Adsorption of cadmium by biochar derived from municipal sewage sludge: Impact factors and adsorption mechanism

Chemosphere 134 (2015) 286–293

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Adsorption of cadmium by biochar derived from municipal sewage sludge: Impact factors and adsorption mechanism Chen Tan a, Zhou Zeyu a, Han Rong b, Meng Ruihong a, Wang Hongtao a,⇑, Lu Wenjing a,⇑ a b

School of Environment, Tsinghua University, Beijing 100084, China School of Environmental Science and Engineering, Chang’an University, Xi’an 710064, China

h i g h l i g h t s  pH and biochar dosage are key factors for adsorption by sludge-derived biochar.  Equilibrium temperature has weak adsorption effect.  The mechanism of adsorption by biochar involves surface precipitation and ion exchange.

a r t i c l e

i n f o

Article history: Received 1 November 2014 Received in revised form 9 April 2015 Accepted 17 April 2015

Keywords: Adsorption Heavy metal Biochar Impact factor Mechanism

a b s t r a c t Static equilibrium experiments were carried out to investigate the impact factors and the mechanism of cadmium adsorption on biochar derived from municipal sewage sludge. An appropriate dosage of biochar is sufficient; in the experiment, 0.2% is the optimal dosage for the largest removal capacity, while the removal capacity of biochar reduces with the increasing dosage. pH is another dominant factor of the adsorption process. The removal capacity of biochar is lower than 20 mgg1 when the solution initial pH is lower than 2 pH units, comparatively retaining more than 40 mgg1 at the solution initial pH higher than 3 pH units. Temperature has weak influence on the adsorptive performance. The main mechanism of the adsorption process of biochar for cadmium mainly involves (1) surface precipitation by forming insoluble cadmium compounds in alkaline condition, and (2) ion exchange for cadmium with exchangeable cations in the biochar, such as calcium ions. Ó 2015 Elsevier Ltd. All rights reserved.

1. Introduction Heavy metals are one category of the most significant contaminants, causing many environmental issues that threaten human health and ecological systems, such as the ‘‘itai–itai disease’’ and ‘‘minamata disease’’ (Taty-Costodes et al., 2003; Wang and Chen, 2009). Both industrial and agricultural effluents discharge large amount of heavy metals into surface water, as well contaminating ground water in trace amount by leaching from the earth surface after rainfall or snow (Di Natale et al., 2008; Kılıç et al., 2013). Cadmium, one of the most toxic heavy metals, has become a significant concern because of its solubility, mobility and biological accumulation (Sud et al., 2008; Belhalfaoui et al., 2009), which is ubiquitous throughout the world, and could lead to bone and

⇑ Corresponding authors. E-mail addresses: [email protected] (H. Wang), [email protected]. cn (W. Lu). http://dx.doi.org/10.1016/j.chemosphere.2015.04.052 0045-6535/Ó 2015 Elsevier Ltd. All rights reserved.

kidney damage after prolonged exposure (Volesky and Holan, 1995; Sud et al., 2008). Traditional treatment methods to remove heavy metals from aqueous solutions include coagulation, chemical precipitation, ion exchange and membrane separation processes. However, given the high cost and significant amount of residual sludge of these traditional methods, adsorption is considered a better choice, especially at low concentration (Wang et al., 2009; Kılıç et al., 2013). Despite the widespread use and perfect adsorption performance, alternative materials for activated carbon as the most popular adsorbent are still needed due to its expensive price (Babel and Kurniawan, 2003; Febrianto et al., 2009; Stavropoulos and Zabaniotou, 2009; Moreno-Barbosa et al., 2013). Biochar, a form of black carbon, is a potential low-price adsorbent with ideal adsorption efficiency, which is always produced as a by-product of biomass pyrolysis for energy recovery (Chun et al., 2004; Peng et al., 2012; Deveci and Kar, 2013; Wang et al., 2013). The mechanism of heavy metal adsorption onto biochar varies, which depends on the properties of both biochar and heavy metals,

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including: (1) surface adsorption via coordination to p electrons (C@C); (2) precipitation as insoluble matters such as hydroxide, phosphate and carbonate; (3) metal exchange with cations; (4) surface complexation with free carboxyl functional groups and hydroxyl functional groups; and (5) electrostatic interactions (Cao et al., 2009; Uchimiya et al., 2010; Harvey et al., 2011; Lu et al., 2012; Kılıç et al., 2013). Pyrolysis is a promising technique that transforms biomass residue of different origin, among others, sludges into fuel (syngas and bio-oil) and emits less pollutants (Shinogi and Kanri, 2003; Yanik et al., 2007; Folgueras et al., 2013; Mašek et al., 2013). Biophysical drying process can reduce sludge moisture content with less energy input, and as well fine and loose particles form from raw sludge by microbial activity; due to the higher thermal conductivity, after the following fast pyrolysis, syngas with high heat value is obtained, which composes of 42.6 vol.% H2 (0.0181 g H2g1 dried sludge) at 900 °C (Han et al., 2012). Our previous work also reported that biochar derived by the above two-step process has much higher adsorption ability for cadmium than commercial activated carbon, and 900 °C is the optimal temperature for both energy recovery and heavy metal removal (Chen et al., 2014). Adsorbent dosage, solution pH and equilibrium temperature are important factors, that could strongly influence adsorption performance. The resolution of adsorption mechanism is the foundation of removal process controlling. Analyzing the impact factors and the adsorption mechanism is the prerequisite of biochar application in heavy metal removal from aqueous solutions.

