Mineral Composition and Elemental Oxide Changes in Heat-Affected Soils and the Implications on Heavy Metal Immobilization by Sewage Sludge

Abstract

This paper investigated how increased soil temperatures affect soil mineralogy and major and trace element oxide concentrations and the implications of these effects on the mobility of potentially toxic elements (PTEs) in heat-affected soils amended with sewage sludge. The aim was to determine the efficiency of sewage sludge as an immobilizer of PTEs in heat-affected PTE-contaminated soils. Soil samples were heated to 150 °C, 300 °C, 500 °C, and 750 °C and later amended with stabilized sewage sludge at different rates. The concentrations of arsenic (As), chromium (Cr), cobalt (Co), copper (Cu), lead (Pb), nickel (Ni), and zinc (Zn) in the different geochemical fractions of the soils were determined before heating, after heating, and after sewage sludge application. Increased soil temperatures affected the mineral assemblage and the concentrations of some major and trace element oxides and the degree of weathering of the soils. These changes were, however, insignificant. The segregation of PTEs into the different soil geochemical fractions before and after heating varied. High soil temperatures resulted in an increase in PTE concentrations in the non-residual fractions of the soil (F1, F2, and F3) with a consequent increase in their mobility. The application of sewage sludge to heated and unheated soils reduced PTE concentrations in the F1 and F2 fractions of both soils, whereas it increased PTE concentrations in the F3 and F4 fractions by up to 30% for As and Cu, 20% for Cd, 25% for Co, 60% for Cr and Ni, 50% for Pb, and 55% for Zn. Significant immobilization of the PTEs was observed in the heat-affected soils that received higher amount of sewage sludge. Fire events could increase the mobility of PTEs in soils, but sewage sludge could still effectively immobilize these PTEs, although it needs to be applied at higher application rates.

1. Introduction

Fire has been successfully used as a soil and vegetation management tool for decades in different parts of the world with benefits that include a reduction in bush encroachment, an increase in biodiversity, the elimination of plant pathogens, and the release of plant nutrients contained in soil organic matter (OM) [1,2,3]. When used for management, the duration and intensity of these fires are controlled, and so the negative effects on the environment, especially on vegetation and soil, are minimal. However, when fires occur accidentally, they tend to burn out of control and may persist for days and even months with various consequences on the environment. The effects of fires on the environment are closely related to the way they burn. Flaming fires, typically observed on trees, shrubs, and branches of trees, are characterized by high temperatures ranging from 1500 °C to 1800 °C, and they tend to spread rapidly. In contrast, smoldering fires, which occur near the soil surface, exhibit lower temperature ranges of 500–700 °C and can persist for extended periods often lasting for months [4,5]. Fuel for the fire plays a crucial role in determining the temperature to which soils could be raised, with coarse-textured soils having a higher tendency to attain high temperatures. However, because of their longer duration, smoldering fires tend to have more effects on soil properties than flaming fires.

Soils comprise minerals, organic matter, liquids, gases, chemical elements and compounds, and various fauna and flora. Through various interactions among these components, they contribute significantly to the sustainability of life and livelihoods. Interactions among soil components are influenced by soil properties and the prevailing environment in the soil, both of which are affected by soil temperature. The effect of temperature changes on soil properties, especially the reduction in soil OM content, has been reported by Certini [6] and Ngole-Jeme [7]. The occurrence of fire events alters the soil environment by introducing volatiles into soil pores and increasing soil temperatures, which in turn influence and recalibrate various chemical and biological properties and processes in the soil environment to different extents. Reports by Xiao et al. [8] show that an increase in soil temperatures is likely to affect the speciation of potentially toxic elements (PTEs) in the soil environment with consequences on their availability. This was shown for cadmium and copper where an increase was observed in their water-soluble fraction with an increase in temperature [9]. Although their studies focused on lower temperature increases of between 0.5 and 1 °C, Sebola and Ngole-Jeme [10] have also shown alteration of the speciation of PTEs at higher temperatures with an increase in PTE mobility observed. The changes in PTE mobility observed, however, are more because of the changes in soil properties caused by increased soil temperatures than the changes in the PTEs and associated compounds present in the soil. Though according to Tan et al. [11], the mineral assemblage of soil is not affected by fires below 500 °C, Zims et al. [12] and Yusiharni and Robert [13] have shown that soil mineralogical composition could be altered at soil temperatures as modest as 200–600 °C. Both primary and secondary soil minerals are affected by increased soil temperatures, with the temperature at which secondary minerals are affected being much lower compared to that at which the primary minerals are [14,15]. Studies on the effect of temperature on secondary soil minerals have been reported by Richardson [16], Ulery and Graham [17], Arocena and Opio [18], Frost et al. [19], Raynard-Callanan et al. [20], and Araya et al. [21]. The alteration of these soil minerals could affect several other soil properties with consequences on various soil processes and interactions. Heydari et al. [22], for example, indicated that as kaolin is destroyed during fire events, oxides and hydroxides of aluminum and silicon accumulate. These changes in oxide and hydroxide contents may mean that the degree of chemical alterations in the soils could change with an increase in soil temperatures. Further, both the minerals and major and trace element oxides present in the soil can bind some PTEs, presenting a major environmental challenge.

Potentially toxic elements have been and remain a major human and environmental health concern globally because of their negative human health effects and their persistence in the environment. Their presence in soil is caused by pedogenic processes which are most often natural but also by anthropogenic activities associated with industrialization, agriculture, waste disposal, and mining among others. The behavior and especially the mobility of these PTEs in the soil environment is influenced by many soil properties with soil pH, OM content, cation exchange capacity (CEC), and mineralogy being the most influential. These properties are themselves affected by an increase in soil temperature [10]. Some of these properties also affect the segregation of PTEs into different geochemical fractions. For example, the concentrations of PTEs bound to the reducible, oxidizable, and residual geochemical fractions of soil are determined by the content of major and trace element oxides, soil OM content, and mineralogy of the soil. Although these geochemical fractions of soil PTEs are less mobile than the exchangeable fraction, they could be remobilized in the environment due to changes in soil elemental oxide concentrations, OM content, and mineralogy.

Remediation and immobilization of PTEs in contaminated soil have been carried out using various methods including physical (soil replacement, isolation, importing, and electrokinetic remediation), biological (phytoremediation and bioremediation), and chemical (leaching, fixation/immobilization, encapsulation, and absorption/adsorption) methods [23]. Chemical remediation methods commonly used for soils contaminated with PTEs such as fixation and adsorption involve immobilizing the elements in the soil with chemicals or substances that may include clays, metallic oxides, biosolids/sewage sludge, animal manures, biochar, and lime, among others [24]. Although there is usually an initial decrease in PTE mobility in sludge-amended soil after sludge addition, long-term application of sewage sludge to soil has been reported to increase the concentrations of PTEs in the soil [25]. The use of sewage sludge as a PTE immobilizer in soils exploits the capacity of the PTEs in the soil to form immobile complexes with the OM contained in the sewage sludge. Additionally, the OM in sewage sludge can form organo-mineral complexes with the minerals contained in the soil. These complexes may increase the soil’s ability to bind the PTEs, thereby reducing their mobility in the environment, or they may mobilize them, causing their spread in the environment. Upon the degradation of the organic material, these complexes disintegrate with a consequent increase in the mobility of these elements. The breakdown of the organic matter bound to minerals and these PTEs is, however, very slow because of they are inaccessible to soil microbes. Sewage sludge application on soil to immobilize PTEs is increasingly being employed, especially on contaminated soils that have arable potential, because it is a good source of macronutrients such as nitrogen, potassium, and phosphorus [26,27], it is able to increase soil OM content, and it also improves soil tilth [28]. Environmental quality regulating bodies therefore endorse the practice of applying stabilized sewage sludge on soils due to its agronomic value as well as its ability to rehabilitate degraded soils like those in mining environments [26,29]. Sewage sludge application on soil not only improves the arable potential of the soil, but it is also a more environmentally friendly sludge disposal technique compared to sludge incineration or its disposal in landfills [27,30].

