Polycyclic aromatic hydrocarbons in dust from rural communities around gas flaring points in the Niger Delta of Nigeria: an exploration of spatial patterns, sources and possible risk

Abstract

Indoor and outdoor dust from three rural communities (Emu-Ebendo, EME, Otu-Jeremi, OTJ, and Ebedei, EBD) around gas flaring points, and a rural community (Ugono Abraka, UGA) without gas flare points, in the Niger Delta of Nigeria, was analysed for the concentrations and distribution of polycyclic aromatic hydrocarbons (PAHs), their sources, and possible health risk resulting from human exposure to PAHs in dust from these rural communities. The PAHs were extracted from the dust with a mixture of dichloromethane/n-hexane by ultrasonication, and purified on a silica gel/alumina packed column. Gas chromatography-mass spectrometry was employed to determine the identity and concentrations of PAHs in the cleaned extracts. The Σ16PAH concentrations in the indoor dust ranged from 558 to 167 000, 6580 to 413 000, and 2350–37 500 μg kg⁻¹ for EME, OTJ and EBD respectively, while those of their outdoor counterparts varied from 347 to 19 700, 15 000 to 130 000, and 1780 to 46 300 μg kg⁻¹ for EME, OTJ and EBD respectively. On the other hand, the UGA community without gas flare points had Σ16PAH concentrations in the range of 444–5260 μg kg⁻¹ for indoor dust, and 154–7000 μg kg⁻¹ for outdoor dust. The lifetime cancer risk values for PAHs in these matrices surpassed the acceptable limit of 10⁻⁶ suggesting a potential carcinogenic risk resulting from human exposure to PAHs in indoor and outdoor dust from these rural communities. Principal component analysis suggested that PAH contamination of dust from these communities arises principally from gas flaring, combustion of wood/biomass, and vehicular emissions.

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Environmental significance

Dust is an important repository for various organic contaminants, and is a gateway for human exposure to them. It can be used as a proxy for evaluating the contamination status of an area, and the consequent possible risk to the ecosystem and humans. This study evaluated the relationships between the concentrations, compositions, sources, and risk of PAHs in indoor and outdoor dust from rural communities around gas flare points in the Niger Delta. Our findings suggest that indoor and outdoor dust from these rural communities around gas flare points were highly contaminated with PAHs and require clean up, remedial actions, and the implementation of stringent pollution control measures with a view to reducing the adverse consequences of PAHs on the ecosystem and humans.

Introduction

Besides oil spillages, gas flaring is a major source of environmental pollution associated with the activities of oil and gas industries in the Niger Delta. Despite the existence of the Nigerian “Associated Gas Reinjection Act of 1979”, which prohibits gas flaring by oil industries beyond January 1984, gas flaring has continued unabated as a cheap and safe way of disposing of gases produced alongside crude oil due to the lack of capacity to (i) enforce this regulation, and (ii) utilize these gases for other purposes.^1,2

In Nigeria, over 70% of the natural gas produced by oil companies is flared.³ Historically, gas flaring has been practiced in Nigeria for over six decades. Apart from the loss of a potential energy source, gas flaring is a typical incomplete combustion process which produces a wide variety of pollutants including polycyclic aromatic hydrocarbons (PAHs).

PAHs constitute a large group of semi-volatile and persistent organic compounds with two or more fused aromatic rings arranged in either an angular, cluster or linear form.⁴ PAHs are released into the environment mainly from anthropogenic processes such as coke production, oil spillages, and incomplete burning of fossil fuels, among others. Natural events, such as oil seeps, volcanic eruptions, and forest fires, also contribute to the PAH load in the environment, but to a lesser extent than anthropogenic inputs.^5,6

PAHs are priority environmental pollutants because of their toxicity, persistent nature, propensity for bioaccumulation, endocrine and immune system disruptions, and capacity to induce carcinogenic and mutagenic effects.⁷

Dust is an indispensable component of the environment and can serve as an effective sink for PAHs and other contaminants. It forms a gateway for human exposure to contaminants in it through inhalation, skin contact, unconscious ingestion, and consumption of foods contaminated with dust.^6,8

The occurrence of significant concentrations of PAHs in dust is a source of concern because dust particles are mobile, and have relatively small particle sizes that can easily be inhaled, unconsciously ingested, or adhere to human skin. The widespread nature of dust makes it possible for humans to be constantly exposed to contaminants in dust particles.

Dust plays a significant role in the geochemical cycling of pollutants and can be used as a proxy for monitoring the contamination status of a site and the possible risk to the ecosystem and humans. The contamination status of indoor dust is of concern because of the long period of time people spend in an indoor environment per day. It has been estimated that people stay in one indoor environment or the other for more than 80% of the day. Infants and toddlers are at higher risk of exposure to contaminants in indoor dust than their adult counterparts because they play close to the floor, through their hand-to-mouth habits, licking of toys, and touching and putting other dust-contaminated household objects into their mouths.^9,10

Outdoor dust has the capacity to exchange pollutants with air and water depending on the weather conditions. For example, during dry weather periods, it interacts with atmospheric aerosols via desorption and re-suspension by turbulence,¹¹ while in wet weather periods dust can cause an upsurge in the pollution level of surface runoff through adsorption and desorption processes caused by washing away of adsorbed contaminants.^12–14 Therefore, a knowledge of the distribution patterns, fate and sources of PAHs in dust is necessary for improving environmental quality, source control, and reducing the adverse impacts on humans and the ecosystem.

There are dozens of reports concerning the distribution of PAHs in indoor and outdoor dust.^15–27 In Nigeria, like other regions, most of the studies were focused on urban areas with high vehicular emissions, and intense commercial and industrial activities.^6,28–36 Most considered outdoor or indoor dust and only a few explored their relationship.

Most gas flare points in the Niger Delta are located in rural communities, and the resultant emissions from this process can potentially contribute to adverse environmental and public health effects on humans residing in these rural areas. Despite extensive studies on the impacts of oil production activities on the environment of the Niger Delta, there is no report on the concentrations of PAHs in dust, indoor or outdoor, from rural communities around gas flare points in the Niger Delta. This study evaluated the concentrations, sources, and associated risk of PAHs in indoor and outdoor dust from rural communities around gas flare points. Such information is necessary for developing pollution abatement programs and risk and environmental quality management.

Materials and methods

The study areas

The study areas included Ebedei (EBD) (latitude 5.8324° N and 6.2482° E), Emu-Ebendo (EME) (latitude 5.67084° N and longitude 6.35168° E), Otu-Jeremi (OTJ) (latitude 5.43818° N and longitude 5.87829° E) and Ugono-Abraka (UGA) (control site) in Delta State, Nigeria (Fig. 1). The estimated population of these communities ranged from one to five thousand. The study areas experience a tropical equatorial climate with a distinct dry season (November to April) and wet (rainy) season (May to October). The mean annual temperature and rainfall of the study areas are 32.8 °C and 2673.8 mm respectively.

