Materials responsible for formaldehyde and volatile organic compound (VOC) emissions
Abstract:
Volatile organic compounds (VOCs) are an important class of indoor air pollutants; with indoor concentrations generally higher than outdoors. Formaldehyde is a priority VOC because of its frequent occurrence in indoor air and the serious health outcomes resulting from exposure. Taking formaldehyde as a representative VOC, this chapter reviews the knowledge necessary to develop solutions to indoor VOC pollution. The toxicology of formaldehyde is briefly reviewed. Then the current understanding of VOC emission behavior is discussed, including experimental techniques for measuring emissions, modeling approaches for predicting emissions, and the impacts of environmental factors on emissions. With a comprehensive understanding spanning emission characteristics and toxicology, it is possible to develop effective strategies to maintain indoor VOC concentrations below a safe threshold.
4.1 Introduction
Since people typically spend over 80% of their time indoors (Klepeis et al., 2001; Adgate et al., 2004), indoor air quality (IAQ) has a substantial effect on occupants’ comfort, health and productivity. Among all the causes of degradation of indoor air quality, volatile organic compounds (VOCs) have been recognized as one of the most important classes of indoor air pollutants (Weschler, 2009). VOCs are a large group of organic chemicals that have a low boiling point; according to the US EPA, VOC means ‘any compound of carbon, excluding carbon monoxide, carbon dioxide, carbonic acid, metallic carbides or carbonates, and ammonium carbonate, which participates in atmospheric photochemical reactions’. Common VOCs occurring in the indoor environment include formaldehyde, benzene, toluene, xylene, styrene, acetaldehyde, naphthalene, limonene, and hexanal (Weschler, 2009; Sarigiannis et al., 2011). Exposure to VOCs may cause reduced worker productivity (Bako-Biro et al., 2004; Fanger, 2006), acute health effects such as eye and respiratory irritations, headaches, fatigue, and asthmatic symptoms (Mølhave, 1989; Wolkoff and Nielsen, 2001; Billionnet et al., 2011; Jie et al., 2011), and chronic illnesses such as cancer (Rennix et al., 2005; Sax et al., 2006; Boeglin et al., 2006). However, the health effect of individual VOCs can vary greatly, ranging from being highly toxic to having little known health effects. For example, benzene is ‘known to be a human carcinogen’ (NTP, 2011) while toluene is much less toxic than benzene although it shares a similar molecular structure. There are many sources of VOCs in the indoor environment, including building materials, consumer products, and furniture (Bluyssen et al., 1996; Missia et al., 2010). Some major sources, such as paints, carpets, composite wood products, and floorings, are used extensively and permanently indoors, leading to ubiquitous and abundant presence of VOCs in the indoor air. It has been reported that indoor VOC concentrations generally far exceed outdoor levels (Ohura et al., 2006; Jia et al., 2008; Missia et al., 2010).
Among all the VOCs in the indoor environment, formaldehyde is one of the most common and best-known compounds and a priority indoor air pollutant due to its wide distribution in indoor air and its highly toxic nature (Salthammer et al., 2010). Different from most VOCs which are liquids or solids, formaldehyde is a colorless gas with a pungent odor at room temperature and pressure (Reuss et al., 2003). It is soluble in water and is generally used in solution or in its polymerized form, paraformaldehyde (Lide, 2003). Formaldehyde is formed in large quantities via the oxidation of hydrocarbons naturally (WHO, 1989). It is also endogenously formed in most life forms and is present in tissues, cells, and bodily fluids (Heck and Casanova, 2004). The most important anthropogenic sources include direct emission from production and use of formaldehyde, and combustion, such as smoking and automotive exhaust from engines (WHO, 1989). Since the 1880s formaldehyde has been produced commercially, and in recent years global industrial production of formaldehyde is over 20 million tonnes (Bizzari, 2000), mainly used for the following (IARC, 2006):
• Production of synthetic resins including urea-formaldehyde, phenolformaldehyde, and melamine-formaldehyde. These resins are used as adhesives and impregnating resins in the manufacture of wood products and curable molding products, and in the textile, leather, rubber and cement industries.
• As an intermediate in the synthesis of other industrial chemicals.
• As a preservation and disinfection agent for human and veterinary drugs, biological specimens, pesticides and cosmetic products.
Under atmospheric conditions, formaldehyde is readily removed by photolysis and reaction with hydroxyl radicals in sunlight to carbon dioxide, resulting in a low background concentration from 0.1 to 2.7 μg/m3 (WHO, 1989). Due to the presence of a large spectrum of formaldehyde emission sources indoors, such as the composite wood products made from urea-formaldehyde resin, and the slow removal rate in the indoor environment, the levels of formaldehyde in indoor air are mostly much higher than outdoors, ranging from 10 to 4000 μg/m3 (WHO, 1989; IARC, 2006). Therefore, although people can be exposed to formaldehyde through other sources such as food, indoor air is the most influential source for the general population while occupational exposures are important for specific populations such as employees in formaldehyde-related industry and sanitary services (Kauppinen et al., 2000).
It has been well recognized that exposure to formaldehyde can cause irritation of the mucosa of the eye and upper respiratory system and may induce allergic contact dermatitis and contact urticaria (Paustenbach et al., 1997; Koss and Tesseraux, 1999; Arts et al., 2008). Chronic formaldehyde exposure can cause cancers, and therefore it has been classified as ‘carcinogenic to humans (Group 1)’ by the International Agency for Research on Cancer (IARC, 2006) and described as ‘known to be a human carcinogen’ by the US National Toxicology Program (NTP, 2011). However, the toxicological mechanisms of formaldehyde are complicated and require further study (Speit et al., 2007; Bosetti et al., 2008). Concerning the health risks associated with indoor formaldehyde exposure, various international guidelines and recommendations have been established for formaldehyde in indoor air (Salthammer et al., 2010). Recently, new legislation on ‘Formaldehyde Standards for Composite Wood Products’ was passed by the US Congress and signed by the President.
The occurrences of formaldehyde and other VOCs in building materials and emissions during the use phase are closely related to the manufacturing procedures. For example, manufacturing of widely used medium-density fiberboard consists of several steps: firstly, wood chips are milled into wood fibers; the wood fibers are then blended with adhesive resins and the resulting mixture is dried by hot air; finally, the wood fibers after resin application are placed on a conveyer belt and hot-pressed into medium-density fiberboard (He et al., 2012). With low cost and good performance, urea-formaldehyde and phenol-formaldehyde resins are the most commonly used adhesives in wood-based panels, both of which would release formaldehyde due to residuals and degradation. The application of adhesives and hot-pressing treatment therefore lead to high emissions of formaldehyde when the products are used indoors.
In this chapter, formaldehyde will be presented in detail as the representative of indoor VOCs. Its toxicology will be first discussed, including its toxicokinetics, major health effects, and the possible mechanisms of its primary toxicities. Then state-of-the-art knowledge for formaldehyde and other VOC emissions from building materials will be reviewed, including the experimental techniques for measuring emissions, modeling techniques for describing and predicting emissions, and impacts of environmental factors on emissions. These understandings could facilitate the screening-level assessment of exposure to indoor VOCs, product reformulation strategies to reduce or prevent VOC emissions, and development of standards for both the VOC concentrations in indoor air and environmental performance of indoor materials.
4.2 Toxicology of formaldehyde
4.2.1 Toxicokinetics
Absorption
Due to its high water solubility, formaldehyde is rapidly absorbed in the respiratory tract once inhaled (WHO, 1989). Various airflow patterns due to anatomical differences in noses may cause different uptake of inhaled formaldehyde by different species of animals (Schreider, 1986; Morgan et al., 1991; Georgieva et al., 2003). A three-dimensional, anatomically realistic computational fluid dynamics (CFD) model of formaldehyde gas transport in the nasal passages predicted that over 90% of inhaled formaldehyde is absorbed in the upper respiratory tract for humans (Kimbell et al., 2001). Formaldehyde can penetrate human skin but dermal absorption is expected to be slight when exposed to airborne formaldehyde (WHO, 1989).
