Introduction

The toxicity of HMs and their accumulation due to the presence in food chain is one of the most important environmental and health problems in modern societies. The biological degradation of the HMs is not possible in the soil, making them one of the most dangerous environmental pollutants, and their removal is considered a serious problem (Tembo et al. 2006; Rashed 2010). Lead is also one of the HMs with no biological function and the potential to produce the toxicity for plants and other organisms. According to the total Pb concentration in soil reported by World Health Organization (WHO), the amount of 35 mg kg−1 is taken into account as the toxic concentration of Pb (Tembo et al. 2006). The availability and harmful effects of the HMs for animals, plants, and microorganisms, their movement towards the underground water, as well as the effects of these metals on the soil chemistry processes depend on the reactions occurring between these ions and the soil particles (Murali and Aylmore 1983). In other words, three processes can control the fate and availability of the HMs in soil. These processes include (1) the removal of the HMs from the soil solution due to the adsorption on the surface of soil particles, (2) desorption of metals from particles to solution, and (3) the dissolution and precipitation of the HMs in the solid soil phase. The nature of the adsorption and desorption processes influence the solubility and availability of the HMs in the soil (Sparks 2003). Adsorption characteristics are influenced by the solubility and availability of HMs, pH, redox, soil organic matter, iron oxides, and calcium carbonate (Antoniadis et al. 2007).

Desorption of the HMs in the soil is subject to three factors, soil properties, HMs properties, and extraction conditions. The soil properties influencing the adsorption and mobility of the HMs include soil pH, texture, cation exchange capacity (CEC), organic matter (OC), Fe and Ca, and pollutants presence in the soil. HMs properties and extraction conditions include the type and concentration of the HMs, distribution of the HMs in soil components and availability of the HMs, solution pH, electrolyte presence, solution ratio, and duration of contact (Zhu et al. 2010).

In the contaminated soils, the HMs usually exist at the same time and compete with each other for adsorbing sites. Therefore, selective adsorption and competition of the HMs by soil have high importance in determining the toxicity potential of these metals in soil, mobility, and their fate (Serrano et al. 2005; Jalali and Moharammi 2007). Individual adsorption of the HMs in soils and different minerals has been widely investigated (Trivedi and Axe 2001; Vasudevan et al. 2002). Also, numerous studies have also been done to understand the competitive adsorption of the HMs in the pure minerals, organic compost, and acidic soils (Zhu et al. 2012; Zhang et al. 2012; Sheikhhosseini et al. 2013). The results of these studies showed that the main factor influencing the competitive adsorption is not only the concentration of the HMs in the soil, but also related to the soil properties, HMs type, and environmental factors (Veeresh et al. 2003).

Due to the spread of the soil contaminated with the HMs, there is a need to develop the soil remediation techniques. These techniques must be cost-effective and reduce the pollution without influencing the soil fertility. For this purpose, studies have been done to reduce the mobility and availability of the HMs (Zhang et al. 2012; Xu et al. 2013; Melo et al. 2016). Nowadays, the biochar is widely used for remediation of the polluted soils (Zhang et al. 2013; Melo et al. 2013, 2016). Biochar is an organic matter obtained from the pyrolysis of the organic compounds. Pyrolysis of the organic matter such as plant residue, forest, and livestock manure causes the release of the volatile C compounds, fixed C compounds, and ash containing a significant amount of Ca and K (Ahmad et al. 2014). Pyrolysis temperature plays an important role in Pb sorption capacity of biochars (Ahmad et al. 2012; Ding et al. 2014). The biochar has a high surface charge, high specific area, and high stability against degradation, and thus has significant power on ion adsorption compared with other organic matter. The biochar may influence the toxicity, mobility, and fate of the various HMs in the soil as a result of improving the adsorption capacity of the soil (Ahmad et al. 2014). Several studies have shown that applying the biochar (produced from different materials or temperatures) in the contaminated soils can significantly reduce the availability of the HMs through the cationic exchange, precipitation, and pH increase (in acid soils) (Zhang et al. 2012; Xu et al. 2013; Melo et al. 2016). However, there is limited evidence regarding the effects of biochar on the mobility of the HMs in calcareous soils. Moreover, Zn is present in many soils of Pb mines (Dayani and Mohammadi 2010); therefore, it is necessary to consider the effect of Zn on the mobility of Pb in the soils.

