1 Introduction

Although the surface of the earth dominates with water, there is a shortage of potable water, and more than 1.2 billion people worldwide do not have access to clean drinking water (Bolisetty et al. 2019). Two and a half percent of fresh water was polluted by organic, inorganic, biological and macroscopic contaminants (Shaikh 2020). The industrial wastewater is one of the sources of the contaminants which often migrate into soil with irrigation. Soil and water pollution, that poses serious risks to human health is particularly true in developing countries such as India and China (Zhao et al. 2014). In 2014, the Chinese government published a national soil survey that showed that 16.1% of all soil sampling sites and 19.4% of arable soil sampling sites exceeded the soil standard (National Soil Survey 2014). It were reported that 2.786 × 109 m2 of agricultural soils are polluted by Cd (Liu et al. 2015a). Unlike organic pollutants, heavy metals usually do not disappear over time in the natural environment, and can accumulate in the organism along with the food chain, causing human kidney damage and lower bone density, posing a threat to human and animal health (Jiang et al. 2012). Compared with other toxic metals such as Cu and Pb, Cd which is potentially toxic to biota at lower concentrations demonstrates greater mobility along the food chain and in the environment (He et al. 2015). Therefore, it is imperative to limit the transfer of Cd from wastewater to soils by reducing the Cd content in wastewater used for irrigation, and decrease the availability of Cd in soils by passivation.

To date, many methods have been used for the removal of heavy metals from aqueous systems including electrocoagulation, membrane filtration, ion-exchange, reverse osmosis, electrodialysis, solvent extraction and adsorption (Gunatilake 2015; Mahmood et al. 2011; Parmar and Thakur 2013). Among these methods, adsorption is considered to be the most efficient approach because of its simplicity and low cost of operation, highly efficient wastewater treatment and environmental friendliness (Mohan et al. 2014). However, most of commercially used adsorbents have high cost. The use of low-cost adsorbents such as biochar produced from the thermal degradation of the carbon-rich agricultural biomass is an alternative method for the removal of heavy metals. Recently, biochar has attracted increasing attention due to its important roles in carbon sequestration (Deng et al. 2017), pollutants removal (Nie et al. 2018; Tan et al. 2015) and soil fertility enhancement (Van Zwieten et al. 2010). Biochar derived from different feedstocks or produced under different conditions possesses various characteristics that directly affect its adsorption capacity for pollutants by controlling the surface functional groups, pore distribution, ash category (Ok et al. 2015; Rajapaksha et al. 2016) and redox capacity (Yuan et al. 2017). The heavy metal adsorption capacity of biochar is always lower than those of active carbon or other commercial adsorbents due to its limited amount of adsorption sites (Mohan et al. 2014). To date, an increasing number of studies have focused on the removal of toxic metals using biochar loaded with different minerals such as magnetite, hematite (γ-Fe2O3), zero-valent Fe (ZVI), hydrous Mn oxide and Fe–Mn oxides (Lu et al. 2016; Wang et al. 2016b, 2015). Magnesium oxide (MgO) is a nontoxic, low-cost and environment-friendly material produced from the naturally abundant and safe Mg element (Tian et al. 2017). It has been proved that MgO is a high-efficiency adsorbent for many types of pollutants such as phosphate, nitrate (Zhang et al. 2012), lead, cadmium (Cai et al. 2017; Chowdhury et al. 2016), nickel (Feng et al. 2015), fluoride (Jin et al. 2016b) and Congo red (Tian et al. 2013). Shen et al. (2019) reported that magnesium oxide (MgO)-coated corncob biochar reduced TCLP-leached lead from 10.63 to 5.24 mg L−1 (reduced by 50.71%). So here come the questions: could MgO-loaded biochar composite remove Cd efficiently in aqueous solutions? what’s the mechanism? would co-existing Pb2+ affect the Cd removal? and what about the effect of MgO-loaded biochar composite on the availability of toxic metals in polluted soil? In this study, MgO-laden corn straw biochar composite was synthesized and used for the immobilization of Cd2+/Pb2+ from aqueous solutions and co-contaminated soil. The objectives of this study are to (1) investigate the adsorption capability and mechanism of MgO-laden corn straw biochar for Cd2+; (2) evaluate the ecological risks of heavy metals (Cd and Pb) in contaminated soils after remediation by MgO-laden biochar composite.