2. Materials and methods Analytical reagent (AR) grade chemicals and deionized water (DW) were used throughout this study. All the labware was soaked in dilute nitric acid at least overnight, thoroughly flushed with tap water, and washed three times with DW.

2.1. Biochar preparation Municipal sewage sludge with an initial moisture content of 82.1% sampled from the Xiaojiahe Municipal Sewage Treatment Plant in Beijing, China, was treated via biophysical drying and fast pyrolysis as described by Han et al. (2012). The moisture content of the sludge after seven days biophysical drying decreased to approximately 25%, and then reduced to lower than 2.5% by air drying, denoting the residue as ‘‘dried sludge’’. The pyrolysis temperature was 900 °C, and the solid result biochar was ground through a 40 mesh sieve (0.45 mm) without further activation, abbreviated as ‘‘BC900’’. The main characteristics of BC900 are as following reported by Chen et al. (2014): ash content percentage 88.07 ± 0.56 wt.%; C element 15.92 ± 2.74 wt.%, H element 0.11 ± 0.11 wt.%, O element 2.439 ± 0.575 wt.%, N element 0.53 ± 0.07 wt.%; Ca element 69.56 gkg1, Mg element 17.52 gkg1; pH 12.15 (S/L = 1:10), pHPZC 10.17, cation exchange capacity 247.51 ± 7.49 cmolkg1; surface area 67.603 m2g1 (Brunauer-Emmett-Teller [BET] model), average pore size 3.840 nm, pore volume 0.09855 cm3g1. To remove exchangeable ions and soluble alkaline substances, 2 g of BC900 was soaked in 800 mL DW overnight followed by filtering with 0.45 lm polysulfone filter membrane and flushing with plenty of DW until the effluent pH reached between 6 and 7, and then oven dried at 60 °C to constant weight. The residue biochar was referred as ‘‘washed BC900’’. To exclude the adsorption distribution of organic matters, the BC900 residue after calcinating to constant weight during approximately 120 min at 650 °C was referred as ‘‘BC900 ash’’.

2.2. Adsorption equilibrium experiment Cd2+ stock solution was prepared by dissolving Cd(NO3)24H2O in DW, and the Cd2+ concentration in Cd2+ stock solution was 2000 mgL1. 20.00 mL nitric acid (15 M) and 9.25 g Ca(OH)2 was mixed and diluted with DW to 500 mL, and Ca2+ stock solution was desired at a Ca2+ concentration of 10 000 mgL1. Afterwards, Cd2+-bearing solutions were prepared by diluting the stock solutions to specific concentrations. In the pH influence experiments, the set pH level of Cd2+-bearing solutions (initial Cd2+ concentration of 200 mgL1) was adjusted by 1 M HNO3 and NaOH solutions before adding the adsorbent; the initial pH of the Cd2+bearing solutions was not adjusted in other cases. Approximately 50 mg of absorbents, such as BC900, dried sludge, washed BC900 and BC900 ash, was placed into a 40 mL glass bottle. Then, 25 mL of Cd2+-bearing solution was added and intensively mixed using a vortex maker. After stirring using a thermostatic box overnight at specific temperature, the suspension was filtered with 0.45 lm polysulfone filter membrane. The residual Cd2+ and released Ca2+ and Mg2+ concentrations were determined by ICP–OES (IRIS Intrepid II XSP Spectrometer, ThermoFisher, USA). All adsorption experiments were run in triplicate, and the blank solution was measured for quality control. To investigate the influence of BC900 dosage, the BC900 mass placed into an equilibrium system was 25, 50, 125, 200 and 250 mg, with the S:L (the solid mass/the liquid volume) value of 0.1%, 0.2%, 0.5%, 0.8% and 1.0%, respectively. The removal percentage and removal capacity of Cd2+ were calculated as follows:

