Indoor particulate matter

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Katleen De Brouwere and Rudi Torfs, VITO

Introduction and relevant exposure metrics

Indoor air quality is recognized as a priority in the European Action Plan on Environment and Health 2004-2010. The importance of indoor air quality is well understood: people spend about 90% of their time indoors, and this becomes particularly important for vulnerable groups like babies, children and aged people or people already suffering from, e.g., respiratory or allergic diseases. Indoor air contains a mix of pollutants, many of which are identified as potential health hazards.

Whereas the harmful effects of outdoor PM have been widely investigated (e.g. the ExternE, CAFE and APHEIS), this is not the case for indoor PM. We can think of 3 reasons why health impact assessment on indoor PM is less elaborated than health impact assessment of outdoor PM:

  • there is in general a longer tradition in the field of ambient air quality compared to indoor air quality
  • there is a substantial body of epidemiological evidence for health effects of ambient air, supported by harmonized methods (like the APHEA study), and EU-wide studies
  • there is a strong policy trigger, with long-term cohort studies making ambient PM a high priority. No such chronic studies exist for indoor PM.

Mass concentrations such as PM2.5 and PM10 are most widely used as measures for PM. This size fractionation is relevant in view of health perspective because smaller particles can penetrate deeper in the lungs. Nevertheless, particle size is only one part of the story. Chemical and toxicological characterization of PM – which is not the standard approach – needs also to be considered to obtain a better understanding of health effects of indoor PM (Fromme et al., 2006).

Indoor concentrations and personal exposures

In contrast to the assessment of ambient air concentrations, where measured or modelled concentrations at fixed-point monitors can be considered as representative for the ambient air exposure for a larger area, this is far not the case for indoor air. Empirical source – exposure models (e.g. Baxter et al., 2007) based on outdoor PM and factors predictive for indoor PM concentrations due to indoor sources (woodstove use, gas stove use, occupants density, humidifiers,…) can be used to predict indoor PM levels. However, since such indoor source – exposure models require information such as the presence of potential indoor sources, their use frequency, ventilation rates of the dwellings,… which is generally unknown – these models have only limited use to obtain a distribution or mapping of prevailing indoor PM concentration for a given region (especially beyond the boundaries of the model). Instead, indoor PM concentrations and personal exposures rely more on measured data than on modelling.

EU wide projects such as EXPOLIS included a measuring campaign of indoor PM and personal exposure to PM (Götschi et al., 2002). The focus is in many cases on indoor PM, however, the measuring of personal PM recently gained attention (Janssen et al., 2005).

Health effects of short-term exposures

The vast majority of epidemiological particulate matter-health studies is related to ambient PM. Much less is known about health effects of indoor generated PM (Jones, 1999). According to Schneider et al. (2003) who reviewed the literature of indoor PM and associated health effects, there is inadequate scientific evidence that PM measurements (mass or numbers) are useful as predictors of health risk of indoor PM, and thus no indoor PM – exposure response function can be derived. More research is necessary to address the indoor PM toxicity and health effects. Extrapolating the ambient PM exposure response functions to indoor PM could be used in a first stage as a substitute for the unknown indoor PM exposure response functions. This approach assumes that ambient and non-ambient PM are equally toxic. However, this assumption is questionable since the composition of indoor PM may deviate of outdoor PM. We could not find any study that compares the concentration - exposure – health effects relationship between indoor and outdoor PM at an epidemiological level. Notwithstanding the lack of epidemiological data, there is an indication from an vitro study of Long et al. (2001) that indoor-generated particles may be more bioactive than ambient particles. In conclusion, there is currently insufficient knowledge on the health effects of short term exposure. Methodological constraints specific for indoor PM concentrations assessment rely on the basis of this gap. While for outdoor PM, daily variation at fixed monitoring stations are used as a proxy for variations in short term exposure, (and can be correlated with daily variations in mortality and morbidity), no such tools are available for indoor.