In this study, the effects of biochar produced at different temperatures were investigated on mobility (adsorption and desorption) of Pb2+ during incubation period. According to the mentioned explanations, the following hypotheses are put forth: (1) desorption and adsorption of Pb2+ is influenced by the presence of Zn2+, (2) biochar produced at different temperatures influences the desorption and adsorption of Pb2+, and (3) mobility of Pb2+ varies during the incubation times.

Materials and methods

Properties of the studied soil

The soil sample was taken from 0 to 30 cm depth from surface of the calcareous soils with semi-arid climate in Central Iran. This soil was classified as Typic Calcixerepts according to the USDA Soil Taxonomy (Soil Survey Staff 1990). The selected sandy soil was air-dried and ground to pass through 2-mm sieve. At the beginning of study, soil subsamples were used to determine some chemical and physical properties. The values of these measurements were as follows: clay 170 g kg−1, silt 300 g kg−1, sand 530 g kg−1, carbonate calcium equivalent (CCE) 150 g kg−1, electrical conductivity (EC, soil:water = 1:2) 0.28 dS m−1, pH (soil:water = 1:2) 7.9, organic carbon (OC) 8.6 g kg−1, available Pb 0.01 mg kg−1, and available Zn 0.36 mg kg−1. The predominant clay minerals in the soil of the present study were micas, smectites, and chlorite, with fewer quantities of kaolinite and vermiculites (Hosseinpur et al. 2012).

Biochar production and properties

The biochar was produced from walnut leaves (Juglans regia L.). Walnut leaves (WL) were oven-dried at 60 °C for 24 h and were placed in the cylindrical pyrolyzers of 608 cm3 inside a muffle furnace, and were sealed and warmed up at 200 (B200), 400 (B400), and 600 (B600) °C at a rate of 10 °C/min, using a batch pyrolysis facility (Khadem and Raiesi 2017). The slow pyrolysis process was applied, the favorable temperature was kept for 2 h, and the temperature of the pyrolyzed feedstock gradually came down to the room temperature. The biochar samples were ground and passed through 1-mm sieve. Some chemical properties of biochars are presented in Table 1.

Table 1 Properties of used biochars in this study
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To study the effect of applying 1% (weight of amendment/weight of soil) amendment on the individual and competitive adsorption and desorption of Pb2+ in the sandy calcareous soils during incubation, a completely randomized design with a 5 × 2 factorial treatment combination was used in three replicates with the following factors: (1) amendments (control, WL, B200, B400, and B600), (2) adsorption system (individual and competitive) during incubation time (30 and 90 days). Therefore, 300 g of soil was put in each jar and mixed with the amendments, then incubated for 90 days at 25 ± 2 °C (Mirzaei Aminiyan et al. 2014; Abbruzzini et al. 2017). To determine the adsorption and desorption of Pb2+, the soils were sampled on 30 and 90 days after incubation. The moisture of all soils was kept at field capacity by weighing during incubation period.

Characteristics of the individual and competitive adsorption

The competitive adsorption of Pb2+ in the presence of Zn2+ was carried out in the amended soils using batch method (Antoniadis et al. 2007; Jalali and Moharammi 2007). In individual system, subsample (2 g) of the amended soils was placed into 50 ml centrifuge tubes, and 20 ml of equilibrating solution (0.05, 0.1, 0.2, 0.4, 0.5, 1, 2, and 3 mM) of Pb2+ (as Pb(NO3)2) containing 10 mM CaCl2 as the background electrolyte was added to each tube. In competitive system, the amount of Pb2+ and Zn2+ (as Zn (NO3)2) was similar to the individual system (mole ratio Pb2+/Zn2+ = 1 to 1). The suspensions were shaken at 200 rpm for 2 h and then kept still for 24 h at a constant temperature of 25 ± 2 °C. At the end of the adsorption period, the suspensions were centrifuged at 4000 revolutions min−1 for 3 min, and the supernatants were filtrated to measure Pb2+ concentrations. Then, concentration of Pb2+ was measured using an atomic absorption spectrophotometer (AAS model G.B.C 932, G.B.C. Melbourne, Australia).

