Water Decontamination from Cr(III) - ACS Publications

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Water decontamination from Cr(III)-organic complexes based on Pyrite/H2O2: Performance, mechanism, and validation Ye Yu-xuan, Chao Shan, Xiaolin Zhang, Hui Liu, Dandan Wang, Lu Lv, and Bing-Cai Pan Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.8b01693 • Publication Date (Web): 22 Aug 2018 Downloaded from http://pubs.acs.org on August 25, 2018

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Environmental Science & Technology

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Water Decontamination from Cr(III)-Organic Complexes Based on Pyrite/H2O2:

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Performance, Mechanism, and Validation

3 4

Yuxuan Ye†,﹟, Chao Shan† ‡,﹟, Xiaolin Zhang† ‡, Hui Liu†, Dandan Wang†, Lu Lv†‡,

5

Bingcai Pan † ‡*

6 7

†

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Environment, Nanjing University, Nanjing 210023, China

9

‡

10

State Key Laboratory of Pollution Control and Resource Reuse, School of the

Research Center for Environmental Nanotechnology (ReCENT), Nanjing University,

Nanjing 210023, China

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﹟

Both authors contributed to this work equally

13 14 15

* To whom correspondence should be addressed E-mail: [email protected] (B.C.P)

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ACS Paragon Plus Environment


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Abstract

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Fenton reaction is a widely used pretreatment technology to degrade toxic metal-

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organic complexes. However, its efficiency is greatly compromised for Cr(III)-organic

19

complexes due to accumulation of more toxic Cr(VI) and pH dependence. Herein, we

20

proposed a combined pyrite/H2O2-precipitation process to efficiently remove Cr(III)

21

(initially at 10.4 mg Cr/L) complexed by various ligands (citrate, EDTA, oxalate, and

22

tartrate). Negligible Cr(VI) and 0.5 mg/L Cr(VI)

24

and >5 mg/L Cr remained after treatment by the ZVI/H2O2-precipitaion process at

25

pH0>5. As for the mechanisms, pyrite/H2O2 produced a considerable amount of

26

aqueous Fe(II) to initiate Fenton reaction, concurrently releasing massive H+ to keep

27

the reaction pH at ~3.0 irrespective of the initial pHs. The generated ·OH radicals

28

oxidized Cr(III) into Cr(VI) and thereby releasing the organic ligands for further

29

mineralization. The generated Cr(VI) was in situ reduced back to Cr(III) by aqueous

30

Fe(II) and FeS2. Subsequently, all the free metal ions including Cr(III), Fe(III), and

31

Fe(II) were removed via precipitation. Kinetic modeling of the pyrite/H2O2 process

32

involving 17 reactions was performed to verify the proposed mechanism. Additionally,

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the effectiveness of the combined process was further validated by its satisfactory

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performance in treating authentic tannery wastewater.

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Table of Contents (TOC) Graphic

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40

INTRODUCTION

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Chromium(III) compounds are widely present in effluents from a variety of industrial

42

processes such as leather tanning, electroplating and dyes manufacture.1 Generally,

43

Cr(III) are less toxic than Cr(VI), however, Cr(III) tends to be transformed into Cr(VI)

44

in the presence of oxidizing agents such as chlorine,2 manganese oxides3 and H2O2.4

45

Thus, stringent regulation has been adopted on the maximum allowable concentration

46

level (MCL) of total Cr and Cr(VI) as well. For instance, China EPA set the MCL values

47

of the total Cr and Cr(VI) as 1.5 mg/L and 0.1 mg/L respectively in discharged tanning

48

effluents.5 In Europe, it is suggested that the total Cr should not be detected in surface

49

water.6 US EPA has also lowered the MCL value of the total Cr in drinking water to 0.1

50

mg/L recently.7

51

In real industrial effluents and/or natural water, Cr(III) tends to complex with

52

organic ligands such as natural organic matters (NOMs), citrate, oxalate, tartrate and

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EDTA. 8-10 Efficient removal of Cr(III) complexes from contaminated water, however,

