The formation of naphthalene (and other PAHs) during combustion has been extensively studied. Combustion is considered to be the single largest emission source of naphthalene [2]. Fuel type, oxygen supply and temperature are the major factors affecting emissions [63]. Naphthalene formed during combustion is predominantly in the vapor phase [64], e.g., 90% was measured from burning of rice and bean straw [58].

Emission factors (EF) for many biomass fuels have been determined in the laboratory under controlled conditions (Table 2). EFs for open fires have been comprehensively reviewed [65]. For biomass fuels, these factors span over three orders of magnitude: medians are 40, 17, and 8 mg kg⁻¹ for wood, coal and crop straw, respectively. Open burning of agricultural residues is a significant source in developing countries. Burning of rice and bean straw in China, for example, emits an estimated 110 to 126 and 13 to 26 metric tons yr⁻¹ of naphthalene, respectively [58]. Emissions from wildfires also can be significant. In China, wildfires release 141 metric tons yr⁻¹, including 68% and 32% from forest and grassland fires, respectively [66]. In Africa, wildfire emissions are projected to exceed anthropogenic emissions [67]. Wildfire emissions can contribute to naphthalene levels in urban areas by long range transport [68,69]. Burning of wood and coal for domestic heating and cooking are additional sources. Overall, naphthalene EFs for biomass combustion vary considerably by region and time, and uncertainties are high.

Vehicle emissions represent an important naphthalene source in urban areas. For gasoline-powered cars and light-trucks, EFs for vehicles with and without catalytic converters are about 1 and 50 mg km⁻¹, respectively [50]. For diesel-powered vehicles, EFs are about 10, 505, 276, and 20 mg km⁻¹ in idle, creep, transient, and cruise modes, respectively [53]. Naphthalene has been found in exhausts of helicopters and ships [55–57].

Second-hand cigarette smoking is among the largest contributor to personal exposures of naphthalene. EFs of naphthalene in the side-stream smoke measured in the laboratory range from 12 to 15 μg cigarette⁻¹ for commercial cigarettes [44], and slightly higher for research cigarettes [45]. Under realistic conditions, reported EFs range more broadly, from 17 to 54 μg cigarette⁻¹ and depend on air exchange rates, furnishing level [47] and smoking condition [70]. All or most of these emissions are due to the side stream smoke [44,70]. Naphthalene EFs for cigarettes across these studies are reasonably consistent.

Mosquito coils may be an important source of naphthalene exposure. These coils are used both indoors and outdoors; in Asia, they are commonly used in sleeping areas during the night. A recent chamber study measured emissions from 4.8 to 19.5 μg h⁻¹ from burning mosquito coils [71].

A final combustion source, also potentially important indoors, is the burning of incense sticks. While very common in Asian residences and temples, EF estimates are lacking.

Naphthalene emissions during pyrolysis have been documented. The mean naphthalene concentration in the exhaust of flares from the pyrolysis of scrap tires is 150 (range of 0.11 to 543) μg m⁻³ [72]. An EF of 20.2 μg g⁻¹ was determined for the pyrolysis of liquid crystal wastes [73].

Food cooking can involve pyrolysis when organic compounds are partially cracked to small fragments, which then recombine with radicals to form relatively stable PAHs. Frying and boiling yielded 0.25 to 4.4 μg m⁻³ of excess naphthalene concentrations per kg of fish or pork chops, while boiling generated only 0.028 to 0.045 μg m⁻³ kg⁻¹ [74].

Naphthalene is a natural component in coal tar and crude oil with typical contents of 11% and 1.3%, respectively [2]. It is also a constituent of gasoline, diesel and jet fuel. In gasoline, the naphthalene content has been expressed as 1.04 mg g⁻¹ [50], 69 to 2600 mg L⁻¹ [75], and 0.15 to 0.18% (w/w) [76]. Naphthalene’s content varies greatly by brand and grade, and “premium” gasoline tends to have higher concentrations than “regular” gasoline [75]. In diesel, the naphthalene content varies from 6.6 to 1,600 mg L⁻¹ [75]. In jet fuel, the naphthalene content is 0.26% (w/w) [77].

