Exposure Data

IARC Working Group on the Evaluation of Carcinogenic Risks to Humans

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IARC Working Group on the Evaluation of Carcinogenic Risks to Humans. Household Use of Solid Fuels and High-temperature Frying. Lyon (FR): International Agency for Research on Cancer; 2010. (IARC Monographs on the Evaluation of Carcinogenic Risks to Humans, No. 95.)

Household Use of Solid Fuels and High-temperature Frying.

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1Exposure Data

1.1. Definition

‘Cooking fumes’ or ‘cooking oil fumes’ is the term commonly used to describe the visible emissions generated during cooking by frying with oil. However, these emissions are not technically ‘fumes’. In occupational and environmental hygiene, ‘fumes’ are defined as submicron-sized solid particles (particulate matter) created by the cooling of hot vapour. During cooking, such vapour is formed when the cooking oil is heated above its boiling point. In addition to this ultrafine particulate matter, cooking, especially frying and grilling, generates aerosol oil droplets, combustion products, organic gaseous pollutants, and steam from the water contents of the food being cooked.

1.2. Constituents of cooking fumes

Cooking, in particular frying, generates substantial amounts of airborne particulate matter (PM), which includes ultrafine particles (UFP) and fine PM (PM_2.5), and is a major contributor to their indoor levels. In addition, particles created during cooking have organic substances adsorbed on their surface. These include polycyclic aromatic hydrocarbons (PAHs) and heterocyclic amines. Certain gaseous pollutants such as formaldehyde (IARC, 2006), acetaldehyde (IARC, 1999), acrylamide (IARC, 1994) and acrolein (IARC, 1995) are also produced during cooking.

The concentrations of these constituents measured in cooking fumes in field and controlled studies are presented below.

1.2.1. Ultrafine and fine particulate matter

The Particle Total Exposure Assessment Methodology (PTEAM) study was carried out by the Research Triangle Institute and the Harvard University School of Public Health in the USA in 1989–90 (Clayton et al., 1993). Particle concentrations were measured for a probability-based sample of 178 nonsmokers who represented the non-institutionalized population of Riverside, CA (∼139 000 persons). Personal samples of PM₁₀ were taken; the indoor and outdoor samples included both PM₁₀ and PM_2.5. Cooking produced both fine and coarse particles. Homes where cooking took place during monitoring (about 55%) had average PM₁₀ concentrations ∼20 µg/m³ higher than those where no cooking took place (Özkaynak et al., 1996a,b). The proportion of PM_2.5 and PM₁₀ due to cooking was 25% for both particle sizes (Figure 1.1). However, when considered as a fraction of particles due to indoor sources alone, the proportion was 65% and 55%, respectively (Özkaynak et al., 1996b).

Figure 1.1

Fraction of PM_2.5 due to cooking (top); fraction of PM₁₀ due to cooking (bottom).

A large-scale study of personal, indoor and outdoor exposures was undertaken for more than 100 persons living in Seattle, WA, USA (Seattle Study; Liu et al., 2003). Based on 195 cooking events, the average PM_2.5 concentration due to cooking was estimated to be 5.5 (standard error [SE], 2.3) µg/m³ (Allen et al., 2004).

A study of personal, indoor and outdoor exposure to PM_2.5 and associated elements was carried out on 37 residents of the Research Triangle Park area in North Carolina, USA (Research Triangle Park Study; Wallace et al., 2006a,b). Burned food added an average of 11–12 µg/m³ to the indoor concentration (Wallace et al., 2006b). In continuous measurements, the mean estimated PM_2.5 personal exposures during more than 1000 h of cooking were found to be 56 µg/m³ higher than background (Wallace et al. 2006b). The 24-h average increase due to cooking was about 2.5 µg/m³. A different analysis of the results from this study concluded that cooking contributed 52% of personal exposure to PM_2.5 and more than 40% of the indoor concentration of PM_2.5 (Zhao et al., 2006).

A long-term study of indoor and outdoor particle concentrations was carried out between 1997 and 2001 in an occupied townhouse in Reston, VA, USA (Reston, VA Townhouse Study). Cooking produced about an order of magnitude higher number of the smallest UFP (10–50 nm) and from 1.2- to 9.4-fold higher levels of the larger particles compared with identical times when no cooking occurred (Table 1.1; Wallace et al., 2004). The mean mass concentration increased at dinner (4-h averages) from 3.7 µg/m³ to 11.8 µg/m³ assuming a density of combustion particles of 1 g/cm³. About 70% of the particles emitted during dinnertime were <0.05 µm.

