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New Evidence for High Sorption Capacity of Hydrochar for Hydrophobic Organic Pollutants Lanfang Han, Kyoung S. Ro, Ke Sun, Jie Jin, Judy A. Libra, and Baoshan Xing Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b02401 • Publication Date (Web): 22 Nov 2016 Downloaded from http://pubs.acs.org on November 22, 2016

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Environmental Science & Technology

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New Evidence for High Sorption Capacity of Hydrochar for

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Hydrophobic Organic Pollutants

3

Lanfang Han,

4

Libra, § and Baoshan Xingǁ

†,ǁ

Kyoung S. Ro, ‡ Ke Sun, †,* Haoran Sun, † Ziying Wang, † Judy A.

5 6

†

7

Beijing Normal University, Beijing 100875, China

8

‡

9

Service, U.S. Department of Agriculture, 2611 West Lucas Street, Florence, South

State Key Laboratory of Water Environment Simulation, School of Environment,

Coastal Plains Soil, Water, and Plant Research Center, Agricultural Research

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Carolina 29501, USA

11

§

12

Potsdam-Bornim, Germany

13

ǁ

14

USA

Leibniz Institute for Agricultural Engineering, Max-Eyth-Allee 100, 14469

Stockbridge School of Agriculture, University of Massachusetts, Amherst, MA 01003,

15 16 17 18 19 20

*Corresponding authors: (86)‐10‐58807493 (phone), (86)‐10‐58807493 (fax), email:

21

[email protected] (K. SUN).

22 1

ACS Paragon Plus Environment


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ABSTRACT: This study investigated the sorption potential of hydrochars, produced

24

from hydrothermally carbonizing livestock wastes, towards organic pollutants (OPs)

25

with a wide range of hydrophobicity, and compared their sorption capacity with that

26

of pyrochars obtained from conventional dry pyrolysis from the same feedstock.

27

Results of SEM, Raman and

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hydrochars mainly consisted of amorphous alkyl and aryl C. Hydrochars exhibited

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consistently higher logKoc of both nonpolar and polar OPs than pyrochars. This,

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combined with the significantly less energy required for the hydrothermal process,

31

suggests that hydrothermal conversion of surplus livestock waste into value-added

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sorbents could be an alternative manure management strategy. Moreover, the

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hydrochars logKoc values were practically unchanged after the removal of

34

amorphous aromatics, implying that amorphous aromatic C played a comparable role

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in the high sorption capacity of hydrochars compared to amorphous alkyl C. It was

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thus concluded that the dominant amorphous C associated with both alkyl and aryl

37

moieties within hydrochars explained their high sorption capacity for OPs. This

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research not only indicates that animal manure-derived hydrochars are promising

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sorbents for environmental applications, but casts new light on mechanisms

40

underlying the high sorption capacity of hydrochars for both nonpolar and polar

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OPs.

13

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C-NMR demonstrated that organic carbon (OC) of

42 43 44 2


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INTRODUCTION

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Biochar, an emerging carbon (C) material derived from renewable resources (e.g.,

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biomass, agricultural crop residues and livestock wastes),1 has triggered a new “gold

48

rush” based on its broad practical applications in a variety of areas, covering C

49

sequestration, potential soil amendments and environmental remediation.2, 3 Biochars

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made via dry pyrolysis (referred to as pyrochars) from woody feedstocks have been

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studied for decades as sorbents in environmental remediation.4, 5 There is a general

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consensus that the physicochemical heterogeneity of biochars is governed by

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feedstock sources along with pyrolysis process conditions.6 Few studies, however,

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have been made on biochars from less traditional feedstocks such as livestock wastes

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or those produced via wet pyrolysis or hydrothermal carbonization (HTC).7,

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Depending upon the process conditions used, the C recovery of HTC can be very

57

high, and the majority of the initial C stays bound in the final carbonaceous material

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(so-called hydrochars).8-10 In contrast to conventional dry pyrolysis, HTC can

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process wet biomass without a prior drying step. Although the solids may need to be

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separated and dried after the HTC process, the ease of dewatering is increased,

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which reduces energy requirements for the drying process. As a result, HTC requires

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significantly less energy than conventional dry pyrolysis.11 Additionally, due to the

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different processing conditions experienced by hydrochars and pyrochars, their

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chemical structures can differ significantly, which has been underscored in numerous

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studies, e.g., Titirici and Antonietti,8 Titirici et al.12 and Cao et al.13 Since chemical

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composition, physical conformation and polarity of natural organic sorbents in soils

8

3



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and sediments as well as man-made biochars have been shown to affect organic

68

compound sorption,14, 15 the diverse structures of hydrochars and pyrochars possibly

69

contribute to their different sorption abilities. Sun et al.16 reported that the

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hydrochars derived from poultry litter and swine solids exhibited similar or higher

71

sorption capacity toward bisphenol A, 17a-ethinyl estradiol and phenanthrene than

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the pyrochars produced from pyrolyzing poultry litter and wheat straw at 400 °C.

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These findings imply that hydrochar could be utilized as valuable sorbents. However,

74

the feedstocks for the hydrochars and pyrochars were different. Therefore, it is

75

difficult to draw any meaningful comparison. More importantly, the potential

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mechanism underlying the high sorption capacity of hydrochars was not addressed.

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The HTC process was first used to simulate coal-forming geological environments.

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Typical experimental HTC conditions (150-350 oC under saturated water vapor

79

pressure) are quite similar to those used to simulate the diagenesis of natural organic

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matter (NOM) in soils.13,

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hydrochar was analogous to naturally occurring kerogen, organic carbon (OC) that is

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not soluble in acid, base and organic solvents. Therefore, hydrochar from the HTC

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process, in essence, could play a role similar to NOMs in soils. The NOMs have

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been already identified as the key soil components in adsorbing a wide range of

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hydrophobic organic compounds (HOCs) in soils.14, 15, 19 In addition, the important

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role of alkyl domains within NOMs in the interaction with HOCs has been

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highlighted in extensive studies.19-22 In previous research,14 we selected humic acids,

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humins and nonhydrolyzable organic carbon (NHC, i.e., nonhydrolyzable residue

17

Wu et al.18 have reported that cellulose-derived

4


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after the treatment with trifluoroacetic acid (TFA) and HCl) as representatives of

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NOMs, and clearly demonstrated that it was the alkyl C rather than aromatic C that

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regulated the sorption of HOCs by NOMs. Based on this, we hypothesized that just

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like NOMs, hydrochars may have the potential to efficiently remove the HOCs in

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soils or wastewaters, and the alkyl moieties of hydrochar may be responsible for its

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sorption behavior.