2 Materials and methods

2.1 Preparation of corn straw biochar and MgO-laden biochar

Corn straw biochar was produced by heating the sieved corn straw (particle size < 2 mm) in a tube furnace under nitrogen at a heating rate of 10 °C min−1 to reach 600 °C for 1 h. The MgO-laden biochar was prepared by mixing corn straw (particle size < 2 mm) with 2 mol L−1 of MgCl2·6H2O solution at a solid-to-liquid ratio of 1:7 g mL−1. The suspension was placed in an ultrasonic vibration instrument for 30 min and then stirred intensely for 1.5 h at 25 °C. Then, the suspension was dried at 80 °C. The dry mixture of corn straw biomass and MgCl2·6H2O was heated to 600 °C for 1 h at 10 °C min−1 under N2 flow. After cooling, the modified biochar was gently crushed and sieved through 40 mesh (< 0.425 mm) but not through 200 mesh and sealed in a container for further use (Shen et al. 2019; Zhang et al. 2012). In this study, the MgO was obtained at a high heating temperature. The MgO synthesis is based on the following reaction (Eq. 1).

$$ {\text{MgCl}}_{2} \cdot 6{\text{H}}_{{2}} {\text{O}} \to {\text{MgO}} + 2{\text{HCl}} \uparrow + 5{\text{H}}_{{2}} {\text{O}} $$
(1)

2.2 Evaluation of adsorption properties

MgO-laden biochar (1 mg or 10 mg) was weighed into a 50 mL plastic vessel, and then 20 mL of heavy metal solutions with different concentrations (10–1000 mg L−1) were added into the vessels. After shaking for 24 h (25 °C) at 200 rpm, the liquid phase was filtered through nylon membrane filter with a pore size of 0.45 μm. After dilution with 1% HNO3, the concentrations of the heavy metals and Mg2+ were determined by Inductively Coupled Plasma Mass Spectrometry (ICP-MS) (NexION 300X, PerkinElmer, USA) and Inductively Coupled Plasma Optical Emission Spectrometry (ICP-OES) (iCAP6300, Thermo, USA), respectively.

MgO-laden biochar (100 mg) was weighed into a 500 mL conical flask containing cadmium solutions (200 mL) with different concentrations. The adsorption kinetics test was carried out in a mechanical shaker at 25 °C with the rotation speed of 200 rpm. The liquid samples were collected at different intervals and filtered immediately. The concentration of the metal was determined by ICP-MS after dilution with 1% HNO3.

The adsorption capacity (Q, mg g−1) of the adsorbent was calculated using the following equation (Eq. 2):

$$ Q = (C_{0} - C_{t} ) \times v/m $$
(2)

where v (L) is the volume of the Cd2+ solution, C0 (mg L−1) is the initial concentration of the Cd2+ solution, Ct (mg L−1) is the concentration of the Cd2+ solution after adsorption at time t (min), and m (mg) is the mass of the adsorbent. All the experimental treatments were performed in duplicate and the average values were used as the results. Additional analyses were conducted whenever two measurements showed a difference larger than 5%.

2.3 Effect of initial pH and co-existing Pb2+ on the removal of Cd2+ by MgO-laden biochar

Solutions with several initial pH ranging from 2.0 to 8.0 were used in order to estimate the effect of pH on the removal of cadmium by the MgO-laden corn straw biochar in aqueous solutions. The pH values of the solutions (at different initial pH) with and without Cd2+ were determined using a pH meter after shaken for 24 h. To explore the tangible impact on the cadmium removal in wastewater, lead was selected as the co-existing metal. The adsorption isotherm and kinetics in the dual-metal system were studied. The initial concentrations of Cd2+ and Pb2+ ranged from 20 to 400 mg L−1.

2.4 Biochar characterization before and after sorption of Cd2+

To investigate the mechanism of Cd removal by MgO-laden corn straw biochar, the properties of MgO-laden biochar before and after cadmium adsorption were examined. The specific surface area of the biochar was measured by analyzing N2 sorption at 77 K (Tristar 3020, Micromeritics Instrument Corp., USA). Prior to the adsorption of N2, the biochar samples were vacuum degassed at 200 °C for 12 h. The specific surface area was calculated according to the Brunauer-Emmett Teller (BET) method. The micropore surface area and volume were calculated by the t-pot method. The surface morphology images of the biochar samples before and after Cd2+ adsorption were obtained using a scanning electron microscope combined with an energy dispersive X-ray detector (SEM–EDX, Quanta FEG650, FEI, USA). An X-ray diffractometer (XRD) (D8 Advance, Bruker, Germany) was employed to analyze the mineralogical characterization of the biochar samples at the 2θ angles in the 10°–80° range with Cu Kα radiation at a voltage of 40 kV and current of 40 mA, with a step size of 0.02° (Li et al. 2018). The functional groups present in the biochar were determined using a Fourier-Transform infrared (FTIR) spectrometer (Nicolet 6700, Thermo Fisher). FTIR spectra were obtained at a resolution of 4 cm−1 from 4000 to 500 cm−1 (Wang and Liu 2018). Biochar pH was determined after shaking a biochar sample (1 g) in deionized water (20 mL) for 1.5 h (Rajkovich et al. 2012). The C, H and N contents of the biochars were determined according to ASTM D 3176 using an elemental analyzer (Vario EL Cube, Elementar). And the oxygen values of biochars were calculated by difference. Ash contents of the biochars were determined according to ASTM D1782-84.

