Introduction

Hexavalent chromium, the most toxic form of chromium found in the environment, requires monitoring as a primary drinking water contaminant. Anthropogenic activities such as chromium plating, leather tanning, wood preserving, and alloy, brick, and dye making can generate Cr(VI) containing wastewater(Park et al. 2006), which endangers the safety of humans and animals as mutagen and potential carcinogen even at low concentration due to its bioaccumulation and biomagnification. According to the World Health Organization (WHO 2004) and the United States Environmental Protection Agency (USEPA 2011), the maximum allowable limit for typical Cr(VI) in drinking water is 0.05 mg/L, and the maximum permissible level of Cr(VI) in industrial wastewater is 0.1 mg/L in China (CEPB 2008), while the treated concentration in the industrial effluents is often about 0.5 g/L (Bertazzoli et al. 1997). Long-term exposure to Cr(VI) above allowable limit can cause damage to liver, kidney circulation, nerve tissue, dermatitis, and even death (Namasivayam and Kadirvelu 1999). Therefore, it is essential to diminish Cr(VI) before it enters into the aquatic environment.

Cr(VI) removal methods include reduction, filtration reverse osmosis, electrolysis, ion exchange, etc. (Uysal and Ar 2007; Karatepe et al. 2010; Feng et al. 2000). Biosorption shows its superiority in flexibility, reutilization, easy operation, low cost, and high efficiency especially for metals at low concentration (Klimmek and Stan 2001; Sud et al. 2008). A variety of natural and synthetic materials such as sawdust, chitosan, mycelium, and agricultural and industrial waste has been used as Cr(VI) adsorbents (Cadaval et al. 2013; Choi et al. 2009; Chen et al. 2011a; Javaid et al. 2011). Spent mushroom substrate is a residue after mushroom harvest, which is directly discarded or burned without comprehensive utilization. This biomaterial contains a variety of celluloses, hemicellulose, lignin, and chitin originated from mycelium and sawdust, on which sufficient function groups, such as hydroxy, carbonyl, carboxyl, amine, etc., are scattered. The large specific area and polyporous structure as well as abundant functional groups may endow spent mushroom substrate with high biosorption capacity (Garg et al. 2007). In our previous research, spent mushroom substrate has been used to remove Zn(II), Cu(II), and Cr(VI) from simulated wastewater (Vaca et al. 2001; Hu et al. 2014; Beveridge and Murray 1980).

As Cr(VI) mainly exists in its anion forms unfavorable for biosorption, it is imperative to improve the biosorption capacity of the adsorbents by surface modification. Cationic surfactants, such as alkyl quaternary ammonium, dodecyltrimethylammonium, and cetyltrimethylammonium, which can provide positively charged sites and boost anion exchanges, have been successfully used for surface modification (Chen et al. 2011b). Didodecyldimethylammonium bromide (DDAB) is a double-stranded molecular with both hydrophilic and hydrophobic groups that not only forms a double polarity layer on the surface enhancing wettability of biomaterials but also intercalates into the lamellar structure of the biomaterials supplying large amount of positively charged sites (Baek and Yang 2004; Fu and Qutubuddin 2001). The DDAB can bind to the nonpolar group of the adsorbent by hydrophobic interaction with its cationic head pointing at the solution, thereby increasing the number of cationic groups on the adsorbent surface and strengthening the biosorption capacity for anionic ions. In addition, it can also spontaneously develop mixed micelles, vesicles, and other self-assembly supramolecular structure in the solution, which can change the dispersion state of biomaterials, reduce the interfacial tension, and gather chromium ions on biomaterials (Zhang et al. 1993). DDAB could also augment the biosorption capacity of chromium by introducing the quaternary ammonium groups into the biomaterials (Wartelle and Marshall 2005). However, there is no report on DDAB modification of spent mushroom substrate used for anionic Cr(VI) removal.

In this study, Auricularia auricula spent substrate (AASS) was modified by cationic surfactant DDAB to increase the biosorption capacity for Cr(VI) in aqueous solution. To begin with, the biosorption conditions were investigated by a single-factor experiment. Subsequently, the key parameters and optimal biosorption conditions were screened out by Box-Behnken design coupled with response surface methodology analysis. Additional, biosorption mechanism was revealed by scanning electron microscope (SEM) and energy disperse X-ray (EDX), surface area analyzer, Fourier transform infrared (FTIR), X-ray diffraction (XRD), and X-ray photoelectron spectroscopy (XPS).

Materials and methods

Chemicals and reagents

All chemicals used in this study were analytical grade. The stock solution (1 g/L) of Cr(VI) was prepared by dissolving 2.827 g dichromate (K2Cr2O7) (Guangfu Fine Chemical, Tianjin, China) in 1-L deionized water, and work solution at desired concentration was prepared by diluting stock solution during the experiment. The initial pH was adjusted with 0.1 M NaOH (Kermel Chemical Reagent Co., Ltd., Tianjin, China) or 0.1 M HCl (Pilot Chemical Co., Shanghai, China) solution.

