Applied Microbiology and Biotechnology

, Volume 96, Issue 3, pp 829–840

Biosorption behavior and mechanism of heavy metals by the fruiting body of jelly fungus (Auricularia polytricha) from aqueous solutions

Authors

  • Haiwei Huang
    • School of Environmental Science and EngineeringSun Yat-sen University
    • Guangdong Provincial Key Laboratory of Environmental Pollution Control and Remediation Technology
  • Lixiang Cao
    • School of Life SciencesSun Yat-sen University
  • Yuxuan Wan
    • School of Environmental Science and EngineeringSun Yat-sen University
    • Guangdong Provincial Key Laboratory of Environmental Pollution Control and Remediation Technology
    • School of Environmental Science and EngineeringSun Yat-sen University
    • Guangdong Provincial Key Laboratory of Environmental Pollution Control and Remediation Technology
  • Wenfeng Wang
    • School of Life SciencesSun Yat-sen University
Environmental biotechnology

DOI: 10.1007/s00253-011-3846-6

Cite this article as:
Huang, H., Cao, L., Wan, Y. et al. Appl Microbiol Biotechnol (2012) 96: 829. doi:10.1007/s00253-011-3846-6

Abstract

The aim of this study was to investigate the biosorption characteristics of Cd2+, Cu2+, and Pb2+ by the fruiting body of jelly fungus Auricularia polytricha. Batch experiments were conducted to characterize the kinetics, equilibrium, and mechanisms of the biosorption process. Optimum values of pH 5, biomass dosage 4 g L−1, and contact time 60 min provided maximum biosorption capacities of A. polytricha for Cd2+, Cu2+, and Pb2+ of 63.3, 73.7, and 221 mg g−1, respectively. The maximum desorption was achieved using 0.05 mol L−1 HNO3 as an elute. The fruiting body was reusable at least for six cycles of operations. The pseudo-second-order model was the best to describe the biosorption processes among the three kinetic models tested. Freundlich and Dubinin–Radushkevich models fitted the equilibrium data well, indicating a heterogeneous biosorbent surface and the favorable chemisorption nature of the biosorption process. A Fourier transform infrared spectroscopy analysis indicated that carboxyl, amine/hydroxyl, amino, phosphoryl, and C–N–C were the main functional groups to affect the biosorption process. Synergistic ion exchange and surface complexation were the dominant mechanisms in the biosorption process. The present work revealed the potential of jelly fungus (fruiting body of A. polytricha) to remove toxic heavy metals from contaminated water.

Keywords

BiosorptionJelly fungusAuricularia ploytrichaHeavy metals

Introduction

The contamination and redistribution of heavy metals have created environment and health hazard problems (Villaescusa et al. 2004). The conventional technologies for heavy metal removal, including precipitation, chemical oxidation or reduction, electrochemical treatment, reverse osmosis, ion exchange, and evaporation, are ineffective and/or costly, especially when concentrations of heavy metals in aqueous solution are as low as 1–100 mg L−1, and produce a large amount of sludge (Volesky 2001). Therefore, it is necessary to explore more efficient, cost-effective, and environment-friendly methods for heavy metal removal from contaminated water (Veglio and Beolchini 1997).

In recent years, the biosorption method has gained increasing attention due to its low cost, high efficiency, minimum chemical use and sludge products, and its capability to remove heavy metals from dilute solutions (Foo and Hameed 2010; Gadd 2009; Volesky 2007). Therefore, cheaper and effective metal biosorbents have been developed, including bacteria (Vijayaraghavan and Yun 2008), fungi (Du et al. 2009; Pan et al. 2009, 2010; Yan and Viraraghavan 2003), and algae (Davis et al. 2003).

Nevertheless, few studies have focused on the use of macrofungi as biosorbents. Several macrofungi, including Ganoderma lucidum (Muraleedharan et al. 1995), Pycnoporus sanguineus (Zulfadhly et al. 2001), and Agaric fungal species (Mathialagan et al. 2003; Vimala and Das 2009), exhibit high biosorptive potentials for selected heavy metals. One of the advantages of using macrofungi as biosorbents is their relatively prolific production and low cost in many parts of the world. Furthermore, their large biomass, tough texture, and other favorable physical characteristics are the desirous properties as biosorbents (Matheickal and Yu 1997). Auricularia polytricha belongs to jelly fungi, a set of species from different taxonomical groups of Phragmobasidiomycetes, which form gelatinous fruiting bodies. The fruiting bodies are able to survive for a long period of drought. With available water, the dried horny texture of the fungi becomes gelatinous because of the polysaccharides that compose 60–70% of the dry fruiting body. Some polysaccharides have shown chelating effects on metallic ions (Mau et al. 2001). The polysaccharide gel matrix of brown algae, such as alginate, is responsible for the higher metal uptake compared with other algae, bacteria, and fungi (Davis et al. 2003; Mata et al. 2009). Alginates are linear polysaccharides composed of (1,4)-linked mannuronic acid and guluronic acid (Cao et al. 2007). The polysaccharide glucuronoxylomannan of jelly fungi consists of a linear backbone of 1,3-linked alpha-d-mannose with mainly xylose and glucuronic acid in side chains (Khondkar et al. 2002; Sun et al. 2010). It is feasible to apply jelly fungi for wastewater treatments (Davis et al. 2003). However, the extent and mechanisms of biosorption of metals by jelly fungi are still unknown.

Pan et al. (2010) reported the biosorption behavior of dried and humid fruiting body of Tremella fuciformi and A. polytricha. While the biosorption rates for the other treatments decreased with increased metal concentrations, the biosorption rates for the treatment of dried A. polytricha eventually increased at higher metal concentrations. In addition, the mechanical strength of the fruiting body of A. polytricha is stronger than that of T. fuciformi. Therefore, it is worthy to further examine the biosorption characteristics of the fruiting body of A. polytricha. The objective of this study was to investigate the biosorption characteristics of Cd2+, Cu2+, and Pb2+ by the fruiting body of A. polytricha related to the physicochemical properties, including surface area, total pore volume, the point of zero charge, and cation exchange capacity (CEC). First, we determined the optimum conditions (i.e., optimum pH, contact time, biosorbent dosage, and heavy metal concentration) for the maximum biosorption yields of the heavy metals by the fruiting body. Secondly, we tested the reusability of the fruiting body as a biosorbent through multiple cycles of biosorption–desorption. Finally, we explored the biosorption mechanisms of the biosorbent using Fourier transform infrared spectroscopy (FTIR), ion exchange experiments, and scanning electron microscope (SEM) imaging.