Fig. 1 Map of study areas.

Sample collection

Dust samples (n = 60) comprising ten samples each were collected from in- and outdoor environments of the three communities at 700 m to 1.5 km from gas flare points in Delta State, Nigeria. An additional 10 dust samples comprising five each from in- and outdoor environments of Ugono-Abraka with no history of gas flaring were employed as the control. The relative distances of the sampling sites from the control site are 23.1, 45.5, and 55.8 km for EBD, EME and OTJ respectively. The prevailing wind direction in the study areas is south-west. A pan and brush method was employed for collection of dust samples.^29,30

Indoor dust samples were collected from living rooms, while outdoor dust samples were collected within 50 m² from where the indoor dust samples were obtained. The indoor samples were collected from houses made of cement blocks and mud. Some of them had no cemented floors and walls, ceilings and paint. The windows in these buildings were predominantly wooden, while a few had glass louvers or sliding aluminium-framed glass windows. Electricity supply in these rural communities is episodic. Very few homes use electricity generators, while kerosene-powered lamps are the predominant source of light in these homes. The indoor dust samples were collected from floor areas, ceiling fans, cabinetry, wooden beams, and other surfaces by gentle sweeping into the dust pan.

The samples were packed in amber-coloured glass bottles previously cleaned with acetone, and carried in a cooler with ice packs to the laboratory. The samples were sieved through a <63 μm sieve and kept in a refrigerator at −20 °C prior to analysis. The pH, total organic carbon (TOC), and electrical conductivity (EC) of the dust were determined following standard methods.³⁷

Chemicals

The solvents for extraction, namely, dichloromethane (DCM, 99%) and n-hexane (96.0%) were purchased from Merck (Darmstadt, Germany), while alumina (70–230 mesh), silica gel (>632 mesh) and anhydrous sodium sulfate were products of Honeywell Fluka (Seelze, Germany). A US EPA 16 PAH (2000 μg mL⁻¹) mix supplied by AccuStandard Inc. (New Haven, CT, USA) was used for the preparation of calibration standards. A mixed standard containing six labelled ¹³C₁₂-PAHs was obtained from Cambridge Isotope Laboratories, Inc. (Tewksbury, MA, USA).

PAH extraction and instrumental analysis

The PAHs were extracted from the dust samples by employing the US EPA-3550C-ultrasonic extraction method.³⁸ Briefly, 10.0 g of dust was spiked with 200 ng of ¹³C₁₂-labelled PAHs and homogenised with 10 g of Na₂SO₄. The PAHs were extracted from the homogenate with a 50 mL mixture of dichloromethane/n-hexane (1 : 1 v/v) by sonication at 30 °C for 30 min. The extract was filtered, and the extraction cycle was carried out two more times on the residue with fresh portions of the mixed solvent mixture. The extracts were combined and reduced to 2 mL by rotary evaporation; followed by purification on an aluminium oxide/silica gel column (1 : 2 v/v). The elution of PAHs from the column was achieved with a 30 mL mixture of n-hexane/dichloromethane (3 : 1 v/v) and subsequently reduced to approximately 2 mL under a mild stream of pure nitrogen gas.

The quantification of individual PAHs in the extracts was achieved by using an Agilent 7890A gas chromatograph coupled to a 5975C mass selective detector operating in selected ion monitoring mode (SIM). A DB-5 column of 30 m length with 0.25 mm internal diameter and 0.25 μm film thickness was used for the separation of the PAHs (Agilent J &W, Folsom, CA). The mobile phase was high purity helium gas with a flow rate of 1 mL s⁻¹. The initial temperature of the column was 50 °C held for 3 min, thereafter increased to 180 °C at 10 °C min⁻¹, and further increased to 250 °C at 5 °C min⁻¹ and finally to 300 °C at 2 °C min⁻¹. The injector port temperature was fixed at 250 °C and 2.0 μL of sample was injected in a splitless mode.

Quality assurance/quality control

Analyses of method blanks, field blanks, sample matrix spikes, and replicates were applied to validate the analytical procedure. The sample matrix spikes were achieved by introducing known concentrations of PAHs (at 3 concentration levels viz 5, 30 and 100 μg kg⁻¹) into fresh portions of already analyzed samples and subjecting them to the entire analysis cycle from extraction to GC-MS analysis. The average recoveries of ¹³C₁₂-labelled PAHs were 78–100% while those of the sample matrix spike were 74 to 101%. None of the target compounds were found in the procedural blanks (n = 5). All samples were analyzed in triplicate with relative standard deviations (RSD) between 1.5 and 8.9%. The PAH concentrations were obtained by an external calibration method and R² of the calibration curves for the PAH compounds ranged from 0.9992 to 0.9998. The method detection limits for the PAHs ranged between 0.01 and 0.1 μg kg⁻¹.

Statistical analysis

The Kruskal Wallis test was used to establish the difference in concentrations and composition of PAHs within the study sites, whereas the Shapiro–Wilk test was employed to evaluate the normality of the data. The Student's t-test was used to evaluate the differences between the PAH concentrations in the indoor and outdoor dust from the same location, while the Duncan multiple range test was used for comparison of mean concentrations of PAHs in dust from the three locations. Principal component analysis (PCA) and regression analysis were employed to establish the relationships between the test compounds. SPSS version 20.0 software was used for the statistical analyses at p = 0.05.

Evaluation of ecological risk of PAHs

The ecological risk of PAHs in these matrices was assessed by utilizing the risk quotient (RQ) method described by Kalf et al.³⁹ The detailed equations for the RQ method are given in ESI S1.† The interpretation of RQ values is given as follows. The values of RQ_(NCs) and RQ_(MPCs) represent a measure of the level of risk of PAHs to the ecosystem. When RQ_NCs < 1.0 it suggests that there is negligible risk associated with exposure to the individual PAHs, while RQ_MPCs ≥ 1 indicates that exposure to the individual PAHs could cause severe risk to the ecosystem. In such a circumstance, source control and remedial actions are urgently needed to minimize the impact. RQ_NCs ≥ 1.0 and RQ_MPCs < 1 suggest a moderate risk for the individual PAHs; therefore, some control and remedial actions are needed to reduce the impact on the ecosystem. RQ_∑PAHs(NCs) ≥ 800 and RQ_{∑PAHs(MPCS)} = 0 suggest moderate-risk₁. RQ_∑PAHs(NCs) < 800 and RQ_{∑PAHs(MPCs)} ≥1 suggest that the ΣPAHs can be of moderate-risk₂ while RQ_∑PAHs(NCs) ≥ 800 and RQ_{∑PAHs(MPCs)} ≥ 1 suggest that the ΣPAH concentrations can cause a high-risk to the ecosystem (ESI Table S2†).