Preparation and modification of AASS adsorbents

AASS flotsam was dried overnight in an oven at 60 °C and ground into powder below 0.3 mm in size. Five grams of dried AASS were mixed with 100-mL 0.065 mol/L DDAB (Tokyo Chemical Industry Co., Ltd) solution in an orbital rotator at a speed of 150 r/min and 25 °C for 24 h. The modified AASS adsorbents were washed with distilled water to remove residual DDAB and dried at 60 °C to constant weight.

Exploration of biosorption mechanism

Different characterization methods were used to explore the biosorption mechanisms. Surface structures and elements of AASS and modified AASS were detected by a SEM (QuANTA200 model, American FEI, American) equipped with EDX analyzer (LEO 1455 VP/EDX Oxford 300, UK). The samples were coated with aurum by a vacuum electric sputter coater and observed at ×500 and ×1000 magnifications. For specific surface area, pore volume, and pore diameter analysis, samples were degassed with nitrogen at 423 K for 4–6 h and detected with a surface area analyzer (ASAP 2020, micromeritics, American). Particle size distribution of modified AASS was measured with a particle size analyzer (Mastersizer 2000, Malvern, America), with ultrapure water as a dispersant and placed a small amount of sample in the sample cell. It uses ultrasonic constantly stirring, which could disperse in water evenly. FTIR spectrometry was applied to determine the functional groups in charge of Cr(VI) biosorption on AASS, modified AASS, and metal-loaded modified AASS. Samples mixed with KBr at a mass ratio of 1:150 were pressed into a pellet and scanned in the range of 4000–400 cm−1 with a FTIR spectrometer (Alpha, Bruker, Germany). XRD patterns of modified AASS and metal-loaded modified AASS were obtained on a diffractometer (D/max2200 model, Hitachi, Japan) using CuKα radiation at a wavelength of 1.5418 Å at 40 kV and 30 mA and a scan rate of 4°/min in the 2θ region of 5–90°. The XPS (VG Scientific EACALab220i-XL) was performed with a spectrometer with Al Ka X-ray source (energy 1486.6 eV) with 100 W and pass energy 100 meV for general scan and 40 eV for core level spectra of each element. The software Peakfit 4.1 was used to deconvolute oxidation status of chromium in the XPS spectra.

Single-factor biosorption experiment

Single-factor biosorption experiment was conducted in a 250-mL Erlenmeyer flask containing 100-mL diluted Cr(VI) solution. Other factors were adjusted to suitable conditions: adsorbent dosage at 1, 2, 3, 4, and 5 g/L; pH at 2, 3, 4, 5, 6, and 7; initial concentration of Cr(VI) at 15, 20, 25, 50, 75, 100, and 125 mg/L; biosorption temperature at 20, 25, 30, 35, and 40 °C; biosorption time of 15, 30, 45, 60, 90, 120, 150, 180, 240, and 300 min; and rotation speed of 90, 120, 150, and 180 r/min respectively (Pavan et al. 2006; Shroff and Vaidya 2011). When the biosorption reached equilibrium, modified AASS adsorbents were separated by filtration and the remaining Cr(VI) concentration in supernatant was determined by atomic absorption spectrophotometer (AA-6800 model, Shimadzu-GL, Japan). All the biosorption experiments were performed in triplicate, and the average values were used in data analysis.

The biosorption efficiency (R) and the biosorption capacity (Q) of modified AASS at equilibrium were calculated using the following equations (Grimm et al. 2008):

$$ R=\frac{\left({C}_{\mathrm{j}}-{C}_{\mathrm{e}}\right)}{C_{\mathrm{j}}}\times 100\% $$
(1)
$$ Q=\frac{\left({C}_{\mathrm{j}}-{C}_{\mathrm{e}}\right) v}{m} $$
(2)

where C j and C e were the initial and the equilibrium concentration of the Cr(VI) solution, respectively (mg/L); V was the volume of reaction solution (we unified them for 100 mL) (mL); and m was the mass of adsorbent in the reaction solution (g).

Optimization of biosorption conditions by response surface methodology

On the basis of the single-factor experiment, three-factor and three-level Box-Behnken design coupled with response surface methodology was used to optimize biosorption conditions with biosorption capacity as the target (Ghanem et al. 2000; Kiran et al. 2007).