The efficiency of sewage sludge as a soil PTE immobilizer is, however, influenced by several factors among which include the concentration of the PTEs in the sludge, methods used for treating the sludge, the properties of the soil on which the sludge is to be applied, and the management techniques employed on the soil. Soil properties such as texture, pH, organic matter content, Al/Fe/Mn oxide content, and mineral assemblage also affect the interactions between the sewage sludge applied to soil and PTEs contained in the soil. There are many studies [12,13,14,15,16,17,18,19,20,21] that have reported on the effect of fires on soil properties including mineralogy, but investigations on the effect of fire-induced changes in soil temperature on changes in the mineral assemblage and the content of major and trace element oxides in soils are rare. It is not clear how these changes may affect the degree of weathering of the soils. Furthermore, studies that focus on how temperature-induced changes in soil properties affect the efficiency of organic materials as PTE immobilizers in soils are scarce. This study, therefore, aimed at (1) investigating how increased soil temperatures affect soil mineralogy, major and trace element oxide concentrations, and the implications on the degree of weathering of the soils and (2) determining the effect of sewage sludge application on the mobility of PTEs in heat-affected soils. The findings of this study are expected to shed more light on whether sewage sludge and other organic amendments could be utilized in the immobilization of PTEs in soils that have experienced fire events.

2. Research Methods

2.1. Description of Study Area

Soil samples were obtained from Carletonville, a mining city located some 79 km from Johannesburg in Gauteng Province of South Africa. Gold and uranium mining are ongoing in the area which falls under the Witwatersrand basin. Mining activities around Carletonville have resulted in several tailings dumps around the city. Studies on these tailings and surrounding soils [10] have revealed PTE contamination in the area. The city experiences about 668 mm of rainfall annually [31] and has maximum and minimum temperatures of 30 °C and <5 °C, respectively [32]. The topography of the area is undulating in nature with a geology that is characterized by the presence of dolomitic terrains and chert of the Malmani subgroup [31]. Carletonville soils are mostly dominated by gold conglomerates and chloritoid shale or black argillaceous quartzite. The soils are reddish and deep and are characterized by a sandy loam texture with relatively low electrical conductivity, high acidity, and OM content and CEC of 0.86% and 43.41 meq/100 g, respectively [10]. The vegetation in the area is grassland-type.

2.2. Sewage Sludge and Soil Sample Collection and Pre-Preparation

The soil samples used in this study were collected around the tailings dumps in the vicinity of Carletonville with a steel spade and a hand trowel. Each soil sample was split into five portions. According to the results of a study by Drooger [33], on average, fire events heat surface soils at the rate of about 43 °C per min, 3.7 °C/min at depths of 1 cm, and 0.23 °C/min at depths of 3 cm. Using the heating velocity of surface soils based on Drooger [33], the temperature of four of the soil portions was raised to temperatures of 150 °C, 300 °C, 500 °C, and 750 °C within a time of 3.5, 7.0, 11.6, and 17.4 min, respectively, using a muffle furnace. The samples were held at these temperatures for two (2), four (4), and six (6) hours. The fifth soil portion was not heated, and so it was used as a control for the temperature experiment.

Stabilized sewage sludge was collected from a wastewater treatment plant East of Johannesburg in South Africa. This plant receives no industrial wastewater, and it is therefore anticipated to have low levels of PTEs. Wastewater in the plant is treated using the activated sludge method with no chemicals added at any stage of the treatment process. The generated sewage sludge is stabilized using anaerobic digesters to reduce the pathogen load in the sludge, after which it is spread on sludge drying beds from where it is piled for disposal. Sewage sludge samples were collected randomly from the sludge drying beds into plastic sampling bags and transported to the laboratories at Unisa Science Campus in Florida, Roodepoort, Gauteng Province where they were air-dried, crushed, and used for further experiments.

2.3. Experimental Set-Up

The soil samples that had been heated were further split into three. The crushed sewage sludge was then mixed separately with two portions of the heated soil samples at two different soil/sludge volume percent (vol%) ratios, namely 80:20 and 60:40. The third portion served as a control. A schematic representation of the experimental setup is presented in Figure 1. All soil/sludge mixtures and their respective controls were kept under natural weather conditions in a protected environment at the UNISA College of Agriculture and Environmental Sciences Horticulture Center for six months. During this period, the samples were periodically moistened and homogenized to allow for equilibration between the sewage sludge and the soil [34].

2.4. Characterization of Soil Samples

2.4.1. Physicochemical, Mineralogical, and Trace and Major Element Oxide Determination

Basic properties of the soil such as the texture, pH, organic matter (OM) content, and CEC were determined using standard methods of soil analyses as described by van Reeuwijk [35]. X-ray powder diffractometry was used in the determination of the minerals contained in the soils as explained by Fitton [36]. In the determination of the minerals contained in the soil samples, each sample was ground to a particle size of between 10 and 15 µm and then scanned at 2θ angles of between 2 and 70° using a step size of 0.5 s. The peak relative height/area proportions and Reference Intensity Ratio (RIR) of each mineral was then used to determine their abundance in the sample. Mineral phases in the samples were identified using The Mineral Powder Diffraction File Data Book (MPDFDB 2001).

Major and trace element oxide analysis of the soil samples were carried out using methods described by Fitton [36]. Each soil sample was crushed and placed in a muffle furnace at 1000 °C for three hours to allow for the oxidation of Fe²⁺ and S as well as for the determination of Loss on Ignition (LOI). A total of nine grams of flux (consisting of 34% LiBO₂ and 66% Li₂B₄O₇) and one gram of heated sample were then fused at 1050 °C to form a stable glass bead which was then analyzed for the major oxides including water (H₂O) contained in each sample. An amphibolite reference material was used for the quality control of the major and trace element oxide data generated from the XRF analyses. A PANAlytical Axios sequential WDXRF spectrometer was used to determine the concentrations of the major and trace element oxide concentration in the samples. The instrumental setup was established according to the manufacturer’s instructions. Details of these methods are described by the Council for Geosciences [37] and Fitton [36].

2.4.2. Determination of PTE Concentrations in Different Soil Geochemical Fractions

The concentrations of As, Co, Cd, Cr, Cu, Pb, Ni, and Zn in the different geochemical fractions (exchangeable and carbonate bound, reducible, oxidizable, and residual fractions) of the soil samples, as well as samples of these soils which had been heated at different temperatures for different durations and amended with sewage sludge at different rates, were determined using the three-step Community Bureau of Reference (BCR) sequential extraction protocol which makes use of different chemical extractants to extract PTEs from the different geochemical fractions of the soils [38]. The PTEs are extracted sequentially from the most mobile to the least mobile fractions. In the extraction of PTEs in the water-soluble, exchangeable, and acid-soluble fraction, 40 mL of 0.11 mol/L CH₃COOH was added to 1 gm of a soil sample, and the mixture was shaken for 16 h at room temperature. The mixture was then centrifuged for 20 min at 3000 rotations per minute (rpm), and the supernatant liquid was decanted. Potentially toxic elements in this extract represented the portion of the PTEs held in the exchangeable and carbonate-bound fractions (F1 fraction) of the soils. PTEs contained in this geochemical fraction of soils usually have an anthropogenic origin, are weakly bonded to soil particles, and are also more labile and available in the soil environment than the other soil fractions [38]. The remaining soil sample residue was washed with ultra-pure water and again centrifuged, and the second supernatant was discarded.

The washed residue was extracted with 40 mL 0.5 mol/L HONH₂.HCl solution adjusted to pH 2 with concentrated HNO₃. The mixture was also processed as step 1, and the concentrations of PTEs in the decanted supernatant were determined. The PTEs present in this fraction of soils (F2 fraction) are referred to as the reducible fraction and are held by oxides of Fe, and Al in the soil. Changes in the major and trace element oxide contents in the soil may influence the amount of PTEs held in this fraction. The washed residue from step 2 was again digested this time with 10 mL 8.8 mol L⁻¹ H₂O₂ for 1 h at room temperature and for 1 h at 85 °C, followed by the addition of 10 mL 8.8 mol L⁻¹ H₂O₂ and heating in a water bath for 1 h at 85 °C. Then, 50 mL of 1 mol L⁻¹ NH₄COOCH₃ was added to the mixture, after which it was shaken for 16 h at room temperature. The mixture was again centrifuged and decanted, and the residue was rinsed as indicated in previous steps. The concentrations of PTEs in the decanted supernatant represented those bound to OM (F3 fraction) and are referred to as the oxidizable fraction of PTEs. Changes in OM content in the soil are likely to be reflected in the concentration of PTEs in this fraction of soils.

