Introduction

Interest in the characterization of urban soils and dust has increased greatly in the last two decades. Owing to the rapid growth of urban zones accompanied by increase in residential areas, streets, commercial, and industrial zones, increasing pollution levels have been recorded in urban environments1,2,3. Prompt industrialization has resulted in the contamination of terrestrial environments with various pollutants such as heavy metals (HMs). Various sources are responsible for very high concentrations of HMs in different environmental compartments of urban and suburban areas. The elevated levels of HMs in soil could be either of lithogenic or anthropogenic origin4. Anthropogenic activities including emissions from vehicles, industrial waste, atmospheric deposition of dust and aerosols, and incinerators have introduced HMs into environments at significant levels5,6. Thus, anthropogenic activities are the main reason for HMs contamination of soil as well as airborne dust7.

Dust storms are a natural phenomenon in desert ecosystems and their frequency have increased in some parts of the world since 1950s8. The dust can be contaminated with HMs from industrial and vehicle emissions9. Therefore, it is an important factor for urban pollution, particularly in arid and semi-arid regions of the world such as Saudi Arabia. Saudi Arabia is generally comprised of desert and is known by its hot and dry climate, with the highest average maximum summer temperature. Owing to the vast desert, harsh weather conditions, and low annual precipitation, Saudi Arabia faces many dust storms every year. The dust serves as a medium to accumulate HMs and other atmospheric contaminants emitted through various anthropogenic activities10.

There are different routes of consumption of airborne particulate metals by humans, including ingestion, dermal contact, and inhalation11. Ingestion of HMs-bearing dust and soil particles can pose a potential risk to human health, especially to children12. Therefore, the finest fractions of dust and soil can be used as a good indicators of the bio-accessibility of metals, and are more relevant to human health than whole soil owing to their capacity to adhere to skin or ingested or inhaled13. Previously, it has been reported that the bio-accessible fractions of manganese (Mn), nickel (Ni), and zinc (Zn) are higher in urban areas, thereby posing a potentially greater health risk14. HMs can have dangerous and toxic effects when they are present above certain concentrations, are not metabolized by the body, and accumulate in the soft tissues13,15.

Among the various pathways, soil and dust ingestion is a dangerous route (especially for children), mainly owing to “hand to mouth” activity during outdoor activities with an intake rate of 200 mg soil day−1, as indicated by the US-Environmental Protection Agency16. For instance, higher blood lead (Pb) values in children have been found to be associated with fine soil and dust particles ingestion17. Fine particles tend to adhere more efficiently to hands and thus can easily be ingested into the body18. It was found in two European cities i.e., Sevilla and Torino that the availability of HMs in the clay fraction (< 2 μm) was higher than that in other fractions (< 2, 2–10, 10–22, 22–50, and > 50 μm) or whole soils. It indicates that the bio-accessibility of HMs is expected to be higher in the fine fraction of soil. Thus, it is more related to human health than whole soil, which is mainly owing to its ability to adhere to skin or be ingested or inhaled as suspended dust13. Therefore, long term exposure to soil and dust contaminated with HMs can cause severe health effects, including lung, kidney, and liver damage as well as cancer. Chromium (Cr) exposure may lead to cancer, Pb exposure could affect cognitive development in children, enzymatic inhibition, as well as nervous and skeletal damage19, and cadmium (Cd) exposure could damage kidney, bone, and lungs20.