Health risk assessment

Carcinogenic and mutagenic potency. The carcinogenic and mutagenic potency (BaP_TEF and BaP_MEQ) of PAHs in dust samples were assessed by comparing the individual PAH carcinogenic and mutagenic potency to that of the reference compound, benzo(a)pyrene (BaP). The BaP_TEQ and BaP_MEQ of PAHs in dust were computed by using eqn (1) and (2)

BaP_TEQ = ∑C_i × BaP_TEF (1)

BaP_MEQ = ∑C_i × BaP_MEF (2)

where C_i represents the concentration of individual PAHs while BaP_TEF and BaP_MEQ are the respective carcinogenic and mutagenic potency of the PAHs relative to BaP (ESI Table S3†).

Non-carcinogenic and carcinogenic risks. The non-carcinogenic risk given by the hazard index (HI) refers to the sum of the hazard quotients (HQs) of individual exposure routes such as inhalation, accidental oral ingestion, and dermal contact. Similarly, the incremental lifetime cancer risk (ILCR) refers to the sum of carcinogenic risks for the three exposure pathways. The US EPA model equations and specified exposure conditions were utilized for assessing the non-carcinogenic and carcinogenic risks related to human exposure to PAHs from these routes.^40,41 The detailed equations for estimation of non-carcinogenic and carcinogenic risk are given in ESI S2.† The significance of the carcinogenic risk values is qualitatively classified as follows: very low, CR ≤ 10⁻⁶; low, CR = 10⁻⁶ to 10⁻⁴; moderate, CR = 10⁻⁴ to 10⁻³; high, CR = 10⁻³ to 10⁻¹ and very high, CR ≥ 10⁻¹. HI values < 1 suggest no non-carcinogenic risk, while HI > 1 suggest probable non-carcinogenic effects on the exposed human population.

Results and discussion

PAH concentrations and composition of dust around gas flare points

The Σ16 PAH concentrations in indoor dust from EME, OTJ and EBD ranged from 558 to 167 000, 6580 to 413 000, and 2350 to 37 500 μg kg⁻¹ respectively, while those of outdoor dust ranged from 347 to 19 700, 15 000 to 130 000, and 1780 to 46 300 μg kg⁻¹ for EME, OTJ and EBD respectively (Tables 1 and ESI S5†). The Σ16 PAH concentrations in the indoor and outdoor dust from these communities with gas flare points were significantly (p < 0.05) higher than those of the control site (UGA: 154 to 7000 μg kg⁻¹). This suggests discernible impacts of gas flaring activities on the PAH contamination status of these areas.

Table 1 Summary statistics of physicochemical properties and PAH concentrations (in μg kg⁻¹) in indoor and outdoor dusts

