1 Introduction

Environmental pollution with potentially toxic elements (PTEs) has since received global attention due to rapid urbanization and industrialization worldwide [1]. PTE contamination of the environment obstructs human health and natural ecosystem functions [2]. The significant concern of PTEs is their wide sources, non-biodegradability, bioaccumulation, and toxicity [3]. The incidence of PTEs in the environment could be from natural or anthropogenic (man-made) sources such as mining, fossil fuel combustion, and cement production [1]. It has been reported that cement plants are the major anthropogenic sources of PTEs, especially Hg, Pb, Ni, As, Cd, and Zn, because of raw materials processing and fossil fuel combustion [4, 5]. The deposition of PTEs occurs at different distances around the cement plant in ambient air and is then deposited on the ground as dry or wet precipitation, which leads to deterioration of soil quality [6].

Soil is a substance obtained from the Earth’s surface and is considered the ultimate sink of anthropogenic PTEs released into the environment through various sources [7]. Soils are generally considered as carriers of most PTEs released during anthropogenic activities [8]. Contamination of soil with PTEs is a worldwide environmental concern and continues to increase due to massive increases in industrialization, urbanization, human population, and consumption patterns [9, 10]. Soil deterioration by PTE can cause lasting problems in the biogeochemical cycle and may influence soil’s natural function, leading to changes in soil fauna [11], agricultural soils, and food crops [12].

Studies have revealed that PTEs are potentially toxic to plants and animals when water and soils polluted with them are used for crop growth [13, 14]. Consumption crops grown in contaminated soil reach humans through the food chain, and these PTEs are likely to bioaccumulate, resulting in health disorders in the human system [15]. Exposure to PTEs is associated with health implications for the human system, such as liver failure, nervous disorders, kidney disease, and neurotoxicity, either by ingestion or inhalation [16, 17]. Chromium and nickel are considered poisonous, particularly with increasing long-term exposure, including their carcinogenic properties and their effects on fetal development. Chromium exposure has been associated with abnormal enzymatic activity, oxidation–reduction derangement, protein denaturation, and various medical disorders. Nickel carbonyl and hexavalent chromium are classified as carcinogens [18]. Humans can be exposed to arsenic through polluted food, water, and dust. Ingestion, inhalation, and skin absorption of food, water, soil, and dust contribute to cumulative arsenic exposure. Both acute and chronic arsenic exposures have been linked to several non-cancer health issues and an increased risk of skin, bladder, liver, and kidney cancer [19, 20].

Lead pollution in street dust and drinking water poses a major public health problem due to its negative impact on several physiological systems. This affects children, pregnant women, and other vulnerable groups. Leaching of lead into potable water from outdated pipes, plumbing fixtures, and soldered connections is the main source of lead exposure. Lead is absorbed into the circulatory system after intake and may accumulate in soft tissues, organs, and the skeleton. Low-level lead exposure in children can cause anemia, cognitive decline, developmental delays, and lower IQ. Lead exposure also raises blood pressure, renal failure, and reproductive issues in adults. Lead in water can have long-term health effects, even in tiny concentrations. Thus, strict regulatory regulations, infrastructural improvements, and public education are essential to decrease exposure hazards and protect human well-being [21,22,23,24]. Furthermore, there is a need for effective monitoring of PTE sources and distribution to reduce environmental damages, checkmate pollution, and public health as a result of anthropogenic activities [12].

Moreover, studies of PTE concentrations and risk assessment on soils have been reported in Nigeria [12, 25,26,27,28]. However, limited or no studies have been reported on geochemical, and risk assessments of the PTE study area. Hence, this study determined PTE concentration, and assessed geochemical, ecological, non-carcinogenic, and carcinogenic risks related to the activities in the Ashaka cement plant, Bajoga, Nigeria, with the view of providing information on the impact of the cement plant activities on geochemical, ecological, non-carcinogenic, and carcinogenic risks.

2 Materials and methods

2.1 Study area

The study area is the Ashaka cement plant located in Ashaka town, 9 km north of Bajoga, Gombe State, Nigeria. The map of the area is shown in Fig. 1, located between longitudes 11o 28ʹ 30″E and 11o 29ʹ 30″E and latitudes 10o55ʹ30″N and 11o56ʹ30″N. The climate is tropical with two distinct seasons (dry and wet). The dry season between November-February is harmattan, and the hot period between March–April is of the dry season with temperature ranging from 31.1 to 42.1 ℃. The maximum rainfall ranges from 800 to 900 mm (August–September) and a minimum of 250 mm to 350 mm (May–June), with relative humidity ranging from 60 to 80% [29].