2.5 Soil sample collection and incubation experiments

The metals contaminated farmland soil with pH value of 5.12 was collected from Shangyu, Zhejiang province (N30.13°, E120.88°) around lead–zinc mining area. The soil was collected from 0 to 20 cm, mixed thoroughly, air-dried and ground to pass through 2 mm sieve. The Cd/Pb content was determined by ICP-MS (NexION300X, PerkinElmer, USA) after digested by HF-HNO3 (3:1) using a microwave digestion system (MARS6, CEM, USA).The soil pH was measured using a pH meter (Seven Excellence TM, S470, Mettler Toledo, Switzerland) at a soil-to-water ratio of 1:2.5 (w/v) and the dissolved organic carbon (DOC) was extracted at a soil-to-water ratio of 1:5 (w/v) and measured by total organic carbon analysis (MULTI N/C 3100, Analytik Jena, GER).

Incubation experiments were carried out in plastic beakers to evaluate the effect of biochar and MgO-laden biochar on heavy metals availability. Different biochars were thoroughly mixed with 200 g soil at different ratios. Four treatments were as follows:

Control (CK, without remediation), 4% corn straw biochar amendment (w/w), 1% MgO-laden corn straw biochar amendment and 4% MgO-laden corn straw biochar amendment. Then, the soil samples were incubated at 25 °C and constant humidity (60% of water holding capacity) for 50 days.

2.6 Soil heavy metal availability and potential ecological risk assessment

Toxicity characteristic leaching procedure (TCLP) was often used to evaluate the activity of heavy metals in soils (Yu et al. 2014). In this study, TCLP method (EPA Method 1311) was used to extract the available toxic elements. Briefly, 1.5 g of soil was mixed with 30 mL of acetic acid solution (pH = 2.88). The supernatants were filtered and the concentrations of Pb and Cd were determined by ICP-MS after dilution. All tests were conducted in triplicate.

In this study, Nemerow pollution index (Nemerow 1971; Sun et al. 2006) and potential ecological risk index (Hakanson 1980) were used to assess the ecological risk of toxic elements of soils treated with different amendments.

(1) Nemerow pollution index (NPI)

$$ P_{i} = C_{i} /S_{i} $$
(3)
$$ {\text{NPI}} = \sqrt {\frac{{P_{i\max }^{2} + P_{{i{\text{ave}}}}^{2} }}{2}} $$
(4)

where Pi is single pollution index, Ci is toxic element concentration that TCLP extracted; Si is the standard toxicity concentration (mg L−1) of a contaminant (The contents of Pb and Cd are 5 and 0.5 mg kg−1, respectively) (Gao et al. 2020; Sun et al. 2006); NPI is pollution index of multi-metals; Pimax is the maximum value of single pollution index, and Piave is the average value of single pollutant index.

(2) Potential ecological risk index (PI)

$$ E_{i} = T_{i} \times P_{i} $$
(5)
$$ {\text{PI}} = \sum\nolimits_{i = 1}^{n} {E_{i} } $$
(6)

where Ei is monomial potential ecological risk index; Ti is the toxicity response coefficient of single heavy metal (Pb is 5 and Cd is 30) (Yu et al. 2014); Pi is single pollution index, PI is integrated potential ecological risk index of multi-metals.

3 Results and discussion

3.1 Adsorption properties

Figure S1 compares the equilibrium concentrations of Cd2+ after sorption by raw biochar and MgO-laden biochars (treated with different concentrations of MgCl2∙6H2O). The removal efficiency of Cd2+ by pristine biochar was far lower than those by the MgO-laden biochars. The equilibrium concentrations of Cd2+ in the solutions in which the MgO-laden biochars were added were below 0.03 mg L−1 when the initial Cd2+ level was 50 mg L−1. Although the pH value of MgO-laden biochar (9.74) was not significantly changed compared to that of unmodified biochar (9.30), the buffering capacity was enhanced considerably (Wang et al. 2018). Therefore, greater amounts of Cd can be removed by precipitation after modification.