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The major objectives of this study were therefore to: 1) evaluate the potential of

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hydrochars in removing selected organic contaminants with a wide range of logKow,

97

and compare the sorption capacity of hydrochars and pyrochars derived from the

98

same feedstock, and 2) explore the mechanisms accounting for their high sorption

99

capacity. To achieve these goals, the physico-chemical structurees of hydrochars and

100

pyrochars were characterized using X-ray photoelectron spectra (XPS), scanning

101

electron microscopy (SEM), Raman spectra and solid-state cross-polarization

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magic-angle-spinning 13C nuclear magnetic resonance spectra (13C CP-MAS NMR).

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Pyrene (PYE), and four pharmaceuticals and personal care products (PPCPs)

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(triclosan (TCS), estrone (E1), carbamazepine (CBZ) and acetaminophen (AMP))

105

were selected as sorbates to span a wide range of hydrophobicity.

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MATERIALS AND METHODS

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Sorbates and Sorbents. PYE (97+%), TCS (99+%), E1 (99+%), CBZ (99+%)

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and AMP (99+%) were used as sorbates and obtained from Sigma-Aldrich Chemical

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Co. Selected physicochemical properties of five chemicals are summarized in Table

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S1. The preparation methods of hydrochars and pyrochars from swine solids and 5



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poultry litter were described elsewhere.23 Briefly, hydrochars were prepared by

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hydrothermally carbonizing swine solids or poultry litter at 250 oC. Dried and

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ground (less than 2 mm) swine solids and poultry litter were mixed with distilled (DI)

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water to obtain a slurry of 20% (w/w) solids. This slurry was placed into a 1-L

115

non-stirred T316 stainless steel reactor with an external heater (Parr Instruments,

116

Moline, IL). The operating pressure of the system ranged from 3.4 to 9.0 MPa,

117

representing subcritical conditions. Afterwards, the reactor was cooled to room

118

temperature before the reaction products were filtered and dried at 105 °C.

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Subsequently, swine solid- and poultry litter-derived hydrochars were washed with

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DI water (denoted as HSS250 and HPL250, respectively). After drying at 105 °C,

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HSS250 and HPL250 were gently ground and homogenized to pass through a 0.25

122

mm sieve. For pyrochars, swine solids and poultry litter with a prior drying

123

treatment were carbonized at 250, 450, 600 °C for 4 h under N2 condition in a

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Lindburg electric box furnace equipped with a gas tight retort (Model 51662;

125

Lindburg/MPH, Riverside, MI). The temperature control of the retort system was

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based on input temperatures allowing for accurate control of pyrolytic exposure

127

temperature. The HTTs were raised to the desired values (250, 450 and 600 °C) at a

128

ramp rate of about 8 °C/min. The pyrochars were washed with DI water,

129

subsequently oven-dried at 105 °C. Then, these pyrochars were gently ground, and

130

homogenized to pass through a 0.25 mm sieve. The pyrochars are hereafter referred

131

to by their abbreviation based on their individual initial capitals of feedstock source

132

and HTTs (i.e., PSS250, PSS450, PSS600, PPL250, PPL450 and PPL600). 6


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The bleaching (BL) procedure, which was used to remove the non-condensed

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aromatic C (e.g., such as lignin-like and polyphenol units) of samples, was described

135

in detail elsewhere.24 Briefly, the bleaching involved treating 1 g of each sorbent

136

three times (7 h for each time) with 10 g of sodium chlorite (NaClO2), 10 mL of

137

acetic acid (CH3COOH) and 100 mL of DI water. All BL samples (i.e., HSS250-BL,

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HPL250-BL, PSS250-BL, PSS450-BL, PSS600-BL, PPL250-BL, PPL450-BL and

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PPL600-BL) were freeze-dried, ground, and stored for their characterization and

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sorption work.

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Sorbent Characterization. The bulk OC, H, N and O compositions of all samples

142

were determined by an Elementar Vario ELIII elemental analyzer via complete

143

combustion. Ash content of biochars was measured by heating samples at 750 °C for

144

4 h.25 X-ray photoelectron spectra (XPS), which can provide specific information on

145

the surface elemental composition (depth: 3-5 nm) of samples and analysis of large

146

sections (about 100 µm) of the particles, were recorded with a Kratos Axis Ultra

147

electron spectrometer using monochromated Al Ka source operated at 225 W.

148

Processing of the spectra was accomplished with the Xpspeak41 software. Four

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components were identified and quantified in the C1s spectra: 284.8 eV

150

(representing C-C and C-H bonds), 286.3 eV (C-O bonds), 287.5 eV (C=O bonds)

151

and 289.0 eV (COOH bonds). The particle structure and surface topography of

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hydrochars and pyrochars were investigated using SEM imaging analysis with a

153

Hitachi S4800 scanning microscope (Japan). 13C CP-MAS NMR spectra of biochars

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were obtained using a Bruker Avance 300 NMR spectrometer (Karlsruhe, Germany) 7



155

for structural composition. The NMR running parameters and chemical shift

156

assignments were depicted elsewhere.19 Raman measurements of samples were

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performed on a Labram aramis exciting with a 532 nm laser. Pore and surface

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characteristics were examined by gas (N2 and CO2) adsorption using an Autosorb-1

159

gas analyzer (Quantachrome Instrument Corp., Boynton Beach, FL). Surface area

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(SA) using CO2 isotherm at 273 K was calculated using nonlocal density functional

161

theory (NLDFT) (Figure S1). The SA using N2 was determined by the

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Brunaer-Emmett-Teller (BET) equation with multipoint adsorption isotherms of N2

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at 77 K.