Biosorption isotherm model

Langmuir and Freundlich isotherm model was used to describe the equilibrium of Cr(VI) biosorption onto the modified AASS (Han et al. 2010).The biosorption isotherms were carried out at 293, 298, and 303 K, other conditions were set according to the results of the single-factor experiment, e.g., 100-mL solution at pH value of 4.0, rotation speed of 150 r/min, and biosorption time of 120 min, adsorbent dosage and initial concentration used in the isotherm studies ranging from 1-–5 g/L to 15-–125 mg/L, respectively. The equilibrium Cr(VI) concentration in the filtrate was determined by an atomic absorption spectrophotometer.

Langmuir isotherm equation (Langmuir 1918):

$$ {q}_{\mathrm{e}}=\frac{q_{\mathrm{m}} k\mathrm{L} c\mathrm{e}}{1+ k\mathrm{L} c\mathrm{e}} $$
(3)

where q e was the equilibrium biosorption capacity of Cr(VI) (mg/g), q m was the maximum biosorption capacity in theory (mg/g), c e was the equilibrium concentration of Cr(VI) solution(mg/L), and k L was the Langmuir equilibrium biosorption constant (L/mg).

Freundlich isotherm equation (Freundlich 1939):

$$ {q}_{\mathrm{e}}={K}_{\mathrm{f}}{c_{\mathrm{e}}}^{1/ n} $$
(4)

where K f represented the Freundlich constant associated with the biosorption capacity; n represented a constant relating to the biosorption intensity.

Desorption experiment

To recover the adsorbed Cr(VI) and to regenerate the adsorbent for further use, the desorption experiment was conducted with 0.2-g modified AASS loading with 20 mg/g of Cr(VI) in a 250-mL Erlenmeyer flask containing 100 mL of 0.2, 0.4, 0.6, and 0.8,1 M HCl at a speed of 150 r/min and 25 °C for 2 h. The reaction mixture was filtered, and the supernatant was used to determine the Cr(VI) concentration. The adsorbents were washed thoroughly by deionized water to pH 6 and reused for next cycle of biosorption. The biosorption-desorption process was continued for 3 cycles. The desorption efficiency was calculated using the following equation:

$$ \eta =\frac{c_2}{c_0-{c}_1}\times 100\% $$
(5)

where c 2 is the desorption Cr(VI) concentration; c 0 and c 1 are the initial and equilibrium concentration of the Cr(VI) solution, respectively (mg/L).

Results and discussion

Biosorption mechanism

BET and particle diameter analysis

Specific surface area, pore volume, and diameter were the crucial parameters that influence biosorption. It could be seen from Table 1 that native AASS itself had a large specific surface area about 50 m2/g and average pore diameter of 48 Å. The interaction of these micropores and mesopores on the surface of adsorbents resulted in a high biosorption capacity. While the surface modification enlarged the specific area of AASS that could be seen clearly from Figs. 1 and 2, the surface area and the average pore diameter of AASS were increased 14.7 and 11.9%, respectively. As AASS is a composite material, its composition is quite complex, so the particle diameter curve was relatively sharp and the particle size distributes in a wide range from 10 to 1000 μm, especially between 100 and 400 μm as shown in Fig. 3. Since the cationic surfactant also had a scavenging effect to remove impurities on the surface and inside of the spent substrate so that the diffusion channels were smooth and provided sufficient active sites for the biosorption of chromium ions. On the one hand, the specific organic bimolecular layer of DDAB macromolecules played a role in metal ion biosorption by ion exchange; on the other hand, it could provide a lot of internal biosorptions (Ren et al. 2016).

Table 1 BET characteristics of AASS and modified AASS
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Fig. 1
figure 1

Native AASS

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Fig. 2
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Modified AASS

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Fig. 3
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Particle size of modified AASS

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SEM-EDX analysis of biomass

Scanning electron microscopy observation could clearly reveal the morphological characteristics of the chemically modified and metal-loaded adsorbents (Akar et al. 2013). Figures 1, 2, and 4 showed the microstructures of native AASS, modified AASS, and metal-loaded modified AASS, respectively. The native spent substrate appeared to be irregularly lamellar and porous structure on the surface of fragmentary sawdust in Fig. 1, which provided a lot of pores and a large surface area for chromium entering or landing. In addition, honeycomb micropores on the sawdust components including cellulose, hemicellulose, and lignin had comparatively stronger biosorption capacity (Li et al. 2007). For DDAB-modified adsorbents, a more regularly ridge-like structure with much more pores inside was observed in Fig. 2. The modification procedure might remove organic and mineral matter as well as sweep away the debris covered on the surface or inside the macropores, which enlarged the surface area, increased the pore quantity and cubage, deepened the ravines, and exposed more sorption sites embedded in adsorbents, thus enhancing metal uptake. The molecules of DDAB could enter into the space between the layers of xylon slice of the spent substrate via their long alkyl chains, which boosted the biosorption potential for chromium ions (Fu and Qutubuddin 2001). After biosorption, a lot of luminous particles were found to distribute on the surface of adsorbents under greater magnification (×1000) and distinctly crystalline lamellar structures were viewed to accumulate on the ridge or inside the pores under greater magnification (×5000) on the right bottom of the image as shown in Fig. 4.