Distribution
After absorption by animals and humans, formaldehyde can be metabolized and distributed rapidly to the entire body. For example, there was no exposure-related increase in the blood concentration of human volunteers after exposure to 2.3 mg/m3 formaldehyde for 40 minutes (Heck et al., 1985). Following a six-hour inhalation exposure of rats to 14C-formaldehyde, radioactivity was extensively distributed in other tissues, indicating that absorbed 14C-formaldehyde and its metabolites are rapidly removed by the mucosal blood supply and distributed throughout the body (Johansson and Tjalve, 1978).
Metabolism
Formaldehyde undergoes rapid biotransformation immediately after absorption (WHO, 1989). Figure 4.1 shows the biological reactions and metabolism pathways of formaldehyde (Bolt, 1987). Formaldehyde can be oxidized into formate/formic acid and then carbon dioxide. It can also undergo nonenzymatic reactions with amino and other groups in DNA, RNA and protein molecules and can further cause crosslinks between two macromolecules. Formaldehyde and its metabolites can also be incorporated into the one-carbon pool for synthesis of certain nucleic acids and amino acids, and eventually, cellar macromolecules.
Fig. 4.1 Biological reactions and metabolism pathways of formaldehyde. (Bolt, 1987)
Figure 4.2 shows the detailed pathways for metabolic detoxification by oxidation and the enzyme systems involved (Hedberg et al., 2002). The primary and generally most important system initially involves alcohol dehydrogenase 3 (ADH3), which oxidizes S-hydroxymethylglutathione, GSH-conjugated formaldehyde spontaneously formed by formaldehyde and reduced glutathione (GSH), to S-formylglutathione. This intermediate is then further metabolized by S-formylglutathione hydrolase to yield formate/formic acid and reduced GSH. Identical to any alcohol oxidation by ADH enzymes, the ADH3-dependent step requires catalysis by zinc and uses NAD+/NADH (coenzymes found in all living cells) as the electron acceptor and donor (Höög et al., 2001). The second enzyme system is the ALDHs – class 1 (cytosolic ALDH; ALDH1A1) and class 2 (mitochondrial ALDH; ALDH2) – which have an affinity for free formaldehyde. Catalase may also contribute to the oxidation of formaldehyde to formate/formic acid, but only when hydrogen peroxide is present (Jones et al., 1978). Formaldehyde may also be reduced to methanol and then reconverted to formaldehyde (Pocker and Li, 1991).
Fig. 4.2 Metabolism of formaldehyde. (Hedberg et al., 2002)
Excretion
Elimination of formaldehyde from the blood takes very short time, with a half-time of about 1–1.5 minutes, via exhalation or renal excretion. Radioactive investigations on rats using 14C-formaldehyde suggest over 80% of the radiolabel was recovered as carbon dioxide in exhalation (Du Vigneaud et al., 1950; Neely, 1964). Small amounts are excreted in the urine as formate salts, methionine, serine, and other metabolites (WHO, 1989).
4.2.2 Health effects of human exposure
Predominant effects of short-term exposure to formaldehyde in humans include irritation of the eyes, nose and throat, and concentration-dependent discomfort, lachrymation, sneezing, coughing, nausea, dyspnoea and finally death (Table 4.1). Formaldehyde is also a sensitizer. There are several studies reporting asthma caused by sensitization effects of formaldehyde (Nordman et al., 1985; Wantke et al., 1996). Skin sensitization can be induced by direct skin contact with formaldehyde solutions, causing allergic contact dermatitis and contact urticaria (WHO, 1989).
Table 4.1
Effects of formaldehyde in humans after short-term exposure
Sources: WHO (1989); IARC (1995).
It has been confirmed that exposure to formaldehyde can cause nasopharyngeal cancer (IARC, 2006). Although a number of studies have found associations between exposure to formaldehyde and occurrences of other cancers, a causal role for formaldehyde in relation to them cannot be established until more evidence is available (IARC, 2006). The reasons for less opportunity to cause remote site cancers and systematic illnesses compared with the chance in the upper respiratory tract may include the fact that formaldehyde is absorbed almost completely in the upper respiratory tract when inhaled and then rapidly metabolized. There are also a variety of studies evaluating the neural and reproductive effects of exposures to formaldehyde in humans but the results are inconclusive overall (Hemminki et al., 1982; Axelsson et al., 1984; Taskinen et al., 1994, 1999; Collins et al., 2001; IARC, 2006).
4.2.3 Mechanisms of action
Reactions with macromolecules
Formaldehyde can form covalent bindings to specific sites on macromolecules by replacing active hydrogen atoms, and a two-step mechanism for the formation of methylene crosslinks between macromolecules has been proposed (Feldman, 1973). The first step is the fast, reversible formation of unstable methylol derivatives by formaldehyde and macromolecules. The irreversible formation of a stable methylene crosslink may then occur by way of nucleophilic attack on the methylene carbon. These crosslinks could occur between two proteins, between DNA and protein, or between DNA and DNA. By involving DNA, the latter two reactions can arrest DNA replication and lead to micronucleus (MN), sister chromatid exchanges (SCEs), chromosomal aberrations (CA), and DNA damage (Speit et al., 2007; Costa et al., 2008). Such damage to DNA, however, can be removed by spontaneous hydrolysis and active repair by cellular DNA repair systems and, therefore, cell division is required prior to DNA repair for mutagenesis to occur. Before formaldehyde reacts with amino groups in RNA, the hydrogen bonds forming the coiled RNA may break; formaldehyde hardly reacts with native double-stranded DNA since the hydrogen bonds holding DNA in its double helix are more stable (Feldman, 1973; Auerbach et al., 1977). Therefore, unwinding of the double helix during cell replication may be required to expose critical sites on the DNA to covalent binding (Singer and Kusmierek, 1982). Convincing evidence supporting this theory is the apparent cell-cycle specificity of formaldehyde: mutagenesis in Drosophila is restricted to the period of chromosome replication preceding meiosis (Auerbach et al., 1977). This theory also explains the experimental results that exponentially growing cultures of yeast have greater sensitivities to lethality and mutagenesis than stationary cultures (Chanet and von Borstel, 1979).
Cytotoxicity
As described in the previous subsection, formaldehyde can react with proteins and result in protein denaturation, precipitation and coagulation necrosis, inhibiting cellular physiologic functions or even killing the cells. Recent in vitro studies suggest oxidative stress increase caused by formaldehyde may be also an important mechanism for its molecular cytotoxicity (Teng et al., 2001; Oyama et al., 2002). High oxidative stress can cause toxic effects through the production of peroxides and free radicals that damage all components of the cell. Inhibiting cell respiration and other cellular enzymatic antioxidant defense systems by formaldehyde, reactive oxygen species are further generated, thus contributing to oxidative stress. Furthermore, high levels of formaldehyde can significantly reduce GSH content, which actually can work to protect cells from oxidative stress (Meister and Anderson, 1983), and the decrease in cellular GSH content therefore increases cell vulnerability to oxidative stress (Kashiwagi et al., 1994). Obviously, the various mechanisms of cytotoxicity of formaldehyde are correlated with each other, making the impacts complex. As a result of its cytotoxicity, formaldehyde can be used for disinfection and conservation in several circumstances.
