Magnesium Oxide Embedded Nitrogen Self-Doped Biochar

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Magnesium Oxide Embedded Nitrogen Self-doped Biochar Composites: Fast and High-Efficiency Adsorption of Heavy Metals in an Aqueous Solution Li-Li Ling, Wu-Jun Liu, Shun Zhang, and Hong Jiang Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.7b02382 • Publication Date (Web): 28 Jul 2017 Downloaded from http://pubs.acs.org on July 29, 2017

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Environmental Science & Technology

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Magnesium Oxide Embedded Nitrogen Self-doped Biochar Composites: Fast and

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High-Efficiency Adsorption of Heavy Metals in an Aqueous Solution

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Li-Li Ling, Wu-Jun Liu*, Shun Zhang, Hong Jiang*

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CAS Key Laboratory of Urban Pollutant Conversion, Department of Chemistry,

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University of Science and Technology of China, Hefei 230026, China

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*Corresponding authors

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E-mail: [email protected] (H. J.), [email protected] (W.-J. L.)

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Tel/Fax: +86-551-63607482

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ACS Paragon Plus Environment


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ABSTRACT

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Lead (Pb) pollution in natural water bodies is an environmental concern due to toxic effects on

14

aquatic ecosystems and human health, while adsorption is an effective approach to remove Pb

15

from the water. Surface interactions between adsorbents and adsorbates play a dominant role in

16

the adsorption process, and properly engineering a material’s surface property is critical to the

17

improvement of adsorption performance. In this study, the magnesium oxide (MgO) nanoparticles

18

stabilized on the N-doped biochar (MgO@N-biochar) was synthesized by one-pot fast pyrolysis of

19

an MgCl2-loaded N-enriched hydrophyte biomass, as a way to increase the exchangeable ions and

20

N-containing functional groups and facilitate the adsorption of Pb2+. The as-synthesized

21

MgO@N-biochar has a high performance with Pb in an aqueous solution with a large adsorption

22

capacity (893 mg/g), a very short equilibrium time (< 10 min), and a large throughput (~4450 BV).

23

Results show that this excellent adsorption performance can be maintained with various

24

environmentally relevant interferences including pH, natural organic matter, and other metal ions,

25

suggesting that the material may be suitable for the treatment of wastewater, natural bodies of

26

water, and even drinking water. In addition, MgO@N-biochar quickly and efficiently removed

27

Cd2+ and tetracycline. Multiple characterizations and comparative tests have been performed to

28

demonstrate the surface adsorption and ion exchange contributed to partial Pb adsorption, and it

29

can be inferred from these results that the high performance of MgO@N-biochar is mainly due to

30

the surface coordination of Pb2+ and C=O or O=C−O, pyridinic, pyridonic, and pyrrolic N. This

31

work suggests that engineering surface functional groups of biochar may be crucial for the

32

development of high performance heavy metal adsorbents.

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Table of Contents

Mg2+

35 36 37 38

3



39

INTRODUCTION

40

Hydrophytes with high capacity N uptake in eutrophicated water provide a feasible

41

approach to the control of eutrophication.1 The harvest of a mature hydrophyte

42

biomass is essential for effectively preventing the release of N in bodies of water

43

during periods of withering.2 A proper method for effectively minimizing and

44

recycling large amounts of hydrophyte biomass is still a great challenge due to poor

45

storage properties and high transportation cost.

46

Fast pyrolysis is the thermal decomposition of organic solid waste with the

47

absence of O2 at a mediate temperature (400-600 oC) in a very high heating rate. This

48

process is considered a promising technology for converting hydrophyte biomass into

49

valuable bio-oil and biochar,3-5 thus minimizing waste and recycling a large amount of

50

the hydrophyte biomass. Bio-oil can be used in the production of biofuel or for

51

chemical feedstocks,6, 7 while biochar (about one-third of feedstock) can be used as a

52

platform carbon material for diverse purposes because of its stable carbon skeleton

53

and abundant O- and N- containing functional groups.8

54

Aside from the eutrophication, heavy metal pollution (e.g., Pb, Hg, Cd, Cu, and

55

Ni) in natural water bodies is an environmental concern due to toxic effects on aquatic