		Emu-Ebendo (EME)					Otu-Jeremi (OTJ)					Ebedei (EBD)					Ugono-Abraka [UGA] (control site)
		Mean	Median	SD	Min	Max	Mean	Median	SD	Min	Max	Mean	Median	SD	Min	Max	Mean	Median	SD	Min	Max
Indoor	pH	5.69	0.79	5.75	4.30	6.70	5.66	0.73	5.70	4.40	6.70	5.69	1.09	5.60	4.20	7.40	6.22	6.64	1.20	4.62	7.38
	EC (μS cm^−1)	864	850	462	132.0	2680	1135	1048	695	320	3190	252	96.6	213	153	448	0.41	0.17	0.59	0.07	1.45
	TOC (%)	1.27	0.37	1.32	0.36	1.69	1.43	0.65	1.48	0.41	2.49	1.56	0.42	1.54	0.98	2.31	34.5	38.4	9.12	22.1	43.4
	Nap	255	1.18	643	<LOQ	2040	8.5	4.6	13	<LOQ	43.7	28.4	1.56	63.3	0.28	190	174	58	200	8.00	484
	Ace	1320	3.61	4040	<LOQ	12 800	26.7	15.7	35.4	2.64	123	490	2.43	1180	1.02	3380	73.6	44	71.2	8.00	94
	Acy	2900	7.59	9070	<LOQ	28 700	5.18	4.32	3.64	0.4	11.2	280	3.93	456	0.54	1050	96.8	48	100	8.00	174
	Flu	4380	2.71	13 800	<LOQ	43 600	8.3	3.08	11.9	0.84	37.6	392	12.5	623	1.30	1650	109	56	115	4.00	198
	Phen	2690	3.85	5340	<LOQ	15 200	7.77	7.38	4.73	0.3	16.0	469	11.8	782	0.56	2160	96	76	108	14.0	94
	Ant	2670	11.5	6040	<LOQ	19 400	3.17	3.15	1.97	0.94	7.70	651	49.2	1030	0.80	2450	64	66	54	4.00	98
	Flt	944	10	2790	0.24	8880	3.81	3.37	2.65	0.36	9.10	191	111	221	1.12	577	190	120	255	6.00	172
	Pyr	406	29	1050	0.24	3380	3.86	3.48	2.74	<LOQ	9.58	624	482	695	0.68	1710	94	98	85.4	10.0	102
	BaA	274	13	752	0.62	2410	25.1	2.22	60.1	<LOQ	194	737	270	924	2.26	2420	285	140	320	24.0	760
	Chry	608	80	1270	9.82	4160	26.4	<LOQ	62.9	<LOQ	192	728	9.05	1190	1.98	2860	191	242	132	24.0	306
	BbF	2890	68	8340	<LOQ	26 600	972	747	886	<LOQ	2560	604	47.6	1030	4.62	3020	176	234	150	18.0	352
	BkF	4520	184	13 000	<LOQ	41 400	1640	<LOQ	2200	<LOQ	5400	1300	235	1940	19.3	5130	461	96	705	32	408
	BaP	1470	558	2030	<LOQ	6300	1440	<LOQ	4550	<LOQ	14 400	3000	1280	3750	515	10 300	548	680	513	6	860
	DahA	15 500	878	31 400	<LOQ	83 400	14 000	15 400	13 500	<LOQ	35 800	1920	466	3300	137	10 700	311	114	442	10	318
	IndP	3780	830	5350	100	14 600	57 200	7700	141 000	2150	405 000	906	258	1420	0.58	4130	429	104	506	22	980
	BghiP	1490	1200	1670	60	5280	14 300	6290	19 000	279	52 100	1130	401	1580	117	4080	357	64	497	10	476
	Total	46 100	12 800	64 800	558	167 000	78 100	46 500	119 000	6580	413 000	13 500	9410	12 800	2350	37 500	3650	2710	3730	444	5260
	2R	364	16	758	<LOQ	2040	8.51	3.54	13.7	<LOQ	43.7	28.4	1.56	63.2	0.3	190	174	58	200	8	484
	3R	14 000	27	37 400	<LOQ	120 000	51.1	39.9	47.8	18.1	182	2280	92.4	3750	6.4	8940	439	290	438	42	658
	4R	2230	276	4760	12.5	15 300	56.5	18.6	84.4	6.08	233	2280	1250	2640	10.7	6390	759	720	611	100	1050
	5R	24 400	1930	49 900	<LOQ	152 000	18 000	24 600	14 000	<LOQ	37 300	6830	2200	9110	803	27 400	1500	1150	1680	68	1940
	6R	5270	2040	6770	168	17 700	60 000	18 800	125 000	6540	413 000	2030	454	2680	262	8210	786	168	969	32	1460
Outdoor	pH	6.35	0.69	6.25	5.4	7.4	6.43	0.8	6.35	5.40	7.70	6.39	0.87	6.50	5.10	7.70	6.08	6.24	0.74	5.08	7.06
	EC (μS cm^−1)	123	70.2	104	28	274	268	138	248	78.0	521	217	161	153	66.0	512	64.2	11.6	62.4	49.5	80.6
	TOC (%)	0.41	0.18	0.34	0.18	0.68	0.72	0.37	0.72	0.21	1.42	0.87	0.45	0.93	0.3	1.6	0.10	0.06	0.10	0.02	0.16
	Nap	152	0.65	480	<LOQ	1520	7.7	2.1	11.4	0.64	29.8	56.1	4.52	147	0.4	449	186	12	368	<LOQ	842
	Ace	130	0.63	408	<LOQ	1290	50.2	8.3	90.5	3.86	297	547	2.64	1720	1.16	5440	20.4	10	25	<LOQ	62
	Acy	253	0.69	799	<LOQ	2530	21.4	3.9	38.9	0.16	126	859	4.79	1800	0.5	4500	70.8	32	100	<LOQ	244
	Flu	329	0.56	1040	<LOQ	3280	13.1	2.3	33.4	0.56	108	674	6.87	1200	0.84	3400	63.6	12	116	<LOQ	270
	Phen	99	0.94	301	0.34	954	162	12.2	457	2.04	1460	1070	18.9	1900	0.66	5500	110	10	206	<LOQ	476
	Ant	225	1.4	654	0.16	2080	144	12.9	412	1.16	1320	598	15.6	1030	0.84	2700	261	20	528	<LOQ	1200
	Flt	113	0.79	355	0.18	1120	108	2.60	233	0.7	674	410	7.88	680	0.4	1700	37.2	12	63	<LOQ	148
	Pyr	369	0.62	1160	0.2	3680	75.5	4.00	223	0.96	710	1120	540	1720	3.54	5590	12.8	10	11	<LOQ	26
	BaA	281	2.13	700	1.2.0	2200	144	6.2.0	339	1.1	1030	650	19.7	851	1.96	2090	13.2	10	12	<LOQ	28
	Chry	52	6.94	130	1.6	420	238	137	253	3.26	627	1190	20.8	2610	5.02	8340	152	32	290	4	670
	BbF	69	18.5	144	8	475	2130	906	2200	218	6280	2460	1100	4380	7.66	14 500	228	22	464	6	1060
	BkF	128	74	140	24.6	454	6040	5160	5850	1050	13 000	1610	264	3100	5.36	9250	142	34	266	6	618
	BaP	228	118	352	38.1	1210	8320	7680	6800	2000	15 940	1570	1760	1260	76.3	3800	352	34	726	6	1650
	DahA	131	103	81.0	20.0	290	31 700	25 400	28 000	2470	73 500	1810	499	2440	80.8	6100	214	26	399	6	926
	IndP	81.0	95.0	43.0	0.70	126	23 800	6390	31 100	622	65 400	1140	233	1870	10.8	5800	270	30	521	22	1202
	BghiP	81.0	82.0	36.0	9.2	138	27 600	11 500	28 000	758	68 200	763	170	1130	3.66	3320	211	54	375	18	880
	Total	2720	633	6010	347	19 700	69 000	72 600	35 100	15 000	130 000	16 500	12 900	15 100	1780	46 300	2340	1310	2800	154	7000
	2R	152	0.65	480	<LOQ	1520	7.66	2.09	11.4	0.6	29.8	50.5	3.52	140	<LOQ	449	186	12	368	<LOQ	842
	3R	1040	5.82	3200	1.88	10 100	391	54.8	947	17	3060	3740	53.1	6730	7.82	18 100	525	100	941	<LOQ	2200
	4R	815	10.1	1890	3.84	5810	480	103	914	7.81	2930	3370	2190	3890	18	10 900	215	88	265	24	670
	5R	556	349	607	123	2240	26 300	22 200	25 400	<LOQ	73 500	7450	3800	9120	177	27 500	936	112	1860	24	4250
	6R	162	175	57	9.9	210	41 900	30 400	42 500	2470	130 000	2000	455	2660	14.4	7770	481	84	896	44	2080

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On average, the Σ16 PAH concentrations in the indoor dust from these sites were significantly higher (p < 0.05) than those of their outdoor counterparts except for EBD. The high concentrations of PAHs in the indoor dust relative to those of the outdoor dust may be related to the fact that outdoor dust suffers dilution effects from the circulating air, photodegradation arising from high intensity sunlight, and leaching actions of rain. In addition, indoor dust receives additional inputs of PAHs from indoor activities such as the use of kerosene- and diesel-powered lamps, smoking, coking, and burning of incense and mosquito coils. Furthermore, the smaller particle size and higher organic content of indoor dust could aid increased sorption of PAHs compared with outdoor dust.^21,30,42,43

The concentrations of PAHs in the indoor and outdoor dust from OTJ were significantly (p < 0.05) higher than those of EBD and EME which may be related to the intensity of gas flaring, age of gas flaring activities in this community, number of gas flare points, and nature of the flare stack (down or up). For example, gas flaring activities started at OTJ in 1999, while those of EME and EBD commenced in 2009. In addition, OTJ and EME have downward flare stacks, while EBD has upward flare stacks.

The site to site concentrations of PAHs in the indoor and outdoor dust within each of these locations showed significant spatial discrepancies (p < 0.05). The difference in PAH concentrations of the indoor dust is related to diversity in house-keeping habits, and nature and intensity of indoor and outdoor anthropogenic activities within the area. For example, the dominant building patterns in these rural communities are the room and parlour systems with the kitchen detached from the living rooms. Samples collected from buildings with the kitchens closer to the living rooms and those with outdoor frying activities or closer to cassava processing engines showed higher PAH concentrations. Also, the concentrations of PAHs in indoor dust from buildings without cement floors were higher than those with cemented floors and tiles. In addition, indoor dust collected from homes with wooden windows and without ceilings showed lower PAH concentrations than those with ceilings and sliding glass windows. This is because wooden windows are wide open without restrictions, as compared to those with louvres and sliding aluminum framed glass, which enhances the circulation and exchange of air between the indoor and outdoor environments. The dilution effects of the circulating air are more pronounced in buildings with wooden windows and without ceilings.