Table 1.1

Number and concentration of PM_2.5 during dinnertime cooking compared with no cooking.

In a more detailed analysis, 44 high-particle-production (frying, baking, deep-frying) cooking episodes on a gas stove were assessed (Wallace et al., 2004). Most of the particles were in the ultrafine range, but the largest volume was contributed by particles between 0.1 µm and 0.3 µm in diameter. The total particle volume concentration created by the 44 high-particle-production cooking events averaged a little more than 50 (µm/cm)³, corresponding to an average concentration of about 50 µg/m³, about an order of magnitude higher than average values for all types of cooking combined.

The size distribution of ultrafine particles during cooking was studied by Wallace (2006) and Ogulei et al. (2006). Stir-frying using one gas burner produced a peak of PM ∼35 nm, whereas deep-frying using one gas burner followed by baking in the oven produced a peak about twice as high and at a diameter of 64 nm (Figure 1.2).

Figure 1.2

Size distribution of ultrafine particles from cooking. n = number of 5-min measurements. Error bars are standard errors. Stir-frying on one gas burner produced a peak at ∼35 nm; deep-frying on one gas burner followed by baking in the oven produced (more...)

Brauer et al. (2000) reported PM_2.5 concentrations in the range of 24–201 g/m³ in residential kitchens during frying, with peak PM_2.5 concentrations above 400 µg/m³. Kamens et al. (1991) estimated that 5–18% of an 8-h personal particle exposure could be attributed to cooking one meal in one of three homes that they studied.

Abt et al. (2000) studied 17 selected cooking events in three homes that provided mean peak volume concentrations of particles between 20 and 500 nm ranging between 29 and 57 (µm/cm)³. Long et al. (2001) studied nine homes for 6–12 days each and found mean peak volume concentrations for UFP (20–100 nm) of 2.2–18.2 (µm/cm)³. He et al. (2004a) studied 15 homes for 48 h during cooking under good and poor ventilation conditions and found a range of peak submicrometer number concentrations for cooking events between 16 000 and 180 000 particles/cm³. Estimates of the emission rate ranged between 0.2–4 × 10¹² particles/min. Finally, 24 cooking events with high concentrations and well-shaped decay curves, including concurrent air exchange rate measurements, were analysed more accurately, taking into account losses due to deposition during the lag time required to reach the peak, for their source strengths (Wallace et al., 2004). A value of 3×10¹² UFP/min was obtained.

A study in Amsterdam and Helsinki found that cooking increased PM_2.5 concentrations by 1.9–3.4 µg/m³ (14–24%) among two groups of 47 and 37 elderly residents in the two cities, respectively (Brunekreef et al., 2005; the ULTRA Study).

Kleeman et al. (1999) used an industrial charbroiling facility to cook >100 hamburgers. The particle mass consisted mainly of organic compounds, with a very small amount of elemental carbon, and a large unknown component. Most of the particle mass came from particles between 0.1 and 0.4 µm in diameter.

Emission rates during cooking with commercial institutional-scale deep-fryers have been reported (Schauer et al., 1998). Professional chefs prepared vegetables by stir-frying in soya bean or canola oil and deep-frying potatoes in oil. Fine particle emission rates were 21.5±1.2, 29.5±1.3 and 13.1±1.2 mg/kg for stir-frying vegetables in the two oils and deep-frying potatoes, respectively. [Emissions during food preparation by a professional chef using large commercial cookers may differ substantially from emissions in a residence.]

In a recent study in a residential setting in Canada (Evans et al. 2008), real-time measurements were taken during frying to estimate the time-integrated exposure to PM associated with frying food. The production rates and concentrations of UFP and PM_2.5 during and at the end of frying a variety of breakfast foods typical of the Canadian diet at medium temperatures were assessed (Table 1.2).

Table 1.2

The production rates and concentrations of UFP and PM_2.5 during and at the end of frying of various types of foods.

1.2.2. Volatile organic compounds

A large proportion of the vapours generated during cooking is steam from the water contents of the food or from the water used to cook the food. However, during frying (with oil), fatty acid esters that are constituents of edible oils and fat can decompose and produce volatile organic compounds, as well as semi-volatile compounds that can condense to form particles. A wide variety of organic compounds have been identified in cooking emissions, including alkanes, alkenes, alkanoic acids, carbonyls, PAHs and aromatic amines. Felton (1995) reported that the main volatile compounds generated during frying were aldehydes, alcohols, ketones, alkanes, phenols and acids. Of particular concern in relation to carcinogenicity are PAHs, heterocyclic amines and aldehydes.