56

ecosystems and human health.9, 10 Pb is the most common heavy metal found in the

57

hazardous waste sites,11 often entering an aquatic environment from mineral ore

58

dissolution and industrial effluents.12,

59

drinking water where it is often diffused in water distribution systems.14 Because of

60

this high physiological toxicity, Pb is considered as a priority pollutant and ecological

13

Pb is also a primary micropollutant of

4


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hazard, and securing a method for quickly and completely removing it from bodies of

62

water is imperative.15, 16

63

Adsorption using biochar is generally regarded as the most effective method for

64

removing Pb contamination from aquatic environments, but raw biochar often

65

exhibits relatively low adsorption capacity and requires long equilibrium time because

66

of its limited surface functional groups and porous structure.4 For example, the

67

biochar derived from sesame straw has been applied for Pb removal with a maximum

68

capacity of 102 mg/g and an equilibrium time of 24 hours,17 while biochar derived

69

from Alternanthera philoxeroides has a maximum Pb adsorption capacity of 257.1

70

mg/g and an equilibrium time of 2.5 hours.18 Other methods such as surface oxidation

71

and amination can endow biochar with abundant surface functional groups (e.g., C=O,

72

COOH, NH2, and OH) and greatly improve adsorption performance. For example,

73

Tan et al. reported that a mesoporous poly-melamine-formaldehyde with high density

74

amine groups can remarkably improve the removal efficiency of low concentrations

75

of Pb.19 Huang et al. synthesized a mesoporous EDTA-modified silica SBA-15 with

76

abundant surface carboxyl groups which effectively removed Pb2+ under experimental

77

conditions.20 Zhao et al. observed enhanced performance with synthesized

78

few-layered graphene oxide nanosheets with abundant oxygen-containing functional

79

groups used to remove heavy metals from large-volume aqueous solutions.21 Cao and

80

Harris22 prepared the P-enriched biochar from dairy manures by heating at low

81

temperature under air-rich condition, and the authors found that the as-prepared

82

biochar shows excellent performance for Pb removal. It has also been discovered that 5



83

surface oxidation with KMnO4 improves the maximum sorption capacity of the

84

engineered biochar for Pb(II) about 2.1-fold compared to pristine biochar.23

85

The incorporation of inorganic nanostructures into biochar is another approach

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for improving adsorption performance. MgO, a typical alkaline earth metal oxide, is

87

an effective and desirable adsorbent for the removal of heavy metals from aquatic

88

environments because it is naturally abundant, environmentally friendly, and an

89

excellent adsorption material. Compared to the bulk material, nanoscale MgO exhibits

90

a higher capacity and faster adsorption rate of heavy metals because it has more

91

surface active sites.24, 25 These MgO nanoparticles have an agglomeration tendency

92

because of their nanoscale size and high surface energy, limiting their wider

93

application as adsorbents. MgO nanoparticle incorporation into the biochar matrix

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may be an effective approach to enhance the stability of MgO nanoparticles.

95

Elaborating on previous research,26-29 in this work, the MgCl2 is introduced into

96

an N-enriching hydrophyte biomass that is easily produced by the adsorption of

97

MgCl2 from seawater (average content 0.45 wt.%) using biomass as sorbents, and

98

then synthesized the MgO nanoparticles and embedded nitrogen self-doped biochar

99

via fast pyrolysis. In the pyrolysis process, the N-enriching hydrophyte biomass

100

decomposed and formed a porous biochar matrix with self-doped N. The resulting

101

biochar continued to support dispersion and stabilization of MgO NPs formed during

102

the hydrolysis and decomposition of MgCl2 during the pyrolysis process. The main

103

objective of this study was to obtain new MgO nanoparticles embedded in nitrogen

104

self-doped biochar (MgO@N-biochar), and to demonstrate the adsorption 6


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performance towards Pb2+ from wastewater by determining the adsorption capacity,

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kinetics rate, and stability under the interference of various environmental factors

107

including pH, natural organic matter, and other metal ions. The mechanisms of

108

interaction between Pb2+ and MgO@N-biochar were investigated using multiple

109

characterizations and comparative adsorption tests. This work offers a new alternative

110

to transform biomass waste into a selective adsorbent for Pb2+ removal and provides

111

mechanism insights of the interaction between heavy metal ions and biochar-based

112

adsorbents.

113 114

EXPERIMENTAL SECTION

115

Synthesis of MgO Nanoparticles Embedded Nitrogen Self-doped Biochar

116

(MgO@N-biochar). The materials used in this work are described in Text S1 of

117

supporting information. The MgO@N-biochar was synthesized though a 2-stage

118

pyrolysis process within the same reactor. The first stage is to convert the raw

119

biomass into biochar under a non-isothermal heating program (fast pyrolysis). The

120

fast pyrolysis of Mg preloaded T. angustifolia biomass at 400-600 oC with heating

121

rates of about 300-800 oC/s were carried out in the reactor described in previous work

122

(Fig. S1 of supporting information, SI).30 The pyrolysis reactor was heated to set

123

values (400-600 oC) and simultaneously purged by 0.4 L/min of N2 flow. Five grams

124

of the Mg preloaded T. angustifolia biomass were fed into the reactor through a piston.

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The biomass was heated for 1-2 s and decomposed to form MgO@N-biochar, and the

126

volatiles in the pyrolysis process were swept out by N2 flow of 0.2 L/min and 7



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127

condensed using a cold ethanol trap to form bio-oil. The design of the trap is

128

presented in Fig. S1 of SI, in which two condenser was filled with cold ethanol

129

(stored in a low-temperature refrigerator before use and the temperature of this cold

130

ethanol is about -20 oC). After pyrolysis, the solid residue was further subjected to an

131

isothermal carbonization process under the same temperature of the fast pyrolysis for

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another 1 hour to further carbonize the remained biochar, then the reactor was moved

133

out and cooled down to room temperature under the nitrogen flow (200 mL/min).

134

The composition and structure of the MgO-biochar are analyzed by various

135

techniques, and the details of the characterization are shown in Text S2 of supporting

136

information.

137

Performance of MgO@N-biochar for Pb Removal. The performance of

138

MgO@N-biochar was evaluated according to the amount of Pb2+ removed from the

139

water system. The experiments were first conducted in a batch model as follows: 25

140

mg of MgO@N-biochar was placed into a beaker flask containing 25 mL of Pb2+

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solution in different concentrations. The adsorption solution was shaken at room

142

temperature for 30 min, then filtered through a 0.22 µm membrane. The initial

143

solution pHs in the range of 2.0 to 7.0 were adjusted by 2.0 mol/L of aqueous HNO3

144

or NaOH solution and monitored with a pH meter. The Pb2+ concentration in the

145

filtrate was measured through an atomic absorption spectrometer (4530F, Shanghai

146

Precision & Scientific Instrument Co., Ltd. Shanghai, China) and Pb adsorption

147

capacities q (mg/g) at time t (min) were calculated with Eq. 1:

148

（C − Ct）V q= 0 m

(1) 8


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where C0 and Ct (mg/L) are the Pb2+ concentration at initial and time t (min); V (L) is

150

the Pb solution volume, and m (g) is the adsorbent amount.

151

The long-term and cycle performance of the MgO@N-biochar were investigated

152

by a fixed-bed column adsorption and an adsorption-desorption cycle, respectively,

153

described in Text S3 of supporting information.

154 155

RESULTS AND DISCUSSIONS

156 157

Mg2+ Loading and Biomass Pyrolysis. The MgO@N-biochar was prepared through

158

an integrated adsorption-pyrolysis process as illustrated in Fig. S2 of SI, and the N

159

balance during the pyrolysis process was calculated in Text S4 of SI. The MgCl2 in

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the solution was adsorbed by the T. angustifolia biomass to produce the MgCl2 loaded

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biomass. Because the interaction between the biomass and metal ions is usually a

162

single-layer process that occurs on the surface functional groups,31 the adsorbed Mg2+

163

and their hydrolyzed forms are mono-layer dispersed on the surface biomass after

164

drying and pyrolysis. In the following pyrolysis process, adsorbed MgCl2 can be

165

converted into mono-layer dispersed MgO and other partial decomposition products

166

(e.g., MgCl2·2H2O and (MgOH)Cl) (Eqs. 2-4)32 under high temperature and reductive

167

conditions.33-35 MgCl2●6H2O → MgCl2●2H2O + 4 H2O↑

(2)

MgCl2●2H2O →(MgOH)Cl + H2O↑ + HCl↑

(3)

(MgOH)Cl → MgO + HCl↑

(4) 9



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Meanwhile, since the temperature of the biomass feedstock increased at a high rate

169

(e.g., 300-800

170

lignocellulosic biomass (e.g., lignin, cellulose, and hemicelluloses) quickly

171

decomposed to produce volatile species which can be condensed to form bio-oil to be

172

used for the production of biofuels or valuable chemicals.7, 36, 37

173

o

C/s) in the pyrolysis process, the main components of the

Characterization

of

the

MgO@N-biochar.

The

Mg

contents

of

174

MgO@N-biochar materials increased from 12.4 to 19.5 wt.% as pyrolysis temperature

175

increased from 400 to 600 oC. Similar results were also observed in the pyrolysis of

176

other metals (e.g., Pb and Cu) in the preloaded biomass.38, 39 This phenomenon is

177

reflected in Eqs. 2-4, where the MgCl2 was easily hydrolyzed and decomposed to

178

form MgO, which does not volatilize at high temperatures though most of the biomass

179

species during the pyrolysis process tended to volatilize as pyrolysis temperature

180

increased.

181

The XRD pattern of MgO@N-biochar materials is shown in Fig. 1a. This pattern

182

shows two different crystalline phases which can be attributed to MgO and

183

(MgOH)Cl.40 As pyrolysis temperature increased to 500 oC, the XRD peaks for

184

(MgOH)Cl decreased remarkably while those for MgO increased significantly. This

185

suggests that high temperature is favorable for the conversion of (MgOH)Cl into MgO.

186

When pyrolysis temperature increased to 600 oC, the crystalline phase of (MgOH)Cl

187

disappeared, leaving only the MgO crystalline phase.

188

Fig. 1b shows the N2 adsorption-desorption isotherms of the MgO@N-biochar and

189

N-biochar. The results show that the surface area and pore volume of 10


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MgO@N-biochar is much higher than those of raw N-biochar. The porous structure of

191

MgO@N-biochar (Fig. S6 of SI) is mainly attributed to the decomposition of MgCl2

192

during the biomass pyrolysis process, with the release of volatile matter like HCl and

193

H2O (Eqs. 2-4) resulting in the formation of pore structure in the biochar matrix. In

194

addition, the newly formed Mg variations in the pyrolysis process, including MgO

195

and MgOHCl, may act as an in situ template for the generation of a porous structure

196

in the biochar matrix. Furthermore, at high temperatures, the MgCl2 itself has a strong

197

dehydration ability for carbohydrate polymers like cellulose and hemicellulose, thus

198

changing the decomposition pathway of the lignocellulosic biomass and suppressing

199

the formation of heavy tars that can block pore structure and facilitating the

200

generation of open pores in the biochar matrix.41, 42

201

The surface morphology and micro-composition of MgO@N-biochar were

202

analyzed with SEM-EDX and TEM. Figure 2a displays the SEM image and EDX

203

spectrum of MgO@N-biochar-400, showing a porous morphology composed of rough

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carbon sheets and rods, while the EDX spectrum shows C, N, O, Mg, and Cl to be the

205

main elements. These results are in agreement with the XRD results showing MgO

206

and (MgOH)Cl to be the main crystalline phases in MgO@N-biochar-400. For

207

comparison, the SEM image of the raw N-biochar is displayed in Fig. 2b, showing a

208

relative smooth carbon sheet without nanoparticles on the surface. The EDX spectrum

209

shows C, N, and O to be the main elements in the N-biochar.

210

Pb Removal Performance of the MgO@N-biochar. MgO@N-biochar

211

performance was evaluated according to the selective adsorption of Pb from the 11



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212

wastewater under different influence factors. Fig. 3 shows the decrease of Pb

213

concentrations in the wastewater over time using different adsorbents. This metric

214

shows us that Pb adsorption by MgO@N-biochar was very fast, with equilibrium

215

being achieved within 10 min and the Pb removal rates reaching 99% at an initial Pb

216

concentration of 100 mg/L. The kinetics of Pb adsorption on the MgO@N-biochar are

217

fitted by four different models, namely pseudo-first-order model, pseudo-second order

218

model, Elovich model, and intraparticle diffusion model, and the results are presented

219

in Text S5 of SI. From the R2 values of these models, the pseudo-second order model

220

is more suitable than other three models to describe the Pb2+ adsorption kinetic

221

behavior by the MgO@N-biochar. This phenomenon suggests that chemical

222

adsorption could be the rate-controlling mechanism for the adsorption of Pb2+ on

223

MgO@N-biochar.