R ¼ ðC 0  C e Þ=C 0  100%

ð1Þ

Q ¼ ðC 0  C e ÞV=m

ð2Þ 2+

Where, R is the removal percentage of Cd (%); C0 and Ce are the initial and equilibrium concentrations of Cd2+ (mgL1); Q is the removal capacity of Cd2+ at equilibrium (mgg1); V is the volume of the solution (mL) and m is the weight of the absorbents (mg). 2.3. Other characterization After adsorption equilibrium, BC900 was picked up onto a carbon-coated copper grid and then air dried. The micromorphology was observed by an S-5500 scanning electron microscope (SEM. Hitachi, Japan). BC900 after equilibrium in Cd2+-bearing solutions at different Cd2+ concentrations were dried at 60 °C. The samples were characterized by X-ray powder diffraction (XRD) on a D8 advance X-ray diffractometer (Bruker/AXS, Germany) at 40 kV and 40 mA for monochramatized Cu Ka (k = 0.15418 nm) radiation with a scanning rate of 8°min1. Fourier-transformed infrared (FTIR) spectra were investigated in the 4000–400 cm1 region under a 4 cm1 resolution using a Spectrum GX spectrometer (Perkin Elmer, USA). The baseline of the raw data was adjusted and then the modified data were normalized, by OMNIC 8.0.342 software (Thermo Scientific, USA). 3. Results and discussion 3.1. Effect of dosage Absorbent dosage is a significant impact factor of adsorption process, determining the adsorbent–adsorbate equilibrium of the system (Deveci and Kar, 2013). The removal percentage of Cd2+ enhances as the dosage increases from 0% to 0.5%, and then maintains at approximately 100% until the dosage reaches to 1.0%.

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200

12

40

60

40

30

20

20

10

0 0.0

0.2

0.4

0.6

0.8

150

pH 8

100 6 2+

Ca

50 4

2+

removal capacity

10

0

2 0.0

1.0

0.2

0.4

(blank)

BC900 dosage (%)

Ca released concentration (mg/L)

50

p H a ft e r e q u i l i b ri u m (p H u i n t )

C d re m o v a l p e rc e n t a g e (% )

removal percentage

80

Cd removal capacity (mg/g)

60

100

0.6

0.8

1.0

BC900 dosage (%)

(A)

(B)

Fig. 1. Influence of BC900 dosage on Cd2+ removal. Equilibrium conditions: Cd2+ initial concentration of 200 mgL1, no Ca2+ stock solution added, 25.0 ± 1.0 °C, overnight. The mass of BC900 added in the 25 mL solution was 25, 50, 125, 200 and 250 mg for the dosage (S:L) of 0.1%, 0.2%, 0.5%, 0.8% and 1.0% respectively. (A) Cd removal percentage and removal capacity versus BC900 dosage; (B) pH level of equilibrium systems after equilibrium and Ca2+ released concentration versus BC900 dosage.

9 8

60 50 40 30 20

15 °C 25 °C 35 °C

10

pH after equilibrium (pH unit)

7

2+

removal capacity (mg-Cd /g-BC900)

70

6 5 4 3

blank, 25 °C BC900, 15 °C BC900, 25 °C BC900, 35 °C

2 1 0

0 1

2

3

4

5

6

1

2

3

4

5

6

initial pH (pH uint)

initial pH (pH unit)

(A)

(B)

3.5

Ca, 15 °C Mg, 15 °C Ca, 25 °C Mg, 25 °C Ca, 35 °C Mg, 35 °C

cation concentration (mmol/L)

3.0 2.5 2.0 1.5 1.0 0.5 0.0 1

2

3

4

5

6

initial pH (pH unit)

(C)

(D)

Fig. 2. Impact effect of equilibrium temperature and initial solution pH on cadmium removal by BC900. Equilibrium conditions: BC900 dosage 0.2%, Cd2+ initial concentration 200 mgL1, no Ca2+ stock solution added, temperature error control range ±1.0 °C, overnight. (A) removal capacity of cadmium by BC900 versus initial pH; (B) pH level of equilibrium systems after equilibrium versus initial pH; (C) net release concentration of cations versus initial pH; (D) three-dimension plot of pH after equilibrium, the sum concentration of net release Ca2+ and Mg2+ versus removal capacity of cadmium.

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However, the removal capacity of Cd2+ does not perform similar to the removal percentage, improving first and declining subsequently with the largest value of 42.80 ± 2.38 mgg1 at the dosage of 0.2% (Fig. 1(A)). To save the amount of adsorbent and make full use of the adsorption capability of BC900, the optimal dosage is 0.2%. At this dosage, multistage adsorption process can be used to reach a higher removal percentage, which is a better way to adsorb more cadmium and consume less biochar simultaneously. Bhattacharya et al. (2006), Chen et al. (2011), Iqbal et al. (2009), Kılıç et al. (2013) and Kołodyn´ska et al. (2012) also reported similar phenomena of absorbent dosage effect. Larger dosage can provide more active sites, thus the removal percentage will be higher; until after a certain adsorbent dosage, the removal efficiency does not increase sufficiently. However, not all adsorption sites are available for binding or exchanging because of overlapping and aggregation, so a portion of the adsorption capacity of adsorbent will not be used, and decreases as the dosage increases after a certain value. In this study, when the dosage is more than 0.2%, the buffer ability of BC900 makes the pH level of adsorbate solutions higher than 7, Cd2+ will start to transform into hydroxide precipitation of an

