Indoor air quality (IAQ) is an important component of indoor environmental quality. It has many facets. This chapter focuses on the chemical and particulate pollutants that can be found suspended in air or deposited on or sorbed to indoor surfaces. It specifically addresses organic and inorganic volatile and semivolatile molecular pollutants, and particulate matter. In the case of particles, abiotic materials are emphasized, but there is a brief discussion of allergens associated with pollen and of respiratory health risks associated with algal blooms after floods. IAQ problems associated with moisture and dampness of buildings are addressed in Chapter 5, and biologic IAQ concerns associated with microbial agents, insects and arthropods, and mammals and concerns that arise because of efforts to control them are discussed in Chapter 6.
With regard to the pollutants considered in this chapter, there is little in the published literature that considers together all the key elements in this committee’s charge: the effects of climate change on IAQ that would influence public health. However, substantial research has been published on many important components. For example, there is a strong emerging literature on the effects of climate change on outdoor air pollutants (Jacob and Winner, 2009), such as particulate matter (Tagaris et al., 2007) and ozone (Bell et al., 2007; Hogrefe et al., 2004a; Racheria and Adams, 2009), and on related health effects (Kinney, 2008; Tagaris et al., 2009). A voluminous literature characterizes health risks associated with pollutants in outdoor air (Bell et al., 2004; Dockery et al., 1993; Jerrett et al.,
2009; Pope and Dockery, 2006; Pope et al., 2009). Considerable published research documents our understanding of indoor–outdoor relationships of important air pollutants, including particles and ozone (Jia et al., 2008b; Monn, 2001; Wallace, 1996; Weschler et al., 2000). Research has explored the extent to which health risks associated with outdoor pollutants are a consequence of indoor exposures (Weschler, 2006; Wilson and Suh, 1997; Wilson et al., 2000). A large body of work reports on how indoor pollution sources influence IAQ and human health (Jones, 1999; Samet et al., 1987, 1988), including a National Research Council report published three decades ago (NRC, 1981).
The following sections discuss how indoor air pollutant levels might be influenced by climate change. The discussion is organized according to pollutant source category and pollutant class, considering first indoor emission sources and second pollutants of outdoor origin. The treatment is not intended to be comprehensive, but rather broadly illustrative of important IAQ concerns that might be influenced by climate change. Although most of what follows is related to conditions in buildings of the types commonly found in the United States, the chapter concludes with a discussion of an important international public-health problem: exposure to smoke from the indoor combustion of solid biomass and coal in developing countries.
Indoor environments detain pollutants that are emitted indoors. This section reviews important IAQ issues that are associated with indoor pollutant sources and explores how climate change might affect these issues. The emphasis is on conditions in the United States but the discussion is relevant for other countries with similar levels of economic development and similar buildings.
Pollutants from Indoor Combustion
Pollutants released into indoor air cause roughly 100–1,000 times greater human inhalation exposure or dose per unit mass emitted than pollutants released into outdoor air (Smith, 1988). That important observation has been expressed in terms of “intake fraction” (Bennett et al., 2002; Nazaroff, 2008), the ratio of the mass of a pollutant inhaled by an exposed population to the mass of the pollutant emitted from a source. The significance of that point in the present context is that sources have a much larger effect on public health if their pollutants are emitted indoors rather than outdoors. The much higher intake fraction for indoor emissions compared to those outdoors leads to the understanding that small-scale combustion
processes that do not burn much fuel can nevertheless raise substantial IAQ concerns and adversely affect public health.
Combustion might be the most important source of air pollution. Indoor combustion for cooking, lighting, and heating has a long and diverse history of contributing to air-pollution exposure. Lopez et al. (2006) ranked “indoor air pollution from [burning] solid fuels” as one of the top 10 leading causes of global mortality and disease. That ranking is based mainly on the use of biomass and coal in rural parts of developing countries. Unvented or incompletely vented combustion also occurs to a substantial extent in developed countries and has demonstrable effects on indoor pollutant concentrations and exposures. Evidence associating those exposures with public-health consequences ranges from suggestive to clear and compelling. Exposures resulting from indoor combustion could be altered in the future in several ways associated with climate change. Influencing factors could include changing prevalence, frequency, or strength of indoor emission rates and also changes in building ventilation conditions.1 The following paragraphs summarize some of the concerns and provide references to document the nature and importance of the current problems.
Accidental Carbon Monoxide Poisoning
Carbon monoxide (CO) is produced by the incomplete combustion of a carbonaceous fuel. Inhaled CO forms carboxyhemoglobin in the blood, whose presence interferes with transport and delivery of oxygen to tissues and organs. Excessive acute exposures result in illness or death. Chronic lower-level exposures may also have health consequences, but the available empirical evidence is weaker than that for acute poisonings.
CO is regulated as a pollutant in ambient air. Mainly through strong improvements in automotive emission-control technology, urban air CO levels have become well controlled, and almost every area of the United States meets the National Ambient Air Quality Standard for CO (EPA, 2010b).
Despite improvement in outdoor levels, CO remains an important air pollutant. Over the past few decades, hundreds of accidental and fatal acute CO poisonings have occurred each year in the United States (Cobb and Etzel, 1991; King and Bailey, 2008; Mott et al., 2002). The incidence has declined substantially. One important factor is improvements in the control of motor-vehicle emissions. Mott et al. analyzed CO-associated mortality statistics and concluded that, “if rates of unintentional CO-related deaths had remained at pre-1975 levels, an estimated additional 11,700 motor-vehicle-related CO poisoning deaths might have occurred by
1 Building tightening and reduced ventilation rates are further discussed in Chapter 8.
1998.” Holmes and Russell (2004) remarked that the reduction in accidental deaths resulting from improvements in motor-vehicle emission controls “is not accounted for in EPA’s [the Environmental Protection Agency’s] recent reports on the benefits and costs of the [Clean Air Act], yet it dwarfs the estimated direct benefits ascribed to CO control.” In a detailed study of CO poisoning deaths in California during the period 1978–1988, Girman et al. (1998) found that alcohol was a factor in 31% of the cases and that important combustion sources other than motor vehicles included heating or cooking appliances, charcoal grills and hibachis, small engines, and camping equipment. An assessment for Florida over the period 1999–2007 revealed that accidental CO poisonings “were primarily due to motor vehicle exhaust (21%–69%) and generator exposure (12%–33%), and the majority (50%–70%) occurred within the home” (Harduar-Morano and Watkins, 2011).
In the context of climate change, a particular concern about CO exposure arises from the use of emergency electricity generators that burn liquid fuels, such as gasoline. The use and reliability of centrally generated power might be degraded because of climate change for several reasons. For example, hotter summer afternoons may lead to more intense use of air conditioners and thus increase the frequency of service-demand overloads that cause brownouts and blackouts. Severe storms can also cause electricity service disruption. In such cases, people may rely more heavily on their own electricity generators. If the generators are used indoors, or even outdoors but too close to an indoor environment, unhealthful CO exposures can result. Increases in emergency-room and other hospital visits caused by CO poisoning have been reported in association with power outages (Muscatiello et al., 2010), major storms (Van Sickle et al., 2007), and floods (Daley et al., 2001).
A staff report from the Consumer Product Safety Commission (Hnatov et al., 2009) indicated that in 2005 an estimated 27 generator-related CO fatalities were associated with five hurricanes (Katrina, Rita, Wilma, Dennis, and Isabelle). And an estimated 21 generator-related CO fatalities were associated with ice storms, including major storms in the midwestern United States in January and in the Carolinas and Georgia in December.
In addition to electricity generators, shifts in fuel-use patterns during power outages may contribute to increased indoor CO levels. Of concern would be the use of natural-gas–fueled and petroleum-fueled stoves for heating, excessive reliance on unvented combustion-based space heaters, and use of charcoal briquettes or wood stoves indoors for cooking (Hampson and Stock, 2006, Hnatov, 2009).
One expects there to be many more poisonings that result in illness than in death. Analyses of the demand for poison control center services reveal a pattern similar to that in emergency rooms. Klein et al. (2007)
noted a nearly 50% increase in suspected CO poisoning calls in the days after a widespread blackout on the East Coast of the United States in 2003, and Forrester (2009) found more such calls in the counties that were in the disaster area declared for Hurricane Ike than in other counties in Texas. It is reasonable to believe that the prevalence of CO-induced illness is larger than that recorded in the emergency-room statistics because illnesses that are not considered severe might not be reported. A recent study evaluating the use of a web-based query system for public health surveillance reported almost 25,000 CO-related hospitalizations across the United States in 2005, of which approximately 4,200 were confirmed CO-related poisonings (Iqbal et al., 2010). These data were intended to exclude intentional and fire-related CO exposures.
Other factors may also contribute to increased public health risks associated with indoor CO exposures. For example, the Department of Housing and Urban Development’s 2009 American Housing Survey found that just 36% of homes nationwide reported having a working CO detector.2 People of lower socioeconomic status may be more likely to use stoves or unvented space heaters as a heat source (CDC, 1997) and less likely to have working CO detectors (Runyan et al., 2005). Some groups may hold mistaken beliefs about CO. For example, a survey conducted among residents of low socioeconomic status in northern Mexico by Galada et al. (2009) found that a large majority of respondents mistakenly believed that CO could be detected by sight or smell.
Cooking causes air-pollutant exposures that have potential public-health significance. The most severe problems occur from burning of solid biomass fuels or coal, especially in unimproved stoves, in the rural parts of developing countries. The relationship of those concerns to climate change is discussed toward the end of this chapter. However, even when relatively clean fuels are used for cooking in developed countries, indoor air-pollutant exposures with potential public-health consequences can arise. For example, the use of natural gas as a cooking fuel is associated with increased indoor exposures to nitrogen dioxide (NO2), a byproduct of the combustion process (Marbury et al., 1988; Spengler et al., 1994). In a study in the United Kingdom, the use of gas cooking appliances, rather than electric, was associated with respiratory morbidity in women (but not men, possibly women had higher exposure) (Jarvis et al., 1996). Exposure of children to higher indoor NO2 levels has also been reported to be associated with re-
2 As of January 2010, 25 states—including Florida, Texas, and California—required some or all residences to have CO detectors (National Conference of State Legislatures, 2010).
spiratory symptoms (such as wheeze) but not pulmonary function (Neas et al., 1991). In a population of infants at risk for asthma, “the frequency of reported respiratory symptoms in the first year of life was associated with NO2 levels not currently considered to be harmful” (van Strien et al., 2004). However, another study did not find an association between NO2 level and respiratory illnesses in infants (Samet et al., 1993). A study of asthmatic children in inner-city environments found that indoor NO2 levels were substantially elevated in homes with gas stoves and that “higher levels of indoor NO2 are associated with increased asthma symptoms in nonatopic children and decreased peak flows” (Kattan et al., 2007). Early life exposure to household gas appliances has also been associated with negative neuropsychological development (Morales et al., 2009). Valero et al. (2009) investigated the determinants of exposure for a cohort of Spanish women and found that personal NO2 levels were “strongly influenced by indoor NO2 concentrations.” They also found that outdoor NO2 levels and the use of gas appliances were important determinants of indoor NO2 levels, whereas no significant association “was found between personal or indoor NO2 levels and exposure to environmental tobacco smoke (ETS) at home.”
