approaches to chemical inventories, which are important because they enable more accurate chemical exposure estimates, followed by a discussion of the wide range of analytical approaches that can be used to measure the chemical contents of an indoor space. Finally, the chapter concludes by highlighting emerging chemicals of concern indoors, current sampling and measurement capabilities and limitations, and areas of research gaps and future directions. This chapter sets the stage for the committee’s subsequent discussions of how these primary emissions are redistributed in the indoor environment (Chapter 3) and chemically transformed into secondary sources (Chapter 4).

As an example, studies have shown that house dust concentrations of some chemicals, such as flame retardants and plasticizers, are correlated to internal dose for home occupants (Percy et al., 2020; Stapleton et al., 2012). Due to this association, there is increased focus and attention on house dust, including a push by researchers to characterize house dust using non-targeted mass spectrometry approaches to support identification of other chemicals that may be of concern for human exposure (Ouyang et al., 2017; Ulrich et al., 2019).

Biological materials such as microbes, endotoxins, and pests can also be considered continuous indoor sources of VOCs, allergens, and irritants. Their emission rates are largely dependent on environmental factors such as temperature and moisture, with warmer temperatures and wetter conditions leading to increased concentrations of fungi indoors. Emission rates of fungal spores can also be influenced by mechanical ventilation and vibration or disturbance by air currents as well as humidity. The abundance of dust mites in dust correlates with higher indoor humidity levels (D’Amato et al., 2020). Cockroaches, rodents, dust mites, and pets are known sources of allergens especially for individuals with asthma and in lower-income, urban homes (Ahluwalia et al., 2013; Eggleston, 2017). Exposure to dust mites and endotoxins in farming communities is also associated with shaping the innate immune response and conveying protective effects especially in terms of lowering asthma prevalence (Stein et al., 2016). Outdoor plant-based allergen concentrations such as weed, grass, or tree allergens can be driven by seasonal patterns, and these temporal patterns may also vary geographically (e.g., peak concentration season varies in timing, duration, and intensity by region) (Schmidt, 2016). Spatial factors, such as proximity to farming facilities and rural versus urban environments, are associated with greater concentrations of endotoxins in indoor dust (Barnig et al., 2013). Ventilation-related factors, such as window-opening, central air conditioning, and particle filtration, can significantly influence infiltration rates of these biologicals indoors. New analytical techniques allow the quantification of microbial VOC (mVOC) emissions in laboratory settings as a function of microbial respiration, life stage, and substrate (Misztal et al., 2018). Recent efforts have demonstrated mVOC composition from sinks and showers may be more stable than the underlying microbial communities, which can be highly variable in space and time in response to environmental conditions (Adams et al., 2017). The need to quantify the diversity of microbial species that influence indoor air across a variety of indoor environments and materials remains (Bekö et al., 2020; Prussin and Marr, 2015).

Infiltration from interstitial and buffer spaces within a residence, or from neighboring units in multi-unit residential or occupational settings, can be considered a continuous source of chemicals indoors. Emission rates in these spaces depend on temperature gradients between outdoor and indoor surfaces as well as proper sealing and mechanical ventilation. Attached garages (or parking structures in occupational settings) have been shown to introduce hydrocarbons and other pollutants in indoor air. Emission rates also depend on building ventilation and design, air change rates, and the relative position of air intakes to these sources (spatial factors). Wind shear and vertical temperature and pressure gradients create chimney or stack effects within multi-unit apartment buildings, especially high rises, and can also play a role in introducing chemicals indoors from underground or outdoor sources, transferring chemicals within and across units, and accumulating chemicals in elevated floor levels (Man et al., 2019). Connected central ventilation systems and structures can also act as sources of chemicals infiltrating from nearby residences, with secondhand environmental tobacco smoke being a well-known example (Snyder et al., 2016). Not all chemicals in indoor air come from visible spaces. Attics, crawl spaces, and dry wall cavities have been found to contain VOCs emitted from insulation materials, paint, and fungal growth into indoor spaces. Most building materials are porous to some extent, allowing the transport of VOCs through the materials (Xu and Zhang, 2011). Building envelope materials can be sources of chemicals that migrate into the indoor air. Aliphatic alkanes only found in a building envelope air barrier have been measured in the indoor air of a house (Poppendieck et al., 2015). The extent to which VOCs migrate through

building materials depends on the material, environmental factors, and the individual chemical. Studies investigating these three factors have been largely limited to wallboard, relative humidity and/or temperature, and a limited number of chemicals (e.g., formaldehyde) (Tran Le et al., 2021).

Outdoor air is a continuous source of particulate matter (PM), ozone, VOCs and SVOCs, and other chemicals. Entry rates depend on natural and mechanical ventilation by opening windows and doors, operating HVAC systems that bring in outdoor air, or infiltrating through small cracks and openings in building structures that, taken together, determine air change rates. Important spatial and temporal factors, including the proximity of the building air intake to outdoor sources of these chemicals (e.g., to street-level car exhaust) and the building’s location relative to nearby outdoor sources (e.g., roads and industrial facilities), also influence emission rates. Spatial gradients or patterns can also vary at regional or global scales (e.g., wildfire smoke, ozone, dust storms). Temporal factors influencing rates of chemicals infiltrating from outdoors into indoor environments include diurnal or weekday/weekend patterns (e.g., traffic-related nitrogen oxides [NO_x] and ozone) and seasonal patterns (e.g., ozone or allergens). Climate change–related factors influencing infiltration of outdoor chemicals (e.g., increased wildfires, increased temperatures) could also operate at longer timescales of years and decades (Field, 2010; IOM, 2011; Lee et al., 2017; Liang et al., 2021).

The chemical composition and size distribution of aerosols changes as particles move from outdoors to indoors. During cool outdoor weather, chemicals on outdoor particles moving into warmer indoor spaces volatilize into the gas phase. During warmer outdoor weather, the opposite occurs, and indoor chemicals partition onto aerosols of outdoor origin (Avery et al., 2019b; Cummings et al., 2021). Recent measurements have shown that variation in partitioning onto outdoor aerosols migrating indoors is dependent upon the aerosol water liquid fraction, which is a function of temperature and relative humidity. Neglecting this phenomenon can result in errors on the order of 40 percent to 400 percent when modeling the chemical composition of aerosols (Cummings et al., 2021).

Intrusion of chemical vapors from underground sources is also a continuous source with significant spatial and geographic variability influencing its entry rates. Pesticides, solvents, chemicals from leaky underground storage tanks, contaminated soil and groundwater, and radon gas are some of the sources falling under this category. Examples include VOCs like toluene, trichloroethylene (TCE), perchloroethylene, formaldehyde, and petroleum products (Provoost et al., 2011). Temperature gradients, vertical space, and building design are important factors influencing emission rates (Ma et al., 2020; Unnithan et al., 2021). Stack effects in buildings can also enhance the intrusion rates of these chemicals indoors, depending on building envelope tightness and weatherization (Francisco et al., 2020; Pigg et al., 2018).

