IronâManganese (Oxyhydro)oxides, Rather than Oxidation of Sulfides

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Environmental Processes

Iron-manganese (oxyhydro)oxides, rather than oxidation of sulfides, determine the mobilization of Cd during soil drainage in paddy soil systems Jing Wang, Ping-Mei Wang, Yi Gu, Peter M Kopittke, Fangjie Zhao, and Peng Wang Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.8b06863 • Publication Date (Web): 11 Feb 2019 Downloaded from http://pubs.acs.org on February 13, 2019

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Environmental Science & Technology

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Iron-manganese (oxyhydro)oxides, rather than oxidation of sulfides, determine the

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mobilization of Cd during soil drainage in paddy soil systems

3 4

Jing Wang1, Ping-Mei Wang1, Yi Gu1, Peter M. Kopittke2, Fang-Jie Zhao1, and Peng Wang1,*

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1

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210095, China

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2

9

4072, Australia

Nanjing Agricultural University, College of Resources and Environmental Sciences, Nanjing

The University of Queensland, School of Agriculture and Food Sciences, St Lucia, Queensland,

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*Corresponding author: Peng Wang, Phone: +86 25 8439 6509, Email: [email protected]

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ORCID IDs: 0000‑0001-8622‑8767 (P.W.); 0000‑0003‑4948‑1880 (P.M.K.); 0000-0002-0164-

14

169X (F.-J.Z.)

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Abstract

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The pre-harvest drainage of rice paddy fields during the grain filling stage can result in a substantial

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mobilization of Cd in soil and consequently elevated grain Cd concentration. However, the

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processes controlling the mobilization of Cd remains poorly understood. Using twelve field-

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contaminated paddy soils, we investigated the factors controlling the temporal changes in Cd

20

solubility in paddy soils that were incubated anaerobically for 40 d followed by a 20 d oxidation

21

period. Soluble and extractable Cd concentrations decreased rapidly upon flooding but increased

22

during the oxidation phase, with Cd solubility (aqueous Cd/soil Cd) largely depending upon

23

porewater pH. Furthermore, inhibiting sulfate reduction or inhibiting oxidation dissolution of Cd-

24

sulfides had little or no effect on the mobilization of Cd in the subsequent oxidation phase. Both

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sequential extraction and X-ray absorption spectroscopy (XAS) analyses revealed that changes in

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Cd solubility were largely dependent upon the transformation of Cd between the Fe-Mn

27

(oxyhydro)oxide fraction and exchangeable fraction. Mobilization of Cd upon soil drainage was

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caused by a decrease in soil pH resulting in the release of Cd from Fe-Mn (oxyhydr)oxides. Taken

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together, Fe-Mn (oxyhydr)oxides play a critical (and prevalent) role in controlling the mobilization

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of Cd upon soil drainage in paddy systems.


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TOC

8

600 450

pH

7

300

6

150

5

Eh (mV)

32

0 Fe-Mn oxides control Cd mobilization upon drainage of paddy soils 4

-150 0

6

pH 4.5 5.5

450

6.5

150

5

-150 0

10 20 30 40 50 60

Time (d)

33

300

0

4

100

600

Cd fraction (%)

pH

7

Drainage

Eh (mV)

Soil flooding

8

Soil flooding

Drainage

amorphous Fe-Mn oxides Cd-sorbed amorphous Fe-Mn oxidesbound Cd

80

Fe-Mn oxides

pH 4.5 5.5

Cd Fe-Mnbound (oxyhydr)oxides

60 40

10 20 30 40 50 60

(unav ail.) mobile Cd Cd mobile

20

mobile Cd mobile Cd

6.5

Cd-sulfides Cd-sulfides

0 0

10

20

30

40

50

60

Time (d)



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Introduction

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Cadmium (Cd), a toxic element to human, is a ubiquitous contaminant found in most foodstuffs.

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Dietary intake contributes approximately 90% to the total Cd exposure in non-smoking

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populations.1-3 Among plant-derived foods, the grain of rice (Oryza sativa), the staple food for

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about half of the world’s population, tends to accumulate higher levels of Cd compared to other

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cereal crops,4 likely due to the higher expression and functionality of the OsNramp5 gene

40

(responsible for Cd uptake by roots) in the rice plant.5 Rice is therefore the largest contributor to the

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dietary intake of Cd in the populations for which rice is a staple food. For example, rice contributes

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40% of the dietary Cd intake for the general population in Japan,6 56% for the general population in

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China,4 and up to 81% for subsistence rice farmers in a Cd-contaminated area of southern China.7

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For such populations, long-term consumption of Cd-contaminated rice can result in serious health

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problems including irreversible renal dysfunction, low bone density, and even the Itai-Itai disease

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(an extreme poisoning case).8-11 Therefore, the transfer of Cd from paddy soil to the food table is an

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important issue of food safety.

48 49

Cultivation of paddy rice is characterized by episodic flooding and drainage phases.12 Typically,

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paddy water is drained during the last two-three weeks of grain filling before plant maturity. This

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practice can result in a marked increase in Cd mobilization in the soil and consequently elevated

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grain Cd concentration. For example, in a pot experiment, the grain Cd concentrations for plants

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grown in continually flooded soils was found to be 0.005-0.01 mg kg-1, compared to 0.27-0.36 mg

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kg-1 when soils were drained after heading.13 In a similar manner, in a field experiment, grain Cd

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concentrations increased from 0.08 mg kg-1 in a flooded soil to 0.40 mg kg-1 in a soil drained after

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heading.14 These results suggest that the majority of Cd in rice grain (e.g. >80% in the two studies

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listed above) accumulates during the phase when the soil is drained prior to harvest. It is therefore


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critically important to investigate the factors controlling Cd solubility in paddy soils, especially

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during soil pre-harvest drainage, in order to develop strategies to control grain Cd accumulation.

60 61

Changes in the soil oxidation status have a large impact upon the speciation and solubility of Cd in

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the soil. A number of studies have reported a decrease in Cd solubility during soil flooding, with

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this effect being attributed to the formation of Cd-sulfides,15-17 although there is little direct

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evidence.18-23 Khaokaew et al. (2011) used synchrotron-based X-ray absorption spectroscopy

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(XAS) to investigate the changes of Cd speciation in a highly contaminated alkaline paddy soil

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(142 mg Cd kg-1) over 150 d of flooding. These authors found that Cd carbonates were the

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dominant Cd species, with a further 15-30% of the Cd being CdS after 30 d flooding.18 In a

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microcosm incubation experiment with a Cd-spiked paddy soil (20 mg kg-1 Cd), Fluda et al. (2013)

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also determined the speciation of Cd using XAS and found that changes in Cd solubility under

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flooded conditions were associated with microbial sulfate reduction.22 Similarly, Hashimoto and his

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colleagues used XAS to investigate the changes in the speciation and solubility of Cd over time in

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Cd-spiked paddy soils (150 mg Cd kg-1).19-21, 23 They showed that the Cd extractability decreased as

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the proportion of CdS increased in flooded soils. These previous results have clearly demonstrated

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that sulfate reduction plays an important role in controlling Cd solubility in reduced soils. Thus,

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prolonged flooding has been proposed to immobilize Cd through the formation of Cd sulfides,

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decreasing grain Cd concentrations compared to drained soils.13, 24 25 However, it remains unclear if

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the formation of Cd-sulfides influences the mobilization of Cd when the reduced soils are drained.

