Rice waste biochars produced at different pyrolysis temperatures for arsenic and cadmium abatement and detoxification in sediment

Journal Pre-proof Rice waste biochars produced at different pyrolysis temperatures for arsenic and cadmium abatement and detoxification in sediment Wei Zhang, Xiaofei Tan, Yanling Gu, Shaobo Liu, Yunguo Liu, Xinjiang Hu, Jiang Li, Yahui Zhou, Sijia Liu, Yuan He PII:

S0045-6535(20)30461-6

DOI:

https://doi.org/10.1016/j.chemosphere.2020.126268

Reference:

CHEM 126268

To appear in:

ECSN

Received Date: 28 October 2019 Revised Date:

17 February 2020

Accepted Date: 18 February 2020

Please cite this article as: Zhang, W., Tan, X., Gu, Y., Liu, S., Liu, Y., Hu, X., Li, J., Zhou, Y., Liu, S., He, Y., Rice waste biochars produced at different pyrolysis temperatures for arsenic and cadmium abatement and detoxification in sediment, Chemosphere (2020), doi: https://doi.org/10.1016/ j.chemosphere.2020.126268. This is a PDF file of an article that has undergone enhancements after acceptance, such as the addition of a cover page and metadata, and formatting for readability, but it is not yet the definitive version of record. This version will undergo additional copyediting, typesetting and review before it is published in its final form, but we are providing this version to give early visibility of the article. Please note that, during the production process, errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain. © 2020 Published by Elsevier Ltd.

Credit Author Statement Wei Zhang: Conceptualization, Methodology, Software. Xiaofei Tan: Supervision, Writing- Reviewing and Editing. Yanling Gu: Data curation, Writing- Original draft preparation. Shaobo Liu: Investigation, Methodology, Software. Yunguo Liu: Supervision. Xinjiang Hu: Conceptualization. Jiang Li: Writing- Reviewing and Editing. Yahui Zhou: Visualization, Investigation. Sijia Liu: Visualization, Investigation. Yuan He: Writing- Reviewing and Editing.

GRAPHICAL ABSTRACT

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Rice waste biochars produced at different pyrolysis temperatures for

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arsenic and cadmium abatement and detoxification in sediment

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Wei Zhanga,b, Xiaofei Tana,b,∗, Yanling Guc, Shaobo Liu d,∗, Yunguo Liua,b, Xinjiang Hue,

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Jiang Li d, Yahui Zhoua,b, Sijia Liua,b, Yuan Hef

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a

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410082, P. R. China

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b

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Ministry of Education, Changsha 410082, P. R. China

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c

College of Environmental Science and Engineering, Hunan University, Changsha

Key Laboratory of Environmental Biology and Pollution Control (Hunan University),

College of Materials Science and Engineering, Changsha University of Science and

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Technology, Changsha 410114, PR China

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d

College of Architecture and Art, Central South University, Changsha 410083, PR China

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e

College of Environmental Science and Engineering, Central South University of

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Forestry and Technology, Changsha, P. R. China

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f

Center of Changsha Public Engineering Construction, Changsha 410013, China

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∗ ∗

Corresponding author: Tel.: +86 731 88822829; E-mail: [email protected] Corresponding author: Tel.: +86 731 88649208; E-mail: [email protected] 1

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Abstract:

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The effectiveness of rice waste biochars on heavy metal and metalloid abatement

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and detoxification was investigated using comprehensive studies based on As and Cd

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immobilization, bioaccumulation in tubifex, and microbial community changes in

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contaminated sediment. The remediation effects of biochars produced at different

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pyrolytic temperatures (400-700 °C) were evaluated. Bioaccumulation of heavy metal

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and metalloid in the tubifex tissue and change of indigenous microbial community under

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treatment of different biochars were assessed. Biochars produced at 700 °C exhibited

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greater effect on decreasing the concentrations of As and Cd in aqueous phase, and TCLP

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extractable and bioavailable metal(loid) in solid phase of sediment. The concentration of

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As and Cd in water phase decreased by 26%-89% and 22%-71% under the treatment of

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straw biochar, and decreased by 13%-92% and 5%-64% under the treatment of rice husk

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biochar, respectively. As and Cd contents in the tubifex tissue were positively correlated

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with their concentrations in aqueous phase. High-temperature biochars significantly

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reduced metal(loid) bioaccumulation in tubifex. The richness and biodiversity of

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microbial community were both greater in all biochars remediated sediment compared to

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non-treated sediment. These results indicated that rice waste biochars could effectively

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inhibit the bio-availability and toxicity of heavy metal and metalloid in sediment, and the

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higher-temperature biochar exhibited better performance.

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Keywords: rice waste; ecotoxicology; remediation; carbonaceous material; metal(loid)

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1. Introduction

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Heavy metals have characteristics of acute toxicity, bioaccumulation, persistence, and

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non-biodegradability (Tan et al., 2015a; Cai et al., 2019), which pose a significant threat

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to human health and the environment (Montgomery and Elimelech, 2007; Wang et al.,

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2018). More than 10 million main polluted sites were existed worldwide, and over 50% of

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them were polluted by heavy metal(loid)s (Khalid et al., 2017; Gong et al., 2018). Aquatic

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sediments are recognized as the ultimate repositories of continuous discharges of heavy

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metals. These heavy metals in contaminated sediments are the long-term sources of

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exposure to the water environment (Ghosh et al., 2011). Remediation of heavy metals

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contaminated sediments remains a technological challenge.