Fig. 4
figure 4

Metal-loaded modified AASS

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Further evidence of chromium ions onto the modified AASS after biosorption came from EDX spectra and elemental composition analysis as shown in Fig. 5a–c and Table 2. In Fig. 5a, b, it was found that both metal-loaded and unloaded adsorbents had element peaks such C, N, O, S, P, Si, Cr, and other metal cations. Table 2 illustrated the variation of elements. The incremental amount of N from 15.15 to 16.04 wt% implied that DDAB modification might introduce amine-based groups in AASS. Furthermore, the increased C content might be attributed to the insertion of alkyl chains in the adsorbents. The distinct Cr characteristic peak in Fig. 5c and increased Cr content between metal-unloaded and metal-loaded adsorbents substantiated that Cr(VI) had been adsorbed onto the modified AASS. While the decreased content of Ca, Na, Mg, and Al proved the occurrence of ion exchange in the biosorption process with reduced trivalent chromium. Bueno et al. had reported the possibility of ion exchange between K+ and Cr3+ in the biosorption process by EDX analysis (Bueno et al. 2008).

Fig. 5
figure 5

SEM–EDX characterization of native AASS (a), modified AASS (b), and metal-loaded (c) of modified AASS

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Table 2 Elemental composition analysis of native AASS, modified AASS, and mental-loaded modified AASS
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FTIR analysis

FTIR spectroscopy was applied to identify the available functional groups on the adsorbent as well as confirm the newly emerged functional groups on the modified adsorbent (Sujatha et al. 2013). The main functional groups and the corresponding wave numbers could be identified by FTIR after comparing with the infrared spectra of other biomass in the literatures (Schmitt and Flemming 1998). As shown in Fig. 6, the peaks at 3340–3330 cm−1 suggested that the stretching vibrations of hydroxyl (−OH) and amine groups (−NH). It was found that the position migrated from 3340 to 3330 cm−1, indicating that the hydrophilicity and hydrophobicity of the adsorbent surface changed after modification. The peak shape was widened and the intensity was weakened, which indicated that the hydroxyl and amine groups might be involved in the Cr(VI) biosorption (Boddu et al. 2008). Two peaks at 2920 and 2850 cm−1 appeared in the modified AASS adsorbents presented the asymmetric vibrations of alkynes and alkyl chains, respectively. After biosorption, the two peaks became more sharp indicating these two functional group participated in chromium binding (Ramrakhiani et al. 2011; Naja et al. 2005). Thus, upon Cr(VI) reaction, the peaks at 1620–1650 cm−1 were assigned to the formation of C=O of carbonyl and carboxyl group vibration in the adsorbents. After modification, the peak at 1620 cm−1 shifts to 1650 cm−1 and the strength slightly weakens after biosorption, indicating the carboxyl and carbonyl groups might further provide the binding sites for Cr(III) resulting from Cr(VI) reduction (Maleki et al. 2015). The sharp peaks between 1400 and 1500 cm−1 were due to the presence of the carboxyl groups (Kazy et al. 2002). Following Cr(VI) biosorption, the spectra exhibited a change in the peak position indicating the strong role of carboxyl groups in Cr(VI) binding. The intensity of peak at 1320 cm−1 got weaker after biosorption, indicating that the carboxyl groups also contributed to the formation of a complex with Cr(VI) (Li et al. 2013). The peak at 1240 cm−1 was attributed to the presence of carboxyl and that at 1030 cm−1 represented a P-OH stretching vibration which was in accordance with change of the P content in EDX (Kazy et al. 2006; Akar et al. 2012). In the chromium-loaded adsorbents, the appearance of a new peak at 930 cm−1 might be assigned to stretching vibrations of weakly bonded oxygen ligands with chromium. Additionally, Nagy et al. reported that the peaks identified in the less than 900 cm−1 region could be assigned to bending modes of aromatic compounds (Nagy et al. 2013). The peak at 789 cm−1 might be assigned to the out-of-plane CH bending vibrations of lignin. In summary, DDAB modification changed functional groups, the carboxyl, amine groups, hydroxyl, alkyl chains, and phosphate groups played primary roles in metal ion uptake in addition to ion exchange.

