Sources and risk evaluation of polybrominated diphenyl ethers in dust and soil from an urban environment in Nigeria

Study area

Port Harcourt (4.815554°N, 7.049844°E) is one of the major cities in Nigeria (Fig. 1). The climate of the city is characteristic of a tropical area with distinct dry (November–March) and rainy seasons (April–October). The average annual rainfall in PHC is 2500 mm with an average daily temperature of 25–28°C.

Fig. 1.

Map of the study area (LGA, local government area).

Sample collection

Twenty composite samples each of soils and indoor and outdoor dust samples were obtained from different sites within the city. The soils were collected at depths of 0–10 cm with the aid of a soil auger, whereas dust samples were collected from surfaces within indoor and outdoor settings by gentle sweeping of the dust into a dustpan with a brush. After each sampling, the brush and dust pan were washed with detergent and rinsed with distilled water and n-hexane before they were used to collect a sample from another site. The indoor dust samples were obtained from homes and shops, whereas the outdoor dust and soil samples were obtained from a 100-m² area of the point where the indoor dust samples were collected. For each site, 7–10 samples were collected for each matrix and homogenised to form a composite. The site characteristics, sampling quality control processes, transportation, sieving and storage prior to chemical analysis, as well as the methods for determination of physicochemical properties, have been previously reported (Ossai et al. 2021, 2023).

Reagents and chemicals

The extraction solvents, such as acetone (≥99.9%), dichloromethane (DCM) (≥99.9%) and n-hexane (≥97.0%), were supplied by Honeywell Research Chemicals (Muskegon, MI, USA). Alumina, anhydrous sodium sulfate and silica gel were products of Merck (Darmstadt, Germany). A mixed standard containing 39 PBDE congeners (Table 1) was supplied by AccuStandard Inc. (New Haven, CT, USA) and a surrogate PBDE mixed standard containing five ¹³C₁₂-labelled BDEs (28, 47, 99, 153, 183) was supplied by Cambridge Isotope Laboratories, Inc. (Tewksbury, MA, USA).

Extraction and analysis

The United States Environmental Protection Agency method 3540C was applied for BDE extraction from dust and soil samples (United States Environmental Protection Agency 1996). That is, 5.0 g of the dust or soil was homogenised with 5.0 g of Na₂SO₄, followed by the addition of 20 ng g⁻¹ of ¹³C₁₂-labelled PBDEs. The homogenate was subjected to Soxhlet extraction with a 100-mL mixture of 1:1:1 (v/v/v) acetone/DCM/n-hexane for 10 h. The extract was concentrated to a volume of 2 mL by vacuum rotary evaporation and subsequently purified on a column loaded with alumina and silica gel. The PBDEs were eluted from the column with 30 mL of 1:1 (v/v) DCM/n-hexane. The eluate was evaporated to 2 mL with the aid of a gentle stream of pure N₂ gas.

The PBDEs in the extracts were separated and quantified by gas chromatography–mass spectrometry (GC-MS) with an Agilent 7890A gas chromatograph coupled to an Agilent 5975C mass selective detector programmed in selected ion monitoring (SIM) mode (Agilent Technologies, Inc., Palo Alto, CA, USA). The PBDEs were separated on a DB-5 capillary column (15-m × 0.25-mm × 0.25-µm film thickness). The mobile phase was ultrapure helium at a constant flow rate of 1 mL min⁻¹. Sample injection was in splitless mode with an injection volume of 2.0 µL. The GC oven was initially programmed at 85°C for 3 min, and then the temperature was increased at 25°C min⁻¹ to 200°C, and finally at 5°C min⁻¹ from 200 to 300°C. The PBDEs were identified by matching the sample retention times with those of pure PBDE standards, and from the m/z values of the target and confirmation ions.