The adsorption isotherm was carried out in two different ranges of initial Cd2+ concentration including 10–80 mg L−1 and 50–800 mg L−1. The biochar dose was 0.05 g L−1 and 0.5 g L−1, respectively. Figure 1 shows the Langmuir, Freundlich and Langmuir–Freundlich adsorption isotherms of Cd2+ on MgO-laden biochar and the fitting parameters are listed in Table 1. The adsorption data fitted Langmuir model (R2 = 0.928/0.973) better than Freundlich model (R2 = 0.886/0.887) for both systems, and kl = 33.419 >> 1, which suggested that surface coverage plays the dominant role in the adsorption of Cd2+. When the biochar dose increased from 0.05 to 0.5 g L−1, the maximum adsorption capacity values increased from 532.17 to 1007.7 mg g−1 estimated by Langmuir equation. It is summarized that the adsorption capacity for Cd2+ by modified biochar ranged in 22.22–322.6 mg g−1 (Wang et al. 2019), which is far lower than the data in this study. The adsorption fitted Langmuir–Freundlich model best and m = 1.711 > 1 when the biochar dose was 0.5 g L−1, suggesting that the adsorbent was heterogeneous and Cd2+ adsorption by MgO-laden biochar was diffusion-controlled at low metal concentration, while monomolecular adsorption occurred with a saturation value at high metal concentration (Mohan et al. 2011).

Fig. 1
figure 1

Langmuir, Freundlich and Langmuir–Freundlich adsorption isotherms of Cd2+ on MgO-laden biochar (a, biochar dose 0.05 g L−1; b, biochar dose 0.5 g L−1)

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Table 1 The parameters from adsorption isotherm models
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Increasing the adsorbent dose improved the maximum adsorption capacity. To explain these observations, the pH was determined after the adsorption of different concentrations of Cd2+ (Fig. 2). In the system with higher Cd2+ concentration, the pH value of the solution decreased with increasing initial Cd2+ concentration after adsorption because that an increasing amount of hydroxyl ions were combined with Cd2+. In the system with lower concentration of Cd2+, the solution pH first decreased sharply, and then decreased slightly and remained at approximately 8.1 at last. Compared to the pH value of the solution with a lower biochar dose (0.05 g L−1) and lower concentration of Cd2+, the pH value of the solution with a dose of 0.5 g L−1 and larger Cd2+ concentration was higher. A higher pH will lead to increased binding of Cd2+ due to the increase of OH− concentration. Therefore, a higher adsorption capacity was observed in the system with a higher dose of MgO-laden biochar. The OH− originated in the ionization of the Mg(OH)2 which generated when the MgO-laden biochar came in contact with water. The reaction in this process is described by the following equations:

$$ {\text{MgO}}\;{ + }\;{\text{H}}_{{2}} {\text{O}} \to {\text{Mg(OH)}}_{2} \downarrow $$
(7)
$$ {\text{Mg(OH}})_{2} \to {\text{Mg}}^{2 + } + 2{\text{OH}}^{ - } $$
(8)
Fig. 2
figure 2

pH values of the solutions after adsorption of different concentrations of Cd2+ within different ranges of concentration at two adsorbent dose levels

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Figure 3 demonstrates that the increase in the Mg2+ concentration was consistent with the decrease in the Cd2+ concentration in the solution, and the Mg2+ concentration increased with the increase with the increases in the initial concentration of Cd2+, which suggested that more Mg2+ were exchanged with the Cd2+ as the concentration of Cd2+ increased, promoting greater transfer of Cd2+ from the solution to the adsorbent. Recent historical studies have shown that Mg contents in cereal seeds have markedly declined over time, and two thirds of people surveyed in developed countries received less than their minimum daily requirement of Mg (Guo et al. 2016). Thus, the release of magnesium into the environment accompanying the capture of a toxic element will be a win–win situation. The pH decreased with increasing Cd2+ concentration (Fig. 2), suggesting an increasing consumption of OH−. Therefore, Cd2+ might be bound with OH− to generate Cd(OH)2 or Cd(OH) +. The correlation coefficient between the reduction in Cd2+ concentration and the increase in Mg2+ concentration was 0.988, indicating that the presence of Cd2+ in the solution promoted the release of Mg2+. This chemical reaction is described by Eq. 9. It means that the amount of Cd2+ that exchanged with Mg2+ should be the same in theory. The content of MgO in composite is 0.5642 mg g−1. If all the Mg exchanged with Cd2+, the amount of Cd that grabbed by the composite should be 1006 mg g−1, which was a bit lower than the experiment data, because that biochar matrix also could adsorb small amounts of Cd2+, and the other reason was that MgO and its hydration products Mg(OH)2 can also aid the removal of heavy metals through physical sorption, cation exchange, and complexation (Jin et al. 2016a). Therefore, MgO-corn straw biochar composite is rather promising used for Cd2+ removal.

