Lung cancer risk by polycyclic aromatic hydrocarbons

Environ Sci Pollut Res DOI 10.1007/s11356-016-7566-4

RESEARCH ARTICLE

Lung cancer risk by polycyclic aromatic hydrocarbons in a Mediterranean industrialized area Anna Cuadras 1 & Enric Rovira 1 & Rosa Maria Marcé 2,3

&

Francesc Borrull 2,3

Received: 27 April 2016 / Accepted: 31 August 2016 # Springer-Verlag Berlin Heidelberg 2016

Abstract This study focuses on characterizing the chronic risk assessment from inhalation of polycyclic aromatic hydrocarbons (PAHs) for people living near the largest chemical complex in the Mediterranean area. Eighteen PAHs were determined in the atmospheric gas and particle phases, counting PM10 and total suspended particles. The lifetime lung cancer risk from PAH exposure was estimated, and the contribution was assessed by phases. The results obtained with the continuous lifetime scenario were compared with those obtained with different chronic scenarios. The estimated chronic risk was also compared with those reported in previous studies. PAHs were present at higher concentration in the gas phase (>84 %) with a major contribution of the most volatile PAHs, and an equitable distribution of heavy PAHs between gas and particle phases was observed. Petroleum combustion and traffic emissions were suggested as the main sources, but the influence of petrogenic sources cannot be ruled out. The estimated average lifetime lung cancer risk in this study ranged

between 3.2 × 10−5 and 4.3 × 10−5. The gas phase accounted for the most significant contribution to the total risk (>60 %). Fluoranthene (FluT), dibenzo(a,h)anthracene (DahA) and benzo(a)pyrene (BaP), as a whole, made the greatest contribution to the total risk (>80 %). BaP-bound PM10 accounted for a small contribution of the total risk (10 %). Chronic exposures lower than total lifetime hours could even pose a risk >10−5. The results also showed that BaP-bound PM10, according to current legislation, may not be a good indicator of the real risk by PAH exposure. Concerning previous studies, the economic situation may have an impact on reducing the cancer risk by PAH inhalation. Keywords Polycyclic aromatic hydrocarbons . Gas phase . Particulate phase . Air quality . Industrial area . Risk assessment

Introduction Responsible editor: Philippe Garrigues Electronic supplementary material The online version of this article (doi:10.1007/s11356-016-7566-4) contains supplementary material, which is available to authorized users. * Rosa Maria Marcé [email protected]

1

Observatory of Health and Environment of Tarragona, Public Health Agency of Catalonia, Health Department, Generalitat de Catalunya, Av. Maria Cristina, 54, 43002 Tarragona, Spain

2

Department of Organic Chemistry and Analytical Chemistry, Universitat Rovira i Virgili, Marcel·lí Domingo s/n, Campus Sescelades, 43007 Tarragona, Spain

3

Chemistry Technology Center (CTQ), Marcel·lí Domingo, s/n. Campus Sescelades, 43007 Tarragona, Spain

Polycyclic aromatic hydrocarbons (PAHs) are mostly formed during the incomplete combustion of fossil fuels and wood, forest fires or tobacco smoke and from the release of petroleum products (Kim et al. 2013; Ravindra et al. 2008). PAHs have been classified as priority pollutants (US EPA 1990, EEA 1999) due to their ubiquitous presence in the atmosphere and their toxicological profile (ATSDR 1995). Although food is considered to be the main route of exposure for nonsmokers, breathing outdoor air containing PAHs can contribute significantly to health outcomes due to the widespread presence of these hydrocarbons in the atmosphere (Kim et al. 2013; WHO 2000). Teratogenic properties, such as adverse birth outcomes (Padula et al. 2014), delayed child development (Jedrychowski et al. 2015), kidney and liver damage and

