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Our understanding
of health effects related to the indoor environment has evolved
over the past decade. In the past, discussions of indoor
environmental quality (IEQ) focused on indoor air constituents
(primarily particles, bioaerosols, and chemicals), and comfort
factors (temperature, air flow, and humidity) (Samet et al.
1998). More recently, we have begun to look at the relationship
between the built environment and humans as a complex interplay
between building occupants (who they are and what they do) and
an array of physical, chemical, biological, and design factors.
This evolution in understanding has profound implications for
the design and operation of buildings, how the buildings are
used, and the prevention and management of health problems that
occur in building occupants.
Source Characterization
Outdoor air pollution is a dynamic
system in which the physical and chemical processes affecting
the accumulation of pollutants in the atmosphere are constantly
changing, largely driven by complex meteorology and
photochemistry. In contrast, the usual approach of modeling
indoor air pollution considers only pollution source strength
and dilution by air exchange, thus treating the indoor
environment as a static box in which physical and chemical
transformations of indoor air pollutants are absent or
negligible. This misconception produces conservative estimates
for primary indoor air pollutant concentrations and ignores the
secondary pollutants. In-depth studies of indoor air have shown
that the concentration of agents in indoor air is a function of
outdoor concentration, indoor source strength, removal and
deposition rate within the structure, indoor mixing, and
chemical reaction. In the following sections, we use real-world
examples to illustrate the dynamic nature of these processes and
to discuss the implication of this dynamic environment in
assessing exposures and health effects associated with indoor
air pollution.
Indoor production. The
generation of pollutants within the indoor environment may come
from primary and secondary sources. Primary sources include fuel
combustion for cooking, heating, and lighting; tobacco smoking;
bioeffluents from humans and animals; floor and wall coverings;
synthetic paints, glues, polishes, and waxes; pesticides; and
building products. Another source is the release of gases from
solvents used indoors or from water that is used daily for
showers, bathing, cooking, and from drinking fountains. Such
sources are important for by-products (e.g., chloroform) of
chlorination-based water disinfection and radon (McKone and
Knezovich 1991; Xu and Weisel 2005). Because of the use of many
types of synthetic materials in our daily lives, concentrations
of many volatile organic compounds (VOCs) are consistently
higher indoors than outdoors in residences and offices in
developed countries. For some VOCs such as limonene, indoor
levels up to 10 times those outdoors are common, even in
locations with significant outdoor air pollution sources, such
as petrochemical plants (Ott and Roberts 1998; Weisel et al.
2005). Secondary sources refer to indoor chemistry that
transforms a set of indoor pollutants, emitted from primary
sources or transported from outdoors, to a new set of indoor
pollutants, as discussed below.
Outdoor-to-indoor transport.
Pollutants of outdoor origin, including those present in the
outdoor air and those released from soil sources, can be
transported indoors via building openings and cracks (Garbesi et
al. 1999; Nazaroff 2004). Attempts have been made to estimate
the fraction of measured indoor concentration contributed by
outdoor air due to the outdoor-to-indoor transport process (Ott
et al. 2000; Thatcher and Layton 1995). One such study, the
Exposures of Adult Urban Populations in Europe Study (EXPOLIS),
compared concentrations of ambient particulate matter ≤ 2.5 μm
(PM2.5), its 16 elemental constituents and black
carbon, 30 VOCs, and carbon monoxide (CO) among urban adult
populations in seven European cities. The study examined
exposures in workplaces, residential outdoor and indoor air, and
separated workday and leisure time (Jantunen et al. 1998).