Aqua regia (3:1 HCl: HNO₃) was then used to digest the remaining residue on a hotplate at 130 °C until the solution was almost dry. The mixture was made up to volume with 2% HNO₃, and the concentrations of the PTEs in the solution were determined. This is the F4 fraction which represents PTEs bound to the silicate structure of the soil. This fraction of PTEs, also referred to as the residual fraction, is believed to be very immobile and relatively inert. A Shimadzu Inductively Coupled Plasma-Optical Emission Spectroscope (ICP-OES) calibrated with MERCK ICP multi-element standards was used to determine the concentrations of the PTEs in all the extracts from the different extraction stages. The concentrations of each PTE in the different geochemical fractions of each soil sample were summed, and the values were compared with values of the same elements obtained from XRF analyses for accuracy.

2.4.3. Determination of the Degree of Chemical Alteration in Soil Samples

The ratio of a group of mobile oxides to immobile oxides in soils is commonly used as an indication of the extent of chemical weathering that a geological material has undergone [39]. Different indices of weathering including the chemical index of alteration (CIA), chemical index of weathering (CIW), and plagioclase index of alteration (PIA) of the heated soil samples were determined to infer whether an increase in temperature affected the degree of weathering of the soils [40]. CIA assumes that the degradation of feldspar and the formation of clay minerals is the dominant process occurring during chemical weathering [41]. PIA, on the other hand, assesses the source of weathering and elemental redistribution in the geologic material. CIA values for the soils were determined according to Equation (1), whereas CIW and PIA values were determined according to Equation (2) and Equation (3), respectively. The CaO considered in the equations is that which is incorporated into the silicate structure of the soils and not the carbonate fraction of Ca [42].

$$\mathrm{C}\mathrm{I}\mathrm{A}={\displaystyle \frac{{\mathrm{A}\mathrm{l}}_2{\mathrm{O}}_3}{{\mathrm{A}\mathrm{l}}_2{\mathrm{O}}_3+\mathrm{C}\mathrm{a}\mathrm{O}+{\mathrm{N}\mathrm{a}}_2\mathrm{O}+{\mathrm{K}}_2\mathrm{O}}}\times 100$$

(1)

$$\mathrm{C}\mathrm{I}\mathrm{W}={\displaystyle \frac{{\mathrm{A}\mathrm{l}}_2{\mathrm{O}}_3}{{\mathrm{A}\mathrm{l}}_2{\mathrm{O}}_3+\mathrm{C}\mathrm{a}\mathrm{O}+{\mathrm{N}\mathrm{a}}_2\mathrm{O}}}\times 100$$

(2)

$$\mathrm{P}\mathrm{I}\mathrm{A}={\displaystyle \frac{{\mathrm{A}\mathrm{l}}_2{\mathrm{O}}_3-{\mathrm{K}}_2\mathrm{O}}{{\left(\right(\mathrm{A}\mathrm{l}}_2{\mathrm{O}}_3-{\mathrm{K}}_2\mathrm{O})+\mathrm{C}\mathrm{a}\mathrm{O}+{\mathrm{N}\mathrm{a}}_2\mathrm{O})}}\times 100$$

(3)

2.4.4. Determination of PTE Mobility in Sludge Amended Heated Soils

The differences in means of PTEs in the soil samples that had been amended with sewage sludge at different rates and those that did not receive sewage sludge were compared using ANOVA with Tukey’s test to determine whether sludge addition had any effect on the total metal concentrations in the samples. The potentially toxic element mobility factor (MF) in the soils was calculated as indicated in Equation (4) [10].

$$\mathrm{M}\mathrm{F}={\displaystyle \frac{\mathrm{F}1}{\mathrm{F}1+\mathrm{F}2+\mathrm{F}3+\mathrm{F}4}}\times 100$$

(4)

where F1, F2, F3, and F4 represent the exchangeable and carbonate-bound, iron and manganese oxyhydroxide-bound, organic matter, and sulfide-bound residual fractions of PTEs in the soils.

High values for MF imply high mobility of the PTEs in the soil environment. The concentrations of PTEs in the different geochemical fractions of the heat-affected soils amended with sewage sludge were also compared using ANOVA with Tukey’s test to determine how sewage sludge addition affected the partitioning of PTEs in these soils. To determine the effectiveness of sewage sludge in binding PTEs in heat-affected soils, the MF values of the PTEs in all soil samples, both heated and unheated as well as those amended with sewage sludge at different soil/sludge ratios, were compared.

3. Results and Discussion

3.1. Physicochemical Properties of the Soil Samples

The soil samples had sand, silt, and clay contents of 61.3, 35.8, and 2.9%, respectively, indicating that they were sandy loam in texture. With a mean pH of 5.62, the soils could be described as acidic with low amounts of OM content (0.86) and CEC (43.41 meq/100 g). The low pH values of the soil are not unexpected considering that the sites from where the soil samples were collected are surrounded by gold mine tailings which are usually associated with acid mine drainage (AMD) generation. AMD is acidic in nature, and contamination of soil by AMD could hinder the growth of vegetation and survival of organisms in the soil, resulting in low vegetation cover and a consequent low soil OM content. Similar outcomes have been reported in other mining areas by Ngole-Jeme and Fantke [43]. Soil CEC is influenced by soil texture, OM contents, and mineralogy. Considering that the soils have low OM content and are dominated by sand and silt particles which Schulten and Leinweber [44] have reported to have low CEC values, the low CEC values of the soil are not unexpected.

3.2. Effect of Soil Temperature Increase on Soil Mineralogy

The mineral assemblage of both the heated and unheated soil samples comprised quartz, hematite kaolinite, mica, K-feldspar, plagioclase, chlorite, and pyrophyllite (

Table 1. Relative abundance of minerals identified in the soil samples.

Sample	Soil Temperature	Minerals Identified in Samples
		Hematite/Goethite	K-Feldspar	Plagioclase	Quartz	Mica	Kaolinite	Chlorite	Pyrophyllite
Sample 1	0	1	-	-	97	-	2	-	-
	150	1	1	-	96	-	2	-	-
	300	1	-	-	96	-	3	-	-
	500	2	-	-	98	-	-	-	-
	750	2	1	-	97	-	-	-	-
Sample 2	0	1	-	-	97	-	2	-	-
	150	1	1	1	95	-	2	-	-
	300	1	-	-	97	-	2	-	-
	500	1	-	-	99	-	0	-	-
	750	1	-	-	99	-	-	-	-
Sample 3	0	1	-	1	87	2	4	4	1
	150	1	1	-	88	2	3	4	1
	300	1	-	1	90	2	2	2	2
	500	1	-	-	93	2		2	2
	750	1	-	-	98	1	-	-	-
Sample 4	0	1	-	-	96	1	2	-	-
	150	2	1	-	96	-	2	-	-
	300	2	-	-	96	-	2	-	1
	500	1	-	-	99	-	-	-	-
	750	2	-	-	98	-	-	-	-

) and is also reflective of their textural properties. The dominant mineral in sand and silt particles is quartz, and so the high percentage of these particles (98.8%) in the samples justifies the dominance of quartz in their mineral assemblage. Kaolinite, the second most abundant mineral in the samples, is a secondary mineral that is commonly encountered in the clay fraction of highly weathered soils. With the mean clay fraction of the sample comprising only 2.9% of the weight percent of sand, silt, and clay, the low amount of kaolinite in the samples is not unexpected. Soil samples 1 and 2 contained the same minerals (quartz, hematite, K-feldspar, and kaolinite), and they both showed the same pattern of changes in mineral abundance with an increase in soil temperature and duration of heating (Table 1). Typical diffractograms of the soil samples heated at different temperatures are presented in Figure 2.

Heating these soil samples to 150 °C and holding the soil temperature at this level for up to 4 h had no effect on the mineral assemblage, but on increasing the duration of heating to 6 h, there were slight increases observed in the abundance of quartz, a disappearance of K-feldspar, a reduction in kaolinite abundance, and no change in hematite content, as shown in Table 1. At temperatures of 300 °C, kaolinite content in the samples remained the same as was the case at 150 °C, but the content of quartz increased, K-feldspar was not detected, and hematite content fluctuated. At 500 °C and 750 °C, quartz and hematite contents increased with an increase in the duration of heating, whereas kaolinite was no longer detected at these temperatures (Table 1). Soil samples 3 and 4 had quartz, kaolinite hematite, mica, chlorite pyrophyllite, and K-feldspar constituting their mineral assemblages (Table 1), though these samples had lower amounts of quartz than samples 1 and 2. The abundance of quartz and hematite in the samples also increased with an increase in soil temperature and duration of heating, as was the case with the other two soil samples, but the abundance of kaolinite, chlorite, and mica showed an opposite trend, decreasing with an increase in soil temperature and duration of heating (Table 1).