Cooking can also substantially increase indoor fine-particle mass concentrations (PM2.5) (Abt et al., 2000; Buonanno et al., 2009; Evans et al., 2008; Olson and Burke, 2006; Wallace et al., 2004). Fumes from Chinese-style cooking with hot oil have been shown to be mutagenic (Chiang et al., 1997), and this cooking style has also been reported to be a risk factor for lung cancer in nonsmoking women in Taiwan (Ko et al., 1997). Exposure to ultrafine particles can be substantially increased by emissions from cooking (Bhangar et al., 2011; Mullen et al., 2010). Emissions of ultrafine particles can be caused not only by the combustion of cooking fuel but from high temperatures associated with electric cooking elements (Wallace et al., 2008).
Climate change could affect the indoor concentrations of cooking-associated pollutants in the United States and other developed countries in several ways. First, it may be that a mitigation response to climate change drives a movement toward smaller per-capita housing space (with lower life-cycle environmental effects) and with lower air-exchange rates (to save heating and cooling energy). If so, emissions from cooking would be diluted into a smaller volume and would persist for longer times, and these changes would tend to increase concentrations and exposures associated with a given level of cooking. Second, climate-change mitigation goals might push cooking away from the use of natural gas and toward a heavier reliance on electricity (assuming that electricity would be generated from lower-carbon sources than today). Such a shift would reduce associated exposures to NO2 and to the ultrafine particles formed in combustion flames. Third, tighter building envelopes resulting from weatherization efforts might reduce the
efficacy of local exhaust hoods and fans for removing cooking-related emissions before they enter indoor air. Dampers have been developed that automatically open when exhaust fans are activated to permit additional ventilation supply air to flow freely into a building, thereby mitigating this otherwise adverse effect of weatherization.
In the United States, combustion for space heating can sometimes be associated with substantial pollutant emissions, especially because of the relatively large amounts of fuel used for home heating compared with, for example, cooking. When on-site combustion is used to generate heat, it is usually the case that the heat is first extracted from the combustion gases and then the byproducts are vented to the outside. Leakage may occur, and some of the generated pollutants can enter the occupied indoor space of the same building for which the heat is being generated. In addition, combustion for heating is sometimes unvented by design, in which case all the byproducts formed are emitted into the indoor environment with the generated heat. The direct evidence that links household heating with health effects is sparse. Household use of kerosene heaters and fireplaces for heating was found to be associated with respiratory symptoms in nonsmoking women in Connecticut and Virginia during the 1990s (Triche et al., 2005). A study of coroners’ reports in California found that unvented combustion heating appliances and cooking indoors with charcoal were associated with CO deaths (Liu et al., 2000).
Climate change could induce several shifts that would affect indoor air-pollutant exposures associated with heating. First, if average temperatures rise, as is expected, less heating may be needed, and—other things being equal—there would tend to be less associated pollution exposure. Climate-change mitigation efforts may lead to better insulation of buildings, which also would lessen heating requirements. Second, there could be shifts in the types of heating sources used. Mitigation efforts could serve as a driving force for substituting electricity (from low-carbon sources) for fossil-fuel combustion—a change that would tend to improve IAQ. In contrast, mitigation goals might also encourage greater use of wood as a household heating fuel. Wood contains contemporary rather than fossil carbon. If grown and harvested sustainably and if burned completely, wood combustion could have little or no net climate impact. However, as practiced today, residential wood combustion is associated with degraded neighborhood air quality owing to emissions exhausted from chimneys and is associated with degraded IAQ in the households that burn the wood owing to leakage of combustion byproducts into the indoor environment (Gustafson et al., 2008; Traynor et al., 1987). If done poorly, increased wood-based
heating could exacerbate IAQ problems associated with residential wood combustion.
Another trend that might emerge and that would tend to degrade IAQ is greater reliance on unvented combustion-based space heaters. Devices of this type have a high thermal efficiency because all the generated heat is discharged indoors. However, their use can cause substantially increased indoor concentrations of NO2, sulfur dioxide (SO2), and particulate matter (Francisco et al., 2010; Leaderer, 1982; Leaderer et al., 1990; Ruiz et al., 2010; Wallace and Ott, 2011).
An additional concern associated with climate change and home heating is building envelope tightness. Efforts to save energy by reducing the leakiness of building envelopes can increase the risk of “backdrafting,” in which air flows into a building through the exhaust flue, instead of flowing out of the building, and carries combustion byproducts with it. The causes and consequences of backdrafting have received some attention in the literature (Nagda et al., 1996), but the prevalence even in current conditions in the building stock has not been well characterized, and it is not clear what to expect in this regard as a consequence of climate change.
Habitual indoor smoking adversely affects IAQ and public health. Sidestream smoke (from the smoldering tobacco product) and exhaled mainstream smoke together constitute the source of environmental tobacco smoke (ETS). Smoking indoors has a strong influence on indoor levels of PM2.5 (Hyland et al., 2008; Nazaroff and Klepeis, 2004). ETS is also an important cause of environmental exposure to some hazardous air pollutants, including acrylonitrile, 1,3-butadiene, acetaldehyde, acrolein, and formaldehyde (Nazaroff and Singer, 2004). Evidence indicates that several severe adverse health effects are associated with ETS exposure, including acute myocardial infarction (Lightwood and Glantz, 2009), lung cancer (Fontham et al., 1994), and a host of respiratory health problems in children (DiFranza et al., 2004). Over the past few decades, there has been a marked reduction in exposure to ETS in the US population, as reflected in lower concentrations of serum cotinine in nonsmokers (Pirkle et al., 2006). The decline is a consequence mainly of declines in the amount of smoking that occurs indoors rather than of changes in the building stock.
In a future influenced by climate change, exposure of nonsmokers to ETS will be determined to a great degree by the prevalence and intensity of smoking in indoor spaces. In the United States, smoking in public places has become uncommon. However, smoking in private residences continues: Singh et al. (2010) estimated that 7.6% of children in the United States are exposed to ETS in their own homes. Exposures to ETS occur not only in the
residence in which smoking occurs but, in the case of multifamily dwellings, in neighboring units (Bohac et al., 2011). Some parts of the US population have a relatively high prevalence of indoor smoking. For example, a study of 100 asthmatic children in inner-city Baltimore revealed an indoor smoking prevalence of 46% and found that average indoor PM2.5 and PM10 levels were 33–54 µg/m3 higher in smoking than in nonsmoking households (Breysse et al., 2005). In another study, fine-particle concentrations were sampled over two-week periods in 294 inner-city homes with asthmatic children (Wallace et al., 2003). In these homes, the average particle mass concentration, 27.7 µg/m3, was considerably higher than the average concurrently measured outdoor concentration, 13.6 µg/m3. Smoking occurred in 101 of the homes (34%) and caused an average increase of 37 µg/m3 for indoor fine particle levels. Other identified sources—frying, smoky cooking events, and use of incense—made smaller contributions, 3–6 µg/m3.
It is unknown how smoking patterns that would affect indoor ETS will evolve. In particular, it is not clear that indoor smoking behaviors would be influenced by climate change. Changes in tobacco or in tobacco products could alter the ETS characteristics associated with indoor smoking, and there is some published evidence that tobacco itself might be altered in response to changing temperature and atmospheric CO2 levels (Ziska et al., 2005).
Changes in the residential building stock that are a consequence of climate-change concerns could influence exposure to ETS. Currently, unintended airflow pathways in multiunit residential buildings can lead to exposures to secondhand smoke in the units of nonsmokers (Kraev et al., 2009; Wilson et al., 2011; Winickoff et al., 2010). Mitigation measures to reduce energy use in buildings could lead to systematically lower ventilation rates and alteration of internal airflows that could cause higher concentrations and exposures to secondhand smoke. For a given characteristic, such as number of cigarettes smoked indoors per day, any of those changes would tend to increase exposures to ETS indoors.
Candles, Incense, and Other Small-Scale Combustion Processes
Pagels et al. (2009) summarize some of the IAQ concerns related to indoor candle use. The local high temperature created by a candle flame can volatilize candle components that are then emitted to indoor air. Some candles have metal-cored wicks that emit lead at rates sufficient to pose health concerns (Wasson et al., 2002). Depending on combustion conditions, the candle flame also produces soot particles and other products of incomplete combustion that are emitted indoors (Fine et al., 1999).
According to the National Candle Association (National Candle Association, 2011), US retail sales of candles are roughly $2 billion per year,
and “candles are used in 7 out of 10 US households.” Given the type and scale of emissions summarized in the previous paragraph, the potential for air-pollutant exposure due to candle use would seem to be substantial, but scientific data that would permit one to quantify the extent of indoor use and the resulting air-pollutant exposures are lacking.
In developing countries, combustion-based technologies, such as candles and kerosene lamps, are commonly used to provide lighting. Those are inherently inefficient in converting chemical energy into light (Mills, 2005). The air-pollutant exposure consequences of combustion-based lighting are expected to be substantial but have only begun to be explored (Apple et al., 2010).
Indoor air-pollutant emissions from other small-scale combustion sources have been investigated, and a few illustrative examples are noted here. Jetter et al. (2002) studied the emissions from burning incense and concluded that “incense smoke can pose a health risk to people due to inhalation exposure of particulate matter.” Liu et al. (2003) characterized emissions and IAQ effects of burning mosquito coils, which are commonly used in households in Asia, Africa, and South America. They concluded that “exposure to the smoke of mosquito coils similar to the tested ones can pose significant acute and chronic health risks.”
As in the case of other indoor combustion activities, climate change would affect IAQ and potentially public health if it were accompanied by a change in the source emission rate (for example, owing to a change in use) or were accompanied by a change in the other factors that influence exposures associated with a given magnitude of emissions. There is no good basis of expectations of use patterns of small-scale combustion sources. As noted in connection with other combustion sources, reduced household volume per occupant and lower air-exchange rates might be consequences of efforts to mitigate anthropogenic effects on climate, and such changes would tend to increase air-pollutant exposures that result from indoor combustion sources.
Radon and Its Decay Products
Indoor radon is a major cause of the public’s health-relevant radiation exposure. Exposure to increased residential radon is an important risk factor for lung cancer. On the basis of a combined analysis of 13 studies that collectively involved 7,148 lung-cancer cases and 14,208 controls, Darby et al. (2005) concluded that residential radon is “responsible for about 2% of all deaths from cancer in Europe.” In a parallel North American effort encompassing 7 studies that collectively assessed 3,662 cases and 4,966 controls, Krewski et al. (2005) reported that their results “provide direct evidence of an association between residential radon and lung cancer risk, a
finding predicted using miner data and consistent with results from animal and in vitro studies.”