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Sorption Experiments. All sorption isotherms were obtained using a batch

165

equilibration technique at 23 ± 1 °C. An appropriate amount of the investigated

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biochar samples was added to the background solution, which contained 0.01 M

167

CaCl2 in DI water with 200 mg/L NaN3 to minimize biodegradation. The amount of

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sorbents was chosen to result in 20-80% uptake of the initial solute amount. The

169

solutes were added to the vials to achieve initial concentrations from 0.001-0.12

170

mg/L for PYE, 0.1-10 mg/L for TCS, 0.2-20 mg/L for E1 and 0.2-100 mg/L for both

171

CBZ and AMP, in order to cover the range between detection limit and aqueous

172

solubility of solutes. Our preliminary test indicated that apparent equilibrium of PYE,

173

TCS, E1 and CBZ was reached within 7 d for hydrochars and pyrochars, while AMP

174

reached apparent equilibrium within 2 d. The blank experiments included solute

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without any sorbent, and sorbent without solute. Headspace was kept to a minimum

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to reduce solute vapor loss. After being shaken on the rotary shaker for 9 d of PYE, 8


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TCS, E1 and CBZ and for 3 d for AMP, all vials were placed upright for 24 h. Then,

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about 2 mL supernatant was taken out and analyzed directly by HPLC (1260 Series,

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Agilent Technologies, Santa Clara, CA) equipped with a reversed-phase C18 column

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(5 µm, 4.6 mm × 250 mm). Concentration of PYE was analyzed with a fluorescence

181

detector at 272 nm (excitation wavelength) and 419 nm (emission wavelength),

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while the investigated PPCPs were quantified on a diode array detector operated at a

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wavelength of 220 nm for TCS, 280 nm for E1, 280 nm for CBZ and 249 nm for

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AMP. The mobile phases for the sorbates were: PYE - methanol/water (90/10, v:v),

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TCS - acetonitrile/water (80/20, v:v), E1 and CBZ - acetonitrile/water (70/30, v:v)

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and AMP - acetonitrile/water (60/40, v:v). The flow rates for all sorbates were 0.8

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mL/min, and the column temperature was set at 35 °C. All samples coupled with

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blanks were conducted in duplicate. The analysis of blank experiments showed that

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the sorption of each sorbate by the vials and the decomposition of each sorbate were

190

insignificant, and the concentration of each sorbate potentially released from

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hydrochars and pyrochars was all below the detection limit.

192 193 194

Data Analysis. The sorption data were fitted to the logarithmic form of the Freundlich isotherm logqe = logKF + n logCe

(1)

195

where qe [µg/g] is the solid phase concentration; Ce [µg/L] is the solution phase

196

concentration; KF [(µg/g)/(µg/L)n] is the Freundlich affinity coefficient; and n is the

197

Freundlich exponential coefficient.

198

The sorption distribution coefficient (Kd) and the OC-normalized Kd (Koc) were 9



199

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obtained by the Eq. (2) and (3).

200

Kd = qe/Ce

(2)

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Koc=Kd/foc

(3)

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where foc is OC content. The logKd and logKoc values (Ce = 0.01, 0.1 and 1Sw, water

203

solubility of solutes) of hydrochars and pyrochars were calculated according the

204

above equations. The investigated correlations among properties of sorbents as well

205

as their sorption coefficients of each sorbate (Pearson correlation coefficients: p and

206

r values) were obtained from the Pearson correlation analysis by SPSS 16.0 software

207

(SPSS Inc., USA).

208

RESULTS AND DISCUSSION

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Characterization of hydrochars and pyrochars. The bulk elemental

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composition, ash content and SA determined by CO2 (CO2-SA) and N2 (N2-SA) of

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hydrochars and pyrochars are tabulated in Table S2. It has often been reported that

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the bulk C content of plant-derived pyrochars generally rises with HTTs.6,

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However, manure-derived pyrochars in our study, which had high ash contents, did

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not follow this pattern. The pyrochars obtained at 250 °C contained more C than

215

ones at 450 °C and 600 °C (Table S2). This was probably due to higher rates of loss

216

for C-containing compounds relative to the ash components during pyrolysis, since

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the ash content was approximately doubled at the higher HTTs (Table S2). The OC

218

contents of hydrochars were similar to (HPL250) or slightly higher (HSS250) than

219

their corresponding pyrochars at 250 °C (PSS250 and PPL250). However, the typical

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van Krevelen diagram showed that they differed in the relationships between the

26

10


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H/C versus O/C atomic ratios (Figure 1). Although the H/C ratios of two hydrochars

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were only slightly lower than corresponding pyrochars prepared at 250 °C, the O/C

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ratios were distinctly lower (Figure 1), indicating a higher degree of decarboxylation

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of hydrochars, which was further confirmed by the

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below. The typical transition from the high ratios of H/C and O/C of the primary

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plant macromolecules (e.g., cellulose) to the low ratios of more carbonized

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structures4,

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suggesting that both feedstocks underwent substantial carbonization during the

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pyrolysis carried out at 450 °C and 600 °C. In contrast to conventional dry pyrolysis,

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the experimental conditions required by HTC are close to the real geological

231

environment with moisture, and are similar to those of the method used to simulate

232

the diagenesis of NOMs in soils.17 From the van Krevelen diagram (Figure 1), it was

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seen that the elemental compositions of HSS250 and HPL250 were similar to those

234

of type II and type III kerogen, respectively, which accorded well with the report by

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Wu et al.18 that cellulose-derived hydrochar was analogous to naturally occurring

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kerogen. In addition, the aforementioned OC content and H/C ratios are important

237

parameters in the International Biochar Initiative (IBI) guidelines27 and European

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biochar guidelines28. Concerning their OC content, all hydrochars and pyrochars

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belonged to Class 2 biochar (30% ≤ OC content < 60%) according to the IBI

240

guidelines. The molar ratios of H/C of 450 and 600 oC pyrochars were also lower

241

than the maximum allowable threshold (0.7) required by IBI guidelines and

242

European biochar guidelines. However, the H/C of hydrochars and 250 oC pyrochars

6, 26

13

C NMR analyses discussed

was seen for the pyrochars at 450 °C and 600 °C (Figure 1),

11



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did not meet the requirements. In general, the ratio of H/C is related to aromaticity,

244

and has thus been commonly proposed as an indicator of the relative stability of

245

biochars in the soil system.29, 30 The biochars with the H/C higher than 0.7 may be

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considered to be labile or be easily decomposed. However, there is emerging