extremely low solubility, whereas the cation-exchange ability of released Ca2+ is wasted (Fig. 1(B)). 3.2. Effect of equilibrium temperature and initial pH Reaction temperature is a quite important factor, attracted much attention of researchers, which may influence reaction rate and process. All of the reaction situations, such as removal capacity of Cd2+, pH level of solutions after equilibrium and cation net release concentration, are analogous at different temperatures (15.0 °C, 25.0 °C and 35.0 °C), without remarkable differences (Fig. 2). The fact that equilibrium temperature does not significantly affect the adsorption process suggests that adsorption of Cd2+ by BC900 occurs mainly as chemical adsorption rather than physical adsorption (Vitela-Rodriguez and Rangel-Mendez, 2013). Adsorption experiments were conducted at six pH levels of 1, 2, 3, 4, 5 and 6 (Fig. 2), and the resulting data show that pH can influence the adsorptive performance of BC900 and cadmium system strongly. Generally, the removal efficiency increases with increasing the initial pH (Fig. 2(A)). For example, at the experimental

40

BC900 dried sludge washed BC900 BC900 ash

100

rem o v al p ercen tag e (% )

80

60

40

30

removal capacity (mg/g)

BC900 dried sludge washed BC900 BC900 ash

20

10

20

0

0 50

100

150

200

50

2+

100

initial Cd concentration (mg/L)

8

6

70 60 50 40 30

BC900 dried sludge washed BC900 BC900 ash

20

2+

4

80

Ca released concentration (mg/L)

pH (pH unit)

(B) BC900, initial BC900, blank BC900, after equilibrium dried sludge, initial dried sludge, blank dried sludge, after equilibrium washed BC900, initial washed BC900, blank washed BC900, after equilibrium BC900 ash, initial BC900 ash, blank BC900 ash, after equilibrium

10

200

initial Cd concentration (mg/L)

(A) 12

150

2+

10 0

2 0

50

100

150 2+

200

initial Cd concentration (mg/L)

(C)

250

0

50

100

150

200

2+

initial Cd concentration (mg/L)

(D)

Fig. 3. Adsorption performance of cadmium onto different adsorbents. Equilibrium conditions: adsorbent dosage 0.2%, no Ca2+ stock solution added, 25.0 ± 1.0 °C, overnight. (A) removal percentage of cadmium versus initial Cd2+ concentration; (B) removal capacity of cadmium versus initial Cd2+ concentration; (C) pH after equilibrium versus initial Cd2+ concentration, ‘‘blank’’ indicates only adsorbate solution without adsorbent; (D) Ca2+ released concentration versus initial Cd2+ concentration.

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Fig. 4. SEM photos of BC900 after equilibrium. Equilibrium conditions: adsorbent dosage 0.2%, no Ca2+ stock solution added, 25.0 ± 1.0 °C, overnight. (A) Equilibrium system DW, magnification 50 000; (B) equilibrium system DW, magnification 100 000; (C) equilibrium system 200 mgL1 Cd2+-bearing solution, magnification 50 000; (D) equilibrium system 200 mgL1 Cd2+-bearing solution, magnification 100 000.

temperature of 25.0 °C, the removal capacity is enhanced from 4.39 ± 0.26 mgg1 at the initial pH of approximately 1 pH unit to 68.87 ± 2.26 mgg1 at the initial pH of approximately 6 pH units with 14.69 times increase. When the initial pH is not higher than 2 pH units, the removal capacity is lower than 20 mgg1; comparatively, the removal capacity maintains at the level of more than 40 mgg1 when the initial pH is not lower than 3 pH units, with the considerable augmentation of more than 20 mgg1. Solution pH is considered the most important factor of metal adsorption onto adsorbent (Qin et al., 2011). The initial pH can affect the adsorption performance in the following ways: (1) electrostatic repulsion and affinity between adsorbent and adsorbate (Rao et al., 2002; Bhattacharya et al., 2006; Papandreou et al., 2011; Vitela-Rodriguez and Rangel-Mendez, 2013); (2) ion exchange process between adsorbent and adsorbate (Iqbal et al., 2009; Su et al., 2010); and (3) metal species distribution, such as soluble or insoluble and cation or anion (Zhang et al., 2008; Qin et al., 2011). In lower pH systems, BC900 with high pHPZC (10.17) charges positive on the surface, and the high electrostatic repelling forces inhibits the contact of Cd2+ and BC900. Meanwhile, in this work, the results also show that pH after equilibrium (Fig. 2(B)) and cation net release concentration (Fig. 2(C)) are functions of initial pH, and the removal capacity of cadmium is strongly related to both of the above dependent variables (Fig. 2(D)). When the initial pH is higher, the pH after equilibrium can improve to an adequate level by the pH buffer capacity of biochar, and the cadmium transforms to hydroxide precipitation to settle. However, precipitation will not form easily at pH lower than 8. Therefore, when the pH is lower, only limited local precipitation can form; although substantial Ca2+ and Mg2+ cations are leachated and released, huge amounts of H+ ion compete with these divalent metal ions, weakening the adsorption performance at lower initial pH. Meanwhile, this phenomenon is consistent with primary analysis of the adsorption mechanism in our previous work (Chen et al., 2014).