The concentration of adsorbed Pb2+ (q, mg kg−1) was calculated as

$$q = \frac{{(C_{{\text{i}}} - { }C_{{\text{e}}} ){ } \times {\text{ V}}}}{{\text{W}}}$$
(1)

where Ci and Ce represent the initial and equilibrium concentrations (mg l−1) of Pb2+ in the solution, respectively. The v/w represents the volume of the solution (l) to mass of the sorbent (kg), respectively (Echeverria et al. 1998; Jalali and Moharammi 2007; Mohan et al. 2007).

Adsorption isotherms were obtained by plotting the amount of adsorbed Pb2+ on the amended soils versus the concentration of Pb2+ in equilibrium solutions. Then, the linear Langmuir (2), and Freundlich (3) (Echeverria et al. 1998; Sparks 2003; Jalali and Moharammi 2007; Hararah et al. 2012; Yang et al. 2019) were used to describe the distribution of Pb2+ between solution and solid phases of the soils.

$$\frac{{C_{{\text{e}}} }}{q} = \frac{1}{{{\text{K}}_{{\text{L}}} \,{\text{qm}}}} + \frac{1 }{{{\text{qm}} }}C_{{\text{e}}}$$
(2)
$$\log q = \log {\text{K}}_{{\text{f}}} \times \,n\log C_{{\text{e}}}$$
(3)

where q is the amount of adsorbed Pb2+ at equilibrium (mg kg−1); Ce is the concentration of Pb2+ in solution at equilibrium (mg l−1); and qm is the maximum adsorption capacity of Pb2+ on soil (mg kg−1) (Echeverria et al. 1998; Sui and Thompson 2000; Jalali and Moharammi 2007; Hararah et al. 2012; Yang et al. 2019). The KL is the Langmuir constant (l mg−1), which increases exponentially with the energy of sorption (energy constant related to the strength of adsorption), Kf is the Freundlich constant (l kg−1) reflecting the adsorption capacity or distribution coefficient, and n is an empirical constant (unit less) (Vega et al. 2006; Jalali and Moharammi 2007; Mohan et al. 2007; Hararah et al. 2012; Yang et al. 2019). Maximum buffering capacity (MBC) (4) indicates the ability of the soil to resist change of Pb2+ concentration in soil solution (Sui and Thompson 2000; Yang et al. 2019). The MBC is an integrated parameter that combines the qm and KL of Langmuir isotherm (Yang et al. 2019). Yang et al. (2019) reported that higher amount of MBC reveals more elements will be adsorbed onto soil.

$${\text{MBC}} = {\text{qm}} \times {\text{K}}_{{\text{L}}}$$
(4)

Desorption of Pb2+

To determine the desorption of adsorbed Pb2+, 20 ml of 10 mM CaCl2 or DTPA–TEA (0.005 M DTPA + 0.1 M triethanolamine + 0.01 M CaCl2, pH 7.3) was added to the residual soils in centrifuge tubes from the adsorption study, and their contents were shaken for 24 and 2 h, at a constant temperature of 25 ± 2 °C (Wang and Harrel 2005; Hararah et al. 2012). At the end of desorption period, the suspensions were centrifuged at 4000 r min−1 for 3 min, and the supernatants were filtrated to measure Pb2+ concentrations. Then, the concentration of Pb2+ was measured using AAS (above model). Percentage of the desorbed Pb2+ was calculated as

$${\text{Desorbed}}\,{\text{Pb}}^{{{2} + }} \left( \% \right) = \frac{{{\text{Pb}}_{{{\text{de}}}} }}{{{\text{Pb}}_{{{\text{ad}}}} }} \times 100$$
(5)

where Pbde is the concentration of desorbed Pb2+ (mg kg−1) and Pbad the concentration of adsorbed Pb2+ (mg kg−1).

Statistical analysis

Linear regression analysis was used to fit the Langmuir and Freundlich isotherms to the adsorption data using the Excel software (Jalali and Moharammi 2007; Remenyi et al. 2009). The goodness of fit of these models to the data was investigated based on the coefficients of determination (R2).