$$ {\text{MgO}} + {\text{H}}_{{2}} {\text{O}}\;{ + }\;C{\text{d}}^{{2 + }} \to {\text{Mg}}^{2 + } + {\text{Cd(OH}})_{2} \downarrow $$
(9)
Fig. 3
figure 3

Molar concentration changes (a) and correlation (b) of the reduction in the Cd2+ concentration and the increase in the Mg2+ concentration in the solution after adsorbing different concentrations of Cd2+

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Adsorption kinetic models are often used to explain the adsorption process and investigate the rate control mechanism of the removal of toxic metal ions from the aqueous phase by various absorbents. First, the effect of the contact time on the removal of Cd2+ by MgO-laden corn straw biochar in solutions with the initial concentrations of Cd2+ ranging from 19 to 168 mg L−1 was studied. Figure S2 demonstrates that the sorption rate of Cd2+ increased initially, and then achieved adsorption equilibrium. More than 98% of Cd2+ in the solutions with various initial concentrations in the 19–168 mg L−1 range were removed by MgO-laden biochar. The adsorption equilibrium was reached more rapidly for lower Cd2+ level in solution. The adsorption equilibrium time increased with increasing initial Cd2+ level. For example, only 40 min were required to remove 98% of Cd2+ for the initial Cd2+ level of 19 mg L−1, while approximately 120 min were necessary for the removal of 98% of Cd2+ when the initial Cd2+ level was changed to 48 mg L−1.

The results of the fits to the pseudo-first order, pseudo-second order and intraparticle diffusion kinetic models of Cd2+ adsorption on MgO-laden biochar are presented in Fig. 4 and the corresponding parameters are listed in Table 2. At a low initial Cd2+ level (19 mg L−1), pseudo-first-order and pseudo-second-order equations could describe the adsorption well (R2 = 0.999/0.979), while the intra-particle diffusion equation failed to describe the adsorption. As the initial Cd2+ concentration increased from 48 to 168 mg g−1, the correlation coefficient of the intraparticle diffusion equation increased from 0.966 to 0.993, and the sorption rate constant changed also in a narrow range from 9.791 to 13.70 mg g−1 min−1 and the line did not pass through the origin, suggesting tha adsorption or sort of boundary layer was controlled by intraparticle diffusion at an initial concentration in a range of 48–168 mg g−1, but the adsorption was not the only rate-limiting step, and the sorption rate was limited or combined with other steps (Kumar et al. 2018). The sorption rate constant showed no dependence on the initial Cd2+ concentration. When the initial Cd2+ concentration ranged from 48 to 168 mg L−1, the pseudo-second-order model could also describe the adsorption process well (R2 = 0.957–0.994) and the sorption rate constant (k2) decreased with the increasing initial Cd2+ concentration. The metal adsorption kinetic evaluation of biomass-based sorbent mostly conformed to the pseudo-second-order model (Akpomie and Conradie 2020; Tao et al. 2019). The pseudo-second-order equation is based on the assumption that the rate-limiting process is chemisorption or physicochemical adsorption including electron transfer and exchange or bond formation (Li 2019). Therefore, the adsorption of Cd2+ onto the MgO-laden biochar is due to either chemisorption or physicochemical adsorption.

