Mechanisms for cadmium adsorption by magnetic biochar composites in an aqueous solution

Journal Pre-proof Mechanisms for cadmium adsorption by magnetic biochar composites in an aqueous solution Zulqarnain Haider Khan, Minling Gao, Weiwen Qiu, Md. Shafiqul Islam, Zhengguo Song PII:

S0045-6535(19)32942-X

DOI:

https://doi.org/10.1016/j.chemosphere.2019.125701

Reference:

CHEM 125701

To appear in:

ECSN

Received Date: 25 September 2019 Revised Date:

14 December 2019

Accepted Date: 17 December 2019

Please cite this article as: Khan, Z.H., Gao, M., Qiu, W., Islam, M.S., Song, Z., Mechanisms for cadmium adsorption by magnetic biochar composites in an aqueous solution, Chemosphere (2020), doi: https:// doi.org/10.1016/j.chemosphere.2019.125701. This is a PDF file of an article that has undergone enhancements after acceptance, such as the addition of a cover page and metadata, and formatting for readability, but it is not yet the definitive version of record. This version will undergo additional copyediting, typesetting and review before it is published in its final form, but we are providing this version to give early visibility of the article. Please note that, during the production process, errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain. © 2019 Published by Elsevier Ltd.

1

Mechanisms for cadmium adsorption by magnetic biochar

2

composites in an aqueous solution

3

Zulqarnain Haider Khana, Minling Gaob, Weiwen Qiuc, Md. Shafiqul Islama, Zhengguo Songb*

4

a

5

China

6

b

7

China

8

c

9

Christchurch 8140, New Zealand

Agro-Environmental Protection Institute, Ministry of Agriculture of China, Tianjin, 300191,

Department of Civil and Environmental Engineering, Shantou University, Shantou, 515063,

The New Zealand Institute for Plant and Food Research Limited, Private Bag 4704,

10

*Corresponding author. Tel: 0086 13920782195

11

Email: [email protected]

1

12

Abstract: There is a demand to develop techniques for the continuous removal/immobilization

13

of heavy metals from contaminated soil and water bodies. In this study, a unique biochar

14

preparation method was developed for the removal of cadmium. First, conventional biochars of

15

corn straw were produced by pyrolysis at two temperatures and then treated using one-step

16

synthesis at different ferric nitrate ratios and different calcination temperatures to produce

17

magnetic biochars. Second, the prepared biochars were used as adsorbents for Cd(II) removal

18

from a solution, and the best one was selected for further evaluation. Various techniques were

19

used to characterize the adsorbents and determine the main adsorption mechanism. The results

20

indicated that the biochars successfully carried iron particles within, which improved the specific

21

surface area, formed inner-sphere complexes with oxygen-containing groups, and increased the

22

number of oxygen-containing groups. The adsorption experiments revealed that MBC800-0.6300

23

had a higher affinity for Cd(II) than the other adsorbents. Batch adsorption experiments were

24

performed to explore the influence of the kinetics, isotherm, pH, thermodynamics, ionic strength,

25

and humic acid on Cd(II) adsorption. The results indicated that the Langmuir model fit the Cd(II)

26

adsorption best with MBC800-0.6300 having the highest adsorption capacity (46.90 mg g−1). The

27

sorption kinetics of Cd(II) on the adsorbent follows a pseudo-second-order kinetics model.

28

Because MBC800-0.6300 is loaded with metal ions, it can be conveniently collected by a magnet.

29

Thus, biochar modification methods with ferric nitrate impregnation provide an excellent

30

approach to eliminating Cd(II) from aqueous solutions. The possible adsorption mechanisms

31

include chemisorption, electrostatic interaction, and monolayer adsorption.

32

Keywords: Magnetic biochar composite, synthesis, adsorption process, solution

33

2

34

1. Introduction

35

Many types of chemical pollutants continuously damage aquatic environments (Sayed et al.,

36

2019). Heavy metals and chemical pollutants strongly affect the environment owing to their

37

characteristics of bioaccumulation, high persistence, and non-biodegradability (Chowdhury et

38

al., 2017). Cadmium is one of the most harmful heavy metals and is discharged into aquatic

39

environments through anthropogenic and natural activities (Chowdhury et al., 2014). It mainly

40

derives from plastics, paint pigments, electroplating, smelting operations, and the nickel–

41

cadmium and silver–cadmium battery industries (Asci et al., 2008). Because of the high toxicity

42

and persistence of Cd(II) in natural water and farmland, it has become an increasing concern

43

over the past decades (Zhou et al., 2018; Liu et al., 2019). Cadmium is considered the seventh-

44

most hazardous chemical by the US Department of Health and Human Services (ATSDR, 2012).

45

In Japan, the main cause of the itai-itai disease was Cd(II) accumulation in the aquatic

46

environment (Chowdhury et al., 2017). Cadmium has harmful effects on the human body and

47

causes carcinogenicity and liver damage (Zhou et al., 2018). Critical illnesses such as hepatic

48

injury, lung damage, hypertension, renal dysfunction, and teratogenic effects can also be caused

49

by Cd(II) (Zhang et al., 2019). The World Health Organization (WHO) and United States

50

Environmental Protection Agency (USEPA) have declared the maximum acceptable

51

concentrations of Cd(II) to be 0.03 and 0.05 mg L-1, respectively (Chowdhury et al., 2014).

52

Because of its extreme toxicity, Cd(II) needs to be removed from contaminated water to ensure

53

access to safe and pure water. Hence, Cd(II) must be eliminated from wastewater before it is

54

disposed in the environment. Various physical and chemical techniques have been employed to

55

lower the Cd(II) concentration to meet environmental standards, including chemical

56

precipitation, ultrafiltration, membrane separation, electrochemical deposition, and adsorption

57

(Zhang et al., 2019). Although these methods perform well at removing Cd(II) from wastewater, 3

58

they still have significant disadvantages, such as high sludge production, high energy

59

requirements, and byproduct formation (Zamri et al., 2017). Among the above methods,

60

adsorption is regarded as a highly efficient, low in cost, simple to operate, and economical

61

approach to removing Cd(II) (Zamri et al., 2017). Many influencing factors play an essential role

62

in the adsorption process, including the surface area, adsorption capacity, and mechanical

63

stability. Therefore, an efficient and low-cost adsorbent for Cd(II) removal needs to be identified

64

to guarantee drinking water and food safety. Various adsorbents are available, but carbon

65

materials have been reported to perform the best owing to their numerous advantages, such as

66

high stability, excellent removal efficiency, and large surface area (Hanigan et al., 2012).

67

Biochar (BC) contains carbon and is produced by the pyrolysis of waste biomass in an

68

oxygen-deficient environment (Joseph and Lehmann, 2009). It can immobilize or remove heavy

69

metals to decrease toxicity. However, pristine BC has limited capability to eliminate pollutants

70

from aqueous solutions and needs to be further improved with well-surface properties and novel

71

structures (Song et al., 2014). Recent literature has shown that the surface features of BC can be

72

modified to enhance the adsorption potential by increasing the number of functional groups and

73

selectivity to remove several pollutants from polluted systems (Ma et al., 2014). In addition, iron

74

oxide nanomaterials have been shown to be helpful for removing heavy metals because of their

75

magnetic recyclability, high reactivity, and natural abundance (Wu et al., 2018). Iron can

76

potentially be used to modify BC to increase the heavy metal sorption.

77

Some modification techniques have been applied to adsorbents in recent years for water

78

treatment. Karunanayake et al. (2018) found that Fe3O4-magnetized Douglas fir BC is effective

79

for Cd(II) and Pb(II) removal and can easily be separated from an aqueous solution. Steam-

80

activated (800 °C for 3 h) and KOH-treated (1.3 M) BCs have been shown to successfully

4

81

remove copper from solutions (Ippolito et al., 2012). Similarly, nanoscale zerovalent, NaOH, and

82

Zn ion-modified BCs produced from various metal ion-treated biomass materials have been used

83

as adsorbents for the removal of heavy metals such as cadmium, lead, and arsenic from water.

84

BC modification via activation can increase the surface area, ligand functional group presence,

85

or number of available electrons (Ippolito et al., 2012), all of which help improve heavy metal

86

sorption. Modifying BC with iron can increase its ability to remove heavy metals by enhancing

87

its sorption ability. Iron BC composites may provide excellent Cd(II) removal efficacy because

88

they are magnetically recyclable, cost-effective, and environmentally friendly. Separating BC

89

from the solution after adsorption is difficult because of its tiny particle size. Recently, a new

90

technique known as magnetic separation has attracted much attention for its ability to facilitate

91

solid–liquid separation (Pan et al., 2012). BCs loaded with magnetic nanoparticles may be able

92

to achieve rapid separation and recovery under an external magnetic field. In recent years,

93

considerable attention has been given to using ferrites to develop magnetic composites (Reddy

94

and Lee 2013). Combining BCs and ferrites should provide both magnetic properties and

95

excellent adsorption performance. To obtain a useful adsorption ability for water treatment,

96

several BC-supported ferrite composites have been prepared and demonstrated good results.

