Introduction

Heavy metals are among the pollutants found in water and soils that have most significant environmental impacts. They can be transported to uncontaminated environmental compartments, such as groundwater, and can be absorbed by plants, hence affecting the trophic chain and causing harmful effects to living organisms and the environment (Mohan et al. 2007). Therefore, it is necessary to develop low-cost and effective methods able to reduce the impact of heavy metal pollution in soil and water.

The use of non-toxic solid by-products, such as quarry waste, biosolids and sewage sludge, both in permeable reactive barriers and as amendment of soils, is an increasingly considered strategy for decontamination of metal-contaminated waters or immobilization of metals in contaminated soils, through the irreversible sorption of metals by the solid phase of the by-product (Mulligan et al. 2001; Guo et al. 2006). In this context, the use of biochars for the treatment of contaminated water and soils has recently received much attention (Mohan et al. 2014; Ahmad et al. 2014). Biochar is a carbon-rich pyrolytic material derived from the thermal decomposition of biomass in a closed system with little or no oxygen (Lehmann et al. 2006). It has been found that biochars derived from materials such as oak bark, crop straw, and manure may be more efficient for reducing heavy metal concentrations in water, compared to commercially available activated carbon (Mohan et al. 2007; Cao et al. 2011; Tong et al. 2011). Although biochars may have lower surface area (Han et al. 2013; Wang et al. 2015a) and lower degree of porosity than activated carbon, its higher content in acid functional groups may explain this pattern (Han et al. 2013), as oxygen-containing functional groups (mainly phenolic, carboxylic, and hydroxyl groups) in biochar surface enhance metal sorption efficiency with respect to activated carbon (Bogusz et al. 2015; Frišták et al. 2015). However, the type of biomass from which biochar is obtained affects its intrinsic properties, such as the metal sorption capacity (Zhao et al. 2013; Frišták et al. 2015; Wang et al. 2015b; Venegas et al. 2015).

Metal sorption onto biochar is usually based on cation exchange (e.g., K, Ca, Mg) followed by complexation between oxygen functional groups and metal (Ding et al. 2014; Wang et al. 2015a; Frišták et al. 2015; Zhang et al. 2015). Additionally, precipitation of metals can occur in the presence of phosphate, hydroxide, and others anions released from biochars (Wang et al. 2015a).

There are numerous candidate materials for the production of biochars due to the high levels of generation of biomass wastes in many countries, which should be managed, revalorized, and reduced (IBÁ 2015; IBGE 2015; Bergier et al. 2012). As the metal sorption by biochars cannot be predicted, but it is extremely dependent on the biomass from which it is produced and on the target metal (Chen et al. 2011; Kim et al. 2014), the present work examines the capacity of five biochar materials produced from biomass of different origin, for the sorption and removal of heavy metals (Cd, Cu, Pb, and Zn) present in aqueous solutions, with the aim of recommending the most efficient materials for treatment of contaminated water.

Materials and methods

Materials

Five biochars, produced by slow pyrolysis of biomass, were used in this work. They were obtained from the following biomasses: sugarcane bagasse (SB), eucalyptus forest residues (CE), castor meal (CM), green pericarp of coconut (PC), and water hyacinth (WH) (Eichornia crassipes).

Details of the biomass origin, as well as the pyrolysis conditions used to obtain the biochars and the methods used to characterize them, are included in the Supplementary material.

Determination of water-soluble metal content in biochars

The water-soluble metal contents were quantified in the extracts obtained after equilibrating known amounts of the materials with Milli-Q water for 16 h, using a liquid/solid ratio of 25 mL g−1, with end-over-end shaking at room temperature.

The concentrations of elements in the extracts were determined by inductively coupled plasma optical emission spectrometry (ICP-OES) and inductively coupled plasma mass spectrometry (ICP-MS). Details of the measurements are described in the Supplementary material.

Sorption isotherms

Stock metal solutions containing 100 mmol L−1 Cd, Cu, Pb, or Zn were prepared by dissolving a weighed amount of Cd(NO3)2·4H2O, Cu(NO3)2·3H2O, Pb(NO3)2, and Zn(NO3)2·4H2O (Merck, Pro-Analysi) in Milli-Q water, respectively. Batch sorption tests were performed in 80 mL polypropylene centrifuge tubes. Two grams of the biochar samples was pre-equilibrated for 16 h with 45 mL of Milli-Q water, using end-over-end shaking. Then, suitable volumes of the corresponding stock solution were added in order to obtain initial metal concentrations within the 0.025–5 mmol L−1 range (a minimum of seven different concentrations) after adjustment of the final volume to 50 mL. The resulting suspensions were shaken end-over-end for 24 h. The solid and liquid phases were then separated by centrifugation at 15,000 rpm for 10 min and the solutions were filtered through 0.45 μm Millipore filters. The filtered supernatants were immediately acidified to pH <2 with concentrated HNO3 and kept at 4 °C until analysis.