224

The adsorption isotherms of different MgO@N-biochar materials are examined

225

by changing the initial Pb concentrations in the range of 10 to 1000 mg/L (Fig. 3b).

226

The results were fitted to the Dubinin Radushkevich (D-R) and Langmuir isotherm

227

models, and the results are shown in Text S6 of SI. The maximum Pb adsorption

228

capacity

229

MgO@N-biochar-400, Such a high adsorption capacity (893 mg/g) and short

230

equilibrium time (10 min) places MgO@N-biochar in a notable position among

231

state-of-the-art adsorbents for Pb removal (Table S6 of SI).

calculated

from

the

Langmuir

model

is

893

mg/g

for

the

232

Influence of Environmental Factors. Generally, in a natural or industrial

233

application, many environmentally relevant interference factors may influence the 12


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behavior of Pb adsorption by an adsorbent. pH is a critical environmental factor

235

affecting the behavior of heavy metal ion adsorption, which not only influences the

236

form of surface functional groups of adsorbents but also the conditions of the heavy

237

metal ions in the aqueous solution. Figure 4a shows the influence of pH on Pb

238

adsorption in MgO@N-biochar-400. The adsorption capacities of the adsorbent

239

remained nearly unchanged in a pH range of 3 to 7, but significantly decreased with a

240

pH of 2, suggesting that MgO@N-biochar-400 has a strong anti-pH-interference

241

ability. This phenomenon is at odds with previous references where pH has had a

242

large influence on Pb adsorption.43,

243

functional groups, including OH, NH2, C=O, COOH, and MgO, on the surface of

244

MgO@N-biochar, which act as buffer agents to maintain pH in a relatively stable

245

region (Fig. S3 of SI). As the results confirm, when the initial pH was in the range of

246

3 to 7, the equilibrium pH remained in the relatively stable range of 6.6 to 9.0. The

247

increase of equilibrium pH can be explained as follows: In the aqueous solution, the

248

free Pb2+ can be hydrolyzed to produce H+ (Eq. 5)

44

This is because of the large number of

249

Pb2+ + H2O Pb(OH)+ + H+

250

During the adsorption process, an ion-exchange interaction between Pb2+ and

251

Mg2+ happens, with more Mg2+ released from the adsorbent to the solution, and the

252

free Pb2+ is adsorbed to the adsorbent, thus less H+ is produced from the hydrolysis.

253

Meanwhile, the considerable amount of MgO in the adsorbent can capture the already

254

existed H+ to form the Mg(OH)+ (Eq. 6)

255

(5)

MgO + H+ Mg(OH)+

(6) 13



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Due to the decrease of H+ production and increase of H+ capture in the adsorption

257

process, the equilibrium pH values can increase in compared to the initial ones.

258

In a practical sense, industrial wastewater and natural bodies of water usually

259

contain various inorganic ions (e.g., Na+, K+, SO42-, Cl-) which may interfere with the

260

adsorption of heavy metals by adsorbents.45, 46 To evaluate the effects of ionic strength,

261

the experiments were carried out by adding Na+ (NaNO3) at different concentrations

262

into the aqueous solution. Figure 4b shows the effects of ionic strength (Na+) on the

263

adsorption of Pb by the MgO@N-biochar-400. It can be deduced that Pb adsorption

264

with MgO@N-biochar-400 was not influenced by Na+ even though its concentration

265

was 1000 times larger than that of Pb. The main reason for this phenomenon may be

266

that Na+ has a low charge density and large ionic size, leading to a stronger interaction

267

between Na+ and the surrounding H2O than the solid adsorbent.47,

268

monovalent ions, the influence of multivalent metal ions (take Ca2+ and Al3+ as the

269

examples for divalent and trivalent ions, respectively) on the Pb adsorption is also

270

investigated. It can be seen that the presence of Ca2+ and Al3+ has no significant

