Copper (Cu) is an essential mineral element for plant growth, but it is toxic to plants and animals, including humans when the concentration of copper reaches or exceeds a certain level (Kim et al. 2020; Rakhra et al. 2017; Shahid et al. 2016). At present, the remediation measures of heavy metal pollution mainly include physical (Yuvaraj et al. 2021), chemical remediation (Rambabu et al. 2019), and biological remediation (Akcil et al. 2015; Ndeddy Aka and Babalola 2016). Biochar has the ability to adsorb and passivate heavy metals in soil, which can be used as a remediation approach for heavy metal contaminated soil (Li et al. 2019; Namgay et al. 2010). Possible mechanisms for biochars contributed to the heavy metal immobilization were precipitation and adsorption (primarily electrostatic interactions, ionic exchange, and complexation) of metal ions on biochars (Salam et al. 2019; Uchimiya et al. 2011) which is due to its large surface area, high porosity and pH, and a large number of active functional groups such as hydroxyl, carboxyl and carbonyl groups (He et al. 2018; Rizwan et al. 2020).

During the process of producing biochar by straw, major researches focus on the passivation of biochar on heavy metals in soil (Chen et al. 2020). However, it’s unclear whether the heavy metals absorbed by biochar are released again. In this study, two kinds of rice straw biochar were prepared at 400°C (BC400) and 600°C (BC600). Their apparent structure, group characteristics, and basic physical and chemical properties were determined. The isothermal adsorption and kinetics adsorption of Cu were analyzed. The biological effects of BC400 and BC600 on the passivation of Cu in the soil were evaluated by a pot experiment. The objectives of this study were to determine the following:(1) the adsorption efficiency to Cu by rice straw biochar prepared at different temperatures; (2) the effect of Copper-loaded biochar on the chemical morphology of copper ions in soil; (3) determine the passivation and sorption–desorption capacity of biochar to Cu2+ in soil.

Materials and Methods

Rice straw, collected from Hai’an, Jiangsu province, China (120° 22′ E, 32° 22′ N), was washed three times in deionized water and dried at 75°C for 48 h, then crushed to pass through a 0.90-mm sieve. The rice straw has the following properties: pH 6.18, nitrogen 6.12 g kg−1, phosphorus 3.52 g kg−1 and potassium 12.71 g kg−1, Cu 5.65 mg kg−1. The rice straw was subjected to oxygen-limited cracking, in which it was placed in a rectangular corundum porcelain boat and pyrolyzed in a tube-furnace under N2 protection. The pyrolysis temperature was set between 400 and 600°C with a heating rate of 10°C min−1 and held at the peak temperature for 4 h. When the biochar had cooled to room temperature (25°C), it was collected, weighed, and crushed to pass through a 0.15-mm sieve. The BC400 and BC600 were then stored for compositional analysis.

The ash content of biochars was determined by burning method, a small amount of biochar placed in the muffle furnace and burned at 750°C for 6 h (Brewer et al. 2014) and the pH of biochar was determined in deionized water at the ratio of 1:2.5 (m:v) with a pH meter (Inesa Inc., Shanghai, China).

The pHpzc, for biochar were determined by the following procedure (Medellin-Castillo et al. 2007): 25 mL of 0.1 mol L−1 KNO3 solution was added to a 50-mL centrifugal tube and bubbled with nitrogen gas to expel dissolved CO2. Next, 0.500 g biochar was added, 0.1 mol L−1 HNO3 or 0.1 mol L−1 NaOH were used to adjust the solution pH to a selected value from 2 to 12, then the samples were shaked at 160 rpm for 24 h at room temperature, Then the samples were filtrated and the final pH was measured separately, the pH difference (ΔpH) between final pH (pHf) and initial pH (pHi) was plotted against the initial pH (pHi), the point at which ΔpH equal to zero was considered to be the pHpzc.

Different biochar materials were scanned by S-4800II field emission scanning electron microscopy (FESEM, HITACHI, Japan) to observe the microstructure and morphology; the elements of biochar were determined by energy dispersive spectroscopy (Arakaki et al. 2006) and Vario EL cube element analyzer (Elementar Analysensystem, GmbH, Germany), the specific surface area of biochar was determined by Autosorb IQ3 specific surface area and pore size were measured based on the Brunauer-Emmet-Teller (BET) (Autosorb IQ3, Quantachrome, USA).