Materials and methods

Materials

Fresh fruiting body of A. polytricha (red ear, Agricultural Culture Collection of China, ACCC50140) was washed with tap water and then deionized water for three times and dried at 40°C until reaching a constant weight. The sample was then powdered and sieved to yield biomass with particle sizes of 0.471–0.985 mm.

Stock metal solutions (1,000 mg L−1 for each solution) were prepared by dissolving appropriate amounts of Cd(NO3)2⋅4H2O, Cu(NO3)2⋅3H2O, and Pb(NO3)2 in deionized water. Solutions of NaOH and HNO3 (0.1 mol L−1 each) were used for pH adjustment.

Physical properties of the biomass

The specific surface area and total pore size distribution were measured using the N2-BET method with the accelerated surface area and porosimetry system (ASAP 2010, Micromeritics, USA). The point of zero charge (PZC) for the fruiting body was determined using the method of Leyva-Ramos et al. (2005), with some modifications as follows: 0.5 g of the fruiting body was added into 10 mL CO2-free water in a 25-mL Erlenmeyer flask. The flask was sealed with a rubber stopper and shaken at 150 rpm for 48 h at 25°C. Then, the solution pH was measured with a pH meter (Mettler Toledo, FE 20, Switzerland) and the pH value taken to be the PZC. This method has been used satisfactorily by Leyva-Ramos et al. (2005).

Concentrations of Ca2+, Mg2+, Na+, and K+ released from acid-washed biomass were measured to determine the ion exchange capacity of the biomass. Fruiting body of 1.0 g was treated with 50 mL of 0.1 mol L−1 HCl. The suspension was stirred for 1 h at 25°C and then filtered through a double-circle filter paper. The fruiting body was resuspended in the same HCl solution. The above procedure was repeated four times. One gram of the fruiting body treated with 50 mL deionized water was used as the control. The filtrates were analyzed for concentrations of Ca2+, Mg2+, Na+, and K+ with inductively coupled plasma–optical emission spectroscopy (ICP-OES; PerkinElmer Optima 5300 DV, USA).

Batch biosorption experiments

Batch biosorption experiments were carried out as a function of pH (from pH 2 to 6), biosorbent dosage (from 0.4 to 12 g L−1), contact time (from 1 to 120 min) of the biosorbent in solution, and initial concentrations of Cd2+, Cu2+, and Pb2+ (each from 10 to 1,000 mg L−1). Controls were performed without the biomass addition. For each treatment with the biosorbent, 4 g L−1 biomass was added into a 250-mL Erlenmeyer flask containing 25 mL of a solution. The flasks were shaken at 150 rpm and 25°C for the entire reaction time. The suspension was filtered through a double-circle filter paper. Aqueous Cd2+, Cu2+, and Pb2+ concentrations before and after biosorption were measured using a flame atomic absorption spectrophotometer (Hitachi Z-5000, Japan) at wavelengths of 228.2, 324.8, and 283.3 nm, respectively. Three replicates were conducted for each experiment condition.

Biosorption–desorption experiments

The heavy metal-loaded biomass after each batch experiment was mixed with different metal-free solutions to test the reusability of the adsorption reactions. The metal-free solutions included deionized water, 0.05 and 0.1 mol L−1 HNO3, and 0.1 and 0.2 mol L−1 NaCl (25 mL for each). Each mixture was agitated at 150 rpm on a rotary shaker for 1 h and filtered. The concentrations of Cd2+, Cu2+, and Pb2+ in the filtrate were measured using ICP-OES and the extent of desorption calculated based on the concentrations.

After the adsorption experiment, the heavy metal-laden biomass was washed with deionized water to remove residual H+ from the biomass until pH rose to 5.0 or higher in the water. Then, the biomass was regenerated by one of the following two methods: by washing the biomass with deionized water until the pH of the wash solution reaches 5.0–6.0 or by soaking and stirring the biomass in 0.1 mol L−1 NaOH solution at a solid/liquid ratio of 4 g L−1 for 5 min. Afterwards, the biomass was washed with deionized water until the pH of the wash solution was between 7.0 and 8.0. The two groups of biomass after the regeneration described above were then reused to investigate the reusability of biomass. Six cycles of biosorption–elution–regeneration experiments were conducted to examine the biomass to retain the metal removal capability.

Ion exchange, FTIR, and SEM approaches

The ion exchange properties, the FTIR (Nicolet 6700, Thermo, USA), and SEM (JSM-6360LV, JEOL, Japan) were used to elucidate the biosorption mechanisms between A. polytricha and the heavy metals at the optimum conditions. The ion exchange properties were determined by measuring the release of Ca2+, Mg2+, Na+, and K+ from the fruiting body during the biosorption of Cd2+, Cu2+, and Pb2+ (Villaescusa et al. 2004). For each experiment, 0.1 g of the biosorbent was added to 25 mL of a metal solution (50 mg L−1) at the optimal pH value of 5.0. The fruiting body suspension was shaken until reaching the equilibrium and then the filtrate was analyzed for concentrations of Ca2+, Mg2+, Na+, K+, Cd2+, Cu2+, and Pb2+ using ICP-OES.

The predominant functional groups of the fruiting body prepared in KBr discs were determined using the FTIR in the 4,000- to 500-cm−1 wavenumber range at a resolution of 4 cm−1. The spectra of the biosorbent before and after the Cd2+, Cu2+, and Pb2+ biosorption processes were determined. The SEM procedure at ×5,000 magnification with an accelerating voltage of 25 keV was also used to investigate the biosorption mechanisms before and after the biosorption process.