The use of naphthalene as a moth repellent has been stated to be the second largest naphthalene exposure source after combustion [2], although its use in this application exceeds rates in the available emission inventories (described later). A variety of materials are used as “moth preventatives” in the form of balls, crystals, flakes, cakes, blocks, bars and “nuggets,” available loose and packaged in porous bags and boxes, clothes hangers, and other niceties, and sometimes scented with cedar, lavender and other fragrances. In the U.S., most of these products are made of essentially pure p-dichlorobenzene [78], which avoids the flammability hazard associated with naphthalene. However, naphthalene remains readily available, e.g., sold as “old fashioned mothballs” or flakes, except in California where it has not been registered under Proposition 65. At the federal level, naphthalene is registered for use on indoor sites as a moth repellant in the form of balls and flakes (it is not formed in blocks and other configurations). Based on industry reports [79], the U.S. moth preventative consumer sales totals $14 million annually, of which 41% can be identified as naphthalene products. However, these figures are incomplete, e.g., many merchandisers and bulk sales are excluded, and actual sales are likely multiples of this value. Also, naphthalene’s actual share will be lower since many retailers only stock items that can be sold nationally and, as mentioned, California sales for this purpose are not permitted.

In chambers, the emission rate from moth repellents made of essentially pure naphthalene is 0.16 to 0.19 mg g⁻¹ h⁻¹ [80], and it tends to last longer than p-dicholorobenezene in the same application. Clothes stored with moth repellents may adsorb naphthalene and subsequently become secondary sources, e.g., a regular cotton shirt absorbed up to 3 mg of naphthalene when indirectly exposed to mothballs in a storage cabinet [81]. US EPA [62] notes that application rates are “imprecise,” ranging from 0.25 to 0.37 lbs per 12 ft⁻³ for mothballs or flakes used indoors as a moth repellant, and 1 lb per 12 ft⁻³ for flakes used indoors as an animal repellant.

More generally, naphthalene is used as a fumigant to repel animals and insects in closets, attics, soils (including gardens), and other applications, and also as a deodorizer in diaper pails and toilets. Outdoors, it is used to control nuisance vertebrate pests (snakes, squirrels, rats, rabbits, bats, etc.) around garden and building peripheries with an application rate from 0.56 to 10.8 lb of granules or flakes per treated area [62].

Many building materials emit naphthalene. A Canadian material emission database listed naphthalene in 41 of 69 commonly used materials [60]. On an area basis, caulking has the highest emission rate, 310 mg m⁻² h⁻¹, among materials tested, followed by carpet pads (installed underneath carpets), 2.1 to 9.9 mg m⁻² h⁻¹. Emission rates fell below 1 mg m⁻² h⁻¹ for other materials tested, which included solid and engineered materials and flooring materials. In a chamber study, naphthalene emissions from rubber floor covering with a loading ratio of 0.4 m² m⁻³ and an air exchange rate (AER) of 0.5 h⁻¹ resulted in a concentration of 3 μg m⁻³ [82]. Overall emission rates from materials in ten urban homes in Chicago, Illinois, estimated using AER measurements and steady-state and full-mixing assumptions, averaged 0.245 mg h⁻¹, and the median was 0.115 mg h⁻¹ [83].

A few countries have compiled estimates of nationwide emissions of naphthalene. Table 3 lists annual releases from the U.S., Canada, Netherlands, Scotland and Switzerland. Of these countries, the U.S. has the highest emissions: releases peaked at about 3,000 metric tons yr⁻¹ in 1988 and 1998; more recent figures are in the 1,500 metric ton yr⁻¹ range. The main source is the chemical industry, including the production of phthalic anhydride, and the manufacture of phthalate plasticizers, resins, dyes and insect repellents [2]. The chemical industry accounted for as much as 41% of (industrial) naphthalene emissions in 2006. The next largest category was the primary metals industry, which accounted for 28% of 2007 emissions. The data in Table 3 suggest downward trends over the last two decades in aggregate emissions in the U.S. and the Netherlands as well as U.S. mobile emissions; the trend for aggregate emissions in Canada appears to be flat. Decreases in aggregate emissions may be reflected in the declining trend of outdoor concentrations (depicted later in Figure 4).