In order to evaluate possible changes of pH and exchangeable cations caused by the addition of the target metal, control samples were generated in parallel by following the same procedure, but without the presence of the target metal.

Oxidizing digestion of the supernatants was carried out, as described in the Supplementary material, for those highly colored solutions indicating high dissolved organic carbon (DOC) contents, as recommended in previous studies (Silva and Ciminelli 2009).

Desorption tests

The residues obtained from selected sorption experiments (those that used 0.025, 0.10, 0.30, and 1.75 mmol L−1 initial metal concentrations) were employed in the desorption tests. The material remaining from the sorption procedure was dried at 105 °C for 48 h, 50 mL of Milli-Q water were added, and the resulting suspension was shaken end-over-end for 24 h. After attainment of equilibrium, the solid and liquid phases were separated by centrifugation at 15,000 rpm for 10 min and the solution was filtered through a 0.45-μm Millipore filter. The filtered supernatants were immediately acidified to pH <2 with concentrated HNO3 and kept at 4 °C until analysis.

Data treatment and sorption isotherm fitting

The sorption and desorption parameters were calculated from the initial metal concentration in the contact solution (C i, mmol L−1), the equilibrium concentration in the supernatant after the sorption experiments (C eq, mmol L−1), and the equilibrium concentration in the supernatant after the desorption experiments (C eq,des, mmol L−1), as follows:

  1. (a)

    Sorption solid–liquid distribution coefficient (K d, L kg−1):

    $$ {\mathrm{K}}_{\mathrm{d}}=\frac{\left[\left({\mathrm{C}}_{\mathrm{i}}-{\mathrm{C}}_{\mathrm{eq}}\right)\kern0.5em \times \kern0.5em \mathrm{V}\right]/\mathrm{m}}{{\mathrm{C}}_{\mathrm{eq}}} $$
    (1)

    where V is the volume of the liquid phase (in L) and m is the biochar mass (in kg).

  2. (b)

    Removal rate from solution (R, %):

    $$ R\%=\frac{\left({C}_{\mathrm{i}}-{C}_{\mathrm{eq}}\right)\kern0.5em \times 100}{C_{\mathrm{i}}} $$
    (2)
  3. (c)

    Desorption solid–liquid distribution coefficient (K d,des, L kg−1):

    $$ {{\mathrm{K}}_{\mathrm{d}}}_{,\mathrm{d}\mathrm{e}\mathrm{s}}=\frac{\left[\left({\mathrm{C}}_{\mathrm{i}}-{\mathrm{C}}_{\mathrm{eq}}\right)\kern0.5em \times \kern0.5em \mathrm{V}+{\mathrm{C}}_{\mathrm{eq}}\times {\mathrm{V}}_{\mathrm{sol},\mathrm{rem}}-{\mathrm{C}}_{\mathrm{eq},\mathrm{d}\mathrm{e}\mathrm{s}}\times \mathrm{V}\right]/\mathrm{m}}{{\mathrm{C}}_{\mathrm{eq},\mathrm{d}\mathrm{e}\mathrm{s}}} $$
    (3)

    where C sorb is the amount of sorbed metal (mmol kg−1) and V sol,rem is the volume of solution (L) that remained in the residue after the sorption test.

  4. (d)

    Desorption rate (R des, %):

    $$ {\mathrm{R}}_{\mathrm{des}}\%=\frac{{\mathrm{C}}_{\mathrm{eq},\mathrm{d}\mathrm{e}\mathrm{s}}\kern0.5em \times \kern0.5em \mathrm{V}\kern0.5em \times \kern0.5em 100}{\left({\mathrm{C}}_{\mathrm{i}}-{\mathrm{C}}_{\mathrm{eq}}\right)\kern0.5em \times \kern0.5em \mathrm{V}\kern0.5em +\kern0.5em {\mathrm{C}}_{\mathrm{eq}}\kern0.5em \times \kern0.5em {\mathrm{V}}_{\mathrm{sol},\mathrm{rem}}} $$
    (4)

The values of C sorb and C eq obtained at different initial metal concentrations were used for the construction of isotherms. The isotherms were then fitted using the Freundlich model:

$$ {C}_{\mathrm{sorb}}={K}_{\mathrm{F}}{C}_{\mathrm{eq}}^N $$
(5)

where K F is the Freundlich constant and describes the partitioning of the solute between the solid and the liquid phases, and parameter N provides information about the heterogeneity of the sorption sites and is indicative of the degree of linearity of the sorption isotherm.