The target and intervention values for PAHs in soils established by the Nigerian regulatory authority are 1000 and 40 000 μg kg⁻¹ respectively.⁴⁴ PAH concentrations in these matrices exceeded the target values, while 70% of the indoor and outdoor dust from EME and OTJ exceeded the intervention value. It implies that the majority of sites from EME and OTJ require clean up and remedial action. However, sustainable environmental management is required for sites from EBD since the PAH concentrations were above the target values but below the intervention value.

Table 2 provides a comparison of PAH concentrations in dust from these rural communities around gas flare points with those of other regions with diverse anthropogenic pressures. Despite the differences in the analytical approach, number of PAHs analyzed, and climatic conditions, such a comparison provides information on global concentration trends. The PAH concentrations in the indoor dust from these areas were higher than those reported for indoor dust from rural, semi-urban and urban areas in Nigeria^28,30,33,34 and other regions,^22,24,27,45 but were comparable to those found for indoor dust from Warri, Nigeria³⁰ and Changchun, China.²² The PAH concentrations in the outdoor dust from these rural communities were higher than those previously reported for similar matrices from the Niger Delta and other parts of Nigeria^{6,31,32,34–36} and some cities with diverse anthropogenic pressures in other regions,^{20,23,25,26,47–49} but were lower than those of outdoor dust from Ulsan, Korea.¹⁷ Kong et al.⁵⁰ reported PAH concentrations of 1410 to 54 780 μg kg⁻¹ in dust from the Dongying oilfield in China, which were lower than the PAH concentrations in dust from OTJ but were higher than those of EBD and EME.

Table 2 Comparison of PAH concentrations (μg kg⁻¹) in this study with those reported for similar matrices in other regions

Location	Environmental media	No. of PAHs	Concentration range	Mean	Median	Reference
Emu-Ebendo, Nigeria	Indoor dust (rural)	16	558–167 000	46 100	12 800	This study
Otu-Jeremi, Nigeria	Indoor dust (rural)	16	6580–413 000	78 100	46 500	This study
Ebedei, Nigeria	Indoor dust (rural)	16	2350–37 500	13 500	9410	This study
Warri, Nigeria	Indoor dust (urban)	15	4531–111 914	42 117	31 604	30
Abraka, Nigeria	Indoor dust (semi-urban)	16	124–2131	1121	1263	30
Emu-Uno, Nigeia	Indoor dust (rural)	16	60–1413	828	870	30
Port Harcourt, Nigeria	Indoor dust (urban)	17	276–9130	2590	1598	34
Ilorin, Nigeria	Indoor dust (urban)	16	3950–8700	6090		33
23 Cities, China	Indoor dust (urban)	16	1000–470 000			17
Cape Coast, Ghana	Indoor dust (urban)	15	nd – 3240			18
Changchun, China	Indoor dust (urban)	16	21 800–329 600			^22
Lagos, Nigeria	Indoor dust (urban)	16	304–7677			28
Jeddah, Saudi Arabia	Indoor dust (urban)	16	22–9150			22
Cities of Nepal	Indoor dust (urban)	16	747–4910		1320	24
Vojvodina province, Serbia	Indoor dust (urban)	16	140–8265	1825	1404	27
Amman, Jordan	Indoor dust (urban)	16	641–65 422			45
Emu-Ebendo, Nigeria	Outdoor dust (urban)	16	347–19 700	2720	633	This study
Otu-Jeremi, Nigeria	Outdoor dust (rural)	16	15 000–130 000	69 000	72 600	This study
Ebedei, Nigeria	Outdoor dust (rural)	16	1780–46 300	16 500	12 900	This study
Lagos, Nigeria	Outdoor dust (urban)	16	289–17 343	3162	1047	31
Benin City, Nigeria	Outdoor dust (urban)	16	153–2303			32
Port Harcourt, Nigeria	Outdoor dust (urban)	17	44–13 200	2488	661	34
Warri, Nigeria	Outdoor dust (urban)	16	165–1012			6
Delta State, Nigeria	Outdoor dust (informal trade site)	16	120–8790			36
Ibadan, Nigeria	Outdoor dust (urban)	16	365 100–430 400			35
Babylon, Iraq	Outdoor dust (urban)	16	556–1890	1060		26
Dongying, China	Outdoor dust (oilfield)	17	1410–54 780			50
Bushehr city, Iran	Outdoor dust (urban)	17	736–5491			51
Rio de Janeiro, Brazil	Outdoor dust (urban)	16	108–8570			20
Jeddah, Saudi Arabia	Outdoor dust (urban)	16	1660–4980			47
Newcastle upon Tyne, UK	Outdoor dust (urban)	16	550–46 000			52
Ulsan, Korea	Outdoor dust (urban)	16	118 000–245 000			17
Greater Cairo, Egypt	Outdoor dust (urban)	16	45–12 300			53
Mashhad, Iran	Outdoor dust (urban)	16	764–8990	2183	1891	48
Bangalore, India	Outdoor dust (urban)	22	670–1800	1100		46
Dresden, Germany	Outdoor dust (urban)	16	950–270			25
Tokyo, Japan	Outdoor dust (urban)	16	250–5260			23
Hanoi, Vietnam	Outdoor dust (urban)	22	530–4700			46
Birmingham, UK	Outdoor dust (urban)	16	200–99 600			15
Nova Sad, Serbia	Outdoor dust (urban)	16	35–2422			49
Bangkok, Thailand	Outdoor dust (urban)	16	300–1100			16

---

The PAH homologue profiles in indoor and outdoor dust from these study areas showed variations. The order of average concentrations of PAH homologues in indoor dust from EME was 5 > 3 > 6 > 4 > 2-ring PAHs and those of outdoor dust followed the sequence of 3 > 4 > 5 > 6 > 2-ring PAHs. In OTJ, the sequence of average concentrations of PAH homologues in the in- and outdoor dust was 6 > 5 > 4 > 3 > 2-ring PAHs. In the case of EBD, the PAH homologue profile in indoor dust was in the order of 5 > 4 = 3 > 6 > 2-rings, while those of outdoor dust followed the order of 5 > 3 > 4 > 6 > 2-ring PAHs (Fig. 2a and b). The homologue distribution patterns in the study areas contrast those obtained in the control site (i.e. 5 > 6 > 4 > 3 > 2-ring PAHs and 5 > 3 > 6 > 4 > 2-ring PAHs for indoor and outdoor dust respectively). The similarity in the compositional patterns of PAHs in the indoor and outdoor dust from OTJ and EBD suggest a common source, while the differences observed in the case of EME reflect dissimilarities in their fate and sources.

