in a stove in an open fireplace and PM collected

1, 2.5, 5, 7.5 and 10 μg/cm2

24 h

Cell viability (Alamar blue and propidium iodide assays)

Cytokine production (IL-6, IL-8)

HO-1 expression (western blot)

DNA damage (phosphorylated histone H2AX (γH2AX) and phosphorylated form of ataxia telangiectasia mutated (p-ATM))

Cell cycle (DNA staining)

Wood and pellet PM stimulated a decrease in cell viability, charcoal PM had no effect

Small changes in cytokine production were observed that were greatest for pellet PM

Increase in HO-1 expression caused by all PM samples

Only pellet PM caused DNA damage and alterations to the cell cycle

The emission source influenced the toxicity of PM

DCFH = 2′,7′-dichlorofluorescin diacetate, DEP = diesel exhaust particles, DTT = dithiothreitol, ELISA = enzyme-linked immunosorbent assay, ESR = electron spin resonance, LDH = lactate dehydrogenase, MDA = malondialdehyde, MIP-2 = macrophage inflammatory protein 2, MTT = (3-(4,5-dimethylthiazol-2-yl)-2,5-diphenyl-tetrazolium bromide), NF-κB = nuclear factor kappa B, NO = nitric oxide, IL = interleukin, PI = propidium iodide, ROS = reactive oxygen species, SCG = single-cell gel electrophoresis, TBA = thiobarbituric acid, TMRE = tetramethylrhodamine ethyl ester, TNF = tumour necrosis factor, WST-1 = water-soluble tetrazolium salt

To help identify the contribution of different emission sources to the adverse health effects of ambient PM, the toxicities of PM samples obtained from different sources have been compared in vitro. For example, ambient PM samples have been collected from different locations and seasons that reflect different emission sources (e.g. [90, 100, 102, 103, 108, 109, 111, 113]). Alternatively, biomass PM samples have been generated in furnaces in the laboratory and their toxicity assessed in vitro (e.g. [98, 106, 107, 110, 112, 114]). Finally, PM samples have been collected during wildfire events and their toxicity assessed in vitro [111–114]. As with in vivo studies, for comparative purposes, studies often, but not always, include PM samples collected from urban locations (e.g. to represent traffic derived PM) in order to identify whether the toxicity of PM from different sources is comparable.

Existing studies have focussed on assessing toxicity to the lung via investigation of responses of bronchial and alveolar epithelial cell lines (e.g. human Beas-2B and A549 cell lines) and macrophages (e.g. RAW264.7 (mouse) and THP-1 (human) cell lines). The cell types selected are relevant as PM is likely to interact with these cells following inhalation. Of benefit is that cells of human origin are frequently used in in vitro studies. However, to date, monocultures of cells have typically been used when investigating PM toxicity in vitro. Advanced in vitro 3D co-culture models are available that better represent the physiology of the lung (e.g. [115]), and therefore, it is recommended that these models are used more widely to assess PM toxicity in the future.

The findings from existing studies indicate that biomass burning derived PM can elicit toxicity in vitro (Table 3); however, there is a lack of agreement regarding whether the toxic potency of biomass burning derived PM is greater than that observed for other emission sources (e.g. traffic). The discrepancies observed between studies are likely to derive from differences in the experimental design employed, which are described in more detail below.

The sampling procedure used for PM collection may account for differences in the toxic potency of PM investigated between studies. Furthermore, the samples selected for investigation are very variable (e.g. use of laboratory generated vs ambient sources, fuel selected for biomass and combustion procedure used (e.g. furnace used, burning time, phase of burning when PM was collected)) which can influence their toxic potency. Of interest is that existing studies have primarily assessed the toxicity of PM derived from wood burning and wildfires. The majority of wildfire samples tested to date have come from one source: PM derived from peat wildfires in North Carolina in 2008. Of benefit is that the toxicity of these samples has been extensively tested across epidemiology, in vivo and in vitro studies (see below). However, ideally the toxicity of PM samples collected from wildfire events associated with the burning of different types of biomass would be tested. Such samples are not always available; thus, it is suggested that toxicity tests are performed which test the toxicity of biomass burning derived PM samples that are generated in the laboratory. This will help reveal the potential adverse health effects of more diverse types of biomass sources. In particular, it is noteworthy that there is a lack of studies which have assessed the toxicity of agricultural burning–derived PM, and this knowledge gap should be addressed in the future. Furthermore, the majority of existing (in vitro and in vivo) studies which have investigated the toxicity of environmental PM samples have been performed in the EU and US, and there is a lack of investigations of samples obtained from other countries. This is important as different emission sources are likely to contribute to PM emissions in different countries, and thus, a wider panel of ambient PM samples which reflects the global diversity of biomass burning should be considered in the future. In vitro studies not represented in Table 3 have assessed the toxicity of PM collected from different urban/traffic-derived and rural sources and frequently find seasonal differences in PM toxicity (e.g. [31, 32, 116]). As the toxicity of biomass burning–derived PM was not assessed in these studies, they are not discussed in-depth here. However, these studies provide additional evidence that the emission source can influence PM characteristics and toxicity.

