1 Introduction

Water quality is affected by persistent or pseudo-persistent organic and inorganic pollutants from natural sources and anthropogenic activities. The persistence of pollutants in water sources is of great concern since they may be able to be transferred to the trophic chain, presenting a risk to the environment and human health [1, 2].

Malachite green (MG) synthetic dye is a cationic organic compound used mainly by the textile industry. Besides, MG is used as an antifungal and antibacterial agent in the aquaculture industry [3]. Discharges of untreated dye effluents pollute surface water and groundwater as dyes tend to be stable, difficult to degrade, cumulative, and toxic [4]. In addition, MG hinders the photosynthesis process in the aquatic ecosystem contributing to high chemical oxygen demand (COD) [5]. The non-biodegradable and genotoxic characteristics of the MG induce carcinogenicity and mutagenicity. In the aquaculture industry, the use of MG is not allowed in many countries for the treatment of food-producing animals, including the USA, the UK, China, Canada, and the European Union. However, food control authorities established a minimum required performance limit (MRPL) of 2 μg/kg of MG residues in aquaculture products and the environmental quality standard limit for the concentration of MG in water was set around 0.5–100 μg/L [6, 7].

On the other hand, toxic inorganic elements such as arsenic tend to release into the environment from the natural weathering of rocks and industrial, agricultural, and mining activities [8]. Arsenic contamination has been reported in many countries because of high concentrations in water sources. Arsenic-contaminated irrigation water represents a major risk to food safety, agricultural production, and soil contamination. Different studies have reported the bioaccumulation of arsenic in different parts of the plant, in the order roots > stem > leaves > edible parts in crops such as rice, wheat, maize, and some fruits and vegetables [9, 10]. Therefore, the recommended permissible limit of arsenic intake in drinking water as set by the World Health Organization (WHO) should be lower than 10 μg/L as a provisional guideline value [11]. Arsenic is a metalloid that can be present as arsenite [As (III)] or arsenate [As (V)]. They are more toxic and mobile than organic arsenic compounds [12, 13]. Additionally, As (III) is more difficult to remove from water than As (V) because at pH < 9 As (III) exists in the uncharged form (H3AsO3). The negatively charged species of As (III), including H2SO3−, HAsO32−, and AsO33−, are found at a pH higher than 9.2. Consequently, the removal of As (III) in water by ion exchange or sorption processes is less effective than for As (V), which can exist in the form of different ions in a wide range of pHs [14, 15].

Different remediation technologies have been used for the removal of inorganic and organic pollutants present in water, such as flocculation-coagulation [16], electrochemical oxidation and photocatalytic processes[17], membrane separation [18], and phytoremediation [19], among others [20]. Currently, special attention is focused on the sorption process as a simple, efficient, and low-cost alternative for effectively remedying different contaminants in water [21]. Activated carbon is the most common and widely used sorbent for removing pollutants. However, the high production price and the use of non-renewable feedstocks have promoted the search for alternative materials of low cost and greater availability [22, 23]. Recent studies have focused on the use of different materials as potential sorbents, such as biochar, obtained by pyrolysis of industrial organic by-products of agricultural wastes, including the use of plant and animal wastes, sewage sludge, and organic municipal solid waste as materials for the removal of dyes, heavy metals, metalloids, and other contaminants [24].

Conventionally, studies have used the slow pyrolysis process to obtain biochar by thermal decomposition of biomass in an inert atmosphere [24]. Generally, biochar from gasification has not been used extensively for environmental management; its use has been studied mainly as a soil amendment agent or growing media component [25, 26]. However, the utilization of the top-lit updraft gasification process has previously been tested to produce biochar with properties for new uses [27]. Therefore, it is an alternative that can be explored to use biochar for environmental management.