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Studies found that carbonaceous particles in the sediment have a vital role in reducing

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the concentration of pollutants and bioavailability in the aqueous phase of the sediment

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(Werner et al., 2005; Gilmour et al., 2013; Pal and Maiti, 2019). Millward et al. (2005)

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proposed a new in-situ remediation technology by adding black carbon to the bio-active

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layer of the sediment to adsorb and immobilize contaminants, thereby promoting the

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natural process of pollutants stabilization. The porous structure of carbonaceous materials

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endows their high affinity and adsorption capacity for pollutants, and thus have become

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potential materials for the remediation of sediments (Chen et al., 2016; Thompson et al.,

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2016; Abel et al., 2017). Activated carbon and various nanomaterials such as zero-valent

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iron (nZVI), carbon nanotubes (CNTs), and titanium dioxide nanoparticles (TiO2 NPs)

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have been used for the remediation of metal(loid) contaminated sediments (Libralato et

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al., 2018; Cai et al., 2019). These materials usually have some drawbacks limiting their

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practical application, due to high cost, hard to operation, and even induce secondary

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pollution to the natural environment. Biochar has exhibited to be a promising carbon

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material used in remediation technology due to its wide availability of feedstocks,

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low-cost and favorable physical/chemical surface characteristics (Tan et al., 2015b).

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Recent studies have found that biochar has certain adsorption ability for various

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pollutants in water and soil due to its available pore structure, appropriate surface

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properties, and high pH and CEC, which can be applied as a feasible in-situ remediation

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material for sediments (Ghosh et al., 2011; Tan et al., 2016; Tan et al., 2017).

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The pyrolytic temperature of biochar has been reported to have large effect on

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properties of biochar and its adsorption ability for heavy metals from water. Kim et al.

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(2013) suggested that the pH and specific surface area of biochar increased significantly

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as the pyrolysis temperature increased, and biochar prepared at higher pyrolytic

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temperature exhibited higher Cd adsorption capacity. Niazi et al., (2018) reported that

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high-temperature perilla leaf biochar (700°C) removed more arsenite at pH 7–9 than

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low-temperature biochar (300°C). Shen et al. (2012) pointed out that higher pyrolytic

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temperature contributed to the lower Cr(VI) removal ability of biochar owing to the

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decrease of surface functional groups. Similarly, pyrolytic temperature may have

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influence on the remediation efficiency of biochar in sediment. Chi and Liu (2016)

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prepared wheat straw biochar at temperature of 400 °C and 700 °C for phenanthrene and

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pyrene treatment in sediment, and the results suggested that high-temperature biochar

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possessed more pore structure and aromaticity, making it manifested stronger

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immobilization performance in sediment. However, relatively little is known concerning

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the pyrolytic temperature-mediated remediation effect of biochars in sediment

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contaminated by different kinds of heavy metals. Therefore, in this study, the

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immobilization effectiveness of As and Cd in sediment using biochars produced at

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different pyrolytic temperatures were tested. Equilibrium concentration in aqueous phase,

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leachability of As and Cd in the sediment, and chemical fractions of two metal(loid)s in

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sediment were determined by standard methods.

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In addition, decrease of the bioavailability of heavy metals is the main goal of a

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successful immobilization remediation technique. Organisms and microbial communities

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in sediment/water system can be the biological indicator for biological availability and

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toxicity of heavy metals, which are of great importance because they represent essential

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part of the aquatic food chain. Tubifex worm is one of the most ubiquitous organisms in

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freshwater ecosystems (Liu et al., 2014), which is usually exposed to environmental

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pollutants via solid and aqueous phase of sediment through ingestion and intimate contact.

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Because of its wide distribution in sediment, tubifex is widely used in assessing the

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biological availability and toxicity of pollutants in sediment (Steen Redeker et al., 2004;

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Liu et al., 2014). Therefore, bioaccumulation of As and Cd in the tubifex tissue and

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change of indigenous microbial community in sediment under treatment of different

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biochars were also assessed in present study.

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2. Materials and methods

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2.1. Sediment collection

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Sediment used in present study was got from the Xiangjiang River (top 20 cm). The

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debris in the collected sediment were eliminated and the resulting fine sediment was well

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mixed to form a uniform sediment mixture. The samples were air-dried, and sieved

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through 60 mesh, then stored in sealed sample bag for later use. Basic properties of

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sediment are shown in Supporting Information Table S1.

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2.2. Preparation of biochars

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Two kinds of feedstocks (rice husk and rice straw) of biochar were collected from

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the outskirts farm of Yiyang city, Hunan province, China. The above biomass was

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air-dried under natural conditions and then crushed by a pulverizer and sieved through 2

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mm mesh. Biomass was pyrolyzed into biochar using a tube furnace. The furnace was

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filled with nitrogen to maintain the anoxic conditions during the pyrolysis process. A

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heating program was set up to heat the furnace to the target temperature (400, 500, 600,

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and 700 °C) at a heating rate of 7 °C min–1 for 1 h. The biochars prepared from rice straw

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and rice husk at different temperatures were labeled as RS400, RS500, RS600, RS700,

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RH400, RH500, RH600, and RH700, respectively.

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2.3. Characterization methods

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The element content of biochars was tested using elemental analyzer (Vario EL Cube,

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Elementar, Germany) under the pattern of CHN+O. The Fourier transform infrared (FTIR)

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spectra was measured on a FTIR spectrometer (NICOLET 5700, USA) at the range of

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400 cm–1–4000 cm–1. The BET surface area of the biochar was measured by Automated

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Gas Sorption Anlyzer (Autosorb-IQ, Quantachrome, USA). The morphological structure

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of biochar was analyzed by scanning electron microscopy (SEM) (TM3000, Hitachi,

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Japan). X-ray diffraction (XRD) spectra was analyzed using a Bruker D8-Advance X-ray

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diffractometer (Bruker, German). X-ray Photoelectron Spectrometer (XPS) analysis was

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conducted using ESCALAB 250Xi (Thermo Fisher, USA). Raman spectra was tested on a

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Raman Microscope (LabRAM HR, HORIBA, France) with a laser excitation wavelength

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of 488 nm. Raman spectroscopy was measured using Raman microscope (LabRAM HR,

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HORIBA, France).