(a) PAHs

Dubowsky et al. (1999) reported peak total particle-bound PAH concentrations in a range from undetectable to 670 ng/m³ during cooking when measured with a Gossen PAS monitor.

A study in Taiwan found several PAHs in the fumes of three cooking oils (safflower, vegetable and corn oil) (Chiang et al., 1999a).

By contrast, Wallace (2000) did not measure increased concentrations of total PAHs during cooking.

(b) Aldehydes

Schauer et al. (1998) reported emissions of 20 100 µg formaldehyde/g of food during stir-frying of vegetables on an institutional-size cooker. They reported emissions of 12 400 µg/g formaldehyde and 20 900 µg/g acetaldehyde during deep-frying of potatoes.

(c) Aromatic amines

One study found the aromatic amines 2-naphthylamine and 4-aminobiphenyl in the fumes of three different cooking oils (sunflower oil, vegetable oil and refined lard) (Chiang et al., 1999b).

(d) Other volatile compounds

Rogge et al. (1991) measured the fine aerosol emission rates for single organic compounds from charbroiling and frying hamburger meat. The compounds detected were n-alkanes, n-alkanoic acids, n-alkenoic acids, dicarboxylic acids, n-alkanals and n-alkenals, n-alkanones, alkanols and furans.

Ho et al. (2006) studied emissions of 13 carbonyl compounds in cooking exhaust fumes from 15 restaurants in Hong Kong Special Administrative Region, China, and developed a new method of analysis using Tenax coated with a hydrazine compound followed by thermal desorption and mass spectrometry. This allowed them to separate three similar compounds: acetone, acrolein and propanal. The most prevalent compounds were formaldehyde (in all but four of the restaurants), acrolein, acetaldehyde and nonanal, which accounted for 72% of all carbonyl emissions. Based on a small sample of restaurants, the authors estimated total annual emissions for acrolein, formaldehyde and acetaldehyde of 7.7, 6.6 and 3.0 tonnes per year from cooking compared with 1.8, 10 and 33 tonnes per year, respectively, from vehicles.

1.3. Effect of different parameters of cooking on emissions

The chemical composition of cooking emissions varies widely depending on the cooking oils used, the temperature, the kind of food cooked, as well as the method and style of cooking adopted.

1.3.1. Effect of the type of oil and temperature

(a) Mixture of volatile components

Studies were undertaken to identify qualitatively the volatile components emitted during the heating of cooking oils to 265–275°C (Li, et al. 1994; Pellizzari et al. 1995; Shields et al. 1995; Chiang et al., 1999a; Wu et al. 1999). The oils tested were rapeseed, canola, soya bean and peanut. The major constituents identified in the oil vapours were saturated, unsaturated and oxygenated hydrocarbons. These studies detected a variety of agents in emissions from heated cooking oils including 1,3-butadiene, benzene, benzo[a]pyrene, dibenz[a,h]anthracene, acrolein, formaldehyde and acetaldehyde. Emissions were highest for rapeseed oil and lowest for peanut oil. In one study, the emission levels of 1,3-butadiene and benzene were approximately 22-fold and 12-fold higher, respectively, for rapeseed oil than for peanut oil (Shields et al., 1995). Compared with rapeseed oil heated to 275°C, fourfold and 14-fold lower levels of 1,3-butadiene were detected when the oils were heated to 240°C and 185°C, respectively.

(b) PAHs and nitro-PAHs

In a study performed in a controlled environment (Air Resources Board of the State of California Study; Fortmann et al., 2001), five untreated cooking oils were extracted and analysed for PAHs (Table 1.3). All were found to contain some PAHs; olive oil and peanut oil contained generally higher concentrations than rapeseed, corn or vegetable oils.

Table 1.3

Concentrations (ng/g) of polycyclic aromatic hydrocarbons in untreated cooking oils.

In a similar study, PAHs levels in samples of five raw cooking oils (canola, olive, corn, soya bean and vegetable oil) were not increased compared with the blank (Kelly, 2001).