To study adsorption isotherms, 0.30 g of biochar received 50 mL of CuSO4 solution with Cu2+ concentration of 200 mg L−1, and samples were collected at 0, 3 min, 12 min, 30 min, 1 h, 2 h, 4 h, 6 h, 8 h, 12 h, 16 and 20 h. The sample was shaken for 24 h on a shaker at 200 r min−1 and 25°C, and then filtered through a 0.45 μm × 150 mm filter. The concentration of heavy metals in the filtrate was determined with an inductively coupled plasma mass spectrometer (ICP-MS; Thermo Fisher Scientific, USA). The amount of heavy metal adsorbed by the biochar was calculated according to the change of the heavy metal content in the solution before and after the treatment. The initial pH value of the mixtures was adjusted to 5.5 ± 0.05 using HCl and NaOH solution (0.1 M) under stirring. All tests were conducted in triplicate. The experiment was conducted at room temperature (25°C).

To investigate the adsorption characteristics of Cu by biochar prepared at different temperatures, the Langmuir and Freundlich isotherm adsorption equations for Cu adsorption were fitted (Haghtalab et al. 2012). The formulas are as follows:

$${\text{Langmuir adsorption isotherm is}}:\,{\text{q}}_{{\text{e}}} = \frac{{{\text{Kq}}_{{\text{m}}} {\text{C}}_{{\text{e}}} }}{{1 + {\text{KC}}_{{\text{e}}} }}$$
$${\text{Freundlich adsorption isotherm can be expressed as}}:\,{\text{q}}_{{\text{e}}} = {\text{K}}_{{\text{f}}} {\text{C}}_{{\text{e}}} ^{{1/{\text{n}}}}$$

where \({\text{q}}_{\text{e}}\) represents the amount of adsorption at equilibrium (mg/g); \({ \text{C}}_{\text{e}}\) represents the solution concentration at equilibrium (mg L−1); \({\text{q}}_{\text{m}}\) represents the maximum adsorption capacity (mg g−1); K is the Langmuir equilibrium parameter constant (L mg−1); \({ \text{K}}_{\text{f}}\) is the adsorption capacity (mg g−1); and n is the Freundlich constant, a constant for a given system at a certain temperature, related to the physical properties and temperature of the adsorbent.

The adsorption kinetics were calculated using the following formula:

$${Q}_{t}={Q}_{e}(1-{e}^{-{k}_{1}t})$$
$${Q}_{t}=\frac{{Q}_{e}^{2}{k}_{2}t}{1+{Q}_{e}{k}_{2}t}$$

where \({Q}_{t}\) and \({Q}_{e}\) (mg g−1) are the amounts of Cu adsorbed on biochar at time \(t\) and equilibrium, respectively; and \({k}_{1}\) are \({k}_{2}\) are the corresponding adsorption rate constants.

The soil was obtained from Yangzhou University agricultural experiment filed, Yangzhou, China (119° 23′ E, 32° 20′ N). The soil properties were as follows: pH, 7.75; organic matter, 5.96 g kg−1; total Cu, 111.10 mg kg−1. 3 kg soil (of air-dry soil and sieved using a 2-mm sieve) were placed in plastic beaker.

Five treatments were prepared in triplicate as follows: control (CK), BC400, BC600, BC400-Cu and BC600-Cu. Except for CK, 2%(W/W) BC400, BC600, BC400-Cu and BC600-Cu were added into a plastic beaker. The mixture was stirred to be well homogenized and then incubated at room temperature (25°C) for 30 days. Base fertilizer (N, 180 mg kg−1; P2O5, 90 mg kg−1; K2O, 120 mg kg−1) were mixed with the soil. Then, after sowing and thinning out seedlings of Ryegrass (Loliumperenne L.), 40 healthy individuals were kept in each pot. During the test, the weighing method was used to maintain the soil moisture content at about 60% of the saturated water-holding capacity. The ryegrass was grown for a total of 120 days, during which the above ground part biomass of the it was harvested once every 30 days for a total of four harvests. After each harvest, it was dried at 105°C for 0.5 h, 65°C for 24 h, and then weighed. Finally, the four harvested ryegrass samples were mixed and crushed to pass through a 0.90-mm sieve, then plant samples were carried out by dry ashing. The different fractions of Cu in the soil samples have been determined by means of BCR sequential extraction method which was developed by the European Bureau of Reference. Cu was measured by inductively coupled plasma mass spectrometry (ICP-MS; Thermo Fisher Scientific, USA).