Methods for data analyses

The desorption efficiency is calculated as follows:
$$ {DE} = 100\% \frac{{{C_d}}}{{{C_b}}} $$
(1)
where DE is the desorption efficiency (in percent), and Cd and Cb are the desorbed and biosorbed metal concentrations (in milligrams per liter), respectively. The removal efficiency (R, in percent) and biosorption capacity (Q, in milligrams per gram) of the biosorbent are calculated as
$$ R = 100\% \frac{{({C_i} - {C_f})}}{{{C_i}}} $$
(2)
$$ Q = \frac{{({C_i} - {C_f})V}}{W} $$
(3)
in which Ci and Cf are the initial and final heavy metal concentrations (in milligrams per liter), respectively, V is the volume of the tested solution (in liters), and W is the amount of dried biomass (in grams).
Three kinetic models, namely, the pseudo-first-order model, the pseudo-second-order model, and the Elovich equation, were used to characterize the biosorption kinetics of Cd2+, Cu2+, and Pb2+ by A. polytricha. The integration form of the pseudo-first-order model is as follows (Lagergren 1898):
$$ \ln ({Q_e} - {Q_t}) = \ln {Q_e} - {K_1}t $$
(4)
where K1 is the rate constant of the pseudo-first-order biosorption (per minute), Qe is the equilibrium biosorption capacity (in milligrams per gram), and Qt is the concentration of metal ion on the surface of the biosorbent (in milligrams per gram) at time t (in minutes). The rate constant and the equilibrium biosorption capacity can be determined from the plot of ln(Qe − Qt) vs. t. The pseudo-second-order model is in the form of (Ho and McKay 1999)
$$ \frac{t}{{{Q_t}}} = \frac{t}{{{Q_{{\rm e} }}}} + \frac{1}{{{K_2}Q_{{\rm e} }^2}} $$
(5)
where K2 is the rate constant of the pseudo-second-order biosorption (in grams per milligram per minute). The values of K2 and Qe can be determined from the plot of t/Qt vs. t. Assuming αβt ≧ 1, the simplified integration form of the Elovich equation is as follows (Elovich 1939):
$$ {Q_t} = \frac{1}{\beta }ln(\alpha \beta ) + \frac{1}{\beta }lnt $$
(6)
where α is the initial biosorption rate (in milligrams per gram per minute) and β is the desorption constant (in milligrams per gram). The coefficients can be obtained from the plot of Qt vs. lnt. The crucial effect of surface energetic heterogeneity on the adsorption equilibrium in a liquid/solid system may be estimated using the Elovich equation (Sen Gupta and Bhattacharyya 2011).
Three commonly used isotherms, i.e., Freundlich, Langmuir, and Dubinin–Radushkevich (D–R) equations, were used to describe the biosorption equilibrium. The Freundlich isotherm is in the form of (Freundlich 1906)
$$ {Q_e} = {K_F}C_{{\rm e} }^{{1/n}} $$
(7)
where KF (L1/n mg−(11/n) g−1) and n are the Freundlich constants representing the biosorption capacity and the biosorption intensity, respectively. Values of n in the range from 1 to 10 indicate favorable biosorption (Basha and Jha 2008); the smaller the value of 1/n, the more heterogeneous (Gadd 2009). The Langmuir isotherm is defined as follows (Langmuir 1918):
$$ \mathop{Q}\nolimits_{{e} } = \frac{{\mathop{K}\nolimits_{{L} } \mathop{Q}\nolimits_{{m} } \mathop{C}\nolimits_{{e} } }}{{\mathop{1}\nolimits + \mathop{K}\nolimits_{{L} } \mathop{C}\nolimits_{{e} } }} $$
(8)
where Ce is the equilibrium concentration (in milligrams per liter), Qe the concentration of metal ion adsorbed (in milligrams per gram), Qm the theoretical monolayer capacity (in milligrams per gram), and KL is the biosorption equilibrium constant (in liters per milligram) related to the biosorption energy. The D–R equation is expressed as (Dubinin et al. 1947)
$$ {Q_e} = {Q_{{D} }}\exp ( - {K_{{D} }}\varepsilon_{{D} }^2) $$
(9)
$$ {\varepsilon_D} = {\rm RT} \ln (1 + 1/{C_{{\rm e} }}) $$
(10)
where QD is the D–R constant (in milligrams per gram), KD is the activity coefficient (in moles per square kilojoule), and εD is the Polyanyi potential (in joules per mole). The mean free energy of biosorption, E (in kilojoules per mole), is calculated as:
$$ E = \frac{1}{{\sqrt {{2{K_D}}} }} $$
(11)

The magnitude of E is used to estimate the type of biosorption mechanism. An E value between 8 and 16 kJ mol−1 indicates that the biosorption process takes place chemically, while E < 8 kJ mol−1 shows the physically biosorption process (Anayurt et al. 2009).

The coefficient of determination (R2) of the linear regression and the root mean square (RMS) were used to determine the best-fitting process of the kinetic and isotherm models. The RMS is estimated as:
$$ {RMS} = \sqrt {{\sum {({Q_{{cal}}} - {Q_{{\exp }}}} {)^2}/N}} $$
(12)
where N is the number of experimental points, and Qexp and Qcal are the experimental data and calculated values with a model for the adsorbed amount. The values of the mean, standard deviation (SD) of the data, and the model parameters were calculated using Microsoft Excel 2007.

Results

Biosorbent characterization

The fruiting body had a slightly acidic PZC (6.39). In the presence of Cd2+, Cu2+, and Pb2+, the PZC decreased due to the specific adsorption of these ions (Babic et al. 1999).