Quality assurance and quality control

Replicate analyses, field and procedural blanks and recoveries achieved for sample matrix spikes and those of ¹³C₁₂-labelled PBDEs were used to validate the method. All glassware was thoroughly washed, cleaned with acetone and baked at 450°C for 4 h in a muffle furnace. The field blanks consisted of the extraction reagents taken to the field and returned to the laboratory alongside the samples and were then subjected to all the analysis steps as for the samples. The concentrations of the targeted PBDE congeners in the field and method blanks (n = 5) were below their respective detection limits. The method precision (n = 3) was ≤9%. The recoveries for the sample matrix spikes and those of ¹³C₁₂-labelled PBDEs were 79.9–97.3 and 78.9–101% respectively. An external calibration method was employed to determine the PBDE concentrations. The limits of detection (LODs) and limits of quantification (LOQ) corresponded to the concentrations that produce a 3:1 and 10:1 signal-to-noise ratio respectively. The LODs for the PBDEs varied from 0.007 to 0.021 ng g⁻¹, and the LOQ values ranged from 0.02 to 0.07 ng g⁻¹ (Supplementary Table S1).

Statistical analysis

Any significant differences between site PBDE concentrations were determined with the Kruskal–Wallis test, whereas the normality of the data was ascertained from the Shapiro–Wilk test. The mean concentrations of PBDEs in these matrices were compared by using the Tukey post hoc test. The relationship between the PBDE contamination status and their possible sources was established from principal component analysis (PCA) and hierarchical cluster analysis (HCA). The concentrations in the samples were analysed using the Kaiser–Meyer–Olkin (KMO) and Bartlett’s Test of Sphericity to confirm their suitability for PCA analysis. All statistical analyses were performed with SPSS software (ver. 15.1, see https://www.ibm.com/products/spss-statistics) at a significance level of P < 0.05.

Ecological and human health risk of PBDEs

The risk quotient (RQ) method (Chokwe et al. 2019) and the model equations of the United States Environmental Protection Agency (1986, 2009) were applied to evaluate the ecological and human health risks of PBDEs. A detailed description of the RQ method and the model equations is provided in the ‘Ecological risk of PBDEs in soil and dust’ section in the Supplementary material. The toxicological parameters and variables for human health risk assessment of PBDEs are provided in Supplementary Tables S2 and S3.

Concentrations and profiles of PBDEs

A summary of the PBDE concentrations in these media is presented in Table 1, whereas the individual site concentrations are presented in Fig. 2 and Supplementary Tables S4–S6. The Σ39 PBDE concentrations ranged from 1.69 to 590 ng g⁻¹ for soil and from 0.45 to 112 and 0.54 to 60.4 ng g⁻¹ for indoor and outdoor dust respectively. The Σ39 PBDEs and individual congeners had skewed concentration distributions with a reasonably high coefficient of variation, which implies that the PBDE concentrations in these matrices varied widely and originated from a wide range of variable sources. On average, the PBDE loads of these media followed the order: soil > indoor dust > outdoor dust. The same trend was reported for PAHs in these media (Ossai et al. 2021). This is related to the higher sorption and retention capacity of soil. In addition, soil receives the impact of most anthropogenic activities directly or indirectly from atmospheric transportation and deposition. Apart from the above, poor waste management practices, disposal of house-swept soils and dust, displays of goods outside shops during sales, discharges from polymeric roofing sheets, obsolete vehicles and household items, and other materials that are kept outside homes can increase the PBDE burden of urban soils compared to that of indoor dust. In addition, the PBDE contamination status of indoor dust depends on cleaning habits, level of furnishing, amount of electronic equipment, and air exchange between the indoor and outdoor environment. The concentrations of Σ39 PBDEs in the indoor dust samples were higher than those of outdoor dust, which may be associated with discharges from household items and the influence of indoor environmental conditions (Wilford et al. 2004; Hazrati and Harrad 2006; Toms et al. 2009).

Fig. 2.

Individual site PBDE concentrations in (a) indoor dust, (b) outdoor dust, and (c) soil from Port Harcourt city.