Fig. 6
figure 6

FTIR spectrometer of AASS (a), modified AASS (b), and after Cr(VI)absorption (c)

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XRD

According to the principle of X-ray diffraction, the diffraction peak of the crystalline phase was sharp and the peak of the amorphous phase was obviously broadened. XRD pattern of the adsorbents loaded with Cr(VI) was not apparently different from that of the unloaded. The diffraction peaks at 22.03° of modified AASS and 26.27° after biosorption represented SiO2, indicating that the active sites and the skeleton of SiO2 were fully exposed as shown in Fig. 7 (Schroden et al. 2002). SiO2 played a strong complexation and coordination role on metal ions, and EDX analysis also revealed that the content of Si was changed. There was no absorption peak of the chromium ion compound before the biosorption, but the characteristic peaks of crystals containing chromium in XRD patterns were found after biosorption. More sharp absorption peaks at 10–35°and the appearance of crystalline state after biosorption illustrated that the adsorbents and chromium ions formed the crystal structure of metal compounds. After biosorption, retrieval analysis was performed with the MDI Jade 5.0 software, and the major crystalline species were displayed in Table 3 by comparing with XRD standards in the PDF Card. The formation of these compounds indicated that a complexation reaction occurred between chromium ions and adsorbents.

Fig. 7
figure 7

X-ray spectra of modified AASS (a) before and (b) after Cr(VI) absorption

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Table 3 The major crystalline species after adsorption
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XPS

Figure 8a illustrates the XPS of Cr 2p of modified AASS after Cr(VI) biosorption. The high-resolution Cr 2p spectrum of the Cr-loaded modified AASS showed that both Cr2p3/2 and Cr2p1/2 were presented at 575–580 and 585–590 eV, respectively (Vankar et al. 2013). The binding energy of 2p3/2 of Cr in Na2CrO4, Na2Cr2O7, K2Cr2O7 was reported to be in the range of 578.4–578.9 eV (Park et al. 2008). The binding energies 578.68 and 588.88 eV were obtained after XPS-peak-differentiation-imitating analysis of Cr 2P3/2, which represented Cr(VI) and Cr(III), respectively. The binding energy of Cr(III) was about 578 eV, and the binding energy of Cr(VI) was generally higher than that of Cr(III) (Watts et al. 2015; Liu et al. 2000). Thus, Cr(VI) adsorbed on modified AASS underwent a reduction reaction to produce trivalent chromium. The ratio of adsorbed Cr(III) to Cr(VI) onto the modified AASS surface was about 3:1 according to the 2p3/2 peak area. Since Cr(III) was a cationic metal ion, it could react with the hydroxyl or carboxyl functional groups on the surface of the biomass adsorbent. There were two peaks in the O1s XPS spectrum of the modified AASS after biosorption of Cr(VI) (Fig. 8b); the binding energy at 531.68 eV was oxygen in the OH or C=O (lipid, anhydride, amide) bounded to the metal ion (Manning et al. 2006). In Fig. 8c, a peak, appearing at 284.82 eV, was the carbon in the C=C or C-H (Wang et al. 2009). The symmetrical peaks in the N1s XPS spectra are shown in Fig. 7d, the binding energy at 401.58 eV is N in the quaternary ammonia and NH4 +, and that at 399.73 eV was the deprotonation peak of N1s in the NH (Pels et al. 1995).

Fig. 8
figure 8

Cr 2p, O1s, C1s, N1s (a–d) high-resolution XP spectrum of modified AASS after Cr(VI) biosorption

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Parameters affecting Cr(VI) biosorption in single-factor experiment

Effect of the adsorbent dosage

Adsorbent dosage determines the number of active sites for metal binding, so it greatly affects the biosorption capacity and efficiency. In Fig. 9, when the adsorbent dosage increased from 1 to 5 g/L, the biosorption capacity decreased from 16.92 to 3.966 mg/g and the biosorption efficiency escalated from 84.63 to 99.16%. The same trend was reported by Onundi et al. (2010). The reason for the increment of biosorption efficiency was that the amount of active sites was proportionally augmented with the increase of adsorbent dosage resulting in additional Cr(VI) binding or ion exchange. When the adsorbent dosage continued to increase, the polymerization and overlap between the adsorbents occurred simultaneously, obstructing the effective sites for Cr(VI) binding or ion exchange (Garg et al. 2003). Therefore, in the following experiment, 2 g/L was selected as the optimum adsorbent dosage.