Cations were released when the fruiting body was treated with an acidic solution (0.1 mol L−1 HCl) and with deionized water (the control, Table 1). The total amount of cations (0.40 meq g−1 dry biomass or 0.20 mmol g−1) released was equivalent to the total ionic content in the fruiting body, which were 0.30, 0.01, 0.07, and 0.02 meq g−1 of Ca2+, Mg2+, K+, and Na+, respectively. This total ionic content was considered as an approximate measure of the CEC of the biosorbent (Villaescusa et al. 2004).
Table 1

Major cation concentrations of the fruiting body of A. polytricha

No. of washing

Ca2+

Mg2+

K+

Na+

Concentration of cations released with 0.1 mol L−1 HCl (mg L−1)

1

110 ± 2.89

15.5 ± 0511

12.4 ± 0.441

17.4 ± 0.784

2

12.0 ± 0.524

1.30 ± 0.0953

1.81 ± 0.0332

2.49 ± 0.0729

3

1.16 ± 0.163

0

1.08 ± 0.0243

1.42 ± 0.103

4

0

0

1.03 ± 0.0349

1.44 ± 0.115

Concentration of cations released with deionized water (mg L−1)

1

3.33 ± 0.296

0.14 ± 0.0249

5.26 ± 0.184

7.95 ± 0.0449

2

0

0

1.51 ± 0.0336

2.12 ± 0.0813

3

0

0

1.12 ± 0.0143

1.45 ± 0.0746

4

0

0

1.12 ± 0.0539

1.49 ± 0.184

Net amount of cation released (meq g−1)a

 

0.30

0.01

0.07

0.02

aDifference between cations released with 0.1 mol L−1 HCl and that by deionized water

Biosorption experiments

As shown in Fig. 1, the metal removal efficiency increased sharply with pH increasing from 2 to 4. The maximum removal efficiencies were 94%, 96%, and 100% for Cd2+, Cu2+, and Pb2+, respectively, at pH 5. At pH 6.0, the removal efficiencies were 86% and 91% for Cd2+ and Cu2+, respectively, and remained 100% for Pb2+. With pH values of 3–6, A. polytricha displayed higher removal rates of Pb2+ than those of Cd2+ and Cu2+. Therefore, the initial pH for the subsequent experiments was set as the optimal pH of 5.0.
https://static-content.springer.com/image/art%3A10.1007%2Fs00253-011-3846-6/MediaObjects/253_2011_3846_Fig1_HTML.gif
Fig. 1

Effect of pH values on the biosorption of Cd2+, Cu2+, and Pb2+ by the fruiting body of A. polytricha biomass (metal concentration, 10 mg L−1; biomass dose, 4 g L−1)

The metal biosorption process was time-dependent with an initially rapid increase of biosorption efficiencies during the first 10–15 min, then much more slowly increase efficiencies until reaching the equilibrium (Fig. 2). The removal efficiencies were in the order of Pb2+ > Cd2+ > Cu2+, corresponding to times of 15, 30, and 60 min, respectively. The removal efficiencies by A. polytricha at 60 min reached 100%, 100%, 93.6%, respectively, for Pb2+, Cd2+, and Cu2+ and remained unchanged thereafter. Therefore, 60 min was considered to be the equilibrium time for all subsequent biosorption experiments.
https://static-content.springer.com/image/art%3A10.1007%2Fs00253-011-3846-6/MediaObjects/253_2011_3846_Fig2_HTML.gif
Fig. 2

Effect of the contact time on the biosorption of Cd2+, Cu2+, and Pb2+ by the fruiting body of A. polytricha biomass (metal concentration, 10 mg L−1; biomass dose, 4 g L−1; pH 5.0)

The removal efficiencies of Cd2+ and Cu2+ increased from 47% to 95% and from 54% to 91%, respectively, with the biosorbent biomass increasing from 0.4 to 4 g L−1 (Fig. 3). The removal efficiency of Pb2+ increased slightly from 93% to 100%, with the biomass increasing from 0.4 to 2 g L−1 or higher. Almost no changes were observed for further increases in biomass dosage up to 12 g L−1. The lowest biosorbent dosage (0.4 g L−1) resulted in the highest biosorption capacity and the lowest removal efficiency. Resulting in both relatively high biosorption capacity and removal efficiency for all three metals, the biomass dosage of 4 g L−1 was selected as the optimal dosage in the following biosorption experiments.
https://static-content.springer.com/image/art%3A10.1007%2Fs00253-011-3846-6/MediaObjects/253_2011_3846_Fig3_HTML.gif
Fig. 3

Effect of the biomass dosage on the biosorption of Cd2+, Cu2+, and Pb2+ by the fruiting body of A. polytricha (metal concentration, 10 mg L−1; pH 5.0)

Biosorption kinetics and isotherms

Fitting results of the kinetic models are summarized in Table 2. The R2 values between the experimental data and fitted values with the pseudo-second-order for the biosorption of Cd2+, Cu2+, and Pb2+ by A. polytricha were 1.00, 0.999, and 1.00, respectively. The comparison of the parameter values, R2, and RMS of the different models indicated that the pseudo-second-order model was the best to characterize the biosorption kinetic of heavy metals by A. polytricha. The equilibrium biosorption capacity values (Qe2,cal) fitted with the pseudo-second-order model (2.59, 2.61, 2.34 mg g−1 for Cd2+, Cu2+, Pb2+, respectively) were almost identical to those from the kinetic data (2.58, 2.62, 2.38 mg g−1 for Cd2+, Cu2+, Pb2+, respectively). The biosorption rates (K2 values) were in the order of Pb2+ > Cd2+ > Cu2+, which were in agreement with the removal efficiencies displayed in Fig. 2.
Table 2

Fitted parameters of the pseudo-first-order, pseudo-second-order, and Elovich models to biosorption kinetic data of Cd2+, Cu2+, Pb2+ by the fruiting body of A. polytricha

Parameter

Cd2+

Cu2+

Pb2+

Pseudo-first-order

Qe1,cal (mg g−1)

0.420

0.499

0.124

K1 (min−1)

0.0345

0.0306

0.0101

R2

0.527

0.349

0.243

RMS

0. 320

0.240

0.0965

Pseudo-second-order

Qe2,cal (mg g−1)

