Introduction

With rapid population growth worldwide, urban areas have become the spaces most affected by intensive human activities including residential, traffic, industrial, and commercial works. Consequently, a steady increase in heavy metal(loid) (HM) contamination has been observed in soils of global cities, resulting in environmental contamination and subsequent high risk to human health. Several investigations have been conducted on soil contamination caused by HMs such as As, Cd, Cu, Pb, and Zn. Because most HM contaminants are undoubtedly toxic to terrestrial organisms, several studies are still being conducted on the metals present in the soil. Nevertheless, there have been relatively rare studies focusing on Cr and Ni contaminants because they exhibit less toxic effects to the ecosystem and human health compared with other HM contaminants (Cheng et al. 2014; Mansourri and Madani 2016). However, there is considerable evidence showing elevated concentrations of Cr and Ni in the surrounding environment, which indicates the urgent need to focus on these two metals. Chromium and Ni are included among the most widespread HMs in soils because they may arise from numerous sources in urban and industrial areas (Adriano 2001; Bodek et al. 1988; Ho and Tai 1988). Both metals are known to exist in the soil at levels as high as 20,000 mg kg−1 in some highly industrialized areas (Kabata-Pendias and Pendias 2000). In roadside soils, twofold concentrations of Cr and Ni were determined compared with those in natural soils (Münch 1993). Also, the concentrations of Cr and Ni were found to be high in the soil around a specific industrial complex (Jeong et al. 2015), and it was reported that the load of Cr and Ni in the soil also increased in the rooftop vegetable garden in Seoul, where there is a lot of human activity (Kim et al. 2015).

The concentrations of Cr and Ni in the environment are generally associated with the natural background concentration and are mostly enriched in soils occurring in ultrabasic terrains (Wang and Zhang 2018). Also, they are naturally present in all types of rocks, particularly in serpentine rocks, and in soils originating from ultramafic rocks (Mrvić et al. 2009; Quantin et al. 2008), and occur in the pedosphere at various concentrations compared to other trace elements. These metals are not very soluble and toxic in these naturally enriched environments, as they are generally bound in resistant minerals such as silicates, oxides, and spinel (Becquer et al. 2006; Oze et al. 2004a). However, Cr and Ni derived from anthropogenic sources are generally toxic and very mobile, which include effluents from the agricultural industry such as intensive fertilization (Papadopoulos et al. 2007; Papastergios et al. 2007), fossil fuels or coal-burning emissions from the local power plants (Sarkar et al. 2006), manufacturing and construction activities, leather factory and tannery (Tariq et al. 2006), cement factory (Banat et al. 2005), vehicle emissions, and ore mining and smelting processes (Salt et al. 2000). Many previous studies have predicted that the significant correlation between Cr and Ni found in urban soils may be due to these contaminant sources simultaneously releasing both metals into the surrounding soil (e.g., Cheng et al. 2018; Krishna and Govil 2005; Yekeen and Onifade 2012; Zhao et al. 2022, etc.). However, studies on the factors causing the co-existence of Cr and Ni in the anthropogenic soils are very limited so far.

Chromium and Ni released into the environment can be absorbed by living organisms and can enter the human body through inhalation, skin exposure, or intake of contaminated drinking water and foods (Zhao et al. 2022). The major harmful effects on humans are vomiting, gastrointestinal bleeding, liver damage, and acute renal failure appear (Jung et al. 2012) and, for Ni, its addiction causes chronic bronchitis, damage to lung function, and nasal cancer (ATSDR 2005). Such harmful effects of Cr and Ni are largely dependent on absorption and exposure time but are also determined by differences in contamination levels according to land use types and history within anthropogenic areas. Today, Cr and Ni are emerging as the most responsible elements for environmental toxicity due to their increased releases derived from anthropogenic activities, especially in areas where urbanization and industrialization are accelerating. Therefore, like other HMs, it is a worrisome issue to determine the present status of these two toxic metals, redress the subsequent environmental problems, and adopt a future mitigation strategy. In the present review, we (1) provide an overview of contamination levels of Cr and Ni and their relationship in different types of anthropogenic soils around the world such as roadside, urban CBD, and industrialized areas, and (2) propose appropriate practices for the remediation of soils co-contaminated with Cr and Ni, along with a description of the chemical nature of the two metals in the soil.