3.3. Analysis of adsorption mechanism To explore the Cd2+ removal mechanism, adsorption equilibrium experiments with different adsorbents were carried out, and the results are shown in Fig. 3. The adsorptive behavior of the four adsorbents can be classified into three groups: (1) washed BC900; (2) dried sludge; and (3) BC900 and BC900 ash. The adsorptive efficiency acts in the following order: washed BC900 < dried sludge < BC900 ash  BC900. It can also be found that when the removal efficiency becomes larger, the pH after equilibrium and the Ca2+ released concentration are higher, which is related to soluble alkaline substances and exchangeable cations of divalent alkaline earth metal (Ca2+). After DW washing, residual Ca element is seldom observed, however the vast majority of organic matters are remained (evidenced by FTIR spectra). Compared with BC900, the removal capacity of washed BC900 is very small, always lower than 15 mgg1, indicating that organic matters have an insignificant role, meanwhile minerals and exchangeable ions are indispensable in the adsorption process. By contrast, dried sludge consists of alkaline and exchangeable components, and thus has a certain degree of pH buffer and ion exchange capacity (Hu et al., 2012). Given that ash content is enriched by almost two times from 48.02 ± 0.46% (in dried sludge, dry basis) to 88.07 ± 0.56% through pyrolysis, the pH buffer and cation-exchange capacity improve observably (Fig. 2(C) and (D)), accordingly the cadmium removal efficiency is enhanced sharply. Due to the similarity of components, BC900 ash performs analogous behavior of adsorption, pH buffer and ion-exchange. Thus, it is found that (1) the contribution of organic matters on adsorption of cadmium is small; (2) alkaline earth metals such as Ca are very important for adsorptive reaction; and (3) precipitation formed by pH rising and ion exchange between Ca2+ and Cd2+ may be two major adsorption routes. Micromorphology of BC900 after adsorption in DW and Cd2+bearing solution was also examined, and the SEM photos are

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60 100 55

2+

X - r a y d i f f r a c t i o n i n t e n s i t y ( a .u .)

washed BC900

80 70

50 45 40

60

35

50 40

30

30

25 20

20 0

BC900, 50 ppm Cd

2+

BC900

2+

removal percentage, ~50 mg/L Cd 2+ removal percentage, ~200 mg/L Cd 2+ removal capacity, ~50 mg/L Cd 2+ removal capacity, ~200 mg/L Cd

Cd removal capacity (mg/g)

Cd removal percentage (%)

90

50

100

150

200

250

2+

initial Ca concentration (mg/L)

(A) 12

BC900, 100 ppm Cd

pH (pH unit)

10

BC900, 200 ppm Cd 0

10

20

30

40

50

60

70

80

8 2+

initial, ~50 mg/L Cd 2+ blank after equilibrium, ~50 mg/L Cd 2+ BC900 after equilibrium, ~50 mg/L Cd 2+ initial, ~200 mg/L Cd 2+ blank after equilibrium, ~200 mg/L Cd 2+ BC900 after equilibrium, ~200 mg/L Cd

6

90 4

2 (degree) Fig. 5. XRD patterns of BC900 after different treatments or equilibrium conditions. Equilibrium conditions: BC900 dosage 0.2%, no Ca2+ stock solution added, 25.0 ± 1.0 °C, overnight.

2 0

50

100

150

200

2+

initial Ca concentration (mg/L)

(B) ~1035

55 ~1420

~780

50

2+

~3435

net released concentration of Ca (mg/L)

BC900

Absorbance (a.u.)

washed BC900

BC900+DW, after equilibrium

45

40

2+

~50 mg/L Cd 2+ ~200 mg/L Cd

35

30

BC900+Cd,

0

~1385

50

100

150

200

2+

initial Ca concentration (mg/L)

after equilibrium

(C) 4000

3600

3200

2800 2000

1600

1200

800

400

-1

wavenumbers (cm ) Fig. 6. FTIR spectra of BC900 after different treatments or equilibrium conditions. ‘‘BC900, Cd’’ indicates BC900 after equilibrium in 200 mgL1 Cd2+-bearing solution. Equilibrium conditions: BC900 dosage 0.2%, no Ca2+ stock solution added, 25.0 ± 1.0 °C, overnight, drying at 60 °C.

shown in Fig. 4. Great microstructural changes are observed on the BC900 surface, which may be attributed to the adsorption of cadmium. After equilibrium in DW, the surface of BC900 is smooth