A repeated measures ANOVA was performed to analyze the effects of the independent factors (treatments) and incubation times on the coefficients of isotherms. Before the ANOVA analysis, Mauchly’s sphericity test was performed. When the test value was significant at 5%, the degree of freedom (df) was multiplied by the epsilon coefficient (Huynh–Feldt correction). The means of the treatments were separated at 5% of significance level using the least significant difference (LSD). Simple linear regression was performed between all the attributes. The statistical analyses of the data were carried out using STATISTICA 8 (StatSoft, Inc. 2007; Moghimi et al. 2018).

Results and discussion

Adsorption characteristics

The relation between the equilibrated concentration of Pb2+ in solution (Ceq) and the amount of adsorbed Pb2+ (q) onto the treated soils is shown in Fig. 1. All relations between Ceq and q are L shape (Sposito 1989; Limousin et al. 2007; Park et al. 2016). This figure showed that concentration of the adsorbed Pb2+ onto the soils treated with biochars was higher than the soil treated with the feedstock and control soil. The amount of adsorbed Pb2+ was also found to be increased in the soil treated with B600 compared with the soils treated with B400 and B200. Applying the biochar to the soil enhanced the concentration of equilibrium Pb2+ in the soil solution through the increase in the pyrolysis temperature.

Fig. 1
figure 1

Relation between equilibrated concentration of Pb in solution and amount of adsorbed Pb onto soils. B0 is feedstock; b200, b400, and b600 are biochars produced at 200, 400, and 600 °C, respectively

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Coefficients of determination (R2) for two models are shown in Table 2. The results of this table revealed that Langmuir and Freundlich isotherms could describe the adsorption of Pb2+ onto all the studied soils. However, the Langmuir model (R2 = 0.983–0.995) fitted slightly better than the Freundlich model (R2 = 0.915–0.977). Trakal et al. (2011), Ding et al. (2014), and Wang et al. (2015) obtained a similar result for Pb2+ adsorption by different biochars.

Table 2 Coefficient of determination (R2) of two models applied to describe Pb2+ adsorption onto different soils
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Isotherm coefficients are useful parameters to compare the capacity of different soils regarding the adsorption of the HMs under the same experimental conditions (including HMs concentration, equilibrium time, temperature, etc.) (Jalali and Moharrami 2007; Trakal et al. 2011). Summary of analysis of variance (ANOVA) results for isotherm models is shown in Table 3. ANOVA showed a significant interaction effect of the biochar, adsorption, and time (p < 0.01) on all coefficients of isotherms except qm in the Langmuir model. Main effects (biochar, adsorption, and time) were only significant on qm.

Table 3 Summary of analysis of variance (ANOVA) results (mean square values) for isotherm models
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The qm (maximum adsorption capacity of Pb2+) augmented in all the amended soils compared with control (Fig. 2A). The qm value in the soils treated with B400 (443–627 mg kg−1) and B600 (814–998 mg kg−1) increased (p < 0.05) compared with the soils treated with the feedstock and B200. Therefore, the adsorption ability of Pb2+ enhanced with the increase in the pyrolysis temperature. Moreover, maximum adsorption capacity of Pb2+ in the soil treated with B200 did not improve (p > 0.05) compared with the soil treated with the feedstock (Fig. 2A). The qm value was higher in the individual system (93 mg kg−1) than the competitive system (Fig. 2A). Therefore, it can be concluded that the maximum adsorption capacity of Pb2+ in the soil will decrease in the presence of Zn. In agreement with this result, Echeverria et al. (1998) demonstrated that the qm value was greater in the individual than competitive systems. The averaged qm value increased (85 mg kg−1) on 90 days compared with 30 days of the incubation. There is limited evidence regarding the effects of biochar on Pb2+ adsorption in calcareous soils. Trakal et al. (2011) reported that the qm value increased in an acidic soil treated with 1% willow biochar (41.4 mg kg−1) compared with control (39.4 mg kg−1). In some studies, Pb2+ adsorption by biochars was determined in aqueous solutions (Mohan et al. 2007; Lu et al. 2012). Mohan et al. (2007) reported that adsorption capacity of Pb2+ from solution by biochars produced at 400–450 °C ranged from 2620 to 1310 mg kg−1. Also, Lu et al. (2012) reported that adsorption capacity of Pb2+ from aqueous solutions by sludge biochar (produced at 550 °C) was 30,900 mg kg−1. Therefore, the qm value in our study was lower than these studies. It might be related to the experimental condition (such as absorbent, soil type, concentration of HMs, equilibrium time, etc.).