Environ Sci Pollut Res

atherosclerosis, have been related to PAH inhalation (ATSDR 1995; WHO 2000). Over the past decade, certain PAHs have been classified as endocrine disruptors (Annamalai and Namasivayam 2015; WHO 2012). However, carcinogenicity and mutagenic effects are the most commonly reported effects of exposure to PAHs (Kim et al. 2013; WHO 2013). According to the International Agency for Research on Cancer (IARC), several PAHs have been classified as carcinogens: benzo(a)pyrene (BaP) as carcinogenic to humans (group 1); dibenzo(a,h)anthracene (DahA) as a probable human carcinogen (group 2A); and naphthalene (Nap), benzo(a)anthracene (BaA), chrysene (Chr), benzo(b)fluoranthene (BbF), benzo(j)fluoranthene (BjF) and indeno(1,2,3-cd)pyrene as possibly carcinogenic to humans (group 2B). Moreover, other PAHs, such as fluoranthene (FluT), may contribute to carcinogenicity due to mutagenic properties, although it was classified as a weak carcinogen (Bostrom et al. 2002). Lung cancer risk was established in relation to PAH inhalation by epidemiological and toxicological studies (WHO 2000; Armstrong et al. 2004). Other cancers, such as skin, bladder and stomach cancers, have also been suggested in literature, but there is less evidence (Kim et al. 2013; IARC 100F 2012). PAHs can be divided into two groups: those with low-molecular-weight (LMW) compounds (two- and three-ring PAHs), which are predominantly in the gas phase and can condensate after emission, and those with high-molecular-weight (HMW, four- and five-ring PAHs), which are found more commonly in the particle phase (Kim et al. 2013; WHO-IPCS 1998). Previous studies showed that a fraction of the PAHs inhaled is absorbed and metabolized in the tracheobronchial tree and t ransported slowly into the blood stream . Therefore, it contributes towards the development of cancer at the site of entry. Another fraction can penetrate down to the alveoli and rapidly be absorbed into the blood, contributing to the development of cancer in the systemic circulation. Because of their lipophilic characteristics, heavy PAHs, such as BaP, are retained in the lung tissue and reach high local concentrations, even at the low levels of exposure (Bostrom et al. 2002). Particles smaller than 10 μm have been shown to cause significant health problems because they can affect gas exchange within the lungs and can even penetrate them. Gas molecules behave similarly to particles smaller than 1 μm. They can therefore penetrate down to the alveoli and can reach the circulatory system (Kim et al. 2015). Exposure to high levels of PAHs is a growing health concern in Europe (Garrido et al. 2014; Gerreiro et al. 2014). BaPbound PM10 is used as an indicator for carcinogenic PAHs, in accordance with the current legislation (D 2004/107/CE). However, some studies suggest considering other PAHs,

besides BaP, from a health point of view (Bostrom et al. 2002; Ravindra et al. 2008). Risk assessment is used to evaluate chronic health effects in order to facilitate governments to make decisions (Robson and Toscano 2007; US EPA 2009). Lifetime lung cancer risk has been defined in relation to BaP inhalation (WHO 2000), and other PAHs may be considered taking into account their cancer potency in relation to BaP (Callen et al. 2014; DelgadoSaborit et al. 2011; Wickramasinghe et al. 2012). In the Tarragona region, a large chemical complex operates close to several urban and suburban areas. In this context, the aims of this study are to estimate the lifetime lung cancer risk due to PAH inhalation for residents living in urban-suburban areas close to this chemical site and assess the contribution to the total risk by PM10, total suspended particle and gas phases. Previous studies have estimated the risk of developing cancer due to PAH inhalation in industrialized urban areas, taking gas and particle phases into consideration (Gaga et al. 2012; Ramírez et al. 2011; Ravindra et al. 2006). However, to our knowledge, this is the first time that risk related to BaP-bound PM10, according to current legislation, is compared with the total risk by total suspended particle (TSP) and gas phases. To this end, PAHs were monitored in two locations from May to October 2013. Eighteen PAHs classified as priority compounds (ATSDR 2013) were determined in gas and particle phases, counting PM10 and TSP. The lifetime lung cancer risk from PAH exposure was estimated, and the contribution was assessed by phases. The results obtained were compared with recommended values. In addition, the risk estimated for the continuous lifetime scenario was compared with different chronic scenarios in order to consider other windows of exposure. The estimated chronic risk was also compared with those reported in previous studies at the same industrial area.

Material and methods Study area The largest chemical complex in Southern Europe and the Mediterranean area is located in the Tarragona region (Catalonia, north-eastern Spain). Two main areas make up this chemical complex: the North Industrial Complex, which includes a large oil refinery and several chemical industries, covering an area of 470 ha, and the South Industrial Complex, which includes a refinery of asphaltic products and several chemical and petrochemical plants, covering 717 ha. As a whole, this chemical complex produces 17.2 million tonnes per year, about 25 % of Spain’s chemical production (AEQT 2013). Petroleum derivatives, such as gasoline, diesel, fuel oil or kerosene, ethylene, polypropylene and styrene, are the major products manufactured in the north complex. In the south complex, the predominant products