Fig. 7. Ca2+ competition effect to Cd2+ adsorption by BC900. Initial Cd2+ concentration 50 mgL1 or 200 mgL1; initial Ca2+ concentration 0 mgL1, 50 mgL1, 100 mgL1, and 200 mgL1. Equilibrium conditions: adsorbent dosage 0.2%, 25.0 ± 1.0 °C, overnight. (A) Cd2+ removal percentage and removal capacity versus initial Ca2+ concentration; (B) solution pH after equilibrium versus initial Ca2+ concentration; (C) Ca2+ net released concentration versus initial Ca2+ concentration.

with evenly distributed small cracks and micropores (Fig. 4(A) and (B)). By contrast, a large number of small particles are observed on the BC900 surface after equilibrium in Cd2+-bearing solution, which fill the pore structures and make the surface

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rough and uneven. These small particles are formed by cadmium adsorption, especially by precipitation of insoluble cadmium hydroxide inducing from BC900 buffer causing local pH increase. There are few differences between each XRD patterns of BC900 after different treatments or equilibrium conditions (Fig. 5), which indicates that the Cd2+ precipitation may be amorphous compounds. However, FTIR spectra (Fig. 6) show that (1) the peaks at 3435 cm1, 1420 cm1, 1035 cm1 and 780 cm1 are attributed to hydrogen bonded hydroxyl, CAH stretching vibration of CH2 and CH3, symmetric CAO stretching and CAH bending for aromatic ring respectively (Uchimiya et al., 2011; Zhang et al., 2011), while these organic surface functional groups are changing minimally, which suggests that the organic matters of BC900 remain and do not participate into the adsorption reaction in all the test conditions; and (2) a peak at 1385 cm1 exists in the spectrum of BC900 after equilibrium in 200 mgL1 Cd2+ solution, which is attributed to nitrate in inorganic salts (Miller and Wilkins, 1952), indicating that ion exchange between adsorbent and adsorbate occurs, and a portion of nitrate ions in Cd(NO3)2 are settled by flocculation because of network capturing of cadmium precipitation. Ca2+ cation competition was also measured to estimate the adsorption mechanism further, and the result curves are displayed in Fig. 7. When the initial Cd2+ concentration is 50 mgL1, the solution pH after equilibrium is raised to higher than 9 pH units by BC900 buffer (Fig. 7(B)), and it is alkaline enough to form precipitation of hydroxide. Therefore, the removal percentage of cadmium is uninfluenced and keeps close to the level of 100% (Fig. 7(A)). However, in the case of initial Cd2+ concentration of 200 mgL1, the solution pH after equilibrium ranges in a small territory from 6.5 to 8.5 (Fig. 7(B)), and the Ca2+ net released concentration remains 50.0 ± 5.0 mgL1 (Fig. 7(C)), but the removal efficiency decreases sharply with the increase in initial Ca2+ concentration (Fig 7.(A)), from 66.05 ± 0.88% (56.91 ± 0.64 mgg1) at the initial Ca2+ concentration of 0 mgL1 to 29.76 ± 2.06% (24.56 ± 1.48 mgg1) of the initial Ca2+ concentration of 200 mgL1, dropping by more than 50%. At this time, adsorption performance is contributed partially by local precipitation of solutions near the BC900 surface, and partially by cation exchange between Cd2+ in the solution and alkaline earth metallic cations in BC900 matrix. If cation exchange did not take place, concentrations of Ca2+ in the solution cannot influence cadmium adsorption efficiency. The excess Ca2+ competes with Cd2+ to occupy the active sites released by Ca2+ dissociation from the BC900 matrix. Thus, increased Ca2+ concentration in solution can weaken Cd2+ adsorption drastically. The adsorption mechanism can be described as the following equations: (1) pH raising

CaO þ H2 O ¼ CaOH2 Ca2þ þ 2OH

ð3Þ

(2) Precipitating 2þ

Cd

þ 2OH ¼ CdðOHÞ2 ðsÞ

ð4Þ

(3) cation exchanging 2þ

M-Ca2þ þ Cd

2þ

M-Cd

þ Ca2þ

ð5Þ

(4) competing and inhibiting

M-Ca2þ þ 2Hþ M-2Hþ þ Ca2þ 2þ

M-Cd

þ Ca2þ M-Ca2þ þ Cd

2þ

where M represents the biochar matrix.