247

evidence that shows the H/C ratio or aromaticity level is not the unique factor

248

determining the stability of biochars, and other factors (e.g., minerals) can exert an

249

additional influence on its stability.31-33 Moreover, there have been studies showing

250

that the OC in hydrochar is more stable than we expect.34 For instance, Malghani et

251

al.34 reported that the roughly two thirds of hydrochar remained

252

after two

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to its behavior in soils as well as longer term field trials are needed.

in

the

soil

cropping seasons. Further studies on matching biochar characteristics

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Table S3 shows the surface elemental compositions of two types of biochars. The

255

surface O/C ratios of HSS250 (0.21) and HPL250 (0.49) were in the same range or

256

lower than that of their corresponding pyrochars (0.22-0.42 and 0.40-0.67,

257

respectively). Also, hydrochars generally contained less surface O-containing

258

functional groups (C-O, C=O and COOH) than the corresponding pyrochars,

259

implying the relatively higher surface hydrophobicity of hydrochars.

260

SEM images in Figure 2 neatly displays the dramatically different morphological

261

characteristics of hydrochars and pyrochars. The hydrochars particles (HSS250 and

262

HPL250) were relatively small in size and occurred as discrete spheres or as

263

aciniform agglomerates (Figure 2), indicative of the amorphous structure of

264

hydrochars.35 A similar morphology was not observed in pyrochars (Figure 2). 12


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Compared with hydrochars, the surfaces of pyrochars were more inerratic, and

266

became flatter with the increase of HTTs. Moreover, the Raman spectra in Figure 3a

267

and b illustrated that HSS250 and HPL250 mainly exhibited the amorphous C peak

268

(1357 cm-1), further demonstrating that the OC of hydrochars was primarily

269

amorphous. For pyrochars, the influence of HTTs upon their crystallinity was

270

observed (Figure 3). Specifically, only a small amorphous C peak was present for the

271

250 °C pyrochars, whereas the 450 and 600 oC pyrochars exhibited two typical peaks

272

namely, amorphous C (A) and graphite peak (G).36 The A peak of the 450 oC

273

pyrochars was comparable to or slightly stronger than their G peak, while the G peak

274

of the 600 oC pyrochars was obviously sharper than their A peak (Figure 3a and b).

275

The reason why only amorphous carbon peak was observed may be related to the

276

alteration of crystallinity of biochars with HTTs. During the pyrolytic process, the

277

crystallinity of precursor carbon materials in feedstocks lost around 250-300 °C.6, 37

278

For example, Keiluweit et al.6 found that as the temperature increased from 100 to

279

300 °C, the crystallinity of precursor materials gradually decreased until it

280

disappeared completely. With the loss of crystalline C, the amorphous C was

281

gradually formed. Largely because of this process, 250 oC pyrochars in this study

282

was mainly composed of amorphous C.

283

Besides the distinct morphology, 13C NMR spectra (Figure 4 and Table S4) further

284

showed that hydrochars and low and high temperature pyrochars were different in

285

functional groups. The NMR spectra of hydrochars, being mainly composed of alkyl

286

C (0-45 ppm) and aryl C (108-148 ppm), were largely similar to those of NHC in 13



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soils and sediments (Figure 4a) studied by Ran et al.19 and those of kerogen isolated

288

from the Borden aquifer,38 from the unweathered Woodford shale39 and from the

289

Altrium shales.40 Moreover, Ran et al.41 have previously proposed that

290

spectrum of NHC was similar to that of kerogen, and NHC was operationally similar

291

to the kerogen. These results well coincided with the above-mentioned similar

292

elemental compositions between hydrochars and type II and type III kerogen.

293

Furthermore, the broad peak of 0-45 ppm of hydrochars belonged to short chain

294

polymethylene (28 ppm) and methyl C in acetyl groups (21 ppm). These peaks all

295

belong to amorphous aliphatic C.21, 42, 43 No clear crystalline alkyl C peak (31.5 ppm)

296

was observed in hydrochars. This was in satisfactory accordance with the amorphous

297

OC structure of hydrochars suggested by SEM and Raman analyses. Additionally,

298

consistent with the different elemental compositions of 250 oC pyrochars and 450

299

and 600 oC ones, they also contained different OC-containing functional groups.

300

Specifically, pyrochars obtained at 250 °C (PSS250 and PPL250) had diverse

301

functional groups including carbohydrates besides alkyl and aryl C, while pyrochars

302

produced at high HTTs (450 and 600 °C) showed a large contribution of aromatic C

303

(108-165 ppm), and a negligible contribution of alkyl C (0-45 ppm) and other

304

structural groups (Figure 4b and c).16, 26 In view of the highly different compositions

305

and structures, the 250 °C pyrochars could be differentiated from those produced at

306

higher temperatures.

13

C NMR

307

Comparison of PYE and PPCPs sorption between hydrochars and pyrochars.

308

The sorption isotherms of PYE and four PPCPs (TCS, E1, CBZ and AMP) by 14


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hydrochars and pyrochars are shown in Figure S2 and the sorption data were well

310

fitted with Freundlich model as indicated by the high R2 values (all > 0.80) (Table

311

S5-S9). Apart from pyrochars produced at 250 °C, all sorption isotherms of PYE and

312

four PPCPs were highly nonlinear as evidenced by the nonlinear coefficient (n)

313

ranging from 0.22 to 0.83 (Table S5-S9). The sorption nonlinearity of hydrochars

314

was higher than pyrochars prepared at 250 °C but weaker than pyrochars at 450 °C

315

and 600 °C (Table S5-S9); for pyrochars, the values of n exhibited a downward trend

316

with increasing HTTs. Figure S3a clearly indicated that the differing nonlinearity

317

could be interpreted as the result of their different levels of aromaticity, with n and

318

aromaticity showing an inverse relationship. It has also been postulated that as more

319

molecules entered into micropores, the isotherms became more nonlinear.44 Using

320

CO2-SA as a measure of the microporosity, it could be seen that there is a

321

significantly negative relationship between microporosity and the nonlinearity index

322

n values of PYE and PPCPs (Figure S3b). This relationship, along with the positive

323

and significant correlation of CO2-SA of pyrochars to their aromaticity, suggests the

324

pore-filling resulted by the nanopores from aromatic C dominates the nonlinear

325

sorption of these selected sorbates by pyrochars.