Results and discussion

Characterization of biochar samples

Table 1 summarizes the main characteristics of the biochars, which mainly depended on the type of biomass used for their production. Most of them showed a neutral or basic pH, with the exception of the SB material, which was slightly acidic. All biochars had high contents of organic matter, as shown by the high loss on ignition (LOI) and total organic carbon (TOC) values. The similar TOC (%) and C (%) values indicated that all the carbon content in the biochars was organic. The DOC contents of the biochars were in the range found in the literature (0.1–110 g kg−1) for biochars produced at low temperature (Gell et al. 2011), except for CM and WH with DOC above this range. Strong correlation was obtained between DOC and fulvic acid (FA) (R 2 = 0.93), as well as between DOC and humic acid (HA) (R 2 = 0.92).

Table 1 Main characteristics of the biochars
Full size table

The acid neutralization capacity (ANC) of CM and WH was similar to that found elsewhere for biochars (Venegas et al. 2015), but considerably higher than the one obtained for SB, CE, and PC. This can be attributed to a lower aromaticity and a higher abundance of carboxylic groups in CM and WH with respect of the rest of biochars, in agreement with the results of the 13C-NMR analyses (see Table S1 in the Supplementary material). Moreover, ANC was correlated with cation exchange capacity (CEC) (R 2 = 0.92) and increased with the Ca+Mg content of the biochars, suggesting that the functional groups responsible for the buffering capacity of the biochars were those that participated in cationic exchange.

The total K, Ca, and Mg contents were highest for WH, being consistent with the highest ash content of this biochar (30 %). In the materials with the highest levels of total P, the soluble fractions of this element were 41 % for WH, 12 % for PC, and 5 % for CM. All the biochars contained high levels of Si, Al, and Fe, suggesting the presence of Fe oxides, aluminates, and silicates, in agreement with previous works (Abdel-Fattah et al. 2015; Zhang et al. 2015).

The total concentrations of Cd, Cu, Pb, and Zn were sufficiently low to be considered negligible for the sorption experiments, with water-soluble fractions lower than the C eq obtained from the sorption isotherms.

Isotherms for metal sorption by the biochars

Description of the sorption isotherms and derived sorption parameters

Figure 1 shows the C sorb vs. C eq sorption isotherms for all the metal-biochar combinations examined. Three contrasted isotherm patterns were observed depending on the metal-biochar combination. CE-Cu and PC-Pb showed a linear behavior, indicating an almost constant K d value for the whole range of concentrations assayed, which is consistent with the interaction of Cu and Pb with organic matter by complexation (Hinz 2001). PC-Cu, WH-Cu, and WH-Pb presented sigmoidal-shaped (S-type) isotherms, characterized by low slopes of C sorb vs. C eq at low surface coverage and higher slopes as the initial metal concentration increased. This type of isotherm suggests the occurrence of competition reactions during sorption processes at low metal concentrations (Hinz 2001). On one hand, the high concentrations of water-soluble Ca, Mg, and K in the WH biochar might affect metal sorption in the low concentration range. On the other hand, the competition at low Cu loadings could be attributed to nitrogen and oxygen ligand groups of DOC that could form soluble complexes with Cu, even at low concentrations of DOC (40 mg L−1). This pattern agrees with the fact that the affinity of DOC for Cu is known to be stronger than for other metals such as Cd, Ni, and Pb (Christensen et al. 1999; Khokhotva and Waara 2010). The rest of isotherms could be classified as high-affinity (H-type) isotherms (Hinz 2001). This reflects a very high affinity of the biochars for the metals at low metal concentrations, with the affinity decreasing at higher metal concentrations. Such behavior suggests the formation of inner-sphere surface complexes at low metal loadings, mainly due to the presence of phenolic and carboxylic groups, whereas ionic exchange may be the main mechanism at higher metal concentrations (Essington 2004; Ahmad et al. 2012).