Fig. 4
figure 4

a Sorption kinetics of Cd2+ at different concentrations on MgO-laden biochar. b Cd2+ sorption data fitted by the intraparticle diffusion model (biochar dose was 0.5 g L−1)

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Table 2 The parameters from adsorption kinetics models
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3.2 Effect of initial pH and coexisting Pb2+ on Cd2+ removal from aqueous solutions by MgO-laden biochar

The effect of pH on the Cd2+ adsorption onto MgO-laden biochar was observed in the pH range of 2.0–8.0 (Fig. 5). Various initial pH values of solutions with and without Cd2+ were adjusted by HCl (0.1 M) and NaOH (0.1 M) solutions at the beginning of the adsorption. After shaking for 24 h, the pH of the solution with and without Cd2+ (only with adsorbent) showed a pronounced increase. However, compared to the solutions without Cd2+, the pH values of the solution with Cd2+ were lower, suggesting that the pH decreased after Cd2+ adsorption. When the initial pH was 2.0, the pH of the solution without Cd2+ changed to 4.90 after shaking for 24 h and changed to 2.92 after the adsorption of Cd2+, and the removal rate of Cd2+ was rather low. However, the removal rate reached 81.5% when the initial pH was 2.2. After adsorption, the concentration of Cd2+ in solutions ranged from 0.03 to 0.06 mg L−1 when the initial pH was in the 3.0–7.0 range. To summarize, the Cd2+ removal by MgO-laden biochar was pH dependent, and the adsorbent changed the solutions pH by the hydrolyzation reaction of MgO and kept the solution alkaline, improving the precipitation of Cd2+.

Fig. 5
figure 5

The effect of initial pH on the Cd2+ removal by MgO-laden biochar from aqueous solutions (initial concentration of Cd2+ was 92 mg L−1; biochar dose was 0.5 g L−1)

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In wastewater, Cd2+ always co-exists with other metal ions. Compared to other metal ions, Pb2+ is more likely to affect the removal of Cd2+ due to its smaller hydrated radius (4.01 Å) and solubility product constant (Ksp) that will give rise to a greater affinity for most functional groups and OH− (Park et al. 2016). Here, we assumed that the concentrations of Pb2+ and Cd2+ in the solution system were the same. In fact, the situation is more complicated in industrial wastewater. Figure 6 shows the adsorption kinetics (a) and isotherms (b) of mixed ions. The adsorption equilibrium time of Cd2+ was delayed from 360 to 480 min, and the maximum adsorption capacity was decreased. The Cd2+ adsorption isotherm was best fitted by the Langmuir model (R2 = 0.947), and the maximum Cd2+ adsorption capacity of 401.85 mg g−1 was estimated by the Langmuir model. Therefore, it is concluded that more adsorbent is required in the presence of Pb2+ coexisting in the system.

Fig. 6
figure 6

Kinetics (a) and isotherms (b) of mixed ions adsorption on MgO-laden biochar (biochar dose, 0.5 g L−1; concentrations of Cd2+ and Pb2+ were 80 mg L−1)

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3.3 Changes of MgO-laden biochar characteristics before and after Cd2+ adsorption

To investigate the mechanism of Cd2+ removal by MgO-laden biochar, SEM–EDS, XRD and FTIR were used to detect the characteristics of pristine biochar, modified biochar and modified biochar after Cd2+ sorption.

Table S1 lists the pH, ultimate composition and pore characteristics of corn straw biochar and modified biochar. After modification, the biochar pH, ash and oxygen contents increased but the special surface area (SSA) calculated by the BET method decreased from 135.91 to 38.32 m2 g−1 and the pore width changed strongly. Previous study also found that surface areas of MgO-coated magnetic sugarcane harvest residue biochars decreased from 118 to 27.2 m2 g−1 with Mg content increased from 2.12% to 20.47% (Shen et al. 2019). The average pore width increased from 1.9 to 5.6 nm. Cai et al. (2017) reported that highly active and commercial MgO particles have the BET surface areas of 47.85 and 5.26 m2 g−1, and average pore diameters of 12.15 and 30.36 nm, respectively. SEM images showed that different size of particles distributed on the biochar surface and the pores of the biochar were filled by the particles (Fig. 7), which was relative to the ultrasonic treatment. It was reported that the MgCl2 with the ultrasonic impregnation method may produce large aggregates during pyrolysis, which may result in the heterogeneity of Mg-containing minerals on biochar and the blockage of biochar pores after synthesis (Zhang et al. 2020). That is why the specific surface area decreased after Mg impregnation.

Fig. 7
figure 7

SEM micrographs of a corn straw biochar, b MgO-laden corn straw biochar and c MgO-laden corn straw biochar after sorption of Cd2+ at different magnifications

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Figure 7 shows that the surface of the raw biochar was smooth with many pores and few particles that were identified as KCl crystals by XRD (Fig. 9). After modification using MgCl2·6H2O, the surface of the biochar became rough and heterogeneous, and the nanoscale and microsized particles were filled in the pores and on the surface of the biochar, so that the specific surface area decreased from 135.91 to 38.32 m2 g−1; however, the adsorption capacity was enhanced. Therefore, the physical adsorption by pores does not dominate the sorption process. After adsorption, the concentration of Cd2+ in the solution decreased drastically, and many spindly particles were generated, suggesting that new Cd-containing particles were generated, achieving the transfer of Cd from solution to the adsorbent.