Fig. 9
figure 9

Effect of adsorbent dosage on Cr(VI) biosorption (25 °C, 150 r/min, contact time of120 min, pH 4,initial Cr(VI) concentration of 20 mg/L)

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Effects of solution pH

The pH was a key factor to influence the state of chromium ions in the solution and the quantity of electric charges on the modified AASS surface. Biosorption capacity varying with pH is shown in Fig. 10; the optimal biosorption occurred at pH 4 due to the electrostatic interaction between positive charges and chromate anions. When the pH rose from 2 to 4, the biosorption capacity increased from 8.659 to 9.753 mg/g. The deactivation of functional groups under acidic circumstance weakened their binding affinities with chromium ions (Khormaei et al. 2007). When pH was higher than 4, the biosorption capacity gradually decreased to the minimum value of 7.839 mg/g. The increased amount of OH− could neutralize some positive charges on the adsorbents and correspondingly increased the electronegativity (Inbaraj and Sulochana 2006). Simultaneously, OH− could compete with Cr2O7 2− for the functional groups on the adsorbent, which reduced biosorption capacity as well. Therefore, pH 4 was favorable for Cr(VI) removal by modified AASS, and the following biosorption experiments were conducted at this pH value. Moreover, on the basis of the optimum pH of 4, the solution pH rose to 5.16 after biosorption, which might be assigned to the redox reaction between HCrO4 − and H+ to form Cr(III).

Fig. 10
figure 10

Effect of solution pH on Cr(VI) biosorption (25 °C, 150 r/min, adsorbent dosage of 2 g/L, contact time of 120 min, initial Cr(VI) concentration of 20 mg/L)

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Effect of initial Cr(VI) concentration

Figure 11 showed that metal uptake went up with the increasing initial Cr(VI) concentration. The biosorption capacity of the adsorbent for Cr(VI) increased from 7.301 to 35.39 mg/g, and biosorption efficiency reduced from 97.34 to 56.63% with initial concentrations ascending from 15 to 125 mg/L. Furthermore, when the initial concentration exceeded 75 mg/L, the upward trend slowed down. This was because that a finite number of surface-functional groups on saturated adsorbents could only accommodate the amount of Cr(VI) ions at this concentration (Zheng et al. 2008). It was noticed that the higher the initial metal concentration was, the lower biosorption efficiency was obtained. Hence, the optimal initial Cr(VI) concentration was set at 20 mg/L.

Fig. 11
figure 11

Effect of initial Cr(VI) concentration (25 °C, 150 r/min, adsorbent dosage of 2 g/L, contact time of120 min, pH 4.0)

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Effect of biosorption temperature

Figure 12 illustrated the effect of temperature on biosorption capacity and efficiency of Cr(VI). As the temperature gradually increased, the biosorption capacity presented upward trend first and then went down. Neither low temperature (20 °C) nor high temperature (40 °C) was favorable for Cr(VI) biosorption. When temperature increased from 20 to 25 °C, the biosorption capacity and efficiency were increased speedily to 9.727 mg/g and 97.27%, and then the maximum capacity and efficiency of 9.746 mg/g and 97.46% were observed as temperature decreased to 30 °C. Moderate temperatures were profitable for a quick and irreversible binding between Cr(VI) and adsorbent. With temperature ascending from 30 to 40 °C, both biosorption capacity and efficiency dropped off dramatically. The reason might be explained that high temperature promoted the molecular thermodynamic movement and the collision between chromium ions and adsorbent. Besides, the nonfunctional groups on the adsorbent would be activated at high temperature. When it reached 40 °C, the biosorption capacity decreased to 8.561 mg/g; it might accelerate solute movement and triggered physical collision and desorption process (Malkoc et al. 2006). Moreover, biosorption operated at high temperature would consume more energy, and extra high temperature would cause the transition of DDAB from the gel phase to the liquid crystal phase and subsequently decreased the hydrophilicity of adsorbents (Walde 2006). Therefore, 25 °C was taken as the appropriate biosorption temperature.

Fig. 12
figure 12

Effect of temperature on Cr(VI) biosorption (adsorbent dosage of 2 g/L, 150 r/min, pH 4, contact time of 120 min, initial Cr(VI) concentration of 20 mg/L)

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Effect of biosorption time and effect of rotating speed

The influence of time on the adsorption of Cr(VI) onto the modified AASS was illustrated in Fig. 13; the biosorption capacity increased with the lengthening of the time, but the upward trend slowed down gradually, which could be divided into three stages. In the first 120 min, the biosorption capacity increased rapidly from 7.643 to 9.647 mg/g, indicating that the initial biosorption was a rapid process that might be due to the relatively high biosorption sites on the surface of modified AASS. When the contact time was during 120 to 240 min, a slow uptrend was observed that the biosorption sites on the surface of the adsorbent have been saturated, and the trend of the biosorption capacity growth was not obvious. Shortly afterwards, the biosorption process stepped into the third stage, the biosorption capacity and efficiency decreased, possibly because the biosorption reached the equilibrium. So the optimal biosorption time was set to 120 min. Compared with other parameters, it was noticed that the effect of the rotation speed was inconspicuous at optimal biosorption time. With the increase of the rotation speed from 60 to 180 r/min, the biosorption capacity of Cr(VI) was enhanced only by 13.7% because of the diminution of mass resistance as shown in Fig. 12. While the biosorption capacity tended to be stable with the rotational speed rose to around 130 r/min, indicating the influence of the mechanical oscillation was not significant in Cr(VI) adsorption. The oscillating process could provide more opportunity for metal ions to contact with the adsorbent, but might also damage the structure of adsorbent under vigorous vibration. Therefore, we selected 150 r/min as the most appropriate speed.

Fig. 13
figure 13

Effect of temperature on Cr(VI) biosorption and effect of rotation speed on Cr(VI) biosorption (adsorbent dosage of 2 g/L, pH 4, contact time of 120 min, initial Cr(VI) concentration of 20 mg/L)

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Optimization of biosorption conditions by response surface methodology

The results of single-factor experiment showed that the maximum biosorption capacity of 9.796 mg/g was obtained at pH 4.0, temperature of 25 °C, biosorption dosage of 2 g/L, initial concentration of 20 mg/L, biosorption time of 120 min, and rotation speed of 150 r/min. Under the same conditions, the biosorption of the native AASS was 6.058 mg/g. All these parameters were taken as variables for optimization of biosorption conditions, and the biosorption efficiency was taken as the response value of the variables on the model of the significant situation. Each factor was divided into three levels to determine the significant conditions of the test, 0 as the optimum condition from the single-factor experiment; +1 as the high level; −1 as the low level. As shown in Table 4, adsorbent dosage (A), pH (B), and initial Cr(VI) concentration (C) were the key parameters. The p < 0.0001 indicated that the quadratic model used in the experiment was remarkable and significant in statistics (Rastogi and Rashmi 1999). Lack of fit represented the difference between the model function and the real function, namely the extent of the difference. In this case, p was 0.0554 > 0.05, which was beneficial to the model. The coefficient of variation (CV) of 1.09% indicated that the data dispersion was normal. The p values of factors A (0.0001) and C (0.0017) were less than 0.005 that illustrated that the initial concentration of Cr(VI) solution on the biosorption efficiency was significant. Both p values of B2 and C2 were less than 0.0001, indicating an observable effect on the biosorption efficiency. The p value of the interaction term AB and AC was 0.6002 and 0.2183, respectively, only BC was less than 0.001, so the interaction term had a slight effect on biosorption efficiency. Hence, the regression analysis and the regression fitting were carried out using Design-Expert 8.05b to obtain the regression equation for the biosorption of Cr(VI):

$$ Y=80.22+39.29 A+4.37 B+1.450 C+1.867 AB+0.2923 AC-0.1289 BC-53.6{A}^2-0.8532{B}^2-0.0663{C}^2 $$
Table 4 Results of Box-Behnken design
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where A represents modified adsorbent dosage (g/L), B represents pH, and C represents the initial Cr(VI) concentration (mg/L).

The correlation coefficient R 2 provides a measure of the model’s variability in the observed response values. The closer the R 2 value to one, the stronger the model is, the better the response is predicted (Kumar et al. 2010). In this study, the correlation coefficient R 2 of 0.9872 showed that the regression was significant, and the equation fitted well. Through variance analysis, we could see that there was a high correlation between the predictive value and the measured value. The optimum conditions were as follows: adsorbent dosage of 1.5 g/L, initial Cr(VI) concentration of 18.97 mg/L, pH 4.01 (for actual operation, the optimal initial concentration and pH were set to 19 mg/L and 4, respectively). Under these conditions, the biosorption capacity reached 12.23 mg/g. Subsequently, experiments were carried out under these conditions and an average biosorption capacity of 12.16 mg/g (very close to the theoretical value) was obtained. The biosorption capacity was surface curve of each factor was obtained as shown in Fig. 14.

Fig. 14
figure 14

3D surface and contour

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Biosorption isotherm of Cr(VI)

Figure 15a, b plotted the removal of Cr(VI) biosorption onto the modified AASS based on the Langmuir and Freundlich isotherm models, respectively. The obtained equilibrium data were fitted for Freundlich isotherm models. Table 5 describes parameters b, q m, K f, and n obtained from the two models at different temperatures. By comparing the regression coefficients (R 2) of the two models, it was found that the experimental data followed the Freundlich model better. The Freundlich model represents physical biosorption that was empirically applicable for adsorbents with heterogeneous surfaces in favor of multi-molecular layer biosorption. The Freundlich constants (n) were related to the biosorption capacity and the biosorption efficiency, respectively. The n value in the range of 1 to 10 indicated that the biosorption strength was large and in favor of physical process, while n = 1 Freundlich equal Henry linear adsorption isotherm and n value less than 1.0 illustrated that biosorption was difficult to proceed (Uçer et al. 2006; Frantz et al. 2017; Khalfa et al. 2016). As shown in Table 5, all the n values approached to 3 which indicated that this biosorption process was easily carried out.