Concentrations of Cr and Ni in different anthropogenic soils

In general, the concentrations of naturally occurring Cr and Ni in soils developed from ultramafic rocks are up to 125,000 and 10,000 mg kg−1, respectively (Adriano 1986; Hseu 2006; Oze et al. 2004b), whereas their concentrations in other soils commonly range from 0 to 100 mg kg−1 (Kabata-Pendias and Pendias 2000; McGrath 1995). The regulation of maximum permissible limits of Cr and Ni in soils has been established in different countries (Table 1), and large variations exist in such regulations according to the different countries. In this review, we evaluated the concentrations of Cr and Ni in different land-use types, which are affected by human activities, including roadside, urban CBD, and industrial area soils (Table 2). For this purpose, we collected scattered literatures that recorded the concentrations of Cr and Ni in anthropogenic soils of 46 global cities exceeding the background global soil data (50 mg kg−1 for Cr and 25 mg kg−1 for Ni) reported by Berrow and Reaves (1984) and the maximum allowable limits (100 mg kg−1 for Cr and 50 mg kg−1 for Ni) recommended by the World Health Organization (1996). As shown in Table 2, the contamination levels of Cr and Ni in soils depend on different land-use types and are generally higher than the global background concentration. The Cr and Ni contamination levels in different land-use types observed in different regions of the world can be arranged in the following descending order: industrial > CBD > roadside soils.

Table 1 Maximum permissible concentration (MPC) and background concentration (BC) for Cr and Ni in soils of different countries
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Table 2 Concentrations of Cr and Ni in different types of anthropogenic soils, such as roadside, central business district (CBD), and industrial area, in various cities around the world
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As shown in Fig. 1, the highest concentrations of Cr and Ni are observed in industrial soils. Legacies of industrial activities are likely the major cause of Cr and Ni released into the urban soils: 91 up to 1840 mg kg−1 for Cr and 45 up to 769 mg kg−1 for Ni. Of 18 industrial soils, the peak concentrations of both metals were observed in soils of Berlin, Germany (Birke and Rauch 2000). The concentrations of Cr and Ni in soils are also dependent on the industrial type. The higher concentrations of Cr and Ni were found in soils close to numerous industries such as cement factory, mining, leather and tannery factory, electroplating, metal, chemical, textile, ceramic, and industrial complexes (Deepali and Gangwar 2010; Kabir et al. 2012; Kashem and Singh 1999; Nieboer and Nriagu 1992; Yaylali-Abanuz 2011).

Fig. 1
figure 1

Distribution of Cr and Ni concentrations in soils of different land use types

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In roadside soils worldwide, the concentrations of Cr and Ni were 73–240 and 39–161 mg kg−1, respectively, and the highest concentrations were detected from roadside soils in Kavala, Greece, for Cr (Christoforidis and Stamatis 2009) and in Eskisehir, Turkey, for Ni (Malkoc et al. 2010) (Table 1). Such contamination of roadside soils and dusts by Cr and Ni is significantly associated with the traffic environment (Christoforidis and Stamatis 2009; Malkoc et al. 2010). In fact, the effluents of long-term traffic activities can be the major source causing the accumulation of various HMs (Zhang et al. 2012). As such, the accumulated Cr and Ni contaminants in the traffic areas were caused by fuel and lubricant oil combustion, tire corrosion, wear of brake linings, discharge from batteries, and emission from gasoline vehicles (Adamiec et al. 2016; Apeagyei et al. 2011; Hjortenkrans et al. 2007; Maeabal et al. 2019; Pulles et al. 2012; Shinggu 2014). Hjortenkrans (2008) also reported that the total metal concentrations, including Cr and Ni, in roadside soils have increased by 3–16 times compared with regional background concentrations during the past few decades.