Despite all these changes, only differences in the abundance of k-feldspar in soils heated at temperatures of between 150 and 300 °C (p = 0.037) and between 150 and 500 °C (p = 0.037); quartz in soil samples heated at 150 and 750 °C (p = 0.009); and kaolinite in soils heated at temperatures of 150 and 500 °C (0.005), 150 and 750 °C (p = 0.000), and 300 and 500 °C (p = 0.005) were significant. The duration of heating had no effect on the changes in mineral abundance (p > 0.05), but the sample type had a significant effect (p < 0.05).

3.3. Effect of Soil Temperature Increase on Major Element Oxides

The concentration patterns of major element oxides in the soils varied with concentrations following the order SiO₂ > Al₂O₃ > Fe₂O₃ > TiO₂ > K₂O > MnO > P₂O₅ > Cr2O3 > CaO > Na₂O > MgO for samples 1 and 4, SiO₂ > Al₂O₃ > Fe₂O₃ > TiO₂ > MnO > K₂O > P₂O₅ > CaO >Cr₂O₃ > Na₂O > MgO for sample 2, and SiO₂ > Al₂O₃ >Fe₂O₃ > K₂O >TiO₂ > MgO > MnO > CaO > Na₂O >P₂O₅ > Cr₂O₃ for sample 3 (

Table 2. Major and trace element oxides in samples with increase in soil temperature.

Sample	Soil Temperature (℃)	Major and Trace Element Oxide Concentrations (%)
		SiO2	TiO2	Al2O3	Fe2O3(t)	MnO	MgO	CaO	Na2O	K2O	P2O5	Cr2O3	LOI.	Total	H2O−
Sample 1	0	82.89	0.51	6.05	5.90	0.173	<0.01	0.05	0.03	0.24	0.069	0.055	4.15	100.08	0.95
	150	84.24	0.51	6.27	5.65	0.175	<0.01	0.05	0.02	0.24	0.071	0.050	2.82	100.06	0.63
	300	83.74	0.52	6.40	5.90	0.189	<0.01	0.05	<0.01	0.25	0.071	0.054	2.72	99.84	0.71
	500	84.23	0.52	6.26	5.87	0.207	<0.01	0.04	0.04	0.25	0.074	0.050	2.59	100.11	0.72
	750	85.48	0.55	6.72	6.23	0.211	<0.02	0.05	0.24	0.28	0.074	0.054	0.21	100.07	0.34
Sample 2	0	88.06	0.36	4.43	3.11	0.264	0.03	0.06	<0.01	0.22	0.054	0.021	3.42	100.04	0.80
	150	89.50	0.36	4.33	3.00	0.306	0.05	0.07	0.10	0.23	0.058	0.019	2.00	100.02	0.48
	300	89.39	0.36	4.49	3.03	0.283	0.04	0.07	<0.01	0.22	0.055	0.021	1.99	99.95	0.53
	500	88.79	0.36	4.37	2.95	0.280	0.04	0.06	0.04	0.21	0.057	0.021	1.84	99.01	0.57
	750	90.74	0.38	4.63	3.20	0.288	0.04	0.07	0.32	0.24	0.059	0.021	0.18	100.01	0.28
Sample 3	0	83.61	0.44	7.02	4.28	0.208	0.25	0.11	0.05	0.58	0.059	0.028	3.46	100.09	0.83
	150	84.71	0.43	6.77	4.16	0.178	0.24	0.11	0.05	0.57	0.057	0.028	2.42	99.72	0.62
	300	84.80	0.44	6.98	4.16	0.209	0.25	0.11	0.06	0.57	0.059	0.028	2.33	100.01	0.63
	500	86.59	0.45	7.19	4.20	0.212	0.24	0.11	0.07	0.59	0.062	0.027	0.39	100.13	0.36
	750	86.72	0.45	7.01	4.25	0.214	0.27	0.12	0.08	0.60	0.062	0.029	0.33	100.13	0.40
Sample 4	0	84.77	0.41	5.36	5.82	0.075	<0.01	0.04	<0.01	0.23	0.079	0.061	3.04	99.87	0.76
	150	84.82	0.42	5.40	6.42	0.078	<0.01	0.04	0.09	0.25	0.081	0.069	2.04	99.68	0.70
	300	85.47	0.42	5.28	6.33	0.083	<0.01	0.04	<0.01	0.24	0.084	0.070	2.06	100.06	0.62
	500	85.69	0.42	5.35	6.16	0.075	<0.01	0.04	<0.01	0.23	0.079	0.073	1.96	100.06	0.63
	750	85.64	0.44	5.95	7.02	0.084	<0.02	0.04	<0.02	0.26	0.089	0.082	0.26	99.83	0.35

). These patterns showed that the oxides of Si, Al, and Fe were the most abundant in the soils. This is not unusual given the mineralogical compositions of the samples which are dominated by silica, kaolinite, and hematite with chemical formulae of SiO₂, Al₂Si₂O₅(OH)₄, and Fe₂O₃, respectively. These chemical formulae confirm the presence of Si, Fe, and Al in the soils. Higher concentrations of MgO and K₂O were observed in sample 3 (Table 2), which can be explained by the presence of a magnesium-bearing mica (KMg₃(AlSi₃O₁₀)(OH)₂) in sample 3 (Table 1).

Except for TiO₂, Cr₂O₃, CaO, and MgO, the concentrations of the major oxides of the samples increased with an increase in soil temperature, whereas values for Loss on Ignition (LOI) and water content in the samples decreased. The duration of heating had no effect. ANOVA analyses, however, showed that only differences in Na₂O content (especially at 750 °C), LOI, and water contents in the samples were significant. The absence of any significant changes in the concentrations of the major and trace element oxides can be explained by the fact that they have high threshold temperatures which were not reached in this study, and which are not likely to be attained during fire events. It is therefore not likely that they would be lost from the soils at the temperatures to which they were heated.

3.4. Effect of Increased Soil Temperature on the Degree of Weathering of the Soils

The values for CIA, CIW, and PIA of the soil samples were in the ranges of 90.46–95.05, 97.99–98.69, and 97.58–99.03, respectively (Figure 3), which indicates that the soils were generally highly weathered, with the degree of weathering following the order sample 4 > sample 1 > sample 3 > sample 2 (Figure 3). The values for CIA and PIA showed a general decrease with an increase in soil temperature for sample 2, with the other samples showing minimal changes. However, these changes were insignificant (p > 0.05), indicating that an increase in soil temperature had no major effect on the degree of weathering of the samples. This is like what was observed in the concentration of the major and trace element oxides.

3.5. Concentrations of PTEs in the Soil Samples

Considering the minimal differences in the mineralogy and elemental concentrations in the samples, only two of the soil samples were evaluated for the effect of sludge application on the mobility of PTEs in heat-affected soils. The percentage recovery of the PTEs from analyses of the reference sample ranged from 77.5% for Co to 102.2% for Ni, reflecting a high level of reliability of the method used and results obtained. The concentrations of the PTEs in the sludge used in this study were all lower than what was contained in the soils as shown in

Table 3. Concentrations of PTEs in the soils, sludge, and sludge-amended unheated soils.