Fig. 2
figure 2

Effect of biochar (A) and/or in competitive or individual system (B) at two incubation times (C) on the maximum adsorption capacity of Pb2+ on soils. b0 is feedstock; b200, b400, and b600 are biochars produced at 200, 400, and 600 °C, respectively. Values are mean and bars indicate SE

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The interaction effects of the biochar, adsorption type, and time on KL (strength of adsorption or energy of sorption) are presented in Table 4. The results showed that energy of sorption was lower (p < 0.05) in the competitive adsorption than the individual adsorption of Pb2+ in all treatments. Park et al. (2016) obtained a similar result for Pb2+ adsorption by different biochars. They reported that the strength of adsorption of HMs from solution by pepper stem biochar ranged from 0.0481 to 0.1267 in individual system and from 0.0413 to 0.0685 in competitive system. In both of the individual and competitive systems, KL increased with the increase in the pyrolysis temperature. This coefficient was higher in the soil treated with B600 than other soils. In agreement with this result, Wang et al. (2015) investigated the effects of biochars produced from peanut shell and Chinese medicine material residues at 300 to 600 °C on Pb2+ adsorption from solution. Their results showed that the KL value increased with the increase in the pyrolysis temperature. Ding et al. (2014) reported that the strengths of adsorption of Pb2+ from solution by bagasse biochars produced at low temperatures (0.051–0.055) were lower than biochar produced at high temperatures (0.11–0.18). In the competitive system, the KL value was not different in the soils treated with B0, B200, and B400 at both incubation times. It might be attributed to surface functional groups, pore size distribution, and concentration of PO43− and CO32− in these amendments (Cheng and Lehmann 2009; Ding et al. 2014; Cao et al. 2017). However, this result deserves further study. Strength of Pb2+ adsorption enhanced on 90 days compared with 30 days of incubation in the soils treated with B400 and B600. Cheng and Lehmann (2009) and Cao et al. (2017) reported that acidic surface functional groups (carboxylic and phenolic structures) developed in biochar aging. Accordingly, negative charge (cation exchange capacity) of biochars could be increased during incubation period. Therefore, total adsorbed Pb2+ onto the soils increased as a result of the soils’ biochar aging.

Table 4 The interactions effects of biochar, adsorption, and time on KL and MBC of Langmuir
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The interactions effects of the biochar, adsorption, and time on the MBC are shown in Table 4. Similar to qm and KL trends, the MBC increased in the soils treated with the biochars compared with control and the soil treated with the feedstock. The MBC was lower (p < 0.05) in the competitive system than the individual system in all treatments. In the individual and competitive systems, the MBC increased in the soil treated with B600 compared with the soil treated with B400. In the individual system, the MBC was lower (p < 0.05) on 30 days of incubation than 90 days of incubation in the soils treated with B400 and B600. The MBC value did not differ in the control soil on 90 days of incubation compared with 30 days of incubation. The soils’ resistance to the change of Pb2+ concentration in the soil solution against adding Pb2+ to the soil has been shown previously by the maximum buffering capacity (Sui and Thompson 2000; Yang et al. 2019). Soil with high MBC (soil treated with B600) is capable to adsorb Pb2+ more than the soil with low MBC. Therefore, this amendment can improve the contaminated soils in the presence of Pb2+ + Zn2+ or Pb2+ alone.

The interaction effects of the biochar, adsorption, and time on Kf (adsorption capacity) are presented in Table 5. In the individual system, the adsorption capacity of Pb2+ increased on 90 days of incubation in the soils treated with B400 and B600. In the competitive system, the adsorption capacity did not change in the soils treated with the biochars (p > 0.05). The Kf decreased (40–61%) in the competitive system compared with the individual system. Park et al. (2016) obtained a similar result for Pb2+ adsorption by pepper stem biochar from solutions. The Kf value increased in the soil treated with B600 than other treatments (22–88%). The trend of Kf was similar to the qm value in the studied soils (r = 0.89, p < 0.01).

Table 5 The interactions effects of biochar, adsorption, and time on constants of Freundlich model
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The interaction effects of the biochar, adsorption, and time on n are given in Table 5. The n coefficient did not change in the soils treated with the biochars.