As noted, emission inventory estimates have many limitations and may not reflect true releases. The quality of inventories, including their accuracy, validation and completeness, remains an issue. In fact, most inventories do not include naphthalene, instead focusing on criteria air pollutants and selected VOCs [84]. Additionally, inventories emphasize industrial and mobile sources, and most or all residential and commercial sources are generally excluded. While individually small, collectively these sources are important. For example, in 1989, about 5,500 metric tons of naphthalene were used as a moth repellent, 7,000 metric tons in 1994 [2], and 3,400 metric tons in 2008 [62]; essentially all of which was emitted to air. These quantities considerably exceed values listed in the national emission inventory (2,215 and 1,624 metric tons in 1989 and 1994, respectively). These values are uncertain, and they do not include off-label uses of naphthalene as a fumigant.

Naphthalene concentrations in residences, along with other VOCs, have been reviewed in several papers. An older U.S. EPA report [96] summarized indoor air concentrations measured from 1987 to 1989 and, interestingly, listed naphthalene separately as a VOC, PAH and pesticide. As a VOC, the reported mean and median concentrations in 230 German homes were 2.3 and 2.0 μg m⁻³, respectively; as a PAH, concentrations were 1.0 and 2.2 μg m⁻³ in homes with and without smokers, respectively; and as a pesticide, concentrations ranged from 0.55 to 4.2 μg m⁻³ in U.S. residences. Brown et al. [37] reviewed 50 studies between 1978 and 1990, and estimated that the weighted average GM concentration was below 1 μg m⁻³ in various non-occupational indoor environments. Holcomb and Seabrook [97] listed average VOC concentrations in various microenvironments from 27 studies; naphthalene concentrations were reported in only two studies conducted before 1990 with quite high concentrations, from 9.2 to 13.9 μg m⁻³. Hodgson and Levin [98] estimated that the most representative median concentration in existing houses was 0.84 μg m⁻³. Wang et al. [99] reported naphthalene concentrations 0 to 1.8 μg m⁻³ in aircraft, but residential concentrations were not noted. Generally, naphthalene received little attention: it was omitted in other reviews of VOCs, e.g., Shah and Singh [100]; Dawson and McAlary [101], as well as in reviews of PAHs, e.g., Srogi [102] and Chang et al. [103].

Naphthalene-specific reviews have focused on its toxicity and health effects, although some have reported indoor and outdoor concentrations. The Report on Carcinogens [16] is limited to concentrations in workplaces. The IARC monograph [14] gives references only prior to 1995. The ATSDR toxicological profile [2], updated in 2003, includes references up to 1999, but omits many articles (which we cite in this review). Pruess et al. [11] reviews indoor and outdoor exposures, and proposes a hygiene-based exposure limit of 1,500 μg m⁻³, which is far above environmental limits. This review lists the means and ranges of concentrations found in 10 indoor residential studies and 21 outdoor studies. A recent summary of naphthalene exposure data [5] used a single study conducted in day care centers to represent indoor levels.

Table 4 summarizes 21 studies that measured indoor concentrations of naphthalene from 1986 to 2006. Concentrations are highly skewed, both within and between studies, as shown by study averages that greatly exceed medians, and by high peak concentrations (up to 144 μg m⁻³). Median concentrations ranged from 0.17 to 4.1 μg m⁻³, and averages from 0.8 to 9.5 μg m⁻³. Two studies showed strikingly high concentrations: the 150-home study in Syracuse, NY in which over 80% of the homes had one or more tobacco smokers indoors [104]; and a study of 27 residences which included many receiving complaints from occupants [105]. After excluding these studies, median concentrations ranged from 0.18 to 1.7 μg m⁻³ in non-smoking residences, which we present as a representative range for residences. Exaggerated by extreme values, means showed a wider range, 0.27 to 4.1 μg m⁻³. These ranges resemble those suggested in the previous reviews [37,96,98], and they encompass the value (0.95 μg m⁻³) used to estimate chronic inhalational doses and cancer risks [16].