97

In this study, magnetic biochar (MBC) composites were developed that are ecofriendly and

98

easy to produce. MBC composites were characterized and synthesized, and the adsorption

99

kinetics and isotherms of Cd(II) with this innovative sorbent were thoroughly investigated. The

100

kinetic and equilibrium parameters were examined, and the influences of the pH, ionic strength,

101

and humic acid (HA) on the adsorption process were analyzed. The thermodynamic parameters

102

were also determined. Based on the results, a unique and exceedingly efficient MBC for removal

103

of Cd(II) from water was identified.

5

104

2. Materials and methods

105

Corn waste straw was collected from a Tianjin suburban farm and crushed into powder

106

before treatment. All chemicals were of analytical grade and bought from Jiuxinyaozheng Co.,

107

Ltd., Beijing, China. Deionized water was used in the experiment (18.25 MΩ/cm) and obtained

108

with a Millipore Milli-Q water purification system.

109

2.1 Adsorbent preparation

110

Following Song et al. (2014), BCs were produced by slow pyrolysis of waste corn straw

111

powder in a muffle furnace for a residence time of 2 h and with a continuous flow of N2 gas

112

(nitrogen flow rate: 300 cm3 min−1) at two different temperatures of 600 and 800 °C. These were

113

designated as BC600 and BC800, respectively. To prepare the MBCs, 5 g of BC600 or BC800

114

was added to 60 mL of ferric nitrate at different concentrations (0.3, 0.6, 0.9, and 1.2 mol L−1)

115

and then stirred for 3 h with a magnetic stirrer. These were designated as MBC600-X and

116

MBC800-X, respectively. These two materials were sonicated for 2.5 h and evaporated to

117

dryness in a water bath at a constant temperature (100 °C) and subjected to calcination in a

118

nitrogen flow (600 cm3 min−1) at three different temperatures (300, 600, and 800 °C) for another

119

2 h. The final samples were labeled as MBC600-Xy and MBC800-Xy, where X represents iron

120

nitrate (mol L−1) and Y represents the calcination temperature (°C). The residue was ground and

121

passed through a sieve with a size of 0.154 mm and rinsed with deionized distilled water several

122

times. The untreated BC was used as a control. Different concentrations of corn straw powder

123

(0.3, 0.6, 0.9 and 1.2 mol L−1) and ferric nitrate (100 mL) were added to Milli-Q water in a 500

124

mL beaker and stirred by sonication. The samples were then evaporated to dryness. Finally,

125

pyrolysis was completed in a muffle furnace with a nitrogen flow of 600 mL min−1 at a

126

temperature of 600 or 800 °C for 2 h.

6

127

2.2 Characterization methods

128

An elemental analyzer (Vario el cube, Elementar, Germany) was used to identify the

129

elements (C, N, H, O) of the absorbents. The Brunauer–Emmett–Teller (BET) surface area was

130

determined by N2 adsorption and desorption measurements (3Flex, Micromeritics, USA). A

131

scanning electron microscope (SEM-Hitachi SU-1510) was used to detect the surface

132

morphologies and elemental contents of the materials. X-ray powder diffraction (XRD) patterns

133

were obtained at a scanning rate of 4°/min in the scanning range of 10–90° to determine the

134

phase composition with a Cu anode (Bruker, Germany, λKɑ = 1.5406 Å). A vibrating sample

135

magnetometer (VSM, Squid-vsm) from American Quantum was used to assess the magnetic

136

properties of the MBCs. To identify the functional groups on the materials, Fourier transform

137

infrared spectroscopy (FT-IR) was performed in the wavenumber range of 400–4000 cm−1 with a

138

spectrophotometer (Nicolet 6700, Thermo Nicolet Co., Waltham, MA, USA). X-ray

139

photoelectron spectroscopy (XPS) was performed to explore the adsorption mechanism of

140

samples with an American Thermo ESCALAB 250Xi XPS System. The zero charges (pHzpc)

141

were measured with a Zetaziser Nano Series (NANO-ZS90, UK). The particle size distribution

142

was measured with a particle size analyzer (Mastersizer-2000, Malvern Instruments Ltd.,

143

Malvern, UK).

144

2.3 Adsorption experiments

145

Batch experiments were conducted to determine the potential of MBC, equilibrium contact

146

time, and effects of relevant factors (pH, HA, temperature, and ionic strength) for cadmium ion

147

adsorption. Adsorption kinetic experiments were performed by adding 0.5 g of a sample to a

148

Cd(II) solution (500 mL, 150 mg L−1) and stirring at 1000 rpm to determine the equilibrium time.

149

Aliquots (0.5 mL) were sampled, filtered through a Whatman No. 42 filters, and analyzed by

7

150

atomic absorption spectrometry (AAS) at different time intervals (1, 5, 10, 20, 30, 60, 120, 240,

151

360, 480, 720, 960, and 1440 min).

152

In all experiments on the adsorption isotherm, 20 mg of the adsorbent was added into 50

153

mL brown vials containing 20 mL of Cd(II) solutions with different initial concentrations (10–

154

150 mg L−1) and 0.01 M NaNO3 as the background electrolyte. The impact of pH on adsorption

155

was analyzed by adjusting the pH value from 3.0 to 8.0 with 0.1 mol/L NaOH and HNO3

156

solutions. Different temperatures (290, 300, and 310 K) were used to study the adsorption

157

thermodynamics. Different concentrations of NaNO3 (0.001–0.1 M) were added to the Cd(II)

158

solution to determine the ionic strength. Different concentrations of HA (10–30 mg L−1) were

159

added to the Cd(II) solution to study their effect.

160

In each experiment, brown glass vials were placed in an air bath, and a thermostatic

161

oscillator end-over-end tumbler was used to shake them for 6 h at 180 r/min and 25 °C to reach

162

adsorption equilibrium. After shaking, all solutions were filtered with Whatman No. 42 filters for

163

analysis, and the Cd(II) concentration in the filtrate was determined by AAS. All experiments

164

were conducted in duplicate. The Cd(II) adsorption capacity of MBCs was determined from the

165

difference in concentration before and after adsorption equilibrium was achieved.

166

The Cd(II) adsorption capacity (Qe) is formulated as follows:

167


 =

(  )

(1)



168

where C0 (mg g−1) and Ce (mg g−1) are the initial and equilibrium Cd(II) concentrations,

169

respectively; m (g) is the mass of the adsorbent; and V (L) is the volume of the Cd(II) solution.

170

2.4 Statistical analyses

8

171

Statistical analyses were performed with SPSS v.18.0 for Windows (SPSS Inc., Chicago,

172

USA). Means were compared by one-way analysis of variance, and Duncan’s multiple range

173

tests were completed at the 5% level (significant difference P < 0.05).

174

3. Results and discussion

175

3.1 Characterization of magnetic biochar

176

Fig. S1 shows SEM images of BC600, BC800, MBC600-0.6300, and MBC800-0.6300 that

177

were acquired to observe the morphology of different adsorbents. These show the structural

178

changes to the BC surface after modification. The surfaces of the MBCs changed from that of the

179

pristine BC to become rough with many different shapes and a porous structure (Figs. S1(a) and

180

(d)). This indicates that the modification effectively enhanced the porosity and made the surface

181

easier for Fe to adsorb onto (Wu et al., 2016). Figs. S1(b) and (e) clearly show that MBC600-

182

0.6300 and MBC800-0.6300 were very different compared to the exposed BC surface in Figs.