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Currently, little is known about the factors controlling Cd mobilization during soil drainage.

79 80

The aim of the present study was to investigate the factors controlling Cd mobilization in paddy

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soils, focusing on the processes occurring upon drainage. We investigated the temporal changes in

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soluble and extractable Cd concentrations and in Cd speciation in the solid phase during the 5|Page ACS Paragon Plus Environment


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reduction and oxidation phases using XAS. Unlike previous studies, we used field-contaminated

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paddy soils (1.36-11.1 mg Cd kg-1) without spiking of large doses of exogenous Cd. Our results

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demonstrate that Fe-Mn (oxyhydro)oxides play a critical role in controlling Cd solubility in paddy

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soils, and that the concomitant decrease in soil pH upon soil drainage is the key factor resulting in

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the enhanced Cd solubility in drained paddy soils.

88 89

Materials and Methods

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Soil collection and characterization

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Twelve paddy soils were collected from the plow layer (0-20 cm) of Cd-contaminated paddy fields,

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located in Guangxi (GX), Fujian (FJ), and Hunan (HN) provinces of southern China. All these

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paddy sites were contaminated with Cd either due to mining activities or due to irrigation with

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contaminated water. After collection, soils were air-dried and passed through a 4 mm sieve. Soil

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pH, the contents of soil organic carbon (SOC) and water-extractable sulfate, and soil metal

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concentrations were determined as described previously26, 27 (Table S1 in the Supporting

97

Information (SI)).

98 99

Soil incubation experiments

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Six incubation studies were conducted to investigate the factors controlling the solubility of Cd in

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paddy soils under anaerobic and aerobic conditions. The incubation procedures simulated the

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occurrence of typical flooded/drained paddy system and followed the microcosm method as

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described by Fulda et al.22 with some modifications (SI). In brief, 30 g aliquots of air-dried soils

104

were placed into 120 ml vials to which 60 ml deionized water was added. The soils were first

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anaerobically incubated for up to 40 d (hereafter denoted as ‘reducing phase’, R). Subsequently, the

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soils were allowed to oxidize for a further 20 d (‘oxidation phase’ O).

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Experiment 1 used the 12 paddy soils to investigate the temporal changes in soil redox potential

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(Eh), porewater pH, soluble Cd, and other major soluble ions during a 40 d anaerobic incubation (R

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phase) and a subsequent 20 d aerobic incubation (O phase). At different incubation periods within

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the R phase (0, 1, 5, 20, and 40 d; hereafter denoted as R0, R1, R5, R20, and R40) and the O phase

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(1, 3, 7, and 20 d; hereafter denoted as O1, O3, O7, and O20), three vials from each incubation

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series were destructively sampled for solution analyses. Once the vials were opened, soil Eh was

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immediately measured at ca. 3 cm underneath the soil surface using a Pt combined electrode (OPR-

115

33C, EAI, UK). The soil suspensions were centrifuged at 1735 g for 10 min. Solution pH was

116

measured with a pH electrode in the unfiltered supernatants. The filtered supernatants were

117

acidified for analyses of major metals using inductively coupled plasma mass spectrometry (ICP-

118

MS, Perkin Elmer NexION 300X, USA) and major anions using ion chromatography (IC, ICS-600,

119

Dionex, USA) (SI). Unless stated otherwise, all sampling was conducted in a glovebox (MBstar,

120

MBraun, Germany) under an Ar atmosphere.

121 122

Based on cluster analysis of the soil properties of the 12 soils used in Experiment 1 (Figure S1),

123

three representative soils (GX, FJ, and HN-1) were chosen for Experiment 2. The three soils were

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incubated as described in Experiment 1. At various incubation periods within the R and O phases,

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three vials from each incubation series were destructively sampled for both solution (as described

126

above) and solid-phase analyses. For solid-phase analyses, soil materials were obtained after

127

centrifugation and divided into two subsamples. One subsample was used immediately as a wet

128

paste for extraction using 0.1 M CaCl2 (denoted hereafter as extractable Cd), reflecting

129

‘exchangeable Cd’ as described previously (SI).28 The other subsample was shock-frozen in liquid

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N2, freeze-dried under an O2-free atmosphere, and sealed in gastight bags for analysis of Cd

131

speciation using synchrotron-based XAS (see later).

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Experiment 3 aimed to determine the extent to which sulfate reduction influenced Cd solubility

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during the reduction-oxidation cycle (with this experiment simultaneously carried out with

135

Experiment 2). The three soils (GX, FJ, and HN-1) were added with 0 (Control) or 20 mM

136

molybdate (denoted as ‘+ Mo’). Sodium molybdate was added as an inhibitor of sulfate reducing

137

bacteria.38 At each time point of sampling during the reduction and oxidation phases, three vials

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from each treatment (Control and +Mo) were sacrificed for solution (i.e. pH, sulfate, soluble Cd)

139

and solid-phase analyses (i.e. extractable Cd and Cd speciation) as described above.

140 141

Experiment 4 investigated whether microbially-mediated oxidation dissolution of the Cd-sulfides

142

formed in reduced soil influenced the mobilization of Cd during the subsequent oxidation phase

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(with this experiment simultaneously carried out with Experiment 2). The oxidation dissolution of

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metal sulfides has been reported to be largely mediated by microbes.29 In this experiment, two soils

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(GX and FJ) were incubated as described in Experiment 2, but additional 18 vials from each soil

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were prepared at the same time with the control. After 39 d of reduction phase, these additional

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vials were then gamma irradiated with 50 k Gray (produced from a Co source, BFT-IV, China) to

148

sterilize the soil.36 These vials were then opened at 40 d of flooding (R40), prior to the beginning of

149

the oxidation phase. At R40, O1, O3, O5, O7, and O20, the suspensions were centrifuged for

150

solution (i.e. pH, sulfate, soluble Cd) and solid-phase (i.e. extractable Cd and Cd speciation using

151

XAS) analyses as described above. At the end of oxidation phase, culturable cells/colonies were

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counted using the plate count method to check the efficacy of gamma irradiation.37

153 154

Experiment 5 investigated whether different Cd species or soil pH influenced the mobilization of

155

Cd during the oxidation phase. It has been reported that Cd can form different species in reduced

156

soils, including CdS, CdCO3 and adsorbed Cd.18, 22, 23 In this experiment, an acid soil (HN-2) was

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chosen and two soil pH treatments were applied, being a control (pH 5.18, unadjusted) and 8|Page ACS Paragon Plus Environment

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amendment of lime to increase initial pH to 6.70 (pH 6.70, adjusted). For each pH soil, the same

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amount of Cd as different Cd species, including CdS, CdCO3 and CdSO4, and a control (only

160

addition of deionized water), were exogenously added into the vials at 15 d flooding (R15) in a

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glovebox and mixed thoroughly with the soils before the vials were sealed for continuing

162

incubation. The amounts of Cd added ranged from 7.5 to 10 mg kg-1, comparable to the background

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Cd in the HN-2 soil (6.8 mg kg-1, the control). At R20, R40, O7, and O20, the suspensions were

164

centrifuged for solution (i.e. pH, sulfate, soluble Cd) and solid-phase (i.e. extractable Cd) analyses

165

as described in Experiment 2.