The main objective of this study is to evaluate the potential of rice husk and wild sugarcane biochars obtained by the gasification process as sorbents in removing the MG and the As (III) from aqueous solutions. Rice husk is a highly available agro-industrial waste biomass; its main components are cellulose, hemicellulose, lignin, and minerals that depend on the variety of rice, climatic conditions, and geographical location of the crop [28]. In Panama, the wild sugarcane (Saccharum spontaneum L.) is considered an invasive species of rapid propagation and persistence since it alters the growth process of native plants [29]. In addition, other studies have focused on using wild sugarcane as an energy resource [30] or for bioremediation of groundwater contaminated with nitrate [31, 32]. Recently, biochar from wild sugarcane was evaluated for removing a synthetic herbicide (atrazine) in aqueous solutions [33].

The main objective of the present work is to study the potential use as sorbents of MG and As (III) of two biochars obtained as by-products of rice husk and wild sugarcane gasification. Batch sorption studies were carried out, and the interaction between sorbent and sorbate was evaluated using kinetic models (pseudo-first-order, pseudo-second-order) and sorption parameters.

2 Materials and methods

2.1 Sorbates

Sodium (meta) arsenite and malachite green oxalate dye with purity ≥ 90% were supplied by Sigma–Aldrich chemical company. These were used to prepare the stock solutions for the batch sorption experiments. The chemical characteristics of both sorbates are shown in Table 1.

Table 1 Chemical parameters of sorbates
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2.2 Preparation of biochars

Two types of biomasses were selected as raw materials to produce biochar, rice husks (RH) and wild sugarcane (WS) (Saccharum spontaneum L.). The rice husk biomass is the agro-residual product of the company Molino el Anhelo in Panama, and the sugarcane wild was obtained from the local area in Chilibre (Panama), according to [33]. A top-lit updraft (TLUD) gasification reactor was used to carbonize the biomasses with an air supply of 16 L/min. The carbonization process used a total of 1107 g of RH and 1741 g of WS with a particle size of less than 4.75 mm. The gasification process was carried out according to previous studies [25, 34]. Reaction temperatures were measured using 2 K-type thermocouples (1/8 in. diameter) distributed along the reactor, and temperatures were recorded at a sampling rate starting from 1 s with a data acquisition system 4-channel K thermometer SD logger (88,598 AZ model, Taichung City 427, Taiwan R.O.C).

2.3 Biochar characterization

The following parameters were determined for each biochar: pH and redox potential (Eh) were determined at a dilution of 0.1:25 (g mL−1) using a Crison micro pH 2000 and a pH 60 DHS, respectively [35, 36]. Ash percentage was determined by loss on ignition (LOI) method [37], and cation exchange capacity (CEC) was determined with the PerkinElmer AAnalyst 400 AA Spectrophotometer according to the standardized protocol of ISO 2347 [38]. Z potential analysis was performed using the Malvern Zetasizer Nano ZS90. C, H, N, and S contents were determined by dry combustion using a LECO CHNS 932 analyzer (SCAI-Málaga University). Oxygen was calculated by difference as 100% − (%C + %H + %N + %S + %Ash). Textural parameters of two biochars were determined by N2 adsorption isotherms, obtained using an ASAP 2420 gas sorption analyzer from Micromeritics (SCAI-Málaga University). The sample (around 120 mg) was previously degassed under dynamic vacuum conditions to constant weight at 150 °C. The apparent specific surface area (SBET) and micropore volume (Smic) were obtained using the MicroActive software (v-4.03) from Micromeritics.

Finally, biochars were analyzed by Fourier transform infrared spectroscopy (FTIR) using a Bruker vertex70 FTIR spectrophotometer (SCAI-Malaga University). The measurements were carried out by transmission with the biochar power sample dispersed in KBr. A standard spectral resolution of 4 cm−1 in the spectral range of 4000–400 cm−1 and 64 accumulations per sample was used for the spectra acquisition.

2.4 Sorption studies

Sorption studies will be performed with MG and As (III) solutions.