As shown in Table 1, the contamination levels of Cr and Ni in urban CBD soils were generally higher than those in roadside soils. This is because urban soil is a complex area affected by diverse contaminants released from sources such as industrial activities, commercial districts, residential areas, and road networks with heavy traffic (Gunawardana et al. 2012; Pal 2012). The concentrations of Cr and Ni in urban soils were 79–870 and 31–910 mg kg−1, respectively. The peak concentrations of Cr and Ni were found in Torino, Italy (Biasioli et al. 2006) and in Jinchang, China (Liao et al. 2006), respectively. The contamination levels of both metals in urban soils have been increasing globally but varying with the age of the settlement, current emissions from traffic and industry, and washout (Sager 2020). The primary causes of urban soil contamination with Cr and Ni are anthropogenic sources such as vehicle emissions, industrial discharges, energy production, domestic wastes and domestic wastewater from household washing products, and various construction wastes (Alloway 1995; Biasioli et al. 2006; Galitskova and Murzayeva 2016; Thornton 1991).

Relationship of Cr and Ni in anthropogenic soils

There is limited information about the strong correlation of Cr and Ni levels in soils. A linear regression analysis conducted using the collected data revealed strongly positive relationships between Cr and Ni levels in roadside (r2 = 0.9604), CBD (r2 = 0.9345), and industrial (r2 = 0.7608) soils (Fig. 2). Similarly, several studies reported significantly positive correlations between Cr and Ni concentrations in various soils (Abdel-Salam and Abu-Zuid 2015; Aziz et al. 2015; Cheng et al. 2018; Luo et al. 2017; Petrotou et al. 2010). Krishna and Govil (2005) reported that Cr–Ni ratio was closely related to a correlation coefficient of 0.72041 by indicating the close relationship of anthropogenic Cr and Ni sources. Yekeen and Onifade (2012) also described similar results of a positive correlation between Cr and Ni concentrations in roadside soils affected mostly by automobile emissions. Such a strong relationship between both metals may be attributed to the same geochemical affinity, origin, and geochemistry (Poznanović Spahić et al. 2018; Wang and Zhang 2018). Occasionally, Cr and Ni contamination in anthropogenic soils is associated with the natural background concentrations (Shen et al. 2019; Shi et al. 2008). Weathering or erosion of ultramafic-derived constituents such as peridotite, dunite, and pyroxenite, particularly rich in serpentinite, is said to be responsible for the enrichment of Cr and Ni contents in the soil, which contributes to their high correlation (Albanese et al. 2015; Kulikova et al. 2019; Oze et al. 2003; Tassi et al. 2018; Venturelly et al. 1997). Furthermore, these two metals have similar geochemical properties such as the ionic size and radius which account for the strongly positive correlation observed between them (Mandic et al. 2002; Nuamah Daniel et al. 2019). In addition, these two elements in the soil are generated together by the same anthropogenic sources which include leather processing (Tariq et al. 2006), cement plant (Banat et al. 2005; Farzadkia et al. 2016), the metal processing factory (Panagopoulos et al. 2015), and combustion of fossil fuels (Sarkar et al. 2006).

Fig. 2
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Linear relationship between Cr and Ni concentrations in human-affected soils such as a roadside soils (n = 10), b urban CBD soils (n = 18), and c industrial soils (n = 18) (logarithm values)

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Chemistry of Cr and Ni in soils

The mobility and distribution of Cr and Ni in the soil environment are controlled by several factors such as soil pH, clay, and organic matter through reactions such as the redox reaction, precipitation–dissolution, and sorption–desorption (Chauhan et al. 2008; Saleh et al. 1989; Zayed and Terry 2003).