Despite the differences in the number of congeners analysed, the PBDE concentrations in our samples are compared with those reported for the same matrices from different regions of the world but with diverse anthropogenic pressures in Supplementary Table S7. The indoor dust from PHC had lower PBDE concentrations than those from China (Sun et al. 2016; Niu et al. 2018), Nigeria (Olukunle et al. 2015; Adeyi et al. 2020; Akinrinade et al. 2021; Ibeto et al. 2021), South Africa (Kefeni et al. 2014; Abafe and Martincigh 2015), Egypt (Hassan and Shoeib 2015; Khairy and Lohmann 2018), Melbourne, Australia (McGrath et al. 2018), Aveiro and Coimbra, Portugal (Coelho et al. 2016), South Korea (Kim et al. 2016), Jeddah, Saudi Arabia (Ali et al. 2016), Basrah, Iraq (Al-Omran and Harrad 2016) and Hanoi, Vietnam (Hoang et al. 2021), but were similar to those from Croatia (Klinčić et al. 2021). The PBDE concentrations in soils from PHC surpassed those reported for urban soils from municipal waste dumps in Abuja, Nigeria (Oloruntoba et al. 2021), India, Vietnam, South Korea, China (Wu MH et al. 2015; Li et al. 2016), Nepal (Yadav et al. 2018), Punjab Province, Pakistan (Syed et al. 2013), Melbourne, Australia (McGrath et al. 2016), and e-waste recycling sites in Tianjin, China (Wu Z et al. 2019), but were lower than those of urban soils of Japan (Li et al. 2016) and e-waste sites in western Nigeria (Folarin et al. 2024). PBDE concentrations in outdoor dust were not as high as those from China (Yu Y-X et al. 2012; Wu MH et al. 2015; Wu Z et al. 2022), Northern Vietnam (Anh et al. 2018) and Alexandria, Egypt (Khairy and Lohmann 2018), and e-waste sites in Nigeria (Folarin et al. 2024).

The PBDE compositions in these media are illustrated in Fig. 3. On average, the PBDE homologue concentrations of dust (indoor and outdoor) followed the order: penta-BDEs > di-BDEs > tri-BDEs > hexa-BDEs > tetra-BDEs > mono-BDEs > hepta-BDEs > deca-BDE, whereas those of soils followed the order: tri-BDEs > di-BDEs > penta-BDEs > tetra-BDEs > hexa-BDEs > mono-BDEs > hepta-BDEs > deca-BDE. This implies that di-, tri- and penta-BDEs are the top three PBDE homologues in these matrices. These patterns are visualised in the dendrogram in Fig. 4 where the homologues are clustered based on their abundance in specific sites. The differences in the concentrations and relative proportions of the PBDE homologues in the individual sites of these matrices are influenced by a number of factors. For example, the concentrations and proportion of BDE homologues in soils reflect the diversity and strength of the input sources and the soil physicochemical properties, which determine its retention capacity. In addition, rainfall aids the leaching of these compounds in the soil, whereas light intensity (sunshine), temperature, microbial communities and presence of iron and iron sulfide in the soils have been known to influence debromination processes, especially for higher molecular weight BDEs (Keum and Li 2005; Song et al. 2015; Pan et al. 2016). Specifically, sites close to electronic repair workshops (e.g. S-7, S-8 and S-14) and those with high traffic density and industrial activities (e.g. S-11, S-12 and S-18) showed elevated concentrations of PBDEs than other sites. In addition to these factors, wind also influences the distribution of PBDEs in dust. In the case of indoor dust, the differences in individual site BDE concentrations and profiles are influenced by types of furnishing, electrical appliances, levels of illumination, and air exchange between indoor and outdoor environments.

Fig. 3.

Congener profiles of PBDEs in indoor dust (ID1–ID20), outdoor dust (OD1–OD20), and soil (S1–S20) from Port Harcourt city.

Fig. 4.

Principal component analysis showing the biplots for PBDEs in (a) indoor dust, (b) outdoor dust, and (c) soil from Port Harcourt city.