183

S1(a) and (d) and featured bunches of aggregates (particles) stuck to the surface with plate-like

184

rough and irregular morphologies containing sharp edges and corners. Iron oxide particles were

185

imbedded within the BC600 and BC800 matrices, which indicates good mechanical bonding (Hu

186

et al., 2015a). Fig. S7 shows that the modifications of BC600 and BC800 reduced the particle

187

size of the adsorbents. The particle size distribution analysis revealed that the mass median (d50)

188

particle sizes of BC600, BC800, MBC600-0.6300, MBC800-0.6300, MBC600-0.6300, and

189

MBC800-0.6300 with Cd(II) adsorption were 76.375, 112.835, 70.988, 83.772, 25.055, and

190

24.085 µm, respectively. An adsorbent with a smaller particle size can have a greater surface

191

area and thus adsorb more Cd(II). Thus, the particle size of adsorbents is another key factor that

192

affects the Cd(II) adsorption capacity (Hidayat et al., 2010). As given in Table 1, the iron

193

contents of MBC600-0.6300 and MBC800-0.6300 were 6.47% and 7.07%, respectively. This

9

194

indicates that the surfaces of the modified BCs were saturated by iron. The carbon, nitrogen, and

195

hydrogen contents of BC600 were 84.35%, 1.36%, and 2.69%, respectively. They were 86.14%,

196

1.11%, and 1.34%, respectively, for BC800; 59%, 2.18%, and 2.12%, respectively, for MBC600-

197

0.6300; and 48.65%, 1.42%, and 1.01%, respectively, for MBC800-0.6300. BET analysis was used

198

to characterize the pores and textures of the adsorbents. The pore volume, pore size, and surface

199

area of the BCs are presented in Table 1. The specific surface areas of BC600, BC800, MBC600-

200

0.6300, and MBC800-0.6300 were 97.18, 93.76, 225.90, and 313.88 m2 g−1, respectively. These

201

results indicate that the Fe impregnation influenced the pore opening activation and pore

202

structure of the pristine BC surface (Zhou et al., 2010). The greater development of the surface

203

area and pore volume of MBC800-0.6300 can be attributed to the high pyrolysis temperature. The

204

sample elemental composition analysis showed that the MBCs had a lower carbon content and

205

higher iron content than pristine BC, which suggests that carbon was successfully blended with

206

iron in the MBC composite. The large surface area of MBC800-0.6300 may have provided several

207

active adsorption sites. MBC600-0.6300 and MBC800-0.6300 had lower surface areas that most

208

commercial adsorbents; for example, activated carbon has a surface area of 1896 m2 g−1

209

(Hameed et al., 2007). The specific surface area is not only a key parameter that influences the

210

Cd(II) sorption capacity but may also enhance the reaction between an MBC and Cd(II).

211

3.1.1 Magnetic properties

212

Fig. 1 shows the magnetic hysteresis loops used to study the magnetic properties of MBC600-

213

0.6300 and MBC800-0.6300. The curves indicated typical super-paramagnetic properties.

214

MBC600-0.6300 and MBC800-0.6300 had magnetization M values of 14.5 and 9.75 emu g−1,

215

respectively. Nevertheless, the M value of MBC800-0.6300 was still sufficiently high for the

10

216

powdered MBC to be separated by a magnet after use. MBC800-0.6300 was easily separated in

217

the presence of an external magnetic field, as shown by the inset.

218

3.1.2 XRD analysis

219

XRD analysis was performed over a range of 2θ = 10°–90° to identify the crystallographic

220

structures of BC600, BC800, MBC600-0.6300, and MBC800-0.6300, as shown in Fig. S2. The

221

BC600 and BC800 patterns demonstrated a broad intense sharp peak at about 26.6°, which

222

indicates amorphous carbon. A high adsorption capacity is achieved with an amorphous system

223

(Kang et al., 2018). The amorphous carbon peak in the MBC600-0.6300 and MBC800-0.6300

224

patterns weakened significantly because the thermochemical reaction between the carbon and

225

iron-containing compounds during pyrolysis directly resulted in crystal lattice defects of graphite

226

microcrystallites. The characteristic peaks indicated that Fe was present on the surfaces of

227

MBC600-0.6300 and MBC800-0.6300. The diffraction peaks at 30.1°, 35.62°, 42.3°, 53.6°, 57.2°,

228

and 62.7° corresponding to the (220), (311), (400), (422), (511), and (440) planes, respectively,

229

indicated the presence of Fe3O4. The peaks at 23.82°. 33.03°, 40.72°, 49.25°, and 63.80°

230

corresponded to α-Fe2O3. The Fe3O4 phase was highly crystalline and deposited on the surfaces

231

of the composites, which was indicated by the sharpness of the XRD peaks. These results agree

232

with those of Wu et al. (2017).

233

3.2 Sorption kinetics

234

Fig. 2 shows the effect of the reaction time (0–1440 min) on Cd(II) adsorption by different

235

adsorbents. The adsorption kinetics with BC600, BC800, MBC600-0.6300, and MBC800-0.6300

236

were studied with an initial Cd(II) concentration of 150 mg/L. To understand the adsorption

237

mechanism and potential rate-limiting steps, the experimental data was fitted to pseudo-first-

11

238

order (PFO) and pseudo-second-order (PSO) models, which are the most commonly used

239

models. The models were formulated as follows:

240

PFO model: =
(1 − (− ))

(2)

241

PSO model: /
= 1/(
  )/
+ /


(3)

242

where t is the time, Qt (mg g−1) is the amount of Cd(II) adsorbed by the adsorbent when

243

equilibrium is reached, and k1 and k2 (min-1 and g mg−1 min) are the equilibrium rate constants of

244

the PFO and PSO models, respectively.

245

The Cd(II) sorption capacity (Qt, mg g−1) improved rapidly for the modified adsorbents

246

within the first 2.5 h. This phenomenon occurred on the outer surface of the adsorbents, and the

247

adsorption mechanism was assumed to be physicochemical. Most Cd(II) was sorbed in the first 4

248

h. Then, Cd sorption slowed down, which may have been caused by Cd(II) slowly penetrating

249

the pores and reacting with internal active sites until equilibrium was reached. Hence, it was

250

appropriate for equilibrium to be attained within 6 h. This period was used for further

251

experiments. MBC800-0.6300 adsorbed more Cd(II) than the other three adsorbents within this

252

time, which may have been because it had many active sites and a greater pore volume (0.22 cm3

253

g−1) and surface area (313.88 m2 g−1). A large number of active sites available for adsorption in

254

the initial stage was another contributor. Over time, these active adsorption sites decreased, and

255

the remaining sites were unsuited for Cd(II) removal.

256

Table S1 indicates that the PFO model had a lower correlation coefficient (R2) for the

257

nonlinear plot features than the PSO model. This implies that the PFO model did not suit the

258

experimental data and that the PSO model was more suited. The higher R2 of the PSO model

259

suggests possible chemical adsorption (Ifthikar et al., 2017; Cao et al., 2017) comprising valence

260

forces from the exchange or sharing of electrons between the adsorbate and adsorbent for the

12

261

rate-determining step. The value of Qt calculated with the PSO model was close to the

262

experimental value, which further validated the PSO model as fitting the kinetic adsorption

263

results.

264

3.3 Sorption isotherms

265

Adsorption isotherm models are essential for understanding the adsorption mechanisms in

266

terms of mathematical derivations and fundamental characteristics. The Langmuir and

267

Freundlich isotherm models are commonly used and were fitted to the adsorption data. The

268

Langmuir isotherm model assumes that every molecule has a constant adsorption activation

269

energy and enthalpy and represents homogeneous adsorption. The Freundlich isotherm model is

270

empirical and considers the surface to be heterogeneous. Nonlinear regression was applied to the

271

Freundlich and Langmuir isotherms to study the initial concentration parameters. The Cd(II)

272

adsorption data were fitted to the Langmuir and Freundlich isotherm models with Origin8.0,

273

which are expressed as follows:   !"

274

Q =

275

Q = K % C/'

(4)

# !"

(5)

276

where Qmax (mg g−1) and Qe (mg g−1) are maximum and equilibrium uptake capacities,

277

respectively. KL (L mg−1) is an equilibrium constant and Ce (mg g−1) is the equilibrium Cd(II)

278

concentration. n (dimensionless) represents the bond distribution, which is the heterogeneity

279

factor and Kf (mL3 g−1) is the Freundlich constant and is related to the adsorption ability.

280

As shown in Fig. 3, the correlation coefficients indicated that the Langmuir model fitted the

281

experimental data better than the Freundlich model for the Cd(II) adsorption isotherms of

282

BC600, BC800, MBC600-0.6300, and MBC800-0.6300 with different initial Cd(II) concentrations

283

(10–150 mg L−1) at pH 6. According to the Langmuir parameters, MBC800-0.6300 had the

13

284

highest Cd(II) adsorption ability (Table 2). This may be because of its large number of active

285

sites, large surface area, and strong affinity towards Cd(II). The positively charged surface of BC

286

from electrostatic repulsion decreases adsorption. The Langmuir parameter Qmax indicates

287

monolayer adsorption capacity. The Langmuir maximum Cd(II) sorption capacity of MBC800-

288

0.6300 (46.90 mg g−1) was more significant than those of BC600, BC800 and MBC600-0.6300.