Chlorinated solvents, such as TCE, are increasingly recognized as potentially hazardous chemicals that can enter the indoor environment through vapor intrusion (Forand et al., 2012). In 2021, California issued new draft vapor screening guidance in an effort to provide more consistency in evaluating exposure and health risks despite the challenges of significant spatial and temporal variability in subsurface and indoor air concentrations. The lack of data and high degree of variability in space and time of both sources and concentrations in indoor air has been described as a challenge by both the U.S. Environmental Protection Agency (EPA) and California guidance documents and in the literature (Johnston and Gibson, 2015).

Radon gas concentrations are usually highest in underground levels and basements, with greater soil permeability and fractures in foundations of homes increasing intrusion rates indoors (Kellenbenz and Shakya, 2021). Although inconclusive, some studies have hypothesized that increased building weatherization and modern construction could potentially lead to higher radon exposure in North American indoor residential environments, despite wide geographic variability in the geologic distribution of radon and its precursor minerals, like uranium, thorium, and radium

(Stanley et al., 2019). Naturally present radon gas (²²²Rn) and its decay products, like lead and polonium, are widely recognized as an indoor air concern and a risk factor for lung cancer (Darby et al., 2005; Krewski et al., 2006). Recent work has shown that ²²²Rn progeny in the form of radioactive nuclides can react with water vapor and gases to form clusters. These clusters attach onto particle surfaces, emit high energy radiation, and get transported into the lungs and human body with inhaled PM, which essentially acts as a delivery vector (Liu et al., 2019). Several recent studies reported that radon gas exposure as a surrogate of gross α-, β-, and γ- particle radioactivity, or directly measured particle radioactivity on filters (Liu et al., 2020), might potentially explain differential toxicity of PM_2.5 seen in health studies (Blomberg et al., 2019; Huang et al., 2020, 2021; Papatheodorou et al., 2021).

Episodic Sources

Humans and human activities, in terms of occupancy and behaviors, are an important episodic source of indoor chemicals. This category encompasses several classes of compounds in gas- and particle-phase emission processes, given the complexity of human activities and their interactions with the indoor environment. They are considered episodic, since emissions only occur when humans are present or performing certain activities. Types of sources falling under this category include metabolites excreted from skin (e.g., ammonia; see Li et al., 2020) and breath (e.g., carbon dioxide [CO₂], VOCs; see Tang et al., 2015) or other human excretions; desquamation and hair; microbes (bacteria and viruses; see Yang et al., 2021); generation of chemicals through activities such as cooking, smoking (Wallace and Ott, 2011), and burning candles (Afshari and Ekberg, 2005); use or manufacture of illicit drugs (Kuhn et al., 2019); use of disinfectants and cleaning products (Afshari and Ekberg, 2005); use of personal care products (e.g., fragrances and lotions); application of pesticides and insecticides; use of air fresheners; operation of HVAC systems; and the act of vacuuming and moving around that might resuspend or re-entrain settled dust, its components, and larger PM size fractions (e.g., PM₁₀) in indoor spaces.

Next, some of the major sources that fall under this category are discussed, recognizing that humans and human activities as an episodic source is too broad of a topic to cover comprehensively here. For example, cooking emits pollutants based on the fuel source, as well as the range of activities that contribute to food preparation, including CO, CO₂, NO/NO₂ (gas combustion or other high temperature combustion), organics, ultrafine particles (UFPs), and fine particles (Bhangar et al., 2011; Patel et al., 2020; Wallace et al., 2004; Wallace and Ott, 2011). Emission rates depend on fuel type (Wallace et al., 2008), cooking method (e.g., frying emits more UFPs than boiling or steaming), food types (e.g., oils or meats), temperature, duration of time, and cooking surface materials. Emission rates can exhibit diurnal or day of the week patterns that mimic typical human time-activity patterns (e.g., meal time peaks). Consumer products (e.g., cleaners, disinfectants, pesticide applications on pets or in the home/garden, and room air fresheners) and personal care products introduce chemicals to indoor environments episodically during and following their use. Near-field exposure depends on product formulations, which chemicals are released, how much, where the chemicals reside (e.g., dust or air), and human presence and activity level. Humans can also introduce “take-home” chemicals into the indoor environment on clothes—for example, lead, asbestos, and pesticides from occupational settings (Kalweit et al., 2020). A recent study in a German movie theater found that third-hand smoke is an important source of exposure in nonsmoking indoor environments, where VOCs, hazardous air pollutants (e.g., benzene and formaldehyde), and nicotine can off-gas from clothing of moviegoers and partition onto aerosols and dust (Sheu et al., 2020). Temporal factors (time of day and day of the week patterns corresponding to work schedules) and environmental factors (temperature, relative humidity) influence these emission rates. Home renovations and repairs can introduce new building materials and cause

novel emissions owing to the materials used. Considerations for emissions are similar to building materials as described above for continuous sources. HVAC operation may appear like a periodic source impacting indoor air chemical composition. Dust coated heating coils can act as a reservoir of SVOCs, taking up and releasing them as a function of temperature or moisture. Water condensing on cooling coils or other parts of the HVAC system can take up and release soluble gases. Wet surfaces in the HVAC system may also promote biological activity and thus act as a primary source of mVOCs or promote chemical transformations that take up some and release other chemicals. Temporal factors, like time of day, and environmental factors, like weather, influence these emission rates.

The personal cloud effect is a term referring to increased concentrations of chemicals found near or around humans in their immediate surroundings. The personal cloud is the result of several factors that include, but are not limited to, bioeffluents, personal care products, reactions with skin and clothes, and the thermal plume (air moved upward over the torso as adults release about 100 W of heat). Hence, human presence is a direct episodic source of indoor chemicals. While the personal cloud effect is well documented (Licina et al., 2017; Wallace, 2000; Weschler, 2016), several recent studies have characterized emission rates of human activities in indoor environments, including the effects of a crawling infant on dust resuspension (Hyytiäinen et al., 2018; Wu et al., 2021a) and the movement and heat from sleeping resulting in VOCs and SVOCs being released from the bedding environment (Boor et al., 2014, 2017). Recent work also shows that SVOCs adsorbed to the surfaces of cooking pans and burner coils can desorb and form UFPs when heated (Wallace et al., 2008, 2017).