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still remains a major challenge. Traditional methods such as chemical precipitation,11

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flocculation/coagulation 12-14 and adsorption 15 exhibit satisfactory removal toward free

56

Cr(III) ions. However, it is not the case for the Cr(III)-organic complexes.16, 17 For

57

instance, after precipitation of a tannery effluent, 10-20 mg/L Cr(III) was still detected

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in the filtrate, mainly in the form of Cr(III)-carboxyl complexes.5 Similarly, it is rather

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difficult to dissociate and remove Cr(III)-azo dye complexes from water via the

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precipitation-flocculation process.18

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In the past decades, extensive studies have been devoted to utilizing advanced

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oxidation processes (AOPs) to destroy heavy metal-organic complexes,19-21 followed

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by precipitation, coagulation, and/or adsorption to further remove the released metal

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ions.22-24 Unlike other heavy metal complexes, possible generation and accumulation

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of more toxic Cr(VI) species intensively compromise the utilization of AOPs for

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eliminating Cr(III) complexes.6, 25, 26 For instance, considerable amount (13-83%) of

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Cr(III)-oxalic and Cr(III)-malonic complexes were eventually transformed into Cr(VI)

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during electrochemical oxidation processes.25 Consequently, AOP-based destruction of

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Cr(III)-organic complexes would usually encounter the dilemma between the oxidative

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decomplexation and the control of Cr(VI) during the oxidation. On the other hand, the

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complexation/decomplexation rates of most Cr(III) complexes are extremely slow due

72

to kinetic inertness of water exchange (e.g., t1/2 = 81.6 h at 298.15 K for Cr(H2O)63+).27,

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28

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not reach equilibrium even after 211 days at pH 2.3.8 Thus, the techniques based on

75

thermodynamic decomplexation, e.g., a recently developed process combining Fe(III)

76

displacement, UV irradiation and alkaline precipitation for the removal of Cu(II)

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organic complexes,24 are kinetically unacceptable when used for eliminating Cr(III)

78

complexes.

For instance, the complexation of natural organic matters (NOMs) with Cr(III) did

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Very recently, we proposed a new UV/Fe(III)+OH process involving ·OH radical

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oxidation and Fe(II) reduction concurrently for efficient removal of Cr(III)

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complexes.29 In brief, UV/Fe(III) generates ·OH radicals to transform Cr(III) into

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Cr(VI), releasing the organic ligand to form Fe(III)-organic complexes simultaneously.

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Then, the photolysis of Fe(III)-organic under UV irradiation produces massive Fe(II)

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species via the ligands to metals charge transfer (LMCT) procedure, in turn in situ

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reducing Cr(VI) back to Cr(III). The free Cr(III) ions are then removed by subsequent

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precipitation. By using this method, Cr(III) complexes with citrate, EDTA, tartrate,

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oxalate and acetate can be steadily eliminated from water. Moreover, it is demonstrated

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that this combined process was applicable to real tanning effluents, resulting in the

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residual total Cr below 1.5 mg/L and negligible accumulation of Cr(VI).29 Nevertheless,

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the scaled-up applicability of UV/Fe(III)+OH process needs further improvement due

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to the energy intensive characteristics of UV irradiation.30, 31 Besides, UV/Fe(III)+OH

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process is highly pH-dependent, and it would result in a poor Cr removal unless the

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initial pH fell in a narrow range 2.5-3.0, where aqueous Fe(III) are mainly present as

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highly photo-reactive FeOH2+ to generate ·OH radicals.32-35

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As inspired by UV/Fe(III)+OH process, we learned that effective removal of

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Cr(III)-organic complexes requires two distinct stages, i.e., oxidation of Cr(III) into

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Cr(VI) to release the organic ligands for AOPs, and subsequent reduction of the formed

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Cr(VI) into free Cr(III) ions to avoid the accumulation of toxic Cr(VI). Herein, we

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proposed a pyrite/H2O2 system followed by chemical precipitation to respond to Cr(III)