271

influence on the Pb adsorption by the MgO@N-biochar when their concentration is

272

the same as Pb2+, and even the presence of all the Ca2+, Mn2+, Al3+, and Fe3+ has no

273

remarkable effect on Pb adsorption when their concentrations are the same with the

274

Pb2+ (Figs. S4 and S5 of SI), indicating a robust anti-interference ability of the

275

MgO@N-biochar in Pb adsorption. Such a high anti-interference ability of the

276

MgO@N-biochar in Pb adsorption can be explained from the view of material

277

structure or surface characteristics of the adsorbent. From the view of material 14


48

Besides the

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structure, as shown in the XRD pattern (Fig. 1a), apart from the MgO, Mg(OH)Cl is

279

also found in the biochar@MgO-400 which can react with Pb2+ to form the PbCl2.

280

Because of the high ion radius and unique electronic configuration ([Xe]4f14 5d10 6s2

281

6p2), the Pb2+ is hard to be polarized, and the formed Pb2+ is insoluble (Ksp=1.7 x 10-5),

282

thus removing the Pb2+ from the water. Meanwhile, due to the relatively low high ion

283

radius, many other metal ions (e.g., Zn2+, Cd2+, Ni2+, Cu2+, Fe3+, Mn2+and so on) are

284

easily to be polarized, and their chlorides usually have high solubility. Therefore, the

285

presence of Mg(OH)Cl in the biochar@MgO-400 can selectively remove Pb2+ from

286

the water with the co-exist of other many other metal ions like Zn2+, Cd2+, Ni2+, Cu2+,

287

Fe3+, Mn2+ and so on.49, 50 From the view of surface characteristics, the abundant

288

surface functional groups (e.g., COOH, NH2, C=O, and OH) on the biochar can

289

coordinate with Pb2+ to form the Pb-organic complexes,51 which can selectively

290

remove Pb2+ from the water with co-exist of some metal ions (e.g., K+, Ca2+, and Al3+)

291

and organic matters (e.g., humic acid). Because of the unique material structure or

292

surface characteristics, the biochar@MgO materials have high selectivity toward Pb2+

293

adsorption.

294

Humic acid is a type of natural organic matter widely found in water systems and

295

reported to have great influence on the adsorption of heavy metals including Pb(II),

296

Cd(II), Cu(II), and Cr(VI).52 Figure 4c shows Pb adsorption performance of

297

MgO@N-biochar-400 under different concentrations of humic acid. It can be deduced

298

that after adding humic acid, the adsorption capacities of MgO@N-biochar-400

299

remained nearly unchanged with a humic acid concentration of 100 mg/L, suggesting 15



300

a high anti-jamming capability in MgO@N-biochar-400 towards Pb adsorption. The

301

main reason for the phenomenon that the Pb adsorption by MgO@N-biochar is

302

insusceptible to humic acid can be explained as follows: (1) ion-exchange between

303

Pb2+ and Mg2+ is one of the main contributions to the adsorption capacity, which

304

cannot be affected significantly by the presence of humic acid; (2) the chemical

305

interactions between Pb2+ and surface functional groups (e.g., NH2, OH, and COOH)

306

are another main contribution to the adsorption capacity. These interactions can also

307

not be influenced by the humic acid since the humic acid does not compete with Pb2+

308

to the surface functional groups of the adsorbent. Indeed, the presence of humic acid

309

may influence the pH values of the solution, thus affecting the adsorption process, but

310

in the case of this work, the robust buffer effects of surface functional groups can

311

maintain the solution pH during the adsorption process, thus minimizing the impact of

312

humic acid on the Pb adsorption by the MgO@N-biochar.

313

Long-term and cycle performance of the adsorbent. To evaluate the potential

314

practical applications of MgO@N-biochar, a fixed-bed column adsorption was carried

315

out by feeding an influent containing 20 mg/L of Pb2+ with a single bed volume (BV)

316

of 2.4 cm3. Based on the state standard for integrated wastewater discharge in China

317

(GB: 8978-1996), the breakthrough point of Pb2+ was set as 1.0 mg/L. As shown in

318

Fig. 4d, the effective treatment volume for Pb2+ containing wastewater by 0.5 g of

319

MgO@N-biochar was 10,070 mL (~4450 BV), with Pb concentration of effluents

320

lower than 1.0 mg/L. These results indicate that MgO@N-biochar can be employed as

321

an adsorbent for long term application in a typical wastewater system. Apart from the 16


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322

long-term performance of the adsorbent, its cycle performance is also very important

323

for practical applications. As shown in Fig. 4e. The MgO@N-biochar shows a

324

favorable recycle performance toward Pb adsorption, during which the Pb removal

325

efficiency remained almost unchanged for 10 times cycle use.