Fig. 1
figure 1

A map of the study area showing sampling sites

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2.2 Sample collection and preparation

A systematic sampling technique was used for sample collection with little modification [30]. The surrounding soil of a cement plant was collected at depths ranging from 0 to 15 cm (topsoil). Four subsamples within each sampling point were mixed to form a composite sample. A total of ninety three soil samples were collected.

Approximately 50 g of soil samples were transferred into polyethylene bags, and subsequently transported to the laboratory and air-dried at room temperature for 2 weeks [31]. Before the determination of PTEs, substantial impurities were detached from the air-dried soil samples, pulverized using a mortar and pestle and sieved via a U.S. No. 10 (2 mm) mesh. Approximately 0.5 g of the soil was transferred into a Teflon cup, and a binary combination of 3.75 mL of HCl and 1.25 mL of HNO3 was added as previously reported [32]. The resulting solutions were then analyzed for PTE concentration using inductively coupled plasma optical emission spectrometry (ICP-OES, Agilent 720-ES).

2.3 Quality assurance

The precision of the analytical procedure was determined by a recovery study. The recovery study was conducted by determining the PTE concentration in triplicate in spiked and un-spiked samples. The ICP-OES (Agilent 720-ES) was first calibrated using a multi-element standard solution (QCSTD-27). The percentage of mean recovery of the PTEs ranged from 98 to 104%. The lowest limit of quantification (LOQ) and limit of detection (LOD) were established using the standard, the response’s standard deviation and the calibration curve’s slope are as follows: LOD = 3.3 σ/S, LOQ = 10 σ/S, where σ represents the response’s standard deviation and S denotes the calibration curve’s slope [33, 34]. The limits of detection for Al, As, Ba, Cd, Co, Cr, Cu, Fe, Hg Mn, Mo, Ni, Pb, Sb, Sc, Se, Sr, Ti, V and Zn were 0.013, 0.003, 0.004, 0.0001, 0.001, 0.0004, 0.0008, 0.01, 0.005, 0.0006, 0.0001, 0.005, 0.001, 0.0001, 0.0001, 0.003, 0.003, 0.003, 0.009 and 0.0003 μg/L, respectively.

2.4 Data analysis

The data obtained were statistically evaluated (simple descriptive and inferential) using SPSS software version 25. The generated data were used to estimate the potential ecological and health risks.

2.5 Geochemical load index (GLI)

The geochemical load index (GLI) was used to evaluate soil pollution by potentially toxic elements, as reported by other relative studies [12, 35, 36].

$$\mathrm{GLI }= {{\text{Log}}}_{2}[\frac{{{\text{C}}}_{{\text{i}}}}{\mathrm{GBV }\times 1.5}]$$
(1)

Where Ci = measured elements (mg kg−1), GBV = geóchemical backgróund vąlue of the element, 1.5 = control vąlues attributed to lithogenic variątion in the soil. The evaluation parameters of the geochemical analysis are presented in Table S1 (in the supplementary information).

2.6 Ecological risk assessment

Ecological risk was used to evaluate the overall soil pollution the expression proposed by Hakanson, [37].

$${{\text{C}}}_{{\text{f}}}=\frac{{{\text{C}}}_{{\text{i}}}}{{{\text{C}}}_{{\text{o}}}}$$
(2)
$${{\text{E}}}_{{\text{r}}}^{{\text{i}}}={{\text{T}}}_{{\text{r}}}\frac{{{\text{C}}}_{{\text{i}}}}{{{\text{C}}}_{{\text{o}}}}$$
(3)
$${{\text{R}}}_{{\text{i}}}=\sum {{\text{E}}}_{{\text{r}}}^{{\text{i}}}$$
(4)

where Ci = measured elements, Co = background value of elements, Tr = toxic response factor, \({{\text{E}}}_{\mathrm{r }}^{{\text{i}}}\) = potential ecological risk factor, and Ri = summation of \({{\text{E}}}_{{\text{r}}}^{{\text{i}}}\). The Tr and Co values are shown in Table S2.