326

The logKoc values of PYE, TCS, E1, CBZ and AMP for each biochar followed the

327

order at all test solute concentrations: PYE > TCS > E1 > CBZ > AMP,

328

corresponding well to the order of their Kow values (Table S1). This suggests the role

329

of hydrophobic effect (van der Waals force) in sorption of tested sorbates byeach

330

biochar.16, 45 Previous work showed that the hydrochars prepared from swine solids 15



Page 16 of 35

331

and poultry litter at 250 °C displayed higher sorption to polar compounds (bisphenol

332

A and 17a-ethinyl estradiol) than the pyrochars produced from wheat straw and

333

poultry litter at 400 °C.16 In the current study, it was found that when using the same

334

feedstock source, hydrochars produced at 250 °C in the presence of water also

335

exhibited higher sorption capacity to both nonpolar (PYE) and polar compounds

336

(TCS, E1, CBZ and AMP) than corresponding pyrochars obtained at low and high

337

HTTs (250, 450 and 600 °C, respectively) (Figure 5). For instance, logKoc values (Ce

338

= 0.01Sw) of CBZ were in the range of 4.19 to 4.31 for hydrochars compared to 2.34

339

to 3.21 for pyrochars (Table S8). Moreover, the logKoc of PYE for hydrochars

340

(6.08-6.19) in this investigation was also higher than those for plant-derived

341

pyrochars at low and high HTTs.46, 47 However, in contrast to the the high sorption

342

capacity of hydrochars for TCS in this study, Zhu et al.48 reported that the TCS

343

adsorption capacity onto the plant-derived hydrochars at similar temperature was low.

344

The different results may be related to the different kind of feedstock used for

345

hydrochar production. It has been largely reported that feedstock source is one of

346

key factors influencing the composition and structure of the final biochars, thereby

347

affecting their sorption capacities.49, 50 Qiu et al.26 observed that due to the much

348

higher surface OC content than bulk OC of animal manure-derived pyrochars, they

349

exhibited higher sorption of dibutyl phthalate and phenanthrene compared to the

350

plant-derived hydrochars. The animal manure-derived hydrochars in this work

351

showed the much lower O contents (8.95-12.51 %) than plant-derived one (27.2 %)

352

reported by Zhu et al.48, which may partly explain the higher sorption capacities of 16


Page 17 of 35


353

hydrochars in this study to TCS. Therefore, animal manure-derived hydrochars could

354

be an effective sorbents for environmental remediation. However, biochar itself may

355

be considered as one type of potential contaminants, metals/metalloids and

356

polycyclic aromatic hydrocarbons (PAHs)) within hydrochars and pyrochars should

357

be compared before using hydrochars instead of pyrochars for remediation purposes.

358

For

359

concentrations depended on feedstocks of biochars, while HTTs exerted very little

360

influence.51 In this respect, hydrochars and pyrochars from swine solids or poultry

361

litter in this study may have comparable metals/metalloids concentrations. For PAHs,

362

their content and composition in the biochars were found to strongly depend on

363

HTTs.51 However, a large amount of PAHs within biochars were observed to be

364

generally produced at relatively high temperature (> 300 oC).51,

365

suggests the use of hydrochars instead of pyrochars for remediation purposes as

366

sorbents would be promising.

metals/metalloids,

it

has

been

demonstrated

that

metals/metalloids

52

This further

367

Role of amorphous alkyl domain of hydrochars in their sorption. The high

368

sorption capacity of hydrochar may be tied to its distinct structure. Wang et al.53

369

proposed that geosorbents of higher polarity (characterized by O/C or O+N/C or

370

total polar functional groups) would have lower sorption for organic compounds,

371

since they have substantial hydrophilic moieties facing outside, which can facilitate

372

the formation of water cluster at their surfaces through H-bonding. Water clusters

373

would reduce the surface hydrophobicity of sorbents and the accessibility of organic

374

molecules to sorption domains, thereby decreasing their sorption capacity. This, 17



Page 18 of 35

375

along with the aforementioned lower contents of surface O-containing functional

376

groups of hydrochars (Table S3), indicates that the higher sorption capacity of

377

hydrochars may be partly explained by their comparatively lower surface

378

O-containing functional groups. However, the surface O-containing functional

379

groups of hydrochars were only slightly lower than the corresponding pyrochars

380

(Table S3), and thus were less likely to be the major factor accounting for their high

381

sorption capacity. As discussed above, HTC resulted in hydrochar with kerogen-like

382

structure. With respect to kerogen, the important role of their alkyl domains in the

383

interaction with HOCs has been highlighted before.19, 20 In our previous research,14 it

384

was clearly demonstrated that the alkyl C dominated the sorption of HOCs by NOMs

385

including kerogen. Hence, just like kerogen, alkyl C of hydrochars may also be the

386

main structural components accounting for the sorption capacity. This was supported

387

by the results for HSS250, which had a higher level of aliphaticity than HPL250 and

388

showed a higher logKoc for each chemical (Table S5-S9). The molecular flexibility of

389

amorphous alkyl domains has been proposed to endow them with better structural

390

compatibility with organic compounds.54 Moreover, an overview by Chefetz and

391

Xing

392

alkyl regions closely resemble that of alkane solvents and they can serve as an ideal

393

medium for the sorption of HOCs. In this sense, it could be assumed that the

394

amorphous alkyl C of hydrochars primarily accounted for their high sorption

395

capacity to PYE and PPCPs. However, hydrochars were demonstrated to be

396

composed of both alkyl and aryl C, the role of aryl C in the sorption of hydrochars

21

has also suggested that the relatively low density of the mobile amorphous

18


Page 19 of 35


397

should be further clarified. The bleaching technique, which has been previously

398

employed to selectively degrade non-condensed aromatic moieties in organic

399

geosorbents,14 was used to treat the hydrochars and pyrochars. It removed substantial

400

amounts of OC (Table S2). For the hydrochars, approximately one third of the

401

original OC was recovered, while the OC recovery of pyrochars ranged from 12.34%

402

up to 55.33%. The NMR results in Figure 4b and c clearly illustrated that bleaching

403

treatment successfully removed part of aryl C of the hydrochars, which was further