Potentially Toxic Element	PTE Concentrations (mg/kg)
	Sludge	100:0 Soil/Sludge Ratio		80:20 Soil/Sludge Ratio		60:40 Soil/Sludge Ratio
		Sample 1	Sample 2	Sample 1	Sample 2	Sample 1	Sample 2
As	6.8	15	11.3	18.8	10.9	12.6	9.3
Cd	4.1	7.1	6.6	9.3	5.2	6.6	5.9
Co	48.9	67	51	71.2	58.4	52.3	45.3
Cr	102.3	185	105.3	172.6	102.5	147.1	107.7
Cu	157.9	188.3	176	203.7	184.8	152.4	161.5
Ni	75.6	87.9	72	91.1	77.6	65.5	63.6
Pb	87.2	96.3	103	98.2	104.4	86.7	91.9
Zn	175.1	187	189	186.8	186.9	182.9	173.5

. This is not unexpected considering that the treatment plant where the sludge came from received mainly municipal wastewater from a small community. The addition of this sludge at soil/sludge ratios of 80:20 volume percent (vol%) to soil samples that had not been heated resulted in slight increases in the concentrations of As, Cd, and Pb in the soils, with significant increases in Co, Cu, Ni, and Zn concentrations, and a decrease in Cr concentrations (Table 3). At a soil/sludge vol% ratio of 60:40, a decrease in metal concentration was observed, possibly because the concentrations of PTEs were higher in the soil sample than in the sludge. This is not uncommon as soils around mines generally have high PTE concentrations because of the high PTE concentrations in mine tailings [43]. The pattern of PTE concentrations observed in this study is therefore not unusual.

3.6. Segregation of PTEs into Different Geochemical Fractions in the Soils

In the soils that had not been amended with sewage sludge, the segregation of PTEs in the different soil geochemical fractions followed the order F4 > F3 > F2 = F1 for Cr, Co, and Ni, whereas for Cd and Pb, the pattern was F1 = F2 > F4 = F3 (Figure 4). Arsenic, Cu, and Zn showed a similar pattern of F4 > F3 = F2 = F1 as shown in Figure 4. These results showed that except for Cd and Pb, the PTEs all had high concentrations in the residual fraction.

The application of sewage sludge to both soil samples resulted in a decrease in the concentrations of PTEs in the F1 and F2 fractions, whereas PTE concentrations in the F3 fraction increased, as shown in

Table 4. Changes in PTE concentrations in the different soil geochemical fractions 6 months after sewage sludge application.

Potentially Toxic Element	Differences in PTE Concentrations (mg/kg)
	80:20 Soil/Sludge
	Soil Sample 1				Soil Sample 2
	F1	F2	F3	F4	F1	F2	F3	F4
As	−0.69	−0.34	0.90	2.31	−0.09	−0.08	2.17	2.02
Cd	−1.60	−1.62	1.19	−0.80	−0.79	−0.78	0.61	−0.31
Co	−0.54	−0.63	3.12	2.17	−0.30	−0.30	1.55	1.60
Cr	−9.12	−9.35	5.78	7.88	−8.96	−9.14	5.51	7.74
Cu	−10.82	−7.62	7.05	4.58	−16.50	−6.10	6.31	2.68
Ni	−2.44	−2.49	8.08	−4.75	−1.37	−1.41	4.17	−6.81
Pb	−3.90	−3.32	1.65	−2.49	−2.06	−3.52	1.52	−2.13
Zn	−2.17	−2.19	2.75	3.58	−1.04	−1.04	1.57	3.87
Potentially Toxic Element	60:40 Soil/Sludge
	Soil Sample 1				Soil Sample 2
	F1	F2	F3	F4	F1	F2	F3	F4
As	−2.68	−0.34	0.90	−2.55	−1.28	−1.58	2.31	−2.18
Cd	−3.50	−1.51	1.15	−0.92	−1.88	−0.87	0.65	−0.25
Co	−2.54	−0.63	3.12	−3.17	−2.30	−0.30	1.55	−1.60
Cr	−10.15	−9.37	6.21	−6.45	−9.23	−9.42	6.01	−8.84
Cu	−31.71	−28.12	8.51	−6.92	−17.13	−16.61	6.31	−8.62
Ni	−2.44	−2.49	8.08	−8.75	−1.37	−1.41	4.17	−6.81
Pb	−3.82	−2.72	1.66	−2.96	−2.05	−3.79	1.60	−1.88
Zn	−2.15	−2.17	2.74	2.02	−1.04	−1.05	1.60	2.90

. The most significant decreases in PTE concentrations were observed in the F1 followed by the F4 and then the F2 fractions for soil sample 1 and the F4, F1, and then F2 fractions in soil sample 2 (Table 4). In the soils that had been amended at a sludge application rate of 80:20 vol%, Cu, Cr, and Ni showed the greatest decrease in concentrations in the F1, F2, and F4 fractions, respectively, whereas the highest increase in concentration in the F3 fraction was displayed by Ni. This pattern was slightly altered in both soils when higher sludge amounts (60:40 vol%) were applied, with the greatest decreases in metal concentrations in the F1, F2, and F4 fractions shown by Cu, Cu, and Ni, respectively (Table 4).

The percentage changes in PTE partitioning observed in the soils after sludge addition were significantly higher for soil sample 1 than for soil sample 2 (p = 0.023), and when sludge was applied at a higher application rate (60:40 vol%) than when applied at lower application rate (80:20 vol%) (Table 4). Sample 1 initially had higher PTE concentrations than sample 2 (Table 3), so the addition of sludge with a lower concentration of PTEs would have diluted the PTE concentrations in Sample 1, resulting in a higher percentage change with sludge addition. The redistribution of PTEs into different geochemical fractions of the soil after sludge addition further highlights the role of OM in the immobilization of PTEs in soil. According to Zhang et al. [45], once OM is introduced into soil, it forms complexes that are rich in carboxyl, hydroxyl, and amine functional groups, which have high adsorption capacities for PTEs. The sewage sludge used in this study came from a municipal wastewater treatment plant receiving mainly domestic wastewater which is generally rich in OM [46] because most of the wastewater comprises sewage. The introduction of this sludge to the soil would have increased the OM contents, allowing for the formation of complexes between the sewage-sludge-derived OM and the soil particles, with a consequent increase in the adsorption sites in the soil. More sites were therefore available for the adsorption of PTEs in the sludge-amended soils, resulting in an increase in the PTE concentrations in the F3 fraction, which is the fraction of the PTEs bound to OM, possibly at the expense of the PTEs in the F1 fraction (Table 4) where decreases were observed.

The formation of OM-PTE complexes, however, varies with specific PTEs [47], the functional group present in the OM [48], and the level of humification of the OM [45]. Cadmium, Cu, and Ni, for example, generally have strong affinities for OM in soils as they are able to form covalent bonds with organic ligands in the soil, keeping them bound to the organic fraction (F3) of the soil [49,50]. The role of the functional group in the formation of these complexes is displayed by Cu. According to Li et al. [47], Cu has a strong affinity for the acid functional groups contained in the humic acids in OM, and so OM rich in humic acids will immobilize Cu in soils. The role of OM humification in PTE binding is also exemplified by Cd, where highly humified OM would favor OM complexation with Cd [47]. These patterns of PTE interaction with organic matter may explain the increase in PTE concentrations in the F3 fraction of the soils after sludge application.

The reduction in PTE concentrations observed in the F1, F2, and F4 fractions may also be linked to the interaction between the exogenous OM introduced into the soils through sludge application and the oxides of Fe and Al present in the soil. According to Kleber et al. [51], isomorphous substitution of metal cations, nucleation, and crystallization inhibition interactions between Fe/Al/oxides and hydroxides with OM in the soil could result in differential adsorption of PTEs. This has been shown to occur for Cu and Zn by Lippold et al. [52]. The blocking of exchange sites on the oxides of Fe and Al and mineral surfaces in soils due to organic matter coating could be another factor affecting the pattern of PTE concentrations in the F2 and F4 fractions, respectively, of the soils after sludge addition. According to Akbarpour et al. [53], organic matter may form a coating on the Fe-Mn oxides (F2 fraction) in soils with high OM, reducing the binding of PTEs to this fraction. A similar observation has also been made for soil minerals by Gao et al. [54], who observed the formation of organo-mineral complexes between soil minerals and organic matter with a consequent reduction in the adsorption sites on the minerals (F4 fraction), and a decrease in their ability to bind PTEs in the soil environment. The decrease in the concentration of some PTEs in the F2 and F4 fractions of the soil observed with sludge application is therefore also explained.

Several studies have looked at the long-term effect of sewage-sludge applications on metal mobility in soils. Most of the observations indicate an initial decrease in metal mobility which may be associated with the high OM content of the sludge. As the sludge decomposes, however, the fraction of the sorption sites contributed by the sludge OM disappears, reducing the sorption capacity of the soil and liberating the PTEs that were initially bound to the OM. This may eventually result in an increase in the mobility of PTEs in sewage-sludge-amended soils.