The cytotoxicity of formaldehyde also plays an important role in its car-cinogenicity. A prominent response to cell loss associated with cytotoxicity is compensatory cell replication, i.e. division of surviving cells to compensate for dead cells and maintain the functions of organisms. Several studies have shown that cell proliferation is stimulated after exposure to formaldehyde (Swenberg et al., 1983; Tyihák et al., 2001). During the increased cell division, the likelihood of interaction of formaldehyde with DNA would increase, as would fixation of adducts before DNA repair could occur. Therefore, carcinogenesis of formaldehyde is related directly to the increased cell division resulting from its cytotoxicity.
Irritation
It is well known that formaldehyde can cause irritation, which is defined as a chemically produced local inflammatory response characterized by edema, erythema or corrosion (Mathias and Maibach, 1982). Respiratory tract irritation and eye irritation can involve a chemosensory effect, i.e., formaldehyde molecules bind to specific receptors on sensory neurons at local nerve endings (nervus trigeminus) and therefore activate an opening ion channel via specific proteins, which is called trigeminal stimulation or sensory irritation. Over a broad range of concentrations, the trigeminal stimulation will not necessarily lead to cell or tissue damage but it may cause reflex responses such as sneezing, lachrymation, rhinorrhea, coughing, vasodilation and changes in the rate and depth of respiration, resulting in a decrease in the total amount of inhaled material, thus protecting the individual (Lang et al., 2008). On the other hand, the irritation occurring at high concentrations is a localized pathophysiological response to formaldehyde involving local redness, swelling, pruritis or pain. These effects are comparable to those induced at the skin and can be termed pathological irritation (Arts, 2006). The pathological irritation of formaldehyde is believed to be similar to that caused by other irritant chemicals, such as toluene diisocyanate (TDI) and cotton dust, which cause histamine release from basophils.
Immunological basis of allergy
Some adverse effects caused by formaldehyde, such as allergic contact dermatitis and asthma, may result from its immunological effects as a sensitizer (Feinman, 1988). Most allergic reactions to formaldehyde can be divided into two basic types: delayed cell-mediated and immediate antibody-mediated, which follow different characteristic time courses and are explained by different mechanisms (WHO, 1989; Roberts and Adams, 2000). Allergic contact dermatitis (ACD), which can be induced by skin contact with formaldehyde, is an example of delayed cell-mediated hypersensitivity. T-cells, one specific group of lymphocytes, and Langerhans cells, one group of dendritic cells in the epidermis, are involved (Merad et al., 2008). Formaldehyde can bind to skin proteins to produce complete antigen complexes (Basketter et al., 2008). Langerhans cells can take up these complexes and present them to T-cells so that sensitization of T-cells occurs. Upon reexposure to formaldehyde, primed T-cells recognize the antigen and are triggered to produce effector cells, which initiate the immune reaction. There is a delay of 24–48 hours before this reaction is elicited. After elicitation, pharmacologically active mediators (such as enzymes and cytokines) are secreted, producing tissue inflammation.
Once established, ACD may persist for years (Feinman, 1988). In contrast, antibody-mediated immediate hypersensitivity reactions have a rapid onset and shorter duration. They are mediated by special antibodies produced in response to allergic stimulation in previously sensitized subjects. IgE is a specific antibody class important in the immediate allergic response to environmental antigens. IgE antibodies can bind to the mast cells, granulocytes, macrophages and platelets. These cells, after binding IgE and antigen, are triggered to release mediators, such as prostaglandins, histamine, leukotrienes, chemotactic factors, and platelet activating factors. These mediators directly or indirectly produce effects such as smooth muscle contraction, bronchoconstriction and capillary vasodilation that result in characteristic clinical syndromes like asthma and contact urticaria. The production of specific antibodies is under genetic control so that there are geographical and demographic differences in the incidence of antibody-mediated immediate hypersensitivity to a given allergen (Feinman, 1988).
4.3 Emission testing of formaldehyde and other volatile organic compounds (VOCs)
4.3.1 Analytical methods for measuring formaldehyde and other VOCs
To measure the concentration of formaldehyde and other VOCs in indoor air, two main analytical methods are generally deployed, i.e., the GC method and the DNPH method.
For the GC method, the VOCs are first collected from the indoor environment using air sampling pumps coupled with sorbent tubes containing appropriate sorbents, such as Tenax® TA sorbent, activated carbon and graphitized carbon blacks. Then the collected VOCs are released by thermal desorption and injected into a gas chromatography (GC) column, a common type of chromatography used in analytic chemistry for separating compounds. Different kinds of detectors can then be used for qualitative and quantitative determination of the compounds. The flame ionization detector (FID), photoionization detector (PID), and mass spectrometric detector (MS) have been successfully used for VOCs (Ulman and Chilmonczyk, 2007) Chung et al. (2003) used GC/FID to simultaneously measure the total non-methane organic carbon and speciated VOCs. Zhang et al. (2000) suggested that GC/PID is an easy, fast and reliable method for analyzing isoprene emissions. GC/MS procedure to determine VOC concentrations has been included in several standards such as ASTM D5466-01 (2001) and ISO 16000-6 (2004). The GC/MS analysis can also be used in conjunction with a solid-phase micro-extraction (SPME) technique. The SPME involves a very thin fiber coated with an extracting phase, which can extract different kinds of VOCs from samples. The extracted VOCs can then be analyzed by GC/MS. The combination of SPME and GC/MS techniques offers spectral analysis and high instrumental sensitivity, and allows simultaneous identification and quantification of several VOCs in a sample (Pecoraino et al., 2008). Furthermore, SPME is generally convenient and fast, and can be performed without solvents, thus having great potential in laboratory and field applications (Pawliszyn, 2009).
The DNPH method is widely used for determining carbonyl compounds, such as formaldehyde, other aldehydes and ketones. It has been recommended by several standards (ISO 16000-3, 2001; ASTM D5197-03, 2003). This method is based on the specific reaction between carbonyl compounds and the 2,4-dinitrophenylhydrazine (DNPH) coated on a silica gel adsorbent, which forms stable hydrazone derivatives by nucleophilic addition of DNPH with the carbonyl group in the presence of a strong acid catalyst. During the sampling process, indoor air is pumped through cartridges containing silica gel coated with an acid solution of DNPH. Then the DNPH-carbonyl derivatives are extracted with acetonitrile and analyzed using high-performance liquid chromatography (HPLC). During the analysis process, the chromatographic separation of the hydrazones is achieved using a C18-column and water/acetonitrile solvent combinations, and an ultraviolet (UV) absorption detector is often used for detection (Vairavamurthy et al., 1992; Salthammer et al., 2010). However, the DNPH method still has some drawbacks. Foster et al. (1996) observed that air humidity would affect the formation rate of hydrazones due to the accumulation of water on the cartridge. The presence of ozone at high concentrations would interfere negatively by reacting with both DNPH and its carbonyl derivatives in the cartridge (ASTM D5197-03, 2003). Ho et al. (2011) suggested that the DNPH method is unsuitable for determining unsaturated carbonyls and alternative derivatizing agents or other analytical methods should be found.
4.3.2 Emission testing methods and chambers
The emission of formaldehyde and other VOCs is an important factor in evaluating the environmental performance and health impacts of building materials. Various emission testing methods have been developed and used, and some of them have been specified in standards. For the emission testing of formaldehyde from wood-based materials, the perforator method, desiccator method, and chamber method are commonly used.
The perforator method, established as European standard EN 120 (1993), determines formaldehyde content in wood-based materials by extraction in a perforator. The test material is first cut into pieces. Then the formaldehyde in the pieces is extracted by boiling toluene and transferred into water. Finally, the formaldehyde content in the aqueous solution is measured by a suitable analytical technique (Risholm-Sundman et al., 2007). The entire procedure requires large and complicated equipment, and takes 2 h for extraction and a total of about 4 h. This method determines the total formaldehyde content in the test material.