2.7 Health risk assessment

The PTEs measured in soil samples were used to evaluate health risk using the United States Environmental Protection Agency (USEPA) model and other previous studies [12, 31, 38]. Exposure through the three pathways was used.

$${{\text{ADD}}}_{{\text{inh}}}=\frac{\mathrm{C }\times \mathrm{ InhR }\times \mathrm{ EF }\times \mathrm{ ED }}{\mathrm{PEP }\times \mathrm{ BW}\times \mathrm{ AT}}$$
(5)
$${{\text{ADD}}}_{{\text{ing}}}=\frac{\mathrm{C }\times \mathrm{IngR }\times \mathrm{EF }\times \mathrm{ED }\times \mathrm{ CF}}{\mathrm{BW }\times \mathrm{ AT}}$$
(6)
$${{\text{ADD}}}_{{\text{derm}}}=\frac{\mathrm{C }\times \mathrm{ SL }\times \mathrm{ SA }\times \mathrm{ ABS }\times \mathrm{ EF }\times \mathrm{ ED }\times \mathrm{ CF}}{\mathrm{BW }\times \mathrm{ AT}}$$
(7)

The non-carcinogenic effect of PTEs using the hazard quotient (HQ) in soil samples using the expression below USEPA, [39].

$${\text{HQ}}=\frac{{\text{ADD}}}{{\text{RfD}}}$$
(8)

The hazard index (HI) represents the summation of HQ multiple routes [40]. When HQ or HI > 1, the greater the probability of non-carcinogenic adverse health effects to occur [39].

$${\text{HI}}=\sum_{{\text{n}}}^{{\text{i}}}{{\text{HQ}}}_{{\text{i}}}$$
(9)

Cancer risk (CR) was evaluated using the expression below [39].

$$\mathrm{CR }=\mathrm{ ADD }\times \mathrm{ CSF}$$
(10)

where ADD = average daily exposure dose of PTEs, RfD = reference dose and CSF = cancer slope factor. The evaluation parameters for the exposure assessment are listed in Table S3.

3 Results and discussion

3.1 Concentrations of PTE occurrence in soils

Table 1 presents the mean concentrations of PTEs in the soil surrounding the cement plant. The obtained data for each sampling point are presented in Table S4. The mean PTE concentrations were 398.44 ± 10.26, 0.15 ± 0.15, 1.89 ± 1.76, 0.07 ± 0.27, 0.001 ± 0.001, 0.39 ± 0.33, 0.64 ± 0.83, 219.73 ± 10.54, 0.08 ± 0.14, 6.23 ± 6.80, 0.11 ± 0.33, 0.26 ± 0.25, 0.68 ± 0.70, 0.01 ± 0.01, 0.12 ± 0.15, 0.03 ± 0.07, 1.63 ± 1.31, 1.97 ± 1.59, 0.44 ± 0.46 and 4.50 ± 4.18 mg·kg−1, for Al, As, Ba, Cd, Co, Cr, Cu, Fe, Hg, Mn, Mo, Ni, Pb, Sb, Sc, Se, Sr, Ti, V, and Zn, respectively. The PTE mean concentrations were in decline order of Al > Fe > Mn > Zn > Ti > Ba > Sr > Pb > Cu > V > Cr > Ni > As > Sc > Mo > Hg > Cd > Se > Sb > Co.

Table 1 Descriptive statistics of potentially toxic elements (mg kg−1) in soil samples
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The most abundant measured PTEs were Al and Fe, whereas the other PTEs showed a minor distribution. From the observed skewness data, the biggest asymmetry was for Hg and Cd, followed by Cu, Co, Se, Ni, Mn, As and Ba. However, compared with the background values of the world average for soil, the mean concentrations of the investigated PTEs were lower than the world average elemental values [41]. The PTE concentrations observed in this study are lower than those reported outside Nigeria [38, 42,43,44] and in Nigeria [26,27,28, 45], but incomparable with those at Karst, Brazil [46], Odajana, Nigeria [47], and Canakale-Ezine, Turkey [48] presented in (Table 2), which indicate that cement production activities had less influence in the study areas. Moreover, it has been reported that an estimation of adverse health effects associated with PTEs, based on concentrations is not sufficient and must be followed with other estimated indices [16].

Table 2 Comparison of mean concentrations (mg·kg−1) of potentially toxic elements in the soil samples and other studies
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3.2 Multivariate analysis

3.2.1 Correlation analysis

The results of CA for PTEs of the soil samples (p < 0.01 and p < 0.05) are presented in Table 3. The correlation shows that pair elements have a positive correlation (r > 0.5) with Cu-Al, Fe-As, Mn-Al, Mn-Ba, Ni–Al, Ni–Cd, Ni–Cr, Ni-Cu, Se-Mo, Sr-Al, Sr-Ba, Sr-Mn, Sr-Ni, Ti-Ba, V-Al, V-Ba, V-Mn, V-Sr, Zn-Cd, Zn-Cu, and Zn-Ni at p < 0.01 and Cd-Al, Cr-Cd, Cu-Al, Cu-Cr at p < 0.05. The strong correlation showed that some PTE pairs at (p < 0.01 and p < 0.05) signify their concurrent discharge from a similar source. The Sc shows non-interaction with other PTEs, this indicates Sc perhaps not originated from cement plant activities.