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complexes in a wide pH range. Pyrite (FeS2) is among the most abundant iron-

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containing minerals in the earth’s crust.36-39 It could adjust the pH of most test solutions

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from 3-11 to ~3 based on the proton-released reactions (eqs. 1-3),37, 39 rendering it

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capable of activating H2O2 to generate ·OH radicals. For instance, pyrite/H2O2 exhibited

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similar degradation efficiency toward nitrobenzene when varying initial pH from 3.0 to

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11.0.40 Similarly, pyrite was highly efficient to catalyze H2O2 decomposition for

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alachlor degradation at a wide range of initial pH (3.2−9.2).41 More importantly, pyrite

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may release considerable amount of Fe(II) into water due to the reactions of FeS2 with

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O2 or Fe(III) (eqs.1 and 3),42, 43 possibly enabling reduction of the generated Cr(VI).

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2FeS2 + 7O2 + 2H2O → 2Fe2++4SO42- + 4H+

(1)

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2FeS2 + 15H2O2 →2Fe3++4SO42- + 2H+ + 14H2O

(2)

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FeS 2 +14Fe3+ +8H 2 O→15Fe2+ +2SO4 2-+16H+

(3)

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Thus, the main objective of this study is to propose a combined pyrite/H2O2-

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precipitation (pyrite/H2O2+OH) process to eliminate Cr(III)-organic complexes and

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examine its applicability under varying conditions, where citrate was employed as a

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model ligand. The classical Fenton-precipitation (Fe2+/H2O2+OH) and heterogeneous

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Fenton using zero valent iron as the Fe2+ source (ZVI/H2O2+OH) were employed for

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comparison in terms of removal efficiency, pH dependence and Cr(VI) accumulation.

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Evolution of Cr(VI) of the pyrite/H2O2 process was particularly concerned, and the

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underlying mechanism was also probed. Furthermore, the application of pyrite/H2O2

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was extended to other Cr(III) complexes with a variety of organic ligands and an

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authentic tannery effluent as well.

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EXPERIMENTAL SECTION

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Chemicals and Materials. Pyrites used in this study were purchased from Sigma-

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Aldrich (325 mesh powder, particle size 90% Cr removal was achieved in 30 min, and less than 0.5 mg/L residual

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Cr was detected in 60 min. Such preliminary results suggested the advantage of

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pyrite/H2O2 in decomposing Cr(III)-organic complexes over other processes. As

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observed in Figure S1b, both pyrite/H2O2 and Fe2+/H2O2 resulted in ~37% removal of

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TOC, implying their efficient decomposition of organic ligands. The formation of

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dissolved Fe(II) and total Fe in pyrite/H2O2 system were also monitored in Figure S1c.

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Obviously, the amount of Fe(III) ion increased rapidly in the first 10 min following the

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reaction in eq. 2. After 30 min, the amount of total Fe increased very slowly, while the

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Fe(II) concentration ascended steadily, suggesting the transformation of Fe(III) into

178

Fe(II) (eq. 3).

179

Effect of the pyrite and H2O2 dosage on the degradation of Cr(III)-citrate by

180

pyrite/H2O2 system were investigated in Figures S2a and S3a, based on which the

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optimal dosage of pyrite and H2O2 were determined as 4 g/L and 20 mM respectively

182

in the subsequent experiments except for otherwise specified. As shown in Figure S2b,

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the dissolved total Fe was roughly linearly correlated with the pyrite dosage, and the

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dissolved Fe(II) increased with the increasing pyrite dosage. It was not beyond

185

expectation since pyrite is the only source of the dissolved Fe(II) and Fe(III). Figure

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S3b depicted that approximately 100% decomposition of H2O2 occurred even at the

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dosage of 50 mM, which could be well explained by the reaction in eq. 2.

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Reactive species responsible for Cr(III)-citrate degradation.

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the ESR spectra of pyrite/H2O2, pyrite and pure water (blank). A weak signal of ·OH

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radical was observed in the pyrite system. The reactions among oxygen, water and

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pyrite would generate a small amount of H2O237 and then produce ·OH radicals (eq 4).