Fig. 2 PAH composition of (a) outdoor dust and (b) indoor dust from rural communities around gas flare points.

There was poor correlation between the ∑16 PAHs in the indoor dust with that of outdoor dust from these areas (ESI Fig. S1†) which suggests differences in the strength of their input sources, transport processes and other environmental variables that affect the fate of PAHs. Similarly, there was poor correlation between the physicochemical properties of dust, such as pH, TOC, and EC, and ∑16 PAHs (ESI Fig. S2–S4†). This implies that the PAH contamination of the indoor and outdoor dust from these areas may have come from recent and continuous contamination events which can induce distortion in the adsorption equilibrium between TOC and PAHs in these matrices. In addition, it indicates that TOC plays little or no discernible role in the fate of PAHs in these matrices.³¹

The high molecular weight (HMW) PAHs represented 76.3 to 99.9% of the ∑16 PAHs in the indoor and outdoor dust from these areas. The dominance of HMW PAHs over LMW PAHs may be due to the lipophilic nature of HMW PAHs which enables them to be preferentially associated with the particulate phase, while the LMW PAHs are linked with gas phase partitioning which makes them more volatile and mobile. In addition, LMW PAHs exhibit multi-hop transport characteristics and are prone to leaching by both microbial and photo-degradation as compared with their HMW counterparts. Again, HMW PAHs are associated with high temperature combustion processes including gas flaring.^31,54

The 5-ring PAHs were the dominant homologues in the indoor and outdoor dust from EBD with significant contributions from BkF and BaP respectively. In the case of EME, 5-ring PAHs were the dominant species in the indoor dust with a significant contribution from DahA, while 3-ring PAHs were the dominant species in the outdoor dust with a significant input from Flu. The 6-ring PAHs were the dominant species in the indoor and outdoor dust from OTJ, with major contributions from IndP and BghiP respectively. The 3-ring PAHs accounted for <25% of the ∑16 PAHs in dust from the study areas except for those from EME. The low concentrations of 3-ring PAHs recorded in this study affirmed the fact that volatilization controls the fate of LMW PAHs.

On average, the prominent 3, 4, 5 and 6 ring PAHs in indoor dust from EME were Flu, Flt, DahA and IndP respectively, while those of the outdoor counterparts were Flu, Pyr, BaP and IndP for 3, 4, 5 and 6-ring PAHs respectively. In OTJ, the dominant 3, 4, 5 and 6 ring PAHs in the indoor dust were Ace, Chry, DahA and IndP respectively, while Phen, Chry, DahA and BghiP were the dominant 3, 4, 5 and 6-ring PAHs in the outdoor dust from OTJ. In the case of indoor dust from EBD, Ace, BaA, BaP and BghiP were the dominant 3, 4, 5 and 6-ring PAHs respectively, while Phen, Chry, BbF and IndP were the respective dominant 3, 4, 5 and 6-ring PAHs in the outdoor dust. This suggests transformation of PAHs arising from changes in the environmental characteristics.

Ecological risk of PAHs

The computed risk quotient (RQ) for both the negligible concentrations (RQ_(NCs)) and maximum permissible concentrations (RQ_(MPCs)) of PAHs in dust are shown in ESI Tables S6 and S7† respectively. The computed RQ_(NCs) values for indoor and outdoor dust varied from 116–98 200 and 43.5–28 400 respectively, while those for ΣRQ_(MPCs) ranged from 0.00–981 and 0.00–283 for indoor and outdoor dust respectively.

The ΣRQ_(NCs) and ΣRQ_(MPCs) values of PAHs were above 800 and 1 in the majority of the indoor and outdoor dust samples investigated. This suggests a high ecological risk associated with exposure to PAHs from these areas. However, of the ten outdoor dust samples from EME, only two sites had ΣRQ_(NCs) values above 800, while four samples of the outdoor dust had ΣRQ_(MPCs) values above 1. This indicates that exposure to PAHs in dust in the majority of sites from EME is of negligible ecological risk. The RQ_(NCs) and RQ_(MPCs) recorded in this study signify the need for remedial actions in order to reduce the ecological risks of PAHs in indoor and outdoor dust of these studied areas. BaP, DahA, IndP and BghiP were among the PAHs in the dust that contributed to the ecological risk of PAHs in the study areas.

Human health risk

BaP carcinogenic and mutagenic properties. The BaP_TEQ and BaP_MEQ concentrations for PAHs in these matrices are shown in ESI Table S8.† The ΣBaP_TEQ concentrations varied from 54–87 900 and 103–73 500 μg kg⁻¹ for indoor and outdoor dust respectively, while the BaP_MEQ concentrations varied from 194–126 000 and 117–21 300 μg kg⁻¹ for indoor and outdoor dust respectively.

On average the BaP_TEQ and BaP_MEQ concentrations of PAHs in dust (in- and outdoor) followed the sequence of OTJ > EBD > EME. The ∑BaP_TEQ concentrations recorded in this study were higher than those reported in previous studies.^6,31,34,55 The BaP_TEQ in these matrices surpassed the Dutch target value of 33 μg kg⁻¹ (ref. 56) and Method B cleanup level of 137 μg kg⁻¹ for soils.⁵⁷ BaP, DahA and IndP were the major compounds influencing the BaP_TEQ and BaP_MEQ concentrations of PAHs in these matrices.

Non-carcinogenic risk. The HQ values for PAHs in indoor and outdoor dust followed the order of HQ_ing > HQ_dermal > HQ_inh, (ESI Table S9†). The HQ_ing was the major contributor to the HI values of PAHs in dust from these areas. The HI values for adults' and children's exposure to PAHs in dust from these area ranged from 10⁻³ to 7.1 (Fig. 3a). The HQ_ing values for children's exposure to PAHs in dust from these areas were greater than the HQ_ing values for adults; this is attributed to the hand-to-mouth habits of children and their smaller body mass.

Fig. 3 (a) Hazard Index (HI) of PAHs in indoor ((A) Emu-Ebendo, (B) Otu-Jeremi, (C) Ebedei) and outdoor dust ((D) Emu-Ebendo, (E) Otu-Jeremi, (F) Ebedei) from rural communities around gas flaring points in the Niger Delta, Nigeria. (b) Total cancer risk (TCR) of PAHs in indoor ((A) Emu-Ebendo, (B) Otu-Jeremi, (C) Ebedei) and outdoor dust ((D) Emu-Ebendo, (E) Otu-Jeremi, (F) Ebedei) from rural communities around gas flaring points in the Niger Delta, Nigeria.