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2.4. Sediment incubation

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A set of 50 g (dry weight) of sediment was added into 100 mL screw mouth bottles, and

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deionized water (60 mL) was added into each bottle, then let them sit for 24 h. Then, 2.5 g

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of biochars were added into the bottles to yield the final concentration of 5% (dry weight

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sediment basis). The bottles were sealed with caps, and then the sediments with biochar

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were thoroughly mixed by shaking. Afterward, the samples were incubated in an artificial

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climate chamber at 20 ± 1 °C for 90 days. Three replicates were prepared for each sample,

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and the sediment without biochar addition was set as control (CK).

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2.5. Equilibrium concentration in aqueous phase

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After incubation, the upper solution in each bottle was withdrew and got through a 0.45

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µm filter. After the upper solution was completely removed using syringes for sample

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pipetting, part of the remained sediment was placed into centrifuge tube and centrifuged

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(4000 r min–1, 15 min). The supernatant in centrifuge tube was taken out and filtered (0.45

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µm filter) to obtain pore water. The concentration of arsenic and cadmium in overlying

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water (COW) and pore water (CPW) were determined by ICP-OES (PerkinElmer Optima

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5300DV). The residual sediment was freeze-dried and ground. Then, the obtained

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samples were passed through a sieve (100-mesh), and kept at 4°C for later analysis.

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2.6. Toxicity characteristic leaching procedure (TCLP)

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The leachability of As and Cd in the sediment was determined by the toxicity

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characteristic leaching procedure (TCLP) based on U.S. EPA Method 1311 (Shen et al.,

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2017). The extraction solution was prepared using CH3COOH (kept pH at 4.93 ± 0.05).

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Then, 2 g sediment and 40 mL extraction solution were mixed and shaken for 18 h (190 r

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min–1, 25°C). The mixture was then centrifuged (4000 r min–1, 15 min), and the obtained

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supernatant was taken out and filtered. The As and Cd amount in the supernatant were

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measured using ICP-OES.

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2.7. Community Bureau of Reference (BCR) sequential extraction

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Four different kinds of As and Cd speciation in sediment were measured using the

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BCR sequential extraction method (Mossop and Davidson, 2003; Jiang et al., 2012). The

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above sediment sample (1.0 g) was used in the four sequential steps. The extractants that

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used for the extraction of four fractions (acid soluble fraction, reducible fraction,

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oxidizable fraction, and residual fraction) in each step were: 0.11 M CH3COOH; 0.5 M

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NH2OH•HCl and 0.05 M HNO3; 8.8 M H2O2 followed by 1.0 M CH3COONH4 (adjusted

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to pH 2 with HNO3); and acid digestion (HNO3, HF and HClO4).

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2.8. As and Cd accumulation from the sediment by tubifex

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Tubifex was obtained from Tianhong cultivation and planting base (Hubei, China),

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which were then cultured in deionized water with continuous aeration in an artificial

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climate chamber (20 ± 1 °C, light/darkness of 12 h/12 h). The worms were acclimatized

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by feeding with tetraMin Flakes (Tetra Werke, Melle, Germany) for 1 week before the

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further experiments (Liu et al., 2014). After the incubation of sediment as mentioned

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above, screw mouth bottles were opened and let them sit for 24 h. Then, approximately 2

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g of tubifex was added to each sediment/water system for As and Cd accumulation by

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tubifex. The test worms were sampled at days 7 using forceps and rinsed with deionized

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water. During the experiment period, the evaporated water in bottle was compensated by

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adding the weighed deionized water every day. The collected worms were digested by

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acid (HNO3, HF and HClO4) and the As and Cd amount were determined by ICP-OES.

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2.9. Analysis of bacterial community

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CTAB/SDS method was applied for the extraction of total genome DNA from

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samples. After the amplification, quantification and qualification of genes results, the

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obtained data was analyzed by a series of kit and software (Wang et al., 2007; Magoč and

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Salzberg, 2011; Caporaso et al., 2010; Jia et al., 2017; Awasthi et al., 2017). The detailed

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measure and analysis methods were provided in the Supporting Information.

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3. Results and discussion

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3.1. Characterization of different biochars

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Physic/chemical properties of rice straw and rice husk biochars are shown in Table 1.

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The results demonstrated that C content increased, while the content of O, H and N

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decreased with the increase of pyrolysis temperature. H/C and O/C molar ratios decreased

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at higher pyrolysis temperature. H/C is often used as a measure of the degree of

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carbonization, the decrease of H/C indicated the increased degree of carbonization of the

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two kinds of biochars. Van Krevelen diagram is widely applied to present the elemental

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content changes of biochar with pyrolysis temperature (Keiluweit et al., 2012; Li et al.,

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2013b). The gradual decrease of H/C and O/C with increasing temperature followed the

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trajectory of dehydration reactions can be seen in Fig. 1i. These trends in elemental

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composition and atomic ratios with the increase of pyrolysis temperatures were

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accordance with the previous researches (Ahmad et al., 2012; Chen et al., 2012b; Kim et

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al., 2013). The results of XPS surveys suggested that carbon and oxygen were the main

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content in biochar (Supporting Information Fig. S1). The change tendencies of C1s and

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O1s spectra of biochars produced from two feedstocks and at different temperatures were

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consistent with the elemental composition analysis, which suggested that the C content

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increased and the content of O decreased with the increase of pyrolysis temperature. The

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O/C molar ratio indicates the amount of oxygen-containing groups in biochar, which may

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influence its ability in pollution control.