The proportions of the mono-BDEs in these matrices relative to the Σ39 PBDEs varied between 0.0 and 53.3%. Apart from ID-8 to ID-10, OD-9 and OD-10, the mono-BDEs captured less than 19% of the Σ39 PBDEs in all the other samples. Of the three mono-BDE congeners analysed, BDE-3 was dominant in the indoor dust and soil, whereas BDE-2 was dominant in the outdoor dust. The di-BDEs constituted 0.0–64.2% of the Σ39 PBDEs in these matrices. The di-BDEs were the major proportion of the Σ39 PBDEs in 27% of these samples. BDE-15 was the dominant di-BDE in these matrices. The tri-BDEs were responsible for 0.0–62.1% of the Σ39 PBDEs in these matrices. On average, the BDE-25 concentrations exceeded those of other tri-BDEs in indoor and outdoor dust, whereas BDE-37 was dominant in soils. The tri-BDEs were the most prominent homologues in ID-3, ID-4, ID-17, S-2, S-8 and S-12. In this study, the mono- to tri-BDEs constituted >50% of the Σ39 PBDEs in 20–60% of these media. The proportion of mono- to tri-BDEs in these media was in the order: indoor dust > soil > outdoor dust. The lower brominated BDEs (LBDEs), such as the mono- to tri-BDEs, could be generated by debromination of higher-brominated BDE congeners (HBDEs) in the environment and organisms (de Wit 2002; Du et al. 2013; Chen et al. 2018; Yao et al. 2020). However, there is limited direct evidence of the fate of LBDEs in the environment. Therefore, the prominence of LBDEs in indoor dust is most likely related to debromination processes within the indoor environment.

The relative proportions of tetra-BDEs to Σ39 PBDEs in these media were less than 29% except for ID-14 and OD-12. The tetra-BDE congeners were not detected in ID-7–10, OD-11 and OD-12. On average, BDE-75 concentrations were above those of other tetra-BDEs in these matrices. The penta-BDEs accounted for up to 83.2% of the Σ39 PBDEs in some of these samples. The penta-BDE homologues were the major congeners in 15% of the indoor dust and 40% of soils and outdoor dust (Fig. 3). On average, BDE-118 concentrations were higher than those of the other five penta-BDE congeners analysed in the indoor dust and soils, whereas the concentration of BDE-100 exceeded those of the other congeners in the outdoor dust. The detection frequency (DF) of the hexa-BDE congeners in these matrices was 85–90%. The hexa-BDEs captured 0.0–57.1% of the Σ39 PBDEs in these matrices. The hexa-BDEs were the most dominant homologues in 10–30% of these samples. BDE-137 was the dominant hexa-BDE in indoor dust, BDE-153 in outdoor dust and BDE-154 in the case of soils. The hepta-BDE congeners accounted for less than 25% of the Σ39 PBDEs in these matrices, except for ID-15 (50.2%). The average concentrations of BDE-181 exceeded those of BDE-183 in these matrices.

The DF of deca-BDE in these matrices ranged between 70 and 85% with concentrations of <LOQ to 17.8 ng g⁻¹. The average BDE-209 concentration in the soils was above those of indoor and outdoor dust. Our results suggest that the deca-BDE is by far less prominent than other homologues in these matrices. The low proportions (<7% except for ID-13 [26.5%]) of deca-BDE relative to the Σ39 PBDEs in our samples suggest rapid debromination or limited use of the deca-BDE mixture in this region. Low DFs and concentrations of BDE-209 have been reported for indoor dust (Iwegbue et al. 2019) and sediments (Iwegbue et al. 2024b) from the region. BDE-209 is susceptible to debromination into LBDEs through microbial and photodegradation processes. In addition, the presence of iron and iron sulfides can cause reductive debromination of deca-BDE (Keum and Li 2005; Song et al. 2015; Pan et al. 2016). Photolytic debromination of BDE-209 has been known to produce tetra- to octa-BDEs (Söderström et al. 2004; Ahn et al. 2006; Stapleton and Dodder 2008) with BDE-183 (hepta-BDE) as the debromination intermediate (Xu et al. 2019).