For the desiccator method (JIS A1460, 2001), the sample with a given surface area is positioned over water in a desiccator at a controlled temperature. The formaldehyde released from the tested sample for 24 h is collected by the water and then determined by the acetylacetone method or the chromotropic acid method (Salthammer et al., 2010). This method provides an accurate measurement for formaldehyde emission from the sample, but does not provide information about total formaldehyde content (Kim et al., 2010).
The chamber method, which has been prescribed in many standards, is widely used for the emissions testing of formaldehyde and other VOCs. The sample is placed in a chamber with clean air passing through it and the chamber concentrations of target VOCs are measured over time, generally for several days to weeks. The emission testing chambers are often made of glass or stainless steel, and in a cylindrical or rectangular shape, with size varying from a few liters (small chamber) to several cubic meters (large chamber or full-scale chamber). Chamber testing conditions are commonly controlled or specified, such as temperature, relative humidity, air exchange rate, air velocity, and material volume or loading factor (ratio of the material surface to the volume of the chamber). For the chamber method specified in most standards, the result is reported as an area-specific emission rate (emission factor) or as a chamber concentration at steady state (EN 717–1, 2004; ISO 16000–10, 2006; ASTM D 6670–01, 2001; ISO/CD 12460, 2007; Salthammer et al., 2010). The main advantage of the chamber method is that it simulates the real scenario of emissions in the indoor environment.
Table 4.2 lists some typical emission testing chambers and their applications. In the table, the Field and Laboratory Emission Cell (FLEC) and the Chamber for Laboratory Investigations of Materials, Pollution and Air Quality (CLIMPAQ) are two specially designed chambers. The FLEC is circular and made of stainless steel with a volume of 0.035 L, which includes a cap and a lower chamber. When testing, the material is placed in the lower chamber, forming a cone-shaped cavity with the inner surface of the FLEC cap. Zhang and Niu (2003) analyzed the mass transfer process of VOC in the FLEC. The CLIMPAQ is made of panes of window glass with a volume of 50.9 L, and other main surface materials are stainless steel and eloxated aluminum. One internal fan circulates the air over the tested material with some metal grids in the flow direction. The air exchange rate and air velocity over the test material surface can therefore be adjusted independently.
Table 4.2
Some typical emission testing chambers and their applications
| Testing chamber | Application | Standard/reference |
| 20 L small chamber | Determination of formaldehyde release, formaldehyde emission by the chamber method | ISO/CD 12460, 2007 |
| 30 L small chamber | Determination of the emission characteristic parameters of formaldehyde and other VOCs from building materials | Xiong et al., 2011 |
| 50 L small chamber | Measurement of VOC adsorption/desorption characteristics of typical building materials | An et al., 1999 |
| Measurement and simulation of VOC emission from materials | Yang et al., 2001 | |
| Study of some environmental factors on VOC adsorption | Huang et al., 2006 | |
| Dual 18.4 L small chamber | Measurement of the VOC diffusion and partition coefficients and study of the similarities between water vapor and VOC diffusion | Xu et al., 2009 |
| Dual 50 L small chamber | Measurement of the VOC diffusion and partition coefficients for building materials | Bodalal et al., 2000 |
| FLEC | Determination of the VOC emissions from building products and furnishing | ISO 16000-10, 2006 |
| Determination of VOC emissions from indoor materials/products Characterization of VOC emissions from building materials | ASTM D 7143-05, 2005 Zhang and Niu, 2003 |
|
| CLIMPAQ | Emission testing of materials and products in a controlled environment | Gunnarsen et al., 1994 |
| Dual CLIMPAQ | Measurement of VOC diffusion and sorption parameters in building materials | Meininghaus et al., 2000 |
| 1 m3 chamber | Determination of formaldehyde release by the chamber method | ISO/CD 12460, 2007 |
| Large chamber (≥ 20 m3) | Determination of formaldehyde release by the chamber method Determination of VOC emissions from indoor materials/products | EN 717-1, 2004 ASTM D 6670-01, 2001 |
| 20 m3 large chamber | Measurement of VOC emissions from new carpets | Little et al., 1994 |
| 30 m3 large chamber | Determination of the emission characteristic parameters of formaldehyde and other VOCs from furniture | Yao et al., 2011 |
4.3.3 Reference materials for emissions testing
As introduced previously, the chamber method is widely used to determine emissions of formaldehyde and other VOCs from building materials. However, very different emission profiles are often obtained for the same material tested in different laboratories (Howard-Reed and Nabinger, 2006). There is thus a compelling need for a reference emission source that can be used to evaluate and calibrate the testing procedures. Till now, two different kinds of reference materials have been developed for VOC emissions testing (Cox et al., 2010; Wei et al., 2011).
For the first reference material, polymethylpentene (PMP) film is selected as the substrate. Toluene, as a representative VOC, is infused into the film so that the loaded film has an emission profile similar to that of a typical building material that can be measured in emission testing chambers. As Fig. 4.3 shows, the toluene-laden dry air is passed through stainless steel vessels containing several PMP films. The effluent air from the last vessel is passed through a high-resolution dynamic microbalance which holds an extra film and monitors its mass throughout the loading process. During the loading process (about 2 weeks), toluene molecules diffuse from the air into the films until sorption equilibrium is reached between the material phase and the gas phase. The mass change data recorded by the microbalance is used to monitor the loading process and determine the material-phase concentration of toluene in the films when loading is complete. The loaded films are then sent to different laboratories for emission testing. Meanwhile, the emission characteristic parameters of the loaded film can be determined independently and a fundamental model can be used to predict its emissions accurately (Cox et al., 2010; Liu et al., 2011). Therefore, the model-predicted emission profile serves as a true reference value for validating the measured results by different laboratories, evaluating the test performance, and identifying the root cause of variability. This reference material has been successfully employed in some interlaboratory studies (Howard-Reed et al.., 2011a, 2011b).
The second reference material is called a LIFE (liquid, inner-tube-diffusion, film and emission) reference (Wei et al., 2011), as shown in Fig. 4.4. The LIFE reference comprises a small Teflon cylinder containing a pure VOC in liquid phase, a thin film covering the opening of the cylinder, and some Teflon fastening pieces (washers and hole-cover). Therefore, VOC can diffuse through the film and the emission rate can be adjusted by changing the film. Toluene is selected as the target VOC to develop the LIFE reference, and the preliminary results show that it has the following features (Wei et al., 2011): (1) its emission rate is constant; (2) it has a long applied life of 1000 hours; and (3) it is easy to store, apply and maintain. Therefore the LIFE reference is very useful for calibrating emission testing procedures. In addition, the LIFE reference also has the equivalent emission characteristic parameters (initial emittable concentration, diffusion coefficient and partition coefficient), and can be treated as an equivalent building material. It can be thus used as a reference to determine whether a method for measuring emission characteristic parameters is reliable or not.
4.4 Emission models of formaldehyde and other volatile organic compounds (VOCs)
Although solid emission data can be obtained from emission chamber tests, they are only applicable for the specific testing conditions. Emission chamber tests are also often expensive and time-consuming. Therefore, much effort has been made to predict emissions using modeling techniques. Generally, existing models can be categorized into two groups (Zhang and Xu, 2003). The first one is empirical models constructed upon statistical analysis of emission test data (Guo, 2002a), such as the first-order decay model and the second-order decay model (Dunn, 1987; Clausen, 1993). Although simple to derive and use, empirical models lack a physical basis and provide little insight into the controlling mechanism, and therefore cannot be easily scaled from the test conditions to other conditions. In contrast, the second group of models, as will be introduced below, is based on mass-transfer mechanisms with model parameters having clear physical meanings, and therefore can predict emissions for various conditions.