3.2.2 Principal components analysis

PCA was conducted to identify the source of PTE at the sampling points [12]. Table 4 presents the results of the PCA and four components demonstrating 64% cumulative total variance (TV) of the diverse PTEs in the samples. Al, Ba, Mn, Sr, Ti, and V with a TV of 20% correlated with the first component. This might be due to anthropogenic sources. Cd, Co, Cr, Cu, Ni, and Zn with a TV of 19% correlated with the second component. The combination of Cd, Ni, and Cu are known to be from crude oil sources and transportation [49]. This could be due to emissions of automobile exhaust and machines used. The third component correlated with As, Sn, and Mo with a TV of 13%, and As, Fe, Hg, Mo, Sb, and Se with a total variance of 13% (Fig. 2), which indicated that the third and fourth components were contributed through paedogenic and lithogenic origin. These observations concur with the correlation result that predicted Sc may not be from cement plant activities.

Table 3 Correlation analysis for the potentially toxic elements in the soil samples
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Fig. 2
figure 2

Rotated component loading plot of component analysis of potentially toxic element of soil samples

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3.2.3 Cluster analysis

Cluster analysis grouped PTEs that were likely from similar sources [41, 50]. The PTEs are grouped into two clusters in Fig. 3, where Mn, V, Al, Sr, Ba, and Ti are in one group, while Cd, Ni, Cr, Zn, Pb, Sb, Se, Mo, Co, Sc, and Hg are in the second group. The cluster analysis showed that Mn, V, Al, Sr, Ba, and Ti, which have a related source and are not inclined by Cd, Ni, Cr, Zn, Pb, Sb, Se, Mo, Co, Sc and Hg, but the presence of Zn, Cd, Cr, Pb, Ni is influenced by Sb, Se, Mo, Co, Sc, and Hg, signifying that they are from a similar source. These observations concur with the PCA result, which showed that Al, Ba, Mn, Sr, Ti, and V may be from the same sources. Thus, the PTEs in groups one and two of the cluster are probably from anthropogenic or paedogenic sources [31].

Fig. 3
figure 3

Dendrograms produced by hierarchical clustering for potentially toxic element of soil samples

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3.3 Geochemical load index

Table 5 presents the geochemical load index (GLI) of the soil samples. The results of GLI show that the samples ranged from unpolluted to moderate for the investigated PTEs. The GLI values of the PTEs were in the following trend of Cd > Hg > Sb > Se > Al > Zn > Mo > Pb > Cu > As > Sc > Mn > Sr > Cr > Ni > V > Ba > Ti > Co. The GLI result in this study was consistent with values reported in previous studies [38, 51], but lower than values reported by [37, 42, 44, 46].

Table 4 Principal component analysis of potentially toxic elements in the soil samples
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3.4 Ecological risk assessments

To quantify the ecological risk associated with PTEs, the integrated potential ecological risk factor (\({{\text{E}}}_{{\text{r}}}^{{\text{i}}}\)) is used [52, 53]. The results of the ecological risk assessment are presented in Table 5. The distribution of potential ecological risk (\({{\text{E}}}_{{\text{r}}}^{{\text{i}}}\)) decreases in the order Hg > Cd > Mo > Pb > As > Cu > Sb > V > Ni > Cr > Mn > Ba > Zn > Ti > Co. The values of \({{\text{E}}}_{{\text{r}}}^{{\text{i}}}\) of PTES were below 40, indicating low \({{\text{E}}}_{{\text{r}}}^{{\text{i}}}\) [7]. Consequently, the results cumulatively deduce that low ecological risk from the investigated elements. Similarly, Handan et al. reported values of \({{\text{E}}}_{{\text{r}}}^{{\text{i}}}\) of As, Cu, Zn Cr, and Ni < 40 in the sediment of coastal estuaries, in Turkey [54]. The results obtained in this study are comparable with values reported in a previous study that \({{\text{E}}}_{{\text{r}}}^{{\text{i}}}\) below 40 in soil at Rampal, Bangladesh [55]. However, the values of \({{\text{E}}}_{{\text{r}}}^{{\text{i}}}\) obtained in the present study were lower than those reported in a similar study for Pb, Cd and Pb from Bese, Nigeria [25], and in western Saudi Arabia [42], respectively. The overall ecological risk of the investigated PTEs was below 150, indicating low ecological risk. The results show that Hg contributed 49.54% to the overall ecological risk.