Maximum concentrations reported in the different studies vary considerably, from around 1 to 144 μg m⁻³ (Table 4). The highest concentration (144 μg m⁻³) was observed in a home in Ottawa, Canada [113]; a nearly comparable level (92 μg m⁻³) was measured in a southeast Michigan, USA home [107]. However, most extrema were below 50 μg m⁻³, in accordance with Preuss et al. [11] who stated that “the majority of investigations found naphthalene levels far below 50 μg m⁻³.” We do not know whether these high concentrations reflect long-term levels and chronic exposure, or short-term excursions that might have occurred due to intermittent use of a naphthalene product or for other some other reason.

Although the effect of tobacco smoking is not always consistent (see below), we separately estimated concentrations in residences containing smokers. Concentrations averaged 1.8 μg m⁻³ in 10 Michigan homes [44], 9.5 μg m⁻³ in over 120 New York State homes [104], and from 1.3 to 2.5 μg·m⁻³ in 6 Ohio homes [111]. One median concentration was available, 2.8 μg m⁻³, in Syracuse, NY [104]. We propose a representative range of 1.8 to 9.5 μg m⁻³ for average (medians not available) naphthalene concentrations in residences with smokers, considerably higher than that just discussed for smoke-free residences.

Very elevated concentrations of naphthalene have been reported in rural homes in developing countries, such as China and African countries, where coal, wood and crop residues are widely used in simple and unvented cook stoves. For example, naphthalene concentrations averaged 28.7 μg m⁻³ in rural houses in Burundi [124]. While beyond our present scope, these studies suggest the need for further investigation given their significance and the scarcity of the data available.

Regional differences. Figure 1 displays ranges of median and mean concentrations for the U.S., Canadian and European studies. Naphthalene concentrations in US and Canada residences are similar. The North American studies show the widest range of concentrations, and levels generally exceed those in European studies, which show median concentrations below 0.6 μg m⁻³.

Urban vs. rural areas. Indoor concentrations did not show significant differences between rural and urban residences in one study [6]. Other (inter-study) comparisons also suggest little if any difference, e.g., the median concentration in a rural area in Missoula, MT, USA was only 0.3 μg m⁻³ [106], while similar or lower concentrations have been found in urban settings, e.g., 0.18 μg m⁻³ in the Chicago area [109]. Thus, excluding the measurements in rural areas [6,106] from the derivation of representative concentration ranges does not significantly affect the indoor concentration ranges.

Long-term trends. Figure 2 shows study averages and median concentrations by year. Indoor levels did not show any clear trend, even after removing observations from smoking and complaint residences, or after stratifying by region or country. This differs from the decreasing trend seen for many indoor VOCs in North America [98] and Europe [118], and it is somewhat surprising given naphthalene’s decreasing share of the moth repellent market. The flat trend may hint that the older studies did not fully characterize naphthalene levels, possibly due to measurement problems as discussed later.

Seasonal variation. Indoor concentrations of naphthalene did not show consistent seasonal patterns. Higher concentrations were expected in winter due to lower air exchange rates and increased emissions from wood-burning fireplaces, which might explain trends seen in the nationwide Canadian study where concentrations averaged 6.7, 4.3 and 2.5 μg m⁻³ for outdoor temperatures ≤0 °C, 0–15 °C and ≥15 °C, respectively [115]. However, the opposite trend may occur due to higher emission rates from materials and outdoor barbecuing in summer, possible reasons for the higher levels in summer seen in ten Chicago area homes [109]. Seasonal cycles of aromatic VOCs (but not naphthalene) have been modeled in two German studies for aromatics [117,118].

Our review indicates that several emission sources influence indoor concentrations, including indoor tobacco smoking, use of moth repellents, and the presence of an attached garage. Additional factors affecting indoor concentrations include season, building characteristics (location, house age, and type of heating system), furniture, and ventilation condition. As discussed below, however, the evidence for these sources and factors is inconsistent, and few studies have quantitatively apportioned sources or the effects of these factors.

Attached garages. Garages containing vehicles or gasoline are known sources of aromatic compounds and naphthalene, although naphthalene concentrations in garages have been rarely reported [125]. Thus, garages attached to residences can significantly elevate indoor concentrations of aromatic compounds, including naphthalene [108,112]. In southeast Michigan homes with attached garages, sources in the garage were responsible for 35% of the indoor naphthalene levels; most of the remainder (65%) arose from indoor sources; and contributions from outdoor sources were negligible [126].