Radon-222 (radon), the most health-significant of the three naturally occurring isotopes, is generated by the radioactive decay of radium-226, a ubiquitous trace element in the earth’s crust. Being an inert gas, radon has the potential to migrate from its parent material during its short lifetime (half-life, 3.8 days) and enter indoor or outdoor air, where humans may encounter it. Radon does not directly pose a substantial health hazard. However, its radioactive decay marks the beginning of a sequence of short-lived products. Those radon decay products—isotopes of bismuth, lead, and polonium—are chemically reactive and, when inhaled, can be retained on respiratory tract tissues; later radioactive decays irradiate lung cells. Of particular health concern are the alpha-particle emissions from the decay of polonium-218 and polonium-214. It is the radiation damage caused by those alpha-particle emissions that creates the lung-cancer risk associated with exposure to residential radon. The epidemiologic evidence is consistent with a linear no-threshold dose–response model. Health risks posed by a given level of radon exposure are much higher in smokers than in nonsmokers (Ginevan and Mills, 1986).
The three main sources of indoor radon are soil near a building’s foundation; earthen building materials, such as concrete; and tap water from underground sources. In aggregate for the entire building stock, soil is the most important radon source, although the other two sources dominate in some buildings. The significance of soil as a source of indoor radon depends on the radium content of the soil, on the permeability of the soil, and on the degree of coupling between the indoor space of the building and the pore air in the underlying and adjacent soil (Nazaroff, 1992). The only important mechanism for removing radon from indoor air is ventilation. However, the effective radiation dose to lung tissue associated with a given level of indoor radon depends on the dynamic behavior of the short-lived decay products (Porstendörfer, 1994), which can be influenced not only by the ventilation rate but by such factors as indoor particle levels, active air filtration, and the intensity of indoor air movement.
Annual average residential radon levels in the United States have been estimated to have an arithmetic mean of 46 ± 4 Bq/m3 (1.25 ± 0.12 pCi/L) with an estimated 6% of dwellings exceeding the EPA mitigation level of 148 Bq/m3 (4 pCi/L) (Marcinowski et al., 1994). EPA has estimated that 20,000 US lung-cancer deaths a year are radon-related (Pawel and Puskin, 2004). Radon-control systems are well established in principle for maintaining low indoor radon concentrations (Rahman and Tracy, 2009). However, challenges remain to identify buildings with high concentrations and to apply effective controls, where appropriate, in both existing and new buildings.
Climate change might induce shifts in indoor radon and decay-product concentrations for several reasons, although the direction and scale of the changes are difficult to predict. Changes that would reduce ventilation rates would tend to increase indoor radon levels and might also alter the effective radiation dose received. Constructing buildings with or near materials that have high radium content should be avoided irrespective of climate-change concerns. The goal of improving the energy performance of buildings might induce increased use of subterranean spaces for habitation or stronger thermal coupling of building interiors to climate-buffered underground zones. Care would be needed in such cases to prevent radon levels from increasing in the occupied spaces.
Volatile Organic Compounds and Semivolatile Organic Compounds
Organic compounds constitute a diverse set of chemicals that have a broad array of properties. For the discussion here, organic compounds are divided into two primary groups. Volatile organic compounds (VOCs) are species that have high enough vapor pressures to volatilize substantially and, when unconfined, to be found predominantly in the gas phase. Semivolatile organic compounds (SVOCs) are preferentially found in the condensed phase, but they still have sufficient volatility to be present in the vapor-phase. For SVOCs, the saturation vapor pressure is roughly in the range 10-9 to 10 Pa; VOCs have vapor pressures higher than 10 Pa. Organic compounds that have extremely low volatility can also be present purely in the condensed phase and could still contribute to IAQ concerns as constituents of particulate matter. An important example of this category would be polycyclic aromatic hydrocarbons (PAHs) that have many rings.
Excessive exposures to volatile and semivolatile organic compounds indoors raise a broad range of public health concerns. For example, many organic chemicals that have been classified by the US federal government as hazardous air pollutants (HAPs) are present at significant levels indoors. This classification applies to pollutants that “cause or may cause cancer or other serious health effects, such as reproductive effects or birth defects, or adverse environmental or ecological effects (EPA, 2009b). Some of these chemicals may be used as constituents of construction and finishing materials, or as ingredients in consumer products used indoors. Seminal research from the 1980s known as the Total Exposure Assessment Methodology (TEAM) studies showed that concentrations in personal air (heavily influenced by indoor conditions, because people spend most of their time indoors) commonly exceeded outdoor levels for several toxic air pollutants, including benzene, carbon tetrachloride, chloroform, dichlorobenzene, tetrachloroethylene, 1,1,1-trichloroethane, trichloroethylene, and xylenes (Wallace et al., 1985, 1987, 1988).
In addition to concerns related to their status as toxic air pollutants, exposure to VOCs has also been associated with such health effects as allergic symptoms, asthma, and symptoms of sick-building syndrome (Garrett et al., 1999; Norbäck et al., 2000; Smedje et al., 1997; Ten Brinke et al., 1998). Establishing definitive links between exposure to those compounds and these types of health effects is challenging because the amount of exposure sustained by study subjects and the conditions under which they are exposed generally are beyond the direct control of the investigator. In addition, human populations are routinely exposed to multiple contaminants whose individual, let alone joint, effects are not known (Cohen and Gordis, 1993).
Research on indoor VOCs began in the 1980s. Research on indoor SVOCs is much less developed. However, many studies of subclasses of SVOCs indoors—including pesticides, polybrominated diphenyl ethers (flame retardants), and phthalates (plasticizers)—have been published. Several review articles have been published on the occurrence and potential health significance of VOCs and SVOCs indoors (Brown et al., 1994; Jones, 1999; Logue et al., 2011; Mendell, 2007; Rudel and Perovich, 2009; Salthammer and Bahadir, 2009; Weschler and Nazaroff, 2008; Wolkoff, 1995; Wolkoff et al., 1997). Logue et al. (2011) compared published concentration data on VOCs in US houses with health-based exposure guidelines and standards. They identified seven organic compounds as “priority hazards based on the robustness of measured concentration data and the fraction of residences that appear to be impacted.” In alphabetical order, the seven are acetaldehyde, acrolein, benzene, 1,3-butadiene, 1,4-dichlorobenzene, formaldehyde, and naphthalene. An important attribute of SVOCs is their potentially long persistence indoors. Available evidence suggests that these may pose legacy pollution concerns (Weschler and Nazaroff, 2008).
Focusing on sensory irritation and other perceived IAQ effects, Wolkoff et al. (1997) have called attention to the importance of secondary pollutant formation indoors due to reactions involving organic compounds and oxidizing agents, such as ozone (O3) and NO2. Mendell (2007) reviewed 21 studies from the “epidemiologic literature on associations between indoor residential chemical emissions, or emission-related materials or activities, and respiratory health or allergy in infants or children.” He found that the most frequently identified risk factors included “formaldehyde or particleboard, phthalates or plastic materials, and recent painting.”
Emissions of VOCs indoors tend to be higher after new construction and renovation activities because of releases from finite-capacity reservoirs in wood-based products, paints, floor finishes, glues, and other construction and finishing materials (Dales et al., 2008; Herbarth and Matysik, 2010). House dust is an important repository for SVOCs and other particle-
bound contaminants (Butte and Heinzow, 2002). Results of studies of house dust have demonstrated the presence of polychlorinated biphenyls (PCBs), PAHs, plasticizers (phthalates and phenols), flame retardants, other organic xenobiotics, and inorganic constituents (Weschler and Nazaroff, 2010). Dust ingestion can also be an exposure pathway of concern for SVOCs (Roosens et al., 2009). Infants are generally affected more by dust ingestion than adults because of their contact with floors and their high level of hand-to-mouth activity.
In subsections that follow, indoor exposure conditions and associated health concerns are summarized for two important examples in the broader category of VOCs and SVOCs: formaldehyde and endocrine-disrupting chemicals. Overall, however, the state of knowledge about VOCs and SVOCs in indoor environments and their consequences for public health is far from complete. The chemicals are diverse in their characteristics and complex in their dynamic behavior. Conditions vary with time in any given building and can also vary markedly among buildings. Concentrations are influenced by a variety of factors, some of which reflect properties of the chemicals, some of which depend on properties of the buildings into which they are emitted, and some of which depend on actions of building occupants, for example, in relation to frequency and intensity of use of products that contain the chemical of concern. Occupant behaviors also affect how concentrations are related to exposures.
Increasingly, biomarkers are being used to measure body burdens of environmental chemicals or chemical byproducts in human tissue (Paustenbach and Galbraith, 2006; Sexton et al., 2006). Some recent work has focused on prenatal life and infancy as highly vulnerable periods of development. Investigations have monitored indoor exposures to multiple chemicals and birth outcomes (Eskenazi et al., 1999; Herbstman et al., 2010; Perera et al., 2003; Rosas and Eskenazi, 2008). However, there is still an inadequate understanding of the relationship between environmental concentrations of VOCs and SVOCs, resulting biomarker levels, and associated health outcomes.
Available information suggests that concerns about the influence of climate change on exposures and public-health risks associated with VOCs and SVOCs are substantial enough to warrant further attention, but it is insufficient to support substantive conclusions. In his review of indoor pollutants over the past 50 years, Weschler (2009) made an important point in this regard, stating that
Many of the chemicals presently found in indoor environments, as well as in the blood and urine of occupants, were not present 50 years ago. Given the public’s exposure to such species, there would be exceptional value in monitoring networks that provided cross-sectional and longitudinal information regarding pollutants found in representative buildings.
Formaldehyde is a chemical of concern as a carcinogen and as an airway irritant. The International Agency for Research on Cancer has classified formaldehyde as “carcinogenic to humans” (IARC, 2006). Formaldehyde is listed by EPA as a hazardous air pollutant (EPA, 2008) and as a toxic air contaminant in the state of California (OEHHA, 2007). Indoor formaldehyde levels can be elevated because of emissions from indoor sources, such as wood-based products, paints, O3-initiated chemical reactions, and combustion (Salthammer et al., 2010). Acute effects, such as eye, nose, and throat irritation have been observed after controlled exposures at levels of at least 1,230 µg/m3 (Kulle, 1993); however, persons who have asthma or allergic sensitization have been shown to respond with bronchial symptoms to exposure at concentrations as low as 100 µg/m3 (Casset et al., 2006). An Institute of Medicine report on IAQ and asthma concluded that there was limited or suggestive evidence of an association between formaldehyde exposure and wheezing and other respiratory symptoms (IOM, 2000). It also concluded that there was inadequate or insufficient evidence to determine whether an association between formaldehyde exposure and asthma exists. At moderate to high exposure, formaldehyde can be irritating, producing such symptoms as sore throat, cough, scratchy eyes, and nosebleed. Some people are more sensitive than others, and an exposure that causes no problems for some people can make other people sick or uncomfortable.