2.4.1 Batch sorption experiments

The essential parameter to consider when designing sorption experiments is the sorption kinetics, which determines the rate at which sorption occurs. In the batch sorption kinetics study, BRH and BWS were used as sorbents of MG and As (III) in a 250-mL solution, composed of a concentration of 30 ppm of sorbate and 250 mg of each biochar, placed in a shaker in a water bath. Sorption experiments were performed at room temperature (21 °C), with controlled agitation, and carried out at the natural pH of the MG and As (III) solutions. Aliquot samples were extracted at specific time intervals of 10, 20, 30, 50, 70, and 120 min for MG and at time intervals of 10, 20, 30, 60, and 120 min for As. The final dye concentration of each aliquot was analyzed at a wavelength of 618 nm using a UV–Vis (Zuzi 4201/50 spectrophotometer), and the final As (III) concentration was quantified by ICP-MS (SCIEX Perkin Elmer model) from SCAI-Malaga University. The equilibrium sorption capacity of MG and As, qe (mg·g−1), was calculated using Eq. 1:

$${q}_{e}=\frac{\left({C}_{i}-{C}_{e}\right)\left(V\right)}{W}$$
(1)

Ci and Ce (mg·L−1) are the liquid-phase initial and equilibrium concentrations of MG and As (III) in solution, respectively. V is the batch volume (L), and W is the mass of dry sorbent used (g).

2.4.2 Kinetic models

The sorption analysis has been carried out using kinetic models, which provide insight into the sorption pathways, the likely mechanisms involved, and possible rate-limiting steps. To complete this analysis concerning the possible nature of the interactions between the sorbent and the sorbate (i.e., physisorption or chemisorption), the experimental data were fitted to kinetic models such as pseudo-first- and pseudo-second-order models [39].

Pseudo-first-order (PFO)

It is also known as the Lagergren model. It assumes that the rate-limiting involves the diffusion process and that the sorption kinetic depends only on the sorbate concentration. Physisorption phenomena control this model (Van der Waal forces, mechanical adhesion, and/or hydrogen bonding). The PFO equation is defined according to Eq. 2

$$\frac{d{q}_{t}}{dt}={k}_{1}\left({q}_{e}-{q}_{t}\right)$$
(2)

After integration, the linearized form of PFO is obtained according to Eq. 3:

$$\mathrm{ln}\left({q}_{e}-{q}_{t}\right)={\mathrm{ln}q}_{e}-{k}_{1}t$$
(3)

where qt is the sorption capacity at time t (mg·g−1), qe is the sorption capacity at the equilibrium (mg·g−1), and k1 is the rate constant of PFO (min−1).

Pseudo-second-order (PSO)

It assumes that the rate-limiting step is mediated by chemisorption (ion exchange, covalent forces, and/or sharing of electrons between sorbate and sorbent), which is related to sharing electrons between the surface of the sorbent and the sorbate. PSO equation is expressed according to Eq. 4:

$$\frac{d{q}_{t}}{dt}={k}_{2}{\left({q}_{e}-{q}_{t}\right)}^{2}$$
(4)

where k2 (g/mg·min) is the equilibrium rate constant of PSO. After the equation is integrated and considering the boundary conditions, the PFO equation is defined according to Eq. 5:

$$\frac{1}{{q}_{t}}=\left[\frac{1}{{k}_{2}{{q}_{e}}^{2}}\right]\frac{1}{t}+\frac{1}{{q}_{e}}$$
(5)

3 Results and discussion

3.1 Physicochemical characterization and elemental composition of biochars

From the preparation of the biochars, 250 g of rice husk biochar (BRH) was obtained with a yield of 22.85%, and 250 g of biochar from wild sugarcane (BWS) with a yield of 15.74%, reaching an average maximum temperature of 1070 °C and 673 °C respectively. Table 2 shows the physicochemical properties of the two biochars. BRH and BWS exhibit alkaline pH values of 9.63 and 10.6, respectively. A high CEC value allows the retention and exchange of positively charged ions, since it influences the mobility or retention of certain contaminants. In the case of BWS, the CEC value is high value compared to BRH (346 and 99 mmolc kg−1, respectively). Other studies [25, 26, 40] also reported similar pH and CEC results for biochar obtained by gasification. Furthermore, the difference between the properties of each biochar could be due to the different feedstocks and thermochemical decomposition method (oxidation rate, temperature, and thermal heating rate) since they have a substantial influence on the intrinsic properties of biochar and directly can affect the selectivity of sorbents in terms of efficiency for maximum sorbate removal [41, 42].