Results and Discussion

The basic physical and chemical properties of BC400 and BC600 were shown in Table 1. The yield of the biochars decreased from 37.6% to 31.7% as the temperature increased from 400 to 600°C, the decrease yield of the biochars with temperature was similar to the reports by Zhu et al. (2018) which was attributed to the loss of volatiles and the condensation of aliphatic compounds because of increasing temperature. In addition, BC600 had a higher ash content compared to BC400. BC600 had a high pH and a low CEC value. Our results were consistent with the reported by Wang et al. (2013).

Table 1 Material properties of BC400 and BC600
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The atomic ratios of H/C, O/C and (N+O)/C of organic components in biochar could reflect the aromaticity, hydrophilicity and polarity of biochar samples. As the atomic ratio of H/C decreases, the degree of aromaticity increased, and as the ratio of O/C and (N+O)/C increased, the aromatic structure and polar groups also increased (Chi et al. 2017; Jung et al. 2013). Table 1 showed that the aromaticity and polarity of BC600 were greater than BC400, indicating that the former had more aromatic structures and polar groups. Chen et al. suggested that high temperatures were beneficial to the thermal transformation of organic matter into carbonized organic matter and the structure containing dense ring carbon (Chen et al. 2014). Surface area of the biochar increased from 13.07 to 29.00 m2 g−1 as temperature increases from 400 to 600°C (Table 1).

Figure 1 showed that BC400 were primarily strip-like, neatly arranged and had a smooth surface, whereas BC600 were more granular in appearance and had a rougher surface. Compared with BC400, pores on BC600 were well developed, and the pore distribution was relatively dense. Hence, it can be concluded that the surface structural changes in biochar are significantly influenced by pyrolysis temperature.

Fig. 1
figure 1

The SEM of surface structure of BC400 and BC600

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Corresponding pore size distribution curve of BC400 and BC600 are shown in Fig. 2a and b. The rice straw biochar prepared at 400 and 600°C belongs to the typical TYPE IV curve, which indicated that biochar contains mesoporous structure. BC600 contains more mesoporous structure and the specific surface area than BC400.

Fig. 2
figure 2

N2 adsorption–desorption isothermal a and Barrett–Joyner–Halenda (BJH) Corresponding pore size distribution curve b of BC400 and BC600

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The infrared spectra showed in Fig. 3 that the surface of biochar contained oxygen-containing functional groups, such as –COOH, –OH, aromatics, lipids, and alkane. The adsorption peaks near 3215 and 3366 cm−1 are the stretching vibration peaks of –OH, while the adsorption peaks near 2913 cm−1 are the stretching vibration peaks of –CH2 in alicyclic hydrocarbons or naphthenic hydrocarbons (Frost et al. 2015). However, BC600 did not contain this peak. The results show that the increase of pyrolysis temperature resulted in the loss of alkane groups; the adsorption peaks near 1701 and 1799 cm−1 were C=O stretching vibration adsorption peaks of carboxylic acid, indicating the presence of acids; the peaks near 776 and 789 cm−1 were C-H bending vibration adsorption peaks of aromatic rings; the peaks at 1455 and 1615 cm−1 were C=C bending vibration peaks; and the peaks near 1320 and 1393 cm−1 were C–O–C stretching peaks. The existence of shrinkage vibration peaks clearly shows the existence of benzene ring substances on the surface of biochar, indicating that rice straw formed a good aromatic structure during carbonization (Mohanty et al. 2013). In general, BC600 contained more aromatic structure substances than BC400.

Fig. 3
figure 3

FTIR spectra of BC400 and BC600

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Comparing the functional group compositions of biochars prepared at the two temperatures, which reveals that BC400 and BC600 have the same types of groups, similar characteristic adsorption peaks, and more oxygen-containing functional groups such as –COOH and –OH on the surface, those cations have the ability of adsorbing. The primary variation between the two temperature biochars is shown in –OH and –CH2. The vibration peak of –OH is weaker than the corresponding vibration peak of BC400, but it contains more aromatic structures, which is consistent with the results of surface element composition (Table 1). That is, with the increase of temperature, the degree of aromatization increases. According to Weidemann et al. (2017) the degree of aromatization of aromatic components increases with the increase of temperature at 400°C and above.