There is no US federal regulation or standard for formaldehyde levels in residential settings. In 2008, however, California established a regulation limiting the emissions of formaldehyde from composite wood products3 that are common indoor sources of formaldehyde (California Environmental Protection Agency, 2011). Also, the California Air Resources Board has published a guideline value of 27 ppb (33 µg/m3) for formaldehyde in residences to avoid irritant effects (Cal/EPA ARB, 2004).
Residential formaldehyde levels are influenced by ventilation rates. In a sample of 122 new homes in California, formaldehyde levels were inversely correlated to the air-exchange rate (Offermann, 2009). Based on 24-h average samples in 105 of these homes, more than 90% exceeded California’s chronic reference exposure level of 9 µg/m3 (OEHHA, 2007) and 59% exceeded the 33 µg/m3 indoor guideline value. A study of 96 homes in Québec City found that higher concentrations of formaldehyde were associated with lower air-exchange rates (Gilbert et al., 2006). A monitoring program that sampled formaldehyde in 252 homes in Germany also found that indoor levels were inversely correlated with air-exchange rates (Salthammer
3 California Code of Regulations §93120. Airborne Toxic Control Measure to Reduce Formaldehyde Emissions from Composite Wood Products.
et al., 1995). That study also found that indoor formaldehyde levels were positively, albeit weakly correlated with higher indoor temperatures and relative humidities. Higher indoor temperatures and relative humidities might be expected in some indoor environments as a consequence of climate change.
In response to the first modern energy crisis in the 1970s, urea-formaldehyde foam insulation (UFFI) was widely used as a retrofit building insulation material. Emissions of formaldehyde, perhaps exacerbated because of improper or inappropriate use in some cases, led to a concern among Canadian authorities, where the use of UFFI was banned in 1980 (CMHC, 2011). In 1982, UFFI was banned by the US Consumer Products Safety Commission (CPSC, 1982); however, that ban was overturned by a legal ruling in 1983 (CPSC, 1983). Research conducted on homes with UFFI in Toronto only showed a moderate increase in formaldehyde levels (median = 38 ppb, N = 571) compared with untreated homes (median = 31 ppb; N = 231) (Broder et al., 1988). Remedial intervention to remove UFFI from some homes showed post-removal improvement in many health status indicators, although the “improvement in health status among the UFFI removal subset was not associated with any significant diminution of formaldehyde exposures” (Broder et al., 1991). L’Abbé and Hoey (1984) reviewed the evidence available at that time and concluded that “epidemiologic studies have not established causation or an association between UFFI exposure and health effects.”
After Hurricane Katrina (in 2005), hundreds of temporary housing trailers provided by the Federal Emergency Management Agency were found to have elevated levels of formaldehyde (Maddalena et al., 2009). After learning of the potential concern regarding the air quality in the trailers used for temporary housing after hurricanes, the Centers for Disease Control and Prevention stated that residents living in temporary trailer housing should open windows as much as possible to let in fresh air, keep indoor temperatures at the lowest comfortable setting, run an air conditioner or dehumidifier to control mold, and spend as much time outdoors as possible. Children and elderly people and those with chronic diseases, such as asthma, were particularly encouraged to spend time outside. However, in the event of unhealthful outdoor temperatures and high outdoor levels of pollutants, vulnerable populations may be forced to choose between unhealthful indoor exposures to formaldehyde and other pollutants from indoor sources and exposures to heat and outdoor air pollution. Persons in buildings that lack adequate ventilation and air conditioning would be particularly vulnerable.
Many of the chemicals classified as SVOCs and used in products found indoors have demonstrated or suspected endocrine-disrupting properties (Rudel and Perovich, 2009). Human exposure to indoor air and to dust enriched with endocrine-disrupting chemicals released from indoor sources has become an issue of increasing concern (Hwang et al., 2008; Rudel et al., 2003, 2010). Semivolatile, endocrine-disrupting chemicals of concern include brominated organics, such as polybrominated diphenyl ethers (PBDEs), used as flame retardants; phthalates, used as plasticizers; PCBs, historically used in caulks and many other products; and pesticides, including organochlorines (such as DDT), organophosphates (such as chlorpyrifos), and pyrethroids (such as permethrin). The following paragraphs provide brief summaries of the concerns for the first three of these classes; pesticides are discussed in Chapter 6.
Epidemiologic studies have demonstrated some associations between indoor levels of SVOCs and adverse health effects, such as childhood leukemia, neurologic disorders, non-Hodgkin lymphoma, and respiratory symptoms (Bornehag et al., 2004; Butte, 2004; Colt et al., 2006). Studies have also produced suggestive evidence that prenatal exposure to those substances may have a deleterious effect on neurodevelopment (Chevrier et al., 2008; Eskenazi et al., 2006; Jacobson and Jacobson, 1996; Jacobson et al., 1990; Rogan and Gladen, 1991; Rogan et al., 1986).
PBDEs, a major class of flame retardants, are ubiquitous environmental contaminants with particularly high concentrations in humans in the United States (Fischer et al., 2006). For the purpose of retarding ignition, these chemicals are added to consumer electronic cases and other materials used indoors. The compounds are slowly released from the products during their life cycles (Alcock et al., 2003). Biomonitoring has shown children’s levels to be 2–5 times higher than those of their parents, perhaps because of children’s greater exposure to and ingestion of house dust into which PBDEs preferentially partition (Fischer et al., 2006). The potential health hazards of PBDEs are attracting increasing scrutiny. They have been shown to reduce fertility in humans at levels found in households (Harley et al., 2010). Children who have higher concentrations of PBDEs in their umbilical-cord blood at birth have been found to score lower on neurodevelopment tests at the ages of 1 and 6 years (Herbstman et al., 2010). Although the pathways through which PBDEs get into people are not fully understood, releases from indoor construction and furnishing materials, aging and wear of consumer products, and direct exposure during use (for example, from furniture) are potentially important contributors.
Phthalates are one of the more abundant contaminants in household dust. At one time, they were considered safe, but studies have revealed
that they may pose endocrine-disruption risks and exhibit reproductive and developmental toxicity (ATSDR, 2002; Duty et al., 2003; Hauser and Calafat, 2005). EPA announced that a review of the safety of these chemicals would begin in fall 2010—an indication of concern about phthalates based on toxicity (particularly in the development of the male reproductive system), prevalence in the environment, widespread use, and resulting human exposure (EPA, 2010a). Adverse effects on the development of the male reproductive system may be the most sensitive health outcomes of phthalate exposure according to studies of laboratory animals. Several studies have shown associations between phthalate exposures and human health, although no causal link in humans has been established. Biomonitoring data from 1999–2000 and 2001–2002 demonstrated that children have the highest exposures to phthalates of all groups monitored; other biomonitoring data have shown in utero exposures to phthalates (CDC, 2005). Indoor environmental exposures may be important contributors to total uptake (Wormuth et al., 2006).
Although the use of PCBs has been banned or restricted for decades, they are still being found indoors in older buildings at levels that are considered to be of concern for human health. Potential health risks posed by PCBs remain high in some indoor environments because of weak removal processes and long-term release from sources (Herrick et al., 2004; Rudel et al., 2008).
Indoor exposures to CO2 are likely to increase as a consequence of climate change. The atmospheric background concentration of CO2 is rising. The preindustrial level was approximately 280 ppm; the level is about 390 ppm now and continues to rise by a few parts per million per year (Keeling, 2009; NRC, 2010). There is also evidence of a rural–urban gradient—levels of CO2 in outdoor air are higher in urban than in rural environments (George et al., 2007)—and the percentage of people living in urbanized areas is increasing. In the United States in 2000, 58% of the population lived in an urbanized area with a population above 200,000 (US Department of Transportation, 2004).
CO2 levels are substantially higher in occupied buildings than outdoors. Unvented combustion sources can contribute, but the main indoor source of CO2 is the exhaled breath of building occupants. The metabolic production of CO2 by humans depends on diet and activity level. A sedentary adult typically generates CO2 at 0.31 L/min (ASHRAE, 2010), corresponding to 34 g/h (at an atmospheric pressure of 1 atm and a temperature of 293 K).
CO2 levels are commonly used to guide ventilation practice in occupied buildings. In that case, CO2 is serving as a marker of human bioeffluents.
Research shows that “maintaining a steady-state CO2 concentration in a space no greater than about 700 ppm above outdoor air levels will indicate that a substantial majority of visitors entering a space will be satisfied with respect to human bioeffluents (body odor)” (ASHRAE, 2010).
In common practice, building ventilation requires energy use. In mechanically ventilated buildings, fan power is required to move air through ducts. When the temperature or humidity of the outdoor air is not suitable for establishing desired indoor conditions, energy is used to condition the ventilation air. The desire to mitigate effects of climate change is creating pressure to reduce ventilation rates in buildings and the use of energy. Hence, one might reasonably expect indoor CO2 to rise in a climatechange–influenced future for two reasons: increased baseline levels due to rising outdoor CO2 levels, especially in cities; and reduced ventilation rates in buildings as part of a mitigation strategy.
CO2 is an acid gas. At high levels, “inhalation of CO2 can produce physiological effects on the central nervous, respiratory, and the cardiovascular systems” (US Department of Labor, 1990). Recognizing its potential for frank adverse health effects, the Occupational Safety and Health Administration maintains an occupational standard for CO2, with a “transitional limit” of 5,000 ppm for the 8-h time-weighted average (US Department of Labor, 1989).
Occupational standards for pollutants are typically set at much higher levels than would be appropriate for the general public. In the United States, there are no health-based guidelines or standards for CO2 itself that would apply for the general public in all indoor environments. In Germany, a governmental work group recommended that “based on health and hygiene considerations: concentrations of indoor air carbon dioxide levels below 1000 ppm are regarded as harmless, those between 1000 and 2000 ppm as elevated and those above 2000 ppm as unacceptable” (Ad-hoc Work Group, 2008; translated in Heinzow and Sagunski, 2009).
Existing literature does not provide a clear answer to the question of whether public-health consequences are associated with exposure to CO2 indoors at the levels at which they might occur in a future influenced by climate change, but it does contain some important clues. A study by Bekö et al. (2010) measured CO2 in the bedrooms of 500 Danish children to characterize ventilation conditions during sleep. They found that 6% of the rooms had levels over 3,000 ppm (20-min running mean) at some time during the night, which is well above that viewed as “unacceptable” by the German indoor-air working group cited above. Shendell et al. (2004) studied the association between student absenteeism and classroom CO2 in Washington and Idaho. They found that 45% of classrooms studied “had short-term indoor CO2 concentrations above 1000 ppm.” They also found a statistically significant association between higher indoor–outdoor
differences in CO2 and student absenteeism. Haverinen-Shaughnessy et al. (2011) measured CO2 in 100 classrooms, inferred that 87 had substandard ventilation rates, and found a positive association between the inferred ventilation rates (0.9–7.1 L/s per person) and student performance on standardized tests. Seppänen et al. (1999) reviewed the literature on ventilation rates, CO2 concentrations, and sick-building syndrome symptoms. They reported that “about half of the carbon dioxide studies suggest that the risk of sick building syndrome symptoms continues to decrease with decreasing [indoor] carbon dioxide concentrations below 800 ppm.” Evaluating data from a study conducted by EPA of 100 US office buildings, Erdmann and Apte (2004) found “statistically significant, dose-dependent associations (P < 0.05) for combined mucous membrane, dry eyes, sore throat, nose/sinus congestion, sneeze, and wheeze symptoms” with the difference between indoor and outdoor CO2 levels.