Chromium

Depending on the pH and redox condition, Cr in soils can occur in two oxidation states, Cr3+ and Cr6+. Chromium(III) is less mobile, toxic, and bioavailable than Cr6+ and occurs Cr(OH)3 precipitation in natural soil (Yang et al. 2022). In contrast, Cr6+ exhibits strong mobility and high bioavailability and is commonly found as forms of HCrO4− and CrO42− oxyanions in contaminated sites (Song and Ma 2017). The Cr mobility and speciation are always controlled by redox reaction, precipitation-dissolution, and sorption-desorption reactions in contaminated soil environments (Zayed and Terry 2003). The fate of Cr in soil is partly governed by oxidation-reduction reactions and soil pH. Chromium(VI) is the dominant form of Cr where aerobic conditions exist. Under anaerobic conditions, Cr6+ can be reduced to Cr3+ in soils by redox reactions with aqueous inorganic species, electron transfer at mineral surfaces, reactions with non-humic organic substances such as carbohydrates and proteins, or reduction by soil humic substances. The latter, which constitute most of the organic fraction in most soils, represent a significant reservoir of electron donors for Cr6+ reduction (Kožuh et al. 2000). In addition, soil properties like soil organic matter and iron (Fe) oxides influence Cr6+ reduction. Soil organic matter is a major reductant as it serves as an electron acceptor and reduces Cr6+ by releasing CO2 (Dong et al. 2014). Iron oxides retain Cr3+ to form (Cr, Fe)(OH)3 by limiting the oxidation of Cr3+ in soil (Rai et al. 1989). Besides, soil pH affects the reduction of Cr6+ and its reaction increase with decreasing pH values. In aerobic soils, the reduction of Cr6+ to Cr3+ is possible even at a slightly alkaline pH if the soil contains organic energy sources to perform the redox reaction (Adriano 1986). On the other hand, the oxidation of Cr3+ to Cr6+ is controlled by the concentration of water-soluble Cr, pH, initial available surface area, and ionic strength (Fendorf and Zasoski 1992). In general, the oxidation of Cr3+ to Cr6+ is a very slow process at pH > 5 (Eary and Rai 1987). Unlike Fe oxides, manganese (Mn) oxide is known for increasing the oxidation of Cr3+ in soils (Rai et al. 1989). In aerobic zones, dissolved Cr3+ can be oxidized on the surface of MnOx under acidic conditions (pH < 6.0), but under alkaline conditions (pH = 8.0–9.4), Mn(II)-catalyzed oxidation plays a dominant role (Liang et al. 2021).

In addition, Cr speciation is governed by precipitation and dissolution reactions. The solubility of chromium(III) is limited by the formation of several insoluble oxides or hydroxides. In neutral and high soil pH, Cr3+ can readily substitute for Fe3+ in minerals that precipitate as insoluble chromic hydroxide, Cr(OH)3 (Saleh et al. 1989). In soils, most Cr occurs as Cr3+ and within the mineral structures or forms of mixed Cr3+ and Fe3+ oxides (Adriano 1986). According to Rai et al. (1987), (Cr,Fe)(OH)3 is a solid phase of Cr3+ that has even lower solubility than Cr(OH)3. Furthermore, Cr solubility in soil may be altered to some extent by root exudates released in the rhizosphere, such as organic acids, which react strongly with metal ions in the soil aqueous and solid phases (Jones and Darrah 1994). As a result of Zeng et al.’s (2008) study, the production of low molecular weight organic acids (e.g., oxalic and malic acids) from rice roots to soil resulted in the enhancement of Cr accumulation in rice plants, suggesting the interaction between Cr and organic ligands which lead to the formation of mobile organically bound Cr3+.

Sorption and desorption reactions are important factors governing Cr solubility and mobility in soil (Zayed and Terry 2003). Several researchers have also described that Cr in soils is associated with Fe oxides as a second important factor controlling metal solubility, mobility, and sorption (Contin et al. 2007; Ge et al. 2000; Hernandez et al. 2003; Manceau et al. 2007). Iron ions are involved in the mechanism of metal sorption by the isomorphic substitution of divalent or trivalent cations (Sipos et al. 2014). Adsorption of Cr6+ onto the surface of Fe oxides can occur only at an acidic or neutral pH. Chromium mobility also depends on other sorption characteristics of the soil including the clay content and amount of organic matter present. Hexavalent Cr (Cr6+) adsorbs onto soil surfaces, especially iron, aluminum oxides, and clay minerals (kaolinite) (Rai et al. 1989). On all of these solids, Cr6+ adsorption increases with decreasing pH (Rai et al. 1989). Cr6+ adsorbs more tightly onto oxide and clay particles than other anions such as chloride, nitrate, or sulfate, but it can be desorbed by the presence of excess amounts of phosphates due to the competition of phosphates with chromates for the same adsorption sites (Adriano 1986; Bartlett and Kimble 1976). Cr3+ is adsorbed 30–300 times more strongly onto soil clay minerals than Cr6+ (Griffin et al. 1977). At low pH (< 4), in general, Cr3+ is the dominant form of Cr and can also form solution complexes with Cl−, CN−, F−, OH−, SO42−, NH3, and soluble organic ligands. As pH increases, the adsorption of Cr3+ onto clay and oxide minerals increases, whereas it decreases for Cr6+, with no further sorption of Cr6+ occurring at pH > 8.5. The increased adsorption of Cr3+ at high pH is attributed to the cation exchange adsorption of Cr3+ hydrolyzed species (Chrostowski et al. 1991; Griffin et al. 1977).