Riyadh has a hot and dry climate with the highest average maximum summer temperature. It in know for vast desert, extreme weather, higher traffic density, and industrial activities, which are causing the accumulation of significant quantities of dust in the city21. Al-Rajhi et al.22 investigated the levels of HMs in indoor and outdoor dust in Riyadh, and found that the old industrial area had higher levels of HMs. Likewise, El-Desoky et al.23 found that outdoor dust is correlated with indoor dust in Riyadh, thereby resulting in greater concentrations of Pb in indoor dust. They further reported that the concentrations of Pb in the blood of 17.8% of children were higher than the global limit (10 µg dL−1). Similarly, another city in Saudi Arabia, Mahad AD’Dahab is also facing higher pollution levels of HMs in its environment, which is mainly owing to its closeness to mining activities. Elevated levels of HMs contamination have been observed in sites next to the ground rock landfill in Mahad AD’Dahab24. Previously, Al-Farraj et al.25,26,27 have demonstrated that the soil of Mahad AD’Dahab contained very high levels of Cd, copper (Cu), Pb, and Zn. In this context, Al Bakheet et al.28 found that the concentrations of Pb, mercury (Hg), and Cd in the blood samples of Mahad AD’Dahab residents were higher than those of Riyadh residents. They also reported that HMs-prone groups are significantly associated with kidney disease, urinary tract disorders, growth disorders, blood diseases, and genetic disorders. Both Riyadh and Mahad AD’Dahab cities have been found to be polluted with HMs through anthropogenic activities such as vehicle transmission, industrial emission, and mining activities; however, the extent of HMs contamination in urban and suburban areas of both cities and their associated ecological and human health risks have not been explored yet.

Therefore, the objectives of this study were to (1) investigate the levels of different HMs in dust as well as different size fractions of soil collected from several urban and suburban areas of Riyadh and Mahad AD’Dahab; and (2) assess the pollution levels, sources, and associated ecological and human health risks posed by HMs in dust and soil (< 63 µm).

Materials and methods

Study area, sampling, and analyses

Soil and dust samples were collected from various urban and suburban areas of Riyadh and Mahad AD’Dahab (Supplementary Fig. S1a,b). Nineteen sites in Riyadh and five sites in Mahad AD’Dahab, including urban and suburban areas, were selected for soil and dust sample collection. Soil samples were taken as a compound samples from the same place dust sample was taken. Soil samples were collected at a depth of 0–3 cm. The soil samples were taken, which were affected by the falling dust as a result of industrial and mining activities. Dust samples were collected through a Marble Dust Collector measuring 52.5 × 31.5 cm (Supplementary Fig. S2). The dust samples were collected on a monthly basis from May 2014 to April 2015 and were divided into four groups according to the seasons of collection (summer, autumn, winter, and spring). The dust samples were collected using a soft plastic brush, stored in plastic bags, and transported to the laboratory for analysis. Soil and dust samples were air dried at room temperature (24–25 °C). Soil samples were passed through a 2 mm sieve and then divided into three fractions using sieves ranging as follow (bulk, 63–250 and < 63 µm). The total HMs (Al, Fe, Mn, Zn, Ti, Cu, Cr, Co, Ni, Pb, and Cd) content in the soil and dust was assessed after digestion in a microwave (MARS, CEM Corporation, USA) using a method reported by USEPA 305129 (total-recoverable). Specifically, 0.5 g of each dust or soil sample was place into a Teflon vessel, to which 10 mL of HNO3 was added. The vessels were then capped, placed in a microwave, and digested for 10 min according to the USEPA 3051 method. The digested samples were filtered through a 0.45 µm membrane and brought to a total volume of 50 mL with deionized water in a volumetric flask. HMs were measured using inductively coupled plasma optical emission spectrometry (ICP-OES; PerkinElmer Optima 4300 DV, USA). For quality control, soil and dust samples were analyzed in four replicates. Qtest was applied to exclude abnormal values at a confidence level of 95%.

Levels of HMs in soil and dust samples

In this study, contamination levels of HMs in soil and dust were characterized using the pollution index (PI), the integrated pollution index (IPI), the geo-accumulation index (Igeo), and the potential ecological risk index (RI). The pollution index and the IPI are used to assess environmental quality30. The PI is calculated as the ratio of the metal concentration in the sample to the background content of the corresponding metal in the lithosphere (content of the earth’s crust). The following classification of PI was used: PI ≤ 1, low level; 1 < PI ≤ 3, middle level; and PI > 3, high level. The IPI of all measured metals in samples was defined as the mean value of the PI of the metals. The classification of IPI was as follows: IPI ≤ 1, low level; 1 < IPI ≤ 2, middle level; and IPI > 2, high level31.