It is important to note that those associations do not demonstrate that CO2 itself is harmful to public health at the levels ordinarily encountered indoors. It may be that the adverse effects reported result from some other contaminant whose concentrations correlate with those of indoor CO2. Alternatively, because CO2 is not only a product of metabolism but a biologic trigger to induce breathing, it is conceivable that levels of CO2 encountered indoors have direct health consequences. Studies of health hazards of CO2 exposure have tended to stress conditions in healthy young adults, such as submariners (Margel et al., 2003), astronauts (Manzey and Lorenz, 1998), and motorcycle riders (Bruhwiler et al., 2005). Studies are lacking of the potential health consequences of chronic or episodic exposures to increased CO2 at levels below 5,000 ppm in the young, the elderly, and the infirm.
Air pollutants of outdoor origin enter buildings with ventilation air. Depending on the pollutant and the building conditions, the indoor proportion of the outdoor pollutant level ranges from zero (perfect sheltering) to 100% (no benefit from being indoors). A building provides virtually no protection against CO2 from the outdoors, for example, but does provide some protection against PM and O3. Filters in the mechanical ventilation system of typical commercial buildings actively remove some portion of the particles in the air that passes through them (Hanley et al., 1994). Particles can also deposit onto indoor surfaces passively, and this phenomenon reduces the indoor proportion of outdoor particles (Riley et al., 2002). Ozone can be removed from ventilation air by using activated carbon (Bekö et al., 2009; Shair, 1981), but the use of activated carbon in building mechanical systems is not common. On the other hand, O3 reacts rapidly with indoor surfaces and with selected chemicals in indoor air (most notably nitric
oxide [NO] and terpenes). Consequently, the indoor level of O3 attributable to its presence outdoors is reduced, commonly to 20–70% of the outdoor level (Weschler, 2000). However, O3 reactions indoors also generate byproducts—including formaldehyde, acrolein, and ultrafine particles—that have potential adverse effects (Weschler, 2006). Those two air pollutants, PM and O3, currently receive the most attention in outdoor air-pollution control policy, and urban environments are furthest from compliance with air-quality standards for them. The relationship of indoor to outdoor concentrations for PM and O3 have complex characteristics but also have been fairly well studied.
Epidemiologic studies indicate that ambient concentrations of PM and O3 are associated with substantial adverse health effects. It is thus important to consider what could happen to those pollutants in indoor air in a climate-change regime. In addition to PM and O3, which are addressed in detail below, other ambient pollutants are worth discussing, at least briefly; several are summarized at the end of this section.
Particulate Matter of Outdoor Origin
Airborne particles are a complex pollutant class, with source attributes, atmospheric dynamics, and health consequences that vary with size and chemical composition. EPA has established and maintains health-based national ambient-air quality standards (NAAQSs) for PM in outdoor air. The standards are based on 24-h average and annual average mass concentrations of particles finer than 10 µm in diameter (PM10) and finer than 2.5 µm in diameter (PM2.5). Particles 2.5–10 µm in diameter are referred to as coarse, and particles smaller than 2.5 µm in diameter are termed fine. The NAAQSs for PM do not consider the chemical composition of the particles. Emerging evidence suggests that inhalation exposure to ultrafine particles, those smaller than 0.1 µm in diameter, also poses health risks in a manner that would not be well captured by the existing NAAQSs (Sioutas et al., 2005). In addition to the ambient overall PM mass-concentration standards, some chemical components (such as lead) that would be found primarily in the particle phase are regulated separately, either under the NAAQS or as hazardous air pollutants (EPA, 2010c).
PM in outdoor air is strongly associated with adverse health outcomes. After a comprehensive review, Pope and Dockery (2006) concluded that “the literature provides compelling evidence that continued reductions in exposure to combustion-related fine particulate air pollution as indicated by PM2.5 will result in improvements in cardiopulmonary health.” In a 2009 study, Pope et al. concluded that “a decrease of 10 µg m–3 in the concentration of fine particulate matter was associated with an estimated increase in mean (±SE) life expectancy of 0.61 ± 0.20 year (P = 0.004).”
Coarse particles (Brunekreef and Forsberg, 2005) and ultrafine particles (Oberdörster, 2001) have also been associated with adverse health effects that can be different from those of the fine-particle fraction.
Atmospheric particles may be classified as primary or secondary. Primary PM is emitted directly in the particle phase from sources. Secondary particles are formed in the atmosphere from the conversion of gaseous precursors to condensed-phase species. Coarse particles are mainly of primary origin and tend to be mechanically generated, for example, by abrasion. Soil dust, sea salt, and fragments of tires, roadways, and vehicle brakes are primary particles that are found mainly in the coarse mode. Coarse particles are commonly removed fairly rapidly (in minutes to hours) from the atmosphere by a combination of gravitational settling and inertial impaction on the earth’s surface. Because of their short atmospheric lifetime, concentrations of coarse particles can be spatially heterogeneous, with elevated concentrations found near emission sources.
The fine mode is a mixture of primary and secondary particles. Much of the primary material results from combustion processes and consists of noncombustible impurities, such as trace metals in coal, or products of incomplete combustion, such as soot (which is largely elemental carbon). Most of the atmospheric fine-particle mass is associated with sizes greater than 0.1 µm. Important secondary contributions to fine particles are associated with emissions of gaseous ammonia (NH3), nitrogen oxides (mainly NO), and SO2. Atmospheric oxidation processes convert the nitrogen oxides (NOx) and SO2 to nitric acid and sulfuric acid, which can then combine with NH3 to form salts, such as ammonium nitrate or ammonium sulfate. These salts condense onto pre-existing particles to contribute to the fine-particle mass concentration.
Another important source of secondary PM derives from the emission of VOCs and SVOCs. Atmospheric oxidation processes tend to increase the polarity and reduce the vapor pressure of those species, causing their partitioning to shift from the gas phase to the condensed phase. Secondary particle formation occurs on a regional scale, in part because the relatively long time (hours to days) required for the atmospheric transformation processes allows substantial transport and dispersal from local and urban sources.
The particle size range 0.1–2 µm in diameter is also known as the accumulation mode because of the relatively long atmospheric persistence associated with these particles (Nazaroff, 2004). They are too big to diffuse and too small to settle rapidly, so they persist for many days in the atmosphere.4 The combined importance of secondary formation and the slow
4 Incorporation into cloud drops that precipitate is a major atmospheric removal mechanism that provides a typical atmospheric lifetime of 1–2 weeks for accumulation-mode particles.
atmospheric removal processes mean that fine-particle mass concentrations exhibit a higher degree of spatial homogeneity than do coarse particles.
Atmospheric ultrafine particles have important primary sources, mainly tailpipe emissions from internal-combustion engines (Kittelson, 1998). They are also formed through secondary nucleation events in the atmosphere (Kulmala et al., 2004). Ultrafine particles have relatively short atmospheric lifetimes. Primary ultrafine particles exhibit high spatial heterogeneity with very high concentrations on and near heavily traveled roadways. In contrast, secondary ultrafine particles are formed on a regional scale and so exhibit more spatial homogeneity. However, the secondary formation events occur as bursts, so the temporal variability associated with secondary ultrafine particles can be high.
The degree to which particles of outdoor origin are present indoors depends on three main factors: particle size, building ventilation rate, and the presence and degree of effectiveness of any filters used for removing particles from an HVAC system or from recirculated air. Particles in the accumulation mode have the greatest ability to penetrate and persist indoors (Bennett and Koutrakis, 2006; Nazaroff, 2004; Riley et al., 2002). Coarse particles have a more difficult time in penetrating infiltration cracks in the building envelope (Liu and Nazaroff, 2001) or penetrating fibrous filters in ventilation systems (Hanley et al., 1994). Coarse particles also deposit more rapidly onto indoor surfaces (Thatcher et al., 2002). Similarly, ultrafine particles penetrate infiltration cracks less effectively, are filtered more easily, and deposit onto indoor surfaces more rapidly (Lai and Nazaroff, 2000) than do accumulation-mode particles. Using a material balance model with empirical data on governing factors, Riley et al. (2002) estimated that an urban residence with a typical ventilation configuration would have indoor proportions of outdoor particles of about 0.45 for particle-number concentration (mainly ultrafine particles), about 0.8 for PM2.5 mass concentration, but only about 0.2 for coarse-particle mass concentration. Those results also illustrate that buildings typically provide occupants some protection, but not extensive, from exposure to particles in outdoor air.
Absent active filtration, higher ventilation rates tend to produce higher indoor concentrations of outdoor particles. The reason is that ventilation serves as the sole source introducing outdoor particles into indoor air but as only one of several removal mechanisms. Higher ventilation rates increase the source term proportionally, but removal rates less than proportionally. Furthermore, higher rates of ventilation provided by open doors and windows (natural ventilation) tend to allow penetration with little attenuation.
In mechanically ventilated buildings, filters are commonly used to treat the supply air. Their efficiency can vary widely, as classified by the “minimum efficiency reporting value” (ASHRAE, 1999). The effectiveness of
filtration in providing protection indoors against particles of outdoor origin depends both on filtration efficiency and on the airflow configuration.
In addition to those considerations, the chemical composition of particles can influence the penetration and persistence of outdoor particles in indoor environments. For example, Lunden et al. (2003) have shown that, under wintertime conditions, aerosol ammonium nitrate levels can be much lower indoors than outdoors. With warmer temperatures indoors, ammonium nitrate has an enhanced tendency to dissociate to its constituent gases, NH3 and HNO3, and the HNO3 is then rapidly scavenged by the chemically basic gypsum wallboard commonly found indoors.
Regarding vulnerable populations, the findings of Hystad et al. (2009) should be noted. They found that “residences with low [economic] building values had higher infiltration efficiencies than other residences, which could lead to greater exposure gradients between low and high socioeconomic status individuals than previously identified using only ambient PM2.5 concentrations.” Results from McCormack et al. (2008) are also noteworthy in this regard. They reported time-integrated measurements of particle levels (PM10 and PM2.5) for three-day periods in the homes of 300 children (ages 2–6) in Baltimore’s inner city. The children were primarily African-American and from lower socioeconomic conditions. Smoking prevalence in the homes was 56%. Average indoor PM2.5 and PM10 levels were higher indoors (39.5 and 56.2 µg/m3, respectively) than the simultaneously measured outdoor levels (15.6 and 21.8 µg/m3, respectively). Evidently because of the importance of indoor sources, open windows were associated with significantly lower indoor PM levels.