2.59

2.61

2.34

K2 (g mg−1 min−1)

0.501

0.248

2.01

R2

1.00

0.999

1.00

RMS

0.000263

0.00197

0.000132

Elovich equation

α (mg g−1 min−1)

138

569

8.31E+20

β (g mg−1)

3.93

4.68

23.6

R2

0.820

0.856

0.699

RMS

0.174

0.128

0.086

Experimental conditions: heavy metal concentration, 10 mg L−1; biomass dosage, 4 g L−1; stirring rate, 150 rpm; temperature, 25°C; contact times, 1, 3, 5, 10, 15, 30, 45, 60, and 120 min

To examine the relationship between the sorbed (Qe) and aqueous concentrations (Ce) at equilibrium, the experimental data were fitted with the Langmuir, Freundlich, and D–R isotherms; the fitting parameters are listed in Table 3. The KF were 2.05, 3.37, and 10.2 and the 1/n values were 0.488, 0.457, and 0.704 for Cd2+, Cu2+, and Pb2+, respectively. The R2 values for the Freundlich model were found to be 0.975, 0.982, and 0.960, and the RMS were 5.14, 5.49, and 8.26 for Cd2+, Cu2+, and Pb2+, respectively. Statistically (with the higher R2 and lower RMS values), the Freundlich model characterized the biosorption isotherm processes better than the Langmuir and D–R models. The calculated E values from the D–R model for Cd2+, Cu2+, and Pb2+ biosorption were within the energy range of an ion exchange reaction (i.e., 8–16 kJ mol−1). The removal efficiency decreased from 100% to about 30%, with the initial concentration increasing from 10 to 1,000 mg L−1. The experimental values were 63.3, 73.7, and 221 mg g−1 for Cd2+, Cu2+, and Pb2+, respectively (data not shown).
Table 3

Fitted parameters of the isotherm models to biosorption equilibrium data of Cd2+, Cu2+, Pb2+ by the fruiting body of A. polytricha

Parameter

Cd2+

Cu2+

Pb2+

Langmuir

Qm (mg g−1)

69.9

41.89.3

250

KL (L mg−1)

0.00561

0.00526

0.0422

R2

0.774

0.880

0.866

RMS

7.76

6.51

18.4

Freundlich

KF (L1/n mg−(11/n) g−1)

2.05

3.37

10.2

n

2.05

2.19

1.42

R2

0.975

0.982

0.960

RMS

5.14

5.49

8.26

Dubinin–Radushkevich

QD (mg g−1)

122

101

1,366

KD (mol2 kJ−2)

0.00600

0.00500

0.00500

E (kJ mol−1)

9.13

10.0

10.0

R2

0.892

0.934

0.969

RMS

8.04

11.5

10.8

Experimental conditions: heavy metal concentrations, 10–1,000 mg L−1; biomass dosage, 4 g L−1; stirring rate, 150 rpm; temperature, 25°C; contact time, 60 min

Elution, regeneration, and reuse of biosorbent

Table 4 lists the desorption efficiencies of various solutions for eluting the biosorbed metal ions. With 85–100% desorption for Cd2+, Cu2+, and Pb2+, the HNO3 solutions of 0.05 and 0.1 mol L−1 were more effective than the 0.1 and 0.2 mol L−1 NaCl solutions, while deionized water exhibited a negligible desorption capability. The mineral acids (i.e., HNO3 and HCl) are proton exchange agents which dislodge high valence metal ions from the biomass. Table 4 shows increasing desorption efficiencies with the increasing NaCl electrolyte concentrations, suggesting an ion exchange process controlling desorption. However, the increased HNO3 concentrations did not increase the metal elution, indicating saturation of proton exchange (Ferraz et al. 2004).
Table 4

Desorption efficiencies (%, mean ± SD) of Cd2+, Cu2+, and Pb2+ by different solutions

Solution

Cd2+

Cu2+

Pb2+

Deionized water

0

0

0

0.05 mol L−1 HNO3

95.2 ± 1.98

99.4 ± 2.21

97.2 ± 2.64

0.10 mol L−1 HNO3

85.0 ± 7.52

100 ± 0

85.7 ± 1.72

0.10 mol L−1 NaCl

77.2 ± 4.65

38.4 ± 1.60

14.2 ± 1.64

0.20 mol L−1 NaCl

81.0 ± 4.45

45.9 ± 4.75

31.9 ± 0.33

When deionized water was used as the regeneration agent, the removal efficiencies (mean ± SD) of Cd2+, Cu2+, and Pb2+ were 91.14 ± 0.95%, 79.16 ± 2.15%, and 99.41 ± 0%, respectively. As 0.1 mol L−1 NaOH was used as the regeneration agent, the removal efficiencies of Cd2+, Cu2+, and Pb2+ were 97.93 ± 0.63%, 86.61 ± 6.86%, and 71.27 ± 0.49%, respectively. Therefore, 0.1 mol L−1 NaOH was selected as the most efficient agent for the regeneration of the desorbed Cd2+- and Cu2+-loaded biomass while deionized water for the regeneration of the desorbed Pb2+-loaded biomass. Similar results were reported in the regeneration of Cd2+- and Pb2+-loaded Mucor rouxii (Yan and Viraraghavan 2003).