Indoor tobacco smoking. Elevated naphthalene concentrations were found in residences with tobacco smoking in an early study [111]. However, this was not the case in more recent studies, including studies that specifically evaluated effects of environmental tobacco smoke (ETS) [6,108,112], probably because ETS contributes only a small amount of naphthalene. Based on an analysis using ETS tracers, Charles et al. [44] estimated only about 3% of naphthalene was due to ETS. Nazaroff and Singer [10] provide a comparable estimate, a concentration increase of 0.1 to 0.2 μg m⁻³ due to smoking. These small contributions likely explain the inconsistent effect seen for indoor smoking.

Moth repellents. Elevated naphthalene concentrations due to indoor storage of mothballs were seen in a study of 10 residences [110]. Several older studies reported high indoor concentrations, 520 to 1200 μg m⁻³, in bedrooms and living rooms adjacent to closed spaces containing mothballs [23]. Price and Jayjock [5] provide some simple calculations of naphthalene concentrations in homes using mothballs, but supporting measurements are not given. Based on a mothball emission rate of 0.175 mg g⁻¹ h⁻¹, a ball or cake weight of 32 g [80], the typical U.S. house volume of 369 m³ and exchange rate 0.63 h⁻¹, and a fully mixed (box) model [5], the indoor concentration of naphthalene is 24 μg·m⁻³. Concentrations would gradually decline, but the product would last about 8 months. Most users probably place a few mothballs in closed closets, plastic clothes bags, clothes chests or drawers about once per year. Such environments have limited exchange to the rest of the home, and emission rates can be affected by mass-transfer limitations and source-sink effects. These processes would tend to lower concentrations, but concentrations could be predicted using more complex, multicompartment models. Presently, p-dichlorobenzene appears to be the predominant ingredient in moth repellents sold in the U.S., however, naphthalene’s use in this application also continues. In many other countries, naphthalene continues to find more extensive use as both a repellent and deodorizer.

Griego et al. [20] state that Recochem, a Canadian company, is the sole U.S. registrant for pesticide use of naphthalene, but that off-label uses of mothballs as area fumigants, e.g., many mothballs placed on open trays in attics or other portions of homes, can elevated indoor levels by 10 to 300 μg m⁻³. The maximum indoor levels noted earlier, 90 and 144 μg m⁻³ in Michigan and Ottawa homes, respectively, fall in the range suggested for off-label uses, however, the specific use of naphthalene in these homes is unknown. Overall, historical exposures that resulted from the use of mothballs as well as off-label uses have not been documented.

Other factors. Domestic wood burning for heating has been reported to increase indoor PAH concentrations [127]. However, no significant difference was found in the single study that specifically examined fuel combustion and naphthalene levels [111]. Incense burning has been found to significantly increase indoor naphthalene concentrations in Asian homes [128].

Air fresheners have been stated as naphthalene sources [2], but Heroux et al. [112] found lower naphthalene concentrations in rooms containing air fresheners. Other chemicals appear to have replaced naphthalene use in this application, at least in the U.S.

Naphthalene concentrations have been positively related to home age [109], the opposite of what has been observed for other VOCs [105]. Lower naphthalene concentrations were found in homes with new furniture (≤12 months old), attributed to the sink effect [112].

In summary, few determinants of naphthalene concentrations that are consistent across the studies have been established. This may be caused by several factors. First, observed naphthalene concentrations are frequently low, often near or below method detection limits (MDLs). Measurements at such levels are prone to large uncertainties that may mask the true variation. Second, as discussed later, current monitoring methods for environmental naphthalene (as opposed to occupational settings) have not been fully validated. Third, the number of studies that have measured naphthalene is small, and most were not designed to identify determinants. Fourth, source-sink effects for naphthalene may be significant, and can complicate the collection of representative samples and the identification of primary (not re-emitting) sources. Lastly, in many cases there are multiple sources of naphthalene and source attribution is not always easy.