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Pyrolytic temperature also affected the surface chemical properties of biochars.

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Some FTIR spectra bands of rice straw biochar decreased slightly at higher pyrolytic

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temperature (Supporting Information Fig. S2), including −OH (3426 cm–1), aliphatic

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chain –CH2 and –CH3 (2948 cm–1), phenolic –OH and aromatic CO– (1250 cm–1), and

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C−O at (1106 cm−1) (Chen et al., 2012a). This was attributed to the enhanced

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carbonization reaction of different components in biomass (Chen et al., 2012a). For rice

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husk biochar, the peak strength including C=O (1631 cm−1), phenolic –OH and aromatic

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CO– (1250 cm–1), and C−O (1106 cm−1) decreased slightly (Tan et al., 2015), but the

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other peak positions did not change obviously. The above results suggested that the

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production temperature of biochar posed a great influence on its surface chemical

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properties.

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The BET surface area of rice straw biochar and rice husk biochar increased from

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0.367 to 51.105 m2 g−1 and 0.226 to 22.479 m2 g−1 with increasing pyrolytic temperature

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(Table 1), respectively. This might be attributed to volatile substances release and

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appearance of vascular bundles structures after pyrolysis (Ahmad et al., 2012; Chen et al.,

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2012b; Kim et al., 2013). The pyrolysis temperature is an important factor affecting the

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surface area of biochar, which may influence the environmental pollution remediation

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ability of biochar. Biochar produced at higher temperature possessed of higher surface

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area, which might favor heavy metal bonding. Therefore, it is especially important to

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select suitable preparation conditions before applying biochar to actual site remediation.

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Scanning electron micrographs of different biochars are shown in Fig. 1a-h. Rice

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straw biochar had a vascular bundle structure, and many micropores were distributed on

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the vascular bundle. At lower temperatures, the vascular bundle structure of straw biochar

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was relatively regular. As the pyrolysis temperature increased, the original pores became

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larger and new micropores appeared, which might be the reason why high-temperature

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biochar had a high specific surface area (Ahmad et al., 2012; Li et al., 2013a). Compared

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with the vascular bundle structure of straw biochar, the surface of the rice husk biochar

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had a uniformly arranged convex block structure. At lower temperatures, these raised

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block structures were relatively regular and smooth. As the pyrolysis temperature

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enhanced, the matrix of the rice husk biochar began to rupture and collapse, creating new

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micropores and exposing the internal pore structure, resulting in a higher specific surface

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area. The formation of new micropores and exposure of the internal pores might also

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provide more active site for heavy metal immobilization.

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Raman spectroscopy was used to evaluate the microstructure of biochars. The

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G-band near 1600 cm−1 related to the in-plane vibrations of the sp2-bonded crystallite

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carbon, while another peak centered at 1360 cm−1 denoted as the defect sites or

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disordered carbons (or D-band) (Zhao et al., 2013). The intensity ratio of D band versus

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G band (ID/IG) is commonly used to evaluate the degree of crystallization (representing

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the disorder degree) of carbonaceous materials (Zhang et al., 2019). As shown in Fig. 1j

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and k, for rice straw and rice husk biochars produced at 400–700°C, there was slight

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increase trend of ID/IG with temperature increasing (0.809–1.008 and 0.794–1.013,

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respectively), suggesting the increase of the proportion of condensed aromatic ring

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structures with defects (Sun et al., 2014; Guizani et al., 2017; Zhang et al., 2019). This

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might be due to the enhancement of aromatic rings and dehydrogenation of

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hydroaromatics at higher pyrolysis temperature, which increased the number of

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aromatic rings (Azargohar et al., 2014). Similar results were also reported by (Guizani

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et al., 2017) and (Azargohar et al., 2014). The results implied that ratio of defect site or

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disordered carbon to ordered graphite crystallite carbon was mainly determined by

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production temperature of biochars (Zhao et al., 2013). The XRD patterns of biochars

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are shown in Fig. 1l. All biochar samples presented a broad diffraction peak at 23–25°,

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attributing to the amorphous carbon property of biochar (Jonidi Jafari et al., 2017). The

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minerals in biochars produced from two feedstocks were significantly different, while

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were less affected by pyrolytic temperature. Three kinds of minerals with peaks labeled:

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C, calcite (PDF-# 05-0586); Q, quartz (PDF-# 46-1045); and S, sylvite (PDF-# 41-1476)

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were existed in these biochars (Zhang et al., 2015). The K content in biochars were

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measured (Table S2) and the results suggested that the K content in rice straw biochar

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(1723.0-1973.0 mg kg−1) was near 2 folds higher than that in rice husk biochar

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(841.3-1239.0 mg kg−1), so that the intensity of sylvite in XRD pattern of rice straw

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biochar was more obvious than that in rice husk biochar. This might be due to the fact

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that K was better accumulated in rice husk than in rice straw, similar results were also

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reported by previous studies (Yang et al., 2016; Deka et al., 2018).