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The addition of H2O2 dramatically increased the signal intensity of ·OH radicals mainly

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arising from two aspects, i.e., the reaction between H2O2 and FeS2 (eq. 2) facilitated

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Figure S4a depicted

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Fe(III) dissolving into water, thereby favoring aqueous Fe(II) formation (eq. 3), and the

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reaction between aqueous Fe(II) and H2O2 (eq. 4) generated massive ·OH radicals. To

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further verify the role of ·OH radicals in Cr(III)-citrate degradation, methanol at

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different levels (0.10 M and 1.0 M) was added into pyrite/H2O2 system to scavenge

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the ·OH radicals. As suggested in Figure S4b, the presence of 0.10 M methanol

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intensively suppressed the degradation of Cr(III)-citrate, while Cr(III)-citrate

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degradation was almost inhibited completely by 1.0 M methanol. Also, the formation

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of Cr(VI) was described in Figure S4c, suggesting that methanol inhibited the

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transformation of Cr(III) into Cr(VI) evidently. According to our previous study,29 the

203

oxidation of Cr(III) into Cr(VI) is essential to the release of organic ligands and thereby

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the removal of Cr(III)-organic complexes. The above results demonstrated the key role

205

of ·OH radicals in both Cr(VI) formation and Cr(III)-citrate degradation. Besides,

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Figure S4b suggested that N2 atmosphere exhibited negligible effect on Cr(III)-citrate

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removal, implying that the amount of ·OH radicals generated by the reactions among

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oxygen, pyrite and water could be ignored, which was also demonstrated by the results

209

in Figure S4a. The subsequent decline in Cr(VI) concentration depicted in Figure S4c

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was mainly attributed to the reduction by Fe(II), as discussed below.

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Fe2+ + H2 O2 → Fe3+ + OH - + . OH

(4)

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Effect of initial pH on Cr(III)-citrate removal. It is widely known that classical

213

Fenton is highly pH dependent, with the efficiency dropping down sharply at pH above

214

4.0.48 Herein, effect of initial pH on Cr(III)-citrate removal by pyrite/H2O2+OH system

215

was evaluated, and an extensively explored heterogeneous Fenton, i.e., zero-valent iron

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ZVI/H2O2+OH system was also performed for comparison (Figure 1a). Attractively, the

217

degradation of Cr(III)-citrate by pyrite/H2O2 remained constant at initial pH range 3.0-

218

9.0, and >95% removal of Cr(III)-citrate achieved in 60 min. Conversely, at the initial

219

pH of 5.0 or higher, the removal of total Cr by ZVI/H2O2+OH was inhibited completely.

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ESR spectra of pyrite/H2O2 and ZVI/H2O2 at different pHs were compared in Figure

221

1b. The strong signals of ·OH radicals were observed in pyrite/H2O2 system at all the

222

tested pHs and in ZVI/H2O2 system at pH=3.0, whereas no signals were detected in

223

ZVI/H2O2 system at pH=5.0 and 9.0. In Fenton system, ·OH radicals were generated

224

by the reaction between Fe(II) and H2O2 at acidic pHs (eq. 4). A relatively high pH

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(e.g., >5.0) suppressed the generation of ·OH radicals by ZVI/H2O2 partially because

226

negligible Fe(II) was released from ZVI (Figure 1c), whereas a considerable amount of

227

Fe(II) (~3 mg/L) was still detected in pyrite/H2O2 system even at pH=9.0 (Figure 1c).

228

As shown in Figure 1d, the pHs of pyrite/H2O2 system declined to ~3.0 rapidly in 10

229

min irrespective of initial pH from 5.0 to 9.0, thereby facilitating the dissolution of Fe(II)

230

and the generation of ·OH radicals. Such pH variation mainly resulted from the

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reactions among FeS2, H2O2 and Fe(III), which could produce massive protons

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following eqs. 2 and 3. Considering that Cr(III)-organic complexes mainly retained in

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the suspension after precipitation treatment,5, 16 the pH independence of pyrite/H2O2

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made it very attractive in water decontamination of Cr(III) complexes.