Fume samples from three different commercial cooking oils commonly used in Taiwan, China (lard oil, soya bean oil and peanut oil), were collected and tested for PAHs. All samples contained dibenz[a,h]anthracene and benz[a]anthracene; extracts of fume samples from the latter two also contained benzo[a]pyrene (Chiang et al., 1997). In a later study, fume samples from safflower, olive, coconut, mustard, vegetable and corn oil were similarly tested (Chiang et al., 1999a). Extracts of fumes from safflower oil, vegetable oil and corn oil contained benzo[a]pyrene, dibenz[a,h]anthracene, benzo[b]fluoranthene, and benz[a]anthracene. Concentrations are shown in Table 1.4.

Table 1.4

The polycyclic aromatic hydrocarbon contents (µg/m³) of fumes from various oils heated to 250±10°C for 30 min.

Wei See et al. (2006) studied three ethnic food stalls in a food court for levels of PM_2.5 and PAHs. PAHs varied from 38 to 141 to 609 ng/m³ at the Indian, Chinese and Malay stalls, respectively. The trend was considered to be related to the cooking temperature and amount of oil used (simmering, stir-frying and deep-frying). Frying provided relatively more high-molecular-weight PAHs compared with simmering, which produced relatively more low-molecular-weight PAHs.

In addition to PAHs, fumes from three different commercial cooking oils frequently used in Chinese cooking (lard oil, soya bean oil and peanut oil) also contained nitro-PAHs such as 1-nitropyrene and 1,3-dinitropyrene (Table 1.5) (Wu et al., 1998).

Table 1.5

Concentrations of PAHs and nitro-PAHs (µg/m³) in fumes from various oils heated to 250±10°C for 30 min.

Zhu and Wang (2003) studied 12 PAHs in the air of six domestic and four commercial kitchens. Mean concentrations of benzo[a]pyrene were 6–24 ng/m³ in the domestic kitchens and 150–440 ng/m³ in the commercial kitchens. Cooking oils were ranked lard>soya bean oil>rapeseed oil. Increases in cooking temperature produced increased PAH concentrations.

Various samples of cooking oil fumes were analysed in an effort to study the relationship between the high incidence of pulmonary adenocarcinoma in Chinese women and cooking oil fumes in the kitchen (Li et al., 1994). The samples included oil fumes from three commercial cooking oils. All samples contained benzo[a]pyrene and dibenz[a,h]anthracene. The concentration of dibenz[a,h]anthracene in the fume samples was 5.7–22.8 times higher than that of benzo[a]pyrene. Concentrations of benzo[a]pyrene and dibenz[a,h]anthracene were, respectively, 0.463 and 5.736 µg/g in refined vegetable oil, 0.341 and 3.725 µg/g in soya bean oil and 0.305 and 4.565 µg/g in vegetable oil.

(c) Heterocyclic amines

Hsu et al. (2006) studied the formation of heterocyclic amines in the fumes from frying French fries in soya bean oil or lard. Lard was more susceptible to form these compounds than soya bean oil heated alone (Hsu et al., 2006). Fumes from soya bean oil heated alone were found to contain three heterocyclic amines, namely, 2-amino-3-methylimidazo[4,5-f]quinoxaline (IQx), 2-amino-3-methylimidazo[4,5-f]quinoline (IQ) and 1-methyl-9H-pyrido[4,3-b]indole (Harman), whereas two additional amines, 2-amino-3,4-dimethylimidazo[4,5-f]quinoline (MeIQ) and 3-amino-1,4-dimethyl-5H-pyrido[4,3-b]indole (Trp-P-1), were generated with lard.

(d) Aldehydes and other volatile organic compounds

Higher aldehydes [C>7] have been detected in emissions from pan-frying beefsteak using four different types of oil (Table 1.6) (Sjaastad & Svendsen 2008). The aldehyde trans,trans-2,4-decadienal (t,t-2,4-DDE) has been found and quantified in both frying oils and fumes generated during frying. The quantity of t,t-2,4-DDE in fried potatoes was considered to be dependent on the oil used, on the frying process and, to a lesser extent, on oil deterioration. The degree of unsaturation of the frying oil was also considered to promote the formation of t,t-2,4-DDE.

Table 1.6

Levels^a of total particles (mg/m³) and higher aldehydes (µg/m³) measured in the breathing zone of the cook during pan-frying of beefsteak using different oils or margarine.