404

supported by the significantly positive relationship between the OC mass loss and

405

the decreased aromaticity of all samples by bleaching (Figure S4). Furthermore,

406

Raman spectra for bleached samples showed that bleaching treatment largely

407

decreased the amorphous C peak of hydrochars (Figure 3). These results revealed

408

that the amorphous aryl C was the main component which was degraded by

409

bleaching procedure. Table S4 presents the percentage of the removed aromatic C

410

(amorphous aromatic C) by bleaching to the total aromatic C of their untreated

411

counterparts. The contribution of amorphous aromatic C accounted for

412

78.40-79.06% to the total aromatic C of hydrochars. Moreover, we determined the

413

logKoc values (Ce = 0.01, 0.1 Sw) of bleached hydrochars to nonpolar (PYE) and

414

polar (E1) organic contaminants. It was seen that after the removal of amorphous

415

aryl C, the logKoc values of PYE and E1 by HSS250 and HPL250 were practically

416

unchanged (Table 1). For example, logKoc (Ce = 0.01 Sw) of PYE by HSS250 and

417

HPL250 changed from 6.19 and 6.08 to 6.27 and 6.12, respectively (Table 1). It was

418

clear that in addition to amorphous alkyl C, amorphous aryl C should also contribute 19



Page 20 of 35

419

to the sorption of hydrochar, otherwise the removal of amorphous aryl C and the

420

decrease of OC after bleaching treatment should result in the increase of logKoc if

421

alkyl C exclusively dominated the sorption; moreover, amorphous aromatic C should

422

play a comparable role in the high sorption capacity of hydrochars compared to

423

amorphous alkyl C.

424

Role of aromatic domains and porosity of pyrochars in their sorption. It has

425

been widely reported that pore-filling mechanism plays a key role in HOCs sorption

426

by microporous solids of geosorbents in soils.5,

427

correlation between logKoc values of each sorbate for pyrochars and their CO2-SA

428

obtained in our case (Figure S3c) further supports that pore-filling is a major

429

mechanism regulating sorption interactions of HOCs-pyrochars. In order to test if the

430

pore-filling mechanism exclusively controlled sorption of these five chemicals, their

431

logKoc values were normalized with the CO2-SA. The values of logKoc/CO2-SA for

432

the specific compounds by pyrochars at low and high HTTs still exhibited great

433

differences, indicating that apart from the microporosity of pyrochars, other factors

434

such as aromatic domains also affected sorption of these chemicals.

14

The significantly positive

435

As discussed above, pyrochars, especially ones at 450 and 600 oC, mainly

436

consisted of aromatics. It has been documented that aromatic π-systems in low-HTTs

437

pyrochars rich in electron-withdrawing functional groups will tend to be

438

electron-deficient and thus can act as π-acceptors, and the aromatic sheets of

439

high-HTTs pyrochars may theoretically act as both electron donors and acceptors

440

towards sorbates.1 Among the sorbate molecules investigated, PYE has been shown 20


Page 21 of 35


441

to be able to behave as π-donors.55 Also, -OH on the aromatic ring of AMP, E1 and

442

TCS as well as -NH2 in amide group on the aromatic ring of CBZ (Table S1) are all

443

electron-donating groups.56, 57 According to the review by Keiluweit and Kleber,58

444

PYE and four PPCPs (as electron donors) could be adsorbed to the pyrochars (as

445

π-accepters) via π-π electron donor-acceptor interactions or/and polar-π interactions.

446

These along with the strikingly positive relationship between the aromaticity of

447

pyrochars and the logKoc values of each compound (Figure S3d) indicate that the

448

sorption capacity of pyrochars for PYE and PPCPS may also result from the

449

contribution of π-π electron donor-acceptor interactions or/and polar-π interactions

450

between the sorbates and the aromatic components within pyrochars. In good

451

agreement with this conclusion, the removal of amorphous aromatic C of pyrochars

452

was noted to lead to the decrease in logKoc of PYE and TCS by pyrochars to varying

453

degrees (Table 1).

454

Environmental Implications. In this work, we investigated the sorption

455

capacities of hydrochars and pyrochars derived from the same manure source

456

towards PYE and four PPCPs. Hydrochars were noted to more effectively adsorb

457

both polar PPCPs and nonpolar PYE than pyrochars. Additionally, we observed that

458

HTC resulted in hydrochars with kerogen-like structure whose alkyl C has already

459

been demonstrated to be responsible for their sorption capacity. Further, it was

460

observed that the logKoc values of hydrochars were almost unchanged after the

461

removal of amorphous aryl C, implying that amorphous aromatic C played a

462

comparable role in the high sorption capacity of hydrochars compared to amorphous 21



Page 22 of 35

463

alkyl C. We thus concluded the dominant amorphous C associated with both alkyl

464

and aryl moieties within hydrochars explained their high sorption capacity for

465

organic pollutants. This research clearly provides evidence for mechanisms

466

underlying the high sorption capacity of hydrochars for both nonpolar and polar

467

organic pollutants. However, it should be noted that although the high sorption

468

capacity of animal manure-derived hydrochar was identified in this work, issues

469

about the long term environmental behavior of hydrochars, such as the release of

470

potential toxic substances from hydrochars into soils and the stability of OC within

471

hydrochars in soils, need to be more clearly addressed in the future before widely

472

applying hydrochars into the environment for remediation purposes.

473

ASSOCIATED CONTENT

474

Supporting Information. Four figures and nine tables are included in the

475

Supporting Information. This material is available free of charge via the Internet at

476

http://pubs.acs.org.

477

AUTHOR INFORMATION

478

Corresponding Author

479

*Phone: 86-10-58807493; fax: 86-10-58807493; e-mail: [email protected].

480

ACKNOWLEDGEMENTS

481

The authors would like to acknowledge the technical support provided by Mr.

482

Melvin Johnson of the USDA-ARS Coastal Plains Soil, Water & Plant Research

483

Center, Florence, SC. This research was supported by National Natural Science 22


Page 23 of 35


484

Foundation of China (41273106 and 41473087) and National Science Foundation for

485

Innovative Research Group (51421065), State Education Ministry, and USDA

486

McIntire-Stennis Program (MAS 00028). L. Han also thanks the China Scholarship

487

Council to support her study at the University of Massachusetts, Amherst. Mention

488

of trade names or commercial products is solely for the purpose of providing specific

489

information and does not imply recommendation or endorsement by the U.S.