Pets also emit chemicals into the indoor environment, though pets are considered a less significant episodic indoor source than humans. These emissions include biologicals (e.g., metabolites in breath or urine, dander, microbes) emitted directly from pets, chemicals applied to pets (e.g., flea and tick treatments), or chemicals brought indoors from the outdoor environment (e.g., allergens, pollen, metals, pesticides) on pet hair or feet, especially for animals that spend time indoors and outdoors. Pets and human occupants are known sources of ammonia (NH₃) indoors (Ampollini et al., 2019; Ito et al., 1998), with one recent study showing a correlation between presence or number of pets indoors and NH₃ concentrations, especially with warmer temperatures (Uchiyama et al., 2015). Cat litter boxes or other forms of indoor pet waste storage can be important sources. Pets’ urine contains urea, which decomposes anaerobically to release ammonia. Very few studies have measured emission rates of NH₃ from pets indoors; however, these are now more feasible with recent developments in on-line ammonia analyzers. Agricultural environments and extreme environments (e.g., hoarding) can also have high concentrations of NH₃.

Water-sourced chemicals released from domestic water use indoors in gas or aerosol form include brominated, chlorinated, and nitrogenated compounds; mineral aerosols; SVOCs; and biologicals. Chlorine and chloramines are disinfectants added during drinking water treatment. They can react with natural organic material to form trihalomethanes, which can be released into indoor air via showers, dishwashers, and washing machines (Gordon et al., 2006). Detergents used in dishwashers can release hypochlorous acid (Dawe et al., 2019) and react with organics to release further chemicals, including chloroform (Olson and Corsi, 2004). In addition, water delivery systems and source waters can become contaminated, delivering chemicals to the indoor environment, especially after forest fires (Omur-Ozbek et al., 2016; Proctor et al., 2020). Depending on geography and source water quality and contamination levels, emission rates can vary significantly between regions. Seasonal trends can also be at play (e.g., concentration of chemicals and dilution rates will depend on seasonal rain, area discharges and runoff, and surrounding land uses, such as agricultural spraying).

Recent studies have reported elevated concentrations of chemicals such as trace metals, elements, and VOCs in water supply systems following wildfires. These chemicals have been detected

in water bodies and water delivery systems in the immediate vicinity of wildfire-impacted areas, as well as further downstream, depending on the environmental transport pathways involved (Burton et al., 2016; Tecle and Neary, 2015). Chemicals detected in water systems post-wildfire also differ depending on the fuels involved, with wildland fires releasing lead, iron, cadmium, mercury, and other metals that have potentially deposited in mature trees over decades. Fires in the wildland-urban interface, on the other hand, can involve furniture, building materials, plastics, and other man-made products, releasing chemicals such as pyrolyzed organic carbon, benzene, and other new and potentially unknown chemicals (Bladon et al., 2014; Proctor et al., 2020). In turn, these chemicals can be aerosolized or volatilized after making their way indoors via the drinking water supply, influencing indoor chemistry in ways that are yet to be fully understood.

CLASSES OF COMPOUNDS IN INDOOR ENVIRONMENTS

While some chemicals are well characterized and studied in indoor environments from the perspective of understanding exposure (e.g., radon, ozone, CO, NO₂, lead, and some VOCs), other chemical classes have only recently been identified in the indoor environment (e.g., disperse azo dyes and ultraviolet [UV] stabilizers). Tables 2-1 and 2-2 present a summary of several chemicals and chemical classes that have primary sources and reservoirs in the indoor environment, in the gas, dust (Table 2-1), and aerosol (Table 2-2) phases. This is not an exhaustive or complete list but rather a list to highlight examples of some of the more common or novel chemicals that stem from sources and/or reservoirs and are of concern for the indoor environment.

It is also important to note that some of the chemicals listed in the tables can have both primary and secondary sources, and the concentration ranges listed may reflect both primary and secondary sources. For instance, the primary source of hypochlorous acid is cleaning solutions, which can directly release, or in some cases react with the indoor environment to produce, chloroform, carbon tetrachloride, chloramines, nitrogen trichloride, and chlorine gas (Mattila et al., 2020; Odabasi, 2008; Wong et al., 2017). Surface reactions of nitrogen dioxide (NO₂) are also considered important secondary sources of nitrous acid (HONO) indoors (Collins et al., 2018).

As shown by the information presented in Table 2-1, concentrations of some chemicals are abundant in indoor air, while others are abundant on surfaces and in indoor dust and airborne particles (refer to Table 2-2 for aerosol phase). These differences are attributable to their physicochemical properties and use patterns. In particular, chemicals with high vapor pressures are often more abundant in the gas phase of indoor air, and chemicals with lower vapor pressures (and thus higher log₁₀(K_oa) values) are more abundant in the condensed phase. The wide range in concentrations within many of these chemical classes is evident, even though most of the reported values are from time-averaged sampling of continuous sources. These concentration differences are related to a number of factors and indicate the high degree of variability observed in indoor environments.

Aerosols

An aerosol is defined as a stable suspension of solid or liquid particles in air. Aerosol particles, which are also known as PM, contain a variety of organic and inorganic chemicals (NASEM, 2016). PM also exhibits a wide size distribution in indoor environments, and some of its components, especially SVOCs, may shift from the particle phase to the gas phase and vice versa as they move indoors from outdoor environments or depending on changes in indoor environmental conditions.

While airborne PM occurs indoors in a wide range of sizes, with aerodynamic diameters ranging from a few nanometers (nm) to ~100 micrometers (μm), referred to as Total Suspended Particles, this section will focus on three size fractions of highest relevance to indoor aerosol exposures and associated health risks. For a more detailed discussion of indoor PM chemical composition,

size distributions, and health risks, refer to Health Risks of Indoor Exposure to Particulate Matter: A Workshop Summary (NASEM, 2016).

These size fractions are defined as follows:

• UFPs are most commonly defined as particles with an aerodynamic diameter ≤100 nm. However, the definition of UFPs can vary across fields and disciplines, with some referring to particles ≤10 nm (e.g., emissions standards) and others to particles ≤200 nm (e.g., health disciplines) as UFPs. These are emitted from primary combustion or formed by conversion indoors. They are similarly formed outdoors and can be transported indoors via infiltration and ventilation. Given their negligible mass, UFPs are usually reported in particle number concentrations (particles/cm³). They span a wide size range and tend to exhibit diffusion-driven behavior in the lungs especially in the smallest size range (<10 nm) (Oberdörster, 2001; Oberdörster and Utell, 2002). Larger UFPs (generally >10 nm) have a much higher surface-area-to-mass ratio compared to larger particles and are thus efficient at transporting chemicals adsorbed onto their surfaces into the alveolar region of the lungs. UFPs are also very efficient at crossing the alveolar epithelial tissue into systemic circulation and reaching other organs (Schraufnagel, 2020). Computational fluid dynamics simulations demonstrate that very small nanoparticles (1-2 nm) will deposit in the olfactory region of the human nose to a larger extent than larger UFPs (>10 nm) (Garcia et al., 2015). Some nanoparticles deposited in the olfactory region during inhalation might undergo direct transport along the olfactory nerve to the olfactory bulb (Elder et al., 2006) raising concern about possible neurological effects (Lucchini et al., 2012).
• Fine particles (PM_2.5), or particles with an aerodynamic diameter ≤2.5 μm, are usually emitted from primary sources like combustion and secondary formation indoors and reported in mass concentrations (μg/m³). If inhaled, PM_2.5 deposits in the tracheobronchial and alveolar regions of the lungs and has been associated with a multitude of adverse health outcomes (Dominici et al., 2006; Pope and Dockery, 2006).
• PM₁₀, or particles ≤10 μm in aerodynamic diameter, are typically primary in nature, emitted from mechanical abrasion, natural sources (e.g., sea salt), and dust resuspension and are reported in μg/m³. Notably, PM₁₀ is considered a subset of larger particles often referred to as “super-coarse” which can reach up to 100 μm in aerodynamic diameter. Coarse PM, typically defined as PM between 2.5 and 10 μm in aerodynamic diameter, primarily deposits in the upper nasopharyngeal region of the lungs and is cleared with mucociliary transport.

Organic compounds as a class usually make up a significant portion (~40–70 percent) of the mass concentration of indoor PM_2.5 (Habre et al., 2014; Polidori et al., 2006; Turpin et al., 2007). Similarly, particulate sulfates, nitrates, ammonium, and water can be present in larger concentrations, as compared to elemental carbon (EC), trace elements, and metals such as nickel, vanadium, and iron, which are usually present at lower concentrations (NASEM, 2016).

The composition of PM will vary based on the source or process generating it (Habre et al., 2014). It is important to note that PM_2.5 in itself is considered a mixture, and the concentration of species within PM_2.5 will depend on the major sources contributing to indoor PM_2.5 concentrations, including those of indoor and outdoor origin, which also vary based on spatial, temporal, and other factors, as described in the next section. For example, concentrations of lead and arsenic in indoor PM_2.5 might be elevated in residences or schools near battery recycling and lead smelter facilities (Meyer et al., 1999). Similarly, concentrations of EC might be higher inside homes that burn solid fuels (e.g., coal or wood) for heating or cooking. Indoor concentrations of UFPs in homes located within 100 m of a major interstate highway in Massachusetts were higher than in homes

>1,000 m away (Fuller et al., 2013). See Chapter 6 for a discussion of the settings and determinants that correlate with higher exposure to indoor PM.

Table 2-2 contains typical mass (or number, in the case of UFPs) concentrations of indoor aerosols across size fractions and typical concentrations of organic and elemental components of PM_2.5 as a more well-studied example of indoor aerosols. Elements and metals are grouped together where they are known to originate from a specific source or source category. Where available, the reported concentrations are for daily, 48-hour, or weekly time-averaged or integrated measurements primarily from studies of real indoor environments (versus laboratory or chamber experiments) and do not capture peak concentrations or episodic sources. References from the House Observations of Microbial and Environmental Chemistry (HOMEChem) study, which simulated a realistic indoor environment, are the one exception to this rule. These are also examples meant to illustrate the wide range of variability in indoor concentrations reported in the literature. They include data from residential, office, and school settings, and from the United States and international studies, but they are not meant to be comprehensive or representative of all types of indoor environments, geographies, or time periods.

As with Table 2-1, many of these aerosols and their components have both primary and secondary sources indoors; however, Table 2-2 does not include secondary organic aerosol (SOA). Formation of SOA indoors is covered in more detail in Chapter 4 of this report.

INDOOR CHEMICAL INVENTORIES

EPA’s Toxic Substances Control Act (TSCA) Chemical Substance Inventory contains more than 33,000 active (and more than 34,000 inactive) registered chemicals for use in the United States (EPA, n.d.). A recent study attempted to catalogue the regional and national chemical inventories for the first time (Wang et al., 2020). An interesting result from this study was that of the ~350,000 chemicals in the inventory, approximately 50,000 were classified as confidential, and about 20 percent (~70,000) were ambiguously described. Global chemical inventories for plastic production alone exceed 10,000 different chemicals (Wiesinger et al., 2021).

Thousands of chemicals are present in indoor environments that potentially can be released into indoor air. Inventorying chemicals and their concentrations in the indoor environment enables more accurate chemical exposure estimates, including for chemicals whose indoor source may be unknown. Cataloguing all the chemicals present in indoor air can theoretically be addressed by two major approaches: bottom-up inventories, which determine the chemical composition and/or emissions of every indoor object; or top-down inventories, which measure the total chemical composition of the gas, particle, dust, and surface phases.

Bottom-Up Inventories of Chemicals in Indoor Environments

A bottom-up approach determines the chemical composition of or emission from every material, item, or product found indoors, including the wide range of examples found in this chapter thus far. Two drivers currently exist for measuring emissions from building materials in a bottom-up manner. First, there are regulatory limits for the emission of certain chemicals in some countries—for example, EPA’s Formaldehyde Emission Standards for Composite Wood Products Act, which limits formaldehyde emission from composite wood products. Second, there are building product certification programs that can be regulatory or voluntary—for example, GREENGUARD, Blue Angel, AgBB, natureplus, and EU Eco Flower. However, these schemes are generally limited to building products, focused on more volatile chemicals, and tested under only one set of chamber parameters (e.g., ASTM International, 2016a), and they can limit reporting to summarizing parameters such as total VOC that lack the detail needed to fully inform indoor chemistry.

shape and density. Direct optical sensors are most accurate when particle diameters are greater than the light wavelength (~100 nm to 300 nm). Systems that use particle growth can quantify particles down to 1 nm in diameter. These methods do not provide information about chemical composition.

Most commonly, PM needs to be captured on a filter first to determine its chemical constituents. Active or passive sampling methods have been employed to collect particles on filters for gravimetric and compositional analysis, with active size-selective samplers often used to collect 50 percent of particles below a specific cut-point in aerodynamic diameter, assuming spherical shape. The filter is extracted with a solvent, and that extract is then treated with a variety of techniques to purify and isolate chemical classes of concern (e.g., water-soluble metals). The final extract is then often evaporated, reconstituted, and analyzed using a variety of mass spectrometry techniques. Non-destructive methods such as X-ray fluorescence or thermal/optical reflectance/transmittance have also been used to characterize elemental composition and elemental versus organic carbon content based on thermal optical properties, respectively, directly on the filter media. Destructive methods like inductively coupled plasma-mass spectrometry have also been used to characterize chemical composition, and many other methods (e.g., radioactivity and scanning electron microscopy) have been used for morphological characterization and counting. However, these techniques only can capture chemical composition of stable particles. Volatilization (e.g., of low molecular weight organics and nitrates) and reactions (e.g., with ozone and nitrogen dioxide) can occur during sampling and storage prior to analysis (ASTM International, 2013). Chemicals are only analyzed if they are in the size range captured by the sampler.