Since VOCs emitted from building materials generally originate from inside the materials, the mass transfer within the materials is important in determining the emission rates and should be considered in modeling. Up to now, two kinds of diffusion models have been proposed, i.e., one-phase solid models (Little et al., 1994; Huang and Haghighat, 2002; Xu and Zhang, 2003; Deng and Kim, 2004), and multi-phase solid models (Lee et al., 2005, 2006; Haghighat et al., 2005; Blondeau et al., 2008; Marion et al., 2011). Multiphase solid models consider the porous structure of the material (pores and solid) and the mass transfer in both the pore-phase and the adsorbed-phase. Instead, one-phase solid models lump the microstructure of the building material into one uniform solid phase and are much simpler than multiphase solid models. However, multi-phase solid models and one-phase solid models can be transformed from one to another and model parameters are also interrelated (Haghighat et al., 2005; Blondeau et al., 2008; Xu et al., 2009). Therefore, typical physically based one-phase solid models for predicting VOC emissions will be presented in this section, and the masstransfer mechanisms of VOC emissions will be discussed.
The first physically based diffusion model was developed by Little et al. (1994) for predicting VOC emissions from carpets, which also presents a widely accepted framework for later development of VOC emission models. With reference to Fig. 4.5, the transient diffusion within the material is described by Fick’s second law,
Fig. 4.5 Schematic representation of VOC emission from a solid material. (Little et al., 1994)
where C(x,t) is the VOC concentration in the slab of material, D is the material-phase diffusion coefficient, t is time, and x is the distance from the bottom of the slab. The boundary condition at the bottom of the material assumes there is no flux, or
The boundary condition at the exposed surface is imposed by a mass balance on the VOC in the chamber air assuming the chamber air is well mixed,
where y is the concentration of VOC in the well-mixed chamber air, Q is the flow rate of air through the chamber, V is the chamber volume, and A is the emission surface area. A linear and instantaneously reversible equilibrium relationship is assumed at the material/air interface, or
where K is the partition coefficient between the material and air. Equation 4.4 implies that external convective mass transfer resistance near the interface is ignored. Assuming concentration-independent D and K and a uniform initial emittable concentration of the VOC, C0, an analytical solution to this equation set was obtained by Little et al. (1994) for calculating C and y at any time directly.
Without neglecting the external mass-transfer resistance, some improved models were then developed. Figure 4.6 shows the principle of VOC emission from a solid material adopted in several models (Huang and Haghighat, 2002; Xu and Zhang, 2003; Deng and Kim, 2004). The difference between Fig. 4.6 and Fig. 4.5 and thus the improvement is the introduction of the convective mass-transfer coefficient, hm, to account for the external convective mass transfer. With the diffusion-governing equation 4.1 and the boundary condition at the bottom, equation 4.2, still valid, the boundary condition at the exposed surface, equations 4.3 and 4.4, should be replaced by
Equation 4.5 shows how external convective mass transfer is taken into account: the emission flux at the material surface is equal to the convective mass transfer through the boundary layer, where y0 is the concentration of air at the material surface while y is the bulk air concentration. Equation 4.6 is the transient mass balance in the chamber air, assuming influent concentration is zero. Similar to equation 4.4, equation 4.7 assumes instantaneous reversible partition equilibrium at the air/material interface. Deng and Kim (2004) obtained the fully analytical solution for this equation set, which can calculate the chamber concentration at any time directly. Generally, this improved model with a fully analytical solution is a desirable approach to predict the emissions from a single layer of homogeneous material. When influent concentration different from zero needs to be considered, the finite difference methods (Huang and Haghighat, 2002; Xu and Zhang, 2003) and the recently developed state-space method (Yan et al., 2009) can be used.
These mass-transfer models discussed above provide not only tools for prediction but also comprehensive insight into the overall emission mechanisms. The mechanisms governing emissions of VOCs from a solid material include the internal diffusion of VOCs within the material (characterized by the diffusion coefficient, D), partition between the material and chamber air at the material/air interface (characterized by the partition coefficient, K), and external convective mass transfer from the air at the material surface to the chamber bulk air (characterized by the external convective mass-transfer coefficient, hm). These three parameters together with the initial emittable concentration, C0, determine the emission strength and duration. Qualitatively, the relative magnitude of the internal mass-transfer resistance due to diffusion and the external convective mass-transfer resistance determines the emission-rate limiting factor. When internal mass-transfer resistance is small, external mass-transfer resistance has noticeable impacts on the emission rate. In contrast, the impact of external convective mass-transfer is negligible when internal mass-transfer resistance is very large, and emissions of VOCs under this condition are internal-diffusion controlled. Therefore, the model developed by Little et al. (1994) is a reasonable simplification for internal-diffusion controlled cases.
Based on this framework, several models have been further developed for various purposes. For example, sorption behaviors are found to be important for some building materials, and models concerning the sink effect of a single material have been developed (Little and Hodgson, 1996; Zhao et al., 2002; Xu and Zhang, 2004; Kumar and Little, 2003a). Several models were also developed to account for more complicated conditions, such as the non-uniform initial emittable concentration and non-zero and time-dependent influent air concentration (Kumar and Little, 2003a). Furthermore, models for predicting emissions from a layered composite material (Kumar and Little, 2003b; Haghighat and Huang, 2003; Yuan et al., 2007a; Hu et al., 2007; Deng et al., 2010) and several layered composite materials (Zhang and Niu, 2004; Li and Niu, 2007) have been developed for more realistic applications.
4.5 Determination of the characteristic emission parameters
The characteristic parameters for physically based models discussed in the previous section include the material-phase diffusion coefficient D, or effective diffusion coefficient for porous material cases, the material/air partition coefficient K, the initial emittable concentration C0, and the external convective mass-transfer coefficient hm. The usefulness of a model is greatly determined by availability of the parameters (Guo, 2002b) and the accuracy of prediction is largely dependent on reliable model parameters (Huang and Haghighat, 2003). Therefore, reliable estimation of these model parameters is prerequisite to applying these models. Before feasible experimental approaches were available, a least-squares regression technique was often employed to estimate model parameters by fitting the emission models to emission chamber test data to obtain the ‘best-fitting’ model parameters. He and Yang (2005) concluded that the regression approach risks the problem of multi-solution when more than one unknown parameters need to be determined, and when this degree of freedom increases, the uncertainties of regression become larger. Therefore it is more rigorous to obtain the model parameters from independent approaches. In this section, primary approaches for determining D, K and C0 are reviewed, while hm is generally obtained based upon heat-mass transfer analogical empirical relations (Kays and Crawford, 1980).
4.5.1 Determination of the diffusion coefficient (D)
Chamber for laboratory investigation of materials, pollutions and air quality (CLIMPAQ) method
The CLIMPAQ method can test several VOCs simultaneously (Meininghaus et al., 2000). The test material is placed between two CLIMPAQs (Gunnarsen et al., 1994), called the primary and secondary chambers. Air flow with a desirable concentration of VOCs and a clean air stream is led into the primary and secondary chambers, respectively, and VOC concentrations in the supply and the exhaust air of each chamber are measured. At steady state, the mass flux through the testing material is equal to the mass leaving the secondary chamber, or
where d is the material thickness, Qsecondary is the flow rate of supply clean air in the secondary chamber, and yprimary and ysecondary are the steady-state concentrations of the VOC in the primary and secondary chambers, respectively. Besides, considering the mass balance from the beginning of the experiment until the steady state is reached, the absorbed mass of VOC into the testing material when the steady state is reached, mab, can be calculated by numerical integration based on discrete measurements of exhaust concentrations:
where Qprimary is the flow rate of supply air in the primary chamber, ysupply is the supply air concentration in the primary chamber, yprimary_i and ysecondary_i are the ith measurements of the exhaust concentration of the primary and secondary chamber, respectively, and Δti, stands for the time interval between single measurements. Assuming that the testing material is free of VOCs at the beginning and the concentration gradient in the material is linear when the steady state is reached, K can be calculated by
which can be combined with equations 4.8 and 4.9 to determine D and K.