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Cr(VI) elimination. As mentioned above, transformation of Cr(III) into Cr(VI) is

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essential to the release of organic ligands.6, 26 However, the accumulation of more toxic

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Cr(VI) would raise higher environmental concern and suppress the final removal of Cr

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in the subsequent precipitation. Herein, the variation of Cr(VI) was particularly

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monitored in pyrite/H2O2 and ZVI/H2O2 systems. As shown in Figure 2a, the Cr(VI)

240

concentration in pyrite/H2O2 increased to 0.5-1.0 mg/L first and then dropped down to

241

nearly zero at pH 3.0-9.0. As for ZVI/H2O2 system, it exhibited similar tendency to

242

pyrite/H2O2 at pH 3.0, i.e., first increased and then reduced, whereas it increased

243

constantly at pH 5.0 and 9.0. The drop of Cr(VI) concentration in pyrite/H2O2 system

244

was accompanied with increased Fe(II) concentration (Figure 1c), implying that

245

aqueous Fe(II) contributed to Cr(VI) reduction. As shown in eq. 3, Fe(III) could be

246

reduced by FeS2 to continuously generate Fe(II), which followed the pseudo-zero order

247

model with rate constant at 4.110-8 M . s-1 (Figure S5) and concurred with the previous

248

study. 42

249

The apparent pseudo-second order rate constant for reduction of Cr(VI) by

250

aqueous Fe(II) was reported to be 0.24 M-1 s-1 at pH 3.0. 49 Hence, the pseudo-first order

251

rate constant for Cr(VI) reduction by aqueous Fe(II) at low concentration (~0.2 mM)

252

was supposed to be quite low (4.810-5 s-1), which could not fully explain the observed

253

rapid elimination of Cr(VI) during the reaction period of 30-60 min (1.210-3 s-1).

254

Therefore, the reduction of Cr(VI) by FeS2 also played an indispensable role in Cr(VI)

255

elimination. The pseudo-first order rate constant of Cr(VI) reduction by FeS2 was

256

determined to be 3.810-3 s-1 (Figure S6), consistent with a previous study. 50

257

As seen in Figure 2b, H2O2 was decomposed rapidly and completely in the

258

pyrite/H2O2 system at all the tested pHs or in ZVI/H2O2 system at pH=3.0 with apparent

259

pseudo-second rate constant of 6.810-3 M-1 . s-1 (Figure S7). The sustained increase of

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260

Cr(VI) concentration in ZVI/H2O2 system at pH 5.0 and 9.0 was mainly attributed to

261

the oxidation of Cr(III) by the undecomposed H2O2,4, 47 as further verified by a separate

262

test of Cr(III) oxidation by H2O2 (Figure 3b). Note that negligible residual H2O2 in

263

pyrite/H2O2 was favorable to avoid re-oxidation of Cr(III) in subsequent precipitation.

264

As shown in Figure 3a, negligible Cr(VI) was detected in all the pyrite/H2O2 systems

265

after precipitation, whereas Cr(VI) concentration increased to ~1.0 mg/L and ~1.7 mg/L

266

in ZVI/H2O2+OH system at initial pH 5.0 and 9.0, respectively.

267

Plausible mechanism for Cr(III)-citrate removal. Based on the above discussions,

268

the plausible mechanism for Cr(III)-citrate removal by the pyrite/H2O2 system was

269

proposed and schematically illustrated in Figure 4. Due to the reaction between FeS2

270

and H2O2 (eq. 2), massive Fe(III) was dissolved into solution, which in turn reacted

271

with FeS2 to generate Fe(II) (eq. 3). The above reactions were accompanied with the

272

release of considerable amounts of protons, keeping a constant pH at ~3.0 and thus

273

making pyrite/H2O2 system applicable for neutral and even alkaline wastewaters. The

274

reaction between Fe(II) and H2O2 (eq. 4) produced massive ·OH radicals, oxidizing the

275

complexed Cr(III) to the free Cr(VI) and concurrently releasing the organic ligands.