166 167

Experiment 6 further investigated the dynamics of Cd associated with various soil solid phases

168

during the reduction-oxidation processes. The three soils (GX, FJ, and HN-1) were incubated as

169

described in Experiment 2. At different time points within the soil reduction and oxidation phases,

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the solid-phase materials were collected for sequential extractions as outlined by Carlsson et al.30

171

with some modifications (SI). Briefly, the suspensions were centrifuged and the Cd in the

172

supernatants measured using ICP-MS as soluble fraction (F1). The five sequential extractions were

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conducted with a 1.0 g soil (dry-weight basis) in 50 ml polypropylene tubes with 0.1 M CaCl2

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(exchangeable Cd, F2), 1.0 M CH3COONH4 at pH 5.0 (specifically-adsorbed Cd, F3), 1.0 M

175

NH2OHHCl (amorphous Fe-Mn oxides-bound Cd, F4), and KClO3 at 12 HCl (organic and sulfides,

176

F5). The concentrations of Cd in the residual fraction were also digested using aqua regia (4:1

177

HNO3/HCl)(F4). The detailed sequential fractionation procedures were presented in Table S2.

178 179

Cadmium speciation using X-ray absorption spectroscopy (XAS)

180

The speciation of Cd was determined in situ using Cd K-edge (26,711 eV) X-ray absorption near

181

edge structure (XANES) spectroscopy, which was collected at the XAS Beamline of the Australian

182

Synchrotron, Melbourne (refer to ref40 for more details). Soil samples were ground using a mortar 9|Page ACS Paragon Plus Environment


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and pestle, and sieved through a 250 µm sieve before analysis. In addition to 52 experimental

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samples, a total of 25 standard compounds of Cd were also analyzed (Table S3 and Figure S2). The

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average spectra were energy normalized using the Athena software package.41 Principle component

186

analysis (PCA) of the normalized spectra indicated that only three components were required to fit

187

the data. Target transformation (TT) was used to identify the relevant standards and standards with

188

a SPOIL value < 2 (Table S3) were selected for the linear combination fitting (LCF) analysis.31 The

189

LCF analysis was performed using Athena software with a fitting range of -30 to +100 eV relative

190

to the Cd K-edge (SI). In the current study, it is difficult to distinguish the XANES spectra for the

191

Cd bound with succinate, citrate, and malate. Therefore, these carboxyl-bound compounds yielded

192

in the LCF analysis were referred to as ‘O-coordinated Cd’, a proxy for Cd sorbed/bound to Fe

193

oxides, nature organic matter, or clay minerals.

194 195

Statistical analysis

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All data are presented as the mean ± standard error (n = 3). Analysis of variance (ANOVA) was

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used to test the significance of treatment effects, followed by comparisons of treatment means using

198

Tukey’s test (p < 0.05). Statistical analyses were performed using IBM SPSS Statistics v. 24.

199 200

Results

201

Characterization of soils

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The 12 paddy soils were contaminated with Cd to varying degree (total Cd concentrations from

203

1.36 to 11.1 mg, Table S1). The concentrations of extractable Cd (0.1 M CaCl2) accounted for 6.0-

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68% of the total Cd in air-dried soils. Water-extractable SO42- ranged from 49.7 to 208 mg S kg-1.

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Other soils properties varied widely with soil pH ranging from 5.0 to 6.7, soil organic carbon from

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1.80 to 27.8 g kg-1. Based on clustering analysis with the basic soil physicochemical properties

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(Figure S1), three soils (GX, FJ, and HN-1) from each cluster were chosen to further examine the

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factors controlling Cd solubility in soils during reduction-oxidation processes.

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Experiment 1: dynamics of soluble Cd varies largely among 12 soils

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The 12 paddy soils were incubated anaerobically for 40 d followed by a 20 d period of oxidation.

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The changes in soil Eh, aqueous pH, Fe, and sulfate showed similar trends in all 12 soils (Figure 1

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and Figure S3), reflecting the typical processes of microbial respiration upon the depletion of O2

214

during reduction and the replacement of O2 during oxidation.12, 22, 32, 33 In brief, solution pH in acid

215

soils increased from their initial values towards near-neutral values during the reducing phase, but

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upon aeriation they gradually decreased to the values similar to (or slightly higher than) the initial

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pH values after 20 d (Figure 1a). Solution pH in alkaline soils remained stable during the entire

218

incubation (Figure 1a). Soil Eh dropped rapidly to values below 0 mV after 20 d flooding and rose

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rapidly to the initial values after 20 d of oxidation (Figure S3a). Soluble Fe and Mn concentrations

220

increased during 3 to 20 d in the reducing phase, reflecting microbially reductive dissolution of

221

Fe(III)- and Mn(III/IV)-(oxyhydr)oxides (Figure S3).44

222 223

The soluble (porewater) Cd concentrations decreased markedly to the values near to (or below) the

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detection limit (0.2 µg L-1) after 5 d flooding in all 12 soils (Figure 1b). Upon soil oxidation,

225

soluble Cd increased substantially, but there were substantial variation among the soils in the rate at

226

which soluble Cd concentration increased (Figure 1b). Interestingly, these temporal changes in

227

soluble Cd concentration were highly related with porewater pH across the 12 soils, irrespective of

228

reduction and oxidation processes (Figure 1c). A decrease in pH by one unit increased Cd solubility

229

(aqueous Cd/soil total Cd) by 4.3-fold, indicating a pH-dependent mechanism which played an

230

important and prevalent role in controlling Cd solubility in paddy soils. The temporal changes in

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soluble Cd were also correlated with soil Eh but the relationships were worse than those with soil 11 | P a g e ACS Paragon Plus Environment


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pH (Figure S4a,b). It is not surprising given that there was a high correlation between porewater pH

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and soil Eh (Figure S4c,d).