Nickel

In nature, Ni is mostly present in the form of nickelous ion, Ni2+, whereas the hydrated form Ni (H2O)62+ is the most common form of Ni found in soil solution. Ni2+ is highly mobile and has phytotoxic effects on plants (Yusuf et al. 2011). The most important oxidation state of Ni is +2, although the +3 and +4 oxidation states are also well known (Greenwood and Earnshaw 1997). Soil pH and redox reaction are the major variables controlling Ni dynamics, mobility, and partitioning in soils. Ni solubility is strongly controlled by soil Ni concentrations and soil pH, implying that the dissolution of Ni increases when the total Ni content increases and the pH decreases, indicating competitive adsorption between Ni and H+ for soil-binding sites. In regions with low pH, the metal exists in the form of nickelous ion, Ni2+. In neutral to slightly alkaline solutions, it precipitates as nickelous hydroxide, Ni(OH)2, which is a stable compound. This precipitate readily dissolves in acid solutions, forming Ni3+, and in very alkaline conditions, it forms nickelite ion, HNiO2, which is soluble in water. In very oxidizing and alkaline conditions, Ni exists in the form of the stable nickelo-nickelic oxide, Ni3O4, which is soluble in acid solutions. Other Ni oxides such as nickelic oxide, Ni2O3, and Ni peroxide, NiO2, are unstable in alkaline solutions and decompose by releasing oxygen. However, in acidic regions, these solids dissolve to produce Ni2+ (Pourbaix 1974).

Ni solubility is also associated with organic matter, oxides, and hydroxides. Because of the susceptibility of Ni2+ to form complexes with organic matter, in several soils its high mobility is maintained even under neutral conditions. This is because Ni2+ creates bonds with organic matter in the form of mobile chelates. The solubility of Ni2+ negatively correlates with Al oxides, as demonstrated by several studies that Ni2+ binds Al oxides or forms a mixture of Ni2+ and Al oxides because of the high affinity of Al oxides for metal (Axe and Trivedi 2002). However, soluble Ni2+ concentration positively correlates with amorphous iron oxides (Zhang et al. 2015). Tipping et al. (2002) showed that Ni2+ could form complexes with dissolved colloidal Fe hydroxides and dissolved organic carbon (DOC), which affected soluble Ni speciation and potentially increased Ni mobility. This is due to the adsorption of humic matter that alters the surface chemistry and colloidal stability of iron oxides. In the case of DOC, it would compete with Ni for the iron oxide surface sites, and consequently, DOC occupies a portion of all iron oxide surface sites, whereas soil absorption capacity decreases relatively and more Ni partitions to the soluble phase.

Numerous Ni sorption mechanisms have been proposed over the past few decades. For acidic soils, cation exchange is perhaps the major mechanism for Ni sorption (Gomes et al. 2001; Papini et al. 2004). Antoniadis and Tsadilas (2007) reported that as the pH increases, Ni sorption on clay lattice edges increases because of the hydrolysis of divalent ions capable of forming inner sphere complexes. Moreover, Ni adsorption by clay is strongly influenced by soil pH as well as silicon and Al oxide surface ratios. The sorption of Ni2+ is also positively related to iron and manganese oxides (Anderson and Christinsen 1988; Tiller et al. 1984).

In soils where Cr and Ni coexist, soil pH is a major contributor to the mobility of both metals. In acidic soils, the migration of Cr3+ and Ni2+ occurs due to their susceptibility to leaching, whereas the Cr6+ adsorption increases with decrease in soil pH (at < pH 3). In soils with high pH, most Cr6+ becomes mobile form because of its low adsorption capacity. For Cr3+ and Ni2+, they precipitate due to the high soil pH environment (Chinthamreddy and Reddy 1999).