The Igeo method was used to calculate the metal pollution levels32. The Igeo is computed via the following equation.

$${I}_{geo}={log}_{2}\frac{Ci}{1.5Bi},$$
(1)

where Ci is the measured concentration of the metal i and Bi is the geochemical background value of the metal. In this study, Bi was the background content of the metal i (background in shale)33. The 1.5 constant was introduced to minimize the variation of background values. The following classifications were carried out according to Igeo: unpolluted, Igeo ≤ 0; unpolluted to moderately polluted, 0 < Igeo ≤ 1; moderately polluted, 1 < Igeo ≤ 2; moderately to strongly polluted, 2 < Igeo ≤ 3; strongly polluted, 3 < Igeo ≤ 4; strongly to extremely polluted, 4 < Igeo ≤ 5; and extremely polluted (5 < Igeo)32.

The potential ecological RI originally mentioned by Hakanson34 was also calculated to assess the degree of HMs pollution in soil and dust samples using the following equations.

$$RI=\sum_{i=1}^{n} Ei,$$
(2)
$$Ei=Ti fi,\mathrm{ and}$$
(3)
$$fi= \frac{Ci}{Bi},$$
(4)

where RI is the sum of all six risk factors for HMs, Ei is the monomial potential ecological risk factor, Ti is the metal toxic factor (with the values for each metal in the order of Zn = 1 < Cr = 2 < Cu = Ni = Pb = 5 < Cd = 30)35, fi is the metal pollution factor, Ci is the concentration of metals in dust, and Bi is a reference value for metals34. Different RI classifications of metal pollution are low ecological risk (RI ≤ 150), moderated ecological risk (150 ≤ RI < 300), considerable ecological risk (300 ≤ RI < 600) and high ecological risk (RI ≥ 600).

The EF is a convenient measure for assessing the degree of metal contamination and determining its probable natural and/or anthropogenic sources36. For normalization, a reference Fe concentration is used because of its natural abundance. The EF was calculated using the following equation37

$${EF}_{m}={\left[{C}_{m\left(soil sample\right)}/{C}_{Fe\left(soil sample\right)}\right]}{}\left[{C}_{m\left(earth crust\right)}/{C}_{Fe\left(earth crust\right)}\right],$$
(5)

where Cm is the content of the examined metal in the soil sample, CFe(soil sample) is the content of the reference metal (Fe) in the soil sample, Cm(earth crust) is the content of the examined metal in the earth’s crust, and CFe(earth crust) is the content of the reference metal (Fe) in the earth’s crust. In general, EF values much higher than 2 are mainly considered to indicate anthropogenic sources, while values less than 2 predominantly an origin in background soil material. Moreover, the EF also assists in determining the degree of metal contamination. Five contamination categories are recognized based on this factor: EF < 2 indicates deficiency to minimal enrichment; EF = 2–5, moderate enrichment; EF = 5–20, significant enrichment; EF = 20–40, very high enrichment; and EF > 40, extremely high enrichment38. In addition, the lithogenic and anthropogenic HMs content was calculated using the following equation39.

$${\left[M\right]}_{lithogenic}= {\left[Fe\right]}_{sample}\times ({\left[M\right]/\left[Fe\right])}_{lithogenic,}$$
(6)

where [M]lithogenic is the metal concentration of lithogenic origin in the sample (mg kg–1), [Fe] sample is the total content of Fe in the soil sample (mg kg–1), and (\(\frac{\left[M\right]}{\left[Fe\right]})lithogenic\) is the ratio of metal concentration to iron concentration in the earth’s crust. Moreover, the anthropogenic HMs content was calculated using the following equation.

$${\left[M\right]}_{anthropogenic}= {\left[M\right]}_{total}-{\left[M\right]}_{lithogenic},$$
(7)

where [M]anthropogenic is the anthropogenic HMs content and [M]total is the total content of HMs measured in soil samples.