Emissions of low-molecular-weight aldehydes from deep-frying with extra virgin olive oil, olive oil and canola oil (control) were investigated at two temperatures, 180 and 240°C, for 15 and 7 h, respectively. Seven alkanals (C-2 to C-7 and C-9), eight 2-alkenals (C-3 to C-10) and 2,4-heptadienal were found in the fumes of all three cooking oils. The generation rates of these aldehydes were found to be dependent on heating temperature, and showed significant increases with increases in temperature. The emissions of low-molecular-weight aldehydes from both kinds of olive oil were very similar and were lower than those observed from canola oil under similar conditions (Fullana et al., 2004a,b).

The composition of the fumes was studied at different temperatures (190–200, 230–240 and 270–280°C). A strong peak was observed within the wavelength range of 260–270 nm in each condensate sample. From gas chromatography–mass spectrometry results, it was tentatively deduced that there were some 2,4-dialkylenaldehydes and other conjugated compounds in the condensates. Large amounts of hexanal and 2-heptenal were present in the cooking oil fumes. The total aldehyde peak areas of the condensates from four kinds of oil were around 30–50% of the total peak area at 270–280°C (Zhu et al., 2001).

Concentrations of ethylene oxide and acetaldehyde were assessed during the simulated frying of soya bean oil without or with flavouring herbs and spices (garlic, onion, ginger, basil) under nitrogen or air at 1atm (Lin et al., 2007). The tests were performed at 130, 150, 180 and 200°C.

The concentration of both ethylene oxide and acetaldehyde in the oil and vapour phases increased with frying temperature within the range of 130 to 200°C. Under air, the amounts of ethylene oxide and acetaldehyde generated in either phase were several times higher when compared with amounts generated under nitrogen. In the oil phase, concentrations of ethylene oxide and acetaldehyde increased linearly from 7.6 ppm at 130°C to 26.2 ppm at 200°C, and from 6.0 ppm to 16.6 ppm, respectively. Similarly, ethylene oxide concentrations in the vapour phase increased from 7 ppm to 85 ppm.

The impact of the combination of flavouring sources and soya bean oil was assessed. Both ethylene oxide and acetaldehyde were distributed between the gas phase and the oil phase after cooking each herb or spice at 150°C for 5 minutes under either atmosphere. In each scenario, the amounts of ethylene oxide and acetaldehyde produced were different when compared with heating soya bean oil alone.

1.3.2. Effect of the type of food, type of cooking or mode of frying

(a) Studies in a controlled environment

In an experimental study, airborne cooking by-products from frying beef (hamburgers), pork (bacon strips) and soya bean-based food (tempeh burgers) were collected, extracted and chemically analysed. 2-Amino-1-methyl-6-phenylimidazo[4,5-b]pyridine (PhIP) was the most abundant heterocyclic amine, followed by 2-amino-3,8-dimethylimidazo[4,5-f]quinoxaline (MeIQx) and 2-amino-3,4,8-trimethylimidazo[4,5-f]quinoxaline (DiMeIQx). No 2-amino-9H-pyrido[2,3-b]indole (AαC) was detected in the food samples fried at about 200°C, although it was present in the collected airborne products. The total amounts of heterocyclic amines in the smoke condensates were 3 ng/g from fried bacon, 0.37 ng/g from fried beef and 0.177 ng/g from fried soya-based food (Table 1.7) (Thiébaud et al., 1995).

Table 1.7

Concentration of heterocyclic amines from frying meat and soya-based patties (ng/g of cooked samples).

One study compared emissions of particles, nitrogen oxides, carbon monoxide, PAHs and formaldehyde in an experimental chamber during seven different types of cooking activity including pan-frying (Table 1.8; Kelly, 2001). Samples were integrated over periods of 1–4 h. [Temperatures were measured but not reported.]. Except for the hamburger cooked on gas, all tests showed an increase in total PAHs, with indoor levels averaging about twice or more the outdoor concentrations. Since the outdoor concentrations would be expected to be roughly half of those indoors in the absence of indoor sources, the increase over normal indoor levels is by a factor of about 3. For seven particle-bound PAHs that are considered to be probably carcinogenic, indoor:outdoor ratios averaged from 1–1.5. Emissions of nitrogen dioxide were found only when the gas stove was used, and were 10 mg/kg for pan-frying of hamburgers. Emissions of formaldehyde remained below 10 ppb (see footnote in Table 1.8).

Table 1.8

Concentrations of PM_2.5, PAHs and formaldehyde in a research house during pan-frying.