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3.2. Equilibrium concentration in aqueous phase

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After treated by different biochars, As and Cd concentration in aqueous phase of the

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sediment is shown in Fig. 2. The addition of different biochars caused As and Cd amount

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in the overlying water to decrease to varying degrees, decreased by 26%-89% and

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22%-71% under the treatment of straw biochar, and decreased by 13%-92% and 5%-64%

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under the treatment of rice husk biochar, respectively. Meanwhile, the concentration of

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As and Cd in the pore water decreased by 36%-73% and 29%-43% under the treatment of

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straw biochar, and decreased by 24%-68% and 20%-54% under the treatment of rice husk

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biochar, respectively (Fig. 2c and d). The amount of As and Cd in water phase of the

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sediment decreased in different levels after biochar treatment, which indicated that

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biochar could effectively increase the As and Cd immobilization in the solid sediment and

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decrease their migration into the water phase. The pyrolysis temperature showed

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significant influence on the performance of biochar. The As and Cd amount in overlying

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water and pore water gradually reduced with increasing pyrolytic temperature for both

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kinds of biochars, which suggested that biochar prepared at a higher pyrolytic

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temperature had a better ability on improving the As and Cd retention efficiency of the

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sediment. Therefore, the biochars prepared at higher temperature were more favorable for

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As and Cd immobilization in the sediment.

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3.3. TCLP leachability

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The concentration of As and Cd in TCLP leaching liquor after treatment of sediment

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with different biochars is shown in Fig. 2e and f. After the treatment of rice straw and rice

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husk biochars, the concentration of As and Cd in TCLP leachate decreased by 3%-30%

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and 12%-25%, and decreased by 34%-67% and 2%-18%, respectively. After treating the

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As and Cd contaminated sediment using biochar prepared at different pyrolytic

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temperatures, the concentration of the As and Cd in TCLP leaching solution decreased

286

gradually with increasing pyrolytic temperature, and it was similar to the variation trend

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of above research. The biochars prepared at higher pyrolytic temperature had a stronger

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ability to fix As and Cd in the tested sediment.

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3.4. Speciation and distribution of As and Cd

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The speciation changes and distribution of As and Cd in different biochars amended

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sediment were analyzed by BCR sequential extraction (Fig. 3). The acid extractable

292

fraction after different biochars treatment significantly decreased compared with that of

293

the untreated sediment. The acid extractable speciation is a bioavailable fraction of heavy

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metals, and has a high mobility. The decrease in the proportion of this part indicated that

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biochars could immobilize As and Cd in the sediment and reduce their bioavailability.

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Acid-soluble Cd accounted for a high proportion of total Cd, indicating that Cd had

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higher mobility, which was usually present in the sediment as a readily available fraction

298

(Beesley et al., 2010). The acid extractable As was relatively less in total amount of As.

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The acid extractable fraction of the As and Cd both showed a decreasing trend with the

300

increasing pyrolysis temperature of the biochar, indicating that the high-temperature

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biochar was more conducive to adsorbing As and Cd and improving their stability, which

302

was in agreement with the As and Cd concentration in aqueous phase.

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The amount of reducible fraction increased at different extent comparing with the

304

original sediment. Reducible fraction is the fraction associated with Fe and Mn oxides. As

305

shown in Table S2, the tested biochars had large amount of Fe and Mn, which might

306

present in biochar as Fe and Mn oxides, resulting in the increase of reducible fraction of

307

the As and Cd in the biochar amended sediment. The metal(loid)s bound to these oxides

308

are unstable and remain available in anoxic conditions (Fuentes et al., 2008). The

309

oxidizable As and Cd was a small fraction in the sediment. It can be seen from the figure

310

that the addition of biochars increased the proportion of oxidizable Cd in the sediment

311

slightly, which was due to the complex between Cd and surface functional groups on

312

biochar (Guo et al., 2006). In addition, when biochar was incorporated, the residual As

313

and Cd enhanced significantly comparing with the original sediment, and increased with

314

the increase of the pyrolytic temperature. Residual heavy metals are forms that are not

315

easily bioavailable, indicating that biochar could effectively promote immobilization of

316

metal(loid)s.

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3.5. Potential immobilization mechanisms

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The results of this study indicated that rice waste biochars could effectively improve

319

the immobilization of As and Cd in sediment, and the higher-temperature biochar

320

exhibited better performance. The content of oxygen-containing groups and the BET

321

surface area of biochar are two important factors influencing biochar performance (Tan et

322

al., 2015b). The O/C molar ratio reflects the amount of oxygen-containing groups in

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biochar to some extent, and the specific surface area reflects the pore characteristics. The

324

correlation coefficient and significance of different parameters were analyzed based on

325

pearman, including the physicochemical characteristics of biochars (O/C and specific

326

surface area), sediments properties (TOC, CEC, and pH), metal(loid)s in water phase,

327

concentration in TCLP leachates, and metal(loid)s speciation (Table 2). It was found that

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both specific surface area and O/C of biochar had significant effects on sediments

329

properties and metal(loid)s immobilization in sediments. Of which, O/C of biochar only

330

had significant influence on the As and Cd in pore and overlying water. Apart from the

331

As and Cd in water phase, specific surface area of biochar also showed significant effects

332

on Cd concentration in TCLP leachates, acid extractable As, residual fraction of As, acid

333

extractable Cd, and residual fraction of Cd. There was a slight increase of pH after

334

biochar treatment, however the difference of pH of different biochars treated sediments

335

was not significant (Fig. S3). Therefore, it was inferred that pH might not be the main

336

influence parameter. It was found that TOC had high correlation coefficient, but it was

17

337

negative correlation with the remediation effect of biochar. Surface area of biochar was

338

highly positive correlation with its remediation efficiency. These results suggested that,

339

as for the biochars produced at different pyrolysis temperatures, the performance of

340

biochars were more determined by their specific surface area than oxygen-containing

341

groups of biochar.