The CLIMPAQ method is in fact a specific representative of the steady-state twin chamber method, which can also be achieved using other types of chambers (Meininghaus and Uhde, 2002; Xu et al., 2009; Farajollahi et al., 2009). The main concern about this method comes from neglecting external convective mass-transfer resistance at the material surfaces when constructing equation 4.8, which may result in underestimation of D (Haghighat et al., 2002).
Microbalance method
A method which could measure D and K separately was developed by Cox et al. (2001a) for vinyl flooring and later successfully applied for polyurethane foam (Zhao et al., 2004), polystyrene foam and oriented strand board (Yuan et al., 2007b). The testing material can be cut into a thin slab and a high-resolution dynamic microbalance is used to monitor the mass of the slab sample. Before testing, the sample is conditioned by sweeping clean air over it until its weight is stable, implying there is no VOC present in the sample. When the experiment begins, an air flow with a known concentration of VOC is passed over the sample for the sorption test, during which sorption of VOC increases its mass until partition equilibrium is reached between the sample and air. Then clean air is passed over the sample again for the desorption test, during which desorption of VOC from the sample decreases the mass of the sample until the sample is clean. Figure 4.7 shows the transient mass gain/loss of a polymeric material sample during the sorption/desorption test.
Based on microbalance test data, K can be determined directly by
where Cequ is the equilibrium concentration in the sample, which can be calculated by dividing the mass increase at the end of the sorption test by the sample volume; and ysorp is the corresponding concentration of the VOC in the air for the sorption test. D is determined by fitting a Fickian diffusion model for a thin slab to experimental sorption and desorption data, which is given by (Crank, 1975)
where Mt is the total mass of the VOC analyte that has entered or left the slab sample in time t, M∞ is the corresponding quantity when equilibrium is reached, and 2L is the thickness of the slab sample. The samples used for the microbalance method are quite small so that the D and K obtained may not be representative for non-uniform materials. Ignoring external convective mass-transfer resistance may also lead to the same problem as for the CLIMPAQ method.
C-history method
An innovative method which can determine D, K and C0 simultaneously was recently developed by Xiong et al. (2011). The test material with uniform distribution of target VOC has to be tested in an airtight chamber for emission profiles. The chamber concentration development, y(t), and the equilibrium chamber concentration, yequ, can be described by an emission model as shown by Fig. 4.6, with chamber flow rate set to be zero. Based on the model, it is found that the logarithm of the dimensionless excess concentration, which is defined as (yequ − y(t))/yequ, is linearly dependent on the emission time, or
where SL and INT are the slope and intercept of the linear relationship, respectively. Furthermore, SL and INT are functions of both D and K as well as the geometries of the tested material slab and the airtight chamber, while yequ is dependent on C0 and K. Therefore, the measured chamber concentration over time can be treated as a form of the logarithm of the dimensionless excess concentration as in equation 4.13 to determine SL and INT by a linear regression over time. Then D and K can be calculated from SL and INT, and C0 can be further determined from yequ and the known value of K. This so-called C-history method has been well employed to determine the D, K and C0 values of formaldehyde in several building materials (Xiong et al., 2011). This promising method could determine three parameters simultaneously under realistic environmental conditions and is quite convenient and time-efficient to employ. The C-history method could be also extended to measure the equivalent emission characteristic parameters of formaldehyde and other VOCs from wood-based furniture (Yao et al., 2011). Table 4.3 summarizes the results using the C-history method for several materials (MDF stands for medium-density fiberboard, and PB stands for particle board).
Porosity test method
Different from the methods above, the porosity test method calculates the diffusion coefficient based on the microstructure of porous materials. In the porous material, molecular, Knudsen and surface diffusion may exist simultaneously. The effect of molecular and Knudsen diffusion can be combined into an overall effective diffusion coefficient, D, which can be calculated as:
where D0 is the mean (reference) diffusion coefficient in the pores of the material, 
is the porosity of the material, and τ is the tortuosity factor of the material. The porosity and tortuosity factors of the building material can be obtained using the mercury intruding porosimetry (MIP) test an the model developed by Carniglia (1986). Therefore, the key to using the porosity test method becomes how to determine D0. Blondeau et al. (2003) developed a parallel pore model, which calculates D0 by a series of sums over every intruding volume difference in the experiment. This is equivalent to treating the pores as many parallel connections in the representative elementary volume (REV) and applies an arithmetic mean method for them. Seo et al. (2005) used a mean pore model to calculate D0 by reducing all kinds of pores into an equivalent pore and using the formula depicting the transitional diffusion zone. In many cases, the parallel or mean assumption cannot reflect the essence of diffusion in porous building materials well, leading to large uncertainties in D0. Based on detailed microstructure analysis of the porous materials, Xiong et al. (2008) developed a macro-meso two-scale model to determine D0. This model categorizes all the pores into macro and meso pores according to the pore diameter, and considers the two kinds of pores to be connected in series. For the macro pores, the diffusion belongs to molecular diffusion, while for the meso pores, transition diffusion applies. Assuming the macro and meso pores have the same tortuosity factor τ, D0 can be determined based on Fick’s law (Xiong et al., 2008). Knowing D0, equation 4.14 can be used to calculate D.
4.5.2 Determination of the partition coefficient (K)
The partition coefficient K describes the thermodynamic status of equilibrium between the material phase and the gas phase. As shown in the previous section, several methods can actually determine both D and K and therefore are not repeated here. The headspace method is simple and straightforward for determining K by measuring the material-phase concentration and the gas-phase concentration respectively in the equilibrium state (Zhang et al., 2007). Analytical methods for measuring gas-phase concentrations of VOCs are very well developed, while methods for measuring the material-phase concentration will be presented in the following section. Furthermore, several delicate methods which employ the basic idea of the headspace method to measure K and C0 simultaneously have been developed recently (Wang et al., 2008; Wang and Zhang, 2009; Xiong et al., 2009).
Multi-injection regression method
In the multi-injection regression method (Wang et al., 2008; Wang and Zhang, 2009), the building material sample with initial emittable concentration C0 is placed in a temperature-controlled airtight chamber. A certain amount of the target VOC is injected into the chamber several times, allowing the sample and chamber air to reach equilibrium after each injection.
For the ith equilibrium state, based on mass conservation and Henry’s linear sorption law, it can be derived that
where C1,i is the equilibrium gas-phase VOC concentration before injection; Ch,i and C2,i are the peak and equilibrium gas-phase VOC concentrations after injection, respectively; and Vm is the building material volume. Equation 4.15 practically defines a linear relationship between the term on the left-hand side, which can be calculated from measured gas-phase concentrations, and the equilibrium gas-phase concentration, C2,i. Therefore, after several discrete injections and corresponding equilibrium states, a linear regression can be carried out to determine K and C0 from the slope and intercept. The accurate measurement of gas-phase concentrations is critical to obtaining reliable K and C0. However, the mass transfer between the chamber air and the material brings large uncertainty to the determination of Ch,i and therefore may reduce the reliability of the regression results.