Individual and competitive desorption

Percentages of Pb2+ desorbed in DTPA–TEA and 10 mM CaCl2 solutions are shown in Table 6. Lead extracted by 10 mM CaCl2 was found to be higher in the competitive system at two incubation times than the individual system. Also, Pb2+ desorbed in 10 mM CaCl2 was higher on 90 days of incubation than 30 days of incubation. The results of this table showed that Pb2+ extracted by 10 mM CaCl2 was lower than 1% of adsorbed Pb2+. Therefore, concentration of the adsorbed Pb2+ was very little on the exchangeable sites in all treatments. The concentration of the lead extracted by 10 mM CaCl2 decreased with the increase in the pyrolysis temperature.

Table 6 Percentage of Pb desorbed in 10 mM CaCl2 and DTPA–TEA solutions
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The amount of Pb extracted by DTPA–TEA was higher than 40% of the adsorbed Pb. However, extractants are not able to extract all amounts of the adsorbed Pb2+. Desorbed Pb by DTPA–TEA in the presence of Zn2+ was higher at two incubation times. Also, Pb2+ extracted by DTPA–TEA was higher (more than 20%) in the competitive system than the individual system. Moreover, the amount of Pb2+ extracted by this extractant decreased with the increase in the pyrolysis temperature.

The DTPA–TEA extractable Pb2+ is the amount of Pb2+ adsorbed in the labile pool of the soils, whereas the DTPA–TEA unextractable Pb is the amount of Pb2+ adsorbed in the nonlabile pool of the soil (Wang and Harrel 2005). Therefore, it can be said that the concentration of Pb2+ adsorbed in the nonlabile pool of the soil will reduce in the presence of Zn2+ on 90 days of incubation.

It seems that the adsorption on the exchangeable sites is not the main mechanism controlling the adsorption of Pb, and the complexation and precipitation are also important mechanisms controlling Pb sorption by the soil treated with the biochar (Lu et al. 2012; Cao et al. 2009; Xu et al. 2013; Ding et al. 2014).

The results of this study indicated that the affinity of soil for sorption of Pb2+ is higher than Zn2+ in all treatments and incubation times, which has also been reported in many studies (Covelo et al. 2004; Vega et al. 2006). The tendency of the HMs cations to form the strong complexes is according to the ionic radius and the ionization potential (Misono softness parameter) (Sposito 1989). Moreover, electronegativity is an important factor in determining the ability of these cations’ chemisorption. Ionic radius of Pb2+ (1.20 \(\dot{A}\)) is greater than Zn2+ (0.74 \(\dot{A}\)) and subsequently, hydrated radius of Pb2+ is smaller than Zn2+. Also, the sequence of electronegativity is Cu2+ (1.9) > Pb2+ (1.8) > Ni2+ (1.8) > Cd2+ (1.7) > Zn2+ (1.6). Therefore, Pb2+ is a suitable metal than Zn2+ for electrostatic adsorption and inner-sphere surface complexation on the soils (McBride 1994). The results of Table 6 indicate the highest amount of added Pb2+ adsorbed as inner-sphere. Meanwhile, Fontes et al. (2000) reported that competition between Cd2+, Cu2+, Pb2+, and Zn2+ had a very small effect on Pb2+ and Cu2+ adsorption on an Oxisol.