289

Hence, MBC800-0.6300 was chosen for further analysis. Table S2 compares the Cd(II) adsorption

290

capacity of MBC800-0.6300 with those of other sorbents. The higher KL value (0.0533 L mg−1) of

291

MBC800-0.6300 further confirmed that it has a greater affinity with Cd(II). The significant Cd(II)

292

adsorption capacity of MBC800-0.6300 may be due to the high Cd affinity of iron oxide on the

293

MBC800-0.6300 surface, which generates stable inner-sphere complexes or strengthens

294

coordination to the surface functional groups (–OH, –COOH) (Tong et al., 2011). The FeOx

295

phases of MBC800-0.6300 may contribute towards Cd(II) sorption (Zhou et al., 2018) through

296

surface complexation, direct electrostatic retention, precipitation, and ion exchange.

297

3.4 Effect of temperature on Cd(II) sorption

298 299

The changes in Gibbs free energy (∆G°), entropy (∆S°), and enthalpy (∆H°) along with other thermodynamic parameters were determined according to the following equations:

300

ΔG° = −RTlnK 0

(6)

301

ΔG° = ΔH° − TΔS°

(7)

302

As shown in Fig. S3, the temperature affected the Cd(II) adsorption of MBC800-0.6300. The

303

ion mobility increased with the temperature. The increased number of ions interacting with the

304

active sites on the adsorbent surface enhanced the adsorption. The thermodynamic parameters

305

were calculated, and the changes in the derivations of the entropy and enthalpy demonstrated that

306

∆S > 0, ∆G < 0, and ∆H > 0. In other words, the adsorption process is endothermic and

14

307

spontaneous (Table 3), as reported by Wan et al. (2014). The Gibbs free energy change increased

308

with the temperature, which indicates that the driving force of the reaction was amplified. This

309

agrees with the above calculation of the positive enthalpy change.

310

3.5 Effect of the solution pH on Cd(II) sorption

311

The adsorption capacity can be affected by the pH in two different ways: the surface electric

312

charge density or metal ion concentration. These affect the metal deposition and ion exchange

313

reaction in an aqueous solution (Jefferson et al., 2015). The adsorption capacity of MBC800-

314

0.6300 decreased with the pH. The lowest adsorption capacity was recorded at pH 3.0. Although

315

the adsorption increased with the pH (Fig. 4), the metal removal rate of BC slowed down when

316

the pH was greater than 6.0 because the precipitation reaction was dominant (Liang et al., 2017).

317

Chen et al. (2015) showed that, in the presence of BC, Cd(II) precipitates as Cd(OH)2 when the

318

solution pH is greater than 8. In this study, when the pH changed from 3.0 to 6.0, the Cd(II)

319

adsorption rose from 28.2 to 46.9 mg g−1. Then, the Cd(II) adsorption increased from 46.9 to

320

48.3 mg g−1 at pH 6.0–8.0.The decline in the adsorption efficiency at low pH may be because of

321

the abundance of H3O+ or H+ ions, which compete with Cd(II) for the available adsorption sites

322

of MBC800-0.6300. However, the adsorption capacity improves with an increase in pH because

323

of the increased negative charges on the adsorbent surface resulting from deprotonation of

324

carboxylic and hydroxyl functional groups. The literature shows that heavy metal adsorption

325

increases with the solution pH (Usman et al., 2016). This may be credited to the reduced

326

competition between protons and Cd(II) to occupy the available sorption sites. Moreover, the

327

number of binding sites increases with the pH, which increases the total metal adsorption on the

328

adsorbent surface (Usman et al., 2016).The effect of the pH on the Cd(II) adsorption can also be

329

described by the pHPZC results. Table 1 and Fig. S6 indicate that the pHpzc for MBC800-0.6300

15

330

was 5.46. If pH < pHpzc, the Cd(II) removal efficiency may be reduced because of possible

331

electrostatic repulsion between the positively charged MBC800-0.6300 and Cd(II). The Cd(II)

332

adsorption of the adsorbent increases with the pH because of the electrostatic attraction between

333

the negatively charged surface of MBC800-0.6300 and positive Cd(II) ions. Fig. S6 shows that the

334

zeta potential values of all samples shifted towards positive after Cd(II) adsorption. This

335

indicates that Cd(II) was adsorbed as inner-sphere surface complexes or cations, which can

336

neutralize the negative charge of the surface (Wang et al., 2012; Wanze et al., 2006).

337

3.6 Effect of the ionic strength on Cd(II) sorption

338

The ionic strength is an important factor that affects the Cd(II) adsorption process. The

339

thickness of the diffused electric double layer and interface potential of the adsorbent may be

340

influenced by the ionic strength of the solution to discriminate between outer-sphere (sensitive to

341

the ionic strength) and inner-sphere (insensitive to the ionic strength) surface complexes (Hu et

342

al., 2015b). NaNO3 concentrations of 0.001–0.1 M influenced the Cd(II) adsorption capacity of

343

MBC800-0.6300 through the solution ionic strength (Fig. S4). The results demonstrated that Na+

344

did not compete with Cd(II) for negatively charged adsorption sites, and the electrostatic

345

repulsion between two positive ions did not impede Cd(II) loading onto surface sites. These

346

results agree with those of Zhou et al. (2017), who found that the removal mechanism mainly

347

relied on inner-sphere surface complexation.

348

3.7 Effect of the humic acid concentration on Cd(II) sorption

349

The influence of HA in the aquatic environment cannot be ignored because it comprises

350

most of the organic matter in natural water bodies. The Cd(II) adsorption capacity of MBC800-

351

0.6300 increased at high HA concentrations, which suggests that the affinity between Cd(II) and

352

MBC800-0.6300 can be enhanced with HA. However, only a small enhancement was observed

16

353

(Fig. S5). The modified functional groups (e.g., COOH and/or OH) on the MBC800-0.6300

354

surface, ternary BC–HA–Cd surface complexes, and electrostatic attraction helped strengthen the

355

Cd(II) adsorption. For example, Sun et al. (2012) observed that increasing the initial natural

356

organic matter concentration at low and high pH enhances Cu(II) adsorption. Yang et al. (2011)

357

found that adding HA improved metal ion adsorption on kaolinite and MWCNT.

358

3.8 Sorption mechanism

359

3.8.1 FTIR analysis

360

The FTIR spectra were analyzed to better understand the possible changes to the functional

361

groups before and after Cd(II) adsorption. Functional groups on the BC surface play a vital role

362

in metal adsorption and mainly depend on the pyrolysis temperature and type of biomass used

363

(Li et al., 2017). Fig. 5 shows that both MBC600-0.6300 and MBC800-0.6300 had stretching

364

vibrations for ‒OH (hydroxyl) at 3325 and 3433 cm−1, respectively (Ahmed et al., 2016), and

365

C=O (carbonyl, lactones, carboxylic and ester, aromatic structures or benzene rings) at 1610 and

366

1581 cm−1, respectively (Zhang et al., 2015). The peaks at 3325  cm−1 for MBC600-0.6300 and

367

3433 cm−1 for MBC800-0.6300 significantly weakened after Cd(II) adsorption, and the peak of

368

MBC800-0.6300 shifted to 3409  cm−1 (Jiang et al., 2018). These may be because of the

369

interaction between the solution Cd(II) and surface hydroxyl groups, which formed –O–Cd

370

groups on the MBC surface (Zhong et al., 2016). The –OH group detected for MBC600 and

371

MBC800 was because of Fe3O4 adsorption (Wu et al., 2018), which formed complexes of metal–

372

ligand composite (2Fe–O–R–OH + Cd(II) → (2H+ + O–R–O–Fe)2Cd) (Lin et al., 2018).