Although the bulk of emissions occur to outdoor air, naphthalene is found at low concentrations in most urban and suburban settings. The 24 studies reviewed reported average outdoor concentrations below 1.0 μg m⁻³ and medians below 0.5 μg m⁻³. A set of studies considered to be representative of urban or suburban levels was developed by excluding the studies conducted in remote or rural areas, or those that used only one sampling site. This left 13 studies. Acknowledging some data quality issues, our estimate of representative urban and suburban concentrations are 0.02 to 0.31 μg m⁻³ for medians, and 0.01 to 0.82 μg m⁻³ for averages (Table 5). This range is within the 0.001 to 1.0 μg m⁻³ brackets suggested in a recent review [5].

Maximum outdoor concentrations ranged from 0.01 to 4.7 μg m⁻³ (Table 5). The upper bound concentration, about 5 μg m⁻³, might be used as a conservative exposure scenario for “general” urban air. Occasionally, higher concentrations have been detected, e.g., 25 μg m⁻³ was observed outside of eight Hangzhou, China homes [123], but this may have resulted from entraining indoor air. Levels may be elevated near industrial and waste disposal sites containing strong sources, e.g., naphthalene levels averaged 11 μg m⁻³ at a landfill site in summer in Guangzhou, South China [129].

Regional differences. Median outdoor concentrations of naphthalene in urban areas ranked by region generally follow the following trend: U.S.A. > Europe > Canada (Figure 3). Median concentrations measured elsewhere (as those summarized in Table 5) fall within the bounds of these studies. It is not feasible to compare ambient concentrations measured in rural or remote areas due to the very low concentrations seen and the small number of studies.

Temporal variation. Like many other pollutants, naphthalene concentrations in outdoor air undergo both diurnal variation. Concentrations tend to peak at night and in the morning, and are lowest at mid-day [40,137]. Nighttime increases are attributable to low mixing heights that build up levels from local sources, while morning peaks are associated with vehicle emissions at rush-hour and diminished dispersion. Levels decrease during midday due to enhanced dispersion, lower traffic and reduced emissions. A simulation study predicted peak concentrations increasing by roughly 0.2 μg m⁻³ due to weak vertical mixing and photo-oxidation [40]. This pattern is similar to that seen for traffic-related VOCs such as benzene, toluene and 1,3-butadiene [138] and some PAHs [139].

Outdoor concentrations also show seasonal variations, and higher levels are seen in winter as compared to summer [40,130,131,137]. Seasonal changes were 0.96 μg m⁻³ in Los Angles and 0.20μg·m⁻³ in Riverside, California, USA [131]. (Other studies did not quantify seasonal changes.) In general, the seasonal changes follow patterns seen for VOCs, i.e., in winter, cool temperatures are often associated with decreased photochemical reaction rates [140], increased emissions from heating sources [39], and decreased dispersion due to more stable air and lower mixing heights [141].

All but one of the ambient studies used sampling periods less than 3 years in length, and thus long-term trends were not investigated. In Leipzig, German, ambient measurements were collected for 8 years, but naphthalene trends were not identified [118]. We attempted to discern trends by aggregating the 12 urban studies, and plotting means and medians against the 15 years spanned (Figure 4). These data suggest a downward trend, especially for median concentrations, but the variation is very large across the studies. Such multi-study comparisons cannot identify trends occurring at specific sites, which are best examined using long-term monitoring at the same sites, consistent methods, and sampling strategies that account for diurnal and seasonal variation.

Rural areas. Unsurprisingly, outdoor concentrations are lower in rural areas that have little traffic and fewer if any industrial facilities (Table 5). Based on measurements in Missoula, MT, U.S. [106] and rural Western Canada [134], rural concentrations typically fall below 0.1 μg m⁻³. Several studies have compared urban and rural areas. In North Carolina, U.S., the average concentration at rural sites was four times lower than at inner city sites in Durham [6]. In southern California, average naphthalene concentrations ranged from 0.06 to 0.27 μg m⁻³ in two rural communities [91]. Occasionally, rural areas experience locally elevated concentrations due to open burning of crop residues and possibly other materials, as shown in an agricultural county in Taiwan where naphthalene levels increased by 1.3 to 2.6 times to 0.38 to 0.44 μg m⁻³ during the burning period [93]. Such events are sporadic but may result in air pollution episodes with unfavorable meteorology [142,143]. Our search failed to find any measurements of naphthalene during air pollution episodes.