EXPOLIS data from Helsinki, Finland, showed the infiltration
factor (the proportion of outdoor PM found indoors) for PM2.5
averaged 0.64 for residential structures, 0.47 for workplaces,
and 0.35 for a subsample of office buildings constructed after
1990 (Hänninen et al. 2004b, 2005). In another study, the
Relationship of Outdoor, Indoor, and Person Air (RIOPA),
fractions of measured indoor concentration contributed by
outdoor air for PM2.5
and each of 24 VOCs including 10 aldehydes and ketones were
estimated for 310 residences located in three U.S. cities (Weisel
et al. 2005). The median fractions of measured indoor
concentration contributed by outdoor air for compounds with
dominant indoor sources were less than 50%, for example, 13% for
d-limonene (a common cleaning solvent), 20% for chloroform (a
byproduct of drinking water disinfection), 31% for α-pinene and
20% for β-pinene (ingredients of synthetic paints), and 19% for
formaldehyde (released from building/furnishing materials). For
the compounds with sole or dominant outdoor sources (e.g.,
methyl tert butyl ether, carbon tetrachloride, and
trichloroethylene), the fractions were about 100%, as expected.
The fractions for PM2.5 had a median of 56%, 25th
percentile of 46%, and 75th percentile of 93% across the RIOPA
homes (Meng et al. 2005; Weisel et al. 2005).
Significant interhome variability
in fractions of measured indoor concentrations contributed by
outdoor air has been observed for PM2.5 and most of the VOCs in
the RIOPA study. This finding has important implications for air
pollution epidemiologic studies using concentrations measured at
outdoor locations. Numerous exposure studies have shown poor
correlations between personal exposure or residential indoor
concentration and outdoor concentrations, indicating the
observed associations between adverse health effects and PM
concentrations measured at fixed outdoor sites do not
necessarily represent the exposure–response relationships (Adgate
et al. 2004; Clayton et al. 1993). Although attempts have been
made to differentiate PM of outdoor origin from PM of indoor
origin, analyses have been complicated because the fraction of
indoor species contributed by outdoor air depends not only on
outdoor concentration but also on home-specific parameters
including air exchange rate [AER; typically expressed as air
exchanges per hour (ach)], indoor generation rate, removal rate,
and house volume (Meng et al. 2005; Thomas et al. 1993; Wallace
et al. 1991).
Outdoor-to-indoor transport of
very reactive chemical species has often been considered
unimportant. An example is ground-level ozone (O3)
that is formed via photochemical reactions and has elevated
concentration in polluted atmospheres during photochemical smog
episodes. O3, like PM, is regulated in the United
States as a criteria pollutant. Because of its high reactivity,
only a fraction of O3 can penetrate a building
envelope. This fraction had been considered insignificant to
cause any exposure concerns until 1989 when Weschler et al.
(1989) showed that indoor exposure to O3 can easily surpass
outdoor exposure. Under moderate AERs (~ 0.5 ach), indoor O3
concentrations may be 20–30% of corresponding outdoor
concentrations. Under high AERs (> 1 ach), indoor O3
levels can be 50–70% of outdoor levels. In a study carried out
in six homes located in suburban New Jersey, indoor O3
concentrations were 22–66% of outdoor levels during afternoon
hours (Zhang et al. 1994). In summer time, 50% of the schools
measured in Mexico City had indoor O3 levels > 113
ppb (Gold et al. 1996). It is reasonably conservative to state
that indoor O3 levels > 20 ppb are common when
outdoor O3 concentrations are elevated. O3
concentration at 20 ppb may not be sufficient to cause health
concerns due to direct O3 exposure, but this O3
level can be sufficient to drive a complex set of indoor
chemical reactions. When O3 generators (so-called air
purifiers) are used at O3 generation rates of tens to
thousands of milligrams per hour, indoor O3
concentrations can be in the parts per million levels in a room
with typical volume and AER.
Particle sources include both
indoor home and residential sources, although recent research
has shown that indoor (workplace and residential) contributions
to total exposures may be underestimated compared with outdoor
sources such as traffic (BeruBe et al. 2004; Koistinen et al.
2004). This appears to depend on the character of the particle;
combustion-derived particles may be due more to outdoor sources,
whereas other particles (for example, soil-derived particles)
may be related to resuspension of particles during a host of
indoor activities (Ferro et al. 2004; Larson et al. 2004).
Recent experiments have shown that a wide range of indoor
activities can result in considerable generation of PM (Afshari
et al. 2005). Models of indoor PM exposure have been developed
to account for both indoor and outdoor sources, as well as
mixing, transport, and removal (Georgopoulos et al. 2005;
Nazaroff 2004).