The summary statistics of pH, electrical conductivity (EC) and total organic carbon (TOC) contents of soils and dust from PHC is given in Supplementary Table S8. A detailed discussion on the occurrence patterns and trends of these physicochemical properties in soils and dust from PHC has been previously reported (Ossai et al. 2021, 2023). The sorption and environmental fates of organic pollutants, such as PBDEs, are influenced by organic matter. The relationships between TOC and concentrations of Σ39 PBDEs and that of the individual homologues are illustrated in Supplementary Fig. S1a–c. The concentrations of Σ39 PBDEs and their homologues showed no relationship with the organic matter contents. This non-equilibrium sorption between TOC, concentrations of Σ39 PBDEs and that of the homologues implies that TOC has little or no influence on the distribution patterns of PBDEs in these matrices. Conversely, the distribution profiles rest on other factors, including recent pollution events, strength and types of input sources, and debromination processes. We observed similar trends for other organic pollutants (PAHs and PCBs) in dust and soils from PHC (Ossai et al. 2021, 2023). There was also a poor correlation between concentration of Σ39 PBDEs in soils and those of indoor and outdoor dust, which implies that the concentrations and profiles of PBDEs in indoor dust are not controlled by outdoor input sources, but rather by indoor input sources and transformations that take place within the indoor environment.

Ecological risk of PBDEs

The risk quotient (RQ) values for exposure of organisms to PBDEs in these media are provided in Supplementary Tables S9–S11. The RQ values for tri‐, tetra‐, penta‐, hexa‐ and deca‐PBDEs in indoor dust varied from 0.0 to 0.88, 0.0 to 0.44, 0.0 to 164, 0.0 to 0.054 and 0.0 to 0.06 respectively, whereas those of outdoor dust varied from 0.0 to 0.33, 0.0 to 0.25, 0.0 to 58.9, 0.0 to 0.02 and 0.0 to 0.05 for tri-, tetra-, penta-, hexa- and deca-PBDEs respectively. In the case of soil, the RQ values for tri-, tetra-, penta-, hexa- and deca-PBDEs ranged from 0.001 to 5.30, 0.01 to 3.54, 1.95 to 253, 0.00 to 0.33 and 0.00 to 0.78 for tri-, tetra-, penta-, hexa- and deca-PBDEs respectively. The RQ values for exposure of organisms to penta-PBDEs in indoor and outdoor dust exceeded 1 in 70% of the samples, whereas those of tri-, tetra-, hexa- and deca-BDEs were less than 1 for all sites. In the case of soils, the RQ values for penta-PBDEs were above 1 for all the sites, whereas the RQ values for tri- and tetra-BDEs were above 1 in 25% of the sites, particularly for S7, S8, S11, S12, S14 and S18. The high ecological risk of tri- and tetra-BDEs in these sites is influenced by discharges from electronic repair works, printing, traffic and industrial activities. The RQ values suggest a potential threat to organisms from exposure to penta-BDEs rather than the other PBDE homologues in these media. The tri- and tetra-BDEs constitute an ecological threat in 25% of the analysed soils.

Human health risk of PBDEs

Exposure of adults and children to BDE-47, BDE-99, BDE-153 and BDE-209 as individual compounds, and for Σ39 PBDEs in dust and soils from PHC gave hazard index (HI) values of 0.0–1.33, and did not exceed 1 for the majority of the sites (Supplementary Table S12). This implies that exposure to BDEs for a considerable number of sites may not result in any adverse effects. However, children’s exposure to BDE-47 in soils from sites S7, S8, S11, S12 and S14 had HI values above 1. This indicates a probable non-carcinogenic risk for children resulting from exposure to BDE-47 at these sites.

The cancer risk (CR) values for exposure of adults and children to BDE-209 in the dust and soils were in the magnitude of 0–10⁻⁷, whereas those of the total PBDEs were in the magnitude of 10⁻¹⁰–10⁻⁶ (Supplementary Table S13). These values are less than the acceptable risk value of 10⁻⁶ and, hence, suggest that exposure to BDE-209 and Σ39 PBDEs in these media results in no carcinogenic risk to humans.