The HI values arising from exposure to outdoor dust were greater than those of the indoor dust in OTJ and EBD, whereas the HI values for exposure to indoor dust from EME were higher than those of outdoor dust. The HI values related to exposure to both indoor and outdoor dust from these areas were majorly <1 for adults and children, which indicate no considerable non-cancer risk resulting from human exposure to PAHs in dust from this area. However, there are a few sites in EME (IN-D1, IN-D2, IN-D3 and OUT-D2) with HI values for adults' and children' exposure greater than 1. Also, the HI values for children's exposure to PAHs in dust from sites IN-D3, IN-D4, IN-D6 and OUT-D1, OUT-D4, OUT-D7 of EBD exceeded 1 (ESI Table S9†). This suggests there is a probable adverse non-cancer risk for children at these sites in EBD and EME.

Incremental life-time cancer risk (ILCR). The ILCR represents the potential age specific cancer risk resulting from human exposure to PAHs in indoor and outdoor dust via the three exposure pathways which includes non-dietary ingestion (ILCR_ing), inhalation (ILCR_inh) and dermal contact (ILCR_derm). The potential cancer risk resulting from human exposure to PAHs in indoor and outdoor dust from these communities is shown in Fig. 3b and ESI Table S10.† The total carcinogenic risk values were in the magnitude of 10⁻¹ to 9.00 for both adults' and children's exposure scenarios.

The ILCR_derm values were higher than those of ILCR_ing and ILCR_inh for both adults and children in the studied areas. This implies that human exposure via non-dietary ingestion contributed more to the total cancer risk. The total cancer risk values for children were higher than those of adults. This may be due to the smaller body weight of children. The total cancer risk values and those related to exposure through unconscious oral ingestion and dermal contact with PAHs in dust by both adults and children in these areas were higher than the acceptable limit of 10⁻⁶. This implies a high potential cancer risk resulting from human exposure to dust from these areas.

The total cancer risk values recorded in this study were higher than those previously reported from human exposure to PAHs in dust from urban and rural areas of Delta State in Nigeria.^6,25 The ILCR results recorded in this study imply that there is an urgent need for remedial actions and source control measures in order to reduce the risk resulting from human exposure to PAHs in these environments.

PAH source analysis

The ratios of Ant/(Ant + Phen) ranged from 0.0 to 0.93 and 0.06 to 0.89 for indoor and outdoor dust respectively. An Ant/(Ant + Phen) ratio < 0.1 is related to petroleum sources, while an Ant/(Ant + Phen) ratio > 0.1 is associated with combustion processes. The ratio values of Ant/(Ant + Phen) in the indoor and outdoor dust from the studied areas advocate the prevalence of PAHs originating from combustion processes. The ratio values of Flt/(Flt + Pyr) in the indoor and outdoor dust from these areas ranged from 0.0 to 0.89 and 0.0 to 1.00 respectively. Flt/(Flt + Pyr) values between 0.4 and 0.5 describe sources linked to burning of liquid petroleum, and values > 0.5 are associated with burning of biomass, wood and coal.⁵² The ratio values of Flt/(Flt + Pyr) were greater than 0.4 in 60 to 70%, 50 to 60% and 30% of the indoor and outdoor dust from EME, OTJ and EBD respectively. The Flt/(Flt + Pyr) values advocate the prevalence of PAHs originating from burning of liquid petroleum and wood in dust from EME and OTJ, and contamination with petroleum-derived PAHs for dust from EBD.

Ratios of BaA/(BaA + Chry) < 0.2 and IndP/(IndP + BghiP) < 0.2 indicate petrogenic sources, while ratios of BaA/(BaA + Chry) between 0.2 and 0.35, and IndP/(IndP + BghiP) ratios between 0.2 and 0.5 describe sources related to the combustion of petroleum such as liquid fossil fuels, and vehicles and crude oil. Ratios of BaA/(BaA + Chry) > 0.35 and IndP/(IndP + BghiP) > 0.5 suggest inputs from biomass and coal combustion.^52,58 The ratios of BaA(BaA + Chry) and IndP/(IndP + BghiP) in indoor and outdoor dust ranged from 0.00–0.99 (Fig. 4 and ESI Table S11†), indicating multiple sources of PAHs which include petrogenic sources, combustion of wood, liquid fossil fuels, crude oil and vehicles. The IndP/(IndP + BghiP) ratios in 40 to 70% of the in- and outdoor dust from EBD, EME and OTJ were greater than 0.5 which advocates the prevalence of PAHs originating from burning of biomass and wood.

Fig. 4 PAH cross plots representing Ant/(Ant + Phen) vs. Flt/(Flt + Pyr) for Emu-Ebendo (A1), Otu-Jeremi (A2), and Ebedei (A3), and BaA/(BaA + Chry) vs. IndP/(IndP + BghiP) for Emu-Ebendo (B1), Otu-Jeremi (B2), and Ebedei (B3) in indoor dust and outdoor dust.

The ΣCOMB/TPAH ratio describes the relationship between the combustion profile of typical organics and their origins.¹⁶ A ΣCOMB/TPAH ratio < 0.3 indicates a contribution from petrogenic sources, while values ranging from 0.3 to 0.7 indicate a contribution from mixed sources, and values > 0.7 indicate PAH contributions from combustion processes at high temperature. The ΣCOMB/TPAH values in these matrices varied from 0.07–1.00. The ΣCOMB/TPAH ratio suggests dominance of PAHs originating from high temperature combustion sources in the studied areas.

The total index values for indoor and outdoor dust ranged from 1.75–13.7 and 4.13–13.6 respectively, with the majority of the samples having total index values greater than 4, which indicates that PAH concentrations in indoor and outdoor dust of these areas arise from high temperature combustion processes such as gas flaring.

PCA like other receptor models such as positive matrix factorization (PMF) and chemical mass balance (CMB) has its strengths and weaknesses. For instance, CMB might misspecify sources and have a collinearity problem which would give rise to negative source contributions. PCA and PMF assume that all sources have been identified and might fail to separate out some factors (sources) because of the high correlation coefficient in these source contributors and the similarity in their source profiles. Therefore, no single model is sufficient enough to give feasible results for sources,^59,60 but a combination of these models gives better output for source apportionment. However, PCA is easier to use, gives qualitative information about the sources, and provides a good visual output of the results.