Fig. 15
figure 15

q e and c e curve at different temperatures (initial Cr(VI) concentration of 15–125 mg/L)

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Table 5 Parameters between two isotherm models at different temperatures (initial Cr(VI) concentration of 15–125 mg/L)
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As for the Langmuir isotherm model, the equation was derived from the assumption that monolayer biosorption on specific homogenous sites would not affect biosorption at an adjacent site (Jalali-Rad et al. 2004). From Langmuir biosorption isotherm, the calculated maximum biosorption of modified AASS for Cr(VI) was 28.10 mg/g at 303 K. A comparison of the adsorbent capacity between various adsorbents and modified AASS is given in Table 6 (Li et al. 2003; Mahmood-ul-Hassa et al. 2015; Babu and Gupta 2008; Arivoli et al. 2008). In the present study, it was found that the adsorption capacity of Cr(VI) onto the modified AASS was 12.16 mg/g at 20 mg/L of Cr(VI) initial concentration, 1.5 g/L of adsorbent dosage, and pH 4. Compared with other agriculture waste adsorbents such as corn cob, maple saw dust, sunflower, banana stalks, and AASS, the modified AASS had higher adsorption capacity in almost all cases, and 34.97 mg/g was the maximum. Generally, an adsorbate diffuses to a solid surface and weakly binding with the adsorbents by intermolecular forces in the adsorption process. The electrostatic, chemical bond, and functional group affect the affinity of an adsorbent for a specific adsorbate. Chemical modification improved the adsorption capacity of adsorbents by introducing active binding sites and forming new functional groups that favored Cr(VI) ions biosorption. The experimental result suggested that ionic exchange and complex formation were one of the major mechanisms for Cr(VI) adsorption onto the modified AASS. The adsorption capacity of neem leaves was higher than that of the modified AASS, neem leaves were activated with concentrated hydrochloric acid, which increased the specific surface area and provided an acidic surface favorable for Cr(VI) adsorption. Figure 16a, b plotted the effect of adsorbent dosage on Cr(VI) biosorption onto the modified AASS based on the Langmuir and Freundlich isotherm models, respectively. Table 7 described the same parameters obtained from the two models at different temperatures. At initial Cr(VI) concentration of 20 mg/L and adsorbent concentration in the range of 1 to 5 g/L, the experimental data followed the Freundlich model better according to the higher R 2 values at all tested temperatures. The n values indicated that this biosorption was easy to carry out just similar as the result in the initial concentration. With the change of temperature, the biosorption might undergo a physical process at first and then entered into the chemical process. In this study, both monolayer biosorption and complex multi-layer biosorption occurred in the biosorption process. Enhancement of biosorption capacity of modified AASS was possibly due to its uniform pore size distribution, which could enhance Cr(VI) accessibility to the active sites, and the highly positive charged property endowed the surface a strong affinity to anions.

Table 6 Adsorption capacity of various adsorbents
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Fig. 16
figure 16

q e and c e curve at different temperatures (adsorbent concentration of 1–5 g/L)

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Table 7 Parameters between two isotherm models at different temperatures (adsorbent concentration of 1–5 g/L)
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Desorption results

The desorption and regeneration of adsorbent were momentous to the application of biosorption technology. It could be seen in Fig. 17 that the desorption rate was the most appropriate when HCl concentration was 0.6 M, and the desorption rate for each cycle was 73.14, 68.56, and 62.33%, respectively. On the one hand, HCl could change the pH of the solution and increase the solubility of Cr(VI), so that the adsorbed chromium ions were easily eluted out; on the other hand, the variation of pH value could also change the protonated state of the functional groups and the electrostatic force of the adsorbent surface, thus undermining the biosorption balance to achieve the purpose of desorption.

Fig. 17
figure 17

Desorption efficiency for each cycle

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Conclusion

In conclusion, we had demonstrated that modified AASS represented an effective material for the biosorption of Cr(VI) in low concentration systems. The optimum conditions of key parameters for Cr(VI) biosorption screened out by response surface methodology were adsorbent dosage of 1.5 g/L, initial Cr(VI) concentration of 19 mg/L, pH value of 4.0, biosorption time of 120 min, rotation speed of 150 r/min, and temperature of 25 °C, under which biosorption capacity could reach 12.16 mg/g and was increased by 2 times relative to that of the unmodified adsorbent. DDAB modification could provide hydrophilic and hydrophobic groups and increase specific surface area of the adsorbent. Furthermore, the ion exchange, complex formation, and Cr(VI) reduced to Cr(III) could also be confirmed by FTIR, SEM-EDX, BET, XRD, XPS, and particle diameter analyses. Both Langmuir and Freundlich isotherm equation could describe the biosorption process of Cr(VI) onto the modified AASS. And after three biosorption-desorption cycles, the final regeneration efficiency was 62.33%.