373

Therefore, MBC600-0.6300 and MBC800-0.6300 had a higher Cd(II) adsorption capacity. After Fe

374

impregnation, a peak at 1341 cm−1 that may be attributed to the Fe3+–OH band of Fe3+–

375

hydroxide or oxyhydroxide (Tabelin et al., 2017) appeared and dramatically weakened after

17

376

Cd(II) sorption. This suggests that Fe significantly contributes to Cd(II) adsorption. The peaks at

377

1049  cm−1 for MBC600-0.6300 and 1022  cm−1 for MBC800-0.6300 can be attributed to the metal

378

hydroxyl (Fe–OH) groups (Liu et al., 2015) and were strengthened and/or produced after

379

magnetization. The peak at 1022 cm−1 for MBC800-0.6300 markedly weakened after Fe

380

impregnation and Cd(II) adsorption (Jiang et al., 2018), which indicates that MBC800-0.6300 has

381

a high Cd(II) adsorption capacity. This adsorption can be assigned to Fe–OH + Cd(II) + H2O →

382

FeOCdOH + 2H+ and the metal hydroxyl group (Fe–OH) (Lin et al., 2018). The bonds at 811

383

and 800 cm−1 can be allocated to the C–H aromatic complexes. Two peaks (633 and 663 cm−1)

384

consistent with the vibration modes of Fe–O (Wang et al., 2014) for magnetite (Zhang et al.,

385

2018) and the peak at 536 cm−1 were attributed to Fe–O stretching vibrations for hematite (Fard

386

et al., 2016). These findings for the MBC600-0.6300 and MBC800-0.6300 surfaces confirmed that

387

iron oxide was loaded onto MBC600-0.6300 and MBC800-0.6300. A similar pattern was detected

388

in the XRD patterns (Fig. 3). Cd(II) can also be adsorbed by electrostatic attraction to form

389

surface complexes (Fe3O4 + Cd(II) → Cd–Fe3O4) (Lin et al., 2018).

390

3.8.2 XPS analysis

391

The XPS spectra were measured to further investigate the surface composition. Fig. 6 shows

392

the XPS high-resolution spectra of Cd 3d, Fe 2p, O 1s and C 1s regions on the adsorbents. Fig.

393

6(a) shows that carbon is present with little oxygen after impregnation and sorption, and the

394

presence of iron and cadmium was observed on the BC. Table 1 indicates that the atomic

395

percentage of oxygen greatly increased after iron impregnation, and the atomic percentage of

396

carbon declined substantially. After Cd(II) sorption, the atomic percentage of carbon increased,

397

whereas those of oxygen and iron declined. The pristine BC and MBC800-0.6300 before and after

398

Cd adsorption showed O/C atomic ratios of 0.11, 0.59, and 0.44, respectively. This implies that

18

399

the functional groups of oxygen declined after Cd(II) sorption but increased after iron

400

impregnation (Fig. 6(a)). To identify the chemical environment of the adsorbed Cd(II), an XPS

401

narrow scan was performed on MBC800-0.6300 after Cd(II) adsorption, as shown in Fig. 6(b).

402

Two sharp peaks centered at around 405 and 411.84 eV were linked to Cd 3d5/2 and Cd 3d3/2,

403

respectively (Hidalgo et al., 2015). The two peaks at 404.4 and 404.6 eV may be assigned to Cd–

404

O (42.61%), and the peaks at 405.2 and 406.1 eV can be attributed to Cd(OH)2 (57.39%). These

405

findings agree with those of Zhou et al. (2019). Cd(II) adsorption onto MBC800-0.6300 was

406

confirmed by XPS analysis.

407

The Fe 2p high-resolution XPS spectra in Fig. 6(c) indicate the presence of Fe 2p3/2 at

408

707.4–711.9 eV, Fe 2p1/2 at 723–725.2 eV. This clearly confirmed the existence of Fe3+ and Fe2+

409

ions, which may indicate the presence of iron oxides (FexOy), iron oxyhydroxide (FeOOH), or

410

iron hydroxides (Fe(OH)x) (Grosvenor et al., 2004). A satellite was found at a binding energy of

411

718.93 eV. Wu et al. (2012) showed the same results for Fe3O4 nanoparticles on graphene oxide

412

surfaces. A significant shift for Fe 2p binding energies was also observed after Cd(II) adsorption

413

onto MBC800-0.6300, which implies that a coordination reaction occurred between Cd(II) and

414

iron. This agrees with the FTIR and XRD results.

415

Fig. 6(d) shows broad peaks in the O1s region that indicate different chemical states of

416

oxygen such as inorganic oxygen (Fe–O/Cd–O) and organic oxygen (C–OH/C–O–C) (Datsyuk

417

et al., 2008). In BC800, elemental oxygen existed only in the form of C–O–C. The binding

418

energies of 530.8 and 530.25 eV in the O 1s region were ascribed to iron-bonded inorganic

419

oxygen (i.e., FeO, Fe2O3, or Fe3O4) (Hu et al., 2015a). The O1s peaks at 532.2 eV were ascribed

420

to organic oxygen (C–OH/C–O–C) (Zhou et al., 2018). The shifts were distinct in the O 1s

421

region of MBC800-0.6300. This indicates that the sorption of cadmium onto the FeO, Fe2O3, or

19

422

FeOOH phases may have been because the energy considerably weakened at the 530.8 eV peak.

423

The new peak at 529.5 eV appeared after cadmium adsorption and may have been due to oxygen

424

bonding with Cd(II) (Cd–O) (Fig. 6d). Cadmium and oxygen binding during the Cd(II)

425

adsorption of MBC800-0.6300 is consistent with the Cd 3d high-resolution region of the XPS

426

spectra.

427

4. Conclusion

428

MBC derived from corn straw powder was doped with iron oxide and analyzed using

429

different techniques (SEM, XRD, FTIR, XPS, and VSM). The resulting MBC composites

430

manifested tremendous physicochemical properties such as more oxygen-containing functional

431

groups, a fine-pore structure, and large surface area. Cd(II) successfully adsorbed onto MBC800-

432

0.6300 because of the iron adhering to the BC surface. After sorption, the used MBC in the

433

aqueous solution could be collected with ease by the application of an external magnetic force.

434

The batch experiments indicated that the Cd(II) removal was pH-dependent. MBC800-0.6300 is

435

suited to Cd(II) removal because of its exceptional adsorption ability of heavy metals and lack of

436

secondary pollution. Further research is required to investigate the recycling rate and

437

toxicological effects of MBC800-0.6300 before it can be used for practical applications.

438

Conflicts of interest

439 440 441 442

The authors declare no conflict of interest. Acknowledgments This study was funded by the National Natural Science Foundation of China (No. 41771525) and STU Scientific Research Foundation for Talents (NTF19025).

443 444

20

445

References

446

ATSDR. Toxicological Profile for Cadmium, US Department of Health and Human Services.

447 448 449 450

Public Health Service, USA, Atlanta, GA (2012) Cao, C.Y., Cheng H. L., Yan Y., Li Y. D., 2017. Thermal activation of serpentine for adsorption of cadmium. Journal of Hazardous Materials 329, pp.222–229. Chen, T., Zhou, Z., Han, R., Meng, R., Wang, H. and Lu, W., 2015. Adsorption of cadmium by

451

biochar derived from municipal sewage sludge: impact factors and adsorption

452

mechanism. Chemosphere, 134, pp.286-293.

453

Chowdhury, P., Athapaththu, S., Elkamel, A. and Ray, A.K., 2017. Visible-solar-light-driven

454

photo-reduction and removal of cadmium ion with Eosin Y-sensitized TiO2 in aqueous

455

solution of triethanolamine. Separation and Purification Technology, 174, pp.109-115.

456

Chowdhury, P., Elkamel, A., Ray, A.K., 2014. Photocatalytic processes for the removal of toxic

457

metal ions. In: Sharma, S.K. (Ed.), Heavy Metals in Water: Presence, Removal and Safety.

458

Royal Society of Chemistry, Cambridge, pp. 25–43.

459

Datsyuk, V., Kalyva, M., Papagelis, K., Parthenios, J., Tasis, D., Siokou, A., Kallitsis, I. and

460

Galiotis, C., 2008. Chemical oxidation of multiwalled carbon nanotubes. Carbon, 46(6),

461

pp.833-840.

462

Fard, G.C., Mirjalili, M. and Najafi, F., 2017. Hydroxylated α-Fe2O3 nanofiber: optimization of

463

synthesis conditions, anionic dyes adsorption kinetic, isotherm and error analysis. Journal

464

of the Taiwan Institute of Chemical Engineers, 70, pp.188-199.

465 466

Grosvenor, A.P., Kobe, B.A., Biesinger, M.C. and McIntyre, N.S., 2004. Investigation of multiplet splitting of Fe 2p XPS spectra and bonding in iron compounds. Surface and

21

467

Interface Analysis: An International Journal devoted to the development and application of

468

techniques for the analysis of surfaces, interfaces and thin films, 36(12), pp.1564-1574.

469

Hameed, B.H., Din, A.M. and Ahmad, A.L., 2007. Adsorption of methylene blue onto bamboo-

470

based activated carbon: kinetics and equilibrium studies. Journal of hazardous

471

materials, 141(3), pp.819-825.

472

Hanigan, D., Zhang, J., Herckes, P., Krasner, S.W., Chen, C. and Westerhoff, P., 2012.