234 Reduction (a)

Oxidation

8.0

Porewater pH

7.5 7.0 6.5 XT CZ ST YS CS MQ

6.0 5.5 5.0

HK FJ-1 GX FJ HN-1 HN-2

4.5

Soluble Cd (µM)

(b)

0

10

20

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40

50

60

0

10

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30

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50

60

0.8

0.6

0.4

0.2

0.0

(c)

Time (d)

Cd solubility (aqueous Cd/soil total Cd)

0.020 y =44.5*exp(-1.77*x) R2 = 0.76, n = 60 p < 0.001

0.015

y =35.9*exp(-1.59*x) R2 = 0.75, n = 48, p < 0.001

0.010

0.005

0.000 5

235

6

7

8

7

6

5

Porewater pH

236

Figure 1. Temporal changes in porewater pH (a) and soluble Cd (b) in 12 paddy soils during 40 d

237

reduction and subsequent 20 d oxidation (Experiment 1). (c) The Cd solubility (aqueous Cd/soil

238

total Cd) as influenced by porewater pH during the reducing phase (left) and the oxidation phase

239

(right). Data are mean ± SE (n=3).

240 241

Experiment 2: Changes in Cd mobilization are not always related to sulfate reduction 12 | P a g e ACS Paragon Plus Environment

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Temporal changes in soil Eh and porewater pH showed similar trends in all three soils (GX, FJ, and

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HN-1) (Figure 2a-c), but the dynamics of soluble sulfate differed markedly (Figure 2d-f).

244

Specifically, the majority of the microbial sulfate reduction occurred between 1 to 20 d in the GX

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soil, 5-20 d in the FJ soil, and 1 to 5 d in the HN-1 soil, with complete sulfate reduction after 5 d

246

(R5) in the HN-1 soil and 20 d (R20) in both the GX and FJ soils. During the soil oxidation phase,

247

dissolved sulfate concentrations increased again, presumably due to the oxidation of sulfides.12, 22, 33

248 249

Soluble and extractable Cd decreased with ongoing soil reduction, but the rate of decrease varied

250

among the soils (Figure 2g-l). The rate of decrease in the extractable Cd was faster in the FJ and

251

HN-1 soils than in the GX soil; the former two soils took 20 d whereas the latter soil took 40 d to

252

reach the minimum value. Interestingly, the decrease in both soluble and extractable Cd (Figure 2g-

253

l) was not always coupled with the decrease in sulfate (Figure 2d-f). For Cd extractability, although

254

the periods during which extractable Cd decreased also encompassed the major periods of sulfate

255

reduction for all three soils, approximately 20% of the reduction in extractable Cd actually occurred

256

prior to the onset of sulfate reduction in the FJ soil and 50% after the completion of sulfate

257

reduction in the HN-1 and GX soils. This decoupling suggests other processes also play a role in

258

controlling Cd solubility during the reduction phase. During the oxidation phase, soluble and

259

extractable Cd increased largely in the three soils (Figure 2g-l).

260 261

Cadmium K-edge XANES was used to examine the changes in Cd speciation in the three soils

262

(Experiment 2), with this approach having not previously applied to field-contaminated soils with

263

such low Cd levels (5.7-11.1 mg kg-1). Visual inspection of the XANES spectra of the three soils

264

indicated that during the reducing phase, there was a shift from O-coordinated Cd (i.e. Cd

265

sorbed/bound to natural organic matter, clay minerals and Fe oxides) to CdS (Figure S5). After 20 d

266

oxidation (O20), the appearance of the spectra shifted back towards to those obtained at the 13 | P a g e ACS Paragon Plus Environment


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beginning of the experiment (R0) (Figure S5). The LCF analysis confirmed these observations. At

268

the beginning of soil reduction (R0), on average ca. 82% of the Cd was presented as O-coordinated

269

Cd in the three soils, with the remaining 18% presented as CdS (Table 1 and Figure S5). After 40 d

270

soil flooding (R40), the proportion of CdS increased to 55-63% in the three soils, with a

271

corresponding decrease in O-coordinated Cd to 37-56% (Table 1 and Figure S5). In the subsequent

272

oxidizing period, the proportion of CdS decreased to 28-46% after 20 d oxidation (O20), suggesting

273

a slow or incomplete oxidation dissolution of CdS, consistent with the previous studies.22, 23

274

Meanwhile, the O-coordinated Cd increased to 54-72%.

275 276

Experiment 3: Inhibition of sulfate reduction does not influence Cd mobilization in the

277

subsequent oxidation phase

278

As expected, the addition of molybdate greatly reduced the loss (reduction) of soluble sulfate,

279

which remained relatively constant across both the reduction and oxidation phases for all three soils

280

(Figure 3a-c). In the reduction phase, molybdate treatment arrested the decrease in the extractable

281

Cd concentration in the GX soil, but had a delaying effect in both the FJ and HN-1 soils (Figure 3d-

282

f). The decrease in the extractable Cd in the molybdate amended FJ and HN-1 soils could not be

283

explained by the formation of Cd-sulfides because sulfate reduction was almost completely

284

inhibited (Figure 3a-c). The XAS analysis also indicated that compared to the control, addition of

285

molybdate inhibited the formation of CdS in the reduction phase, with the percentage of CdS

286

remaining 20-28% in the GX soil and 20-34% in the HN-1 soil over the entire incubation period

287

(Table 1 and Figure S6). Notably, the complete inhibition in microbial sulfate reduction almost did

288

not influence the levels of extractable Cd during the soil oxidation phase, indicating that the

289

formation of Cd-sulfides is likely not the critical factor controlling Cd mobilization in the oxidation

290

phase.

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Environmental Science & Technology Soil: GX

Oxidation

Reduction

(a)

Oxidation

Reduction

(b)

(c)

7.0

400

pH

6.5 6.0

200

5.5

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-200

Soluble SO42- (mM)

0 1.0

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(d)

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(e)

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(f)

0.8 0.6 0.4 0.2 0.0 0 0.8

Soluble Cd (µM)

600

Eh (mV)

7.5

HN-1

FJ

Oxidation

Reduction

10

20

30

40

50

60

0

10

20

30

40

50

60

0

h)

(g)

10 (i)

0.6 0.4 0.2 0.0

Extractable Cd (mg kg-1)

0

292

8

10

20

30

40

50

60

0

(j)

10

20

30

40

50

60

0

(k)

(l)

6 4 2 0 0

10

20

30

40

Time (d)

50

60

0

10

20

30

40

50

60

0

Time (d)

Time (d)

293

Figure 2. Changes in porewater pH (open triangles) and soil Eh (solid circles) (a, b, c), aqueous

294

concentrations of sulfate (d, e, f), soluble Cd (g, h, i), and CaCl2-extractable Cd (j, k, l) in three

295

paddy soils (GX soil, left panels; FJ soil, middle; HN-1 soil, right) during 40 d reduction and

296

subsequent 20 d oxidation (Experiment 2). The periods of major sulfate reduction are marked as

297

gray shaded areas. Data are mean ± SE (n=3).