The Distribution factor (DF) has been widely used to assess the distribution of HMs and environmental risks in the different particle size fractions5,40. The (DF) index was calculated by Eq. (8)41:

$$DF= {C}_{fraction}/{C}_{bulk},$$
(8)

where Cfraction and Cbulk (mg kg−1) are concentration of HMs in a given fraction and bulk samples, respectively.

Risk assessment

Exposure assessment

Risk assessment is a multi-step procedure of estimating the nature and probability of adverse human health effects that are caused by HMs in an environmental medium32. Risk assessment is based on the consideration of human exposure to soil or dust via three different pathways, namely oral intake (ingestion), inhalation, and intake through skin exposure (dermal intake). The average daily doses (ADDs) through ingestion and dermal contact for dust and soil (< 63 µm) were calculated according to the following equations:

$${ADD}_{ingestion}=\frac{{C}_{dust}\times {IR}_{ingestion}\times F\times EF\times ED}{BW\times AT},$$
(9)
$${ADD}_{dermal}=\frac{{C}_{dust}\times SA \times AF\times ABS\times F\times EF\times ED}{BW\times AT}.$$
(10)

All the definitions of the parameters and values of the variables for human health risk assessments are presented in Supplementary Table S1.

Non-cancer risk assessment

The non-carcinogenic quotients of exposure to HMs in dust and soil (< 63 µm) were calculated. Non-cancer risks are expressed as a hazard quotient (HQ). The HQ is the quotient of the ADD divided by the reference dose of a specific HMs and for the exposure through each pathway. The HQ of each metal was determined by the following equation42:

$$HQ=\frac{ADD}{RFD}.$$
(11)

To assess the overall potential non-cancer risk, the hazard index (HI) was calculated, as follows5:

$$HI=\sum HQ={HQ}_{ingestion}+{HQ}_{dermal}.$$
(12)

The value of HI ≤ 1 indicates that there is no significant risk of non-carcinogenic effects. On the other hand, there is a chance that non-carcinogenic effects may occur when HI > 1, and the probability increases as the value of the HI increases43.

Cancer risk assessment

The incremental lifetime cancer risk (ILCR) for an individual is estimated by multiplying the slope factor (SF) with the ADD over a lifetime exposure, as determined by Eq. (13).

$$ILCR= ADD \times SF.$$
(13)

An ILCR value of < 1.0E−06 is considered small, whereas an ILCR of 1.0E−06 to 1.0E−04 is in the range of the acceptable limit and an ILCR of > 1.0E−04 is likely to be harmful to humans44.

Quality assurance and quality control

Great care was taken to avoid any metal contamination during the process of sampling, digestion, and analyses from the beginning of the study. All equipment and containers were soaked in 10% HNO3 for 24 h and then rinsed thoroughly in deionized water prior to use. Each soil and dust sample was replicated four times during digestion and analyses. Qtest was used for the identification and rejection of outliers and was applied to exclude abnormal readings at a confidence level of 95%. Moreover, three standard reference soils (Till-1, Till-2, and Till-4) were employed for quality control in HMs analyses in soil (Supplementary Table S2). Standard solutions with known concentrations were simultaneously analyzed in the experiments after each set of 10 samples, to ensure the analytical performance of the ICP-OES apparatus. The imprecision of the method was computed as the relative standard deviation targeted at ≤ 5%. The detection limits of ICP-OES were < 0.1 µg L–1 for Cd and Fe and 1 µg L–1 for Al, Ti, Co, Cu, Cr, Pb, Mn, Ni, and Zn. Values of the studied metals that were below the detection limits of ICP-OES were rejected. The recovery % of HMs was calculated according to the following equation.

$$Recovery\%=({C}_{ex}/{C}_{ref})\times 100,$$
(14)

where Cex is the HMs concentration extracted by solution (mg kg–1) and Cref is the concentration of HMs in the reference soil “Till” (mg kg–1).

Supplementary Table S2. Recovery of HMs content in the three certified reference materials (Till 1, Till 2, and Till 4) digested using the EPA 3051 method.

Results and discussion

Distribution of heavy metal concentrations in the soil particle size fraction

Supplementary Table S3a,b show the minimum, maximum, and average concentrations of HMs in the different particle size fractions of urban and suburban soils. The HMs in the different fractions of urban and suburban soils were regularly distributed. Among the particle size fractions, the total content of most of the investigated HMs tended to increase as the size of the soil particles decreased. The highest content was generally pronounced for the particle size fraction of < 0.63 μm. The highest amount of metals that accumulated in the fine particles of < 63 μm could be explained by their high reactivity and their affinity toward HMs45. According to some researchers, HMs are often accumulated in the fine fraction, such as clay particles that act as metal sorbents, which is mainly due to their high surface area and negative surface charge40,45. It was generally observed that the total HMs concentrations in the mined soils of Mahad AD’Dahab were higher than those in the industrial activity-impacted soils of Riyadh. In this context, urban soils are often contaminated with HMs owing to anthropogenic sources. Previously, it has been reported that mining activities can result in significant metal accumulation in environmental compartments46. Alike, several studies found that mining operations are significant sources of HMs contamination in soils and higher contents of HMs in mining-impacted soils result from long-term activities47,48.

The Distribution factor (DF) has been widely used to assess the distribution of HMs and environmental risks in the different particle size fractions5,40. Supplementary Fig. S3a,b show the minimum, maximum, and average DFs for HMs in the different particle size fractions of urban and suburban soils. These obtained DF values indicated a greater metal accumulation in the finer fraction (< 63 µm) compared with that of larger size fraction (63–250 μm). Our data were in line with previous findings on the preferential partitioning of HMs to fine soil particle size fractions5,40,49. This can be explained by the larger surface area of the fine particles, which enhances the adsorption capacity of the fine fraction. Additionally, finer soil particles can have higher contents of secondary clay minerals, which are very strong sorbents for HMs49,50. On the contrary, the coarser fractions of sand and silt can have a higher content of the primary mineral quartz (e.g., SiO2), thereby leading to lower sorption capacity. It could be concluded that the fine particle fractions, especially the clay fraction, accumulated higher concentrations of HMs than the coarse fractions, thereby causing potential harm to human health and the environment50.

Supplementary Table S4 shows the Pearson correlation between the HM concentration in dust and soil < 63 µm over all the investigated areas. The Pearson correlation showed that there was a significant correlation between the total concentration of most HMs in dust and soil < 63 µm. The Pearson correlation was 0.55 for Cd, 0.67 for Cu, 0.53 for Fe, 0.66 for Mn, 0.52 for Pb, 0.55 for Ti, and 0.44 for Zn.

Heavy metals content in soil and dust samples

Table 1 shows the minimum, maximum, and mean concentrations of total HMs in the bulk soil samples collected from the investigated sites in the urban and suburban areas of Riyadh and Mahad AD’Dahab. The results indicated that the HMs content varied according to the metal type, sampling site, and study area. Generally, it was observed that the total HM concentrations in urban areas were higher than those detected in suburban areas. Among the two localities, the soil samples collected from Mahad AD’Dahab sites had the highest metal concentrations. For instance, the average HMs concentration in soil samples collected from urban Riyadh and Mahad AD’Dahab amounted to 5410 and 14,100 for Al, 0.012 and 0.197 for Cd, “nd” and 0.273 for Co, 9.08 and 13.2 for Cr, 0.702 and 18.5 for Cu, 4710 and 12,900 for Fe, 81.5 and 308 for Mn, 4.41 and 8.37 for Ni, 5.68 and 13.9 for Pb, 117 and 588 for Ti, and 13.7 and 54.3 for Zn (all in mg kg–1), respectively.

Table 1 Minimum, maximum and average content of heavy metals in soil samples of study area and average content of heavy metals values for world and Netherland soil.
Full size table

In the Riyadh area, the highest total concentrations of most HMs were found at site 16 (with the exception of Cu, Cd, and Co). In the Mahad AD’Dahab area, the highest total concentrations of Cd, Cu, Pb, and Zn were detected at site 3, which is close to the mining area. Moreover, the highest concentrations of Al, Co, Cr, Fe, Ni, and Ti were recorded at site 2.

In Saudi Arabia, quality guidelines for soil HMs have not been established. Therefore, in the current study, the concentrations of HMs in soil samples were compared with other guidelines, including the common range in the earth’s crust, the average concentrations in world soils, the average shale values, and the Dutch optimum and Act target values, as shown in Table 1. The concentrations of most investigated HMs were lower than their corresponding values of the common range in soil according to Ref.51. However, in urban Mahad AD’Dahab sites, the average concentrations of Cd, Pb, and Zn were higher than their corresponding values of the common range (0.06, 10, and 50 mg kg–1, respectively). Moreover, the maximum and average concentrations of Cu in urban Mahad AD’Dahab sites were higher than the average concentration in world soils.

Supplementary Table S5 shows the average metal concentrations in dust samples in relation to season. Generally, the highest average values of most of the HMs were detected in the spring season in urban and suburban areas of Riyadh and suburban areas of Mahad AD’Dahab. However, in urban areas of Mahad AD’Dahab, the highest average values of most HMs were found in the winter season. The variation in the order of the highest metal levels between these two seasons (winter and spring) could be explained by changes in meteorological conditions.

Table 2 shows the comparison of the obtained average metal concentrations with those reported for other countries. The comparison of HMs levels in the collected dust samples with those of other countries showed that the average content of HMs in the investigated sites in the current study were lower than those of most other cities (e.g., Riyadh (Saudi Arabia), Khamees-Mushait (Saudi Arabia), Jeddah (Saudi Arabia), Middle and South of Iraq, Kermanshah (Iran), Rafsanjan SE (Iran), Shangqing (China), ** cancer via HMs intake, followed by dermal contact and inhalation. Moreover, it was observed that carcinogenic risks for children via ingestion and dermal contact were more than that of adults64.

Conclusions

Levels and sources of HMs present in dust and soil of urban and suburban areas in Riyadh and Mahad AD’Dahab cities were studied and associated ecological and human health risks were estimated in this study. Overall, the levels of studied HMs were higher in dust samples than that of soil samples collected from the studied areas. The enrichment factor and PCA analyses exhibited that Al, Co, Cr, Fe, Mn, Ni, Ti, and Zn were of lithogenic origin, whereas, Cu, Zn, and Pb in Riyadh and Cd, Cu, Pb, and Zn in the Mahad AD’Dahab were of anthropogenic activities. The HI values for dust and soil (< 63 µm) samples were as: children up to 6 years > children from 6 to 12 years > adults. The HI values for all human categories for dust in urban areas of Mahad AD’Dahab and soil samples (< 63 µm) in urban and suburban areas of Mahad AD’Dahab were > 1, indicating non-carcinogenic risk. The average values of ILCR for dust and soil (< 63 µm) samples from both urban and suburban areas of Riyadh and Mahad AD’Dahab were close to the recommended threshold for Cr; however, were lower for Pb. Higher levels and ecological and human health risks of HMs in residential areas of Riyadh and Mahad AD’Dahab could be owing to industrial and mining activities in these cities. Therefore, ecological and human health risk assessment showed that soil and dust of Mahad AD’Dahab were more polluted as compared to Riyadh. It is therefore recommended to monitor HMs pollution in dust and soil particles of such residential environments for sustainable ecosystem and human health.