473

Adsorption of N-nitrosodimethylamine precursors by powdered and granular activated

474

carbon. Environmental science & technology, 46(22), pp.12630-12639.

475

Hidalgo, K.T.S., Guzmán-Blas, R., Ortiz-Quiles, E.O., Fachini, E.R., Corchado-García, J.,

476

Larios, E., Zayas, B., José-Yacamán, M. and Cabrera, C.R., 2015. Highly organized

477

nanofiber formation from zero valent iron nanoparticles after cadmium water

478

remediation. Rsc Advances, 5(4), pp.2777-2784.

479

Hidayat, D., A. Purwanto, W.N. Wang, K. Okuyama. 2010. Preparation of size-controlled

480

tungsten oxide nanoparticles and evaluation of their adsorption performance. Mater. Res.

481

Bull., 45, pp. 165-173

482

Hu, X., Ding, Z., Zimmerman, A.R., Wang, S. and Gao, B., 2015a. Batch and column sorption of

483

arsenic onto iron-impregnated biochar synthesized through hydrolysis. Water Research, 68,

484

pp.206-216.

485

Hu, R.; Wang, X.; Dai, S.; Shao, D.; Hayat, T.; Alsaedi, A. Application of Graphitic Carbon

486

Nitride for the Removal of Pb(II) and Aniline from Aqueous Solutions. Chem. Eng. J.

487

2015b, 260, 469 −477.

22

488

Ifthikar, J., Wang, J., Wang, Q., Wang, T., Wang, H., Khan, A., Jawad, A., Sun, T., Jiao, X. and

489

Chen, Z., 2017. Highly efficient lead distribution by magnetic sewage sludge biochar:

490

sorption mechanisms and bench applications. Bioresource technology, 238, pp.399-406.

491

Ippolito, J.A., Strawn, D.G., Scheckel, K.G., Novak, J.M., Ahmedna, M. and Niandou, M.A.S.,

492

2012. Macroscopic and molecular investigations of copper sorption by a steam-activated

493

biochar. Journal of environmental quality, 41(4), pp.1150-1156.

494

Jefferson, W.A., Hu, C., Liu, H. and Qu, J., 2015. Reaction of aqueous Cu–Citrate with MnO2

495

birnessite: Characterization of Mn dissolution, oxidation products and surface

496

interactions. Chemosphere, 119, pp.1-7.

497

Jiang, L., Ye, Q., Chen, J., Chen, Z. and Gu, Y., 2018. Preparation of magnetically recoverable

498

bentonite–Fe3O4–MnO2 composite particles for Cd (II) removal from aqueous

499

solutions. Journal of colloid and interface science, 513, pp.748-759.

500 501 502

Joseph, S. and Lehmann, J., 2009. Biochar for environmental management: science and technology (No. 631.422 B615bi). London, GB: Earthscan. Karunanayake, A.G., Todd, O.A., Crowley, M., Ricchetti, L., Pittman Jr, C.U., Anderson, R.,

503

Mohan, D. and Mlsna, T., 2018. Lead and cadmium remediation using magnetized and

504

nonmagnetized biochar from Douglas fir. Chemical Engineering Journal, 331, pp.480-491.

505

Li, H., Dong, X., da Silva, E.B., de Oliveira, L.M., Chen, Y. and Ma, L.Q., 2017. Mechanisms of

506

metal sorption by biochars: biochar characteristics and modifications. Chemosphere, 178,

507

pp.466-478.

508 509

Liang, J., Li, X., Yu, Z., Zeng, G., Luo, Y., Jiang, L., Yang, Z., Qian, Y. and Wu, H., 2017. Amorphous MnO2 modified biochar derived from aerobically composted swine manure for

23

510

adsorption of Pb (II) and Cd (II). ACS Sustainable Chemistry & Engineering, 5(6), pp.5049-

511

5058.

512

Lin, J., Su, B., Sun, M., Chen, B. and Chen, Z., 2018. Biosynthesized iron oxide nanoparticles

513

used for optimized removal of cadmium with response surface methodology. Science of The

514

Total Environment, 627, pp.314-321.

515

Liu, L., Peng, Q., Qiu, G., Zhu, J., Tan, W., Liu, C., Zheng, L. and Dang, Z., 2019. Cd2+

516

adsorption performance of tunnel-structured manganese oxides driven by electrochemically

517

controlled redox. Environmental pollution, 244, pp.783-791.

518

Liu, M., Yang, L. and Zhang, L., 2016. Functionalization of magnetic hollow porous oval shape

519

NiFe2O4 as a highly selective sorbent for the simultaneous determination of five heavy

520

metals in real samples. Talanta, 161, pp.288-296.

521

Liu, R., Liu, F., Hu, C., He, Z., Liu, H. and Qu, J., 2015. Simultaneous removal of Cd (II) and Sb

522

(V) by Fe–Mn binary oxide: positive effects of Cd (II) on Sb (V) adsorption. Journal of

523

hazardous materials, 300, pp.847-854.

524

Ma, Y., Liu, W.J., Zhang, N., Li, Y.S., Jiang, H. and Sheng, G.P., 2014. Polyethylenimine

525

modified biochar adsorbent for hexavalent chromium removal from the aqueous

526

solution. Bioresource technology, 169, pp.403-408.

527

Mohan, D., Kumar, H., Sarswat, A., Alexandre-Franco, M. and Pittman Jr, C.U., 2014. Cadmium

528

and lead remediation using magnetic oak wood and oak bark fast pyrolysis bio-chars.

529

Chemical Engineering Journal, 236, pp.513-528.

530

Pan, S., Zhang, Y., Shen, H. and Hu, M., 2012. An intensive study on the magnetic effect of

531

mercapto-functionalized nano-magnetic Fe3O4 polymers and their adsorption mechanism

24

532

for the removal of Hg (II) from aqueous solution. Chemical Engineering Journal, 210,

533

pp.564-574.

534

Reddy, D.H.K. and Lee, S.M., 2013. Three-dimensional porous spinel ferrite as an adsorbent for

535

Pb

(II)

removal

from

aqueous

536

Research, 52(45), pp.15789-15800.

solutions. Industrial

& Engineering

Chemistry

537

Sayed, M., Tabassum, S., Shah, N.S., Khan, J.A., Shah, L.A., Rehman, F., Khan, S.U., Khan,

538

H.M. and Ullah, M., 2019. Acid fuchsin dosimeter: a potential dosimeter for food

539

irradiation dosimetry. Journal of Food Measurement and Characterization, 13(1), pp.707-

540

715.

541

Son, E.B., Poo, K.M., Chang, J.S. and Chae, K.J., 2018. Heavy metal removal from aqueous

542

solutions using engineered magnetic biochars derived from waste marine macro-algal

543

biomass. Science of the Total Environment, 615, pp.161-168.

544

Song, Z., Lian, F., Yu, Z., Zhu, L., Xing, B. and Qiu, W., 2014. Synthesis and characterization of

545

a novel MnOx-loaded biochar and its adsorption properties for Cu2+ in aqueous

546

solution. Chemical Engineering Journal, 242, pp.36-42.

547

Sun, W.L., Xia, J., Li, S. and Sun, F., 2012. Effect of natural organic matter (NOM) on Cu (II)

548

adsorption by multi-walled carbon nanotubes: relationship with NOM properties. Chemical

549

engineering journal, 200, pp.627-636.

550 551

Tong, X.J., Li, J.Y., Yuan, J.H. and Xu, R.K., 2011. Adsorption of Cu (II) by biochars generated from three crop straws. Chemical Engineering Journal, 172(2-3), pp.828-834.

552

Trakal, L., Veselská, V., Šafařík, I., Vítková, M., Číhalová, S. and Komárek, M., 2016. Lead and

553

cadmium sorption mechanisms on magnetically modified biochars. Bioresource technology,

554

203, pp.318-324.

25

555

Usman, A., Sallam, A., Zhang, M., Vithanage, M., Ahmad, M., Al-Farraj, A., Ok, Y.S.,

556

Abduljabbar, A. and Al-Wabel, M., 2016. Sorption process of date palm biochar for

557

aqueous

558

Pollution, 227(12), p.449.

559

Cd

(II)

removal:

Efficiency

and

mechanisms. Water,

Air,

&

Soil

Wan, S., Ma, M., Lv, L., Qian, L., Xu, S., Xue, Y. and Ma, Z., 2014. Selective capture of

560

thallium

(I)

ion

from

aqueous

solutions

by

amorphous

561

dioxide. Chemical Engineering Journal, 239, pp.200-206.

hydrous

manganese

562

Wang, Y., Sun, H., Ang, H.M., Tadé, M.O. and Wang, S., 2014. Magnetic Fe3O4/carbon

563

sphere/cobalt composites for catalytic oxidation of phenol solutions with sulfate

564

radicals. Chemical Engineering Journal, 245, pp.1-9.