The endpoints selected for assessment of toxicity, as well as the approaches used to assess the biological response of interest, may influence the toxicity of PM observed in vitro. Of benefit is that a range of endpoints are frequently assessed when investigating PM toxicity in vitro. By using a battery of approaches, a better understanding of the mechanism of toxicity can be gained to help identify potential adverse impacts on health. Interestingly, common markers of toxicity are assessed across in vitro studies, namely cytotoxicity, cytokine expression/production, oxidative stress (e.g. ROS production, antioxidant levels/activity, oxidant damage to proteins and lipids) and DNA damage. These endpoints have been prioritised due to existing knowledge regarding the mechanism of PM toxicity to the lung, as the activation of inflammation and oxidative stress is key to its ability to elicit adverse health outcomes (e.g. [35, 36, 70, 117]). Existing studies have shown that biomass PM can elicit toxicity via the same mechanism to traffic-derived PM (Table 3). This suggests that the detrimental impacts on health that have been widely studied for traffic PM may be applicable to biomass-derived sources. It is noteworthy that different experimental approaches can be used to assess each endpoint of interest and that a consistent approach was not used across studies for each individual endpoint. This is important when making comparisons between studies as the different approaches used to assess the same biological response may vary in their sensitivity. Standardisation of protocols for assessing specific outcomes using particular methods would therefore improve our ability to compare toxicity of different types of PM.

The concentrations of PM and time points tested are variable across in vitro studies. Ideally, a range of PM concentrations would be tested in each study in order to generate dose–response curves, but on several occasions, only one particle concentration is tested. It is therefore recommended that several PM concentrations are tested in each study. The two most commonly used dose metrics are μg/cm2 or μg/mL. On some occasions, but not always, researchers express the particle concentration in both ways, which allows comparisons of PM toxicity to be made between studies. There is debate regarding how particle concentration should be expressed in vitro, and other dose metrics may also be relevant (e.g. surface area) [118, 119]. In vitro studies performed to date have focussed on assessing toxicity following a single exposure, with cell responses typically assessed at 24 h post-exposure. There are fewer longer-term studies which have assessed PM toxicity following repeated exposure, and it is recommended that this knowledge gap is addressed in the future, as this would be more representative of how people are really exposed to biomass-derived PM.

The endpoint being investigated often dictates which time points are selected in each study. For example, earlier time points (e.g. 2–4 h) may be selected to investigate cytokine gene expression, whereas longer time points (e.g. 24 h) are often selected when assessing cytokine production. However, consistent time points are not used between different studies for the same markers of toxicity which makes it difficult to make comparisons between studies. Of interest is that investigation of toxicity of PM to the lungs has been prioritised in studies performed to date, but it would be beneficial to investigate PM toxicity to a more diverse range of target sites in vitro (e.g. cardiovascular system, liver, CNS).

To summarise, the results of in vitro studies performed to date suggest that biomass burning–derived PM can elicit toxicity to the lung. Existing studies have consistently demonstrated that that biomass burning–derived PM can activate inflammatory, oxidative and genotoxic responses. As this aligns with the mechanism of toxicity of traffic PM, it is suggested that biomass-derived PM may stimulate similar adverse health impacts (outlined in Fig. 1). However, it is evident that the PM source can influence its toxicity potency and thus not all biomass sources will emit PM that is equally toxic. In the future, a greater diversity of biomass-derived PM samples needs to be assessed within in vitro studies.