Fig. 1
figure 1

Isotherms for the sorption of Cd, Cu, Pb, and Zn by the biochars. The red lines indicate the Freundlich fitting (CE eucalyptus forest residues, CM castor meal, PC green pericarp of coconut, SB sugarcane bagasse, WH water hyacinth)

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Table 2 shows the K d,min and K d,max values for all the metal-biochar combinations tested. Differences between the K d,max and K d,min values often exceeded 1 order of magnitude. On the basis of the K d,max values, the WH biochar showed the highest affinity for all the metals studied, with a very high value for Cd, followed by CE (for Cd and Pb) and PC (for Cd, Pb, and Zn). The removal percentages deduced from the K d,min (R min, %) and K d,max (R max, %) values were always higher than 95 %, thus confirming the efficiency of the biochars for the treatment of metal-contaminated waters. The K d values obtained in this work were generally higher than the range reported in the literature (see Table S2 in the Supplementary material), especially for Cd (for all biochars) and for the WH biochar (for all metals). The high metal affinity of the WH biochar could be explained by its high pH and ANC, as well as its high CEC and elevated Ca+Mg content, which could enhance ionic exchange-based sorption mechanisms. The inner-sphere sorption mechanism was also favored in the case of the WH biochar, which presented the highest relative concentrations of phenolic and carboxylic functional groups among the biochars examined (Table S1 in the Supplementary material). Moreover, the unexpected high value of K d(Cd) with respect to the K d of less mobile metals such as Cu or Pb, for most biochars, but especially for WH, might be attributed to the formation of insoluble Cd salts such as K4CdCl6 as suggested by Zhang et al. (2015). The high content of water-soluble phosphorus in WH might also enhance metal phosphate precipitation in the solid phase, especially for Pb (Kołodyńska et al. 2012; Inyang et al. 2012). Tong et al. (2011) also observed that high phosphorus contents in biochars may lead to the formation and precipitation of Cu phosphate.

Table 2 Sorption parameters of the biochars and fitting parameters derived from the Freundlich equation
Full size table

Fitting of sorption isotherms

The Freundlich model was selected to fit the experimental data, as described previously for the sorption of metals by other sorbents (Pellera et al. 2012; Shaheen et al. 2013). The results of the fitting exercise are shown in Table 2 (N and K F parameters) and Fig. 1 (fitted curves). In general, the Freundlich model provided a satisfactory description of the metal sorption, with values of the correlation coefficients often higher than 0.9. The values of N obtained from the fitting corroborated the deductions made above from the shapes of the isotherms: N ≈ 1 and a value of K F close to K d for CE-Cu and PC-Pb, which is consistent with linear isotherms; N > 1 for PC-Cu, WH-Cu, and WH-Pb, in agreement with S-type isotherms; and N < 1 for the rest of biochar-metal combinations.

Due to the low C eq in the obtained isotherms (lower than 1 mM), the K F values were always extrapolated. Thus, K F values were lower than the K d, min for the high-affinity isotherms (N < 1), whereas they were higher than the K d, max for the sigmoidal-shaped isotherms (N > 1). Because the K F values are only fully comparable when the N values are the same (Coles and Yong 2006) and they often relate to a metal concentration too high to be representative of environmental scenarios, two additional sorption parameters were thus calculated here in order to improve the comparison. Considering that the isotherms showed linearity between C sorb and C eq, at low C eq, K d,linear values were derived from the slopes of the linear regressions in the region of low metal concentrations. In addition, a K d,Freundlich value was calculated at a C eq of 0.002 mmol L−1, being a concentration level of environmental interest within the range of maximum metal concentration thresholds often applied to drinking water (4.10−5 and 0.08 mmol L−1, depending on the metal considered) (U.S.EPA 2011). Values of K d,linear and K d,Freundlich are also shown in Table 2. Good correlation was obtained between the K d,Freundlich and K d,linear values (R 2 = 0.90), and the these two sorption parameters were also reasonably correlated with K d,max (R 2 = 0.83 with K d,Freundlich and R 2 = 0.79 with K d,linear). This indicated that both K d,Freundlich (at a moderate metal concentration of 0.002 mmol L−1) and K d, linear satisfactorily described the sorption capacities of the biochars, and permitted to identify those with the maximum sorption capacity at metal concentrations of environmental interest. Thus, WH and PC had the highest (or second highest) K d, Freundlich and K d, linear for all of the metals examined.