565

Wang, Z., S.W. Lee, J.G. Catalano, J.S. Lezama-Pacheco, J.R. Bargar, B.M. Tebo, D.E.

566

Giammar. 2012. Adsorption of uranium(VI) to manganese oxides: X-ray absorption

567

spectroscopy and surface complexation modeling. Environ. Sci. Technol., 47, pp. 850-858

568

Wazne, M., X. Meng, G.P. Korfiatis, C. Christodoulatos. 2006. Carbonate effects on hexavalent

569

uranium removal from water by nanocrystalline titanium dioxide. J. Hazard. Mater., 136,

570

pp. 47-52

571

Wu, H., Feng, Q., Yang, H., Alam, E., Gao, B. and Gu, D., 2017. Modified biochar supported

572

Ag/Fe nanoparticles used for removal of cephalexin in solution: Characterization, kinetics

573

and mechanisms. Colloids and Surfaces A: Physicochemical and Engineering Aspects, 517,

574

pp.63-71.

575

Wu, H., Gao, G., Zhou, X., Zhang, Y. and Guo, S., 2012. Control on the formation of Fe3O4

576

nanoparticles on chemically reduced graphene oxide surfaces. CrystEngComm, 14(2),

577

pp.499-504.

26

578

Wu, J., Yi, Y., Li, Y., Fang, Z. and Tsang, E.P., 2016. Excellently reactive Ni/Fe bimetallic

579

catalyst supported by biochar for the remediation of decabromodiphenyl contaminated soil:

580

reactivity, mechanism, pathways and reducing secondary risks. Journal of hazardous

581

materials, 320, pp.341-349.

582

Yang, S., Hu, J., Chen, C., Shao, D. and Wang, X., 2011. Mutual effects of Pb (II) and humic

583

acid adsorption on multiwalled carbon nanotubes/polyacrylamide composites from aqueous

584

solutions. Environmental science & technology, 45(8), pp.3621-3627.

585

Yap, M.W., Mubarak, N.M., Sahu, J.N. and Abdullah, E.C., 2017. Microwave induced synthesis

586

of magnetic biochar from agricultural biomass for removal of lead and cadmium from

587

wastewater. Journal of Industrial and Engineering Chemistry, 45, pp.287-295.

588

Zamri, M.F.M.A., Kamaruddin, M.A., YusofSf, M.S., Aziz, H.A. and Foo, K.Y., 2017. Semi-

589

aerobic stabilized landfill leachate treatment by ion exchange resin: isotherm and kinetic

590

study. Applied Water Science, 7(2), pp.581-590.

591

Zhang, M.M., Liu, Y.G., Li, T.T., Xu, W.H., Zheng, B.H., Tan, X.F., Wang, H., Guo, Y.M.,

592

Guo, F.Y. and Wang, S.F., 2015. Chitosan modification of magnetic biochar produced from

593

Eichhornia crassipes for enhanced sorption of Cr (VI) from aqueous solution. RSC

594

Advances, 5(58), pp.46955-46964.

595

Zhang, S., Cui, M., Chen, J., Ding, Z., Wang, X., Mu, Y. and Meng, C., 2019. Modification of

596

synthetic zeolite X by thiourea and its adsorption for Cd (II). Materials Letters, 236,

597

pp.233-235.

598

Zhang, X., Sun, C., Zhang, L., Liu, H., Cao, B., Liu, L. and Gong, W., 2018. Adsorption studies

599

of cadmium onto magnetic Fe3O4@FePO4 and its preconcentration with detection by

600

electrothermal atomic absorption spectrometry. Talanta, 181, pp.352-358.

27

601

Zhong, L.B., Yin, J., Liu, S.G., Liu, Q., Yang, Y.S. and Zheng, Y.M., 2016. Facile one-pot

602

synthesis of urchin-like Fe–Mn binary oxide nanoparticles for effective adsorption of Cd

603

(II) from water. Rsc Advances, 6(105), pp.103438-103445.

604

Zhou, J., He, J., Li, G., Wang, T., Sun, D., Ding, X., Zhao, J. and Wu, S., 2010. Direct

605

incorporation of magnetic constituents within ordered mesoporous carbon− silica

606

nanocomposites for highly efficient electromagnetic wave absorbers. The Journal of

607

Physical Chemistry C, 114(17), pp.7611-7617.

608 609

Zhou, L., Huang, Y., Qiu, W., Sun, Z., Liu, Z. and Song, Z., 2017. Adsorption properties of nano-MnO2–biochar composites for copper in aqueous solution. Molecules, 22(1), p.173.

610

Zhou, Q., Liao, B., Lin, L., Qiu, W. and Song, Z., 2018. Adsorption of Cu (II) and Cd (II) from

611

aqueous solutions by ferromanganese binary oxide–biochar composites. Science of the total

612

environment, 615, pp.115-122.

613

Zhou, Q., Liao, B., Lin, L., Song, Z., Khan, Z.H. and Lei, M., 2019. Characteristic of adsorption

614

cadmium

of

red

soil

amended

with

615

composite. Environmental Science and Pollution Research, 26(5), pp.5155-5163.

28

a

ferromanganese

oxide-biochar

616

Figures

15

MBC600-0.6300 MBC800-0.6300

-1

Magnetization M (emu g )

10

5

0

-5

-10

-15 -15000

617 618 619

-10000

-5000

0

5000

10000

15000

Intensity of magnetic field H (Oe)

Fig. 1. Magnetization curves of MBC600-0.6300 and MBC800-0.6300. The inset shows their attraction to a permanent magnet.

29

40

Qt (mg g -1)

30

20

10

BC600 MBC600-0.6300

0

0

200

400

600

800

BC800 MBC800-0.6300 1000

1200

1400

1600

Time (min) 620 621

Fig. 2. Adsorption kinetics of Cd(II) with MBC600-0.6300 and MBC800-0.6300.

30

50 BC600 BC800 MBC600-0.6300 40

MBC800-0.6300

Qe (mg g -1)

Langmuir Freundlich 30

20

10

0 0

622 623

20

40

60

80

-1

100

120

140

Ce (mg L ) Fig. 3. Sorption isotherms of Cd(II) with MBC600-0.6300 and MBC800-0.6300.

31

50

pH 3 pH 4 pH 5 pH 6 pH 7 pH 8

-1

Qe (mg g )

40

30

20

10

0 0

20

40

60

80

100

120

-1

624 625

Ce (mg L )

Fig. 4. Effect of pH on Cd(II) adsorption by MBC800-0.6300.

32

BC600 MBC600-0.6300

451

633

811

1049

1341

1610

3326

75

536

663

3306

80

1022

MBC600-0.6300 adsorbed Cd

1341

Transmitance (a.u.)

90 85 80 75 70 65 60 55 50 45 40 35 30 25 20 85

70

1699

60 55 50 45 40 4000

BC800 MBC800-0.6300 MBC800-0.6300 adsorbed Cd 3500

3000

2500

2000

1581

65

1500

1000

500

-1

Wavenumber (cm )

626 627

Fig. 5. FTIR spectra of BC600, BC800, MBC600-0.6300 and MBC800-0.6300 before and after

628

Cd(II) adsorption.

33

a20 x10

4

b

(a) BC600 (b) BC800 (c) MBC600-0.6300(d) MBC800-0.6300

Cd(OH)2

900

(e) MBC600-0.6300 adsorbed Cd (f ) MBC800-0.6300 adsorbed Cd

C1s

3d3/2

15

800

C/S

C/S

O1s

Fe2p

10

Cd-O 3d3/2

700

600

Cd3d

5

500

1400

1200

1000

800

600

400

200

414

0

412

410

629

C

d

5200 5000

Fe2p1/2 FeOOH

Fe2p

408

406

404

Binding energy (eV)

Binding energy (eV) Fe2p3/2 3+ Fe

Fe2p3/2 2+ Fe

4800

4000 3500

C-O

O1s MBC800-0.6300

Fe-O

3000

4600

2500

4400

Fe2p3/2 metal

4200 4000

2000

2+ 3+ Satellite peak of Fe and Fe

1500

3800

C/S

1000 3600

C/S

3400

MBC800-0.6300 730

725

720

715

710

500 536

705

534

532

530

8000 3500

Fe2p1/2 FeOOH

Fe2p3/2 3+ Fe

7000

MBC800-0.6300 Cd adsorbed Fe-O

3000

C/S

2500

C-O

6000

Fe2p3/2 2+ Fe

5000

Cd-O

4000 2000

2+ 3+ Satellite peak of Fe and Fe

3000

1500

MBC800-0.6300 Cd adsorbed

2000

1000 730

725

720

715

533

710

532

531

530

529

528

Binding energy (eV)

Bindinng energy (eV)

630 631 632 633 634

Fig. 6. (a) XPS analysis of BC600, BC800, MBC600-0.6300, and MBC800-0.6300 before and after Cd(II) adsorption; (b) Cd3d spectrum of MBC800-0.6300 after Cd(II) adsorption; (c) Fe2p spectrum of MBC800-0.6300 before and after Cd(II) adsorption; (d) O1s spectrum of MBC8000.6300 before and after Cd(II) adsorption.