The elementary composition (Fig. 8 and Table S2) of the particles on the surface of the modified biochar and postsorption biochar was analyzed use a SEM–EDX. After Cd2+ adsorption, the atom fraction of magnesium decreased from 31.09% to 3.46% and the atom fraction of oxygen only decreased slightly, but the atom fraction of cadmium increased from 0 to 35.03%, verifying the transfer of Mg from the adsorbent to the solution and the transfer of Cd from the solution to the adsorbent. In addition, there were almost no changes in the atom fractions of carbon and chlorine. Therefore, the most likely reactions were the hydrolyzation of MgO particles on the surface of biochar and then Mg2+ and OH− were released to the solution due to the ionization of Mg(OH)2. The released OH− and Cd2+ in the solution finally combined to generate Cd(OH)2, because that the Ksp of Cd(OH)2 is smaller than that of Mg(OH)2; this is consistent with the results of the adsorption tests. On the other hand, the generation of Cd(OH)2 promoted the ionization of Mg(OH)2.

Fig. 8
figure 8

SEM–EDX of a MgO-laden biochar, b postadsorbed biochar

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The result of XRD patterns of biochars (Fig. 9) suggested that only KCl crystal was detected in the corn straw biochar. After the modification, the MgO crystal generated, which was consistent with the SEM–EDX result. Therefore, the solid particles on the surface of the biochar (SEM) were MgO. After sorption of Cd2+, the intensity of MgO decreased from approximately 1700 to 350 and Mg (OH)2 crystals were generated, suggesting the hydrolyzation of MgO. Although Cd was detected in the biochar surface by SEM–EDX, no clear peaks involving Cd-related crystal was observed. It is possible that the structures of the Cd (OH)2 particles adhering to the biochar surface or in the pores were amorphous. The lack of Cd-related crystal may be due to the destruction of the Cd(OH)2 crystals by the biochar matrix, because for the pure MgO nanoparticles, Cd(OH)2 and CdO crystals can be identified by XRD after Cd2+ sorption (Cai et al. 2017).

Fig. 9
figure 9

XRD patterns of biochars

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The FTIR spectra of corn straw biochar, and MgO-laden biochar before and after Cd2+ adsorption are shown in Fig. 10. For corn straw biochar, the peaks at 3398–3476 cm−1 (H-bonded O–H stretching vibrations of hydroxyl groups of phenols, alcohols, and organic acids) (Keiluweit et al. 2010) were observed. The peaks at 1585 cm−1 and 1099 cm−1 were assigned as C=C/C=O stretching vibrations and C–O–C vibrations (Chi et al. 2017; Wang et al. 2016a). The peaks at 740–885 cm−1 were frequently observed from corn straw biochar and other biochars and are due to the out-of-plane deformation of aromatic C–H (Wang and Liu 2017).

Fig. 10
figure 10

FTIR spectra of biochars

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After modification, the peaks at 3409 cm−1 changed to 3398 cm−1 and the peak intensity increased; however, the peak intensity decreased after adsorption of Cd2+, suggesting that O–H functional groups were bound with Cd2+. Moreover, a new peak at 3237 cm−1 corresponding to free OH− that may be generated by the hydrolyzation of MgO appeared after the modification, and it disappeared after the sorption of Cd2+. In addition, the modified biochar showed a pronounced band at 1635 cm−1 due to the protein carboxyl C=O groups (Liu et al. 2015b), and the position of the peak shifted to 1594 cm−1 after sorption, indicating that the carboxyl on the surface of the biochar was complexed with Cd2+. The coverage of micron- and nano-sized MgO particles gave rise to the disappearance of the peak at 1099 cm−1 after the modification and the peak appeared after Cd2+ sorption; however, the peak position changed from 1099 cm−1 to 1052 cm−1, because the hydrolyzation reaction of MgO, the ionization reaction of Mg(OH)2 and the precipitation reaction between Cd2+ and hydroxyl led to the exposure of the C–O–C functional groups, which then complexed with the remaining Cd2+.