ð6Þ ð7Þ

4. Conclusions In this work, static equilibrium experiments under different conditions have been conducted to evaluate the impact factors and to analyze the mechanism of cadmium adsorption on biochar derived from biophysical dried sewage sludge through pyrolysis at 900 °C. Adsorption percentage rises as the dosage of BC900 increases, whereas the dosage of 0.2% is enough, receiving the largest removal capacity of 42.80 ± 2.38 mgg1. pH is the most important factor affecting the adsorption process, with the removal capacity of lower than 20 mgg1 when the solution initial pH is lower than 2 pH units, and more than 40 mgg1 remaining when the solution initial pH is higher than 3 pH units. By contrast, temperature (ranging from 15 °C to 35 °C) is not dominant and weakly affects the adsorptive performance. The mechanism of cadmium adsorption onto biochar is mainly ascribed to (1) surface precipitation by forming insoluble cadmium compounds in alkaline condition induced by buffer capacity of biochar, and (2) ion exchange between cadmium ion in solution and exchangeable alkaline earth cations in the biochar, such as calcium. Acknowledgements This research was supported by the fund from Natural Science Foundation of China (No. 41371472) and the Major Science and Technology Program for Water Pollution Control and Treatment of the Ministry of Environmental Protection of China (No. 2011ZX07317-001). References Babel, S., Kurniawan, T.A., 2003. Low-cost adsorbents for heavy metals uptake from contaminated water: a review. J. Hazard. Mater. 97, 219–243. Belhalfaoui, B., Aziz, A., Elandaloussi, E.H., Ouali, M.S., De Ménorval, L.C., 2009. Succinate-bonded cellulose: a regenerable and powerful sorbent for cadmiumremoval from spiked high-hardness groundwater. J. Hazard. Mater. 169, 831– 837. Bhattacharya, A.K., Mandal, S.N., Das, S.K., 2006. Adsorption of Zn(II) from aqueous solution by using different adsorbents. Chem. Eng. J. 123, 43–51. Cao, X., Ma, L., Gao, B., Harris, W., 2009. Dairy-manure derived biochar effectively sorbs lead and atrazine. Environ. Sci. Technol. 43, 3285–3291. Chen, T., Zhang, Y., Wang, H., Lu, W., Zhou, Z., Zhang, Y., Ren, L., 2014. Influence of pyrolysis temperature on characteristics and heavy metal adsorptive performance of biochar derived from municipal sewage sludge. Bioresour. Technol. 164, 47–54. Chen, X., Chen, G., Chen, L., Chen, Y., Lehmann, J., McBride, M.B., Hay, A.G., 2011. Adsorption of copper and zinc by biochars produced from pyrolysis of hardwood and corn straw in aqueous solution. Bioresour. Technol. 102, 8877– 8884. Chun, Y., Sheng, G.Y., Chiou, C.T., Xing, B.S., 2004. Compositions and sorptive properties of crop residue-derived chars. Environ. Sci. Technol. 38, 4649–4655. Deveci, H., Kar, Y., 2013. Adsorption of hexavalent chromium from aqueous solutions by bio-chars obtained during biomass pyrolysis. J. Ind. Eng. Chem. 19, 190–196. Di Natale, F., Di Natale, M., Greco, R., Lancia, A., Laudante, C., Musmarra, D., 2008. Groundwater protection from cadmium contamination by permeable reactive barriers. J. Hazard. Mater. 160, 428–434. Febrianto, J., Kosasih, A.N., Sunarso, J., Ju, Y.-H., Indraswati, N., Ismadji, S., 2009. Equilibrium and kinetic studies in adsorption of heavy metals using biosorbent: a summary of recent studies. J. Hazard. Mater. 162, 616–645. Folgueras, M.B., Alonso, M., Díaz, R.M., 2013. Influence of sewage sludge treatment on pyrolysis and combustion of dry sludge. Energy 55, 426–435. Han, R., Liu, J., Zhang, Y., Fan, X., Lu, W., Wang, H., 2012. Dewatering and granulation of sewage sludge by biophysical drying and thermo-degradation performance of prepared sludge particles during succedent fast pyrolysis. Bioresour. Technol. 107, 429–436. Harvey, O.R., Herbert, B.E., Rhue, R.D., Kuo, L.-J., 2011. Metal Interactions at the biochar–water interface: energetics and structure–sorption relationships elucidated by flow adsorption microcalorimetry. Environ. Sci. Technol. 45, 5550–5556. Hu, J.-L., He, X.-W., Wang, C.-R., Li, J.-W., Zhang, C.-H., 2012. Cadmium adsorption characteristic of alkali modified sewage sludge. Bioresour. Technol. 121, 25–30. Iqbal, M., Saeed, A., Zafar, S.I., 2009. FTIR spectrophotometry, kinetics and adsorption isotherms modeling, ion exchange, and EDX analysis for understanding the mechanism of Cd2+ and Pb2+ removal by mango peel waste. J. Hazard. Mater. 164, 161–171.