Table 2 Main properties, elemental analysis, and surface characteristics of biochars
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Table 2 shows the specific surface area of 197.64 and 16.03 m2 g−1 for BRH and BWS, respectively. Additionally, the macropore volume was 3.053 and 6.84 cm3 g−1 for BRH and BWS, respectively. However, BWS had a lower micropore volume than BRH. The N2 adsorption–desorption isotherms of BRH and BWS are displayed in Fig. 1. As seen in the graphs (Fig. 1), there is a considerable difference between both isotherms. BRH showed type IV isotherms related to mesoporous development, whereas BWS behaves as type V, which is also associated with macroporous or non-porous materials. Differences at low relative pressures (P/P0 < 1) indicated high development of microporosity in BRH compared to BWS.

Fig. 1
figure 1

N2 adsorption–desorption isotherms of biochars BRH and BWS

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3.2 Elemental analysis

The two biochars (BWS and BRH) showed different contents of oxygen (Table 2), which was higher for the BWS biochar (19.05%). Carbon content was also higher in BWS (54.13%) than in BRH (36.59%). The H/C, O/C, and (O + N)/C ratios were calculated in view of the values of C, H, O, and N in samples. BWS shows a higher H/C ratio (0.24), indicating lower aromaticity than BWS (H/C ratio of 0.09). Additionally, BWS showed slightly higher O/C and (N + O)/C values than BRH. Similar to the results obtained previously, James et al. [34] reported carbon, hydrogen, and nitrogen contents of 36.99%, 5.14%, and 0.58%, respectively, for rice husk biochar obtained in a TLUD gasifier. Similarly, Peterson and Jackson [43] used a TLUD gasifier to obtain biochar from pelletized wheat straw and corn stover with a carbon content of 74.04% and 40.66%, respectively. In addition, Hansen et al. [44] obtained less H/C and O/C atomic ratios in pine wood biochar than in wheat straw biochar, indicating increased dehydration and decarboxylation concurrently and high aromaticity and stability resulting from the carbonization process of pine wood.

3.3 FTIR analysis

FTIR analysis of two biochars was performed to characterize the surface oxygenated functional groups (Fig. 2). The two biochars show similar spectra with main differences related to the relative band intensity. The broad band centered at 3400 cm−1 can be attributed to -OH stretching vibrations of hydroxyl and carboxyl groups. The small bands at 2900 and 2850 cm−1 indicated that both materials had a low content on aliphatic structures. Their intensity was similar for the two biochars. The band centered at 1630 cm−1 can be ascribed to C = O vibrations in carboxylic, ester, lactones, or quinone functional groups, becoming more intense in BWS than in BHR. This result was according to high O/C content of BWS (0.26) than of BHR (0.19). Probably, the higher gasification temperature used for BHR leads to the loss of some oxygenated groups. The slight band at 1380 cm−1 can be related to C-O in phenolic and ether groups and the presence of C-N groups. This band is more intense for BWS. Additionally, the broadband between 1000 and 1200 cm−1 can be attributed to C-O bonds due to the presence of alcohols (C-O) and aliphatic ethers (C–O–C). Finally, Si–O stretching can be observed in this FTIR region, indicating the presence of SiO2.

Fig. 2
figure 2

FTIR spectra of biochars BHR and BWS

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3.4 XRD analysis

Figure 3 shows the XRD analysis of the two biochars, BRH and BWS. The intensity of the diffraction beam has been expressed as a function of 2 theta. As can be seen in the graph, BRH and BWS are amorphous in nature, showing a broad band between 15 and 25° characteristic in these materials. On the other side, BWS also exhibits the crystallographic planes 28 and 40° that can indicate the presence of calcite that does not decompose during gasification at 673 °C.