Sorption mechanisms

In order to further understand the mechanisms of sorption of the metals by the biochars, measurements of pH and the concentrations of Ca, Mg, and K were conducted for selected equilibrium solutions (from the experiments with C i of 0.1, 1, and 5 mmol L−1). An increase in the initial metal concentration generally led to a decrease in the pH of the equilibrium solution, compared to the control solution without addition of the metal (see Table S3 of the Supplementary material). For several metal-biochar combinations, the difference ranged from tenths of a pH unit (for the lowest metal concentration) to nearly 3 pH units (for the highest metal concentration).

In addition, there was a general increase in the Ca+Mg and K concentrations in the equilibrium solutions when the initial metal concentrations of the sorption solutions were higher, which could be explained by a displacement of these cations from the solid phase due to ionic exchange-based heavy metal sorption. These results are shown in Table S3 of the Supplementary material. For the WH biochar, with an extremely high initial K concentration in the solution, the effect was negligible for this cation and could not be observed. These results were in agreement with those observed previously for biochars and other organic materials (Sastre et al. 2007; Mohan et al. 2007). With the exception of the WH biochar, the ratio between the amount of sorbed metal and the net increase in the concentrations of Ca+Mg and K released to the solution tended towards 1, for an initial metal concentration of 1 mmol L−1. This supported the role of cationic exchange as the predominant sorption mechanism, together with a contribution of metal sorption due to the formation of inner-sphere surface complexes at the low metal concentration range (Fiol et al. 2006; Mohan et al. 2007), suggested in the previous section. For a few metal-biochar combinations, the ratios tended to increase at high heavy metal loadings (5 mmol L−1), indicating that other sorption mechanisms, such as precipitation, could not be disregarded (Mohan et al. 2007; Kołodyńska et al. 2012).

Evaluation of sorption reversibility

Desorption parameters such as desorption solid–liquid distribution coefficient (K d,des) and desorption yield (R des) were calculated from the desorption tests. The values of these parameters are summarized in Table S4 of the Supplementary material. The K d,des values were systematically higher than the corresponding K d values, exceeding 106 L kg−1 in many cases, hence confirming that the sorption process was essentially irreversible. The K d,des,max and, thus, the minimum desorption yields (R des,min, %) generally corresponded to the material residues obtained from the sorption tests carried out using lower initial metal concentrations, with most values below 2 %, while the desorption yields increased up to 19 % (R des,max, %) for the sorption residues derived from the tests at the highest C i.

It is difficult to predict desorption data from sorption parameters. However, for the Cd and Zn isotherms, with N values clearly lower than 1, indicating site heterogeneity, an inverse relation was observed between the desorption yields and the K d values (Fig. 2). An explanation is that at low concentrations, sorption occurred at specific sites, with low reversibility, while at higher concentrations, the metal was mainly sorbed at low-affinity sites, with higher reversibility. However, for other metals such as Pb and Cu, with N values generally close to 1, the desorption yields were independent of the K d values.

Fig. 2
figure 2

Relation between desorption yields and K d, for Cd and Zn. The dotted frames indicate the most efficient points for the sorption process with high K d and low reversibility (CE eucalyptus forest residues, CM castor meal, PC green pericarp of coconut, SB sugarcane bagasse, WH water hyacinth)

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The desorption results were in agreement with previous reports (Regmi et al. 2012; Trakal et al. 2014). The data confirmed that sorption of the metals by the biochars was largely irreversible and was controlled by inner-sphere sorption mechanisms, especially at low initial metal concentrations. Therefore, metal sorption by the biochars examined here was characterized not only by a relatively high K d but also by low sorption reversibility.

Conclusions

The biochars examined in this work presented high sorption of the target metals investigated, with removal rates from solution exceeding 95 % for the whole range of concentrations assayed, and even higher than 99.9 % for the low metal concentration range. Thus, the high sorption capacity of the biochars tested, together with low sorption reversibility, confirms them as suitable materials for the treatment of water contaminated with heavy metals. In addition, sorption K d values higher than 104 L kg−1 were obtained for all metals in the case of water hyacinth and green coconut pericarp, and for a few metals in the rest of biochars. These high K d values indicate that the addition of these biochars as amendments in soils contaminated with metals can contribute, besides improving the quality of the soil, to soil remediation.