Indoor-to-outdoor transport.
Ventilation is the primary factor affecting
indoor-to-outdoor transport of indoor generated pollutants.
Ventilation is necessary to reduce concentrations of pollutants
generated indoors, but it is also necessary to reduce the time
available for chemical reactions among indoor pollutants. One
reason offered to support the conventional view of indoor
chemistry being insignificant is that chemical reactions among
indoor pollutants are too slow to complete with air exchange
processes. Although this may be true when the AER is high, a
variety of chemical reactions can take place at AERs typical of
today’s residences and offices. Since the late 1970s, the
airtight design of buildings, driven mainly by energy
conservation, has resulted in reduced AERs. Based on
approximately 4,590 measurements of residential AERs conducted
across the United States, Pandian et al. (1998) reported that
the mean, median, and SDs of AERs were 0.55, 0.42, and 0.47 ach,
respectively, for the northeastern region, and 0.71, 0.62, and
0.56 ach for the southeastern region of the United States. AERs
of this magnitude are undesirable for removing air pollutants
that originate indoors and are low enough for certain chemical
reactions to occur.
Indoor chemistry.
Pollutants can be removed from indoor air through both physical
and chemical processes. Physical processes that can result in
pollutant removal (in addition to transport outdoors) include
phase change, adsorption or absorption, or dissolving in water
or organic films. Recently there has been considerable research
interest in removal of pollutants through chemical reactions.
“Indoor chemistry” has been
defined as reactions involving indoor pollutants, occurring
either in the gas phase or on surfaces (Weschler et al. 2006).
For a chemical reaction to influence the indoor environment, the
rate of the reaction must be sufficient to compete with AERs.
These chemical reaction processes represent sinks for the
reactants (primary indoor pollutants) and sources of new
reaction products (secondary indoor pollutants). The products
may predominate in the air or on the surface. Removal does not
necessarily occur in a simple linear fashion; for example,
semi-volatile organic compounds can undergo an initial removal
followed by a secondary increase due to resuspension of the
compounds adsorbed on particles (Lioy 2006).
Both gas-phase reactions and
surface reactions that can occur under typical indoor conditions
have been identified. The most extensively studied gas-phase
reactions are oxidation reactions involving O3 and
free radicals. O3 drives most indoor oxidation chemistry because
it can react at meaningful rates with nitric oxide, nitrogen
dioxide, and unsaturated organic compounds (e.g., terpenes,
terpenoids, sesquiterpenes, unsaturated fatty acids) to yield
reactive intermediates, the hydroxyl radical (OH), the nitrate
radical (NO3) and oxygenated organic compounds (Weschler
and Shields 1996). Reactions of O3 with NO2,
in the absence of sunlight, form the NO3 radical that
further reacts with VOCs, leading to the formation of indoor
nitric acid. The NO3 radical can also react with NO2
to form dinitrogen pentaoxide (N2O5) that
undergoes hydrolysis, another pathway of nitric acid formation (Weschler
et al. 1992). When O3 and NO2 are present
simultaneously, indoor NO3 may be the dominant indoor
oxidant that effectively reacts with nearly all indoor VOCs. The
role of indoor NO3 chemistry in transforming indoor
air pollutants remains to be evaluated.
Several terpenes, especially
d-limonene and α-pinene, are present at substantially higher
concentrations indoors compared those with outdoors. These
terpenes react readily with O3 under typical or
realistic indoor conditions to initiate a series of complex
chemical reactions, for example, at an O3
concentration of 20 ppb, the rate constant for O3
reaction with d-limonene and α-pinene is approximately 0.36 ach
and approximately 0.15 ach (Fan et al. 2003). Products of these
reactions are found in both the gas and particle phases.
Gas-phase–stable products include aldehydes, carboxylic acids,
potentially allergenic peroxides and hydroperoxides (Fan et al.