are betum, naphtha, kerosene, diesel, propane, polypropylene, polyethylene, styrene, halogenated organic compounds, chlorine and fungicides. Figure 1 shows the map of the study area, the location of the main city (Tarragona) and the two sampling sites. Site 1 is located in La Canonja, and site 2 is located in La Pobla de Mafumet. Both locations are suburban areas, located less than 1 km from the chemical complex, with moderate traffic density (16,272 and 14,642 vehicles per day at site 1 and site 2, respectively) (Spanish Ministry of Public Works, personal communication). Both sites are in the area of maximal environmental impact of the industrial emissions with a major influence of southerly wind in warm period. Wind speeds are light (annual average 96 %. DCM used in pressurized liquid extraction (PLE) and cleanup of the PUFs was purchased from Prolabo (VWR, Llinars del Vallès, Spain) with >99 % purity, and dimethylformamide (DMF) used in the rotary evaporator concentration process was provided by Merck (Darmstadt, Germany). Helium (Carburos Metálicos, Barcelona, Spain) was used for the chromatographic analysis, and Hyflo Super Cel diatomaceous earth (Sigma-Aldrich) was used to the fill extraction cells of the PLE equipment. Analytical method The samples were extracted using an ASE 200 (Accelerated Solvent Extraction equipment from Dionex, Sunnyvale, CA, USA) equipped with 11 mL stainless steel extraction cells for TSP or PM10 filters and 33 mL cells for PUF. TSP filter or half of PM10 filter, cut into small pieces and mixed with diatomaceous cell, or PUFs cut into two pieces were placed in the extraction cell. DCM was used as the extraction solvent. The extraction began with a preheating step of 5 min followed by a static time of 5 min at 100 °C (120 °C for PM10) and 1500 psi. The flush volume was 50 % (100 % for PM10), and the purge time was 120 s. Subsequently, one cycle under the same conditions was applied to extract PUF and two cycles for PM10 and TSP filters. PLE conditions were optimized in a previous study (Ramírez et al. 2011). The extract was placed in a 100-mL round-bottom flask, and then, 400 μL of DMF was added to prevent the evaporation of the most volatile PAHs. The extracts were reduced using a rotary evaporator and then transferred into a 1-mL volumetric flask by adding DCM. Finally, 20 μL of 50 mg/L deuterated PAH solution was added to the volumetric flask, filled to the mark with DCM. Recoveries of the extraction were >90 % for all of the compounds, in line with previous studies (Ramírez et al. 2011), and repeatability values, expressed as % RSD (n = 5), were lower than 9 %. The compounds were determined with a gas chromatography-mass spectrometry (GC-MS) instrument (QP2010 Ultra High-Performance Gas Chromatograph Mass Spectrometer, Shimadzu Corporation, Izasa S.A. Madrid, Spain), with an automatic injector and using a Zebron ZB-5 capillary column (5 % phenyl, 95 % dimethylpolysiloxane, 30 m × 0.25 mm i.d. × 0.25 μm), provided by Phenomenex (Le Pecq Cedex, France). One microlitre of solution was injected in the GC-MS instrument at 270 °C in splitless mode, using helium as the carrier gas with a constant flow of

1 mL min−1. The oven temperature programme started at 100 °C for 4 min before increasing to 290 °C at 6 °C min−1. The GC-MS interface was set at 280 °C, and MS detector operated at an electronic impact energy of 70 eV. The MS detection was in SIM mode, selecting one quantifier ion and two qualifier ions per each compound. Qualitative identification was performed based on the retention time and the ratios of the quantifier and qualifier ions. To quantify each compound, the internal standard calibration method for GC-MS was applied. The standards were prepared in DCM, ranging between 0.01 and 50 mg L−1, adding a constant concentration of 1 mg L−1 of each internal standard. In line with section 14 of method TO-13 (US EPA 1999), field blanks, process blanks and solvent blanks were also performed. All of these steps gave no blank signals for the target PAHs. Limits of detection (LODs) ranged between 0.006 and 0.04 ng m−3 for an average air preconcentration of 1000 m3 and 0.01 and 0.06 ng m−3 for an average air preconcentration of 300 m3. Limits of quantification (LOQs) ranged from 0.012 to 0.10 ng m−3 (air preconcentration of 1000 m3) and from 0.030 to 0.17 ng m−3 (air preconcentration of 300 m3) (for further information about LODs and LOQs, see Table S1 (online resources)). Health risk assessment Toxic equivalent factors: BaP equivalency Toxic equivalence factors (TEFs) were used in order to rank the 18 target PAHs according to their cancer potency relative to BaP. Taking into account the current knowledge of PAH carcinogen effects, the TEFs proposed by Larsen and Larsen (1998) were used, with the exception of volatile PAHs, such as Nap, AcPy and AcP, in which case the TEFs proposed by Nisbet and Lagoy (1992) were applied. TEFs used in the study are shown in Table S2 (online resources). BaP equivalents (BaP-eq) were calculated by multiplying each individual PAH concentration with its corresponding TEF, and the concentration of total PAHs was expressed as BaP-eq. EC based on the duration and pattern of exposure In accordance with US EPA procedure for inhalation chronic exposure (US EPA 2009), the exposure concentration (EC) was calculated as EC ¼ ðCA ET EF EDÞ=AT where EC is the exposure concentration (ng m−3), CA is the PAH concentration in air (ng m−3), ET is the exposure time (h day−1), EF is the exposure frequency (day year−1), ED is the exposure duration (year), and AT is the averaging time