Source apportionment of PBDEs

To identify the pattern of occurrence and potential sources of PBDEs, PCA and HCA were performed on the datasets. The KMO results for BDE homologues in indoor dust, outdoor dust and soil were 0.609, 0.708 and 0.664 respectively. Additionally, Bartlett’s test is significant (P < 0.05), confirming the presence of sufficient correlations among the variables. The PCA biplots for the three environmental matrices are shown in Fig. 4, and the data are given in Supplementary Table S14. The PCA for the indoor dust has three components with a total variance of 88.5%. Factor 1 is characterised by positive factor loadings for mono-, di-, and tri-BDEs, which accounted for 37.6% of the variance. The LBDEs could originate from the debromination of higher brominated BDEs in the environment (de Wit 2002; Du et al. 2013; Chen et al. 2018; Yao et al. 2020). Factor 2 has a variance of 31.7% with high positive factor loadings for tetra-, penta-, and hexa-BDEs. The main components of the penta-BDE formulation are BDE-47, BDE-99 and BDE-100 (La Guardia et al. 2006). Hexa-BDEs, such as BDE-153 and BDE-154, are also constituents of the penta- and octa-BDE technical formulations. Penta-BDE products were generally used to soften polymers such as flexible polyurethane foam, which is used in making household furniture, bed mattresses, couch cushions, paints and lacquers, textiles, rigid polyurethane foam, adhesives, protective coatings for electrical wires and circuit boards, etc. (Kostenko et al. 2024). Factor 3 has a positive factor value for hepta- and deca-BDEs, and this explains 19.1% of the variance. Thus, the BDE homologues in factor 2 are related to discharges from household furniture, bed mattresses, paints, textiles, etc., whereas factor 1 is likely related to debromination processes of higher brominated congeners since mono- to tri-BDEs are not constituents of any of the technical PBDE formulations. Factor 3 is associated with discharges from common plastics used in electrical and electronic equipment such as computers, televisions, air conditioners, washing machines, refrigerator components, stove hoods, audio and video equipment housings, cell phones, remote controls, computers, computer monitors, water storage tanks, plumbing pipes (Kostenko et al. 2024).

The PCA factor loadings for BDE homologues in soils and outdoor dust are similar. The PCA for BDE homologues in outdoor dust and soil has two components, each accounting for 91.5 and 90.4% of the total variances respectively. For the outdoor dust, factor 1 was associated with reasonably high positive factor loadings for tetra-, penta-, hexa-, hepta- and deca-BDEs, and explained 56.0% of the variance. The homologues in factor 1 are those associated with the constituents of penta- and deca-BDE commercial formulations as well as the debromination products of deca-BDE. As stated earlier, deca-BDEs undergo photolytic debromination to produce tetra- to octa-BDEs with hepta-BDE (BDE-183) as an intermediate (Söderström et al. 2004; Ahn et al. 2006; Stapleton and Dodder 2008). Factor 2 was responsible for 35.5% of the variance with positive factor loadings for mono-, di-, and tri-BDEs. The mono-, di-, and tri-BDEs are not constituents of the BDE commercial mixtures but are BDE degradation products of higher BDEs. Therefore, the BDE homologues in factor 2 are those associated with degradation of higher brominated BDEs.

For soils, factor 1 is characterised by reasonably high positive loading values for mono- to penta-BDEs, and this depicts 52.9% of the variance. Mono-, di- and tri-BDEs are degradation products of HBDEs, whereas tetra- and penta-BDEs are components of the penta-BDE commercial formulation (La Guardia et al. 2006; Cheng and Ko 2018). Therefore, the compounds in factor 1 are those associated with debromination processes and the use of the penta-BDE commercial mixture. Factor 2 has characteristic positive loading values for hexa-, hepta-, and deca-BDEs. Hexa- and hepta-BDEs are components of the octa-BDE commercial mixture, whereas the deca-BDE commercial formulation has deca-BDE as a major constituent (97–98%) (D’Silva et al. 2004). Thus, the homologues in factor 2 reflect those linked with octa-BDE and deca-BDE commercial formulations.

