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4
Air Quality
INTRODUCTION
Indoor air quality (IAQ) is an important component of indoor environ-
mental 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 arthro-
pods, 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 emerg-
ing 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 pollut-
ants in outdoor air (Bell et al., 2004; Dockery et al., 1993; Jerrett et al.,
79
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80 CLIMATE CHANGE, THE INDOOR ENVIRONMENT, AND HEALTH
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 pollu-
tion 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 emis-
sion 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 SOURCES OF POLLUTANTS
Indoor environments detain pollutants that are emitted indoors. This
section reviews important IAQ issues that are associated with indoor pol-
lutant 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 observa-
tion 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 signifi-
cance 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
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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. In-
door 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 lead-
ing 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.
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82 CLIMATE CHANGE, THE INDOOR ENVIRONMENT, AND HEALTH
1998.” Holmes and Russell (2004) remarked that the reduction in acciden-
tal 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 ve-
hicle 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 ex-
posure 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 over-
loads 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 out-
ages (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 con-
cern would be the use of natural-gas–fueled and petroleum-fueled stoves
for heating, excessive reliance on unvented combustion-based space heat-
ers, 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)
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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 evaluat-
ing the use of a web-based query system for public health surveillance re-
ported almost 25,000 CO-related hospitalizations across the United States
in 2005, of which approximately 4,200 were confirmed CO-related poison-
ings (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 asso-
ciated 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 work-
ing 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
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 ex-
ample, the use of natural gas as a cooking fuel is associated with increased
indoor exposures to nitrogen dioxide (NO2), a byproduct of the combus-
tion 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).
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84 CLIMATE CHANGE, THE INDOOR ENVIRONMENT, AND HEALTH
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 expo-
sure 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 con-
centrations (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). Expo-
sure 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
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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.
Space Heating
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 heat-
ing 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 driv-
ing 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 combus-
tion 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 leak-
age of combustion byproducts into the indoor environment (Gustafson
et al., 2008; Traynor et al., 1987). If done poorly, increased wood-based
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86 CLIMATE CHANGE, THE INDOOR ENVIRONMENT, AND HEALTH
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 heat-
ing 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 lit-
erature (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.
Smoking
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 pollut-
ants, 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 chil-
dren (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
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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 smok-
ing 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 con-
currently 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 prod-
ucts 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, unin-
tended airflow pathways in multiunit residential buildings can lead to expo-
sures 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 condi-
tions, 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 As-
sociation, 2011), US retail sales of candles are roughly $2 billion per year,
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88 CLIMATE CHANGE, THE INDOOR ENVIRONMENT, AND HEALTH
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 can-
dles 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 fac-
tor 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
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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 nonsmok-
ers (Ginevan and Mills, 1986).
The three main sources of indoor radon are soil near a building’s foun-
dation; 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.
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