(lifetime years (70 years) × 365 day year−1 × 24 day−1). The study focused on the continuous lifetime scenario (ET = 24 h, EF = 365 day, ED = 70 years) in accordance with recommendations for carcinogenic compounds (ATSDR 2005; US EPA 2009). In addition, taking into account the fact that repeated exposures for ≥10 % of a receptor’s lifespan can be considered as chronic, other chronic-subchronic scenarios and exposures were considered: residential receptor type 1 (ET = 24 h, EF = 350 days, ED = 30 years), residential receptor type 2 (ET = 12 h, EF = 350 days, ED = 30 years), commercial receptor (ET = 8 h, EF = 240 days, ED = 30 years), construction worker (ET = 8 h, EF = 240 days, ED = 2) and recreational receptor (ET = 2 h, EF = 100 days, ED = 10) (US EPA 2009). Lifetime lung cancer risk The excess of lifetime lung cancer risk from PAHs was estimated with the following equation: LCR ¼ EC UR where UR is the inhalation unit risk (ng m−3)−1 and EC is the exposure concentration (ng m−3). For the UR, a WHO unit risk (WHO 2000) of 8.7 cases per 100,000 people (UR = 8.7 × 10−5) was used with chronic inhalational exposure to 1 ng m−3 BaP over a lifetime of 70 years. In line with US EPA criteria (US EPA 2000) for risk assessment estimations, half of the LOD were substituted for measured values 99.7 %) may have been ignored in this study. Comparing with previous studies around the world (Table 2), the total PAH levels found in this study were much higher than those obtained in European background sites or remote areas and higher than those recorded in urban areas such as Venice and Zaragoza. Site 1 (15.2–16.6 ng m−3, gas + PM10 and gas + TSP) displayed total PAH levels comparable with suburban areas in Madrid or Athens, and site 2 (34.2– 34.8 ng m−3, gas + PM10 and gas + TSP) showed levels similar to those obtained in some urban and industrial areas, such as Izmir. However, significant compositional differences were observed between the gas and particle phases. Both sites evaluated in the present study showed values lower than those reported in other urban areas with industrial influence, such as Prato, Oporto and Flanders, and very much lower than those found in Shanghai. With respect to previous studies reported in the same Mediterranean area (Ramirez et al. 2011), the levels found in this study were lower than those previously reported.

Phase distribution of PAHs The distribution of PAHs between the gas and particle phase depends on the vapour pressure of the compound and on temperature, humidity and precipitation, as well as the PAH concentration and the nature and the composition of the particles. In general, two- and three-ring PAHs are mainly identified in the gas phase, four-ring PAHs are present in both phases, and five- and six-ring PAHs are mainly attached to particles (Bostrom et al. 2002; Ravindra et al. 2008). In accordance with this generalization, two- and three-ring PAHs were the dominant compounds in the gas phase (69 and 87 % at site 1 and site 2, respectively), followed by four-ring PAHs (28 and 11 %), whereas five- and six-ring PAHs were the minority groups. Figure 2 shows the average individual PAH concentrations obtained during the sampling period by phase and site. It should be noted that despite heavy PAHs making a lower contribution in the gas phase, their concentrations were comparable with those obtained in the particle phase. Therefore, a more equitable distribution for these compounds between phases was shown. Previous studies reported a similar pattern for warm periods (Ramírez et al. 2011; Ravindra et al. 2006). One explanation for this could be that the high temperature changes the gas-particle partitioning, favouring the volatilization of PAHs from the particle to gas phase (Hassan and Khoder 2012). An overestimate of PAHs

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