Remediation technology for Cr- and Ni-contaminated soils

In recent years, the use of in situ immobilization technology has received considerable interest for treating contamination caused by HMs such as Cr and Ni in soils. Immobilization of Cr and Ni through additives is a very promising technique due to its simplicity and high effectiveness, in situ applicability, and low cost (Bolan et al. 2014; Shaheen et al. 2017a). The in situ immobilization technology is a remediation approach based on readily available amendments such as inorganic and organic soil additives that are inexpensive and environmentally safe. To immobilize HMs in contaminated soil, we normally use immobilizing agents like cement, lime, clay, iron oxides, and composts. However, the type and application rate of soil amendments are always determined according to the soil characteristics, pollution characteristics, pollution level, and pollution type (Yoon and Kim 2006). Soil amendments render metals to form insoluble or immobile, low-toxic matters by metal complexation, precipitation, redox reactions, or sorption (Negim 2009). Consequently, such effects of soil amendments decrease the migration of HMs to water, plants, and other environmental media without altering their total concentrations (Lee et al. 2013; Rinklebe et al. 2017; Zhou et al. 2004). Also, the addition of soil amendments improves soil fertility and increases plant growth in contaminated soils (Calace et al. 2005; Clemente et al. 2006; Lwin et al. 2018). The in situ immobilization of Cr and Ni depends on soil characteristics, concentrations, and forms of Cr and Ni in soils, as well as the mobilization and bioavailability of soil Cr and Ni.

Vink et al. (2010) demonstrated that the addition of gypsum (CaSO4) significantly decreased the dissolved concentration of Cr by 24% due to sulfide formation in soils. Shanableh and Abu-Zer (2001) also reported that the addition of lime in the range of 1%–8% resulted in immobilization that exceeded 97% for Cr. Lee et al. (2016) indicated that the application of rice straw charcoal enhanced the overall Cr6+ immobilization of soils, which is primarily attributed to the Cr6+ reduction capacity of rice straw compost with increasing pH or anion content in the soil solutions. Furthermore, the addition of Enteromorpha prolifera biochar to Cr-contaminated soils transformed 94.22% of Cr6+ into Cr3+ and the Cr content of residue fraction increased by 63.38% compared with that in the control soil (Chen et al. 2021). Moreover, the addition of sulfur compounds to contaminated soils decreased the Cr mobility by 55.8%, respectively, due to the reduction of Cr6+ to Cr3+ by reduced sulfur species and the formation of insoluble salts (Shaheen et al. 2017a). The addition of industrial wastes such as brick factory residual and ceramic powder also reduced the Cr availability by reducing Cr6+ to Cr3+ with reduced sulfur species (Shaheen et al. 2017a). Shi et al. (2016) also reported that the application of elemental S amendment decreased the bioavailable Cr6+ by converting less toxic Cr3+ in soils impacted by leather tanneries. Moreover, the addition of sulfur-based inorganic agents, including calcium polysulfide, pyrite, and sodium sulfide, reduced the Cr6+ concentration in Cr-spiked soils (Mahdieh et al. 2016). Yang et al. (2021) reported that reductive materials such as iron-bearing reductants, sulfur-based compounds, and sulfur-containing organic compounds were the most promising amendments for the remediation of Cr-contaminated soils due to their high efficiency and adaptability. Kulczycki and Sacała (2020) also demonstrated that sulfur application (60 mg kg−1 of soil) alleviated Cr stress in maize and wheat and also improved the total biomass and plant growth. Therefore, reductive materials such as sulfur-containing amendments have significant effects on Cr immobilization through the formation of insoluble forms and by reducing the mobile form of Cr ion (Cr6+) to the immobile form (Cr3+).

Increasing the pH through liming has been found to enhance Ni sorption (Harter 1979), thereby reducing the solubility and mobility of Ni2+ in acid soils. It was also found that activated charcoal and potassium humate decreased the mobility of Ni by 38% and 13%, respectively, which indicated that these organic amendments could adsorb Ni from the soil solution and reduce its mobility by forming stable complexes with humic substances (Lee et al. 2013). The application of phosphate rock also reduced the mobility of Ni by 21.9% by forming complexes with Ni (Shaheen et al. 2017a). The addition of animal wastes (bone meal) also decreased the Ni mobility by 23.7% by increasing the soil pH and formation of metal carbonates (Shaheen et al. 2017a). Shaheen et al. (2017b) also found that the application of compost and sulfur altered the distribution of Ni in dry and wet soils. Compost addition increased the organic fraction of Ni by relatively increasing the content of organic matter. Furthermore, an increase in soil pH by compost application is responsible for the changes in Ni fractionation in soils.