298



FJ

GX Reduction

1.4

GX + Mo

1.2

Soluble SO42- (mM)

Oxidation

FJ + Mo

1.0

1.0

0.8

0.8

0.8

0.6

0.6

0.6

0.4

0.4

0.4

0.2

0.2

0.2

0.0

10

10

20

30

40

50

60

10

GX + Mo

8

10

20

30

40

50

60

0

10

(e) FJ + Mo

8

6

6

4

4

4

2

2

2

0 0

10

20

30

40

50

60

10

20

30

40

50

60

40

50

60

(f) HN-1 + Mo

0 0

Time (d)

HN-1 + Mo

8

6

0

Oxidation

0.0 0

(d)

Reduction

(c)

1.2

1.0

0


Oxidation

1.4

(b)

1.2

0.0

299

HN-1

Reduction

1.4

(a)

Page 16 of 28

10

20

30

40

50

60

Time (d)

0

10

20

30

Time (d)

300

Figure 3. The effect of exogenous addition of molybdate (+Mo) on soluble SO42- (a, b, c) and

301

CaCl2-extractable Cd (d, e, f) in GX soil (left), FJ soil (middle), and HN-1 soil (right) (Experiment

302

3). Data are mean ± SE (n=3).

303 304

Experiment 4: Inhibition of oxidation dissolution of Cd-sulfides does not influence Cd

305

mobilization in the oxidation phase

306

Gamma irradiation was applied just before the beginning of the oxidation phase. Following gamma

307

irradiation, no visual microbial cells or colonies were observed in plate culture (Figure S7a,b). The

308

XAS analysis showed that during the oxidation phase, the proportion of Cd decreased from 55% to

309

31% in the GX soil and from 63% to 42% in the FJ soil upon the gamma irradiation treatment, but

310

there was no significant difference in the proportion of CdS between the control and gamma

311

irradiation treatment (Table 1 and Figure S8), indicating that gamma irradiation was effective in 16 | P a g e ACS Paragon Plus Environment

Page 17 of 28


312

inhibiting oxidation dissolution of Cd sulfides. Interestingly, inhibition of oxidation dissolution of

313

Cd-sulfides did not influence the mobilization of Cd in the subsequent oxidation phase (Figure

314

S7c,d), indicating that the increase in extractable Cd in the oxidation phase was not controlled by

315

the oxidation of CdS.

316 317

Experiment 5: Different species of Cd formed in reduced soil does not influence Cd mobilization

318

in the oxidation phase

319

The exogenous applications of the various Cd species (CdS, CdCO3, CdSO4) had no impact on

320

temporal changes in porewater pH (Figure S9). Interestingly, there were also no significant

321

differences in extractable Cd concentrations during the oxidation phase despite the addition of these

322

various Cd species (Figure 4a). In order to test if pH played an important role, lime (CaCO3) was

323

added to adjust soil pH from 5.18 to 6.70 before incubation. An increase in pH substantially

324

decreased the mobilization of Cd in the oxidation phase, with no significant differences observed

325

among the different treatments (Figure 4b). Again, these results indicate that the formation of

326

various Cd species during the reducing phase cannot explain the variation in Cd release during the

327

oxidation phase among soils. Rather, it is soil pH that controls the mobilization of Cd in the

328

oxidation phase.

329 330 331



Page 18 of 28

Soil pH 6.70

Soil pH 5.18 Reduction

Oxidation

5

(a)



5

Addition of various Cd species

4 3 2 1 0

(b)

+CdS +CaCO3

4

+CdSO4

3 2 1 0

0

332

Oxidation

Reduction

10

20

30

40

50

60

0

Time (d)

10

20

30

40

50

60

Time (d)

333

Figure 4. The effects of different Cd species on CaCl2-extractable Cd in the HN-2 soil under

334

different pH conditions. The initial pH of HN-2 soil was 5.18 and parts of the soil was adjusted to

335

pH 6.70 with lime before incubation. Various Cd species including CdS, CdCO3, and CdSO4, was

336

exogenously added into the soil after 15 d flooding and then the vials were closed for continuing

337

incubation. Data are mean ± SE (n=3). The legend in (b) is also applied to (a).

338 339

Experiment 6: Sequential extractions indicate Fe-Mn oxides controlling Cd mobilization

340

Experiment 6 used sequential extractions to determine changes in Cd fractions in soil phases over

341

time. The six extractions recovered 93.3 ± 2.9% of the total Cd in the three soils. For all three soils,

342

the soluble and exchangeable Cd (F1+F2) decreased rapidly in the reduction phase, with a

343

concomitant increase in the Cd fraction associated with Fe-Mn oxides (F4) (Figure 5a,b,c). In the

344

subsequent oxidation phase, exchangeable Cd (F2) increased markedly with a corresponding

345

decrease in the Fe (hydro)oxides fraction (F4). The changes in exchangeable Cd (F2) (Figure

346

5a,b,c) were in broad agreement with the results of CaCl2-extractable Cd (Figure 2j,k,l). In

347

comparison, the Cd associated with acid volatile sulfide and organic matter sulfides (F5) followed

348

the same pattern as F4, but the magnitude was much smaller than the changes in the Cd bound with

349

Fe-Mn (hydro)oxides (F4). 18 | P a g e ACS Paragon Plus Environment

Page 19 of 28


Extracted Cd (%)

(a)

(b)

GX

(c)

FJ

100

100

100

80

80

80

60

60

60

40

40

40

20

20

20

0

0

5

R

R

20

R

40

O

3

O

7

O

20

HN-1

F1 F2 F3 F4 F5 F6

0

5

R

R

20

40 R

3

O

7

O

20

O

5

R

20

R

40

R

3

O

7

O

20

O

350 351

Figure 5. Changes in solid phase Cd in three paddy soils during 40 d reduction and subsequent 20 d

352

oxidation. Data are mean ± SE (n = 3). F1: soluble Cd; F2: exchangeable Cd; F3: specifically-

353

adsorbed Cd; F4: Fe-Mn oxides-bound Cd; F5: organic matter adsorbed/sulfides; F6: residual Cd.

354 355

Discussion

356

The processes controlling the mobilization of Cd during the pre-harvest drainage phase remain

357

unclear despite their important in influencing Cd accumulation in rice grain. Consistent with

358

previous studies,16, 20, 22, 23 we found that soluble and extractable Cd changed markedly over the

359

experimental period, decreasing rapidly upon flooding, but increasing again during the oxidation

360

phase (Figures 1-4). Previous studies have emphasized the role of the sulfate reduction in

361

controlling Cd solubility in paddy soils.20, 22 In contrast, our data indicate that a pH-dependent

362

mechanism is a key factor controlling Cd mobilization, especially during the oxidation period.