The PCA of PAHs in indoor and outdoor dust of EME was resolved into three components which represented 92.9% and 95.6% of the variance respectively (ESI Tables S12 and S13†). For the indoor dust, Factor 1 captured 55.5% of the total variance and was dominated by Nap, Ace, Acy, Flu, Phen, Ant, Flt, Pyr, and BaA. Apart from Nap, Factor 1 was dominated by 3- and 4-ring PAHs. Nap is a marker for incomplete combustion processes,^16,31,61 while Chry, BaA, Pyr and Phen are considered as markers for natural gas combustion.^62,63 Acy is believed to come from liquefied petroleum gas, natural gas and coal combustion.^64,65 Factor 2 contributed 23.1% of the total variance with Chry, BbF, BkF and DahA as dominant PAH compounds. Chry, BkF and DahA are characteristic compounds for diesel emissions,^30,66 while BbF is a tracer of vehicular emissions.^30,67 Factor 3 was dominated by BaP, IndP and BghiP and contributed 14.4% of the total variance. These are tracers of gasoline/gas engine emissions and combustion of heavy oil.^68,69

For outdoor dust from Emu-Ebendo, Factor 1 captured 77.7% of the total variance and consisted of NaP, Ace, Acy, Flu, Phen, Ant, Flt, Pyr, Chry, BbF, BkF and BaP. NaP is a marker for incomplete combustion.^16,30,61 As stated earlier, Chry, Pyr, and Phen are tracers for natural gas combustion.^62,63 Acy is associated with liquefied petroleum gas, natural gas and coal combustion,^64,65 while Ace, Acy, Flu, Phen, Ant, Flt and Pyr are associated with burning of biomass, wood and liquid petroleum. Chry and BkF are indicators of diesel emission related sources,³⁰ while BaP and BbF are indicators of automobile emission related sources.^30,67 Therefore, the PAHs from Factor 1 came from mixed sources including combustion of natural gas (gas flaring), wood and vehicular emissions. Factor 2 accounted for 10.3% and contained BaA and BghiP as the dominant PAHs. These are indicators of gasoline emission related sources. Component 3 contributed 7.6% of the total variance with DahA as the only dominant compound. DahA is an indicator of diesel emission sources.⁶⁴

The PCA for PAHs in indoor and outdoor dust of OTJ was resolved into 5 components each, explaining 92.3% and 92.7% of the total variance for indoor and outdoor dust respectively. In indoor dust, Factor 1 accounted for 23.5% of the total variance with Ant, BaA and BaP as dominant compounds. These are indicators of coal combustion and vehicular emissions.^30,70 Factor 2 was dominated by Ace, Pyr, BkF and BghiP and accounted for 23.3% of the total variance. Pyr and Ace are indicators of low temperature combustion (wood/biomass combustion) related sources,^68,69 while BkF and BghiP are tracers of diesel combustion (diesel vehicle emissions).³⁰ Factor 3 contributed 19.2% of the total variance and was dominated by Flu, Flt and IndP. Flu and Flt are characteristic of wood/biomass and fossil fuel combustion,^68,69 while IndP is an indicator of gasoline emissions. Component 4 accounted for 13.3% of the total variance and was dominated by Phen. This is related to emissions from fossil fuel combustion. Factor 5 was dominated by NaP and Acy. NaP is a marker for incomplete combustion, while Acy is derived from liquefied petroleum gas, natural gas and coal combustion.^64,65

For the outdoor dust from OTJ, Factor 1 accounted for 46.9% of the total variance with NaP, Acy, Flu, Phen, Ant, Flt, Pyr and BaA as the dominant PAHs. NaP as earlier stated is an indicator for incomplete combustion, and BaA, Pyr and Phen are typical markers for natural gas combustion,^62,63 while Acy, Flu, Phen, Ant, Flt, Pyr and BaA are typical markers for wood/biomass and fossil fuel combustion.^30,71 Therefore, Factor 1 consists of PAHs arising from gas flaring and biomass combustion. Factor 2 was dominated by Ace, Chry and BaP and contributed 15.5% of the total variance. These are compounds from automobile emission related sources.^16,72 Factor 3 was dominated by BbF and accounted for 10.9% of the total variance. BbF is a tracer for vehicular emissions.⁶⁷ Factor 4 contributed 10.1% of the total variance and was dominated by IndP which is an indicator of gasoline emission related sources. Factor 5 accounted for 9.4% of the total variance with BkF as the dominant compound. This is an indicator of diesel emission.³⁰

The PCA for PAHs in EBD was resolved into 3 components for indoor dust and 4 components for outdoor dust, explaining a total variance of 89.0% and 92.7% for indoor and outdoor dust respectively. In indoor dust, Factor 1 accounted for 37.7% of the total variance and was dominated by NaP, Ace, Acy, Flu, Phen, Ant, and Flt. NaP is an indicator for incomplete combustion, while the rest of the PAH compounds are markers of burning of biomass, wood and liquid petroleum. Factor 2 contributed 31% of the total variance with Chry, BkF, BaP, DahA and BghiP as dominant PAH compounds. BaP and BghiP are tracers of automobile emissions,^16,72 while BkF, Chry and DahA are compounds of diesel emissions.⁶⁷ Factor 3 was dominated by Pyr, BaA, BbF and IndP and accounted for 20.2% of the total variance. These compounds are markers for coal combustion and vehicular emissions.⁷⁰ BaA and Pyr are tracers for natural gas combustion.^62,63

In outdoor dust, Factor 1 accounted for 26.4% of the total variance with Chry, BbF, DahA and IndP as dominant compounds. BbF and IndP are compounds of emissions from vehicular transportation,^30,55 while Chry and DahA are markers for diesel emissions as stated earlier. Factor 2 contained NaP, Ace, Ant, Flt and BaA as dominant compounds and contributed 26.0% of the total variance. NaP as earlier stated is a marker for incomplete combustion. Factor 3 accounted for 23.1% of the total variance and was dominated by Acy, Flu, Phen and Pyr. These compounds are markers for wood/biomass and fossil fuel combustion.^68,69 Factor 4 contained BkF, BaP and BghiP as dominant compounds and accounted for 17.2% of the total variance. These compounds are tracers of automobile emissions. The PCA results showed that the incomplete combustion resulting from gas flaring was one of the major sources of PAHs in the indoor and outdoor dust of the studied locations.

Conclusions

The results from this study revealed that there was a discernible impact of gas flaring on the PAH concentrations in indoor and outdoor dust from the study areas. The PAH homologue distribution in these media showed dominance of 5- and 6-ring PAHs which are associated with high temperature combustion processes. PAH source analysis suggests that the PAH contamination of dust from these rural communities arises from gas flaring, combustion of wood/biomass, liquid fossil fuels and vehicular emissions. Exposure to PAHs in these matrices is of potential risk for both humans and the ecosystem. The PAH contamination status of indoor and outdoor dust from these areas was in the sequence of OTJ > EME > EBD. The PAH concentrations in dust from the majority of sites from OTJ and EME exceeded intervention values and therefore, clean up, remedial actions and stringent pollution control measures are necessary for these areas in order to ameliorate the adverse consequences of exposure to PAHs.

Conflicts of interest

There are no conflicts of interest to declare.

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Footnote

† Electronic supplementary information (ESI) available. See DOI: https://doi.org/10.1039/d3em00048f

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