New Analytical Approaches

One new analytical approach for PM sampling is the use of HVAC filters as a collection media for SVOCs associated with particles. HVAC systems move large volumes of air through filters, creating a lengthy, time-integrated sample of airborne particles. Dust on HVAC filters has been shown to have equivalent or higher concentrations of polybrominated diphenyl ethers (PBDEs), phthalates, and organophosphates than settled dust (Bi et al., 2018; He et al., 2016; Xu et al., 2015).

A wide range of instruments can be applied to determine continuous particle composition in outdoor PM, including mass spectrometry, chromatography, and particle-into-liquid sampling coupled to on- or off-line analysis (Farmer, 2019). To date, however, the limited indoor particle studies have relied on aerosol mass spectrometers (AMS) and thermal desorption aerosol gas chromatography (TAG). AMS instruments can provide elemental composition and oxygen/carbon and hydrogen/carbon ratio information (Farmer, 2019). PMF and other source apportionment models have also been used with chemical composition data to identify and apportion sources that contribute to indoor PM concentrations (Habre et al., 2014; Hasheminassab et al., 2014; Molnár et al., 2014). Using PMF, AMS data have been used to examine how particles change as they transport through the building envelope during varying seasons (Avery et al., 2019a,b) and change after cooking (Katz et al., 2021a), as well as the impact of particles on third-hand smoke transport in indoor air (DeCarlo et al., 2018). Yet, AMS is limited to analysis of non-refractory PM smaller than 1 micron; larger particles, metals, soot, and salts are not analyzed by AMS. TAG has been applied in residential studies to measure the dynamics and evolution of speciated organic chemicals under natural ventilation conditions (Fortenberry et al., 2019). Semivolatile TAG (SV-TAG) has been applied to study residential sources, dynamics, and gas/particle partitioning of SVOCs on an hourly basis (Kristensen et al., 2019), including analysis of phthalate partitioning dynamics (Lunderberg et al., 2019), and the role of surface emissions and re-partitioning of SVOCs to PM_2.5 (Lunderberg et al., 2020). AMS and SV-TAG have been used together to study emissions and partitioning of siloxanes during cooking events (Katz et al., 2021b). Due to cost and complexity, the application of AMS, TAG, and similar instruments to the indoor environment has been limited. Other continuous particle composition

techniques have not been widely applied to the indoor environment (Farmer, 2019). These include mass spectrometry using soft ionization, extractive electrospray techniques, coupling rapid filter collection with CIMS, and single-particle techniques. Like for gas-phase sampling, facilitating the greater application of these instruments and methods to a wide range of indoor spaces and environmental conditions could make top-down approaches to inventorying indoor aerosols more robust.

Settled Dust Sampling

Traditional Techniques

The collection and analysis of dust has been occurring for decades and has been important to understanding this reservoir of indoor chemicals. Identifying and measuring chemical burdens in dust samples has provided insight into potential exposure to occupants, particularly as it relates to estimating exposure from hand-to-mouth activities, which is an important pathway not only for children but also for adults. Dust is often collected using a vacuum, although some studies have reported using brooms and brushes to collect settled dust in high-traffic areas. In some cases, clean sampling films are left in a horizontal position (e.g., on a bookshelf or countertop) to passively collect airborne dust that settles over time. The mass accumulating on these films can be quantified, and the films can be extracted and analyzed to characterize their chemical composition. Additionally, dust is often extracted and analyzed in a similar fashion to PM on filters.

Unfortunately, there is no standardization of sampling methods for dust, and the various methods currently used can lead to uncertainty in measurements and comparability among studies. For example, differences in the types of vacuums used by researchers vary considerably and are not standardized. EPA has recommended standardized methods and equipment such as the High Volume Surface Sampler; however, this instrument is heavy and expensive compared to commercial vacuums. Researchers have a tendency to use dust collected either by participants using their home vacuum cleaner or with a smaller vacuum or a handheld device that can be used by researchers from home to home. Differences in vacuums can potentially lead to differences in the size and type of dust particles collected and thus differences in the chemical concentrations reported in dust. Dust often needs to be sieved to remove large particles before chemical analysis; however, researchers typically use different sized sieves that result in different sized particles included in the final analyses and further contribute to differences in measurements across studies.

In addition, differences have been found based on the location (room) in the house where the dust is collected, or between settled dust on the floor and elevated dust (e.g., on a bookshelf). Factors that often influence chemical measurements of SVOCs in dust samples were recently highlighted in a meta-analysis in which flame retardants were used as a case example (Al-Omran et al., 2021). Box 2-3 discusses challenges related to assessing exposure to chemicals in dust.

New Analytical Approaches

Moving forward, it will be important to more thoroughly identify and characterize chemicals present in dust samples. Non-targeted approaches that utilize high-resolution mass spectrometry to identify unknown chemical features or responses are now being used to characterize dust samples (Ouyang et al., 2017; Rager et al., 2016; Rostkowski et al., 2019). For example, a recent study used a non-targeted method to identify brominated azo disperse dyes in house dust (Peng et al., 2016). These methods provide more insight into the complex and diverse suite of chemicals that are present in indoor environments, yet they are not currently at a point where every chemical in the thousands in a mass spectrum of an indoor air or dust sample can be identified—let alone quantified. Chemical space is vast, and the number of chemicals that are on the market and their potential

degradation products span tens of thousands of chemicals (EPA, n.d.). Investments in chemoinformatic approaches and non-targeted analyses may improve our ability to fully characterize complex chemical mixtures in indoor dust in the future.

Surface Sampling

Traditional Techniques

Surfaces in indoor environments, including windows, countertops, and walls, accumulate organic films and act as sinks, reaction sites, and reservoirs for SVOCs (Weschler and Nazaroff, 2008). The Occupational Safety and Health Administration’s Guidelines for Surface Sampling Methods are available and widely utilized (OSHA, 2007). Surfaces have traditionally been sampled by taking a clean wipe, often made of cotton or a similar material, and wiping down a specific surface area. Surface wipe samples can then be extracted and analyzed in a similar fashion to PM on filters and may experience similar issues like loss of volatile chemical species (Liu et al., 2003).