363 364

Fe and Mn (oxyhydro)oxides play a critical role in controlling Cd mobilization in soils

365

For the first time, we show that the amount of Cd sorbed to the Fe-Mn (oxyhydro)oxides increased

366

in acid paddy soil during the reduction phase, with concomitant decreases in the soluble and

367

extractable Cd concentrations. When soils are flooded, soil pH increases towards a near-neutral

368

value. This provides more sorption sites in the surface of Fe-Mn (oxydro)oxides, removing Cd from 19 | P a g e ACS Paragon Plus Environment


Page 20 of 28

369

the mobile fractions and decreasing Cd extractability (i.e. become un-extractable using 0.1M

370

CaCl2). However, upon drainage and oxidation, there was a decrease in the proportion of Cd sorbed

371

to Fe-Mn (oxyhydro)oxides and a concomitant increase in soluble and extractable Cd (Figure S10).

372

These changes were brought about primarily by the changes in soil pH during the reduction-

373

oxidation cycle. This mechanism is evidenced by the following findings in the present study. First,

374

the decrease in both the soluble Cd and the CaCl2-extractable Cd during the reduction period was

375

not directly coupled with the major sulfate reduction (Figure 2). Rather, the immobilization of Cd

376

(either before or after the major sulfate reduction) could be explained by i) the enhanced adsorption

377

on the surface of Fe-Mn (oxyhydro)oxides or ii) the co-precipitation of the new formation of

378

‘secondary’ Fe (hydro)oxides from the anaerobic oxidation of Fe(II).34 The sequential extraction

379

results revealed that the specifically-adsorbed Cd (F3) did not increase with increase in soil pH,

380

thereby suggesting the latter possibility, although further studies are required in this regard. In

381

addition, the addition of Mo prevented sulfate reduction, but the decreases in extractable-Cd were

382

still observed after 20 d of reduction (R20) for both the FJ and HN-1 soils (Figure 3), thereby

383

supporting the importance of the adsorption/co-precipitation mechanism apart from sulfate

384

reduction. Second, the sequential extraction showed that changes in Cd mobilization were largely

385

dependent upon the transformation of Cd between the Fe-Mn (oxyhydro)oxide fraction and the

386

exchangeable fraction during both the soil flooding and oxidation phases (Figure 5). Third, in situ

387

analyses of Cd speciation using XAS revealed that a large proportion of Cd was present as O-

388

coordinated Cd (i.e. sorbed Cd) (37-82%), with the remaining Cd being CdS (18-63%) during soil

389

flooding and oxidation (Table 1 and Figure S5). Gamma irradiation and +Mo treatment had no

390

significant effect on Cd speciation (and its associated mobilization) during the oxidation phase

391

(Table 1, Figure 3, Figures S6 and S8), again confirming that the mobilization of Cd during the

392

oxidation phase is likely due to the exchange between strongly-bound/co-precipitated Cd and

393

mobile Cd, rather than the oxidation of Cd sulfides. It should be noted that the ‘O-coordinated Cd’ 20 | P a g e ACS Paragon Plus Environment

Page 21 of 28


394

identified in the XAS analyses actually included the Cd associated with fractions 2 to 4 in

395

sequential extractions. This may be the reason why the large changes in Fe-Mn oxide-bound Cd

396

(F4) were not observed in O-coordinated Cd in the XAS analysis. Fourth, the temporal changes in

397

Cd solubility among the 12 paddy soils varying widely in soil properties were mainly influenced by

398

pH, irrespective of reduction and oxidation processes (Figure 1c), strongly indicating the

399

importance of a pH-dependent mechanism for controlling Cd solubility in paddy soil systems.

400 401

The role of sulfate reduction

402

Sulfate reduction has been reported to be an important process controlling the dynamics of Cd

403

solubility in reduced soils.16, 22 In the present study, addition of molybdate almost completely

404

inhibited sulfate reduction during the reduction phase, preventing the further formation of Cd-

405

sulfides (Table 1 and Figure S6). Interestingly, inhibition of sulfate reduction slowed the decrease

406

in extractable Cd in soils during the reduction period (Figure 3). However, inhibition of oxidation

407

dissolution of Cd-sulfides during the oxidation phase through gamma irradiation had no impact on

408

the mobilization of Cd during the oxidation phase (Figure S7c,d). These results demonstrate that

409

sulfate reduction played an important role in immobilizing Cd in reduced soils. However, it has

410

little or no impact on the mobilization of Cd upon soil drainage. These results could explain why

411

additions of sulfate were not always effective in reducing grain Cd,2, 35, 36 given that the majority of

412

Cd accumulation in rice grain occurs during soil drainage prior to harvest (i.e. soil oxidation phase).

413 414

Environmental implications

415

The role of Fe (oxyhydro)oxides or ‘secondary’ Fe(III) (hydro)oxides in immobilizing Cd has only

416

been tested previously in groundwater systems with bacteriogenic iron oxides.37 In the present

417

study, we provide evidence for the first time confirming the role of Fe-Mn (oxyhydro)oxides in Cd-

418

contaminated paddy soils, with the model presented in Figure S10. It has been reported that the 21 | P a g e ACS Paragon Plus Environment


Page 22 of 28

419

majority (>80%) of the Cd in rice grain accumulates during the grain-filling phase of rice growth,

420

generally coinciding with pre-harvest soil drainage. This increase in Cd accumulation during grain

421

filling stage is due to an increased Cd mobilization during soil drainage, which we attributed here to

422

the decrease in soil pH upon soil oxidation (Figure S10a). Therefore, as an approach to decrease Cd

423

accumulation, it may be feasible to increase pH during this oxidation (drainage) phase to increase

424

the sorption/co-precipitation of Cd by Fe-Mn (oxyhydro)oxides and reduce Cd mobilization in soil,

425

providing a strategy to reduce the mobilization of Cd into mobile fractions. Agronomic practices

426

that are able to increase soil pH, such as applications of lime or applications of alkaline biochar,

427

have been shown to be effective in immobilization of Cd and subsequent accumulation in rice

428

grain.38, 39 On the other hand, the applications of sulfate-based practices have contradictory effects

429

on reducing Cd solubility in soil and subsequent Cd accumulation in rice grain.2, 35, 36 The addition

430

of sulfate may be important in the soils in which the amounts of chalcophilic elements significantly

431

exceed reducible sulfate. In these soils, increasing sulfate is able to facilitate the formation of Cd

432

sulfides in reduced soils.22 Finally, enhancing the retention of Cd by Fe-Mn (oxyhydro)oxides such

433

as by liming the soil to increase pH provides a potential strategy to mitigate the transfer of Cd from

434

soil to humans through the consumption of rice.

435 436

ASSOCIATED CONTENT

437

Supporting Information

438

The Supporting Information is available free of charge on the ACS Publications website.