Fig. 3
figure 3

X-ray diffraction pattern of biochars BRH and BWS

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3.5 Zeta potential

Zeta potential (electro potential kinetics in colloidal systems) is the potential that is formed between the biochar and the medium. This potential provides important information about the behavior of the materials and stability in an aqueous solution. A large positive or negative Z potential indicates good stability of the suspensions due to repulsion between particles. Figure 4 depicts the zeta potential as a function of the pH of BRH and BWS. Both materials show a negative Z potential in the utilized pH range, highlighting that BWS presents the most negative Z potential in magnitude.

Fig. 4
figure 4

Zeta potentials of biochars BRH and BWS under different pH values

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3.6 Batch sorption kinetic studies

The kinetic study allows identifying the possible sorption mechanisms and the effect of the contact time in the process, which is necessary to evaluate the sorption efficiency of the dye. The influence of the biochars BRH and BWS on the removal of MG and As (III) was analyzed in terms of contact time at intervals of 10 min up to 120 min, at an initial concentration of 30 mg L−1 of the dye (MG), 250 mg of each biochar, room temperature (21 °C), constant agitation, and with an initial pH of 4.88 MG/BRH, 9.63 As/BRH, 9.95 MG/BWS, and 10.50 As/BWS for the solution.

Figure 5 shows that the equilibrium state was established at 120 min when the removal of MG and As (III) was kept constant. A 97.46% removal of MG dye was obtained using BWS, and a 61.99% removal of MG dye using BRH. Furthermore, it was observed that more than 90% of MG removal occurred in the first 10 min when using BWS, as opposed to using BRH. The great affinity of BWS by MG can be related to high functionality, high negative zeta potential, and high porous size (instead of lower total porosity). Similar to the results obtained, Rubio-Clemente et al. [40] obtained MG removal capacity of 84.60% at the equilibrium point (30 min) when using palletized pine wood biochar. Pathy et al. [22] obtained a removal efficiency of 80% MG in the first 60 min of the sorption process when using algae biochar. However, the equilibrium point was reached at 180 min with a removal efficiency of 80 to 85%. In contrast, Tsai et al. [45] reported a maximum removal capacity of MG dye of 22.86% after 120 min of reaction when using rice husk biochar. On the other hand, 82.79% and 82.36% removal of As (III) was observed using BRH and BWS, respectively. Previous studies reported similar results for arsenic removal efficiency, specifically arsenite. Sattar et al. [46] reported that sorption was achieved at 120 min (equilibrium point) with 90% removal when using peanut shell biochar. Also, Ali et al. [47] reported removing up to 84% of As (III) when using almond shell biochar.

Fig. 5
figure 5

Comparison of the removal efficiency of MG dye and As (III) by BRH and BWS at 120 min

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Table 3 shows the relevant kinetic parameters and correlation coefficients for MG and As (III) sorption on each biochar sample employing the pseudo-first-order (PFO) and pseudo-second-order (PSO) kinetic models. The linearization indicates the correlation coefficient (R2) values for the sorption kinetics fitted better to PSO (R2@BRH = 0.922 and R2@BWS = 1) than to PFO (R2@BRH = 0.828 and R2@BWS = 0.784) suggesting that the sorption of MG on both biochars occurred by the chemisorption mechanism. In the case of arsenite, the correlation coefficient (R2) values were best fitted to PSO, suggesting that arsenic sorption on both biochars also occurs by the chemisorption mechanism. Therefore, chemisorption was found to be the rate-determining step, which controls the sorption of MG, and as for As (III), possibly both mechanisms (chemisorption and physisorption) were exhibited on biochars.

Table 3 Kinetic parameters for MG dye and As (III) sorption on each biochars (BRH/BWS)
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The sorption process involves a combination of diverse types of interactions between biochar and sorbate. Electrostatic interaction, ion exchange, complexation, and precipitation are the main mechanisms involved in the sorption of inorganic contaminants such as arsenic [48]. For organic contaminants such as MG, different mechanisms involved in sorption are observed, such as hydrogen bonding interaction, pore filling, electrostatic interaction, and π-π interactions [48, 49]. In addition, the sorption mechanism depends mainly on factors such as porosity, specific surface area, pH value, and functional groups on the surface of biochar [50]. Different studies attributed that the functional groups on the surface of biochar (-OH, -COOH, thiol (-SH), phenolic) strongly attract As [51] and MG [24, 52] through the sorption mechanism.