Growth of organic films on indoor surfaces has been modeled theoretically (Weschler and Nazaroff, 2017). In addition, many studies have been done to understand how specific chemicals react with different surfaces, including ozone oxidation reactions of organic chemicals present on surfaces (Shu and Morrison, 2011; Weschler and Nazaroff, 2017). Indoor surface films are also studied by placing clean substrates or coupons (made of materials commonly found indoors, such as glass, metal, or tile) indoors and allowing films to grow, followed by analysis in the laboratory by surface chemistry or extraction techniques.

New Analytical Approaches

A variety of new surface sampling techniques have been applied in the past decade. Some analytical advancements have focused on improving the chemical characterization of surface wipes to evaluate chemical residues using spectrophotometer techniques (Deming and Ziemann, 2020). Surface chemistry is also being probed by applying portable surface reactors to organic chemical reactions on surfaces (Algrim et al., 2020), flux chambers for emissions from surface films on

coupons (Adams et al., 2017) or surfaces (Wu et al., 2016), atomic force microscopy coupled to infrared spectroscopy to measure film coverage and functional group analysis (Or et al., 2018), and analysis of glass collection plates using Fourier Transform Ion Cyclotron Resonance Mass Spectrometry and off-line AMS (O’Brien et al., 2021). Application of more standard, advanced, and emerging techniques for surface chemistry analysis to relevant indoor surfaces holds significant promise for advancing understanding of surface films and their emissions and chemical interactions.

CONCLUSIONS

Chemical sources in the indoor environment are numerous and varied. Potential human exposure to these chemicals is dynamic owing to the complexities inherent not only in these sources but also in the environment and human behaviors. Chemical emissions and abundances are linked to both chronic and episodic sources and are modified by behavioral, environmental, and physical factors. Building materials, consumer products, infiltration of outdoor air, and human behavior (particularly cooking) can strongly influence the chemicals detected in the indoor environment. Dust and indoor surfaces also serve as a sink and reservoir of chemicals that are slowly released from primary sources, and they play a role in understanding human exposure, which is discussed further in Chapter 6. New research also suggests that clothing mediates our exposure to chemicals in the indoor environment, acting as a sink for chemicals present in indoor air (Licina et al., 2019).

Over the past few decades, new materials containing unique chemicals have emerged for use in the indoor environment—for example, new types of building insulation and greater use of electronics and smart devices in indoor environments, particularly in homes. These devices and materials can also be sources of chemicals to the indoor environment, such as plastics, plasticizers, antioxidants, UV stabilizers, and flame retardants. Recycling of materials could lead to increased exposure to hazardous chemicals that have been phased out due to concerns about human exposure. Increased use of cleaning and disinfection agents, particularly during the COVID-19 pandemic, is leading to an increase in the levels of some chemicals, such as quaternary ammonium compounds (Zheng et al., 2020), in the indoor environment. Little information is available on the health effects or distribution of these chemicals in the indoor environment. Over time, some chemicals that were emitted by primary sources in the indoor environment have been phased out of use owing to concerns about their elevated exposure and/or toxicity (e.g., PBDEs, benzyl butyl phthalate). These unfortunate chemical substitutions are often replaced with chemicals that have less data available on their emissions, exposure, and potential hazards.

A wide range of analytical techniques are currently being used to identify new chemicals, in both bottom-up and top-down approaches, that may be released into the indoor environment, but these approaches are costly and time-consuming. The lack of transparency in chemical use (i.e., confidential business information) and challenges in identifying chemical sources in the indoor environment will continue to be major obstacles to chemical management and risk evaluation.

RESEARCH NEEDS

Given its findings about the current state of the science, the committee has identified priority research areas to help drive future advances in understanding primary sources and reservoirs in indoor environments:

• Prioritize acquisition of actionable data and research to link sources with exposures and understand impacts of mixtures on health. Research is needed that provides greater resolution on spatial and temporal trends in chemical emissions indoors, and how these vary for both chronic and episodic sources. Specific needs include improved understanding of the combined influence of ventilation rates, humidity, and temperature on chemical

• emissions indoors, as well as the interaction of the indoor and outdoor environment and its influence on indoor sources and exposures, especially for SVOCs. Better understanding of sources would provide more actionable information to control sources and, consequently, exposures. In addition, the indoor environment, where humans spend the majority of their time, contains a highly complex chemical mixture found in multiple phases and originating from many sources.
• Increase transparency in chemical applications/use in building materials to minimize time and effort needed to establish evidence of exposure and health risks. New or improved analytical methods have helped to identify new chemicals in indoor environments and their transformation products. However, every year new materials are introduced to the market that may become new sources (new types of flooring, new surface treatments for water and grease repellency, etc.). More research is needed to evaluate chemical emissions, and thus potential human exposure, from new materials before they are broadly introduced to the market. Greater transparency would expedite the process of establishing evidence for potential harmful exposures.
• Improve analytical methods and non-targeted approaches to support discovery. In the past, a number of chemicals of human health concern have been identified in indoor settings using targeted or screening methods. Use of non-targeted methods has the potential to identify chemicals of concern earlier. Thus, the committee recommends that greater emphasis is placed on improving non-targeted approaches that utilize high-resolution mass spectrometry to identify previously unknown chemicals and characterize the complex chemical mixtures present in the indoor environment.
• Develop and maintain harmonized chemical information databases. Information databases now exist for a broad range of volatile chemical products and other consumer products (e.g., from SDSs and ingredient labels, and the Chemical and Products Database [CPDat]). These can support the use of non-targeted methods that are key to improving our understanding. However, maintaining and updating these databases requires resources. Those who develop and maintain databases (i.e., regulatory bodies, scientific research groups, or consumer and health protection agencies) would benefit from harmonization of terminology, tracking of data provenance, consistent identification of standard product categories, and consistent processes for data updates.
• Expand research further into nonresidential settings and underrepresented countries and contexts. The vast majority of the information presented in this chapter refers to research conducted in residential settings and in high-income countries. However, people spend time in other indoor environments, which may have different sources and emission rates. For example, less information is available on chemical sources in schools, hospitals, government buildings, retail stores, restaurants, and occupational and recreational buildings (hair and nail salons, rock-climbing gyms, trampoline parks, indoor pools, etc.) across a wide range of socioeconomic strata (Mandin et al., 2017). In addition, most research is conducted in high-income countries or in low-income countries with households that rely on solid fuels for indoor cooking and heating (typically termed “household air pollution” by the World Health Organization). Populations living in low- to middle-income countries likely experience widely variable chemistries indoors due to differences in regional activities, human behaviors, climates, fuels used in cooking and heating, and the use of different materials in building construction. Similarly, significant variability exists among developed countries based on social and economic factors (e.g., public housing, lower-income residences, or disadvantaged neighborhoods). Researchers could consider fostering community-based partnerships to increase the amount of information available and better incorporate local knowledge and context when investigating these understudied environments.

homes and risk of lung cancer: Collaborative analysis of individual data from 13 European case-control studies. BMJ 330:223. https://doi.org/10.1136/bmj.38308.477650.63.