342

The specific surface area of biochar is an important parameter determining its

343

performance. The biochars produced at higher pyrolysis temperatures processed of higher

344

surface area, which might provide more opportunities for metal(loid) bonding by pore

345

filling (Tan et al., 2015). As shown in Table S2, the tested biochars had large amount of Fe

346

and Mn, which might present in biochar as Fe and Mn oxides, resulting in the increase of

347

As and Cd immobilization in the biochar treated sediment. Fe and Mn oxides act as

348

sorption sites for As to be adsorbed onto the biochar via surface complexation (Alkurdi et

349

al., 2019). As could also interact with Ca and Mg in biochar (Table S2) by forming

350

relatively insoluble Ca-As and Mg-As precipitates to immobilize As in sediment (Niazi et

351

al., 2018). In addition, oxygen-containing functional groups on biochar surface (Fig. 2)

352

could contribute in adsorption of As and Cd onto biochar through ion exchange or surface

353

complexation mechanisms (Niazi et al., 2018). Large amount of P in these biochars

354

suggested the potential presence of PO43-, which might form ion exchange with As anion

355

and insoluble precipitate with Cd (Vithanage et al., 2017). It was noted that the content of

356

above-mentioned elements in the tested biochars all showed an increase trend with the

18

357

increase of pyrolysis temperature, which might be another reason for better performance

358

of high-temperature biochar apart from surface area.

359

3.6. Metal(loid)s accumulation from sediment by tubifex

360

Accumulation of As and Cd in the tubifex tissue under treatment of different biochars

361

is shown in Fig. 4a. The As content in original tubifex tissue was 0.053 mg kg−1, and the

362

Cd content was below the limits of detection. After the tubifex was exposed to the

363

contaminated sediment/water system, As and Cd amount in tubifex tissue significantly

364

increased by uptake from the sediment. The tubifex has a close contact with the solid

365

phase and water phase of the sediment, its former body is excavated in the sediment and

366

the latter body is undulated in the overlying water. Therefore, tubifex was exposed to

367

metal(loid)s through sediment, pore water, and overlying water by ingestion and/or

368

epidermal contact (Liu et al., 2014). The addition of biochars had obvious effect on the

369

As and Cd accumulation from sediment by tubifex, which decreased the As and Cd

370

amount in tubifex tissue to different extent. The pyrolytic temperature showed a

371

significant effect on the performance of biochar, and the efficiency of biochar prepared at

372

higher temperature was more significant, which was in agreement with the above results.

373

As and Cd concentrations in the tubifex tissue relative to their corresponding

374

concentrations in water phase of sediment are shown in Fig. 4b and c. It was found that As

375

and Cd contents in the tubifex tissue were positively correlated with their amount in

376

aqueous phase. These results indicated that biochar could effectively inhibit the As and

19

377

Cd accumulation from sediment by organisms through decreasing the bioavailable

378

fraction of As and Cd, and the higher-temperature biochar exhibited better performance.

379

3.7. Influence on indigenous microbial composition

380

The statistical comparisons of the ACE and Shannon indices between biochars

381

amended sediment and original sediment (CK) are shown in Supporting Information Fig.

382

S4. It was observed that there were larger ACE and Shannon indices in sediments

383

remediated by different biochars. ACE and Shannon index values indicated that the

384

richness and biodiversity of microbial community were both greater in all biochars

385

treated sediment compared to non-treated sediment (Fang et al., 2017; Mattei et al., 2017).

386

The bacterial communities in sediment were analyzed on genus level, the heatmap of the

387

relative abundance for genus in sediments treated by different biochars is shown in

388

Supporting Information Fig. S5. Bacteroides, Leptospirillum, Blautia, Akkermansia,

389

Parasutterella, Halomonas, Acinetobacter, Exiguobacterium were the most abundant

390

microbial communities in the non-treated sediment, which was lack of other species. The

391

result was consistent with the previous ACE and Shannon index values, which indicated

392

the lower richness and biodiversity of microbial community in the original sediment. The

393

pristine strains of dominant species were remained after the treatment by different

394

biochars, but they were no longer the dominant strains and some new dominants were

395

formed (Liu et al., 2018). The results suggested that biochars could change the microbial

396

composition in sediment and increase the microbial richness and biodiversity.

20

397

The contributions and distributions of microbial composition similarity were

398

classified into different groups by applying principal coordinates analysis (PCoA) (Fig.

399

5) (Jia et al., 2017). On the PCoA plot, PC1 and PC2 axis explained 37.70% and 20.01%

400

of the variation, respectively. The analysis revealed the similar microbial composition in

401

biochars treated sediments, as communities from the biochars treated sediments tended

402

to be gathered together and obviously separated from the non-treated sediments (Jia et

403

al., 2017; Chen et al., 2019). Canonical correspondence analysis (CCA) plot was

404

constructed to identify the correlation between the environmental parameters and

405

microbial community in sediment. The length of the arrow of environmental parameter

406

shows its influence strength to the overall microbial communities (Sun et al., 2015;

407

Weisener et al., 2017; dos Reis Oliveira et al., 2018). The axis 1 accounted for 41.27%

408

of the variation and axis 2 explained 17.17% of the variation. The environmental

409

parameters including concentration of As and Cd in the pore water (PW As and PW Cd),

410

acid extractable fraction of As and Cd (F1 As and F1 Cd), pH, TOC, and CEC appeared

411

to be the key environmental parameters influencing microbial communities in sediment.

412

The results suggested that the amount and speciation of As and Cd significantly affected

413

the microbial communities in sediment, which was changed after the remediation of

414

biochars as discussed in the previous sections. In addition, pH, TOC, and CEC of

415

sediment increased after biochars treatment (Supporting Information Fig. S3), which

416

also played key role in changing microbial communities in sediment (Wan et al., 2017;

21

417

Chen et al., 2019).