Multi-emission/flush regression method
To overcome the drawbacks of the multi-injection method, an improved method was developed (Xiong et al., 2009; Xiong and Zhang, 2010). This method also involves multiple equilibrium cycles but entails flushing the chamber once the sample and chamber air have reached equilibrium in the airtight chamber. At each equilibrium state, the relationship between the equilibrium gas-phase concentration, Ca,i, and the equilibrium materialphase concentration, Cm,i, conforms to Henry’s law, i.e., Cm,i = KCa,i. For the ith cycle, it can be deduced that
Therefore, after several emission/flushing cycles, K and C0 can be determined from the linear regression of measured equilibrium gas-phase concentration and i, according to equation 4.17.
4.5.3 Determination of the initial emittable concentration (C0)
Traditional methods for measuring concentrations of VOCs in solid materials have used solvents or heat to extract target compounds. However, as discussed by Cox et al. (2001b), total VOC concentration can be apportioned to mobile and partially immobilized fractions. C0 used in the emission models should be the concentration of the readily emitted compound or the mobile fraction, which cannot be distinguished by solvent or heat extraction methods. Cox et al. also found that the emittable concentration of several VOCs in vinyl flooring measured by a new CM-FBD method is 30–70% lower than by a direct thermal desorption method. Experimental measurements of emittable formaldehyde concentrations in building materials using a multi-injection regression method (Wang and Zhang, 2009) are also much lower than the total concentrations measured by a thermal extraction method. Recently, it is reported that the emittable concentration of formaldehyde in building materials increases significantly with increasing temperature (Xiong and Zhang, 2010; see Section 4.6.1). Therefore, it is necessary to measure C0 under environmental conditions resembling those of the actual indoor environment. In addition to the multi-purpose methods discussed above, two valuable methods measuring C0 exclusively are discussed here, both of which actually aim to extract all the emittable analytes from the test materials under moderate environmental conditions and to measure the total amount.
Cryogenic milling – fluidized bed desorption (CM–FBD) method
As discussed previously, internal diffusion controls the emission rate quite often so that complete emission of VOCs from building materials may require a very long time. Therefore, Cox et al. (2001b) pulverized randomly selected small vinyl flooring samples into small particles in a ball mill under a liquid nitrogen bath at a temperature of − 140°C. This cryogenic milling (CM) process reduces the diffusion path lengths and increases emission surface areas, reducing the VOC extraction time, while the low temperature significantly reduces VOC vapor pressure, minimizing VOC loss during the grinding process. Then the extraction of VOCs from the milled particles can be accomplished at room temperature by fluidized-bed desorption (FBD). The particles were placed in a fluidized-bed reactor ventilated continuously with clean air, which extracted VOCs out of the particles. The effluent concentration was measured at suitable time intervals until all the analytes were extracted and the total amount of emitted analytes can therefore be determined based on the air flow rate. The combination of CM and FBD accelerates the extraction greatly under room temperature; however, the experimental system is complicated and there is some concern that the properties of the materials may change when ground into particles.
Multi-flushing extraction method
For the multi-flushing extraction method, the building materials were first ground into powder and then placed in an airtight chamber, and the subsequent procedures were similar to those of the multi-emission/flush regression method (Smith et al., 2008). The test continues until the equilibrium chamber air concentration of the last cycle is less than 10% of that of the first cycle so that most of the VOCs are extracted from the materials. Therefore, a series of equilibrium chamber air concentrations can be obtained during the entire process, and the initial emittable mass, Mt, can be calculated by
The first term on the right-hand side is the amount of VOCs emitted during the entire experimental period, and the second term estimates the mass remaining in the material after the experiment, where V is the volume of the chamber, Ca,i is the equilibrium VOC concentration in air for cycle i, and n is the last cycle of the experiment. Since this method requires many cycles to reach the required condition for the equilibrium concentration of the last cycle, a very long experimental period is needed.
4.6 Influence of environmental factors on emissions of formaldehyde and other volatile organic compounds (VOCs)
4.6.1 Temperature
The temperature generally has significant impacts on the emissions of formaldehyde and other VOCs from building materials. The research on this topic can be classified into two categories: (1) direct study of the impact of temperature on the emission rate or chamber concentration; and (2) analysis of the impact of temperature on the emission characteristic parameters (initial emittable concentration, diffusion coefficient and partition coefficient).
The increase of emission rate or chamber concentration with increasing temperature has been frequently reported. Andersen et al. (1975) observed that the emission rate of formaldehyde from chipboard was doubled for every 7°C temperature rise within the temperature range 14–31°C, and a relationship was formulated between the chamber concentration and temperature and other environmental factors. Myers (1985) reported an exponential relationship between formaldehyde emission rate from wood-based products and temperature, with the emission rate from particle board increasing by a factor of 5.2 when the temperature increased from 23 to 40°C. The experimental results of Lin et al. (2009) suggested that when the temperature increased from 15 to 30°C, the VOC specific emission rates and chamber concentrations increased by 1.5–12.9 times. The same phenomenon was also found by Crawford and Lungu (2011) when studying the styrene emission from a vinyl ester resin thermoset composite material. However, little or negligible effect of temperature on the emissions was found for some VOCs in several materials (Sollinger et al., 1994; Wolkoff, 1998; Wiglusz et al., 2002) and an adverse trend was also reported (Haghighat and Bellis, 1998). The reason for these phenomena is unclear. A possible explanation is different interaction patterns between different types of materials and VOCs (Wolkoff, 1998).
As far as the impact of temperature on the emission characteristic parameters is concerned, many interesting results are obtained. The first attempt at experimentally studying the impact of temperature on the initial emit-table concentration (C0) of formaldehyde in medium-density fiberboard was made by Xiong and Zhang (2010). It was observed that C0 increased significantly with increasing temperature. When the temperature increased by 25.4°C, C0 increased by about 507%, as shown in Fig. 4.8. However, the C0 at room temperature is far less than the value measured by the perforator method recommended by European standard EN 120 (1993) or Chinese national standard GB/T 17657–1999 (1999) which measures the total concentration of formaldehyde in the material. A possible reason is that there is an interaction force between formaldehyde and the material matrix and only the formaldehyde molecules with sufficiently high kinetic energy can overcome the interaction force (binding force) and emit from the material. This part of formaldehyde forms the C0, which is obviously smaller than the total amount (Xiong and Zhang, 2010).
Based on the Langmuir equation (1918), a theoretical correlation between the partition coefficient (K) and temperature (T) is established as (Zhang et al., 2007):
where P1 and P2 are constants for a given material–VOC pair. This correlation improves the equation proposed by Goss and Eisenreich (1997) by means of analyzing the experimental data when studying the VOC sorption process. Good agreement between equation 4.19 and experimental data of formaldehyde emission from some building materials is obtained (Zhang et al., 2007), which indicates a reduction of partition coefficient with increasing temperature. This is reasonable, because the desorption rate increases more rapidly than the adsorption rate when temperature increases, which leads to the decrease of the partition coefficient. However, some inconsistent results also emerged, while the reasons are unclear (Zhang et al., 2002).
An Arrhenius-like behavior is sometimes used to describe the temperature dependence of the diffusion coefficient (D) (Yang et al., 1998). However, the correlation lacks solid theoretical foundations. For VOC diffusion in porous building materials, when the molecular diffusion is dominant, a theoretical correlation can be applied to describe the temperature impact on D (Deng et al., 2009):
where B1 and B2 are constants for a given material–VOC pair. This correlation agrees well with the experimental data (Deng et al., 2009), which reveals a gradual increase in D when temperature increases. The parameters B1 and B2 can be obtained by fitting equation 4.20 to experimental data. Then the correlation can be used to predict the diffusion coefficient at other temperatures.