Health effects of long-term exposures

Although an overview of health effects of long-term exposure to indoor PM is not feasible given the low numbers of such epidemiological studies, there is a substantial number of studies associating one specific source of indoor PM, namely woodstove use, with health effects. These studies could serve as an example for the magnitude of health effects related to indoor PM, though it should be kept in mind the indoor PM source – exposure – health effect relationship of woodstoves might deviate from that of other indoor PM sources. In addition, it is not so illogical to treat the exposure – health effects relationship source by source (here source: woodstove) since the source will affect the composition and hence the toxicity and health effect relationship of the indoor PM. In the major part of these woodstove indoor air – PM health effect studies, no direct measurements of PM are performed, but instead, the presence and eventually the use frequency of woodstoves are used as a proxy for indoor PM levels. Exposure to indoor sources of particulate matter, from combustion of biofuels (wood, charcoal, agricultural residues, dung) have been clearly associated with respiratory disease in developing countries (Smith et al., 2000) However, exposures in developed countries – and related health effect- are far below that of developing countries given the large differences in housing conditions (e.g. presence of chimneys, other ventilation systems,…). European studies on health effects of wood smoke exposure are very scare. A Spanish study reports that among 120 women hospitalized in the Hospital de Mar in Barcelona for an exacerbation of chronic obstructive pulmonary disease (COPD), there was a strong association between exposure to wood and charcoal exposure and COPD, after adjustment for age and smoking (Orozco-Levi et al., 2006). The cone of the other large-scale (5979 subjects) recent European study on this topic (Lissowska et al., 2005) was in Central and Eastern Europe (Czech Republic, Hungary, Poland, Romania, Russia and Slovakia), though the United Kingdom was also included. In that study, the odds ratio (OR) of lung cancer associated with solid fuel use was 1.22 (95 % CI :1.05-1.44) for cooking or heating, 1.37 (95 % CI: 0.90-2.09) for solid fuel used only for cooking, and 1.24 (95 % CI:1.05-1.47) for solid fuel used for both cooking and heating. Among the persons that used solid fuels for cooking during their whole lifetime, the OR increased to 1.8 (95 % CI: 1.36-2.41). In a review paper, Naeher et al. (2007) compiled a large number of northern U.S. and Canadian studies related to health impacts of residential wood smoke. In most of these studies, the occurrence of health problems (asthma, lower respiratory illness, shortness of breath, otitis media,… is compared between exposed and control persons. For example, Honicky et al. (1985, cited by Naeher et al., 2007) compared respiratory symptoms in 31 children who lived in homes with wood stoves with 31 children who lived in homes without wood stoves in Michigan. The two groups did not different with respect to mild symptoms but differed significantly for severe symptoms A similar study was conducted in Boise ID in 1989 by Butterfield et al. (1989, cited by Naeher et al., 2007). Respiratory symptoms were tracked in 59 children under the age of 5 1/2 years during a winter season. Symptoms such as wheeze, cough, and nocturnal awakening were associated with wood smoke exposure. Health effects of indoor air pollution including woodstove use on adult persons in Denver (18 to 70 years) were reported by Ostro et al. (1994, cited by Naeher et al., 2007). The use of wood fires was associated with an increase in daily moderate or severe shortness of breath. On the other hand, there are also some studies that have failed to find associations between woodstove emission exposure and respiratory health. Eisner et al. (2002) found no clear relationship between gas stove use or woodstove exposure and asthma health outcomes for a sensitive population group, i.e. 349 asthma patients. In conclusion, no general indoor PM – health effect relationships can be established. Nevertheless, based on the health effects associated with one single indoor PM source (woodstove), we can bring the following health effects forward as likely being affected by indoor PM:

  • exacerbation of chronic obstructive pulmonary disease (COPD),
  • lung cancer
  • respiratory symptoms: asthma, lower respiratory illness, shortness of breath, otitis media, wheeze, cough,…

What health endpoints might be quantified?

A rather wide range of severe (COPD, chronic bronchitis) and milder respiratory morbidity endpoints and allergies such as cough, atopy, asthma, wheezing have been quantified for the PM indoor source of woodstove. The effects of these symptoms might be quantified for other indoor PM sources, or for total indoor PM concentrations (originating from the sum of all indoor sources and from infiltration of ambient PM).

Is there evidence of a threshold or ‘safe’ level of indoor pm?

No, there is no evidence of a threshold or ‘safe’ level of indoor PM. This is not deduced from direct evidence of indoor PM- health effects studies since data are lacking but instead as a likely parallel to the evidence of no safe level for ambient PM. Applying the precautionary approach, and given that in vitro studies (e.g. Long et al., 2001) suggest that indoor particles are more bioactive that ambient particles, there is no evidence of a threshold or ‘safe’ level of indoor PM.

In addition, no threshold or ‘safe’ level of indoor PM has been published in the literature or suggested by any expert group.

What sub-groups of the population are most susceptible or otherwise will need special consideration in quantification?

A major part of the woodstove indoor PM studies has focused on the sub-groups of (small) children and asthmatic patients. These groups will need special consideration in quantification.

The odds ratio is defined as the ratio of incidence of a health effect (here: lung cancer) in the exposed group (e.g. solid fuel users, namely wood and coal) to the incidence of the control group (here: non-users of coal and wood).


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