3.7. Effect of Soil Temperature Increase on PTE Concentrations in Soils

For both soil samples, an increase in soil temperature and duration of heating caused mostly decreases in PTE concentrations, with the greatest decreases observed in As and Cr and the least in Cd and Cu. These differences were, however, insignificant (p > 0.05) except for As. The lack of difference in PTE concentration with an increase in soil temperature can be justified by the fact that the temperatures to which the soils were heated were below the volatilization temperatures (616 °C, 767 °C, 2900 °C, 2670 °C, 2595 °C, 2730 °C, 1750 °C, 906 °C, for As, Cd, Co, Cr, Cu, Ni, Pb, Zn, respectively) of the elements. The pattern of PTE concentration with an increase in soil temperature observed in this study is, however, contrary to what has been widely reported in fire-affected soils by Stankov et al., [55], Costa et al. [56], Campos et al. [57], and Abraham et al. [58], who all reported increases in soil PTE concentrations after fire events. They attributed the increases to the liberation of PTEs contained in organic matter and vegetation during combustion, as well as the interaction between soil and the ash produced from the combustion of organic materials in the soil. This study was conducted in the laboratory where little or no vegetation was present in the soil samples. The amount of ash formed during the combustion of OM was therefore also minimal. There was no additional source of PTEs to the soil. The observed decreases in PTE concentrations with an increase in soil temperature could therefore be due to loss of some soil particles during analyses. Studies which focus on the amount of PTEs contained in ash produced from vegetation during fires are, however, needed to ascertain the contribution of vegetation ash to PTE contents in fire-affected soils.

3.8. Effect of Sludge Addition on PTE Redistribution into Different Geochemical Fractions of Heat-Affected Soils

The application of sewage sludge to soils that had been subjected to high temperatures also had an impact on the segregation of PTEs into the different soil geochemical fractions as shown in Figure 5. At a sludge application rate of 80:20 soil/sludge vol%, the concentrations of As, Cd, Cu, and Pb in the F1 and F2 fractions of soil sample 1 decreased (p = 0.003), whereas those of Co, Cr, Ni, and Zn showed no significant change (p = 0.08). An increase in concentration was observed in the F3 fraction of all PTEs (Figure 4). In soil sample 2, sludge application at a rate of 80:20 vol% showed varied patterns, but the concentrations of all PTEs in the F3 and F4 patterns also increased, as shown in Figure 5.

Applying sludge at a higher rate of 60:40 soil/sludge vol% resulted in decreases in PTE concentrations in the F1 and F2 fractions, especially for As, Co, Cu, Ni, and Pb, with increases observed in the F3 fractions (Figure 5). This increase is probably at the expense of the PTEs which were found in soil solution (F1 fraction) as a decrease in metal concentration was recorded in this fraction after sludge addition. A similar pattern was observed in the soils that had not been heated. These patterns are similar to what Awad et al. [59] obtained when OM-rich biochar was added to soils. They attributed the increase to the increased adsorption sites contributed by the biochar to the soil. In addition, studies by Malinowska [60] and Malinowska and Jankowski [61] have shown that under alkaline conditions, PTEs tend to be sorbed to organic matter. Zhang et al. [62] have attributed this to the fact that OM contains functional groups like HO-, COO-, and C=O on its surface, which enhance their complexation to PTEs as temperature increases due to increasing electronegativity. Considering that the pH of these soils increased with an increase in soil temperature [10], the complexation of PTEs and OM contained in the applied sludge would have contributed to the increased concentrations of the PTE in the F3 fraction as was also observed in the soils which had not been heated.

Some of the patterns of PTE partitioning observed among the different soil geochemical fractions of the heat-affected soils in this study differ from what has been reported in other studies that have been carried out on sewage-sludge-amended soils. For example, studies by Cao and Ma [63] showed that As segregates preferentially into the oxidizable fraction (F3), which was not the case in this study as As concentrations in the different fractions of both soil samples amended with sewage sludge followed the order F4 > F3 > F1 > F2. The greatest changes in As concentrations in the heat-affected soils after sludge addition at 80:20 vol% were observed in the F1 fraction of soil sample 1 that had been heated at 150 °C and soil sample 2 that had been heated at 500 °C and 750 °C (Figure 5).

In heated soils amended with sewage sludge at the rate of 60:40 vol%, the greatest changes were observed in the F1 and F3 fractions for sample 1 and sample 2, respectively, that had been heated at 750 °C and 300 °C (Figure 5). The sludge application rate affected the concentrations of As in the different geochemical fractions, but the patterns of As concentrations in the samples with respect to the different geochemical fractions remained the same as was observed in sludge-amended soil that had not been heated.

Cadmium concentrations in the soils amended with sludge followed the order F1 = F2 > F3 > F4, indicating that most of the Cd contained in the soils was found in the F4 fraction (Figure 5). Sánchez-Martín et al. [64] have indicated that Cd mainly segregates into the F2 and F3 fractions at lower sludge application rates, but with an increased rate of sludge addition, segregation into the F4 and F1 fractions mainly occurs, with a decrease in the F3 fraction. This is different from what was observed in this study where Cd preferentially segregated into the F1 and F2 fractions at both high and low rates of sludge application. Cadmium concentrations in the different geochemical fractions of the sludge-amended soils showed a significant increase with an increase in temperature, with the most significant increase observed in the F3 fractions of soil samples that had been heated at 150, 300, and 750 °C (Figure 5). Although the concentrations in the different geochemical fractions of the soil changed, the pattern remained like what was observed in the soil that had been amended with sludge without being heated.

The pattern of Cr segregation into the geochemical fractions of the sludge-amended heat-affected soils (F4 > F3 > F2 = F1) was comparable with what was reported by Sánchez-Martín et al. [64] and Malinowska [60] where an increase in Cr concentration in the F3 and F4 fractions was reported with sludge application. The increase in Cr in these fractions could be due to the formation of complexes with OM and binding onto aluminosilicates which dominate the F3 and F4 fractions, respectively. The chromium concentration in the F3 fraction increased in sludge-amended soils that had been heated between 150 and 500 °C but reduced in soils that had been heated at 750 °C. An opposite pattern was observed in the F4 fraction (Figure 5). The pattern of concentration of Cr in the different geochemical fractions of the heated soils after sludge addition was slightly modified compared to the unheated soil which had been amended with sludge.

The addition of sewage sludge to soils results in the preferential segregation of Co into the F3 and F4 fractions with an increase in sludge application rates according to studies carried out by Malinowska and Janowski [61]. This is like what was observed in this study where Co concentrations in the different patterns followed the order F4 > F3 > F2 > F1 in soils amended with sewage sludge at both sludge application rates. The changes in Co concentration observed in the different fractions of the soil with sludge application became more obvious as the temperature of the soil and sludge application rate increased for both soil samples (Figure 5). No clear patterns were observed in changes in Co concentration in the different fractions with sludge addition, which may indicate that Co has a low affinity for soil organic matter and would preferentially sorb to the silicates that are dominant in the F4 fraction of soils. This is seen in the reduction in Co concentration in the F4 fraction with sludge addition (Figure 5). The addition of sludge to the soils would have reduced the soil particles that are rich in soil minerals and increased the organic matter fraction, resulting in a reduction in the binding sites in the F4 fraction and a consequent reduction in Co concentration in the F4 fraction with an increase in sludge addition as shown in Figure 5. Despite these changes, the differences in metal concentrations in the heated and unheated soils amended with sewage sludge were insignificant (p = 0.087).

The concentrations of Cu in the different geochemical fractions of the heated soils after sludge addition followed the order F4 + F3 > F2 > F3 for soils that received sewage sludge at the rate of 60:40 vol%, whereas for the soils that received a lower amount of sewage sludge (80:20 vol%), the pattern of Cu concentration was F4 > F3 > F1 > F2. This is different from the soil that was not heated, as shown in Figure 5. Increases in Cu concentration were observed in the F3 fraction, although not to the levels where Cu concentration was in the soils that had not been heated, whereas decreases were observed in the F1 and F2 fractions of both soils that had received sewage sludge (Figure 5). These increases were more pronounced in soils that had been heated to 300 °C (Figure 5). An increase in Cu concentration in the F3 fraction with an increase in sludge application rate was also observed by Chaudhuri et al. [65], Ngole [66], and Rosen and Chen [67], and this was attributed to the formation of complexes between Cu and OM contained in the sludge-amended soils. The complexation of Cu with organic matter has also been reported by Lippold et al. [52] and Li et al. [47], among others. Copper has a strong affinity for OM, which was destroyed in the heated soils. The addition of sewage sludge to the heated soils where OM had been destroyed at temperatures of 300 °C could have increased the adsorption sites on the soil for Cu but not to the levels of the unheated soils which showed the highest percentage increase in Cu concentration in the F3 fraction. This may explain the lower levels of Cu in the F3 heated soils relative to the unheated soils (Figure 5). Heating the soils therefore affected the effectiveness of OM in binding the Cu contained in the soils.