439

Details regarding soil characterization, microcosm experimental method, Cd speciation analysis and

440

results using XAS, physicochemical properties of the soils tested, target transformation analysis of

441

Cd standard compounds, results of LCF analysis of Cd XANES, temporal changes in soil Eh and

442

porewater soluble ions in 12 paddy soils, the effect of gamma irradiation on soil microbes and Cd


Page 23 of 28


443

extractability/speciation in the GX and FJ soils, and the changes in porewater pH as affected by

444

additions of various Cd species in the HN-2 soil.

445 446

AUTHOR INFORMATION

447

Corresponding Author

448

*Email: [email protected]; tel: +86 (0)25 843909562; fax: +86 (0)25 843909562.

449

Notes

450

The authors declare no competing financial interest.

451 452

Acknowledgements

453

The study was supported by the Natural Science Foundation of China (grant No. 41671309,

454

21661132001), the National Key Research and Development Program of China

455

(2016YFD08004002), the Special Fund for Agro-Scientific Research in the Public Interest (grant

456

no. 201403015), the Natural Science Fund for Jiangsu Distinguished Young Scholar

457

(BK20180025), the fundamental research funds for the Central Universities (Grant

458

No.KJJQ201902), and the Innovative Research Team Development Plan of the Ministry of

459

Education of China (Grant No. IRT_17R56). Parts of this research were undertaken on the XAS

460

beamline at the Australian Synchrotron (AS181/XAS/12932), part of ANSTO (Australian Nuclear

461

Science and Technology Organization).



Page 24 of 28

References 1.

Clemens, S.; Aarts, M. G. M.; Thomine, S.; Verbruggen, N., Plant science: the key to preventing slow cadmium poisoning. Trends Plant Sci 2013, 18, (2), 92-99.

2.

McLaughlin, M. J.; Whatmuff, M.; Warne, M.; Heemsbergen, D.; Barry, G.; Bell, M.; Nash, D.; Pritchard, D., A field investigation of solubility and food chain accumulation of biosolidcadmium across diverse soil types. Environmental Chemistry 2006, 3, (6), 428-432.

3.

Satarug, S.; Garrett, S. H.; Sens, M. A.; Sens, D. A., Cadmium, environmental exposure, and health outcomes. Environmental Health Perspectives 2010, 118, (2), 182-190.

4.

Kunze, G.; Song, Y.; Wang, Y.; Mao, W.; Sui, H.; Yong, L.; Yang, D.; Jiang, D.; Zhang, L.; Gong, Y., Dietary cadmium exposure assessment among the Chinese population. Plos One 2017, 12, (5), e0177978.

5.

Sui, F.-Q.; Chang, J.-D.; Tang, Z.; Liu, W.-J.; Huang, X.-Y.; Zhao, F.-J., Nramp5 expression and functionality likely explain higher cadmium uptake in rice than in wheat and maize. Plant Soil 2018.

6.

Watanabe, T.; Zhang, Z.-W.; Moon, C.-S.; Shimbo, S.; Nakatsuka, H.; Matsuda-Inoguchi, N.; Higashikawa, K.; Ikeda, M., Cadmium exposure of women in general populations in Japan during 1991–1997 compared with 1977–1981. Int. Arch. Occup. Environ. Health 2000, 73, (1), 26-34.

7.

Chen, H.; Yang, X.; Wang, P.; Wang, Z.; Li, M.; Zhao, F. J., Dietary cadmium intake from rice and vegetables and potential health risk: A case study in Xiangtan, southern China. Sci Total Environ 2018, 639, 271-277.

8.

Nordberg, G.; Jin, T.; Bernard, A.; Fierens, S.; Buchet, J. P.; Ye, T.; Kong, Q.; Wang, H., Low bone density and renal dysfunction following environmental cadmium exposure in China. Ambio 2002, 31, (6), 478-81.

9.

Jin, T.; Nordberg, G.; Ye, T.; Bo, M.; Wang, H.; Zhu, G.; Kong, Q.; Bernard, A., Osteoporosis and renal dysfunction in a general population exposed to cadmium in China. Environ. Res. 2004, 96, (3), 353-359.

10. Zhang, W.-L.; Du, Y.; Zhai, M.-M.; Shang, Q., Cadmium exposure and its health effects: A 19-year follow-up study of a polluted area in China. Sci Total Environ 2014, 470–471, 224228. 24 | P a g e ACS Paragon Plus Environment

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11. Nordberg, G. F.; Nogawa, K.; Nordberg, M.; Friberg, L. T., Chapter 23 - Cadmium. In Handbook on the Toxicology of Metals (Third Edition), Academic Press: Burlington, 2007; pp 445-486. 12. Kögel-Knabner, I.; Amelung, W.; Cao, Z.; Fiedler, S.; Frenzel, P.; Jahn, R.; Kalbitz, K.; Kölbl, A.; Schloter, M., Biogeochemistry of paddy soils. Geoderma 2010, 157, (1), 1-14. 13. Arao, T.; Kawasaki, A.; Baba, K.; Mori, S.; Matsumoto, S., Effects of water management on cadmium and arsenic accumulation and dimethylarsinic acid concentrations in Japanese rice. Environ Sci Technol 2009, 43, (24), 9361-7. 14. Inahara, M.; Ogawa, Y.; Azuma, H., Countermeasure by means of flooding in latter growth stage to restrain cadmium uptake by lowland rice. Japanese Journal of Soil Science and Plant Nutrition 2007, 78, 149-155. 15. Cornu, J. Y.; Denaix, L.; Schneider, A.; Pellerin, S., Temporal evolution of redox processes and free Cd dynamics in a metal-contaminated soil after rewetting. Chemosphere 2007, 70, (2), 306-14. 16. de Livera, J.; McLaughlin, M. J.; Hettiarachchi, G. M.; Kirby, J. K.; Beak, D. G., Cadmium solubility in paddy soils: Effects of soil oxidation, metal sulfides and competitive ions. Sci Total Environ 2011, 409, (8), 1489-1497. 17. Huang, J.-H.; Wang, S.-L.; Lin, J.-H.; Chen, Y.-M.; Wang, M.-K., Dynamics of cadmium concentration in contaminated rice paddy soils with submerging time. Paddy and Water Environment 2013, 11, (1-4), 483-491. 18. Khaokaew, S.; Chaney, R. L.; Landrot, G.; Ginder-Vogel, M.; Sparks, D. L., Speciation and release kinetics of cadmium in an alkaline paddy soil under various flooding periods and draining conditions. Environ Sci Technol 2011, 45, (10), 4249-4255. 19. Hashimoto, Y.; Furuya, M.; Yamaguchi, N.; Makino, T., Zerovalent iron with high sulfur content enhances the formation of cadmium sulfide in reduced paddy soils. Soil Sci. Soc. Am. J. 2016, 80, 55-63. 20. Hashimoto, Y.; Kanke, Y., Redox changes in speciation and solubility of arsenic in paddy soils as affected by sulfur concentrations. Environ. Pollut. 2018, 238, 617-623.