Dawe, K. E. R., T. C. Furlani, S. F. Kowal, T. F. Kahan, T. C. VandenBoer, and C. J. Young. 2019. Formation and emission of hydrogen chloride in indoor air. Indoor Air 29(1):70–78. https://doi.org/10.1111/ina.12509.

DeCarlo, P. F., A. M. Avery, and M. S. Waring. 2018. Thirdhand smoke uptake to aerosol particles in the indoor environment. Science Advances 4(5):eaap8368. https://doi.org/10.1126/sciadv.aap8368.

Deming, B. L., and P. J. Ziemann. 2020. Quantification of alkenes on indoor surfaces and implications for chemical sources and sinks. Indoor Air 30(5):914–924. https://doi.org/10.1111/ina.12662.

Dhungana, B., H. Peng, S. Kutarna, G. Umbuzeiro, S. Shrestha, J. Liu, P. D. Jones, B. Subedi, J. P. Giesy, and G. P. Cobb. 2019. Abundances and concentrations of brominated azo dyes detected in indoor dust. Environmental Pollution 252:784–793. https://doi.org/10.1016/j.envpol.2019.05.153.

Dominici, F., R. D. Peng, M. L. Bell, L. Pham, A. McDermott, S. L. Zeger, and J. M. Samet. 2006. Fine particulate air pollution and hospital admission for cardiovascular and respiratory diseases. JAMA 295(10):1127–1134. https://doi.org/10.1001/jama.295.10.1127.

Duncan, S. M., S. Tomaz, G. Morrison, M. Webb, J. Atkin, J. D. Surratt, and B. J. Turpin. 2019. Dynamics of residential water soluble organic gases: Insights into sources and sinks. Environmental Science & Technology 53(4):1812–1821. https://doi.org/10.1021/acs.est.8b05852.

Eggleston, P. A. 2017. Cockroach allergy and urban asthma. Journal of Allergy and Clinical Immunology 140(2):389–390. https://doi.org/10.1016/j.jaci.2017.04.033.

Elder, A., R. Gelein, V. Silva, T. Feikert, L. Opanashuk, J. Carter, R. Potter, A. Maynard, Y. Ito, J. Finkelstein, and G. Oberdörster. 2006. Translocation of inhaled ultrafine manganese oxide particles to the central nervous system. Environmental Health Perspectives 114(8):1172–1178. https://doi.org/10.1289/ehp.9030.

EPA (U.S. Environmental Protection Agency). n.d. Toxic Substances Control Act (TSCA) Chemical Substance Inventory. Accessed January 19, 2022. https://www.epa.gov/tsca-inventory.

EPA. 2017. Chapter 5: Soil and dust ingestion. Exposure Factors Handbook. https://www.epa.gov/expobox/exposure-factors-handbook-chapter-5.

EPA. 2021a. Controlling Pollutants and Sources: Indoor Air Quality Design Tools for Schools. https://www.epa.gov/iaq-schools/controlling-pollutants-and-sources-indoor-air-quality-design-tools-schools.

EPA. 2021b. Premanufacture Notices (PMNs) and Significant New Use Notices (SNUNs) Table. https://www.epa.gov/reviewing-new-chemicals-under-toxic-substances-control-act-tsca/premanufacture-notices-pmns-and.

Fan, X., C. Kubwabo, F. Wu, and P. E. Rasmussen. 2019. Analysis of bisphenol a, alkylphenols, and alkylphenol ethoxylates in NIST SRM 2585 and indoor house dust by gas chromatography-tandem mass spectrometry (GC/MS/MS). Journal of AOAC International 102(1):246–254. https://doi.org/10.5740/jaoacint.18-0071.

Farmer, D. K. 2019. Analytical challenges and opportunities for indoor air chemistry field studies. Analytical Chemistry 91(6):3761–3767. https://doi.org/10.1021/acs.analchem.9b00277.

Ferguson, P. L., and H. M. Stapleton. 2017. Comment on “Mutagenic azo dyes, rather than flame retardants, are the predominant brominated compounds in house dust.” Environmental Science & Technology 51(6):3588–3590. https://doi.org/10.1021/acs.est.7b00372.

Ferro, A. R., R. J. Kopperud, and L. M. Hildemann. 2004. Source strengths for indoor human activities that resuspend particulate matter. Environmental Science & Technology 38(6):1759–1764. https://doi.org/10.1021/es0263893.

Field, W. R. 2010. Climate change and indoor air quality. Contractor Report prepared for U.S. Environmental Protection Agency, Office of Radiation and Indoor Air. https://www.epa.gov/sites/default/files/2014-08/documents/field_climate_change_iaq.pdf.

Finewax, Z., D. Pagonis, M. S. Claflin, A. V. Handschy, W. L. Brown, O. Jenks, B. A. Nault, D. A. Day, B. M. Lerner, J. L. Jimenez, P. J. Ziemann, and J. A. de Gouw. 2021. Quantification and source characterization of volatile organic compounds from exercising and application of chlorine-based cleaning products in a university athletic center. Indoor Air 31(5):1323–1339. https://doi.org/10.1111/ina.12781.

Forand, S. P., E. L. Lewis-Michl, and M. I. Gomez. 2012. Adverse birth outcomes and maternal exposure to trichloroethylene and tetrachloroethylene through soil vapor intrusion in New York State. Environmental Health Perspectives 120(4):616–621. https://doi.org/10.1289/ehp.1103884.

Fortenberry, C., M. Walker, A. Dang, A. Loka, G. Date, K. Cysneiros de Carvalho, G. Morrison, and B. Williams. 2019. Analysis of indoor particles and gases and their evolution with natural ventilation. Indoor Air 29(5):761–779. https://doi.org/10.1111/ina.12584.

Francisco, P. W., S. Gloss, J. Wilson, W. Rose, Y. Sun, S. L. Dixon, J. Breysse, E. Tohn, and D. E. Jacobs. 2020. Radon and moisture impacts from interventions integrated with housing energy retrofits. Indoor Air 30(1):147–155. https://doi.org/10.1111/ina.12616.

Fridén, U. E., M. S. McLachlan, and U. Berger. 2011. Chlorinated paraffins in indoor air and dust: concentrations, congener patterns, and human exposure. Environment International 37(7):1169–1174. https://doi.org/10.1016/j.envint.2011.04.002.

E. von Mutius, D. Vercelli, C. Ober, and A. I. Sperling. 2016. Innate immunity and asthma risk in Amish and Hutterite farm children. New England Journal of Medicine 375(5):411–421. https://doi.org/10.1056/NEJMoa1508749.