Nickel concentrations in the F3 and F4 fractions of soil sample 1 increased with an increase in sludge application rate. The pattern of Ni segregation observed in the sludge-amended heated soil (F4 > F3 > F2 > F1) is comparable with what Chaudhuri et al. [65], Shrivasta and Bernerjee [68], and Sánchez-Martín et al. [64] reported. Nickel forms covalent bonds with organic ligands [50,69], which may explain the high Ni concentration found in the F3 fraction with an increase in sludge application rate. The changes in metal segregation into the different fractions with sludge addition were more obvious in the F4 fractions of both soil samples and at both sludge application rates, and in the F3 fraction of both soils that had received a higher amount of sludge (Figure 5). The extent of the changes observed also increased with an increase in soil temperature and were higher for sample 2 relative to soil sample 1 (Figure 5).

Lead concentrations in the different geochemical fractions of the heat-affected sludge-amended soils followed the order F1 > F2 > F3 > F4 for both soil samples and both sludge application rates (Figure 5), which is contrary to what Ngole [66] and Illera et al. [70] reported. With the addition of sludge at rates of 80:20 and 60:40 vol%, Pb content in the F1 fraction of soil sample 1 was reduced, especially in samples that had been heated at 150 °C (Figure 5). Despite the reduction, the F1 fraction was still the dominant fraction of Pb in both soil samples. The preferential segregation of Pb into the F1 fraction could be attributed to the formation of lead carbonate [70], which was possibly facilitated by the increased soil pH after temperature increase [10].

Sludge addition to the contaminated soil caused significant changes in the concentrations of Zn in the F1, F3, and F4 fractions of both soils, with the temperature of the soil having a minimal effect on the degree of change in Zn concentrations in the different geochemical fractions (Figure 5). Soil sample 2 showed a slightly different pattern, with an increase in temperature having the greatest effect in the F1 fraction (Figure 5). Zinc in the heat-affected sludge-amended soils preferentially sorbed into the F4 fraction followed by the F3 fraction, then the F2, and finally the F1 fraction. Sánchez-Martín et al. [64] and Mendoza et al. [71] have also reported the segregation of Zn into the F4 fraction after sludge addition. According to Parkpian et al. [50] and Vulkan et al. [72], a high percentage of the Zn present in both contaminated and natural soils is bound to silicate lattice, and so the high concentrations of Zn in the F4 fraction of these samples is not unusual. This binding is also favored by soil pH, since high soil pH increases the potential of soils to sorb Zn, rendering it unavailable in soils with high pH values, which was the case in these soils after heating [10].

The changes in PTE partitioning into the different geochemical fractions in the heat-affected soils after sludge addition were more evident in soil sample 1 than soil sample 2, and in the samples that received higher sludge application rates (60:40 vol%) than in those that received sludge at a lower rate (80:20 vol%). These results show that sludge application, especially at high rates, affects the redistribution of PTE in soil geochemical fractions even after the soils have been exposed to high temperatures. Sludge application decreased the percentage of PTEs in the available fraction (F1) compared to the residual fractions (F2, F3, and F4) with consequences on the mobility of the PTEs in the soil environment.

3.9. Effect of Temperature-Induced Changes in Soil Mineralogy and Elemental Oxides on PTE Segregation into Different Geochemical Fractions

Potentially toxic elements present in the F2 fraction are bound to Fe, Al, and Mn oxides, which have high volatilization temperatures and are not likely to have been affected as a direct consequence of increases in soil temperatures. This was reflected in the fact that the changes in the concentration of major and trace element oxides were insignificant. However, increased soil temperature could have indirectly affected this fraction of PTE through temperature-induced changes in the redox conditions, as was found by Qi et al. [73]. Other reports on the effect of temperature on soil redox conditions have been presented by Dorau et al. [74]. The increases in the F2 fraction of some PTEs could have been caused by absorption on soil constituents. Studies by Trehan and Sekhon [75] and Almås et al. [76] have shown that a decrease in soil OM content (which in this study occurred with an increase in soil temperatures) results in greater interaction between PTEs present in the soil and soil particles, which may result in their binding onto these particles through sorption on the soil exchange sites. This may explain the increase observed in the F2 fraction of some PTEs with an increase in soil temperatures.

The F4 fraction comprised PTEs that were bound to the minerals contained in the soils and were not available. Although the changes in mineral composition caused by temperature changes were insignificant, the concentrations of some PTEs in the F4 fraction reduced, whereas PTE concentrations in the other fractions increased as the soil temperature increased. Secondary soil minerals contribute towards the surface charge density of soils and consequently their sorption capacity. A reduction in the abundance of secondary minerals in the soils with an increase in soil temperature as observed in this study (Table 1) would have liberated the PTEs bound by these minerals with a resultant decrease in their concentration in the F4 fraction and a possible increase in the F1 fraction. This would also be possible because the organo-mineral complexes that were formed in the soils would have been destroyed by the high soil temperatures. Whereas this is likely for PTEs with high affinity for these complexes, those with high affinity for the silicates (F4) would have increased because of the sorption of some PTEs by soil particles. The reduction in soil OM observed with an increase in soil temperatures would also have increased the contact between the PTEs and soil particles because particles that were previously coated by OM would have been exposed and become available for sorption of the PTEs. This has been reported for Zn [75,76].

The elimination of the secondary minerals in the soils at the expense of the primary minerals meant that the potential of the added sludge to form organo-mineral complexes with a capacity to reduce the amount of PTEs in the mobile fraction of the soils was significantly reduced. This was reflected in the pattern of metal concentrations in the F2 and F4 fractions of the soil with sludge addition where most PTEs experienced a decrease in concentration in the F2 fraction and an increase in the F4 fraction. No clear pattern could be identified, however, which may be associated with the fact that the changes in the mineral content and major and trace element oxides were minimal.

3.10. Effect of Soil Temperature Increase and Sludge Application on PTE Mobility in the Soils

The percentage of total concentrations of PTEs in the mobile fraction of both soil types that had been heated and amended with sewage sludge was in the range of 2.69%–8.98% for As, 13.26%–30.88% for Cd, 1.08%–9.25% for Co, 5.96%–19.23% for Cr, 4.39%–26.80% for Cu, 2.01%–7.67% for Ni, 6.91%–29.94% for Pb, and 1.40%–12.04% for Zn (

Table 5. Mobility factors of PTEs in heated and sludge-amended soils.