276

The released ligands were partially mineralized by ·OH radicals, resulting in ~37%

277

TOC removal (Figure S1b). Also, the generated Cr(VI) was reduced to Cr(III) by

278

aqueous Fe(II) and FeS2. Finally, the free metal ions, including Fe(III), Fe(II) and

279

Cr(III), were removed effectively during the subsequent precipitation, as elucidated

280

below. The above mechanism was examined by a computational kinetic model (detailed

281

in Text S1 and Table S2). As shown in Figure S8, the predicted results based on the

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model agreed with the general trend of the experimental data, further verifying the

283

proposed mechanism. In addition, the kinetic modeling confirmed that the reduction of

284

Cr(VI) by FeS2 played an indispensable role in Cr(VI) elimination other than by

285

aqueous Fe(II) (Figure S8c). In brief, the oxidative decomplexation of Cr(III)-citrate

286

into Cr(VI) dominated the overall process during the presence of H2O2. Whereas, upon

287

the depletion of H2O2, the accumulated Cr(VI) was subsequently eliminated via

288

reduction by Fe(II) and FeS2, which was a vital procedure to accomplish safe treatment

289

of Cr(III)-complexes.

290

Precipitation. Effect of pH on the precipitation efficiency of Cr and Fe was depicted

291

in Figure S9. The residual metal concentration declined to ~0.5 mg/L when the pH

292

values increased to 7.0, while further increasing pH to 10 resulted in ~1.7 mg/L Cr and

293

~3.0 mg/L Fe in the supernatant. Cr (III) hydroxides exhibited amphoteric behavior and

294

soluble Cr(OH)4- became the predominant species with increasing pH to above 10.51

295

Besides, it was reported that stable FexCr1-x(OH)3 sediment formed at pH 7.0. 52, 53 Thus,

296

the precipitation pH was set as 7.0 except for otherwise specified in this study. As

297

expected, negligible Cr(VI) was detected in the sediments after HCl dissolution.

298

Practical applicability. After the first batch run of the pyrite/H2O2 system to degrade

299

Cr(III)-citrate, the residual solid sample was filtered out and characterized by XPS and

300

XRD analysis. Figure 5(a-b) suggested that most Fe of the virgin and used pyrite

301

samples was present as Fe(II) bound to sulfur, with the binding energy at 707.5 eV. 54

302

Besides, no Cr 2p signals were detected in the utilized pyrite (data not shown),

303

suggesting that the adsorption of Cr onto pyrite could be ignored. XRD patterns in

15



304

Figure 5c also suggested similar crystalline structure of both solid samples. To examine

305

the reusability of the used pyrite, it was filter out and added in solution containing H2O2

306

and Cr(III)-citrate to initiate cyclic runs. The results in Figure 5d depicted a constant

307

Cr removal (~98%) of the pyrite/H2O2 system during 4-cyclic runs.

308

As shown in Figure S3a, pyrite/H2O2 system exhibited almost 100% removal

309

efficiency toward Cr(III)-citrate when H2O2 dosage reached 20 mM, while ~2 mg/L

310

Cr(III) could be detected in Fe2+/H2O2 system even with 40 mM H2O2. Such results

311

indicated that pyrite/H2O2 might exhibit a much higher H2O2 utilization efficiency than

312

Fe2+/H2O2.

313

To demonstrate the wide applicability and advantage of the pyrite/H2O2 system

314

toward other Cr(III)-organic complexes, we further examined the removal efficiency of

315

pyrite/H2O2 toward the complexed Cr(III) by other ligands including EDTA, citrate,

316

oxalate and tartrate. Two other processes, i.e., direct precipitation and Fe(III)

317

coagulation, were employed for comparison and the results were described in Figure

318

6a. As expected, pyrite/H2O2 was capable of reducing Cr from 10.4 mg/L initially to

319

Water Decontamination from Cr(III) - ACS Publications

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