Tan, H., L. Yang, Y. Huang, L. Tao, and D. Chen. 2021. “Novel” Synthetic Antioxidants in House Dust from Multiple Locations in the Asia-Pacific Region and the United States. Environmental Science & Technology 55(13):8675–8682. https://doi.org/10.1021/acs.est.1c00195.

Tang, X., P. K. Misztal, W. W. Nazaroff, and A. H. Goldstein. 2015. Siloxanes are the most abundant volatile organic compound emitted from engineering students in a classroom. Environmental Science & Technology Letters 2(11):303–307. https://doi.org/10.1021/acs.estlett.5b00256.

Tecle, A., and D. G. Neary. 2015. Water quality impacts of forest fires. Journal of Pollution Effects and Control 3:1–7. https://doi.org/10.4172/2375-4397.1000140.

Tran, T. M., K. O. Abualnaja, A. G. Asimakopoulos, A. Covaci, B. Gevao, B. Johnson-Restrepo, T. A. Kumosani, G. Malarvannan, T. B. Minh, H.-B. Moon, H. Nakata, R. K. Sinha, and K. Kannan. 2015. A survey of cyclic and linear siloxanes in indoor dust and their implications for human exposures in twelve countries. Environment International 78:39–44. https://doi.org/10.1016/j.envint.2015.02.011.

Tran Le, A. D., J. S. Zhang, and Z. Liu. 2021. Impact of humidity on formaldehyde and moisture buffering capacity of porous building material. Journal of Building Engineering 36:102114. https://doi.org/10.1016/j.jobe.2020.102114.

Turpin, B., C. Weisel, M. Morandi, S. Colome, T. Stock, S. Eisenreich, and B. Buckley. 2007. Relationships of Indoor, Outdoor, and Personal Air (RIOPA): Part II. Analyses of concentrations of particulate matter species. Research Report (Health Effects Institute) 130:1–77.

Uchiyama, S., T. Tomizawa, A. Tokoro, M. Aoki, M. Hishiki, T. Yamada, R. Tanaka, H. Sakamoto, T. Yoshida, K. Bekki, Y. Inaba, H. Nakagome, and N. Kunugita. 2015. Gaseous chemical compounds in indoor and outdoor air of 602 houses throughout Japan in winter and summer. Environmental Research 137:364–372. https://doi.org/10.1016/j.envres.2014.12.005.

Ulrich, E. M., J. R. Sobus, C. M. Grulke, A. M. Richard, S. R. Newton, M. J. Strynar, K. Mansouri, and A. J. Williams. 2019. EPA’s non-targeted analysis collaborative trial (ENTACT): Genesis, design, and initial findings. Analytical and Bioanalytical Chemistry 411(4):853–866. https://doi.org/10.1007/s00216-018-1435-6.

Unnithan, A., D. N. Bekele, S. Chadalavada, and R. Naidu. 2021. Insights into vapour intrusion phenomena: Current outlook and preferential pathway scenario. Science of the Total Environment 796:148885. https://doi.org/10.1016/j.scitotenv.2021.148885.

USHUD. 2018. Water Piping in the Home. https://www.huduser.gov/portal/pdredge/pdr-edge-trending-091018.html. Accessed May 13, 2022.

Vardoulakis, S., E. Giagloglou, S. Steinle, A. Davis, A. Sleeuwenhoek, K. Galea, K. Dixon, and J. Crawford. 2020. Indoor exposure to selected air pollutants in the home environment: A systematic review. International Journal of Environmental Research and Public Health 17(23):8972. https://doi.org/10.3390/ijerph17238972.

Venier, M., O. Audy, Š. Vojta, J. Bečanová, K. Romanak, L. Melymuk, M. Krátká, P. Kukučka, J. Okeme, A. Saini, M. L. Diamond, and J. Klánová. 2016. Brominated flame retardants in the indoor environment - Comparative study of indoor contamination from three countries. Environ Int 94:150–160. https://doi.org/10.1016/j.envint.2016.04.029.

Vykoukalová, M., M. Venier, Š. Vojta, L. Melymuk, J. Bečanová, K. Romanak, R. Prokeš, J. O. Okeme, A. Saini, M. L. Diamond, and J. Klánová. 2017. Organophosphate esters flame retardants in the indoor environment. Environ Int 106:97–104. https://doi.org/10.1016/j.envint.2017.05.020.

Wallace, L. 2000. Correlations of personal exposure to particles with outdoor air measurements: A review of recent studies. Aerosol Science and Technology 32(1):15–25. https://doi.org/10.1080/027868200303894.

Wallace, L., and W. Ott. 2011. Personal exposure to ultrafine particles. Journal of Exposure Science & Environmental Epidemiology 21(1):20–30. https://doi.org/10.1038/jes.2009.59.

Wallace, L., F. Wang, C. Howard-Reed, and A. Persily. 2008. Contribution of gas and electric stoves to residential ultrafine particle concentration between 2 and 64 nm: Size distributions and emission and coagulation rates. Environmental Science & Technology 42(23):8641–8647. https://doi.org/10.1021/es801402v.

Wallace, L. A., S. J. Emmerich, and C. Howard-Reed. 2004. Source strengths of ultrafine and fine particles due to cooking with a gas stove. Environmental Science & Technology 38(8):2304–2311. https://doi.org/10.1021/es0306260.

Wallace, L. A., W. R. Ott, C. J. Weschler, and A. C. K. Lai. 2017. Desorption of SVOCs from heated surfaces in the form of ultrafine particles. Environmental Science & Technology 51(3):1140–1146. https://doi.org/10.1021/acs.est.6b03248.

Wang, L., A. G. Asimakopoulos, H.-B. Moon, H. Nakata, and K. Kannan. 2013. Benzotriazole, benzothiazole, and benzophe-none compounds in indoor dust from the United States and East Asian countries. Environmental Science & Technology 47(9):4752–4759. https://doi.org/10.1021/es305000d.

Wang, Z., G. W. Walker, D. C. G. Muir, and K. Nagatani-Yoshida. 2020. Toward a global understanding of chemical pollution: A first comprehensive analysis of national and regional chemical inventories. Environmental Science & Technology 54(5):2575–2584. https://doi.org/10.1021/acs.est.9b06379.

Weschler, C. J. 2016. Roles of the human occupant in indoor chemistry. Indoor Air 26:6–24.

Weschler, C. J., and W. W. Nazaroff. 2008. Semivolatile organic compounds in indoor environments. Atmospheric Environment 42(40): 9018–9040. https://doi.org/10.1016/j.atmosenv.2008.09.052.

This page intentionally left blank.