326

Mechanism for Pb Adsorption. The above mentioned results have demonstrated

327

the high Pb adsorption capacity and ultrafast adsorption kinetics of as-synthesized

328

MgO@N-biochar. This excellent adsorption performance is closely related to

329

abundant surface functional groups of the material. Chemical adsorption plays an

330

important role in the removal of Pb,53, 54 especially for biochar-based materials with

331

limited porous structure. To experimentally demonstrate the ion-exchange interaction,

332

the concentration change of Mg2+ during Pb2+ adsorption (Fig. 4f) has been

333

determined. Compared to raw biochar, MgO@N-biochar released a small quantity of

334

Mg2+ into the solution without the addition of Pb2+, while significantly high levels of

335

Mg2+ were released during Pb2+ adsorption. Based on these results, the Pb adsorption

336

amount owing to the ion-exchange between Pb2+ and Mg2+ (ηion-exchange) is calculated,

337

with the hypothesis of one Mg2+ can be exchanged by one Pb2+ (Eq. 7)

338

η ion −exchange = (mMg / WMg ) ×

WPb ×100% QPb

(7)

339

where mMg is the amount of Mg released from the adsorbent to the solution, WMg and

340

WPb are the molecular of Mg and Pb, respectively, and QPb is the total adsorbed Pb.

341

According to Eq. 6, the Pb adsorption amount owing to the ion-exchange between

342

Pb2+ and Mg2+ is calculated as about 42%, indicating that the ion-exchange is a main

343

contribution to the Pd adsorption by the MgO@N-biochar. Similar trends are also 17



344

observed in the adsorption of Cd2+ by the same adsorbent (Figs. S7 and S8). The

345

ion-exchange mechanism can be also confirmed by the XPS results. As is shown in

346

Fig. 5a, two peaks at binding energies of 307 and 352 eV can be attributed to the Mg

347

Auger photoelectrons, which became significantly weak after Pb adsorption. The

348

weakening of these two peaks was primarily due to Mg2+ ion-exchange on the surface

349

of MgO@N-biochar with the Pb2+ in the aqueous solution. In the XRD pattern of

350

MgO@N-biochar after Pb adsorption, most of the XRD peaks attributed to the MgO

351

and MgOHCl became weak or even disappeared, further confirming the involvement

352

of Mg2+ in Pb adsorption (Fig. S9). Similar results were reported in previous works.55,

353

56, 57

354

which can be assigned to the Pb 4f photoelectron, confirming the adsorption of Pb by

355

MgO@N-biochar.

A new peak of binding energy is found at 137 eV in the Pb adsorbed material

356

Apart from the ion-exchange which contributes about 42% of total Pb adsorption,

357

additional interactions should also be responsible for the high Pb adsorption capacity.

358

Figure 5b shows the XPS C 1s spectra of MgO@N-biochar before and after Pb

359

adsorption, where the C 1s spectrum of MgO@N-biochar before Pb adsorption

360

comprises five peaks attributable to the sp3 C−C (284.3 eV), sp2 C=C (284.9 eV),

361

C−O/C−N (285.6 eV), C=O (286.7 eV), and O=C−O (289.6 eV), respectively,58, 59

362

while after Pb adsorption, the binding energies for the peaks of sp2 C=C and sp3 C−C

363

remain nearly unchanged but the binding energies for the peaks of C−O/C−N, C=O,

364

and O=C−O decrease substantially. Such changes can be attributed to the formation of

18


Page 18 of 29

Page 19 of 29


365

carbonyl/carboxyl-Pb complexes.51 Therefore, C−O/C−N, C=O, and O=C−O are the

366

main functional groups involved in Pb adsorption by [email protected]