The question of what might be expected with respect to indoor particles of outdoor origin in a future in which climate change occurs is best addressed in two parts: What is expected to happen to outdoor particles? How might the indoor proportion of outdoor particles shift because of changes in building design and operation?
Regulated Particulate Matter
With respect to the first question, it is useful to consider the possibilities sorted into several categories of outdoor particles. Particles in outdoor air are subject to air-pollution control regulations. Given the strong regulatory, public-policy, and technology momentum and given that many areas in the United States are out of compliance with existing NAAQSs for PM, one might expect some overall improvement over the coming decades with regard to ambient particle levels, at least for the PM10 and PM2.5 particle mass concentrations for which the regulatory machinery is the strongest.
- Sulfate from coal-fired power plants. Important contributions to improved ambient particle levels could be achieved by reducing SO2 emissions from coal-fired power plants. A transformation in the direction of lower sulfur emissions from coal combustion could be accelerated because of climate-change concerns, in that coal-fired electricity in the United States accounts for a large proportion (5–10%) of global anthropogenic emissions of fossil carbon to the atmosphere.
- Tailpipe emissions from motor vehicles. A shift away from the use of petroleum as a transportation fuel would also have important benefits for reducing ambient particle concentrations. Because of the proximity of urban roadways to buildings, tailpipe emissions from vehicles have a higher effectiveness in causing indoor-air pollutant exposure per unit mass emitted than do central-station power plants, which emit their pollutants from tall stacks, often on the edge of or remote from populous regions. As with coal-fired electricity, an effective response to climate change in the transportation sector might yield cobenefits in reducing indoor exposure to PM. For example, a shift from vehicles powered by internal-combustion engines to plug-in hybrid vehicles, to electric vehicles, or to fuel-cell–powered vehicles could lead to a substantial net reduction in outdoor particle levels near buildings and consequent improvements in IAQ.
- Distributed electricity generation. A trend may emerge toward more distributed generation of electricity in the form of combustion close to the point of use (Pepermans et al., 2005). There are potential efficiency benefits if electricity generation is combined with use of waste heat (for example, to heat water in buildings). A potential disadvantage is that such technologies risk being more poorly controlled than central-station power plants, so emissions per unit of useful energy output may be higher. Furthermore, and more importantly, the efficiency in causing exposure may grow markedly higher in moving from central-station to distributed power generation because emissions and people will be in closer proximity in the case of distributed generation (Heath and Nazaroff, 2007; Heath et al., 2006).
- Residential wood combustion. Climate-change concerns might lead to increased use of wood combustion and the burning of other contemporary carbon fuels for home heating. Most US home heating is accomplished by burning fossil fuels either directly (e.g., natural gas or fuel oil at the home site) or indirectly (e.g., by use of electricity that is generated from burning coal or natural gas). Mitigation strategies to reduce greenhouse gas emissions will aim to reduce
societal reliance on fossil fuels. Among the potential strategies could be increased use of renewable fuels such as wood and other biomass sources as a substitute for fossil fuels. Residential wood smoke is an important contributor to ambient particle levels in the winter in many communities (McDonald et al., 2000; Naeher et al., 2007). It is also possible that wood-combustion technologies could be improved to the point where excessive emissions are limited or avoided (Olsson and Kjallstrand, 2006; Ward et al., 2010). And, if improved wood-combustion technologies are not sufficiently effective, it may be that community concerns about the adverse health risks from wood smoke exposure would constrain any increase in use.
- Wildfires. Climate change is expected to increase the frequency of wildfires. Higher ambient temperatures combined with episodes of drought could lead to periods with a higher tendency for forests to burn. Park et al. (2007) have evaluated the importance of burning of biomass of all types as a source of fine PM in the US ambient atmosphere. They have estimated that wildfires, other fires, and residential and industrial biofuel use currently account for 20% (eastern United States) to 30% (western United States) of total observed fine-particle concentrations in outdoor air. Furthermore, they have estimated that annual carbon emissions from open fires were about twice as high (0.7–0.9 Tg per year) as those from biofuel use (0.4 Tg per year). They concluded that biomass burning is “an important contributor to US air quality degradation, which is likely to grow in the future.” Spracklen et al. (2009) have estimated that the annual mean area burned in the western United States will be about 50% larger in 2050 than in 2000, owing to climate-change effects; they also predict increases in summertime organic carbon and elemental carbon aerosol concentrations over the western United States of 40% and 20%, respectively, with most of the change attributable to increased wildfire emissions. Because wood-smoke particles are primarily in the fine mode, ordinary indoor environments, especially residences, do not provide much protection from them. However, Barn et al. (2008) have shown in an experimental study that using a recirculating, high-efficiency filter indoors can provide some protection against exposure to wood smoke associated with forest fires.
- Windblown dust. Another projected effect of climate change is increased frequency of drought in semi-arid regions. If water resources become further strained, changes in water allocations could increase the dryness of land surfaces, for example owing to reduced irrigation of crops and declining reservoir or lake levels. Conditions
such as these would have a tendency to increase the emissions of windblown dust into the atmosphere. Results of several studies illustrate the nature of the concern. Chan et al. (2008) reported that Asian dust storms were associated with an increased frequency of emergency-room visits for ischemic heart disease, cerebrovascular disease, and chronic obstructive pulmonary disease (COPD). Kuo and Shen (2010) showed that PM2.5 and PM10 levels in an office building increased during a dust storm. Hefflin et al. (1994) reported very high PM10 levels (more than 1,000 µg m-3) during seasonal dust storms in southeastern Washington state. However, on the basis of daily emergency-room visits, they concluded that, “the naturally occurring PM10 in this setting has a small effect on the respiratory health of the population in general.” In contrast, Ostro et al. (2000) studied daily mortality in relation to particulate air pollution in the Coachella Valley, California, where “coarse particles of geologic origin are highly correlated with and comprise approximately 60% of PM10, increasing to >90% during wind events.” Their results demonstrated “associations between several measures of particulate matter and daily mortality in an environment in which particulate concentrations are dominated by the coarse fraction.” Malig and Ostro (2009) assessed mortality statistics in 15 California counties for 1999–2005 in relation to coarse-particle monitoring data and found “evidence of an association between acute exposure to coarse particles and mortality” and that “lower socioeconomic status groups may be more susceptible to its effects.”
Indoor Proportion of Outdoor Particles
The building stock in the future may substantially differ from current conditions. The body of evidence is weak for predicting how such changes may affect the infiltration and persistence of particulate matter from outdoor air. The basis is even weaker for attributing a portion of whatever evolution occurs specifically to climate change. Available information on conditions in the United States indicates that residential buildings have tended to become more airtight (Chan et al., 2005), which reduces air infiltration rates. Measurements in new single-family dwellings in California suggest that low ventilation rates are common in that portion of the building stock: 67% of 108 homes monitored had ventilation rates lower than the California building-code requirement of 0.35 air change per hour (Offermann, 2009). As noted before, lower air-exchange rates tend to provide some protection for building occupants against particles of outdoor
origin. However, with lower air-exchange rates, concentrations of particles and other pollutants emitted from indoor sources would be higher.
The US housing stock seems to be moving toward more widespread use of mechanical systems to provide ventilation (Offermann, 2009; Russell et al., 2007). Mechanical systems that provide supply air can be equipped with filters to remove particles, and high filtration efficiency is available at modest cost (Bekö et al., 2008; Fisk et al., 2002). On the other hand, exhaust-only systems, such as continuous bathroom exhaust fans, do not provide the opportunity to deliberately filter supply air as a means of protecting occupants from outdoor particles. Furthermore, there are concerns that the presence of used filters in ventilation supply systems contributes to degraded IAQ and, for example, may increase the occurrence of sick-building syndrome symptoms (Bekö, 2009). Further technological innovation might be warranted to achieve economical and reliable high-performance mechanical ventilation systems in residences that provide good protection for occupants against particles of outdoor origin.
Ozone and Its Byproducts
Ozone is a secondary pollutant that is formed in the atmosphere by photochemical reactions involving NOx and VOCs. Ozone concentrations in outdoor air have declined slowly in the United States, resisting relatively vigorous efforts to control precursor emissions. The background level of O3 in the clean troposphere also has risen. As health-science information has improved, the NAAQS standard for O3 has become more stringent.
Several modeling studies have explored the consequences of climate change for outdoor O3 concentrations. Hogrefe et al. (2004b) combined a global circulation model, a mesoscale regional climate model, and an air quality model to simulate summertime ozone levels in the eastern United States for the 2020s, 2050s, and 2080s. As compared with the 1990s, and considering only the effects of climate change (using the Intergovernmental Panel on Climate Change’s A2 scenario), the maximum 8-h ozone level increased by 2.7, 4.2, and 5.0 ppb, respectively, for the three future time periods. In a related study, Bell et al. (2007) estimated hourly concentrations in 50 eastern US cities in the 1990s and also for the 2050s, taking account of the predicted change in climatic conditions (again using International Panel on Climate Change [IPCC] Scenario A2) but not accounting for changes in anthropogenic precursor emissions. A key finding of their study is that “on average across the 50 cities, the summertime daily 1-h maximum [O3 level] increased 4.8 ppb, with the largest increase at 9.6 ppb.” Tagaris et al. (2009) reported on the results of a detailed modeling study of outdoor PM2.5 and O3 levels in the United States. Like the study by Bell et al., this study did not account for changes in emission sources or
population. Tagaris et al. estimated that climate-change–induced shifts in PM2.5 levels would cause roughly 4,000 additional deaths per year in 2050 compared with 2001, and 300 additional deaths per year would be caused by increasing O3 concentrations. In an earlier study, Tagaris et al. (2007) reported model predictions of regional concentrations of O3 and PM2.5 over the whole United States, incorporating not only the direct effects of climate change (using IPCC Scenario A1B) but expected emission reductions for the year 2050. They estimated that emitted NOx and SO2 would be reduced by more than 50%. They found that “impacts of global climate change alone on regional air quality are small compared to impacts from emission control-related reductions.” Overall, they predicted a 20% decrease in the mean summer maximum daily 8-h O3 levels and that mean annual PM2.5 levels would be an average of 23% lower. Racheria and Adams (2009) published an analogous study in which they concluded that “climate change, by itself, significantly worsens the severity and frequency of high O3 events over most locations in the US, with relatively small changes in average O3 air quality.”
Buildings offer some protection from O3 exposure because O3 irreversibly decomposes on indoor surfaces and reacts with some gas-phase species (primarily NO and terpenes) that may be found indoors. However, some O3 that penetrates does persist. Given common residual O3 levels indoors and the fact that people spend most of their time indoors, most O3 exposure occurs indoors (Weschler, 2006). New evidence from research on O3-initiated chemistry raises a potentially important question: To what extent are the health risks that are ascribed to ozone exposure influenced by the coincident exposure of the products of ozone-initiated chemistry? Ozone-initiated chemistry, producing potentially health-relevant volatile byproducts such as aldehydes and organic acids, can occur on indoor surfaces (Weschler, 2004), on clothing (Coleman et al., 2008), on hair (Pandrangi and Morrison, 2008), and even on human skin (Wisthaler and Weschler, 2010).