In Vivo Versus In Vitro Models

Of benefit is that some studies have investigated the toxicity of the same biomass burning derived PM samples across different experimental models. For example, epidemiology, in vivo and in vitro studies have been performed with PM samples obtained from the same wildfire event in North California in June 2008, with pulmonary and cardiovascular toxicity observed across all models used [52, 73, 89, 92, 104, 105]. Briefly, epidemiology studies identified that an increase in hospital admissions (for respiratory disease (e.g. asthma, COPD, infectious disease) and heart failure) was observed during the time of this wildfire incident [73]. In vitro and in vivo studies demonstrated that wildfire PM could elicit toxicity via inflammation and oxidative stress. Therefore, there was agreement across models that wildfire PM could elicit adverse health effects.

However, the findings from in vitro and in vivo studies which investigate biomass burning derived PM do not always agree. For example, Mirowsky et al. assessed the toxicity of ambient PM samples from urban and rural sites in California (which represented different emission sources) in vivo (mice exposed via oropharyngeal aspiration) and in vitro (pulmonary endothelial and bronchial epithelial cells). The in vivo studies assessed whether PM samples activated inflammatory responses, and the in vitro studies investigated ROS production [93]. There was a lack of correlation between the in vitro and in vivo response. However, whilst oxidative stress and inflammation are linked, the endpoints being assessed in each model were not directly comparable, which may explain the lack of correlation between models. For example, it may have been more advisable to compare cytokine production in vitro with the in vivo inflammatory response.

Accordingly, it is recommended that more studies are performed that compare the response of in vitro and in vivo models to biomass burning–derived PM, to identify the limitations of each model and to inform what models should be prioritised when investigating PM toxicity in the future. Ideally, a tiered testing strategy would be employed in the future when assessing biomass burning–derived PM toxicity that promotes the use of alternative (non-rodent) models before progressing to animal testing.

Toxicity of Biomass Burning Derived PM in Thailand: Directions for Future Research

Exposure to PM emitted from biomass burning from many sources (e.g. domestic cooking and heating, wildfires, agricultural) is anticipated to increase in both developing and developed countries. There is a growing body of epidemiological and toxicological (in vivo and in vitro) evidence that demonstrates biomass burning derived PM can elicit adverse health effects and that the biomass source can influence its toxic potency. Existing experimental studies to date have typically focussed on impacts on the lungs and demonstrated that biomass derived PM can stimulate oxidative stress, inflammation and genotoxicity. Thus, it is likely that biomass-derived PM will stimulate similar adverse health outcomes to urban (e.g. traffic)-derived PM (outlined in Fig. 1), which is known to cause toxicity via similar mechanisms. However, more comparative studies, which assess the toxicity of biomass burning and traffic-derived PM, are required as it is still unclear as to whether urban derived PM is more or less toxic than biomass burning derived PM as inconsistent findings have been observed to date in the available literature. The discrepancies observed are likely to arise due to differences in the experimental design employed between studies. Of interest is that studies conducted to date have focussed on investigation of acute effects, thus longer-term, repeat dose studies are required to fill knowledge gaps. Furthermore, the toxicity of different sources of PM are often assessed at the same concentration in laboratory toxicity tests, but the duration and frequency of exposure to biomass burning derived PM can vary, depending on the source. In addition, concentrations of biomass burning derived PM in the environment (e.g. during wildfires) can be much higher than PM from other sources. Thus, investigation of the toxicity of biomass PM over a wider range of concentrations would be advisable when performing experimental studies. To date, a focus has been placed on investigating potential impacts of biomass PM on the lungs and CV system. However, there is emerging evidence that traffic-derived PM can negatively impact on other body systems, and thus, the ability of biomass-derived PM to cause wide ranging health effects requires consideration in the future.