The parameters in Table 2 showed that the fitting degree between BC400 and BC600 for Cu adsorption and Langmuir adsorption model (R2 of 0.957 and 0.971, respectively) is better than that with the Freundlich adsorption model (R2 of 0.857 and 0.804, respectively). The results reveal that the adsorption of Cu by BC400 and BC600 was mainly monolayer adsorption (Li et al. 2017) (Fig. 4).

Table 2 Parameters in the fitting model of Langmuir and Freundlich isothermal adsorption of Cu2+ on biochar surface
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Fig. 4
figure 4

Langmuir and Freundlich adsorption isotherms for the adsorption of Cu2+ by biochar

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Moreover, the maximum adsorption capacity of BC600 (Qm = 43.75 mg g−1) was higher than that of BC400 (Qm = 30.70 mg g−1), and the Cu adsorption capacity of BC600 was higher than that of BC400 in the Cu concentration range of 0 ~ 300 g kg−1.

The kinetic sorption data were described using the pseudo first-order and pseudo second-order models. The effect of transfer rate on Cu absorption by BC400 and BC600 was represented in Fig. 5. Table 3 showed the fitted model parameters for these models. The Cu sorption by BC400 and BC600 reached equilibrium after 4 h. The equilibrium adsorption capacity of BC600 was higher than that of BC400. Compared with the pseudo first-order kinetic model (R2 of 0.9215 and 0.94545, respectively), the kinetic process of Cu absorption by BC400 and BC600 was better fitted by the pseudo second-order kinetic model (R2 of 0.97075 and 0.95437, respectively), which was used to describe chemisorption (Ho and Mckay 1999; Liu and Fan 2018; Wei et al. 2015). The pseudo second-order kinetic model fitted the sorption data with a higher consistency between theoretical and experimental Qe values. The k2 value of BC600 was higher than that of BC400, indicating that the rate of Cu adsorbed by BC600 was faster than that adsorbed by BC400, which shows higher ash content, functional groups and pH of biochar contribute to the adsorption of Cu in aqueous solution (Sun et al. 2014).

Fig. 5
figure 5

Adsorption kinetics for Cu adsorption by BC400 and BC600

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Table 3 Kinetic fitting parameters for Cu adsorption by BC400 and BC600
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Energy spectrum analysis of biochar surface elements can be used to indicate whether the biochar surface can adsorb heavy metals. The energy spectrum shown in Fig. 6 reveals that Cu was not detected on the surfaces of BC400 and BC600. The characteristic peaks of Cu appeared on the surfaces of BC400-Cu and BC600-Cu. This confirms that Cu2+ had been adsorbed on the surface of BC400 and BC600.

Fig. 6
figure 6

EDS of BC400 and BC600

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Figure 7 shows the effect of biochar on the Cu concentration of different fractions and pH values in soil. The pH values of the BC400 and BC600 treatment were significantly higher than CK (pH 8.57), increasing by 0.18 and 0.14 units, respectively. It has to do with that biochar have high alkalinity (Xu et al. 2016), the pH values of BC400 and BC600 is 11.73 and 12.61, respectively. The pH of biochar increases with the increase of pyrolysis temperature, this trend is consistent with many reported studies (Zhang et al. 2015). The mineral elements (K, Na, Ca, Mg, etc.) of biochar complexed with organic matter gradually, then changed to oxidation or phosphate state during pyrolysis, and the higher the pyrolysis temperature, the higher the content of inorganic mineral components (such as carbonate) (Yuan et al. 2011). While the pH values of the BC400-Cu and BC600-Cu treatments decreased, which were 8.31 and 8.29, respectively. Because of CuSO4 adsorbed by biochar, the pH values of BC400-Cu and BC600-Cu is low than BC400 and BC600.