Another major controlled study of cooking emissions was sponsored by the Air Resources Board of the State of California (Fortmann et al., 2001). PM_2.5 and PM₁₀ particles, carbon monoxide, nitrogen oxide, nitrogen dioxide, PAHs and aldehydes were measured. Cooking activities included wok stir-frying of chicken and vegetables, deep-frying of French fries and pan-frying of bacon, tortillas or hamburgers. The cooking activities were studied under standard conditions or worst-case scenarios. Wok stir-frying was performed with 65 g peanut oil for 1 or 3 min at high temperatures, using chicken and vegetables as food. The concentrations of PM_2.5 particles emitted during the cooking activities under different conditions are given in Table 1.9. Of the 13 PAHs targeted for analysis, pyrene, benzo[e]pyrene, benzo[a]pyrene and benzo(b+j+k)phenanthrenes were detected in more than 60% of the samples. Duplicate samples collected during the worst-case stir-fry test showed that the precision of the PAH sampling method was poor. [Because of the short test, the mass of PAHs in the samples was low, and there was large analytical uncertainty associated with the measurement.]

Table 1.9

PM_2.5 concentrations under different cooking conditions in a research house.

(b) Field studies

Samples of cooking oil fumes from three catering shops were analysed (Li et al., 1994). All samples contained benzo[a]pyrene and dibenz[a,h]anthracene. PAH concentrations at the three catering shops showed levels of benzo[a]pyrene of 41.8 ng/m³ at a Youtiao (deep-fried twisted dough sticks) shop, 22.8 ng/m³ at a Seqenma (candied fritters) workshop and 4.9 ng/m³ at a kitchen of a restaurant; concentrations of dibenz[a,h]anthracene were 338, 144 and 30.3 ng/m³, respectively.

Another study in China showed that the cooking method affected the concentration of benzo[a]pyrene in kitchen air (Du et al., 1996). In the same kitchens, the level of benzo[a]pyrene was elevated in indoor air from the baseline value of 0.41 µg/100m³ to 0.65 µg/100m³ when meat was boiled, and was further increased to 2.64 µg/100m³ when meat was stir-fried.

Li et al. (2003) measured PAHs emitted from the rooftop exhausts of four types of restaurant in Taiwan, China. Although gaseous PAHs outweighed particle-bound PAHs by about 4:1, when expressed in benzo[a]pyrene-equivalents, the ratio was reversed. Chinese food contributed the majority of the level of benzo[a]pyrene-equivalents, while western food contributed about seven times less and fast food and Japanese food contributed negligible amounts. Compared with traffic in the city, restaurants contributed somewhat less total PAHs but about 10 times the benzo[a]pyrene-equivalent amount.

Zhu and Wang (2003) studied 12 PAHs in the air of six domestic and four commercial kitchens. Mean concentrations of benzo[a]pyrene were 6–24 ng/m³ in the domestic kitchens and 150–440 ng/m³ in the commercial kitchens. Cooking practices produced PAHs in the rank order broiling>frying>>boiling.

The influence of frying conditions (deep-frying, pan-frying) was studied (Boskou et al., 2006). In all cases tested, the highest concentration of trans,trans-2,4-decadienal was detected during deep-frying.

Studies have shown that the total amount of organic compounds per milligram of particulate organic matter is much higher in western-style fast food cooking than in Chinese cooking; however, Chinese cooking has a much greater contribution of PAHs to particulate organic matter (Table 1.10) (Zhao et al., 2007a, b).

Table 1.10

Concentrations of organic compounds from western-style fast food and from Chinese cooking (ng/mg of particulate organic matter).

1.4. Human exposure

Neither occupational nor non-occupational exposure to emissions from cooking has been characterized systematically. Most of the available studies examined the nature and amount of emissions produced during different types of cooking in different settings, including the release of emissions from kitchens into the ambient environment. As the substances measured varied widely among studies, it is difficult to summarize quantitatively exposures in different settings. Furthermore, co-exposures were not specifically mentioned. Results from various field studies, carried out primarily in South-East Asia, are summarized in Tables 1.11 and 1.12.

Table 1.11

Occupational exposures to emissions from high-temperature frying.

Table 1.12

Environmental exposure to cooking emissions from commercial restaurants.

Only one recent study provided information on biological monitoring of exposure and effect in the occupational setting (Table 1.11) (Pan et al., 2008).

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