The toxicology studies that have been conducted to date on biomass burning–derived PM can help infer potential health impacts associated with biomass burning in Thailand. However, in order to better understand the health implications of biomass burning in that country, more research is required, and we advocate that this should encompass epidemiology studies, as well as experimental studies. In particular, there is a lack of epidemiology studies which have investigated the health impacts of different sources of biomass burning–derived PM in SE Asia. Accordingly, it would be beneficial to identify whether agricultural burning episodes elicit similar adverse health impacts to those associated with traffic air pollutants. Such studies should encompass consideration of different regions in Thailand and be performed over different seasons to capture different factors which may influence the toxic response (e.g. the biomass source being burnt including, for example, the type of crop being burnt, and whether forest or agricultural land is being burnt, and whether burning is performed pre- or post-harvest). Undertaking epidemiological studies examining well-established health impacts with traffic-related PM during the burning periods would help better understand which health impacts may be associated with PM emitted from biomass burning–derived sources. Such findings could also help provide more precise estimates in health impact assessments from medium- to longer-term exposure at a population level in Thailand. For experimental studies, there is a need to better capture PM sources that are relevant to Thailand in toxicology studies. This could be achieved via the collection of ambient PM samples from a variety of locations in Thailand which represent different emission sources of biomass burning–derived PM. For example, it is recommended that samples should be taken during haze and non-haze periods for comparative analysis of both PM physico-chemical characteristics and PM toxicity and, ideally, in concert with an epidemiological study. It is also advised that the toxicity of these biomass-related PM samples should be compared to traffic-related PM samples collected in Thailand. To better understand whether specific types of biomass have differential toxicity, PM samples could also be generated in the laboratory, by burning biomass samples that are relevant to Thailand (e.g. Bambusa vulgaris, Echinochloa crus-galli, Tectona grandis, Diptero carpusalatus, Zea mays, Oryza sativa, Dimocarpus longan, and Litchi chinensis [120]) and testing the toxicity of the emitted PM in vitro and in vivo.


There is currently a drive to reduce the level of ambient PM (globally) to improve air quality and protect human health. However, assessment of air quality does not currently distinguish between the varied toxicity of PM from different sources. Epidemiology, in vivo and in vitro studies have consistently demonstrated that biomass-derived PM can elicit toxicity. Consideration of the properties of PM which confer toxicity was out with the scope of this review but has been discussed previously (e.g. [10]). It is likely that differences in the toxic effects of PM emitted from different sources/locations derive from differences in the properties of the emitted particles (e.g. particle size (and surface area) and composition (e.g. PAH, metal and endotoxin content)). Therefore, a better understanding of how the emission source influences the toxicity and physico-chemical properties of PM, as well as exposure patterns, could better inform intervention strategies to improve air quality. This will require that a thorough assessment of the physico-chemical properties of PM is performed in parallel to toxicology studies. Given the current state of knowledge, it is appropriate for health impact assessment studies to assume that all particles, regardless of their source, have the same potential to cause disease.

Emissions of PM from biomass burning are rising globally, and thus, it is timely to consider what the impacts of this are for human health. There are many sources of biomass burning that can influence ambient PM levels. In Thailand, the current permanent air quality monitoring is limited in some areas, and thus, it is difficult to quantify the true impact of air pollution from different sources across the country. Additional monitoring, combined with modelling and satellite observations, would improve our understanding of biomass burning in the region. In Thailand, and SE Asia as a whole, large populations are affected by biomass-derived haze events, and thus, it is noteworthy that few epidemiological and experimental (in vitro and in vivo) studies have been conducted in this region to identify what risks may be posed to health; further research is therefore needed.



The authors would like to thank Prof Vicki Stone (Heriot-Watt University) for her valuable feedback on the manuscript.

Funding Information

This review was prepared as part of the TAPHIA (Thailand Air Pollution and Health Impact Assessment) project which is jointly funded by the Thai Research Fund (TRF) and the UK Medical Research Council (MRC) through the Newton Fund (Grant number MR/R006210/1).

Compliance with Ethical Standards

Conflict of Interest

On behalf of all authors, the corresponding author states that there is no conflict of interest.

Human and Animal Rights and Informed Consent

This article does not contain any studies with human or animal subjects performed by any of the authors.


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© The Author(s) 2019

Open Access This article is distributed under the terms of the Creative Commons Attribution 4.0 International License (, which permits unrestricted use, distribution, and reproduction in any medium, provided you give appropriate credit to the original author(s) and the source, provide a link to the Creative Commons license, and indicate if changes were made.

Authors and Affiliations

  • Helinor J. Johnston
    • 1
    Email author
  • William Mueller
    • 2
  • Susanne Steinle
    • 2
  • Sotiris Vardoulakis
    • 2
  • Kraichat Tantrakarnapa
    • 3
  • Miranda Loh
    • 2
  • John W. Cherrie
    • 1
    • 2
  1. 1.School of Engineering and Physical Sciences, Institute of Biological Chemistry, Biophysics and BioengineeringHeriot-Watt UniversityEdinburghUK
  2. 2.Institute of Occupational MedicineEdinburghUK
  3. 3.Department of Social and Environmental Medicine, Faculty of Tropical MedicineMahidol UniversityBangkokThailand

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