Table 5 Geochemical load index and potentially ecological risk indices of the soil samples
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3.5 Health risk assessment

The ADD values of PTEs in soil samples for non-carcinogenic and carcinogenic risk are presented in Tables S5 and S6, respectively. The ADD exposure pathway of the PTEs was greater for ingestion, followed by dermal and inhalation (Table S5). The trend in the ADD values of the investigated PTEs is comparable with similar previous studies [29, 35]. Conversely, the acceptable daily intake (ADI) represents the total amount of a chemical that can be consumed over a lifetime without causing a health risk [56] As reported elsewhere, an estimation of the daily exposure to the general population, including sensitive subpopulations that are thoughts to carry a low lifetime risk of negative consequences is called a “reference dose” (RfD) [56, 57].

The HQ values are in the subsequent trend Cd > As > Al > Fe Pb > Cr > Mn > Ba > Hg > Sb > Mo > Cu Zn > Ni > Se > Sr > Co, Zn > Cd > Cr > Hg > Mn > As > Pb > Ba > Fe > Sb > Cu > Mn > Mo > Co and Mn > Cr > Ba > Cd > As > Al > Hg > Pb > Co > Sb > Cu > Zn > Ni > Se for ingestion, dermal, and inhalation, respectively, for both children and adults. The estimated HQs are in descending order of ingestion > dermal > inhalation. The greater value ingestion may pose a substantial risk compared with the dermal and inhalation pathways. A higher value of the ingestion pathway has been reported through soil exposure compared with dermal and inhalation [58,59,60,61,62,63].

The estimated HI values for the investigated PTEs were (6.65 × 10–6 to 9.33 × 10–2) in children and (8.35 × 10–7 to 1.17 × 10–2) in adults through ingestion, (7.35 × 10–10 to 2.18 × 10–4) in children and (3.68 × 10–10 to 1.11 × 10–4) in adults through inhalation, and (2.75 × 10–8 to 3.30 × 10–2) and adult (5.83 × 10–9 to 7.00 × 10–3) children through the dermal pathway (Table 6). In general, the estimated values obtained were higher for children than for adults, implying the vulnerability of children to PTEs in the soil samples. However, the HI values obtained for ingestion, inhalation, and dermal pathways were less than one (< 1), which signifies that there is negligible or no non-carcinogenic adverse effect likely to occur. Correspondingly, the previous study by Jafari et al. [38] reported that HI < 1 in the soil around the Douroud cement factory, in Iran.

Table 6 Non-carcinogenic hazard quotient (HQ) and hazard index (HI) of potentially toxic elements in soil samples
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The estimated carcinogenic risk (CR) of PTEs in soil samples is presented in Table 7. The estimated all the investigated PTEs are in children (1.12 × 10–10 to 3.55 × 10–2) and adults (1.39 × 10–11 to 4.45 × 10–3) through ingestion, children (2.71 × 10–13 to 6.89 × 10–10) and adult (5.59 × 10–14 to 1.43 × 10–10) through inhalation, and children (2.55 × 10–7 to 7.27 × 10–5) and adult (2.25 × 10–7 to 2.89 × 10–5) through dermal pathway. The CR of Cd for the ingestion pathway exceeded the permissible limit of (1 × 10–6 to 1 × 10–4) signifying probable cancer development. Taiwo et al. [63] reported that Cd contributes 82–86% of the total carcinogenic effect. It has been reported that Cd causes persistent toxicity and acute adverse health effects on the kidney, liver, vascular, and immune systems [64]. The higher CR for ingestion obtained in this study is similar to trends reported in similar studies [38, 44, 48].

Table 7 Carcinogenic risk of potentially toxic elements in soil samples
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4 Conclusion

The assessment of PTEs in the surrounding soil of the cement plant was evaluated using the geochemical load index, ecological and health risk indices. The concentrations of the investigated PTES were lower than the background values. The relative abundance declined in the order of Al > Fe > Mn > Zn > Ti > Ba > Sr > Pb > Cu > V > Cr > Ni > As > Sc > Mo > Hg > Cd > Se > Sb > Co. The geochemical load index revealed that the investigated samples ranged from unpolluted to moderate with PTEs. Ecological risk categorized the investigated soil under the low ecological risk class, with Hg contributing 49.5% to the overall ecological risk. The health risks of ADDs and HQs revealed similar trends for ingestion, dermal, and inhalation pathways for both children and adults. The values of HI for ingestion, inhalation, and dermal pathways obtained are less than one (< 1), which signifies negligible or no non-carcinogenic adverse effects likely to occur. The CR for the ingestion exposure pathway exceeded the permissible limit (1 × 10–4) for both children and adults, and Cd was the major contributor to carcinogenic effects, which suggests possible cancer development in residents.