The distinction is important in the context of climate-change effects on IAQ and health. Changes in building design and operation can be anticipated owing to development of new materials, resource limitations, changing economic conditions, changing fashion, and other factors (Weschler, 2009). Such changes might deliberately or inadvertently alter the indoor–outdoor relationship for O3, for example, through the introduction of active or passive controls (Kunkel et al., 2010; Lee and Davidson, 1999; Shair, 1981). Such changes could also deliberately or inadvertently alter the nature, degree, and importance of O3-initiated indoor chemistry. These considerations overlap but are not coincident. Overall, if ambient O3 concentrations increase while ventilation rates decrease, the net effect on indoor O3 concentrations is uncertain, because changes in these two factors have opposing influence on indoor ozone levels. However, both of these factors
tend to increase the indoor concentrations of the byproducts of O3-initiated chemistry.
Researchers have suggested that pollen levels in outdoor air may rise as a consequence of higher CO2 levels, warmer temperatures, and concomitant longer growing seasons resulting from climate change (Ziska et al., 2009), which would have consequences for health outcomes like allergic rhinitis, asthma, and atopic dermatitis (Reid and Gamble, 2009). Intact pollen grains are relatively large (a few tens of micrometers in diameter). Thus, they should neither effectively penetrate into nor persist in indoor air (Liu and Nazaroff, 2001; Nazaroff, 2004; Sippola and Nazaroff, 2003). Nor should they penetrate further into the respiratory system than the head if inhaled (Yeh et al., 1996). Consideration of those factors suggests that buildings would provide good protection against whole pollen grains and that the biologic insult associated with exposure to whole grains should be concentrated in the extrathoracic regions (eyes, nose, and throat). The tracking of pollen grains into buildings (for example, on clothing) might constitute an IAQ and health concern if the grains are later resuspended indoors. Furthermore (and perhaps more important), pollen grains can fracture, generating much smaller particles (0.5–3 µm in diameter) (D’Amato et al., 2007) that carry allergenic proteins. The smaller particles could penetrate both the building envelope and the upper respiratory tract.
Ziska et al. (2011) studied the effects of the rise of frost-free days during 1995–2009 on the length of the ragweed (Ambrosia spp.) pollen season in the United States. They found that the duration of the season increased by 13–27 days at latitudes greater than about 44°N.5 They noted that longer pollen seasons and higher exposure to pollen may intensify allergic sensitization and increase the duration and severity of allergy symptoms. The committee did not identify any literature specifically regarding climate change and indoor exposure to pollen.
Algal Blooms After Floods
Harmful algal blooms (HABs) occur when saltwater or freshwater reservoirs accumulate algae or other protozoa to abundances at which their biomass or toxins lead to adverse effects on aquatic life or humans. Human activity can affect the frequency and severity of HABs, for example, through
5 The 44th parallel passes through the northern United States, including Oregon, Idaho, Wyoming, South Dakota, Minnesota, Wisconsin, Michigan, New York, Vermont, New Hampshire, and Maine.
increased fertilizer runoff and aquaculture that leads to eutrophication in rivers and coastal areas. Climate also affects the appearance and distribution of HABs. Two factors with known effects are change in water temperature and changes in nutrient levels. Changing patterns of the types of species involved in HABs and their timing can occur with increases in mean water temperature (Glibert et al., 2005). Increased Saharan dust storms have been shown to be rich in iron, a limiting nutrient, which can lead to increases in some species of algae that proliferate in the Caribbean (Lenes et al., 2001).
The red-tide alga, Kernia brevis, produces brevetoxin. These cyclic polyether molecules “become part of the marine aerosol as the fragile, unarmored cells are broken up by wave action. Inhalation of the aerosolized toxin results in upper and lower airway irritation” (Milian et al., 2007). In Florida, increased respiratory irritation has been reported almost annually during red-tide events. Researchers have found that residents, lifeguards, and tourists report many more respiratory symptoms after exposure to red-tide events (Backer et al., 2003, 2005; Fleming et al., 2005). During two red-tide exposure periods in 2005 and 2006, a cohort of asthmatic children and adults in Sarasota, Florida, was studied (Fleming et al., 2009). Their exposure to brevetoxin was assessed via personal air sampling, and their symptoms via questionnaire (Cheng et al., 2010). Researchers observed associations between brevetoxin exposure and increased respiratory symptoms. Other researchers have also reported increases in emergency-department visits for asthma, pneumonia, and bronchitis in residents during red-tide events (Kirkpatrick et al., 2006). There is also evidence that brevetoxin can affect the mucociliary escalator in animals (Abraham et al., 2005). Therefore, exposure to brevetoxin in red-tide events could affect the respiratory tract’s ability to clear other inhaled particles, such as allergens, endotoxins, and fungal spores. The ramifications would be increased residence time of the particles in the airways and a higher biologically relevant dose. A monitoring system is in place for red-tide events and other HABs, but it is mainly for ensuring seafood safety; it is not linked to respiratory health protection.
Flooding caused by extreme precipitation events, which may increase in number and severity under climate change conditions, are commonly followed by disease clusters (IOM, 2008) and may lead to both more frequent HABs and increased exposure to potentially harmful agents associated with them. The committee could not, however, identify any literature specifically addressing changes in risk associated with exposure to these agents in indoor environments.
Ambient SO2 is primarily a result of coal combustion and originates from the presence of sulfur as an impurity in coal. EPA data show that about 68% of nationwide atmospheric sulfur emissions in 2002 were from “electricity generating units” and that the other important sources were “industrial/commercial/residential fuels” (about 16%) and “industrial processes” (about 8%) (EPA, 2009a). Ambient NAAQSs for SO2 and acid-rain legislation (in the 1990 Clean Air Act Amendments) have led to substantial reductions (about 50%) in SO2 emissions from power plants. The largest remaining emissions are from older power plants whose high emission rates continue to be allowed. New coal-fired power plants are required to have good emission controls for SO2 that are achieved, for example, with flue-gas desulfurization.
Future ambient SO2 levels might rise or fall depending on changes in the use of coal as an energy source and on emission controls. The indoor environment provides some protection against SO2 because, as an acid gas, it reacts on indoor surface materials (Biersteker et al., 1965; Grøntoft and Raychaudhuri, 2004; Walsh et al., 1977).
NOx (mainly NO and NO2) are emitted primarily as a result of combustion. To some extent, the presence of nitrogen in fuel (as in coal) leads to NOx emissions. However, any high-temperature combustion process that uses air as the oxidizer can also produce NOx emissions, with the nitrogen originating from N2 in the combustion air. Important sources of NOx in ambient air are mobile sources (both on-road and off-road), fossil-fueled power plants (using coal and natural gas), and other stationary combustion of (mainly) fossil fuels. For 2002, EPA national emission inventory data indicate that mobile sources were responsible for about 60% of NOx emissions. Because NOx is a precursor of O3 and other photochemical smog components, it has been and continues to be subjected to strong emission-control efforts, and continuing progress in reducing emissions in the near future can be expected. A high level of scrutiny and emission control is expected especially for diesel emissions, which are becoming progressively more important (Dallmann and Harley, 2010). Less future reliance on fossil fuels in particular and combustion in general suggests that NOx emissions may also decrease in a future climate-change regime. The indoor environment provides modest to moderate protection against NO2 of outdoor origin (Quackenboss et al., 1986).
Hazardous Air Pollutants
There is a long list of species known as HAPs. In the United States, about 190 HAPs were designated under the 1990 Clean Air Act Amendments. In contrast to the criteria pollutants,6 HAPs are regulated only with respect to emissions from major sources; there are no ambient concentration standards, and the concentrations of these pollutants are not routinely monitored. However, summary appraisals have combined emissions data with dispersion modeling and risk factors to discern which pollutants are most prevalent and where the health risks posed by HAPs are highest. For example, in one study, the median hazard ratio (average ambient concentration divided by a cancer benchmark value) was highest for 1,3-butadiene, formaldehyde, benzene, carbon tetrachloride, chromium, methyl chloride, and chloroform (Woodruff et al., 1998). For chronic noncancer toxicity, acrolein had the highest median hazard ratio in the study. Studies investigating the indoor–outdoor relationships of HAPs reveal that for many species in many buildings indoor concentrations are higher than those outdoors (e.g., Jia et al., 2008a). That characteristic demonstrates the importance of indoor emission sources in contributing to indoor levels. On the other hand, levels of indoor HAPs that are attributable to their presence in outdoor air have not been well studied. For some important species, such as benzene and the chlorinated organics, it is reasonable to expect that indoor environments provide little or no protection from outdoor concentrations. For other species that have higher reactivity, such as acrolein and aldehydes, the penetration and persistence of outdoor pollutants into indoor environments is not known. Future trends in the outdoor levels of the pollutants in a climate-change regime are not clear, but the scrutiny that they are receiving as HAPs suggests that emissions might decline.
One of the dominant environmental health concerns in developing countries results from the use of solid biomass fuels for cooking. Around 2.7 billion people are thought to rely on burning of biomass (dung, crop residues, and wood) or coal as their household fuel, and about 82% of them live in rural areas (IEA/UNDP/UNIDO, 2010). In sub-Saharan Africa, around 80% of the population relies on the traditional use of biomass for cooking, and the electrification rate is only 31%. Combustion of such fuels in open fires or cookstoves is generally inefficient and leads to very high concentrations of products of incomplete combustion, which have
6 The criteria pollutants defined under the NAAQS are O3, particulate matter, CO, NOx, SO2, and lead.
serious health implications for those exposed. The products of incomplete combustion form a complex mixture of pollutants, including fine PM, NO2, sulfur oxides (particularly in the case of coal), CO, and polycyclic aromatic compounds, such as benzo[a]pyrene (Smith, 1993). Many of the compounds are known to pose health hazards. Research also has demonstrated that indoor air pollution from solid biomass fuels may contribute to climate change as a result of emissions of black carbon, methane, CO, and nonmethane VOCs (which are O3 precursors). The overall effect depends on the balance between warming aerosols, such as black carbon, and other types of particles that may be cooling; the net effect is a likely contribution to warming. Black carbon may also accelerate the melting of glaciers when it is deposited in mountainous areas, such as the Himalayas (Ramanathan and Carmichael, 2008). Even if the fuel contains contemporary carbon (for example, the carbon associated with crop residue, wood, or animal dung), there may be a greater climate-change effect than would be the case for an efficient stove burning a fuel that contains fossil carbon, such as liquefied petroleum gas. Improving the cooking conditions for this large population offers the potential for cobenefits: improved public health and reduction in climate-change effects. Wilkinson et al. (2009) quantify the cobenefits for rural India, demonstrating large potential benefits in improved public health and not insignificant improvements in climate-change effects.