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21. Hashimoto, Y.; Yamaguchi, N., Chemical speciation of cadmium and sulfur K-Edge XANES spectroscopy in flooded paddy soils amended with zerovalent iron. Soil Sci. Soc. Am. J. 2013, 77, (4), 1189-1198. 22. Fulda, B.; Voegelin, A.; Kretzschmar, R., Redox-controlled changes in cadmium solubility and solid-phase speciation in a paddy soil as affected by reducible sulfate and copper. Environ Sci Technol 2013, 47, (22), 12775-12783. 23. Furuya, M.; Hashimoto, Y.; Yamaguchi, N., Time-course changes in speciation and solubility of cadmium in reduced and oxidized paddy soils. Soil Sci. Soc. Am. J. 2016, 80, (4), 870-877. 24. Hu, P.; Huang, J.; Ouyang, Y.; Wu, L.; Song, J.; Wang, S.; Li, Z.; Han, C.; Zhou, L.; Huang, Y.; Luo, Y.; Christie, P., Water management affects arsenic and cadmium accumulation in different rice cultivars. Environ. Geochem. Health 2013, 35, (6), 767-778. 25. Bingham, F. T.; Page, A. L.; Mahler, R. J.; Ganje, T. J., Cadmium availability to rice in sludge-amended soil under “flood” and “nonflood” culture. Soil Sci. Soc. Am. J. 1976, 40, (5), 715-719. 26. Liu, F. Z., Practical Handbook of soil monitoring and analysis Chemical Industry Press: 2012. 27. McGrath, S. P.; Cunliffe, C. H., A simplified method for the extraction of the metals Fe, Zn, Cu, Ni, Cd, Pb, Cr, Co and Mn from soils and sewage sludges. Journal of the Science of Food and Agriculture 1985, 36, (9), 794-798. 28. McGrath, S. P.; Cegarra, J., Chemical extractability of heavy metals during and after longterm applications of sewage sludge to soil. J. Soil Sci. 1992, 43, (2), 313-321. 29. Schippers, A., Biogeochemistry of metal sulfide oxidation in mining environments, sediments, and soils. In Sulfur Biogeochemistry - Past and Present, Amend, J. P.; Edwards, K. J.; Lyons, T. W., Eds. Geological Society of America: 2004. 30. Carlsson, E.; Thunberg, J.; Öhlander, B.; Holmström, H., Sequential extraction of sulfiderich tailings remediated by the application of till cover, Kristineberg mine, northern Sweden. Sci Total Environ 2002, 299, (1), 207-226. 31. Malinowski, E. R., Factor analysis in chemistry. John Wiley: New York, 1991. 26 | P a g e ACS Paragon Plus Environment

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32. Borch, T.; Kretzschmar, R.; Kappler, A.; Cappellen, P. V.; Ginder-Vogel, M.; Voegelin, A.; Campbell, K., Biogeochemical redox processes and their impact on contaminant dynamics. Environ Sci Technol 2009, 44, (1), 15-23. 33. Kirk, G., Reduction and Oxidation. In The Biogeochemistry of Submerged Soils, Kirk, G., Ed. John Wiley & Sons, Ltd: Chichester, 2004. 34. Yu, H. Y.; Li, F. B.; Liu, C. S.; Huang, W.; Liu, T. X.; Yu, W. M., Chapter Five - Iron redox cycling coupled to transformation and immobilization of heavy metals: Implications for paddy rice safety in the red soil of south china. In Advances in Agronomy, Sparks, D. L., Ed. Academic Press: 2016; Vol. 137, pp 279-317. 35. Hassan, M. J.; Wang, F.; Ali, S.; Zhang, G., Toxic Effect of Cadmium on Rice as Affected by Nitrogen Fertilizer Form. Plant Soil 2005, 277, (1), 359-365. 36. Sarwar, N.; Malhi, S. S.; Zia, M. H.; Naeem, A.; Bibi, S.; Farid, G., Role of mineral nutrition in minimizing cadmium accumulation by plants. J. Sci. Food Agric. 2010, 90, (6), 925-37. 37. Martinez, R. E.; Pedersen, K.; Ferris, F. G., Cadmium complexation by bacteriogenic iron oxides from a subterranean environment. Journal of Colloid and Interface Science 2004, 275, (1), 82-89. 38. Chen, H.; Zhang, W.; Yang, X.; Wang, P.; McGrath, S. P.; Zhao, F.-J., Effective methods to reduce cadmium accumulation in rice grain. Chemosphere 2018, 207, 699-707. 39. Zhu, H. H.; Chen, C.; Xu, C.; Zhu, Q. H.; Huang, D. Y., Effects of soil acidification and liming on the phytoavailability of cadmium in paddy soils of central subtropical China. Environ. Pollut. 2016, 219, 99-106.



Page 28 of 28

Table 1. Results of the linear combination fitting of the Cd K-edge XANES spectra of three paddy soils under different treatments before being incubated for various periods. GX CdS 19 (1.0) 55 (0.6) 55 (0.6) 43 (0.8) 31 (0.9)

GX + Mo

R0 R20 R40 O7 O20

O-coordinated Cd 81 (1.0)a 45 (0.6) 45 (0.6) 57 (0.8) 69 (0.9)

R0 R20 R40 O7 O20

82 (0.9) 58 (0.9) 56 (0.6) 66 (0.8) 72 (0.8)

HN-1 18 (0.9) 42 (0.9) 44 (0.6) 34 (0.8) 28 (0.8)

0.0003 0.0004 0.0002 0.0003 0.0002

R0 R20 R40 O7 O20

82 (0.7) 46 (0.5) 37 (2.0) 41 (0.8) 54 (0.7)

FJ 18 (0.7) 54 (0.5) 63 (2.0) 59 (0.8) 46 (0.7)

0.0002 0.0001 0.0009 0.0001 0.0001

GX + Gamma sterilization

R-factorb 0.0004 0.0001 0.0002 0.0003 0.0003

O-coordinated Cd 80 (0.7)

CdS 20 (0.7)

R-factor 0.0002

72 (0.9) 72 (0.8) 76 (1.1)

28 (0.9) 28 (0.8) 24 (1.1)

0.0004 0.0003 0.0005

77 (0.7) 80 (0.9) 66 (0.7) 63 (0.5) 73 (0.6)

HN-1 + Mo 23 (0.7) 20 (0.9) 34 (0.7) 37 (0.5) 27 (0.6)

0.0002 0.0003 0.0001 0.0001 0.0001

O-coordinated Cd

CdS

R-factor

56 (0.7) 69 (1.2)

44 (0.7) 31 (1.2)

0.0002 0.0005

FJ + Gamma sterilization

41 (0.5) 58 (0.7)

59 (0.5) 42 (0.7)

0.0001 0.0002

a The values in parentheses show the percentage variation in the calculated values. b The goodness of fit is indicated by the R factor. R factor = Σi(experimental − fit)2/Σi(experimental)2, where the sums are over the data points in the fitting region.


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