418

4. Conclusions

419

Rice waste biochars presented obvious effectiveness for arsenic and cadmium

420

abatement in sediment. Biochars produced at higher pyrolytic temperatures showed

421

more significant effect on As and Cd immobilization. Equilibrium concentration in

422

aqueous phase and TCLP leachability of As and Cd were decreased after biochars

423

treatment. Biochars decreased bioavailable fraction and increased residual fraction of As

424

and Cd. High-temperature biochars significantly reduced As and Cd bioaccumulation in

425

tubifex. The biochars could increase the microbial richness and biodiversity. These

426

results indicated that biochar could effectively inhibit the bioavailability and toxicity of

427

As and Cd in sediment, and the high-temperature biochar exhibited better performance.

428

Further research is required to determine the efficacy of biochar in long-term and field

429

application.

430 431

ACKNOWLEDGEMENTS

432

The present work was funded by the National Natural Science Foundation of China

433

(Grants Nos. 51521006, 51809089 and 51609268), the Key Project of Technological

434

Innovation in the Field of Social Development of Hunan Province, China (Grant Nos.

435

2016SK2010 and 2016SK2001), the Natural Science Foundation of Hunan Province,

436

China (Grant Nos. 2018JJ3040, 2018JJ3096, and 2019JJ50409), and the Science and

22

437

Technology Plan Project of Hunan Province (No. 2019NK2062, 2018SK2047).

438

23

439

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440 441 442 443 444 445 446 447 448 449 450 451 452 453 454 455 456 457 458 459 460 461 462 463 464 465 466 467 468 469 470 471 472 473 474 475 476 477 478

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610

28

611 612

Table 1. Physico-chemical characteristics of biochars produced at different pyrolysis temperatures. Biochar

Elemental composition based) C H O

(%,

mass

Atomic ratios

N

O/C

H/C

(N+O)/C

BET surface −1 area (m2 g )

RS400

51.79

4.49

21.21

1.25

0.31

1.04

0.33

0.367

RS500

52.06

3.66

17.48

1.11

0.25

0.84

0.27

1.841

RS600

54.70

2.65

11.58

0.98

0.16

0.58

0.17

8.598

RS700

55.18

2.00

10.10

0.77

0.14

0.43

0.15

51.105

RH400

49.34

4.43

26.57

0.69

0.40

1.08

0.42

0.226

RH500

53.33

3.44

15.20

0.76

0.21

0.77

0.23

1.356

RH600

57.26

2.69

9.93

0.74

0.13

0.56

0.14

3.356

RH700

57.68

1.99

6.73

0.65

0.09

0.41

0.10

22.479

29

613

Table 2. Correlation coefficient and significance analysis based on pearman.

O/C

SA

TOC

pH

CEC

OWAs

O/C

SA

TOC

pH

CEC

OWAs

PWAs

OWCd

PWCd

TCLPAs

TCLPCd

F1As

F4As

F1Cd

F4Cd

Correlation coefficient

1

-.857**

.905**

-0.683

0.214

.976**

.810*

.905**

.830*

0.643

0.59

0.524

-0.619

0.667

-0.286

Significance

.

0.007

0.002

0.062

0.61

0

0.015

0.002

0.011

0.086

0.123

0.183

0.102

0.071

0.493

-.857**

1

-.976**

.946**

0.19

-.929**

-.905**

-.881**

-.927**

-0.286

-.868**

-.857**

.810*

-.929**

0.69

Significance

0.007

.

0

0

0.651

0.001

0.002

0.004

0.001

0.493

0.005

0.007

0.015

0.001

0.058

Correlation coefficient

.905**

-.976**

1

-.898**

-0.024

.952**

.881**

.857**

.952**

0.381

.831*

.786*

-.738*

.881**

-0.619

Significance

0.002

0

.

0.002

0.955

0

0.004

0.007

0

0.352

0.011

0.021

0.037

0.004

0.102

Correlation coefficient

-0.683

.946**

-.898**

1

0.323

-.778*

-.790*

-.731*

-.834**

-0.096

-.952**

-.934**

.874**

-.994**

.790*

Significance

0.062

0

0.002

.

0.435

0.023

0.02

0.04

0.01

0.821

0

0.001

0.005

0

0.02

Correlation coefficient

0.214

0.19

-0.024

0.323

1

0.119

-0.31

0.024

-0.171

.762*

-0.41

-0.595

0.429

-0.286

.714*

Significance

0.61

0.651

0.955

0.435

.

0.779

0.456

0.955

0.686

0.028

0.313

0.12

0.289

0.493

0.047

.976**

-.929**

.952**

-.778*

0.119

1

.857**

.952**

.878**

0.571

0.663

0.619

-0.643

.762*

-0.405

0

0.001

0

0.023

0.779

.

0.007

0

0.004

0.139

0.073

0.102

0.086

0.028

0.32

Correlation coefficient

.810*

-.905**

.881**

-.790*

-0.31

.857**

1

.833*

.952**

0.238

.711*

.786*

-0.667

.762*

-.714*

Significance

0.015

0.002

0.004

0.02

0.456

0.007

.

0.01

0

0.57

0.048

0.021

0.071

0.028

0.047

Correlation coefficient

.905**

-.881**

.857**

-.731*

0.024

.952**

.833*

1

.781*

0.548

0.566

0.571

-0.595

.714*

-0.357

Significance

0.002

0.004

0.007

0.04

0.955

0

0.01

.

0.022

0.16

0.143

0.139

0.12

0.047

0.385

Correlation coefficient

.830*

-.927**

.952**

-.834**

-0.171

.878**

.952**

.781*

1

0.244

.790*

.805*

-0.659

.805*

-.732*

Significance

0.011

0.001

0

0.01

0.686

0.004

0

0.022

.