Exposures to such pollutants as PM in houses that burn biomass or coal tend to be very high. For example, PM levels can be one or even two orders of magnitude above the EPA NAAQSs, depending on type of fuel, stove characteristics, and housing.
Exposure to air pollution associated with indoor use of solid fuels has been implicated, with various degrees of certainty, as a causative factor in several adverse health outcomes. In total, it is thought that around 1.6 million deaths a year can be attributed to indoor air pollution of this type globally (Ezzati et al., 2002).
Increased incidences of acute respiratory infections in children and of chronic obstructive airways disease, particularly in women, comprise the most compelling evidence for adverse health effects (Bruce et al., 2000). Exposure to these types of indoor air pollution, particularly when coal is used, may also cause lung cancer in women. In low- to middle-income countries—such as China, India, and Mexico—two-thirds of women who have lung cancer are nonsmokers. Exposure to so-called smoky coal seems to be strongly related to lung cancer in China (Mumford et al., 1995). The situation is less clear with regard to wood-smoke exposure, although exposure to known carcinogens, such as benzo[a]pyrene, is likely to be equivalent to smoking several cigarettes per day, so such a health risk cannot be ruled out. There is also some epidemiologic evidence of a causal association
between indoor exposure to smoke and upper aerodigestive tract7 cancers in Latin America (Pintos et al., 1998).
There is reasonably consistent evidence of an association between particulate air pollution and both hospital admissions and cardiovascular-disease mortality, but the association specifically with myocardial infarction is less consistent (Bhaskaran et al., 2009). And although it is plausible that indoor air pollution increases the risk of cardiovascular disease events and mortality, there is a lack of direct evidence (Smith, 2000).
Low birth weight has been associated with use of wood fuel by mothers in Guatemala, perhaps mediated through CO exposure (Boy et al., 2002), and there may be an association with perinatal mortality. Several studies in India have shown an association between biomass fuel and cataracts (cortical, nuclear, and mixed but not posterior subcapsular in one study) (Mohan et al., 1989), and animal studies support this association. A possible association between exposure to wood smoke and tuberculosis has been indicated in a few studies (e.g., Mishra et al., 1999), but a 2010 review concluded that “available original studies looking at this issue do not provide sufficient evidence of an excess risk of tuberculosis due to exposure to indoor coal or biomass combustion” (Slama et al., 2010).
Wilkinson and colleagues (2009) report that advanced biomass stoves available in India at prices of US$20–50 can achieve around 15 lower particle emissions per meal compared with traditional stoves. Hybrid gasifier stoves with small electric blowers can achieve good performance with a range of fuel characteristics but require a source of electricity. There has been experience with large national stove programs, such as the provision of 180 million improved stoves in China over 12 years starting in 1983. A modeling exercise indicated that a 150 million–stove program in India, to be implemented over 10 years, that would provide improved-efficiency cookstoves through various delivery mechanisms, including to poor women receiving antenatal care, could result in the prevention of around 2.2 million premature deaths from acute respiratory infections in children and from COPD and ischemic heart disease in adults (Wilkinson et al., 2009). However, because of the slow evolution of COPD, there may be a substantial lag between the introduction of the new stoves and the reduction in deaths. Wilkinson and colleagues also estimated that such a program could result in a reduction in greenhouse pollutants equivalent to 0.5–1.0 billion tons of CO2 over the decade. Many other benefits could accrue, such as reductions in time spent looking for fuel, in expenditure for fuel, and, in some locations where wood is not harvested sustainably, in deforestation.
7 The aerodigestive tract comprises the upper respiratory and digestive tracts, including the tissues associated with the nose, lips, tongue, mouth, throat, vocal cords, and upper portions of the esophagus and windpipe.
A randomized trial in Guatemala of improved woodstoves with chimneys compared with traditional open fires showed encouraging preliminary results, particularly with regard to acute respiratory infections in infants up to 18 months old, but did not have the power to assess differences in mortality or birth weight (WHO, 2007). Other larger trials were planned or under way at the time this study was completed.
Although levels of indoor air pollutants may be lowered by the use of biogas or liquefied petroleum gas (LPG) or, of course, by provision of electricity, these are all more expensive alternatives and may not be suitable for some communities. The development and implementation of culturally acceptable highly efficient biomass cookstoves will be needed for the indefinite future to meet the requirements of many poor communities for inexpensive, clean, and convenient sources of household energy. Such efforts could yield cobenefits in contributing to the mitigation of climate change.
There is little published information about how climate change might affect IAQ in low-income countries. In theory, climate change could affect the availability of biomass fuel, but the direction and magnitude of the change may depend on the region and on the relative effects of carbon dioxide fertilization and changes in temperature and precipitation. More work is also needed on how changes in forest management, such as the trend toward greater use of forest plantations to supply roundwood, and the commercial exploitation of crops for biofuel, may affect the availability of biomass for use by poor households and thus indirectly affect the indoor environment.
On the basis of its review of papers, reports, and other information presented in this chapter, the committee has reached the following conclusions regarding the effects of climate change on IAQ and its consequent influence on public health.
- The elements that influence important outcomes in the climate-change–IAQ–public-health nexus are numerous and diverse. They are interconnected in a complex manner that includes feedback loops and interweaves natural processes with technology, individual human behavior, and social systems. It is those systemic features, rather than the nature of individual elements themselves, that pose the greatest challenges for understanding and effectively addressing the effects of climate change on IAQ and public health.
- We understand relatively little about how climate change will affect IAQ and thereby public health. More information is available on the factors that influence indoor concentrations of health-relevant
pollutants and how concentrations might shift as a consequence of climate change.
- Three classes of factors have important influences on indoor pollutant concentrations: pollutant attributes (including source properties), building characteristics, and human behavior.
- The concentration of any indoor air pollutant in any indoor space is governed by a balance between emissions and removal. Concentrations in combination with human occupancy govern exposures. Excessive exposures confer health risks on those exposed. Climate change can affect this system in numerous particular ways. For example, by causing an increase in the outdoor concentrations of some pollutants in some places and at some times, climate change is likely to increase indoor concentrations and associated exposures in buildings in those places and at those times.
- In addition to direct shifts caused by climate change, there are likely to be shifts in IAQ that are mediated by human responses to climate change. For example, mitigation measures to reduce energy use in buildings could lead to systematically lower ventilation rates in buildings that would cause higher concentrations of and exposures to pollutants that are emitted from indoor sources. Another example is the increased use of air conditioning, an expected adaptation measure, which could exacerbate anthropogenic emissions of greenhouse gases and, if accompanied by reduced ventilation rates, increase the indoor concentrations of pollutants emitted from indoor sources.
- Reactions to weather emergencies pose public-health risks related to IAQ, such as the potential for poisoning from exposure to CO emitted from emergency electricity generators. Such emergencies may increase in frequency if climate change results in more frequent or more severe climate events like storms.
- Actions taken by individuals can profoundly influence the IAQ in individual buildings. There is a public interest as well as an individual interest in seeing that the system for establishing and maintaining good IAQ works well. Negligent or ill-informed behavior by individuals can cause serious harm.
- Indoor pollutant concentrations can be separated into contributions from indoor sources and from outdoor air. Combustion is a major source of both outdoor and indoor air pollution and is arguably the most important source of indoor air pollutants with respect to health risks. Important combustion-related issues associated with indoor emissions are CO exposures from emergency-generator use and IAQ problems associated with cooking, heating, smoking, and small-scale activities, such as use of candles and in-
cense. Other important pollutants associated primarily with indoor sources include radon, VOCs, and SVOCs. Outdoors, the main pollutants of concern are PM and O3. Specific PM concerns that may be exacerbated by climate change include increases in smoke from wild-land fires, pollen, and windblown dust.
- There is a large gap between what is known about IAQ and public health and what needs to be known. However, the gap between what is known and what is done to address problems is even larger. One of the risks associated with how climate change will affect IAQ is that the gap between what is known and what is done will grow and have adverse consequences for public health.
- In formulating responses to the challenges posed to IAQ and public health by climate change, it is important to recognize and account for the diversity in subpopulations, in part because of variability in susceptibility to the effects of indoor air-pollutant exposure. It is also important to take account of variability within populations in the knowledge and resources with which to take effective action in response to changing conditions.
- It is important to take proper account of the different issues of concern and appropriate responses for different building types, such as single-family dwellings, multifamily apartment buildings, schools, health-care facilities, and offices.
- Efforts to save energy by improving building performance need to be accompanied by strong caution with respect to changing building ventilation rates. Two driving forces are apparent. First, as a mitigation measure, efforts to save energy in buildings are gaining momentum. Energy is required to condition the temperature and humidity of ventilation air, so individuals and organizations may seek to save energy by reducing the rate of ventilation of indoor spaces. Second, as temperatures rise during the warm parts of the year, there may be a progressive shift to greater reliance on air conditioning and away from cooling by means of open windows. The effect of ventilation on IAQ has multiple facets that operate in different directions, so one cannot be certain a priori of the net effect for each building. A lower building ventilation rate will tend to provide enhanced protection against some pollutants from outdoors, such as PM. But reduced ventilation rates tend to cause concentrations of pollutants that originate primarily from indoor sources to increase. Reducing ventilation rates does not automatically mean that IAQ problems will become worse; nor is it appropriate to assume that no problems will be associated with reducing ventilation rates.
- Attention to controlling indoor emission sources is warranted. For many pollutants with indoor sources, it has been found that variability in emissions, not variability in ventilation rates, is the primary determinant of whether indoor air-pollutant levels are excessive. Put another way, when the indoor emission rates are high, ventilation in a normal range is unlikely to be sufficient to avoid a problem. There is no evidence that clearly links increased indoor pollutant emission rates to climate change. However, there are several potential concerns that deserve attention, including CO from emergency generators, emissions from cooking, emissions from heating systems (including unvented combustion appliances, backdrafting, and increased use of wood as a fuel), emissions from smoking, emissions from use of candles, radon from intimate contact of indoor spaces with earthen materials, and VOCs and SVOCs from various indoor sources. Special attention is needed to ensure that life-cycle impact assessments aimed at improving the environmental performance of buildings take proper account of the disproportionately large effects that emissions from indoor materials and processes can have on IAQ and public health.
- Attention ought to also be directed toward improving understanding of the effectiveness of indoor environments as a shelter against pollutants of outdoor origin that may be altered owing to climate change. To date, the scientific literature on the affects of climate change on outdoor air quality has focused on criteria pollutants, especially PM2.5 and O3. There are good regulatory and technologic systems in place that are striving to reduce emissions from anthropogenic sources. The momentum is expected to continue to yield improvements in reducing ambient pollutant concentrations that are clearly associated with anthropogenic sources. A greater concern would apply to pollutants that lie outside the regulatory structure of managed emissions, such as smoke from wildfires, pollen from weeds, and windblown dust. Pollutant emissions from sources like these might be substantially worsened by climate change. If so, indoor environments will be used as imperfect shelters that could be improved with proper attention and a commitment of appropriate resources.
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