Remote locations. Naphthalene concentrations measured at remote locations, such as rural areas of Western Canada, typically range from below method detection limits (<MDLs) to 0.01 μg m⁻³ [134]. This is similar to the 0.0001 to 0.003 μg m⁻³ range suggested by Price and Jayjock [5]. These “background” naphthalene concentrations are small and generally negligible compared to urban levels.

Traffic. Although a known constituent of vehicle exhaust, few studies have examined the effect of traffic on naphthalene concentrations. Naphthalene has been rarely reported in tunnel studies, a common way to characterize traffic emissions. One study in two highway tunnels in Chicago reported a concentration of 8.0 μg m⁻³ [144]. In a multi-community monitoring program in Southern California, naphthalene concentrations varied from 0.06 μg m⁻³ in a community with light traffic to 0.58 μg m⁻³ in a community traversed by 200,000 vehicles day⁻¹ [91]. The naphthalene gradient due to traffic-related emissions has not been characterized.

Proximity to industrial facilities can increase naphthalene concentrations. In Sarnia, Canada, a clear concentration gradient was seen around a large cluster of industrial and chemical facilities [133]. In southeast Michigan, higher concentrations were observed in an industrial city than a suburban community in Michigan [107]. A modeling study of Los Angeles showed a similar pattern [40]. As mentioned earlier, quite high concentrations (11 μg m⁻³) were noted at a landfill site in China in summer [129]. A slightly elevated concentration, 1.1 μg m³, was seen near a former manufactured gas plant [145]. (Note that this concentration was incorrectly reported (as 1,100 μg m³) in this paper due to a unit problem.)

In summary, ambient concentrations of naphthalene vary spatially and temporally. Concentrations in urban areas typically range from 0.02 to 0.31 μg m⁻³; levels are much lower in rural and remote areas, below 0.1 μg m⁻³. Particularly in urban areas, concentrations show diurnal and seasonal patterns, reflecting variability in emission sources and meteorological influences. Rural areas may experience occasional elevated concentrations due to open biomass burning. While several studies used a large number of measurements, no long-term trends were discerned, and this was not investigated in any of the studies. Outdoor concentrations form a “floor” for indoor concentrations, i.e., they represent the level in residences that do not contain naphthalene-containing products.

Personal exposure to naphthalene have been reported in only three European studies (Table 6). These show median concentrations from 0.5 to 2.0 μg m⁻³ and averages from 0.78 to 2.3 μg m⁻³. These levels are within the ranges discussed for indoor residential concentrations (and much higher than outdoor concentrations), suggesting that personal exposures mainly result from exposures and sources indoors. The maximum personal exposures varied from 2.7 to 12.7 μg m⁻³. The maxima ranked as urban > suburban > rural in the U.K.; median and mean concentrations did not depend on urbanization [146]. Edwards et al. [121] made simultaneous indoor, outdoor and personal measurements, and found that personal exposures (including means, medians, and maxima) were lower than indoor concentrations. This suggests that personal contact with naphthalene sources is rare, and that the “personal cloud” effect seen for some other pollutants may not apply to naphthalene. This seems reasonable given that individuals spend minimal amounts of time in a closet or bathroom where mothballs, deodorizers and air fresheners might be used, while monitoring in nearby rooms might end up with high concentrations, and since individuals are outdoors or in other environments where levels are low.

A recent UK study [146] thoroughly analyzed determinants of personal exposure, and found that the presence of an attached garage and exposure to ETS were determinants of VOC exposures, and ETS was the primary determinant of PAH exposures. However, these factors did not significantly affect naphthalene exposures. Other factors, including community setting and property value, also were not influential.

The lack of studies examining personal exposures in North America, where naphthalene has found extensive indoor use and where both indoor and outdoor concentrations are higher, represents a significant gap in the literature. While largely beyond our scope, the same gap appears to exist in the occupational setting, e.g., Rappaport et al. [29] used exclusively European studies to establish the relationship between personal naphthalene exposure and total PAH exposures. Information is also lacking for Australia and other countries where indoor use of naphthalene is common.