0.56

0.02

0.016

0.076

0.016

0.039

Correlation coefficient

0.643

-0.286

0.381

-0.096

.762*

0.571

0.238

0.548

0.244

1

-0.108

-0.214

0.024

0.143

0.405

Significance

0.086

0.493

0.352

0.821

0.028

0.139

0.57

0.16

0.56

.

0.798

0.61

0.955

0.736

0.32

Correlation coefficient

0.59

-.868**

.831*

-.952**

-0.41

0.663

.711*

0.566

.790*

-0.108

1

.964**

-.916**

.928**

-.831*

Correlation coefficient

Correlation coefficient Significance

PWAs

OWCd

PWCd

TCLPAs

TCLPCd

30

F1As

F4As

F1Cd

F4Cd

614 615 616 617

O/C

SA

TOC

pH

CEC

OWAs

PWAs

OWCd

PWCd

TCLPAs

TCLPCd

F1As

F4As

F1Cd

F4Cd

Significance

0.123

0.005

0.011

0

0.313

0.073

0.048

0.143

0.02

0.798

.

0

0.001

0.001

0.011

Correlation coefficient

0.524

-.857**

.786*

-.934**

-0.595

0.619

.786*

0.571

.805*

-0.214

.964**

1

-.881**

.905**

-.929**

Significance

0.183

0.007

0.021

0.001

0.12

0.102

0.021

0.139

0.016

0.61

0

.

0.004

0.002

0.001

Correlation coefficient

-0.619

.810*

-.738*

.874**

0.429

-0.643

-0.667

-0.595

-0.659

0.024

-.916**

-.881**

1

-.857**

0.667

Significance

0.102

0.015

0.037

0.005

0.289

0.086

0.071

0.12

0.076

0.955

0.001

0.004

.

0.007

0.071

Correlation coefficient

0.667

-.929**

.881**

-.994**

-0.286

.762*

.762*

.714*

.805*

0.143

.928**

.905**

-.857**

1

-.762*

Significance

0.071

0.001

0.004

0

0.493

0.028

0.028

0.047

0.016

0.736

0.001

0.002

0.007

.

0.028

Correlation coefficient

-0.286

0.69

-0.619

.790*

.714*

-0.405

-.714*

-0.357

-.732*

0.405

-.831*

-.929**

0.667

-.762*

1

Significance

0.493

0.058

0.102

0.02

0.047

0.32

0.047

0.385

0.039

0.32

0.011

0.001

0.071

0.028

.

O/C: oxygen/carbon molar ratios; SA: specific surface area; TOC: total organic carbon; CEC: cation exchange capacity; OWAs: As in overlying water; PWAs: As in pore water; OWCd: Cd in overlying water; PWCd: Cd in pore water; TCLPAs: As concentration in TCLP leachates of sediment; TCLPCd: Cd concentration in TCLP leachates of sediment; F1As: acid extractable As; F4As: residual fraction of As; F1Cd: acid extractable Cd; F4Cd: residual fraction of Cd. ** Very significant at P< 0.01. * Significant at P< 0.05.

31

618 619

Fig. 1. SEM photograph of biochars produced at different pyrolysis temperatures: (a)

620

RS400, (b) RS500, (c) RS600, (d) RS700, (e) RH400, (f) RH500, (g) RH600, and (h)

621

RH700. (i) The van Krevelen plot of elemental ratios for biochars produced at different

622

pyrolytic temperatures. (j, k) Raman spectroscopy of biochars produced at different

623

pyrolysis temperatures. (l) XRD patterns of biochars produced at different pyrolysis

624

temperatures. Minerals with peak labeled: Q, quartz; C, calcite; S, sylvite.

625

32

626 627

Fig. 2. Heavy metal concentrations in overlying water and pore water of sediment under

628

treatment of different biochars: (a) As in overlying water, (b) Cd in overlying water, (c)

629

As in pore water, (d) Cd in pore water; Heavy metals concentrations in TCLP leachates of

630

sediment after treatment of different biochars: (e) concentration of As, (f) concentration

631

of Cd. Results shown in the figures are mean values ± standard deviations (n = 3) and

632

error bars indicate standard deviations. * significant at p < 0.05 compared with control

633

(CK, un-treated sediment).

634

33

635 636

Fig. 3. Fraction of As and Cd in sediment determined by BCR sequential extraction after

637

treatment of biochars produced at different pyrolysis temperatures: (a) fraction of As, (b)

638

fraction of Cd.

639

34

640 641

Fig. 4. (a) Accumulation of heavy metals in the tubifex tissue under treatment of

642

different biochars; (b and c) Heavy metal concentrations in the tubifex tissue relative to

643

their corresponding concentrations in overlying water and pore water of sediment under

644

treatment of different biochars. Results shown in the figures are mean values ± standard

645

deviations (n = 3) and error bars indicate standard deviations. * significant at p < 0.05

646

compared with control (CK, un-treated sediment).

647

35

648 649

Fig. 5. (a) The contributions and distributions of microbial composition similarity

650

classified into distinct groups by applying principal coordinates analysis (PCoA); (b) The

651

correlation between the environmental parameters and microbial community in

652

sediment by applying canonical correspondence analysis (CCA) plot.

36

Highlights •

Rice waste biochars were effective for metals abatement and detoxification in sediment.

•

High-temperature biochars exhibited greater effect on heavy metal immobilization.

•

High-temperature biochars reduced heavy metal bioaccumulation in Tubifex tubifex.

•

Biochars could increase the microbial richness and biodiversity of sediment.

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