4.6.2 Relative humidity
The impacts of relative humidity on emissions of formaldehyde and other VOCs from building materials are more complicated. Relative humidity may affect the emission rate and the emission characteristic parameters. Table 4.4 summarizes some typical results from the literature. When relative humidity increases, the emission rate or chamber concentration would increase significantly for some material–VOC pairs. However, for some other cases, the change is negligible, and a decreasing trend is even found for a specific case (Wolkoff, 1998). The results indicate that the impact of relative humidity depends on the types of building materials and VOCs. Broadly speaking, possible causes of enhanced emission at a higher relative humidity may include smaller sorption capacity or increased generation of VOCs due to hydrolysis (Xu and Zhang, 2011). But the detailed mechanism of relative humidity effect is still not clear.
The relative humidity seems to have no significant impact on the diffusion coefficient, as shown in Table 4.4, while the partition coefficient is dependent on relative humidity. This may be due to the competition of adsorption sites between water vapor and VOCs. In addition, the capillary condensation phenomenon can also cause some differences in partition coefficient. In Table 4.4, the increase of partition coefficient of formaldehyde was probably because some formaldehyde molecules are absorbed by the adsorbed water under a high relative humidity level, while the decrease of partition coefficient of toluene was possibly due to the competition of water molecules for available adsorption sites with toluene molecules (Xu and Zhang, 2011).
4.6.3 Air velocity
The air velocity across the surface of building materials will affect the convective mass transfer coefficient (hm) of formaldehyde and other VOCs. Higher air velocity can decrease the boundary layer thickness and increase hm, and therefore facilitate the emissions. From the mass transfer point of view, the significance of this impact is dependent on the relative magnitude between the convection effect and the diffusion effect. If the latter is dominant, the impact of air velocity can be neglected. For most emission processes, the air velocity has some impacts on the initial-period emissions but insignificant impacts on the long-term emission. Neglecting convective mass transfer coefficient, i.e., taking hm as infinite, tends to overestimate emission rate and chamber concentration at the initial stage (Xu and Zhang, 2003). Studies for VOC emissions from five commonly used materials have shown that the air velocity does not influence the emission rate after a few days to a week to any great extent (Wolkoff, 1999). For some chambers, the air velocity can be controlled by the mixing fan inside the chamber, while for other chambers that have no special design for controlling air velocity over the material surface, the air velocity in the test chamber is directly related to the chamber flow rate (Zhang et al., 2002).
It should be pointed out that some other factors, such as the material use age, the physical and chemical properties of the materials and VOCs, the VOC concentration in the chamber, and the state of the material (dry or wet), can also influence the emission characteristics. Further experimental and theoretical work is necessary to clarify these questions.
4.7 Conclusion and future trends
Due to its wide distribution and high incidence of contamination, such as the serious formaldehyde contaminations in government-provided trailers for victims of hurricane Katrina in the US, the toxicology of formaldehyde has been the subject of much concern. As reviewed in this chapter, the toxicokinetics and health effects caused by exposure to formaldehyde are quite well understood and some biochemical understanding of the toxicity mechanisms has been acquired. Some other common VOCs, such as benzene, have also been extensively studied for their toxicology. However, as reviewed by European project HealthyAir, the toxicological data are still very poor or lacking for most VOCs, and knowledge about the health effects due to inhalation exposure to them is very limited, especially those caused by simultaneous exposure to various VOCs (Bluyssen et al., 2010). In addition, guidelines or recommendations for indoor VOC concentrations are unavailable for most VOCs. Further investigations into the toxicokinetics and health effects of individual VOCs at indoor-relevant concentration levels as well as those of combinations of common VOCs are therefore urgently required. As suggested by the US EPA ToxCast and ToxPi projects and the European Community Regulation on chemicals and their safe use (REACH), a feasible strategy for further studies of the health effects and risks of VOCs in indoor air should include prioritizing individual compounds or classes of compounds and combining knowledge about the nature of VOC, the degree of human exposure, and toxicology (Judson et al., 2010; Reif et al., 2010; Rudén and Hansson, 2010).
Appropriate ventilation has been considered the primary means of achieving good indoor air quality, and discussions on how much ventilation is sufficient to prevent noxious odors and the spread of disease started at the beginning of the nineteenth century and are still going on (Fig. 4.9). It should be pointed out that the minimum ventilation rates do not prevent occupants of a space developing health symptoms, and the threshold levels for all compounds seem unrealistic due to the numerous compounds and complex mixtures of compounds (Bluyssen, 2009). Generally speaking, a sufficient ventilation rate is effective in removing indoor air pollutants, including VOCs originating indoors. Given the prevailing trend of airtight buildings for less energy use, some energy-saving ventilation strategies have been developed (Laverge et al., 2011). However, source control is regarded as a more straightforward and probably better approach: preventing rather than curing (Bluyssen, 2009). Experimental techniques for emission testing and understanding of emission mechanisms of VOCs presented in this chapter are therefore highly valuable for source control. It seems the overall mechanisms of emissions from individual homogeneous sources are quite well understood and the emission models reviewed above are capable of predicting emissions with good accuracy. The emission models can be used to characterize the source-to-effect continuum for VOCs by linking source types, emissions, transport in indoor air, exposure, and toxic effects. However, there are still many gaps in our understanding of indoor VOC emissions.
Fig. 4.9 Changes in the recommended minimum ventilation rate over the years. (Bluyssen, 2009)
For example, the emission characteristics of sources are much more complicated in the real indoor environment with several sources and sinks present at the same time; indoor and surface chemistry could generate secondary sources; and the impacts of various environmental factors on the emissions are not fully understood (Bluyssen, 2009). Therefore, further studies on the emission mechanisms and indoor chemistry for more realistic scenarios are urgently needed.
Finally, as shown by the HealthyAir project which aims to define, initiate and develop activities that will improve indoor air quality and reduce exposure of people to indoor air pollution, the awareness and education of manufacturers, architects, and end-users about the problems of indoor VOCs are generally poor (Bluyssen et al., 2010). Therefore, in addition to further research on the toxicology of VOCs and emission characteristics of indoor sources, better education of stakeholders and appropriate policies and regulations (such as regulations of indoor VOC concentrations, product emission standards, and product labeling programs) are urgently needed to solve indoor VOC problems and improve indoor air quality.
4.8 Sources of further information
Suggested general standards or books on formaldehyde and other VOCs are as follows.
ASTM D5197-03. Standard test method for determination of formaldehyde and other carbonyl compounds in air (active sampler methodology), 2003. [American Standard].
ASTM D5466-01. Standard test method for determination of volatile organic chemicals in atmospheres (canister sampling methodology), 2001. [American Standard].
ASTM D6007-02. Standard practice for full-scale chamber determination of volatile organic emissions from indoor materials/products, 2002. [American Standard].
ASTM D7143-05. Standard practice for emission cells for determination of volatile organic emissions from indoor materials/products, 2005. [American Standard].
GB/T 17657-1999. Test methods of evaluating the properties of wood-based panels and surface decorated wood-based panels, 1999. [Chinese National Standard].
EN 120. Wood-based panels – determination of formaldehyde content – extraction method called perforator method, 1993. [European Standard].
EN 717-1. Wood-based panels – determination of formaldehyde release – Part 1: Formaldehyde emission by the chamber method, 2004. [European Standard].
ISO 16000-6. Indoor air – Part 6: Determination of volatile organic compounds in indoor and test chamber air by active sampling on Tenax TA sorbent, thermal desorption and gas chromatography using MS/FID, 2004. [International Organization for Standardization].
ISO 16000‐10. Indoor air – Part 10: Determination of the emission of volatile organic compounds from building products and furnishing – Emission test cell method, 2006. [International Organization for Standardization].
ISO/CD 12460. Wood-based panels. Determination of formaldehyde release – formaldehyde emission by the chamber method, 2007. [International Organization for Standardization].
JIS A 1460. Building boards. Determination of formaldehyde emission-desiccator method, 2001. [Japanese Industrial Standard].
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