3.4 Effect of different biochars’ addition on toxic element activity in soil and potential ecological risk assessment based on TCLP leaching

The paddy soil properties which are listed in Table S3 have been described in previous research (Li et al. 2020). The total contents of Cd and Pb in the paddy soil severely exceeded the class II standard for agriculture land of China (NEPA 2018). The effect of corn straw biochar and MgO-laden biochar on TCLP-leaching (a) lead and (b) cadmium is shown in Fig. 11. The results showed that adding corn straw biochar and MgO-laden biochar decreased the concentrations of TCLP-leaching Cd and Pb. TCLP-leaching Cd in soils only reduced by 0.6%, and TCLP-leaching Pb reduced 19.3% after treated with 4% corn straw biochar, because that the active site on the biochar surface was occupied by Pb2+ which had a smaller hydrated ionic radius (Park et al. 2016) and a higher concentration in this soil. MgO-laden biochar presented more effective in Cd and Pb immobilization than pristine biochar, which might be because that MgO-laden biochar increased soil pH more obviously. The soil pH was 6.2 and 7.7 after treated with 4% of biochar and 1% MgO-laden biochar, respectively. And Cd and Pb are easily sequestrated at a higher pH value. Therefore, MgO-laden biochar was more effective to sequestrate Cd2+ and Pb2+. The concentrations of TCLP-leaching Cd and Pb decreased with the addition rate of MgO-laden biochar increasing. Similarly, the Pb2+ was easier to be sequestrated by MgO-laden biochar than Cd2+ due to the smaller solubility product constant of Pb(OH)2, which was consistent with the result of the adsorption test in aqueous solution. Adding 1% of MgO-laden biochar only reduced 7.0% of TCLP-leaching Cd, but TCLP-leaching Pb reduced 21.8%. Adding 4% of MgO-laden biochar decreased the concentrations of TCLP-leaching Cd and Pb by 22.4% and 29.0%, respectively. Previous study also reported that MgO-coated biochar could reduce TCLP-leached Pb (Shen et al. 2019).

Fig. 11
figure 11

Effect of corn straw biochar and MgO-laden biochar on TCLP-leaching a lead and b cadmium

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Potential ecological risk was assessed by Nemerow pollution index (NPI) and Potential ecological risk index (RI). Table 3 lists the calculation results of Nemerrow pollution index and Potential ecological risk index. Table S4 lists the classification of index. The results showed that biochar or MgO-laden biochar addition decreased the Pi (Cd) and Pi (Pb), and single pollution index of Cd decreased from 1.16 to 0.9 after 4% MgO-laden biochar amendment, which suggested that the Cd pollution rank decreased from low polluted to warning limit level. The values of NPI reflected the integrated pollution status of the total heavy metals. The NPI decreased by 28.9% after 4% of MgO-laden biochar application.

Table 3 Nemerrow pollution index and Potential ecological risk index of heavy metals in soil (CSB means ‘corn straw biochar’, MgO-CSB means ‘MgO-laden corn straw biochar’)
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The potential ecological risk index was often used to assess the potential ecological risk of single and multi-pollutants in the environment. The results showed the order of Ei value was Pb > Cd, and the ecological risk of Cd was low (Ei (Cd) < 40). The value of Ei (Pb) was between 300 and 600, suggested that the ecological risk of Pb was high at this soil sample collection site. Both corn straw biochar and MgO-laden biochar addition decreased the Ei (Cd), Ei (Pb) and the comprehensive pollution status (PI) of the two toxic metals. However, MgO-laden biochar presented more efficient in Cd and Pb immobilization than unmodified biochar. Adding 4% of MgO-laden biochar, the value of Ei (Pb) decreased from 434.94 to 309.01, which means that the risk of Pb lowered by one rank, and the PI also decreased from 469.77 to 336.07.

4 Conclusions

In summary, MgO loading enhanced Cd removal in aqueous solution by precipitation with OH− that was released via the ionization of Mg (OH)2 derived from the MgO hydrolysis on the surface of biochar. In addition to precipitation, the adsorption also involved other mechanisms such as complexation with oxygen-containing functional groups and electrostatic adsorption. The maximum adsorption capacity of Cd2+ was 1058.8 mg g−1, but a larger dose is required when there are other coexisting toxic metals like Pb2+ in the system. Adding MgO-laden corn straw biochar into contaminated soil could enhance the immobilization and decrease the potential ecological risks of Cd and Pb assessed by NPI. Considering its advantages of low cost, high adsorption capacity and environmental friendliness, MgO-laden biochar is promising for Cd2+and Pb2+ immobilization in wastewater and contaminated soil.