367

Figure 5c shows the XPS N 1s spectra of MgO@N-biochar before and after Pb

368

adsorption. The N 1s spectrum of MgO@N-biochar presents 4 peaks which can be

369

assigned to pyridinic N (398.9 eV), pyridonic N (399.9 eV), pyrrolic N (400.8 eV),

370

and quaternary N (401.5 eV), respectively.61, 62 After Pb adsorption, only the binding

371

energy of quaternary N remained effectually unchanged while the binding energies of

372

the other three peaks decreased markedly. Such decreases are mainly due to the N

373

atoms in these functional groups sharing their spare electrons with Pb resulting in

374

reduction in electron density. These results suggest that pyridinic N, pyridonic N, and

375

pyrrolic N are also main functional groups that contributed to the high adsorption

376

capacity of MgO@N-biochar.

377

In summary, the high adsorption performance of MgO@N-biochar toward Pb

378

may be attributed to interactions between abundant functional groups on

379

MgO@N-biochar and Pb or Cd ions. The ion-exchange interaction and surface

380

adsorption also played an important role in Pb adsorption. These results suggest that

381

engineering surface functional groups may be a more plausible solution for high

382

performance adsorbents toward heavy metals. The environmental implications of this

383

work are analyzed in Text S7 of SI.

384 385

ASSOCIATED CONTENT

386

Supporting Information 19



387

Texts S1-S7, Tables S1-S5, Figures S1-S9, and references S1-S18. These materials are

388

available free of charge on the ACS Publications website

389 390

AUTHOR INFORMATION

391

Corresponding Authors:

392

*(H. J.) Fax: +86-551-63607482, E-mail: [email protected];

393

*(W.-J. L.) E-mail: [email protected]

394

Notes

395

The authors declare no competing financial interest.

396 397

ACKNOWLEDGEMENTS

398

The authors gratefully acknowledge financial support from National Natural Science

399

Foundation of China (21677138, 21607147), China Postdoctoral Science Foundation

400

(2015M580553), and the Fundamental Research Funds for the Central Universities

401

(WK2060190063).

402 403

REFERENCES

404 405 406 407 408 409 410 411 412 413

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565 566 567

24


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Page 25 of 29


568

569

570 571

Fig. 1. (a) XRD patterns of the MgO@N-biochar synthesized at different pyrolysis

572

temperatures and the raw N-biochar; and (b) N2 adsorption-desorption isotherms of

573

the MgO@N-biochar-400 and the raw N-biochar.

25



574

575 576

Fig. 2 (a) SEM image of the MgO@N-biochar-400 and its EDX spectrum; (b) SEM

577

image of the raw biochar and its EDX spectrum (the Cu element is derived from the

578

Cu grid supporter used in the SEM test, and the Pt is the metal spraying used to

579

improve the conductivity the materials)

26


Page 26 of 29

Page 27 of 29


580 581

Fig. 3. (a) The time dependent adsorption of Pb by the MgO@N-biochar and

582

N-biochar (initial Pb concentration: 100 mg/L; adsorption dosage: 1.0 g/L); (b) the

583

adsorption isotherms of the Pb adsorption by the the MgO@N-biochar and N-biochar

584

(adsorption dosage: 1.0 g/L; time: 30 min).

27



585

586

587 588

Fig. 4. (a) Influence of pH on the adsorption of Pb by the MgO@N-biochar; (b)

589

Influence of ionic strength the adsorption of Pb by the MgO@N-biochar; (c)

590

Influence of humic acid the adsorption of Pb by the MgO@N-biochar. (initial Pb

591

concentration, 100 mg/L; adsorption time, 30 min). (d) Breakthrough curves of Pb2+

592

adsorption by MgO@N-biochar-400. (e) Cycle performance of the MgO-biochar

593

toward Pb adsorption. (f) The released concentration of Mg2+ during Pb2+ adsorption

594

of different biochar materials

28


Page 28 of 29

Page 29 of 29


595

596

597 598

Fig. 5. XPS spectra of MgO@N-biochar before and after Pb adsorption, (a) XPS

599

survey spectra; (b) C 1s spectra before and after Pb adsorption; and (c) N 1s spectra

600

before and after adsorption. 29


Magnesium Oxide Embedded Nitrogen Self-Doped Biochar

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