Cation exchange, complexation, and precipitation are the main mechanisms controlling Pb2+ sorption through the use of the biochar (Lu et al. 2012). Biochar generally has a higher surface area and greater cation exchange capacity than soil; it has been used in trials aimed at decreasing the solubility and toxicity of the HMs and organic compounds in the soils (Beesley et al. 2011). However, their sorption mechanisms depend on the characteristics of the biochar, influenced by the feedstock and pyrolysis temperature (Li et al. 2017). Melo et al. (2013) reported that sorption capacity of Zn2+ and Cd2+ onto the sugarcane straw biochar produced at 700 °C was higher (four times) than that of the sugarcane straw biochar pyrolyzed at 400 °C. Biochars pyrolyzed at high temperature have higher stable carbon compared with the biochars produced at low temperature (Melo et al. 2016). Therefore, biochars pyrolyzed at high temperature are able to keep the HMs longer than the biochars produced at low temperature (Melo et al. 2016) because these amendments are stable. Lu et al. (2012) studied the Pb sorption capacity of the sludge biochar from acid solution. They reported that 38.2 to 42.3% of the total adsorbed Pb2+ sorbed on the organic hydroxyl and carboxyl functional groups at different pH (pH 2 to 5). Moreover, they also argued that the coprecipitation and complexation of Pb on the mineral surfaces (57.7 to 61.8%) are the main mechanisms governing Pb2+ adsorption capacity using the sewage sludge biochar. In contrast, Mohan et al. (2007) investigated the potential of wood/bark biochars as adsorbents for As3+, Cd2+, and Pb2+ from water. They reported that the cation exchange was the main mechanism for Pb2+ adsorption using the wood/bark biochars, indicating that the trend of Pb adsorbed onto the biochar was similar to that of the desorbed cations. Meanwhile, in agreement with our findings (Table 1), Ding et al. (2014) reported that the cation exchange in hydroxyl and carboxyl functional groups played a dominant role in Pb adsorption onto the biochar produced at low temperature (250 °C). However, the cation exchange was not the important mechanism for Pb sorption by the biochar produced at 500 °C, which may be attributed to the reduction of the surface functional groups (low CEC, Table 1). Ding et al. (2014) stated that the intraparticle diffusion mechanism was the predominant mechanism of Pb sorption by the biochar produced at high temperature. The results of the current study indicated that the qm value was correlated with increasing biochar surface area (R2 = 0.943, p < 0.05, Fig. 3), suggesting that the addition of biochars produced at high temperature retain Pb2+ more effectively and thus would be more favorable for adsorption of Pb2+ than the biochars produced at low temperature. In this study, a significant relation found between the strength of adsorption of Pb2+ (KL) and the surface area of biochar (R2 = 0.938, p < 0.05, Fig. 4) indicates that KL increased with the increase in the surface area of biochar.

Fig. 3
figure 3

Simple linear regressions between maximum adsorption capacity of Pb2+ on soil (qm) and P of biochar and biochar surface area when averaged across treatments

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Fig. 4
figure 4

Simple linear regressions between strength of adsorption (KL) and biochar surface area when averaged across treatments

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Phosphorus has been applied for Pb2+ immobilization by forming the low soluble Pb–P minerals (Bolan et al. 2003). Lu et al. (2012), Cao et al. (2009), and Xu et al. (2013) reported that the high concentration of PO43− and CO32−, and high precipitation as Pb2+ phosphate and Pb carbonate minerals (84 to 87%) were the main mechanisms of the dairy manure biochar. Garcia-Perez et al. (2002) reported that P was one of the macronutrient in the bagasse biochar with a concentration of 1330 mg kg−1. Therefore, Pb2+ precipitation also occurred on the surface of biochars, especially for biochars with high P content (Ding et al. 2014). Hence, the amount of P in the biochars increased with the increase in the pyrolysis temperature (419 to 1092 mg kg−1) (Table 1). This is supported by the strong correlation between the qm values and the increase in the P amount of the biochar (R2 = 0.996, p < 0.05, Fig. 3).

Conclusion

The results clearly revealed that the adsorption and desorption of Pb2+ are influenced by the presence of Zn2+ in the soil. Also, biochar produced at different temperatures influenced the adsorption and desorption of Pb2+. In the presence of Zn2+, maximum adsorption capacity and strength of adsorption of Pb2+ decreased. The results showed that maximum adsorption capacity and strength of adsorption of Pb2+ increased with the increase in the pyrolysis temperatures. Strength of Pb2+ adsorption enhanced on 90 days of incubation compared with 30 days of incubation in the soil treated with B400 and B600. The amount of the desorbed Pb2+ decreased with the increase in the pyrolysis temperature. Biochar produced at 600 °C has higher capacity and strength of adsorption of Pb2+, while it had lower desorption of Pb2+ compared with other treatments. Also, adsorption capacity of Pb2+ increased as a result of biochar aging. Our findings demonstrated that walnut leaf biochars produced at high temperatures could be used for immobilization of Pb2+ and reduce toxicity of Pb2+ in sandy calcareous soils.