The use of a fruiting body as a potential biosorbent depends not only on the biosorptive capacity but also on the biomass reusability. The reusability of the biosorbent examined using the consecutive biosorption–desorption experiments is shown in Fig. 4. The maximum desorption efficiencies of Cd2+-, Cu2+-, and Pb2+-loaded biomass were 96%, 98%, and 96%, respectively. Without significant mass loss during the cycles, the removal efficiencies of the biomasses were always higher than 95%. In addition, the treatment with the fruiting body to be immersed in the NaOH solution only for 5 min should have minimal effect on the degradation of the original biomass within the cycles. The removal efficiencies in cycles 2–6 were even higher than that in cycle 1, indicating that biomass could be repeatedly subjected to alkaline treatment without losing its biosorption properties.
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Fig. 4

Desorption efficiencies of 0.05 mol L−1 HNO3 for Cd2+-, Cu2+-, and Pb2+-loaded fruiting bodies of A. polytricha

Biosorption mechanisms

Table 5 lists the amount of released exchangeable ions (Ca2+, Mg2+, K+, and Na+) due to the biosorption of Cd2+, Cu2+, and Pb2+ by the fruiting body. The pH values of the biosorption system increased from 5 to 5.35 ± 0.0493, 5.13 ± 0.0153, and 5.41 ± 0.0862, respectively, for Cd2+, Cu2+, and Pb2+; that is, 5.57 × 10−3, 2.53 × 10−3, and 6.08 × 10−3 mmol L−1 of H+ was absorbed by the fruiting body at the equilibrium. The results were two orders of magnitude lower than the released and adsorbed results. Therefore, the effect of the exchange mechanism on the biosorption was negligible. There were significant amounts of Ca2+, K+, and Na+ released from the fruiting body due to the biosorption of Cd2+, Cu2+, and Pb2+. It appeared that more Ca2+ was released than the other cations, except for the biosorption process of Pb2+, which was inconsistent with the fact that Ca2+ was the major cation in the CEC (Table 1). A coefficient of Rr/b was defined as the ratio of the amount of cations released during biosorption to that of the ions bounded by the fruiting body. The Rr/b values of Cd2+, Cu2+, and Pb2+ were 0.61, 0.65, and 0.63, respectively, at the optimal biosorption condition.
Table 5

Amount of released exchangeable ions (Ca2+, Mg2+, K+, and Na+) due to the biosorption of Cd2+, Cu2+, and Pb2+ by the fruiting body of A. polytricha

Treatment

Cd2+

Cu2+

Pb2+

Ca2+

Mg2+

K+

Na+

Rr/b

 

Total metal bound (mmol L−1)

Amount of cations released (mmol L−1)

Control

   

0

0

0.0282 ± 0.00602

0.0788 ± 0.00221

 

Cd2+

0.419 ± 0.00174

  

0.105 ± 0.0610

 

0.121 ± 0.00687

0.290 ± 0.0431

 

Cu2+

 

0.523 ± 0.00692

 

0.198 ± 0.0139

 

0.122 ± 0.00791

0.269 ± 0.00929

 

Pb2+

  

0.212 ± 0.000100

0.00193 ± 0.00333

 

0.110 ± 0.00450

0.261 ± 0.0215

 
 

Total metal bound (meq g−1)

Amount of cation released (meq g−1)

Cd2+

0.419 ± 0.000980

  

0.105 ± 0.0607

 

0.0465 ± 0.00333

0.105 ± 0.0213

0.61

Cu2+

 

0.523 ± 0.00669

 

0.198 ± 0.0137

 

0.0470 ± 0.00340

0.0953 ± 0.00461

0.65

Pb2+

  

0.212 ± 0.000100

0.00193 ± 0.00335

 

0.0408 ± 0.00216

0.0909 ± 0.0106

0.63

Experimental conditions: heavy metal concentration, 50 mg L−1; biomass dosage, 4 g L−1; stirring rate, 150 rpm; temperature, 25°C; contact time, 60 min

To determine the main functional groups of the biomass participating in the adsorption of Cd2+, Cu2+, and Pb2+, the original biomass and metal-loaded biomass were characterized using FTIR. Figure 5 displays the FTIR spectra of the raw (metal-unloaded) and the metal-loaded fruiting body. Absorbance peaks of 3,396, 2,918, 1,729, 1,649, 1,078, 1,037, 889, and 561 cm−1 were observed in the raw biomass spectra. The broad strong band at 3,396 cm−1 corresponded with the overlapping of amine (–NH) or hydroxyl (–OH) groups. The observed peak at 2,918 cm−1 exhibited the C–H stretching vibrations of the –CH3 and >CH2 functional groups attributed to fatty acids in membrane phospholipids (Sari and Tuzen 2009). The membrane phospholipids also contained ester functional groups, which formed characteristic peaks at 1,729 and 1,649 cm−1 due to the C=O stretching frequencies. Additional information on phospholipids along with the phosphodiester, free phosphate, and monoester phosphate functional groups were determined in the region between 1,250 and 1,200 cm−1, which correspond to >P=O double-bond asymmetric stretching frequencies. The >P=O phosphodiester functional groups also contributed to the spectral features in this region (Yee et al. 2004). The band observed at 1,078 cm−1 was assigned to C–O stretching of alcohols and carboxylic acids. The band at 1,037 cm−1 indicated the C–O–C and C–O–P stretching vibrations of polysaccharides. The small peak at 889 cm−1 represented out-of-plane –CH vibration in aromatic compounds (Balasubramanian et al. 2009) or out-of-plane –NH2 bending. The band at 561 cm−1 for the fungal preparation represented the C–N–C scissoring and was only found in the polypeptide protein structure (Sanghi et al. 2009).
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Fig. 5

FTIR spectra of unloaded (a), Cd2+-loaded (b), Cu2+-loaded (c), and Pb2+-loaded (d) fruiting bodies of A. polytricha biomass