2003). In one experiment where O3 (~ 41 ppb) was
mixed with a VOC mixture comprising 23 commonly found VOCs, the
resulting peak concentration of ultrafine and fine particles was
approximately 100 μg/m3 (Fan et al. 2005). Although
attempts have been made to chemically identify the resulting
particles, the majority of the particle mass could not be
explained by the compounds identified thus far (Fan et al.
2003). It will be even more challenging to identify the
short-lived, highly reactive, thermally labile or highly
oxidized species that are formed in this complex reaction
system. Unstable products of the ozone–terpene reactions include
reactive intermediates and the hydroxyl radical. Hydroxyl
radicals resulting from these indoor reactions can reach levels
higher than typical nighttime outdoor concentrations, and thus
react with other indoor VOCs with which ozone reacts too slowly
to be of any practical significance (Weschler and Shields 1996).
The formation of particles via O3-driven
indoor chemistry has two implications. First, in an analysis of
indoor particles measured in residences located in several
United States cities, 25% of indoor PM2.5 could not
be explained with known sources (Wallace 1996). Indoor chemistry
was not considered in the analysis, which might explain at least
part of the unknown sources. Second, because O3 and
fine particles are co-generated outdoors during photochemical
episodes, indoor particles resulting from indoor O3/VOC
reactions can vary coincidently with the variations of outdoor
summertime fine particles. This will certainly complicate the
effort to separate PM of outdoor origin from PM of indoor
origin. It should also be noted that source characterization may
vary significantly, depending on the size of the particles (Koistinen
et al. 2004).
A second type of indoor chemistry
involves surface reactions. Outdoor aerosol surfaces play an
important role in atmospheric chemistry. The importance of
surface reactions indoors is easily recognized, given that
surface-to-volume ratios indoors are much larger than outdoors
(roughly 3 vs. 0.01 m2/m3). Indeed, indoor
surfaces may be ideal for substance sorption and for water
condensation. Surface water film can react with indoor NO2,
a major product of natural gas combustion, to form nitrous acid
(HONO) and nitric acid (HNO3). The resulting nitrous
acid is released into the air as gas-phase HONO, whereas nitric
acid remains on surfaces as an HNO3–H2O
complex (Dubowski et al. 2004). The latter yields possible
acidic, oxidizing, and nitrating surface films on interior
walls. O3 reacts with unsaturated VOCs contained in
surface coatings at a faster rate than when it reacts with the
same compounds in the gas phase (Reiss et al. 1995).
Indoor surfaces, including
building materials, wall cavities, ducts, skin, clothing, dust,
and airborne particles are very diverse and are a determining
factor of indoor surface chemistry. They affect HONO formation
via surface-NO2 chemistry (Wainman et al. 2001).
Complex physical and chemical processes involving surfaces
include sorption, redox reactions, acid-base chemistry and
hydrolysis (Nazaroff and Singer 2004). For example, diphthalate
esters (plasticizers contained in polyvinyl chloride flooring
materials) can undergo hydrolysis to form alcohols and
monoesters. Aldehydes are emitted, at concentrations exceeding
their odor thresholds, when O3 interacts with carpets
(Morrison and Nazaroff 2002).
Building materials contain a large
number of reactive constituents that can be released into the
indoor air along with secondary products, including terpenoids,
aliphatic aldehydes, phthalates, phenol, mono- and dicarboxylic
acids, diisocyanates, and various photoinitiators.
Photoinitiators, contained in ultraviolet curable coatings, can
undergo decomposition to generate free radicals, and some (e.g.,
benzaldehyde and cyclohexanone) are precursors of odorous
products (Salthammer et al. 2002). In a study conducted in
German houses constructed with wooden studs treated with
pentachlorophenol (PCP), it was found that over time PCP had
been transformed to tetra-chloroanisole, a compound of highly
undesirable odor (Gunschera et al. 2004).
Indoor oxidation chemistry is
largely driven by O3 reactions with unsaturated VOCs
and perhaps with NO2 as well. Given that ozone levels have been
rising in many areas, that indoor use of unsaturated VOCs (e.g.,
terpenes) has been on the rise, and that AERs have been
decreasing, indoor oxidation chemistry has likely increased over
the past several decades.
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