Figure 6 represents the plot of t/qt versus t showing a linear relationship. Values of equilibrium sorption capacity qe and k2 were calculated from the intercept and slope of the plot. Figure 7 shows that the experimental sorption capacity for both samples fits better than theoretical sorption capacity for the case of PSO, indicating that the sorption process follows the pseudo-second-order kinetic model.

Fig. 6
figure 6

(t/qt) versus t. For a MG and b As (III), linearization of PSO

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Fig. 7
figure 7

Pseudo-second-order kinetic model. a MG. b As (III). Initial concentrations = 30 mg/L, 250 mg of biochars (BRH and BWS), 250 mL of volume, and room temperature

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3.7 FTIR and Z potential analysis of BWS samples after sorption of MG and arsenite

To verify the sorption mechanisms, FTIR analysis of BWS sample after sorption of MG and As (III) was performed (Fig. 8). BWS biochar showed an evident weakening in the band related to O–H bonds in As post-sorption material (BWSAs), related to carboxylic and hydroxylic groups (3000–3500 cm−1) that indicate the contribution of OH groups in the uptake of As (III). Additionally, the C = O stretching bands were shifted to higher wavelengths suggesting chemical interactions between the As (III) and the surface functional groups, such as surface complexation, precipitation, and electrostatic interactions [46, 53, 54]. Finally, the new bands at peaks at 1380 and 1070–1100 cm−1 were related to the adsorbed arsenite (AsO–H) [55]. Furthermore, as can be observed in Fig. 8, similarly, there is an evident weakening of the O–H band found between 3000 and 3500 cm−1 in the MG post-sorption biochar spectra (BWSMG). This fact indicates that there might be a strong sorption mechanism, such as a chemical bond through electron sharing between the nitrogen part of the amine functional group in the MG structure and the O–H functional group in the biochar [53, 56].

Fig. 8
figure 8

FTIR analysis post-sorption of BWSMG and BWSAs

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Table 4 shows the zeta potential and hydrodynamic diameter of BWS post-sorption of As (III) and MG. As observed, the Z potential of BWS biochar increases to less negative zeta potential values with the sorption of MG and As (III). This difference between the zeta potential of BWS and the post-sorption material suggests that MG and As (III) molecules had been attached to the BWS sorbent. It is noticed that hydrodynamic diameter increases; nevertheless, the polydispersity index is 1, which indicates that the samples are heterogeneous in size, which can be due to their agglomeration or aggregation.

Table 4 Z potential and hydrodynamic diameter analysis post-sorption of BWSMG and BWSAs
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4 Conclusions

Biochars obtained as by-products of rice husks and wild sugarcane gasification showed adequate characteristics to be used as sorbent of malachite green and arsenic from water solutions. The main conclusions are summarized as follows:

Sorption capacity values of biochars were 18.57 and 29.19 mg·g−1 for MG dye and 24.84 and 24.71 mg·g−1 for arsenite.

Both biochars showed fast sorption kinetics reaching equilibrium at 120 min and, in turn, resulted in a high sorption capacity, with an initial concentration of 30 mg·L−1 of both sorbates. The fits of the kinetic results indicated more accurately that the pseudo-second-order model best describes the sorption of two biochars, suggesting that the removal of MG dye and arsenite mainly supported the chemisorption. Moreover, according to post-sorption analysis, the functional groups -OH, C = O, and C = C were involved in the sorption of MG dye and arsenite on biochar. From the characterization data, it was observed that biochar obtained as by-product of wild sugarcane gasification is the biochar with the highest sorption capacity. This can be attributed to the combination of properties: functional groups, elemental composition, zeta potential, and cation exchange capacity.