490

Department of Agriculture.

491

492 493 494 495 496 497 498 499 500 501 502 503 504 505 506 507 508 509 510 511 512 513 514 515 516 517 518 519

1. Jin, J.; Sun, K.; Wu, F.; Gao, B.; Wang, Z.; Kang, M.; Bai, Y.; Zhao, Y.; Liu, X.; Xing, B. Single-solute and bi-solute sorption of phenanthrene and dibutyl phthalate by plant-and manure-derived biochars. Sci. Total. Environ. 2014, 473, 308-316. 2. Lattao, C.; Cao, X.; Mao, J.; Schmidt-Rohr, K.; Pignatello, J. J. Influence of molecular structure and adsorbent properties on sorption of organic compounds to a temperature series of wood chars. Environ. Sci. Technol. 2014, 48, 4790-4798. 3. Lehmann, J.; Gaunt, J.; Rondon, M. Biochar sequestration in terrestrial ecosystems–a review. Mitig. Adapt. Strat. Gl. 2006, 11 (2), 395-419. 4. Chen, B.; Zhou, D.; Zhu, L. Transitional adsorption and partition of nonpolar and polar aromatic contaminants by biochars of pine needles with different pyrolytic temperatures. Environ. Sci. Technol. 2008, 42 (14), 5137-5143. 5. Chun, Y.; Sheng, G.; Chiou, C. T.; Xing, B. Compositions and sorptive properties of crop residue-derived chars. Environ. Sci. Technol. 2004, 38 (17), 4649-4655. 6. Keiluweit, M.; Nico, P. S.; Johnson, M. G.; Kleber, M. Dynamic molecular structure of plant biomass-derived black carbon (biochar). Environ. Sci. Technol. 2010, 44 (4), 1247-1253. 7. Cao, X.; Ro, K. S.; Libra, J. A.; Kammann, C. I.; Lima, I.; Berge, N.; Li, L.; Li, Y.; Chen, N.; Yang, J. Effects of biomass types and carbonization conditions on the chemical characteristics of hydrochars. J. Agric. Food. Chem. 2013, 61 (39), 9401-9411. 8. Titirici, M.-M.; Antonietti, M. Chemistry and materials options of sustainable carbon materials made by hydrothermal carbonization. Chem. Soc. Rev. 2010, 39 (1), 103-116. 9. Titirici, M. M.; Thomas, A.; Yu, S.-H.; Müller, J.-O.; Antonietti, M. A direct synthesis of mesoporous carbons with bicontinuous pore morphology from crude plant material by hydrothermal carbonization. Chem. Mater. 2007, 19 (17), 4205-4212.

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652 653 654 655 656 657 658 659 660 661 662 663 664 665 666 667 668 669 670 671

Solvent-extractable polycyclic aromatic hydrocarbons in biochar: influence of pyrolysis temperature and feedstock. Environ. Sci. Technol. 2012, 46 (17), 9333-9341. 53. Wang, X.; Guo, X.; Yang, Y.; Tao, S.; Xing, B. Sorption mechanisms of phenanthrene, lindane, and atrazine with various humic acid fractions from a single soil sample. Environ. Sci. Technol. 2011, 45 (6), 2124-2130. 54. Zhang, D.; Pan, B.; Cook, R. L.; Xing, B. Multi-walled carbon nanotube dispersion by the adsorbed humic acids with different chemical structures. Environ. Pollut. 2015, 196, 292-299. 55. Chefetz, B.; Deshmukh, A. P.; Hatcher, P. G.; Guthrie, E. A. Pyrene sorption by natural organic matter. Environ. Sci. Technol. 2000, 34 (14), 2925-2930. 56. Oleszczuk, P.; Pan, B.; Xing, B. Adsorption and desorption of oxytetracycline and carbamazepine by multiwalled carbon nanotubes. Environ. Sci. Technol. 2009, 43 (24), 9167-9173. 57. Lin, D.; Xing, B. Adsorption of phenolic compounds by carbon nanotubes: role of aromaticity and substitution of hydroxyl groups. Environ. Sci. Technol. 2008, 42 (19), 7254-7259. 58. Keiluweit, M.; Kleber, M. Molecular-level interactions in soils and sediments: the role of aromatic π-systems. Environ. Sci. Technol. 2009, 43 (10), 3421-3429.

672 673 674

27



Page 28 of 35

675 676 677 678 679 680 681

Figure Captions:

682

Figure 2. Scanning electron microscope (SEM) of swine solid- and poultry

683

litter-derived hydrochars and pyrochars.

Figure 1. Van Krevelen plot of elemental ratios for swine solid- and poultry litter-derived hydrochars and pyrochars.

684 685

Figure 3. Comparison of Raman spectra of original and bleached swine solid- (a)

686

and poultry litter-derived (b) hydrochars and pyrochars.

687 688

Figure 4. Comparison of solid-state 13C CP-MAS NMR spectra of nonhydrolyzable

689

carbon (NHC) (a), and original and bleached swine solid- (b) and poultry

690

litter-derived (c) hydrochars and pyrochars.

691 692

Figure 5. Comparison of logKoc between hydrochars and pyrochars for five sorbates

693

including pyrene (PYE), triclosan (TCS), estrone (E1), carbamazepine (CBZ) and

694

acetaminophen (AMP).

695 696

Table Caption:

697

Table 1. The logKoc values (Ce =0.01 and 0.1 Sw) of pyrene (PYE) and estrone (E1)

698

original and bleached hydrochars and pyrochars.

699

28


Page 29 of 35


720 721 722 723 724

4.0

demethylation dehydration decarboxylation

3.0 H/C atomic ratio

700 701 702 703 704 705 706 707 708 709 710 711 712 713 714 715 716 717 718 719

2.0

HPL250 PPL250 PPL450 PPL600

HSS250 PSS250 PSS450 PSS600

Ⅰ Ⅱ

Ⅲ

1.0

0.0 0.0

0.2

0.4 0.6 O/C atomic ratio

0.8

1.0

Figure 1. Van Krevelen plot of elemental ratios for swine solid- and poultry litter-derived hydrochars and pyrochars. Note that I, II and III are three evolution paths of kerogen; H and P represent hydrochars and pyrochars, respectively; SS and PL refer to the biochars obtained from swine solids and poultry litter, respectively; 250, 450 and 600 are the heating treatment temperature (oC).