Potentially Toxic Elements	Temperature of Heating	PTE Mobility (%)
		No Sludge		80:20		60:40
		Sample 1	Sample 2	Sample 1	Sample 2	Sample 1	Sample 2
As	no heating	8.72	6.10	4.81	4.29	3.98	3.02
	150	9.52	5.33	7.44	4.11	5.16	3.68
	300	8.61	6.66	5.80	4.50	5.30	2.69
	500	8.98	6.67	3.75	5.24	5.03	5.00
	750	8.97	5.52	4.95	3.64	5.89	3.24
Cd	no heating	28.34	22.70	19.28	19.50	13.68	14.66
	150	28.79	24.26	19.44	18.91	13.26	15.73
	300	29.56	28.57	19.50	18.84	18.23	15.51
	500	29.84	29.81	27.76	27.02	23.20	25.23
	750	30.88	24.39	24.32	20.69	20.40	17.34
Co	no heating	5.97	2.10	4.14	2.35	6.50	1.08
	150	6.11	5.09	6.07	5.12	5.17	4.84
	300	5.49	5.06	5.41	4.93	5.01	3.89
	500	5.74	5.07	4.34	4.87	6.50	2.63
	750	5.85	5.08	4.41	4.29	9.25	5.65
Cr	no heating	14.14	16.10	8.98	19.72	7.46	15.70
	150	19.23	16.91	17.60	12.06	15.40	6.90
	300	11.42	18.92	7.97	12.23	5.96	9.00
	500	14.31	19.75	9.04	11.59	7.55	7.85
	750	15.11	18.61	10.60	9.38	9.91	10.20
Cu	no heating	26.07	23.01	19.29	5.47	15.02	4.90
	150	25.78	24.50	19.56	5.22	15.64	5.83
	300	26.80	22.37	23.20	20.93	17.48	4.80
	500	25.55	24.16	18.97	14.73	15.04	4.39
	750	26.15	25.63	19.57	22.50	16.98	4.69
Ni	no heating	8.39	7.67	6.47	3.83	4.47	2.73
	150	9.48	5.38	7.32	4.45	5.12	3.14
	300	7.45	5.37	4.39	4.80	2.19	3.60
	500	8.58	5.35	3.63	4.04	2.01	2.94
	750	8.79	5.32	4.57	4.00	2.73	3.20
Pb	no heating	26.32	22.33	12.02	7.58	18.55	6.91
	150	19.80	29.05	14.83	7.01	13.51	7.73
	300	18.75	29.94	14.93	26.61	11.99	24.03
	500	23.31	29.15	17.90	22.31	13.95	19.56
	750	16.74	29.48	13.51	21.38	10.30	16.49
Zn	no heating	6.79	12.04	2.07	4.33	3.26	1.80
	150	6.72	5.19	2.37	5.77	4.98	1.40
	300	6.14	5.19	4.16	3.79	3.78	3.16
	500	6.76	6.18	0.51	4.89	1.57	0.70
	750	6.94	7.15	6.11	5.80	5.65	6.80

), indicating that Cd and Cu and Pb were the most mobile in the soil. The mobility of the PTEs in the soils that had not been heated followed the order Pb > Cd > Cu > Cr > As > Ni > Zn > Co as shown in Table 5, with the PTEs being more mobile in soil sample 1 compared to soil sample 2. Up to a soil temperature of 300 °C, there was an increase in the mobility of As, Co, Pb, Ni, and Zn, but thereafter, a decrease in mobility was observed (Table 5). These findings are in alignment with what Jing et al. [77] and Latosinska and Gawdzik [78] reported for Zn, Pb, and Cd where temperatures above 20 °C resulted in an increase in their mobility in soils. The pattern was, however, slightly different for Cu, where little changes in mobility were observed with an increase in soil temperature (Table 5).

In the unheated soil samples which received sludge at application rates of 80:20 and 60:40 vol%, the patterns of PTE mobilities were Pb > Cd > Cu > As = Cr > Ni > Zn > Co and Pb > Cd > Cu > Cr > As > Ni > Zn > Co, respectively, indicating minimal changes in the pattern of metal mobilities with sludge application although the mobilities for specific PTEs decreased (Table 5). At a sludge application rate of 80:20 vol%, the mobilities of Cd, Cr, Ni, Pb, and Zn increased with an increase in temperature, with As, Co, and Cu showing random patterns or no change, as indicated by the values in Table 5. When the sludge application rate was increased to 60:40 sludge:soil vol%, the mobilities of Cd, Cr, and Pb increased, whereas Ni mobility decreased, with As, Co, Cu, and Zn showing random patterns (Table 5).

Decreases in MF of As, Cu, and Pb were observed between heated soils that received no sludge and those that had been heated and received sludge at an application rate of 80:20 vol%, whereas for soils that had received sludge at an application rate of 60:40 vol%, differences in MF were observed for Cu, Ni, and Pb. The MF values of all PTEs studied were lower for soils that had received sludge at an application rate of 60:40 vol% than for those that had received sludge at an application rate of 80:40 vol% (Table 5). The highest PTE mobilities were recorded for Cd, Cu, and Pb in both samples regardless of whether they were heated or not heated. The mobilities of the PTEs in the soils generally showed an increase with an increase in soil temperature up to 500 °C, with significant increases recorded for As, Cd, Cr, Ni, and Pb as reflected by the p-values in

Table 6. p-values of the effect of temperature and sludge application on the mobility of PTEs in heat-affected soil.

Potentially Toxic Elements	Temperature	Sludge Application Rate	Sludge Application Rate * Temperature
As	0.01	0.01	0.02
Cd	0.004	5.07	0.03
Co	0.559	3.63	1.17
Cr	0.02	2.60	0.10
Cu	0.23	5.19	0.39
Ni	0.04	0.01	0.03
Pb	0.04	0.05	0.01
Zn	1.19	1.24	0.53

. Sludge addition reduced the mobilities of all PTEs in both soil types, with the percentage decrease in mobility being very significant in soils that had been heated to a temperature of 300 and 500 °C. The differences in metal mobility with sludge addition between soils heated to a temperature of 500 °C and 750 °C were insignificant, especially for As, Cr, and Zn. An increase in Co mobility was observed with an increase in sludge addition to soils that had been heated to 500 and 750 °C.

ANOVA analysis showed that the differences in the values of MF of the different PTEs caused by sludge application rate were significant for As, Ni, and Pb, whereas temperature significantly affected the MF of As, Cd, Cr, Ni, and Pb (Table 6). The MF values for As, Cd, Ni, and Pb were influenced by interactions between temperature and sludge application rate (Table 4). Zhang et al. [62] have shown that the humic acids contained in OM tend to form insoluble complexes with soil PTEs, which reduces their mobility, whereas complexes formed between PTEs and fulvic acids are soluble and therefore more mobile. The reduced mobility of the PTEs in the soil with increased sludge application could be an indication that the sludge was rich in humic acids. Considering the affinity of many PTEs including Cd, Ni, Cu, and Pb for OM, one would have expected sludge application to significantly affect the mobility of these PTEs in soils, but the fact that these soils had been exposed to high temperatures would have resulted in a possible alteration in various soil properties affecting the ability of these PTEs to form metal organo-mineral complexes in the soil with consequences on their mobility in the soil environment. These effects become visible at soil temperatures of 200 °C and above, which is also the temperature at which OM begins to disappear from the soil.

4. Implications on Fire-Affected Soil

When fire events occur, the soil could experience changes in temperature which could last for several days [79,80]. Changes in soil mineral composition and elemental oxides are not significantly influenced by soil fires, especially if soil temperatures are not significantly changed. Various materials including manures, biochar, and sewage sludge have been used to immobilize PTEs in the soil environment, but in heat-affected soil, the efficiency of these materials is compromised. A high sludge application rate would be recommended for the immobilization of As, Ni, and Pb in fire-affected soils, but applying sludge at a high rate may result in high nutrient contents in the soil and their possible leaching into surrounding water resources, compromising their quality. Considering that high temperatures affect the mobility of most PTEs, attention in fire-affected ecosystems should be directed not only to the vegetation dynamics, microbial activities, and soilwater interactions, but also to the mobility of PTEs because whether contaminated or not, high soil temperatures tend to mobilize PTEs in the soil environment. The use of sewage sludge as a PTE immobilizer in such soils should also be given careful consideration because a higher-than-recommended application rate may be required to achieve success, which is likely to compromise the surrounding environments, especially water resources, because of the potential of high nutrient introduction.

5. Conclusions

An increase in soil temperatures causes changes in soil physicochemical properties that impact the interaction between soil components and PTEs. Increased soil temperatures tend to increase the concentrations of the non-residual fractions (F1, F2, and F4) of PTEs in soils relative to the residual fraction (F4), with a consequent increase in the mobility of PTEs in these soils. An increase in mobility may imply an increase in the bioavailability of the PTEs for uptake by plants and mobilization in the environment. This may result in the contamination of surrounding water bodies with PTEs. Using sewage sludge as a soil PTE immobilizer has proven to be an environmentally friendly and efficient method of sludge disposal that also retards the migration of PTEs in the soils, but this efficiency is compromised in soils that have experienced high temperatures as sludge–soil interactions are compromised by changes in soil properties induced by the increased soil temperatures. For heat-affected contaminated soils, higher rates of sludge application may be required to immobilize PTEs compared to soils that have not experienced fires. The reaction is also PTE-dependent as PTEs have different affinities for different soil geochemical fractions. Therefore, attention needs to be given to the mobility of PTEs in fire-affected soils because they are mobilized in soils that have experienced high soil temperatures. Further studies are needed to determine whether these changes are reversible or irreversible.