The metal-absorbed biomass exhibited stretching band shifts from 3,396 cm−1 to 3,390, 3,400, and 3,410 cm−1, respectively, corresponding to the chemical interactions of Cd2+, Cu2+ and Pb2+ toward –OH/–NH bonds. The peak at 2,918 cm−1 related to –CH absorbance shifted to 2,922 cm−1 for Cd2+ and to 2,924 cm−1 for Cu2+ and Pb2+ after biosorption. The carboxyl peak at 1,729 cm−1 were shifted to 1,732 cm−1 for Cd2+- and to 1,736 cm−1 for Cu2+- and Pb2+-loaded biomass. After biosorption, the peak at 1,037 cm−1 of C–O–C and C–O–P bonds shifted to 1,041, 1,040, and 1,040 cm−1 for Cd2+-, Cu2+-, and Pb2+-loaded biomass, respectively. The spectral analysis before and after ion binding indicated that the phosphoryl group was involved in the Cd2+, Cu2+, and Pb2+ binding. There were peaks shifting in the absorption intensity of >P=O stretching (1,250–1,200 cm−1) and C–O–C and C–O–P stretching (1,040 cm−1) after Cd2+, Cu2+, and Pb2+ biosorption. When the biomass was loaded with Cd2+, Cu2+, and Pb2+, peaks at 2,918 and 889 cm−1 were related to –NH stretching and –NH bending in the amino groups, respectively, indicating the participation of amino groups in the metal uptake. It was noted that in the region of lower wavenumbers (under 700 cm−1), a sharp peak at 561 cm−1 diminished after metal biosorption in comparison with that of the raw fruiting body. This change was attributable to an interaction between C–N–C in polypeptide protein and the metal ions. Therefore, carboxyl, amino/hydroxyl, amine, phosphoryl, and C–N–C on the biomass surface might be the functional groups to provide binding sites.

SEM images of the fruiting body surface before and after metal ion biosorption are shown in Fig. 6. The surface of the raw biomass (Fig. 6a) was smooth and uniform with a regular and plain structure. The surfaces of Cd2+- and Cu2+-loaded biomass changed somewhat (Fig. 6b, c), while the surface of Pb2+-loaded biomass was much rougher (Fig. 6d). The metal ions as spot-like particles distributed on the surface of the Cd2+- and Pb2+-loaded fruiting body, while extra flake-like substances distributed on the surface of the Cu2+-loaded biomass. The chemical interactions between the functional groups of the biomass and metal ions might be responsible for the surface structure changes.
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Fig. 6

SEM images (magnification, ×5,000) of non-loaded (a), Cd2+-loaded (b), Cu2+-loaded (c), and Pb2+-loaded (d) fruiting bodies of A. polytricha

Discussion

Affecting factors of biosorption

The removal of heavy metals is strongly affected by the initial pH of solution (Gadd 2009; Villaescusa et al. 2004). The maximum removal of Cd2+, Cu2+, and Pb2+ often happens under an acid condition (Anayurt et al. 2009; Muraleedharan et al. 1995) and depends on the nature of the biomass and metals (Subbaiah et al. 2009). As shown in Table 6, for initial pH values of 2 and 5, the equilibrium pH values were larger than the initial values, and for the initial pH of 6, the equilibrium pH value was smaller. It was also noticed that the amount of H+ adsorbed onto the fruiting body decreased with increasing initial pH values. As the pH increases, more ligands, such as carboxyl, amino, and phosphoryl groups, are exposed and carry more negative charges with a subsequently higher attraction of positively charged metal ions, resulting in more metal ions being adsorbed onto the surface of the fruiting body. In the initial pH range between 2 and 5, H+ adsorbed onto the fruiting body by competing with Cd2+, Cu2+, and Pb2+, resulting in a pH increase in the equilibrium solution and reducing the metal biosorption. However, for the initial pH of 6, the release of H+ into the solution was attributed to the formation of soluble hydroxylated complexes of the metal ions and the ionized nature of the cell wall surface of the biomass (Balasubramanian et al. 2009; Subbaiah et al. 2009), ascribing to the decrease of biosorption efficiencies. Several studies evidenced the similar pH effect on biosorption of heavy metals by using different kinds of fungal biomass (Anayurt et al. 2009; Sari and Tuzen 2009).
Table 6

Changes of H+ concentrations for different initial pH values in Cd2+, Cu2+, and Pb2+ solutions

Heavy metal

Initial pH

pH at equilibrium

H+ adsorbed by biomass (mmol L−1)

Cd2+

2

2.23 ± 0.0100

4.11 ± 0.136

3

3.96 ± 0.0289

0.891 ± 0.00738

4

4.66 ± 0.0173

0.0781 ± 0.000863

5

5.35 ± 0.0493

0.00557 ± 0.000489

6

5.20 ± 0.0808

−0.00526 ± 0.00115a

Cu2+

2

2.11 ± 0.00577

2.18 ± 0.104

3

3.84 ± 0.0321

0.854 ± 0.0110

4

4.75 ± 0.0153

0.0824 ± 0.000617

5

5.13 ± 0.0153

0.00253 ± 0.000264

6

5.17 ± 0.0346

−0.00576 ± 0.000552 a

Pb2+

2

2.08 ± 0.0100

1.68 ± 0.192

3

3.89 ± 0.0173

0.871 ± 0.00508

4

4.97 ± 0.0802

0.0892 ± 0.00199

5

5.41 ± 0.0862

0.00608 ± 0.000763

6

5.50 ± 0.0681

−0.00214 ± 0.000477 a

aThe release of H+ ion was attributed to the formation of soluble hydroxylated complexes of the metal ions and the ionized nature of the cell wall surface of fruiting body of A. polytricha

The initial fast metal biosorption is related to the abundance of active metal binding sites on the biosorbent surface and the diffusion process from the bulk solution to the surface. For the following slower stage, the biosorption is likely an attachment-governed process due to the ion competition for the less available binding sites on the sorbent surface. The same qualitative behavior was also reported by Sari and Tuzen (2009) and Amarasinghe and Williams (2007).

An increase in the removal efficiency with the biomass dosage is attributable to the increase in surface area, the binding sites, and/or the functional groups (Akar et al. 2006). However, the biosorption efficiency was nearly the same for biomass dosages higher than 4 g L−1. The high biomass dosages resulted in biomass aggregates, which might cause interference between the binding sites and insufficient metal ions in the solution. The results were consistent with those by Ghorbani et al. (2008) and Sari and Tuzen (2009). On the other hand, the biosorption capacity decreased with the increased biomass dose. The result was due to the fact that at a higher biosorbent dosage, the solution ion concentration dropped to a lower value and the system reached equilibrium at a lower value of Q, indicating unsaturated adsorption sites.