Fig. 7
figure 7

Effect of different treatments on Cu concentrations of different fractions and pH in soils (mean ± SD, n = 3). Different letters above the blocks indicate significantly differences (p < 0.05)

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The toxicity of heavy metals in the soil environment is related not only to the total amount but also determined by the distribution of their fractions (Kabiri et al. 2019). Different fractions produce different environmental effects, which directly affect the toxicity, migration and circulation of heavy metals. Acid extractable of heavy metal is absorbed the most easily, can better reflect the soil heavy metal biological effectiveness and mobility, through the determination of various forms of heavy metals in soil, especially the content of acid soluble state of heavy metal can be evaluation into the effect of improved material (Jamali et al. 2008). Figure 7. shows that the residual is mainly fraction of CK, the treatments of BC400 and BC600 in the soil and the acid extractable concentrations Cu in the treatment of BC600 significantly reduces, compare with CK. Due to the large specific surface area, pore structure and functional groups, biochar can adsorb heavy metals or form metal complexes with them (Dong et al. 2013), and Cd is gradually transferred from a more active form to a relatively stable form (Guo et al. 2017). The treatments of BC400-Cu and BC600-Cu are mainly in acid-soluble and reducible states, mainly because CuSO4 adsorbed by biochar enters the soil.

Cu concentration in ryegrass is shown in Fig. 8. Compared with 10.80 mg kg−1of CK, Cu concentration in ryegrass of the BC400 and BC600 treatments decreased significantly, which were 6.77 and 4.24 mg kg−1, respectively. This shows that biochar has a “passivation” effect on Cu in the soil, and reduces the absorption of Cu by ryegrass, and BC600 had a more passivation ability than BC400. Studies have shown that bichar can reduce bio-accessibility of Cu (Li et al. 2020) and the content of Cd in rice and available Cd in soil decreased gradually with the increase of biochar application (Chen et al. 2016). The treatments of BC400-Cu and BC600-Cu are treated by absorbing Cu and have two sources of Cu, one being soil Cu and the other is Cu absorbed by biochars. Compared with the BC400 and BC600 treatments, the average concentration of Cu (16.30 and 18.60 mg kg−1, respectively) in ryegrass increased significantly because of the abundant sources of Cu.

The ratio of Cu concentration in ryegrass to in soil was defined as the bioconcentration factor of Cu. The bioconcentration factor in CK was significantly higher than that of the BC400 and BC600 treatments. The results confirm that biochar can reduce the bioconcentration factor. Some studies indicate that when 1.5% bamboo biochar was applied to cd-contaminated soil while growing pakchoi and maize, the bioavailability of cadmium was significantly reduced by 60.64% and 64.04%, respectively (Mohamed 2015). For the BC400-Cu and BC600-Cu treatments, the uptake of Cu by ryegrass can be attributed not only to soil Cu, but also to Cu absorbed by biochar. Although Cu concentration in ryegrass increased (Fig. 8), the bioconcentration factor of the BC400-Cu and BC600-Cu treatments was significantly lower than that of the BC400 and BC600 treatments, respectively, indicating that the release of Cu adsorbed by BC400 and BC600 in soil was greatly limited, and BC600 has comparatively stronger passivation than BC400.

Fig. 8
figure 8

Cu concentration in ryegrass and the bioconcentration factor of Cu (mean ± SD, n = 3). Different letters above the blocks indicate significantly differences (p < 0.05)

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Biochars produced from rice straw at two pyrolysis temperatures exhibited variable characteristics and efficiency towards Cu immobilization in the contaminated soils. This study demonstrated that BC600 and BC400 have more aromatic and polar groups than BC400. The adsorption curves of BC400 and BC600 were more suited to the Langmuir isothermal adsorption model than the Freundlich model, indicating that the Cu adsorption is mainly monolayer. The saturated theoretical adsorption capacity of BC600 for Cu2+ can reach 43.75 mg g−1. Kinetic analyses implied that the adsorption fitted well with prseudo-second-order model.

Moreover, the pot experiment shown that BC400 and BC600 not only have strong “passivation” effect on Cu in soil but also prevents the release of adsorbed Cu. The existence of these two effects can greatly reduce the bioconcentration factor. Due to the more aromatic polar groups, and higher pH, the passivation and release inhibition of BC600 are comparatively stronger than those of BC400. Therefore, the application of BC600 may be effective to reduce Cu pollution in Cu-contaminated soil.