T. Chen et al. / Chemosphere 134 (2015) 286–293 Kılıç, M., Kırbıyık, Ç., Çepeliog˘ullar, Ö., Pütün, A.E., 2013. Adsorption of heavy metal ions from aqueous solutions by bio-char, a by-product of pyrolysis. Appl. Surf. Sci. 283, 856–862. Kołodyn´ska, D., Wne˛trzak, R., Leahy, J.J., Hayes, M.H.B., Kwapin´ski, W., Hubicki, Z., 2012. Kinetic and adsorptive characterization of biochar in metal ions removal. Chem. Eng. J. 197, 295–305. Lu, H., Zhang, W., Yang, Y., Huang, X., Wang, S., Qiu, R., 2012. Relative distribution of Pb2+ sorption mechanisms by sludge-derived biochar. Water Res. 46, 854–862. Mašek, O., Budarin, V., Gronnow, M., Crombie, K., Brownsort, P., Fitzpatrick, E., Hurst, P., 2013. Microwave and slow pyrolysis biochar—comparison of physical and functional properties. J. Anal. Appl. Pyrol. 100, 41–48. Miller, F.A., Wilkins, C.H., 1952. Infrared spectra and characteristic frequencies of inorganic ions. Anal. Chem. 24, 1253–1294. Moreno-Barbosa, J., López-Velandia, C., Maldonado, A., Giraldo, L., Moreno-Piraján, J., 2013. Removal of lead(II) and zinc(II) ions from aqueous solutions by adsorption onto activated carbon synthesized from watermelon shell and walnut shell. Adsorption 19, 675–685. Papandreou, A.D., Stournaras, C.J., Panias, D., Paspaliaris, I., 2011. Adsorption of Pb(II), Zn(II) and Cr(III) on coal fly ash porous pellets. Miner. Eng. 24, 1495–1501. Peng, F., He, P.-W., Luo, Y., Lu, X., Liang, Y., Fu, J., 2012. Adsorption of phosphate by biomass char deriving from fast pyrolysis of biomass waste. Clean-Soil Air Water 40, 493–498. Qin, Q., Wang, Q., Fu, D., Ma, J., 2011. An efficient approach for Pb(II) and Cd(II) removal using manganese dioxide formed in situ. Chem. Eng. J. 172, 68–74. Rao, M., Parwate, A.V., Bhole, A.G., 2002. Removal of Cr6+ and Ni2+ from aqueous solution using bagasse and fly ash. Waste Manage. 22, 821–830. Shinogi, Y., Kanri, Y., 2003. Pyrolysis of plant, animal and human waste: physical and chemical characterization of the pyrolytic products. Bioresour. Technol. 90, 241–247. Stavropoulos, G.G., Zabaniotou, A.A., 2009. Minimizing activated carbons production cost. Fuel Process. Technol. 90, 952–957. Su, Q., Pan, B., Wan, S., Zhang, W., Lv, L., 2010. Use of hydrous manganese dioxide as a potential sorbent for selective removal of lead, cadmium, and zinc ions from water. J. Colloid Interface Sci. 349, 607–612.

293

Sud, D., Mahajan, G., Kaur, M.P., 2008. Agricultural waste material as potential adsorbent for sequestering heavy metal ions from aqueous solutions – a review. Bioresour. Technol. 99, 6017–6027. Taty-Costodes, V.C., Fauduet, H., Porte, C., Delacroix, A., 2003. Removal of Cd(II) and Pb(II) ions, from aqueous solutions, by adsorption onto sawdust of Pinus sylvestris. J. Hazard. Mater. 105, 121–142. Uchimiya, M., Lima, I.M., Klasson, K.T., Chang, S., Wartelle, L.H., Rodgers, J.E., 2010. Immobilization of heavy metal ions (Cu-II, Cd-II, Ni-II, and Pb-II) by broiler litter-derived biochars in water and soil. J. Agric. Food Chem. 58, 5538–5544. Uchimiya, M., Wartelle, L.H., Klasson, K.T., Fortier, C.A., Lima, I.M., 2011. Influence of pyrolysis temperature on biochar property and function as a heavy metal sorbent in soil. J. Agric. Food Chem. 59, 2501–2510. Vitela-Rodriguez, A.V., Rangel-Mendez, J.R., 2013. Arsenic removal by modified activated carbons with iron hydro(oxide) nanoparticles. J. Environ. Manage. 114, 225–231. Volesky, B., Holan, Z.R., 1995. Biosorption of heavy-metals. Biotechnol. Prog. 11, 235–250. Wang, B., Li, C., Liang, H., 2013. Bioleaching of heavy metal from woody biochar using Acidithiobacillus ferrooxidans and activation for adsorption. Bioresour. Technol. 146, 803–806. Wang, C., Li, J., Sun, X., Wang, L., Sun, X., 2009. Evaluation of zeolites synthesized from fly ash as potential adsorbents for wastewater containing heavy metals. J. Environ. Sci. 21, 127–136. Wang, J., Chen, C., 2009. Biosorbents for heavy metals removal and their future. Biotechnol. Adv. 27, 195–226. Yanik, J., Kornmayer, C., Saglam, M., Yüksel, M., 2007. Fast pyrolysis of agricultural wastes: characterization of pyrolysis products. Fuel Process. Technol. 88, 942– 947. Zhang, B., Xiong, S., Xiao, B., Yu, D., Jia, X., 2011. Mechanism of wet sewage sludge pyrolysis in a tubular furnace. Int. J. Hydrogen Energy 36, 355–363. Zhang, Y., Jiang, J., Chen, M., 2008. MINTEQ modeling for evaluating the leaching behavior of heavy metals in MSWI fly ash. J. Environ. Sci. (China) 20, 1398– 1402.