The HCA analysis grouped the targeted compounds into five clusters each for indoor dust, outdoor dust and soil samples, as shown in the dendrogram in Fig. 5. In indoor dust, cluster one is characterised by the highest concentration of penta-BDEs, influenced by high levels at sites ID1, ID16 and ID19. Cluster two includes di- and tri-BDEs, with sites ID5, ID6, ID11, ID17, and ID19 having the highest concentrations of di-BDEs, and sites ID3, ID4 and ID7 having the highest concentrations of tri-BDEs. Similar cluster patterns were observed in outdoor dust, where di- and penta-BDEs formed the first two clusters, representing the highest sum of homologue concentrations. The di-BDE concentrations were highest at sites OD4, OD8, OD9, OD10, OD16 and OD17, and penta-BDE concentrations were highest at OD1, OD2, OD3, OD7, OD8, OD18, OD19 and OD20. In soil samples, cluster one was dominated by tri-BDEs, representing the highest concentration of all homologues, followed by tetra- and di-BDEs, which was indicated in cluster two of the dendrogram in Fig. 5.

Fig. 5.

Dendrogram of the occurrence of PBDE homologues in (a) indoor dust, (b) outdoor dust, and (c) soil from Port Harcourt city.

Implications for environmental and human health

Annex A of the Stockholm Convention was amended to prohibit the global production, trade and usage of tetra- to hexa-PBDEs and deca-BDE in 2009 and 2017 respectively (United Nations Environmental Program 2023). At this time, a number of countries have also made plans to eliminate PBDEs from commerce and applied regulations necessary to avert the adverse consequences on humans and the ecosystem. For example, the production of penta- and octa-BDE formulations was regulated in the US in 2004, and that of the deca-BDE formulation was restricted in 2013 (La Guardia et al. 2024). However, PBDEs were never manufactured in Nigeria but were imported into the country as component chemicals used in the manufacture of plastics, polyurethane forms, textiles and other household items or from imports of electrical equipment, electronics and automobiles. Nigeria is one of the major importers of used electrical and electronic equipment (UEEE), vehicles and household equipment in Africa, which ends up as waste electrical and electronic equipment (WEEE), thus making Nigeria a dumpsite for e-waste, obsolete household equipment and vehicles. However, Nigeria banned the importation of used electronics in 2011 and specified that the importers of UEEE have to register with the National Environmental Standards and Regulations Enforcement Agency before commencing with importation as contained in the regulation (National Environmental [Electrical/Electronic Sector] Regulations 2010 S.1.No 23, National Environmental Standards and Regulations Enforcement Agency 2016). In addition, Nigeria is a signatory to the Basel Convention of 1992 designed to regulate the trans-boundary transportation of hazardous waste; and the Bamako African Treaty of 1991, specifically to restrict the importation and transport of hazardous waste both into and within Africa (Shittu et al. 2021; Maes and Preston-Whyte 2022; Okeke et al. 2024), but the implementation of these treaties is questionable. Despite the restrictions, the importation of these materials has continued, and inflow of e-waste has increased (Iwegbue et al. 2024b). To date, there is no inventory of PBDEs associated with electronic and electronic equipment (EEE), fire retardant mixtures used by firefighting industries, and those of other polymeric wastes in Nigeria. Additionally, Nigeria does not have adequate infrastructure for recycling and management of WEEE and other wastes, along with the non-existence of legislation on maximum permissible limits of PBDEs in wastes, which implies that there is no regulatory incentive to measure PBDEs in wastes. The management of WEEE in Nigeria involves mainly direct burning, direct disposal along with mixed household waste, and informal collection by waste brokers and scrappers who may only be interested in the metallic components of the WEEE and resort to open burning and extraction of metals by acid leaching (Shittu et al. 2021). Thus, the prevalence of UEEE, used vehicles and other household equipment, which are potential sources of PBDEs, along with the absence of legislation for PBDEs in Nigeria, could lead to proliferation and a wide range of PBDE concentrations in various environmental compartments and their consequent impacts on public health and the ecosystem.