Conversely, the high solubility of Ni with sulfur treatments can be explained by the decrease in soil pH after sulfur treatment (Wang et al. 2008). The application of sulfur was found to increase soil acidification, which can enhance the solubility of Ni in contaminated soils (Catherine et al. 2006; Kaplan et al. 2005; Salati et al. 2010). This result indicated that the mobile fraction of Ni is significantly influenced by soil pH and generally increases as the soil pH decreases (Kayser et al. 2000). Moreover, Shaheen et al. (2015) investigated the impact of various organic and inorganic soil amendments (including activated carbon, bentonite, biochar, cement bypass kiln dust, chitosan, coal fly ash, limestone, nano-hydroxyapatite, organoclay, sugar beet factory lime, and zeolite) on the water-soluble and exchangeable Ni contents in contaminated soil. They reported that the application of different amendments (except organoclay) significantly decreased the water-soluble Ni fraction. Sugar beet factory lime, cement bypass kiln dust, limestone, bentonite, activated carbon, and biochar were the most effective amendments, resulting in a 58%–99% decrease in water-soluble Ni fraction. These results infer that alkaline and carbonate-rich materials can be used to treat Ni-contaminated soils, reducing their solubility due to their alkalinity and high contents of calcium and carbonates, which increase the soil pH and precipitation/sorption of metals in the treated soils.

Furthermore, there are the experimental previous studies that simultaneously alleviate Cr and Ni contamination by applying reducing agents and soil amendments as follows. In Cr and Ni spiked soils, the addition of sulfides resulted in low migration of Cr and Ni by occurring significant reduction of Cr6+ to Cr3+ and increase in soil pH that caused the precipitation of these two metals (Reddy and Chinthamreddy 1999). They also revealed the more pronounced effect of adding sulfides on the migration of both metals rather than other reducing agents including humic acid and ferrous iron. In addition, biochar addition reduced bioaccumulation of Cr and Ni in tomato plants grown in co-polluted soil of Cr and Ni by 93–97% compared to biochar-unamended soil (Herath et al. 2015). Similarly, the application of miscanthus and woodchip ashes showed the effectiveness to stabilize Cr and Ni in contaminated soil (Kang et al. 2016). Cheng-Kim et al. (2016) also reported the remediation effect on Cr and Ni contamination in multi-metals contaminated soil by applying vermicompost of spent mushroom compost.

Table 3 summarizes the results of previous studies on the purification of Cr and Ni contaminants in soil by adding soil amendments described above. A more in-depth understanding of the effects and mechanisms of soil amendments on the geochemistry, dynamics, and distribution of Cr and Ni can contribute to the development of novel remediation strategies and management options for Cr- and Ni-polluted soils as well as shorten the remediation period. In addition, more experiments are required in future studies to confirm the effects of soil amendments on Cr and Ni immobilization at field levels, as the majority of previous experiments were conducted at a laboratory scale.

Table 3 Remediation effects of applying soil amendments into contaminated soil with Cr and Ni reported in previous studies
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Conclusions

Although the database of this review was built based on limited research cases, it could demonstrate that Cr and Ni concentrations in different soil environments generally reflect the influence of various anthropogenic activities such as traffic emissions, industrial discharge, long-range transport, and municipal activities. Observing the generally enhanced metal levels in different anthropogenic soils can indicate that the concentrations of Cr and Ni released by industrial activities are higher than those released by other land-use activities. In addition, the high correlations between Cr and Ni concentrations in all different types of anthropogenic soils indicate that co-pollution of both metals may occur by release from the same sources. Various agents including liming, organic compounds, and phosphates can be used as effective amendments for immobilizing both Cr and Ni in contaminated soil. However, there is a counter-effect of adding sulfur to the chemistry of Cr and Ni; it enhances Ni solubility while it decreases Cr solubility through the reducing process. The findings of our review summarized in Fig. 3 could help in obtaining awareness about Cr and Ni pollution in the roadside, urban CBD, and industrial areas caused by different anthropogenic activities, which would therefore suggest the need for regular monitoring and suitable remediation or management practices to ensure a sustainable environment and reduction of metal contamination in soils, plants, and water.

Fig. 3
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Schematic diagram of sources, chemistry, and immobilization processes of Cr and Ni in anthropogenic soils

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