29



725 726 727 728 729 730 731 732 733 734 735 736 737 738 739 740 741 742 743 744 745 746 747 748 749 750 751 752

a. HSS250

b. HPL250

c. PSS250

d. PPL250

e. PSS450

f. PPL450

g. PSS600

h. PPL600

Page 30 of 35

753 754 755 756 757 758 759 760 761 762 763 764 765 766 767

Figure 2. Scanning electron microscope (SEM) images of swine solid- and poultry litter-derived hydrochars and pyrochars. Note that H and P represent hydrochars and pyrochars, respectively; SS and PL refer to the biochars obtained from swine solids and poultry litter, respectively; 250, 450 and 600 are the heating treatment temperature (oC).

30


Page 31 of 35


a

A

PSS600-BL

Intensity (a.u.)

768 769 770 771 772 773 774 775 776 777 778 779 780 781 782 783 784 785 786 787 788 789 790 791 792 793 794 795 796 797 798 799 800 801 802 803 804 805 806 807 808 809 810 811

G

b PPL600-BL

PSS600

PPL600

PSS450-BL

PPL450-BL

PSS450

PPL450

PSS250-BL

PPL250-BL

PSS250

PPL250

HSS250-BL

HPL250-BL

HSS250

HPL250

1000 1200 1400 1600 1800 2000 -1

Raman Shift (cm )

A

G

1000 1200 1400 1600 1800 2000

Raman Shift ( cm -1)

Figure 3. Comparison of Raman spectra of original and bleached swine solid- (a) and poultry litter-derived (b) hydrochars and pyrochars. Note that A and G refer to amorphous and graphite peak, respectively; H and P represent hydrochars and pyrochars, respectively; SS and PL refer to the biochars obtained from swine solids and poultry litter, respectively; 250, 450 and 600 are the heating treatment temperature (oC); “BL” denotes that the bleached samples.

31



812 813 814 815 816 817 818 819 820 821 822 823 824 825 826 827 828 829 830 831 832 833 834 835 836 837 838 839 840 841 842 843 844 845 846 847 848 849 850 851 852 853 854 855

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a NHC5 NHC4 NHC3 NHC2 NHC1 HPL250 HSS250

250 200 150 100 50 13

-50

C Chemical Shift (ppm)

Aromatic C

b

0

28

Aromatic C

c

28

21

0 -50 250 200 150 100 50

0

21

PSS600-BL

PPL600-BL

PSS600

PPL600

PSS450-BL

PPL450-BL

PSS450

PPL450

PSS250-BL

PPL250-BL

PSS250

PPL250

HSS250-BL

HPL250-BL

HSS250

HPL250

250 200 150 100 50 13


-50

13


Figure 4. Comparison of solid-state 13C CP-MAS NMR spectra of nonhydrolyzable carbon (NHC) (a), and original and bleached swine solid- (b) and poultry litter-derived (c) hydrochars and pyrochars. Note that NHC1 and NHC2 were NHC extracted from surface soils, NHC3, NHC4 and NHC5 were NHC extracted from sediments (data of NHC are cited from the report by Ran et al.19); H and P represent hydrochars and pyrochars, respectively; SS and PL refer to the biochars obtained from swine solids and poultry litter, respectively; 250, 450 and 600 are the heating treatment temperature (oC); “BL” denotes that the bleached samples. 32


Page 33 of 35


8.0 7.0

LogKoc, mL/g

PYE

TCS

AMP

CBZ

E1

6.0 5.0 4.0 3.0 2.0 1.0

PPL600

PPL450

PPL250

PSS600

PSS450

PSS250

HPL250

0.0

HSS250

856 857 858 859 860 861 862 863 864 865 866 867 868 869 870 871 872 873 874 875 876 877 878 879 880 881 882 883 884 885 886 887 888 889 890 891 892 893 894 895 896 897 898 899

Figure 5. Comparison of logKoc between hydrochars and pyrochars for five sorbates including pyrene (PYE), triclosan (TCS), estrone (E1), carbamazepine (CBZ) and acetaminophen (AMP). Note that H and P represent hydrochars and pyrochars, respectively; SS and PL refer to the biochars obtained from swine solids and poultry litter, respectively; 250, 450 and 600 are the heating treatment temperature (oC).

33



Page 34 of 35

900

Table 1. The logKoc values (Ce =0.01 and 0.1 Sw) of pyrene (PYE) and estrone (E1) by original and bleached hydrochars and pyrochars logKoc (mL/g) Samples

Pyrene (PYE)

Estrone (E1)

Ce = 0.01Sw Ce = 0.1Sw Ce = 0.01Sw Ce = 0.1Sw HSS250 6.19 5.97 4.78 4.58 PSS250 5.56 5.56 3.19 3.05 PSS450 5.85 5.36 3.63 3.16 PSS600 5.71 5.08 4.24 3.46 HPL250 6.08 5.82 4.53 4.29 PPL250 5.33 5.19 3.45 3.20 PPL450 5.70 5.22 3.76 3.11 PPL600 5.78 5.23 4.15 3.37 HSS250-BL 6.27 6.05 4.85 4.64 PSS250-BL 5.15 4.38 2.75 2.50 PSS450-BL 5.48 4.86 3.23 2.94 PSS600-BL 5.61 5.01 4.18 3.37 HPL250-BL 6.12 5.88 4.60 4.35 PPL250-BL 5.19 4.96 3.11 2.87 PPL450-BL 5.57 5.15 3.71 3.03 PPL600-BL 5.71 5.15 4.13 3.33 Koc is the organic carbon (OC) normalized sorption distributed coefficient; Sw: aqueous solubility of pyrene (PYE) and estrone (E1). Note that H and P represent hydrochars and pyrochars, respectively; SS and PL refer to the biochars obtained from swine solids and poultry litter, respectively; 250, 450 and 600 are the heating treatment temperature; BL: bleached samples (oC).

34


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901 902 903 904 905 906 907 908 909 910

Graphic for manuscript (Abstract Art)

911 912 913

35


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