Kinetic and equilibrium models

The kinetics data provided an insight into the possible mechanism of adsorption along with the reaction pathways. The biosorption generally follows a three-stage process: boundary layer diffusion, intraparticle diffusion, and biosorption on binding sites (Basha et al. 2008; Zolgharnein and Shahmoradi 2010). Our measured specific surface area of the fruiting body was 5 m2 g−1, indicating that the micropores available for biosorption inside the biomass were limited and the intraparticle diffusion stage was negligible.

In general, the pseudo-second-order kinetic describes well the long process period, while the pseudo-first-order model fits the experimental data well for an initial period of the first reaction step. The assumption of the pseudo-second-order model is that metal ions are adsorbed onto two surface sites and that chemisorption occurs involving valency forces through the sharing or the exchange of electrons between the fruiting body and the divalent metal ions (Ho 2006; Zolgharnein and Shahmoradi 2010), as indicated by the following equation:
$$ \mathop{{2BS}}\nolimits_{{(solid)}} + \mathop{M}\nolimits_{{(aqueous)}}^{{2 + }} \leftrightarrow \mathop{{(BS)}}\nolimits_2 \mathop{M}\nolimits_{{(adsorbedphase)}}^{{2 + }} $$
(13)
where BS is the biosorption site and M2+ represents the binary metal ion. In this study, the best-fitting results of the pseudo-second-order for all the biosorption data supported the assumption that the biosorption rate was controlled by the chemical biosorption or chemisorption. Another reason for the best-fitting results of the pseudo-second-order model was the heterogeneous nature of the surface sites (Reddad et al. 2002), which implied different biosorption kinetic stages of the metal with the respective chemical moieties (Fig. 2).

Among the three isotherm models, the Freundlich model was the best to characterize the biosorption isotherm processes, indicating the heterogeneous surface process. The slope 1/n values <1 suggested that Cd2+, Cu2+, and Pb2+ were favorably adsorbed by the fruiting body. The range of calculated E values from the D–R model indicated that the biosorption occurred via an ion exchange or chemisorption process (Anayurt et al. 2009).

The metal biosorption process was in the order of Pb2+ > Cu2+ > Cd2+. This order might be related to the metal interactions with ligands, determined by the metal ion charge z, radius r, and electronegativity Xm. The values of metal ion charge z and radius r are 2,147 pm and 2.33 (Pauling scale) for Pb2+, 2,138 pm and 1.90 for Cu2+, and 2,148 pm, and 1.68 for Cd2+, respectively. The Xm2r values (798, 498, and 422 for Pb, Cu, and Cd, respectively) represent the valence orbital energy and are a measure of the strength of covalent bonding relative to ionic bonding. While the values of the ion charge density are similar for the three metals, both the electronegativity and the strength values are in the order of Pb2+ > Cu2+ > Cd2+. Similar results were shown by Pan et al. (2010) and Papageorgiou et al. (2006).

Mechanisms

The mechanisms involved in the biosorption are usually complicated, dependent on the type and nature of biomass employed and the type of metals. The best-fitting results of the pseudo-second-order model suggested that the biosorption of heavy metals by the fruiting body was via chemisorption. Furthermore, the results of FTIR indicated the chemical interactions through the ion exchange mechanism during the biosorption process. The results of Rr/b < 1 suggested that the ion exchange was the major mechanism of metal uptake and the donor sites were oxygen-dominated from the structural polysaccharides present on the cell wall of A. polytricha. Among various proposed mechanisms, the ion exchange and complexation mechanisms have been considered the most important process (Liu, et al. 2011):
$$ \mathop{{R} }\nolimits^{{n - }} \mathop{{E} }\nolimits^{{n + }} + \mathop{{M} }\nolimits^{{n + }} \leftrightarrow \mathop{{R} }\nolimits^{{n - }} \mathop{{M} }\nolimits^{{n + }} + \mathop{{E} }\nolimits^{{n + }} $$
(14)
$$ {{{M}}^{{n + }}} + {{mBL}} \leftrightarrow {{MBL}}_m^{{n + }} $$
(15)
where RnEn+ is the biomass with the exchangeable site (usually carboxyl acid or its salt), Mn+ represents the bivalent metal ion, En+ denotes the exchange ion (H+ or other metal ions), and BL signifies the biomass with ligand groups (usually N or O in the amino or hydroxyl groups). An early study has found evidence that peat takes up metal ions through ion exchange and surface complexation (Balasubramanian et al. 2009). The ion exchange and complexation mechanisms can be rapid and reversible, with biomass properties analogous to the conventional ion exchange resins (Gadd 2009). The rapid biosorption by the fruiting body and almost thoroughly regenerated metals from A. polytricha in our study also demonstrated the synergistic mechanism of ion exchange and complexation. The SEM images shown in Fig. 6 displayed the interactions of metal ions with the surface structure of the cell wall of A. polytricha. Therefore, a synergistic mechanism involving ion exchange and surface complexation should be the most responsible mechanism for metal uptake by the biosorption process.

In summary, the biosorption processes of the fruiting body of A. polytricha were a function of pH, time, biomass dosage, and metal concentrations. The processes were also metabolism-independent, rapid, favorable, and reversible, allowing easy desorption of the adsorbed metal ions and reuse/regeneration of the biosorbent. The biosorption mechanism of synergistic ion exchange and surface complexation of the fruiting body was demonstrated by the experiments of ion exchange, FTIR, SEM, and the fitting parameters of the kinetics and isotherm models. The results should provide a better understanding of the process and are beneficial to apply A. polytricha as a biosorbent for toxic metal removal from contaminated water.

Acknowledgments

This work was partly supported by grants from the Chinese National Natural Science Foundation (nos. 51179212 and 51039007), the Fundamental Research Funds for the Central Universities, and the Research Fund Program of Guangdong Provincial Key Laboratory of Environmental Pollution Control and Remediation Technology (no. 2011K0004). The authors are grateful for the insightful reviews of Jeremy B. Fein at University of Notre Dame and three anonymous reviewers.

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© Springer-Verlag 2012