635

34

636

Tables Table 1. Selected physical and chemical properties of adsorbents

637

Adsorbents

C (%)

N (%)

BC600 BC800 MBC6000.6300 MBC8000.6300

84.35 86.14 59.05 48.65

H (%)

O (%)

Fe (%)

SBET (m2 g−1) 97.2 93.7 225.9

Vtotal (cm3 g−1) 0.17 0.05 0.11

Pore volume (nm) 1.84 2.27 3.85

pHPZC

1.36 2.69 8.89 1.11 1.34 8.25 2.18 2.12 20.7 6.47

Ash content (%) 7.23 10.79 13.58

2.62 3.50 3.87

Pore size (nm) 1.84 2.28 3.85

1.42 1.01 29.4 7.07

17.66

313.9

0.22

2.86

5.46

2.86

638

35

Table 2. Parameters of Cd(II) adsorption with different adsorbents

639

Adsorbents BC600 BC800 MBC600-0.6300 MBC800-0.6300

Langmuir Qmax (mg g−1) 16.44 14.32 28.71 46.90

−1

KL (L mg ) 0.0115 0.0363 0.0465 0.0533

640 641 642 643 644 645 646 647 648 649 650 651 652 653 654 655 656 657 658 659

36

2

R 0.970 0.897 0.994 0.991

Freundlich Kf (mL 3 g−1) 0.5747 1.7531 3.9463 6.5562

1/n 0.6138 0.4028 0.3931 0.4012

R2 0.995 0.989 0.964 0.955

660

Table 3. Thermodynamic parameters for the adsorption of Cd(II) by MBC800-0.8400 at different

661

temperatures. T (K)

Qe (mg g−1)

lnKL

∆G (kJ mol−1)

288

43.02

8.986

-21.5166

298

45.36

9.066

-22.4624

308

46.15

9.287

-23.7822

∆H (kJ mol−1) 11.17

662 663 664 665 666 667 668 669 670 671 672 673 674 675 676 677 678 679 680 681

37

∆S (kJ mol−1 K−1) 0.1133

682

Supplementary information

683 684

Fig. S1. Scanning electron microscopy (SEM) images of (a) BC600; MBC600-0.6300 (b) before

685

and (c) after adsorption; (d) BC800; and MBC800-0.6300 (e) before and (f) after adsorption.

38

♦Fe

O3

2


Fe3O4


intensity (a.u)







♦








MBC600-0.6300

BC600 ♦







♦









♦

MBC800-0.6300

BC800 10 686 687

20

30

40

50

60

70

80

2-Theta/degree Fig. S2. XRD patterns of BC800, MBC800-0.6300, BC600, and MBC600-0.6300.

39

90

50

288K 298K 308K

-1

Qe (mg g )

40

30

20

10

0

20

40

60

80

100

-1

Ce (mg L ) 688 689

Fig. S3. Effect of temperature on Cd(II) adsorption by MBC800-0.6300.

690

40

50

0.001 NaNO3 0.01 NaNO3 0.1

NaNO3

-1

Qe (mg g )

40

30

20

10

0

20

40

60

80

100

-1

691 692

Ce (mg L )

Fig. S4. Effect of ionic strength on Cd(II) adsorption by MBC800-0.6300.

41

50

-1

HA = 10 mg L -1 HA = 15 mg L -1 HA = 20 mg L

-1

Qe (mg g )

40

30

20

10

0

20

40

60

80

100

-1

693 694

Ce (mg L )

Fig. S5. Effect of humic acid on Cd(II) adsorption by MBC800-0.6300.

695

42

BC600 BC800 MBC600-0.6300

15 10

MBC800-0.6300 MBC800-0.6300 Cd adsorbed

5

Zeta potential (mV)

0 -5 -10 -15 -20 -25 -30 -35 -40 -45 -50 2

4

6

696 697

8

10

pH

Fig. S6. Zeta potential of magnetic and non-magnetic biochars and treated MBC800-0.6300.

698

43

b7

4

6 Volume (%)

Volume (%)

a5 3 2 1

5 4 3 2 1 0

0 0.02

0.2

2 20 200 Particle size (µm)

699

0.02

2000

5

5 Volume (%)

d6

Volume (%)

4 3 2

0

e

2000

2

0

700

200

3

1

2 20 200 Particle size (µm)

20

4

1

0.2

2

Particle size (µm)

c6

0.02

0.2

0.02

2000

f

4

0.2

2 20 200 Particle size (µm)

2000

4

3.5 3 Volume (%)

Volume (%)

3 2.5 2 1.5 1

2 1

0.5 0 0

0.02 0.02

701

0.2

2 20 200 Particle size (µm)

0.2

2

20

200

2000

2000 Particle size (µm)

702

Figure S7. Particle size distributions of (a) BC600, (b) BC800, (c) MBC600-0.6300, (d) MBC800-

703

0.6300, (e) MBC600-0.6300 with adsorbed Cd, and (f) MBC800-0.6300 with adsorbed Cd.

44

Table S1. Parameters for the dynamic fit of BC and modified materials.

704

Adsorbents BC600 BC800 MBC600-0.6300 MBC800-0.6300

Pseudo-first order Qe (mg g−1) K1 10.09 0.0262 12.12 0.0425 22.29 0.0182 35.31 0.0133

2

R 0.968 0.969 0.992 0.987

705 706 707 708 709 710 711 712 713 714 715 716 717 718 719 720 721 722 723

45

Pseudo-second order Qe (mg g−1) K2 (g mg−1min−1) 10.82 0.0033 13.02 0.0027 24.32 0.0009 39.26 0.0004

R2 0.990 0.986 0.995 0.990

724 725

Table S2. Comparison of Cd(II) adsorption capacities of MBC800-0.6300 and different absorbents in the literature. Surface area (m2/g) 459

Cd adsorption capacity (mg/g) 11

7 5

-

13.51 16.64

Karunanayake et al., 2018 Zhang et al., 2018 Liu et al., 2016

-

5

127

38.3

Trakal et al., 2016

25

5

8.80

7.40

Mohan et al., 2014

25

5

6.10

2.87

Mohan et al., 2014

25

4.8

834

4.77

Yap et al., 2017

25

6

0.97 63.33 313.9

23.16 19.40 46.90

Son et al., 2018 Son et al., 2018 Present study

Adsorbents

Temp. pH

MBC (Douglas fir biochar) Fe3O4@FePO4 magnetic hollow porous oval shape NiFe2O4 Grape husk/FeSO4.7H2O Magnetic oak bark biochar Magnetic Oak wood biochar Magnetic coconut shell biochar KBCmag − 0.05 HBCmag − 0.05 MBC800-0.6300

25

5

25

726 727

46

References

Highlights 1. Magnetic biochars were produced by using a one-step synthesis. 2.The maximum adsorption capacity of MBC800-0.6300 for Cd(II) is 46.9 mg g-1. 3. The Langmuir model is a good fit for adsorption process of Cd(II) to define monolayer sorption. 4. The Cd(II) -loaded MBC800-0.6300 can be conveniently collected by a magnet.

Author Contribution Statement Zhengguo Song conceived of the idea of this study and provided financial means. Weiwen Qiu and Md. Shafiqul IslamS provided significant input on experimental design. Zulqarnain Haider Khan preformed laboratory experiments. Minling Gao and Zulqarnain Haider Khan interpreted histological data. Zulqarnain Haider Khan and Minling Gao designed image analysis methods. Zulqarnain Haider Khan and Zhengguo Song analysed the data and prepared the manuscript, all authors contributed substantially to revisions.

Declaration of interests ☐ The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. ☐The authors declare the following financial interests/personal relationships which may be considered as potential competing interests: