MICROBIAL OXIDATION OF ARSENITE IN ANOXIC ENVIRONMENTS: IMPACTS ON ARSENIC MOBILITY by

MICROBIAL OXIDATION OF ARSENITE IN ANOXIC ENVIRONMENTS: IMPACTS ON ARSENIC MOBILITY  by
MICROBIAL OXIDATION OF ARSENITE IN ANOXIC ENVIRONMENTS:
IMPACTS ON ARSENIC MOBILITY
by
Wenjie Sun
_____________________
Copyright © Wenjie Sun 2008
A Dissertation Submitted to the Faculty of the
DEPARTMENT OF CHEMICAL AND ENVIRONMENTAL ENGINEERING
In Partial Fulfillment of the Requirements
For the Degree of
DOCTOR OF PHILOSOPHY
WITH A MAJOR IN ENVIRONMENTAL ENGINEERING
In the Graduate College
THE UNIVERSITY OF ARIZONA
2008
2
THE UNIVERSITY OF ARIZONA
GRADUATE COLLEGE
As members of the Dissertation Committee, we certify that we have read the dissertation
prepared by Wenjie Sun entitled “Microbial Oxidation of Arsenite in Anoxic
Environments: Impacts on Arsenic Mobility”, and recommend that it be accepted as
fulfilling the dissertation requirement for the Degree of Doctor of Philosophy.
_______________________________________________________________________
Date: 11/10/08
Maria R. Sierra Alvarez
_______________________________________________________________________
Date: 11/10/08
James A. Field
_______________________________________________________________________
Date: 11/10/08
Wendell P. Ela
_______________________________________________________________________
Date: 11/10/08
Raina M. Maier
Final approval and acceptance of this dissertation is contingent upon the candidate’s
submission of the final copies of the dissertation to the Graduate College.
I hereby certify that I have read this dissertation prepared under my direction and
recommend that it be accepted as fulfilling the dissertation requirement.
________________________________________________ Date: 11/10/08
Dissertation Chair: Maria R. Sierra Alvarez
________________________________________________ Date: 11/10/08
Dissertation Chair: James A. Field
3
STATEMENT BY AUTHOR
This dissertation has been submitted in partial fulfillment of requirements for an
advanced degree at the University of Arizona and is deposited in the University Library
to be made available to borrowers under rules of the Library.
Brief quotations from this dissertation are allowable without special permission, provided
that accurate acknowledgment of source is made. Requests for permission for extended
quotation from or reproduction of this manuscript in whole or in part may be granted by
the copyright holder.
SIGNED: Wenjie Sun
4
DEDICATION
Dedicated to my beloved wife, Ling Pan, my son, Kevin Y. Sun, and to my proud
parents, Tianzhen Sun and Aiqin Cao.
5
ACKNOWLEDGEMENTS
First and fore more I am especially grateful to my advisors Dr. Reyes Sierra Alvarez and
Dr. Jim A. Field, not only for the invaluable opportunity to conduct this research but also
for their constant encouragement and unconditional friendship.
I would like to thank my wife Ling Pan and my parents Tianzhen Sun and Aiqin Cao for
all their love and support during my graduate studies and whole life, and my son for the
fun he brings to me and my family.
I am thankful to the faculty members of the Chemical and Environmental Engineering
Department for providing me with the tools and guidance for my professional career,
especially to my committee members, Dr. Ela and Dr. Maier for their helpful discussion
and advice on this research work.
To all staff of the CHEE department, particularly Rosemary Myers, Arla Allen, Jo
Leeming, Judee Atten, Eric Case and Tommy Maynard. Their administrative and
technical support has made this dissertation possible. I also want to acknowledge Mike
Kopplin from the Pharmacology and Toxicology Department for performing arsenic
analysis.
I am especially grateful to the USGS-National Institute for Water Resources, the National
Science Foundation, the UA Water Sustainability Program, and the NIEHS-supported
Superfund Basic Research Program for the funding provided and for the financial support
to complete my PhD degree. Also arsenic analyses were performed by the Analytical
Section of the Hazard Identification Core (Superfund Basic Research Program grant
NIEHS-04940).
6
TABLE OF CONTENTS
LIST OF TABLES………………………………………………....................................17
LIST OF FIGURES…………………………………………………………………......19
ABSTRACT……………………………………………………………………………. 25
DISSERTATION OVERVIEW ………………………………………………………...28
CHAPTER 1 INTRODUCTION ………………………………………………………..33
1.1 Environmental significance of arsenic in the environment……………………...33
1.2 Sources of arsenic……………………………………...…………….………......34
1.2.1
Natural sources of arsenic...…………..……………………………….....35
1.2.2
Anthropogenic sources of arsenic .…….…………………………….......35
1.3 Arsenic speciation and toxicity…………………….…………………………….36
1.3.1
Inorganic arsenic species ………………………………………………..36
1.3.2
Organic arsenic species ………………………………............................37
1.3.3
Toxicity of arsenic ……………………………………………………....38
1.4 Arsenic and microorganisms ……………………………………………………40
1.4.1
Resistance to arsenic toxicity……………………………………………41
1.4.2
Microbial reduction of As (V) to As (III) ………………………………44
1.4.3
Microbial oxidation of As (III) to As (V)……………………………….47
1.4.3.1 Mechanism of microbial oxidation of As (III) to As (V)……………….47
1.4.3.2 Oxidation of As (III) to As (V) by hetertrophic arsenite oxdizers……....47
1.4.3.3 Oxidation of As (III) to As (V) by chemolithotrophic arsenite oxidizers
……………………………………………………………………………………49
7
TABLE OF CONTENTS—Continued
1.5 Environmental significances of biogeochemical cycles of As associated
with Fe ……………………………………………………………………….....52
1.6 Objectives .……………………………………………..………………………..57
1.7 References………………………………………………………….…………....59
CHAPTER 2 ANOXIC OXIDATION OF ARSENITE LINKED TO
DENITRIFICATION IN SLUDGES AND SEDIMENTS ...…………………………....66
2.1 Abstract ………………………………………………………………………….66
2.2 Introduction ……………………………………………………………………...67
2.3 Material and Methods …………………………………………………………...70
2.3.1
Microorganisms …………………………………………………………70
2.3.2
Basal medium …………………………………………………………....71
2.3.3
Batch bioassays ………………………………………………………….72
2.3.4
Batch assays to determine the terminal product of denitrification ……...73
2.3.5
Analytical methods ……………………………………………………...74
2.4 Results …………………………………………………………………………..76
2.4.1
Screening ………………………………………………………………..76
2.4.2
Kinetics ………………………………………………………………….79
2.4.3
Toxicity ………………………………………………………………….80
2.4.4
Consumption of NO3- and formation of N2 linked to As(III) oxidation ...81
8
TABLE OF CONTENTS--Continued
2.5 Discussions ……………………………………………………………...............86
2.5.1
Evidence of bioconversion ………………………………………………86
2.5.2
The occurrence of anoxic As(III) oxidizing bacteria ……………………87
2.5.3
As(III) substrate inhibition ………………………………………………89
2.5.4
Sorbed As(III) as a substrate for denitrification ………………………...90
2.5.5
Environmental significance ……………………………………………..91
2.6 Conclusions ……………………………………………………………………...91
2.7 Acknowledgements …………………………………………………………...…93
2.8 References ……………………………..………………………………………...94
CHAPTER 3 MOLECULAR CHARACTERIZATION AND IN SITU
QUANTIFICATION OF ANOXIC ARSENITE OXIDIZING DENITRIFYING
ENRICHMENT CULTURES …………………….……………………………………100
3.1 Abstract ………………………………………………………………………...100
3.2 Introduction …………………………………………………………………….102
3.3 Material and Methods ………………………………………………………….105
3.3.1
Enrichment cultures ……………………………………………………105
3.3.2
Medium composition …………………………………………………..107
3.3.3
Experimental incubations ………………………………………………107
3.3.4
Most Probable Number (MPN) ………………………………………...108
3.3.5
Batch assays to determine the terminal product of denitrification …….109
3.3.6
Analytical methods …………………………………………………….110
9
TABLE OF CONTENTS--Continued
3.3.7
16S rRNA gene clone libraries ………………………………………...111
3.3.8
FISH analysis …………………………………………………………..114
3.4 Results ………………………………………………………………………….115
3.4.1
Anoxic As(III) conversion by enrichment cultures ……………………115
3.4.2
Community composition of enrichment cultures ………………………120
3.5 Discussions …………………………………………………..………...............126
3.5.1
Evidence of As(III) oxidation linked to dentirification ………………..126
3.5.2
The As(III) oxidizing bacteria community – Azoarcus and
Comamonadaceae ……………………………………………………...127
3.5.3
The As(III) oxidizing bacteria community – less dominant members ...130
3.6 Conclusions …………………………………………..………….……………..133
3.7 Acknowledgements ………………………………………………………….....133
3.8 References……………………………………………………………..………..135
CHAPTER 4 ANOXIC OXIDATION OF ARSENITE LINKED TO
CHEMOLITHOTROPHIC DENITRIFICATION IN CONTINUOUS BIOREACTORS
………………………………….……………………………………………………….142
4.1 Abstract ………………………………………………………………………...142
4.2 Introduction …………………….……………….……………………………...143
4.3 Material and Methods …………………………………………...……………..146
4.3.1
Microorganisms ………………………………………………………..146
4.3.2
Basal medium ………………………………………………………......147
10
TABLE OF CONTENTS--Continued
4.3.3
Continuous columns …………………………………………………...147
4.3.4
Batch bioassay …………………………………………………………152
4.3.5
Batch assays to determine the terminal product of denitrification …….153
4.3.6
Analytical methods …………………………………………………….154
4.4 Results ……………………………………………………...…………………..155
4.4.1
Nitrate-dependent oxidation of As(III) to As(V) in continuous bioreactor
under anoxic conditions …………………………………..……………155
4.4.2
Nitrate reduction coupled to oxidation of As(III) to As(V) ……………160
4.4.3
Kinetics of chemolithotrophic As(III) oxidizers in the sludge in the
column R1 ……………………………………………………………...161
4.4.4
Terminal products of autotrophic denitrification linked to As(III) oxidation
to As(V) ………………………………………………………………..165
4.4.5
Growth with various electron acceptors ……………………………….171
4.4.6
Anoxic oxidation of As(III) to As(V) linked to denitrification in
continuous bench-scale UASB bioreactor ……………………………..171
4.5 Discussions …………………………………………………………………….176
4.5.1
Microbial anoxic oxidation of As(III) linked to chemolithotrophic
denitrification in the continuous bioreactors …………………………...176
4.5.2
End product of denitrification linked to As(III) oxidation in the continuous
bioreactors ……………………………………………………………...178
4.5.3
As(III) substrates inhibition on anoxic oxidation of As(III) in the
continuous bioreactors …………………………………………………180
11
TABLE OF CONTENTS--Continued
4.6 Conclusions …………………………………………………………………….181
4.7 Acknowledgements …………………………………………………………….182
4.8 References…………………………………………………………………..…..183
CHAPTER 5 THE ROLE OF DENITRIFICATION ON ARSENITE OXIDATION AND
ARSENIC MOBILITY IN ANOXIC SEDIMENT COLUMN MODEL WITH
ACTIVATED ALUMINA…….…………………………………………………….…189
5.1 Abstract ………………………………………………………………………..189
5.2 Introduction …………………………………………………………...……….191
5.3 Material and Methods ………………………………………………………….193
5.3.1
Microorganisms ………………………………………………………..193
5.3.2
Basal medium …………………………………………………………..194
5.3.3
Adsorption isotherms of As(III) and As(V) on activated alumina …….195
5.3.4
Attenuation of As(III) under chemolithotrophic denitrifying conditions in
sediment columns packed with activated alumina ……………………..196
5.3.5
Arsenic extraction ……………………………………………………...198
5.3.6
Analytical methods …………………………………………………….200
5.4 Results….………………………………………………………………….........201
5.4.1
Adsorption isotherms of As(III) and As(V) on the Activated Alumina .201
5.4.2
Attenuation of As(III) in AA packed bed columns …………………….203
5.4.3
Residual arsenic in activated alumina and sludge………………………209
5.5 Discussions …………………………………………………………………….212
12
TABLE OF CONTENTS--Continued
5.5.1
Bioremediation of arsenic by the addition of nitrate .…………………212
5.5.2
Microbial nitrate-dependent oxidation of As(III) to As(V)…………….213
5.5.3
Adsorption of As(III) and As(V) on AA………………………………..214
5.5.4
Implications……………………………………………………………..216
5.6 Conclusions …………………………………………………………………….217
5.7 Acknowledgements …………………………………………………………….218
5.8 References...………………………………………………………...…………..219
CHAPTER 6 ANOXIC OXIDATION OF As(III) AND Fe(II) LINKED TO
CHEMOLITHOTROPHIC DENITRIFICATION FOR THE IMMOBILIZATION
OF As IN ANOXIC ENVIRONMENTS…………………………………………….....223
6.1 Abstract ………………………………………………………………………...223
6.2 Introduction …………………….……………….……………………………...225
6.3 Material and Methods ………………………………………………………….228
6.3.1
Microorganisms ………………………………………………………..228
6.3.2
Basal medium ………………………………………………………......229
6.3.3
Formation of biogenic iron (III) oxides coated sand (ICS) linked to
denitrification …………………………………………………………..230
6.3.4
Sorption isotherms of As(III) and As(V) on biogenic iron oxides coated
sands (ICS) …………………………………..…………………………231
13
TABLE OF CONTENTS--Continued
6.3.5
Attenuation of As(III) on freshly biogenic iron(III) oxides coated sands
under chemolithotrophic denitrifying conditions in sediment columns
packed with sand ………………………………………………………232
6.3.6
Batch assay …………………………………………………………….233
6.3.7
Iron and Arsenic extraction from ICS …………………………………234
6.3.8
Mineral analysis ………………………………………………………..235
6.3.9
Scanning electron microscope/ Energy dispersive spectroscopy
(SEM/EDS) analysis …………………………………………………...236
6.3.10 X-ray photoelectron spectroscopy (XPS) analysis …………………….237
6.3.11 Analytical methods……………………………………………………..237
6.4 Results ……………………………………………………...…………………..239
6.4.1
Sorption isotherms of As(III) and As(V) on the nitrate-dependent iron(III)
oxides coated sands …………………………………………….……...239
6.4.2
Microbial nitrate-dependent oxidation of Fe(II) and subsequent
precipitation of Fe(III) oxides formed on sand surface to form iron(III)
coated sands in sediment columns …………………………………….242
6.4.3
Microbial nitrate-dependent oxidation of As(III) and subsequent
adsorption of As(V) on iron(III) oxides coated sands in sediment
columns....................................................................................................246
6.4.4
Residual iron and arsenic in the sand at the end of reactor operation …250
6.4.5
Mineralogy of iron and arsenic on the surface of sand ………………..256
6.5 Discussions …………………………………………………………………….260
14
TABLE OF CONTENTS--Continued
6.5.1
Bioremediation of As in the presence of nitrate ……………………….260
6.5.2
The mechanisms of immobilization of As on iron (hydr)oxides in anoxic
environments …………………………………………………………...260
6.5.3
Implications……………………………………………………………..265
6.6 Acknowledgements …………………………………………………………….266
6.7 References .……………………………………………………………………..267
CHAPTER 7 NOVEL STUDY OF ANOXIC OXIDATION OF As(III) TO As(V)
LINKED TO CHLORATE REDUCTION ………………………….....................……276
7.1 Abstract ………………………………………………………………………...276
7.2 Introduction …………………………………………………………………….278
7.3 Material and Methods ………………………………………………………….281
7.3.1
Microorganisms ………………………………………………………..281
7.3.2
Basal medium ………………………………………………………….282
7.3.3
Batch bioassays ………………………………………………………..282
7.3.4
Enrichment cultures ……………………………………………………283
7.3.5
Most Probable Number (MPN) ………………………………………..285
7.3.6
Continuous column…………………………………………..…………286
7.3.7
Analytical methods …………………………………………………….287
7.4 Results ………………………………………………………………………….288
7.4.1
Screening ……………………………………………………………….288
15
TABLE OF CONTENTS--Continued
7.4.2
Kinetics ………………………………………………………………...290
7.4.3
MPN assays for enrichment cultures …………………………………..293
7.4.4
Rates of anoxic As(III) oxidation linked to autotrophic and hetertrophic
chlorate reduction ………………………………………………………294
7.4.5
End products of autotrophic and heterotrophic chlorate reduction
linked to anoxic As(III) oxidation ……………………………………...298
7.4.6
As(III) oxidation by autotrophic As(III) oxidizing chlorate reducing
bacteria utilizing oxygen as sole electron acceptor …………….……...301
7.4.7
Microbial oxidation of As(III) to As(V) utilizing chlorate as electron
acceptor in continuous bioreactor under anoxic conditions.....................303
7.4.8
Chlorate reduction coupled to anoxic oxidation of As(III)
to As(V) in bioreactor ……………………………………………...…..309
7.4.9
As(III) substrates inhibition to As(III)-oxidizing
chlorate-reducing bacteria ……………………………………………..311
7.5 Discussions …………………………………………………………….............313
7.5.1
Evidence of biological oxidation of As(III) to As(V) linked to
chlorate reduction……………………………………………..…..……313
7.5.2
The occurrence of As(III) oxidizing bacteria utilizing chlorate as electron
acceptor in anoxic environments……………………………………….315
7.5.3
The characteristic of anoxic As(III) oxidizing
chlorate-reducing bacteria……………………………………………...316
16
TABLE OF CONTENTS--Continued
7.5.4
Anoxic oxidation of As(III) linked to chemolithotrophic chlorate
reduction in continuous bench-scale UASB bioreactor…………..….…318
7.5.5
As(III) substrates inhibition on chemolithotrophic As(III)-oxidizing
chlorate-reducing bacteria……………………………………………...319
7.5.6
Bioremediation potential………………………………………………..321
7.6 Acknowledgements ………………………………………………………….…321
7.7 References ……………………………………………………………………...322
CHAPTER 8 CONCLUSIONS ……………………………………….……….............329
8.1 Microbial oxidation of As(III) to As(V) in anoxic environments ……………...329
8.1.1
Anoxic oxidation of As(III) to As(V) is linked to chemolithotrophic
denitrification in sludges and sediments ……………………………….329
8.1.2
Molecular characterization and in situ quantification of anoxic arsenite
oxidizing denitrifying enrichment cultures ………………………….....330
8.1.3
Anoxic oxidation of arsenite linked to chemolithotrophic denitrification in
continuous bioreactors ………………………………………………....331
8.1.4
Novel Study of the Anoxic Oxidation of As(III) to As(V) linked to
Chlorate Reduction …………………………………………………….333
8.2 The role of denitrification on arsenite oxidation and arsenic mobility in anoxic
sediment column model with activated aluminum ………………………….....334
8.3 Anoxic oxidation of As(III) and Fe(II) linked to chemolithotrophic denitrification
for the immobilization of As in anoxic environments ………………………....336
17
LIST OF TABLES
Table 2.1
Summary of microbial As(III) oxidation under denitrifying conditions...78
Table 2.2
Summary of kinetics† of microbial As(III) oxidation (0.5 mM) under
denitrifying conditions ……………………………………………..........79
Table 2.3
Summary of sorbed As(III) oxidation linked to denitrification …………85
Table 3.1
Summary of denitrification linked to As(III) oxidation when adsorbed
onto AA (50 g l-1) …………………………… ………….......................119
Table 3.2
Group-specific oligonucleotide probes used in this study ……………..126
Table 4.1
Operational periods for the two continuous columns of the first continuous
experiment ……………………………………………………………...148
Table 4.2
Operational periods for the UASB reactor of the second continuous
experiment ……………………………………………………………...150
Table 4.3
Summary of As(III) oxidation linked to denitrification in bioreactor
R1 ………………………………………………………………………166
Table 4.4
Nitrate-N balance for the denitrification linked to As(III) oxidation in
batch experiment to determine terminal product of N …………………169
Table 4.5
Results summary of operation periods for the R3 bioreactor ………….175
Table 5.S1
Alkaline extraction protocol for solid-phase As on the AA …...............199
Table 5.1
Freundlich and Langmuir isotherm model parameters for the adsorption
of As(III) and As(V) on AA …………………………………………...202
Table 5.2
Mass balance of arsenic at the end of the experiment ……………….210
18
LIST OF TABLES—Continued
Table 6.1
Freundlich and Langmuir isotherm model parameters for the adsorption
of As(III) and As(V) on iron oxides …………………………………...241
Table 6.2
Mass balance of arsenic and iron at the end of the experiment ………..252
Table 6.3
Mass of arsenic and iron on the ICS of the two continuous sediment
columns at different depth intervals ………………………………...….253
Table 7.1
Summary of microbial oxidation of As(III) (0.5 mM) to As(V) linked to
(per)chlorate ……………………………………………………………291
Table 7.2
Summary of kinetics of microbial As(III) oxidation (0.5 mM) under
chlorate reducing conditions ………………………..………………….292
Table 7.S1
MPN estimations of cell density and cell yield of four enrichment
cultures………………………………………………………………….293
Table 7.3
Summary of kinetics of microbial As(III) oxidation (1.0 mM) under
chlorate reducing conditions by ECC1 ………………………………...297
Table 7.4
Mass Balance for the autotrophic (ECC10 and heterotrophic (ECC3)
chlorate respiration linked to anoxic As(III) (0.5 mM) Oxidation by
ADS.........................................................................................................300
Table 7.5
Results summary of operation periods for the UASB reactor .………...308
19
LIST OF FIGURES
Figure 1.1
Distribution of arsenic in groundwater in major aquifers as well as water
in the world ……………………………………………..…………….....34
Figure 1.2
Schematic representation of the role of microorganisms in arsenic cycling
n the environment ………………………………………………………41
Figure 1.3
Phylogenetic
diversity
of
representative
arsenic-metabolizing
prokaryotics................................................................................................43
Figure 1.4
Biogeochemical redox cycles of arsenic and iron as well as
their interaction …………………………………………………………56
Figure 2.1
The removal of As(III) (panel A) and formation of As(V) (panel B)
by NGS and DPS under denitrifying condition ………………………...77
Figure 2.2
Substrate toxicity as shown by a decrease in the rate of As(V) formation
from the oxidation of As(III) at different initial As(III) concentrations
…………………………………………………………………………....80
Figure 2.3
Elimination of As(III) and formation of As(V) by NGS under
denitrifying condition ………….……………………………….………..81
Figure 2.4
Denitrification at a high initial As(III) aqueous concentration (3.5 mM)
……………………………………………………………………………83
Figure 2.5
Denitrification with As(III) (3.5mM) adsorbed to activated alumina
....................................................................................................................84
Figure 3.1
Formation of As(V) by cultures under denitrifying condition ………....116
20
LIST OF FIGURES--Continued
Figure 3.2
The influence of As(III) on the formation of N2-N, consumption of
NO3- and accumulation of NO2- by EC1 and EC3 cultures under
denitrifying conditions.……………………………………...................118
Figure 3.3
Phylogenetic distributions in the four cultures …………..…………….121
Figure 3.4
Phylogenetic relationships of the EC and MC clones recovered
in this study ………………….……………….. ………………………122
Figure 3.5
FISH analysis of the EC and MC communities using universal
(Eub338) and group-specific (Azo644 and DEN220) probes ………….124
Figure 3.6
Evaluation of representative clones from three ECs and one MC
by rarefaction analysis………………………………………………….125
Figure 4.1
Anaerobic bioreactor R1 and R2 for the first continuous experiment,
and schematic diagram of the UASB bioreactor R3 for the second
continuous experiment ……………………..………………………….151
Figure 4.2
Concentration of As(III) and As(V) in the continuous bioreactor
R1 and R2 as a function of time ………………………………………..156
Figure 4.3
Comparison of the period average As(III) and As(V) concentrations
in the influent and effluent of bioreactors R1 and R2….........................159
Figure 4.4
The influent and effluent concentrations of NO3- determined in the
continuous bioreactor R1 as a function of time ......................................161
Figure 4.5
Elimination of As(III) and formation of As(V) by R1 bio-film at 0.5
and 1.5 mM As(III) under denitrifying conditions……………………..162
21
LIST OF FIGURES--Continued
Figure 4.6
Panel A, formation of As(V) under denitrifying conditions; panel B,
formation of N2 in all the treatments link As(III)
oxidation to denitrification ……………………………………………164
Figure 4.7
Elimination of As(III) and formation of As(V) under denitrifying
conditions …………………………..…………………………………..167
Figure 4.8
Formation of N2 (Panel A) and N2O (Panel B) under
denitrifying conditions.............................................................................170
Figure 4.9
Influent and effluent concentrations of As(III)) and As(V) during
the operation of a 2-L bench-scale up-flow anaerobic sludge
blanket reactor (R3) ……………………………………………………174
Figure 5.1
Adsorption isotherms of As(V) and As(III) on AA …….. …………….203
Figure 5.2
Effluent pH of two AA packed columns fed with a mineral medium
containing 6.67 µM As(III) … …………………………………………204
Figure 5.3
Removal of total arsenic in two activated alumina packed columns
fed with a mineral medium containing 6.67 µM As(III). ……………...205
Figure 5.4
Concentrations of As(III) and As(V) in the influent and effluent of
AA packed column S1 and S2 as a function of time …………….…….207
Figure 5.5
Speciation of arsenic in the influent and effluent of AA packed column
S1 and S2 as a function of time ………..................................................208
Figure 5.6
The profile of sorbed arsenic in two AA packed columns
at the end of the continuous experiment ……………………….………211
22
LIST OF FIGURES—Continued
Figure 6.1
Adsorption isotherms of As(V) and As(III) on biogenic iron oxides ….240
Figure 6.2
Concentrations of Fe(II) and Fe(III) in the influent and effluent of
column SF1 and column SF2 as a function of time ……………………243
Figure 6.3
Iron speciation in the influent and effluent of sand packed columns ….244
Figure 6.4
Removal of total arsenic in two sand packed columns fed with a
mineral medium containing 6.67 µM As(III) …………….……………245
Figure 6.5
Removal of total arsenic in two sand packed columns fed with
a mineral medium containing 6.67 µM As(III).………………….…….247
Figure 6.6
Concentrations of As(III) and As(V) in the influent and effluent
of column SF1 (Panel A) and column SF2 (Panel B) as a
function of time …………………………………………………………249
Figure 6.7
Arsenic speciation in the influent and effluent of sand packed
columns ………………………………………………………………....250
Figure 6.8
The profile of sorbed iron on the ICS in two sands packed columns
at the end of the continuous experiment ……………………………….254
Figure 6.9
The profile of sorbed arsenic on the ICS in two sands packed columns
at the end of the continuous experiment ………………………………255
Figure 6.10
SEM-EDS for column profile………………………………………….257
Figure 6.11
XPS for SF1 and SF2 column profile …………………………………259
23
LIST OF FIGURES--Continued
Figure 7.1
The removal of As(III) and formation of As(V) by ADS, RAS and
NGS in the presence or absence of chlorate under anoxic
condition ………….................................................................................289
Figure 7.2
The removal of As(III) and formation of As(V) by ADS
autotrophic enrichment (ECC1) in the presence of chlorate
under anoxic condition. ………………………………………………...295
Figure 7.3
The removal of As(III) and formation of As(V) by ADS
heterotrophic enrichment (ECC3) in the presence of chlorate
under anoxic condition............................................................................296
Figure 7.4
Chlorate reduction at a initial aqueous As(III) concentration (1.0 mM).
The consumption of ClO3- and the formation of Cl- by
ECC1 under anoxic condition.……………………………………….....299
Figure 7.5
The removal of As(III) and formation of As(V) by ECC1
in the presence of elemental oxygen …………………………………...302
Figure 7.6
Activity of A(III) oxidation in the presence of O2 by ECC1…………...303
Figure 7.7
The removal of As(III) and the formation of As(V) in the continuous
bioreactor linking the anoxic oxidation of As(III) to chlorate respiration
as a function of time ……………………………...…………………….305
Figure 7.8
The influence and effluence concentration of As(III) and As(V) for each
period of the total operation in the continuous bioreactor link the anoxic
oxidation of As(III) to chlorate respiration …………..………………...306
24
LIST OF FIGURES--Continued
Figure 7.9
The removal of ClO3- and the formation of Cl- in the continuous
bioreactor link the anoxic oxidation of As(III) to chlorate respiration as
a function of time ………………………………………………………310
Figure 7.10
The summary of molar ratios of As(III) removed compared
to ClO3- consumed and Cl- formed, and As(V) formed compared
to ClO3- consumed and Cl- formed in the continuous bioreactor
link the anoxic oxidation of As(III) to chlorate respiration as a
function of time ………………………………………………………..311
Figure 7.11
Activity and substrate toxicity as shown by a decrease in the rate
of As(V) formation from the oxidation of As(III) by AOCRB
from bioreactor at different initial As(III) concentrations ……………..312
25
Abstract
Arsenic (As) contamination of groundwater and surface water is a worldwide problem.
Exposure to arsenic in drinking water is an important current public health issue. Arsenic
is well known for its carcinogenic and teratogenic effects. The U.S. Environmental
Protection Agency (USEPA) has recently enacted a stricter drinking water standard for
arsenic that lowers the maximum contaminant level (MCL) from 50 to 10 µg l-1.
Localized elevated As concentrations in groundwater or surface water have been
attributed to the natural release of As from the weathering of As bearing minerals.
Microbial reduction of arsenate (As(V)) to arsenite (As(III)) and ferric (hydr)oxides to
Fe(II) is hypothesized to be the dominant mechanisms of As mobilization in subsurface
environments. If oxidizing conditions can be restored, As can be immobilized by the
formation of As(V) and ferric (hydr)oxides. As(V) is more strongly adsorbed than As(III)
at circumneutral conditions by common non-iron metal oxides in sediments such as those
of aluminum. Ferric (hydr)oxides have strong affinity for both As(III) and As(V) in
circumneutral environments. Oxygen can be introduced into the anaerobic zone by
injection of gaseous O2 to promote oxidation reactions of As(III) and Fe(II), but O2 is
poorly soluble and chemically reactive and thus difficult to distribute in the subsurface.
Nitrate or chlorate can be considered as alternative oxidants with advantages over
elemental oxygen due to their high aqueous solubility and lower chemical reactivity
which together enable them to be better dispersed in the saturated subsurface.
26
The objective of this study is to evaluate the importance of anoxic oxidation of
As(III) to As(V) by anaerobic microorganisms such as chemolithotrophic denitrifying
bacteria and chlorate respiring bacteria in the biogeochemical cycle of arsenic. This study
also investigated an arsenic potential bioremediation strategy based on injecting nitrate or
chlorate into contaminated groundwater and surface water under anaerobic conditions.
In this study, denitrification or chlorate reduction linked to the oxidation of
As(III) to As(V) was shown to be a widespread microbial activity in anaerobic sludge and
sediment samples that were not previously exposed to arsenic contamination. The
biological oxidation of As(III) utilizing nitrate or chlorate as sole electron acceptor was
feasible and stable over prolonged periods of operation in continuous-flow anaerobic
bioreactors. Evidence for the complete denitrification was demonstrated by direct
measurement of N2 formation dependent on As(III) addition. Also complete chlorate
reduction to chloride was attributable to the oxidation of As(III). A 16S rRNA gene clone
library characterization of enrichment cultures indicated that the predominant phylotypes
responsible for As(III) oxidation linked to denitrification were from the genus Azoarcus
and the family Comamonadaceae. A bioremediation strategy was explored that is based
on injecting nitrate to support the microbial oxidation of Fe(II) and As(III) in the
subsurface as a means to immobilize arsenic. Two models were utilized to illustrate the
mechanisms of As removal.
27
1) Sediment columns packed with activated alumina were utilized to demonstrate the
role of nitrate in supporting microbial As(III) oxidation and arsenic mobility in
anoxic sediments containing mostly non-iron oxides;
2) Sand-packed columns were used to simulate natural anaerobic groundwater and
sediment systems with co-occurring As(III) and Fe(II) in the presence or absence
of nitrate. Microbial oxidation by denitrifying microorganisms lead to the
formation of ferric (hydroxides) which adsorbed As(V) formed from As(III)oxidation.
The studies presented here demonstrate that anoxic microbial oxidation of As(III)
and Fe(II) linked to denitrification significantly enhance the immobilization of As in the
anaerobic subsurface environments.
28
DISSERTATION OVERVIEW
This dissertation work is divided in eight chapters. A brief overview of the content in
each of the chapters is given below.
Chapter 1 Introduction
This chapter presents a general introduction on the topic of arsenic in the environment
and provides the motivation and importance to perform this research work. A brief
background section on the significant impacts of arsenic in the environment and the main
microbial processes implicated in the transformation and mobilization of this metalloid
are also included.
Chapter 2 Anoxic Oxidation of Arsenite Linked to Denitrification in Sludges and
Sediments
This chapter provides an assessment to determine whether the microbial oxidation of
As(III) linked to denitrification is a widespread process occurring in anoxic
environments. To answer this question, inocula from various sources was investigated,
including samples of pond sediments from lakes and rivers, wastewater sludge from
municipal and industrial treatment facilities, as well as groundwater. Most of the inocula
29
sources tested were obtained from environments that are not contaminated with As. The
research results from this chapter were published in a paper entitled “Anoxic Oxidation
of Arsenite Linked to Denitrification in Sludges and Sediments” by Wenjie Sun, Reyes
Sierra and Jim A. Field. Water Research (2008) 42, 4569-4577.
Chapter 3 Molecular Characterization and In Situ Quantification of Anoxic
Arsenite Oxidizing Denitrifying Enrichment Cultures
To explore microorganisms involved in the microbial oxidation of arsenite (As(III))
under denitrifying conditions, three enrichment cultures and one mixed culture that
originated from anaerobic environmental samples were characterized. The research
results from this chapter were submitted as a publication entitled “Molecular
Characterization and In Situ Quantification of Anoxic Arsenite Oxidizing Denitrifying
Enrichment Cultures” by Wenjie Sun, Reyes Sierra-Alvarzez, Nuria Fernandez, Jose Luis
Sanz, Ricardo Amils, Antje Legatzki, Raina M. Maier and Jim A. Field. FEMS Microbial
Ecology 2008 (under review).
Chapter
4
Anoxic
Oxidation
of
Arsenite
Linked
to
Chemolithotrophic
Denitrification in Continuous Bioreactors
The main objective of this study was to determine whether the anoxic microbial oxidation
of As(III) coupled to denitrification could be sustained for long periods of time in
bioreactors with the biomass immobilized as granular biofilms. This section will be
30
submitted in 2008 as a paper entitled “Anoxic Oxidation of Arsenite Linked to
Chemolithotrophic Denitrification in Continuous Bioreactors” by Wenjie Sun, Reyes
Sierra-Alvarez, Pieter Rowlette and Jim A. Field.
Chapter 5 The Role of Denitrification on Arsenite Oxidation and Arsenic Mobility
in an Anoxic Sediment Column Model with Activated Alumina
The objective of this study was to evaluated the potential of injecting nitrate into
anaerobic groundwater as a bioremediation method to promote the anoxic oxidation of
As(III) by denitrifying bacteria. The hypothesis was tested using two sediment columns,
which were packed with activated alumina and fed with As(III) in the presence or
absence of nitrate. The columns simulated anaerobic subsurface environments containing
mostly non-iron oxides. The results of this research work will be submitted for
publication as a paper entitled “The Role of Denitrification on Arsenite Oxidation and
Arsenic Mobility in an Anoxic Sediment Column Model with Activated Alumina” by
Wenjie Sun, Reyes Sierra-Alvarez, Ronald S. Oremland and Jim A. Field.
31
Chapter 6 Anoxic Oxidation of Arsenite and Ferrous Iron Linked to
Chemolithotrophic Denitrification for the Immobilization of Arsenite in Anoxic
Environments
The objectives of this study were to evaluate the addition of nitrate as a means to
immobilize arsenic via the oxidation of As(III) and co-occuring Fe(II) to As(V) adsorbed
on biogenic iron(III) hydroxides. The study was conducted with two continuous-flow
sand packed columns, which were fed with As(III) and Fe(II) in the presence and absence
of nitrate. The research results from this chapter will be submitted for publication as a
paper entitled “Anoxic Oxidation of Arsenite and Ferrous Iron Linked to
Chemolithotrophic Denitrification for the Immobilization of Arsenite in Anoxic
Environments” by Wenjie Sun, Reyes Sierra-Alvarez, Ronald S. Oremland and Jim A.
Field.
Chapter 7 Novel Study of Anoxic Oxidation of As(III) to As(V) linked to Chlorate
Reduction
This chapter assessed the feasibility of utilizing chlorate as an electron acceptor for the
biological oxidation of As(III) under anoxic conditions. To answer this question, inocula
from various sources were investigated, including samples of wastewater sludge from
municipal and industrial treatment facilities not known to be contaminated with As. This
study provided clues for a potential bioremediation technology based on converting
32
As(III) to As(V) in subsurface environments. As(V) is expected to be more strongly
adsorbed than As(III) by certain metal oxides such as those of aluminum. The research
results from this chapter will be submitted for publication.
Chapter 8 Conclusions
The environmental significance and main conclusions from this dissertation are
highlighted in this chapter.
33
CHAPTER 1
INTRODUCTION
1.1. Environmental significance of arsenic in the environment
Arsenic (As) contamination of groundwater and surface water is a worldwide problem
(Figure 1.1). Exposure to arsenic in drinking water is an important and current public
health issue, which contributes to non-cancer and cancer diseases [WHO, 1993; Smedley
and Kinniburgh 2002]. Arsenic is a natural occurring element of rocks and will
eventually enter the groundwater or surface water reservoirs through processes of
weathering and mobilization [Welch et al., 2000; Smedley and Kinniburgh 2002].
Industrial, agricultural activities and mining are also important sources of arsenic into the
environment [Thelin et al., 2000; Oremland and Stolz, 2003]. The U.S. Environmental
Protection Agency (USEPA) has recently enacted a stricter drinking water standard for
arsenic that lowers the maximum contaminant level (MCL) from 50 to 10 µg l-1 [USEPA,
2001] since 2006. Approximately 10% of over 30,000 groundwater samples taken in the
USA contain As at levels in excess of the newly enacted MCL [Welch et al., 2000]. The
new regulation is expected to affect approximately 4,000 water utilities nationwide,
which will be required to remove As from drinking water [USEPA, 2001]. The USEPA
and the AWWARF estimate that it will cost between $102 million and $550 million to
34
treat the As by less than 10 µg l-1 in drinking water per year [USEPA, 2001; Frey et al.,
2000].
In the next section of this, a more detailed description of the arsenic sources, speciation
and microbial interactions will be reviewed.
Figure 1.1 Distribution of arsenic in groundwater in major aquifers as well as water in
the world [Smedley and Kinniburgh 2002].
1.2. Sources of arsenic
The occurrence of arsenic in the environment can be attributed to natural processes and to
anthropogenic activities [Winski et al., 1997]. A brief description of the main sources of
arsenic is given in the follow paragraphs.
35
1.2.1. Natural sources of arsenic
Arsenic compounds occur naturally in various reservoirs including ocean, soils, biota,
atmosphere and different kinds of crustal rocks and minerals [Oremland and Stolz.,
2003]. Among these reservoirs, rocks and minerals, such as iron sulfides (e.g.
arsenopyrite,), manganese oxides, sulfides (e.g. orpiment), copper (e.g. enargite), silver,
(e.g. proustite), iron (e.g. loellingite) gold and lead, account for almost 99% of the total
arsenic in the environment. According to the US Environmental protection Agency
[USEPA, 2001] the concentration of arsenic in soil ranges from 0.1 to 1000 mg kg-1, in
fresh water 0.4 ppb, sea water 2.6 ppb, atmospheric dust 50 to 400 ppm, and marine
animals (fish and mussels) 0.001 to 0.3 mg g-1 dwt mass. Through the processes of
dissolution, weathering and erosion, these minerals ultimately release arsenic into the
natural sediments, groundwater and surface water systems. Volcanic activity, geothermal
waters, hot springs and forest fires are also natural sources of arsenic.
1.2.2. Anthropogenic sources of arsenic
Industrial, agricultural activities and mining are also important sources of arsenic into the
environment, for example, smelting of sulfides ores, coal combustion, mine tailings and
wood preservation (e.g., copper and chromium arsenate formulations) [Roussel et al.,
36
2000; Katz et al., 2005]. Organic arsenicals like roxarsone (4-hydroxy-3-nitrophenyl
arsenic acid) are used as a feed additive to promote growth of poultry [Rutherford et al.,
2003]. In recent years, arsenic compounds (e.g. indium arsenide, gallium arsenide) are
finding increasing application in the semiconductor industry. The agricultural use of
arsenic-containing herbicides (monosodium methanearsonate, disodium methanearsonate
and cacodylic acid) [Sierra-Alvarez et al., 2006] and insecticides (calcium arsenate, lead
arsenate) [Thelin et al., 2000; Bednar et al., 2002] also contribute to the release of
arsenic.
1.3. Arsenic speciation and toxicity
Arsenic is a metalloid that occurs in five different oxidation states: -3, 0, +1, +3 and +5.
As in the environment can be found as inorganic or organic species. When carbon is
present within the chemical structure of the compound it is classified as an organic As
species.
1.3.1. Inorganic arsenic species
H3AsO4, H2AsO4-, HAsO42- and AsO3- are the dominant arsenate (As(V)) species in the
aqueous phase, the pKa values for those species are 2.3, 6.8 and 11.6.; while H3AsO3,
H2AsO3-, and HAsO32- are the main arsenite (As(III)) forms with pKa values of 9.2 and
37
13.1 [Welch et al., 2000]. Redox potential of the environment has a great impact on the
transformation of arsenic species between As(III) and As(V). For example, the Eh of
O2/H2O and NO3-/N2 are +0.82 and +0.74 V [Rhine et al., 2006; Oremland et al., 2004]
respectively, which are much higher than the Eh (+0.139 V) of As(V)/As(III) [Madigan
and Martinko, 2006]. It is feasible for As(III) to be oxidized to As(V) under aerobic and
anoxic conditions with higher Eh. In contrast, if Eh is lower than +0.139 V (e.g.
sulfate/sulfide (-0.22 V)), As(V) tends to be reduced to As(III) [Cullen and Reimer, 1989;
Carbonel-Barrachina et al., 1999].
As(V) is the predominant arsenic form under aerobic conditions. In contrast, As(III) is
prevalent under anaerobic conditions with low Eh. As(V) tends to be much strongly
adsorbed onto the surface of solid phases containing clay and aluminum oxides than
As(III) [Lin and Wu, 2001]. Iorn(hydr)oxides exhibit similar affinity on both As(III) and
As(V) [Raven et al., 1998; Fukuoka et al., 2002; Dixit and Hering, 2003]. Thus, As(III) is
easily mobilized and released into the aqueous in subsurface environments with non-iron
metal oxides [Carbonel-Barrachina et al., 1999; Kumaresan et al., 2001].
1.3.2. Organic arsenic species
Organic arsenic compounds are detected in the environment when microorganisms
metabolize As arsenic, even in the human body, as a result of enzyme activity (e.g.
38
methylation of inorganic arsenic species). Organic As compounds can be classified as
arsenobetain, methylated arsenic, arsenosugars, arsenolipids and other organoarsenic
compounds [Bentley and Chasteen, 2002]. Most of these compounds were used, and
some are still used, as effective agents against various pests, e.g. insects, weeds and even
human parasites.
The speciation of As has great impact on the fate and toxicity of As in the environment.
Inorganic As species are considered as the most important arsenic compounds due to
their abundance. The transformations between As species affect not only the cycling of
As in the natural environment, but also the efficiency of existing technologies for their
control and treatment.
1.3.3. Toxicity of arsenic
It is well known that arsenic is one of the most commonly used poisons since antiquity.
Humans are exposed to As from various pathways including ingestion, inhalation, and
possible dermal adsorption. Ingestion of arsenic from water and food is the main route of
exposure to As by the general population. The toxicity of As can affect a wide range of
organisms, including humans, and cause numerous chronic and acute effects [WHO,
1993; ATSDR, 2007]. It is well established that exposure to high As concentrations
increases the incidence of cancers of lung, kidney, liver and bladder. Arsenic may cause
39
damage to many organs of humans, such as the gastrointestinal tract, circulatory system,
liver, kidney, skin, and other tissues very sensitive to arsenic [WHO, 1993; ATSDR,
2007].
Arsenic is also very toxic to microorganisms and aquatic organisms. For example,
sodium arsenite was applied to control the rooted aquatic plants at a dosage of 10 mg l-1
[Moore et al., 1984]. As toxicity to fish is well known. The LC50 values of sodium
arsenite range from 0.05 to 59 mg l-1 for different fish, depending on the species, age and
conditions of the test [Tisler et al., 2002]. The 50% inhibitory concentrations of As(III) to
methanogenic activity are reported to be as low as 15 µM [Sierra-Alvarez et al., 2004].
For different arsenic compounds, the mode of toxicity depends on the chemical form of
As. As(III) is more toxic than As(V) according to various studies, because it can enter the
cell through aqua-glycerolporins, bind to sulfhydryl groups and inactivate/denature the
normal functions of many proteins and enzymes, even finally causing cell damage
[Mukhopahyay et al., 2002]. As(III) is known to interfere with essential enzymatic
functions and transcriptional events and inhibit more than 200 enzymes in the human
body, and it also binds to the vicinal thiols in pyruvate dehydrogenase and 2-oxoglutarate
dehydrogenase to affect respiration chains [WHO, 1993; Oremland and Stolz, 2003;
Oremland et al., 2005]. As(V) is similar to phosphate in molecular structure and usually
enters the cell via phosphate transporters. As(V) can inhibit oxidative phosphorylation,
which is the main energy-generation system for living organisms [Oremland and Stolze,
2003; Oremland et al., 2005].
40
In the past, organic arsenic species were thought to be the result of an As detoxification
mechanism. However, recent studies have shown that in the methylation process of
inorganic As (i.e., Monomethylarsonic acid (MMAA), dimethylarsinic acid (DMAA))
may not only be a detoxification process, but may also enhance the toxicity of As. For
example, in vitro studies in animal and human cells have illustrated that trivalent
methylated As compounds are more toxic than As(III) [Yamanaka et al., 1991].
Due to the negative effect of arsenic to human health and to the environment it is
important to gain a better understanding of the different microbiological processes
involved in the mobilization and transformation of this metalloid.
1.4. Arsenic and microorganisms
Unlike organic compounds, arsenic, as a metalloid compound, cannot be degraded
biologically to environmental benign products. Although it is well known that As is toxic
to most microorganisms, some microorganisms still can acclimate to As toxicity and
evolve some resistant mechanisms to survive, and even transform or utilize As
[Oremland and Stolz, 2003; Inkeep et al., 2007]. Because chemical transformation
processes (e.g oxidation of As(III) by element oxygen) [Eary et al., 1990] are relatively
slow, microorganisms play an important role in the biogeochemical cycling of As by
catalyzing rapid conversions between different forms of As. Figure 1.2 illustrates the
41
major pathways between reduction, oxidation and methylation of inorganic arsenic
mediated by microbes. These processes can change the speciation of arsenic and thus
influence the mobilization and toxicity of arsenic species in the environment.
methylation
(Challenger Mechanism)
arsenite oxidation
O2
NO3-
CH3
H
Fe3+
OH
HO As OH
HO
OH
arsenate
As
OH
arsenite
CH3
volatile
methyl
arsines
O
O
As
HO As OH
CH3
MMA(V)
organic matter
CH3
arsenate reduction
mineralization
reduction HO
As
OH
MMA(III)
Figure 1.2 Schematic representation of the role of microorganisms in arsenic cycling in
the environment (Field, unpublished).
1.4.1. Resistance to arsenic toxicity
Certain microbes in arsenic-enriched environments can adapt to and develop several
mechanisms of resistance to arsenic toxicity. Physiologically, large diversity of
microorganisms can utilize As methylation, As(III) oxidation and As(V) reduction as
components of the resistance systems [Oremland and Stolz, 2003; Mukhopahyay et al.,
42
2002; Silver and Phung, 2005] (Figure 1.3). These metabolic systems that confer
resistance to As(III) and As(V) are widely found in live cells of microorganisms encoded
on chromosomes and on small plasmids. Chromosomal and plasmid-based arsenicresistance genes are clustered in an operon, commonly known as the ars operon [Silver
and Phung, 2005]. For example, bacterial plasmids conferring As toxicity resistance
encode specific efflux pumps able to extrude As from the cell cytoplasm, thus lowering
the intracellular concentration of the toxic ions. The genes and enzymes involved in the
resistance systems have been well studied by Silver [Silver, et al., 2004]. In this chapter,
we will just discuss the main processes of inorganic arsenic transformation catalyzed by
microorganisms.
43
Figure 1.3 Phylogenetic diversity of representative arsenic-metabolizing prokaryotics.
Dissimilitory arsenate-respiring prokaryotes (DARPs) are indicated by yellow circles,
Heterotrophic arsenite oxidizer (HOAs) are indicated by triangles, and chemoautotrophic
arsenite oxidizer (CAOs) are indicated by red squares. [Oremland and Stolze, 2003]
44
1.4.2. Microbial reduction of As(V) to As(III)
Microbial reduction of As(V) to As(III) is an important process in the biogeochemical
cycling of As in the environment. Because As(III) is more toxic and soluble than As(V),
reduction of As(V) to As(III) will often contribute to increase the release of As in
reducing environments. There are two strategies for microorganisms to reduce As(V) to
As(III).
The first strategy is As(V) reduction associated with a specific detoxification
mechanisms involving the As(v) reductase protein, ArsC, which catalyzes the reduction
of As(V) to As(III) that is then transported out of the cell by a secondary uniporter (ArsB)
[Mukhopahyay et al., 2002; Silver and Phung, 2005; Inskeep et al., 2007]. ArsC is an
enzyme that reduces As(V) to the more toxic As(III) in the cytoplasm. Reduced
glutathione serve as electron donor for the As(V) reduction. Although As(III) is more
toxic, it serves as substrate for the ArsB transport protein. As(V) conversion seems
counter-productive, but is a mechanism to differentiate As(V) from PO43- and avoid
extrusion of PO4-3 from the cell. ArsC activity is coupled with efflux from the cells
[Silver and Phung, 2005]. The As(III) produced is pumped out from the cytoplasm trough
an adenosine 5’-triphosphate (ATP)-dependent arsenite transported formed by Ar-SaB
[Oremland and Stolz, 2003]. ArsC is also found in many other strict anaerobic bacteria
like Clostridium sp. and Desulfovibrio sp. Macur [Macur et al., 2001] also reported
reduction of pentavalent arsenic in aerobic bacteria isolated from mine tailings. ArsC
45
arsenate reductase is a small monomeric protein (13 to 16 kD), composing of about 135
amino acid residues and containing three essential cysteine residues that are involved in a
cascade sequence of enzyme activity [Oremland and Stolz, 2003; Silver and Phung,
2005]. The arsC genes can be divided mainly into three families according to different
reluctant, including glutaredoxin-glutathione-coupled enzyme, a less-well-defined
glutaredoxin-dependent arsenate reductase and a group of thioredoxin-coupled arsenate
reductases [Silver and Phung, 2005]. Although they share a common biochemical
function but have no evolutionary relationship [Mukhopadhyay et al., 2002]].
The second important reduction mechanism is dissimilatory arsenate reduction.
The dissimilitory respiratory reduction of As(V) is a ubiquitous process that requires a
suitable electron donating substrate. Various anaerobic bacteria, including γ, δ, and
ε−Proteobacteria, low-GC Gram positive bacteria, thermophilic Eubacteria and Archaea
have been discovered that can utilize As(V) as the electron acceptor in respiration [Macy
et al., 2000; Stolz et al., 1999; Oremland and Stolz, 2003; Oremland et al., 2005; Hoeft et
al., 2004]. A wide range of electron donors are also utilized to reduce As(V) to As(III)
including organics, such as acids (e.g., acetate, lactate), alcohols and sugars (e.g., ethanol,
glycerol), aromatics (e.g., syringate, phenol, benzoate, toluene), and inorganic
compounds (e.g., hydrogen, sulfide) [Macy et al., 1996, 2000; Newman, et al., 1997,
1998; Takai et al., 2003; Jensen et al., 2005; Stolz et al., 1999; Oremland et al., 2005;
Hoeft et al., 2004; Liu et al., 2004]. Microbes also gain energy for growth from these
processes of oxidation of electron donors [Oremland et al., 2004]. Microbial respiration
of As(V) is carried out by terminal reductases (AsR). Arsenate reductase have been
46
characterized for some microorganisms e.g. Staphylococcus aureus [Mukhopahyay et al.,
2003], Chrysiogenes arsenatis [Krafft and Macy, 1998], Geospirillum barnessi [Newman
et al., 1998] and Bacillus arsenicoselenatis [Stolz et al., 1999]. The well studied arsenate
respiration systems contains the arrAB operon, which is a heterodimer periplasmic or
membraneassociated protein consisting of a larger molybdopterin subunit (arrA), and a
smaller [Fe-S] center protein (arrB) [Silver and Phung, 2005]. The arrA gene encodes a
molybdenum-containing enzyme within the dimethyl sulfoxide reductase family, which
contains an iron-sulfur center, perhaps a highpotential [4Fe-4S] cluster. ArrB encodes an
iron-sulfur protein. This was demonstrated for the first time by the purification and
characterization of the respiratory arsenate reductase from Chrysiogenes arsenatis [Krafft
and Macy, 1998]. A periplasmic enzyme is heterodimer with an 87-kDa subunit (arrA)
and a 29-kDa subunit (ArrB) [Krafft and Macy, 1998]. Malasarn [Malasarn et al., 2004]
identified a gene for a dissimilatory arsenate reductase arrA and showed that microbial
respiration of As is a more important mechanism for As(V) reduction in the environment.
47
1.4.3. Microbial oxidation of As(III) to As(V)
1.4.3.1. Mechanism of microbial oxidation of As(III) to As(V)
Microbial oxidation of As(III) to As(V) by aerobic heterotrophic microorganisms was
first reported in 1918 [Green et al., 1918]. As(III) can be oxidized to As(V) by a variety
of microbes in the environment, which include more than nine genera, i.e, α, β, and γProteobacteria; Deinocci and Crenarchaeota [Oremland and Stolz, 2003]. Based on the
physiological diversity, the As(III) oxidizers can be divided into two classes:
heterotrophic arsenite oxidizers (HAOs) [Santini et al., 2000; Rhine et al., 2007], which
have been studied for a long time, and chemolithoautotrophic arsenite oxidizers (CAOs)
[Oremland and Stolz, 2003; Inskeep et al., 2007; Stolz et al., 2006], which were described
and have raised interest in recent years.
1.4.3.2. Oxidation of As(III) to As(V) by heterotrophic arsenite oxidizers
Heterotrophic arsenite oxidation was generally considered an important detoxification
process of As(III) in the environment. So far, most isolated As(III)-oxidizing bacteria are
HAOs [Anderson et al., 1992, 2002; Muller et al., 2003; Oremland and Stolz, 2003;
48
Silver and Phung, 2005]. Large numbers of heterotrophic arsenite-oxidizing bacteria have
been isolated from various environments such as freshwater, soil, sediments and marine
[Ehrlich et al., 2002]. Bacillus arsenooxydans was first isolated from an arsenical cattle
dip in South Africa by Green in 1918 [Green et al., 1918]. The mechanisms of As(III)
oxidation by Alcaligenes faecalis and the characteristics of the arsenite oxidase of this
microorganism have been studied in detail [Anderson et al., 1992, 2002]. Subsequently,
15 As(III)-oxidizing bacterial strains were also isolated from cattle dip in Australia, and
all were found to be HAOs [Ehrlich et al., 2002].
Most of HAOs require organic carbon compounds to support microbial growth
and couple the oxidation of As(III) as electron donors, but some HAOs seem to oxidize
As(III) to As(V) as an auxiliary source of energy. The enzyme involved in the oxidation
of As(III) by HAOs is located on the periplasm and it is distinct from the dissimilatory
arsenate reductases [Oremland and Stolz, 2003; Mukhopahyay et al., 2002; Silver and
Phung, 2005]. This enzyme is a mononuclear molybdenum enzyme and belongs to the
dimethyl sulfoxide (DMSO) reductase family of proteins [Mukhopahyay et al., 2002;
Silver and Phung, 2005]. As(III) oxidase genes have been characterized for the following
microorganisms Alcaligenes faecalis, Agrobacterium tumefaciens [Kyshyap et al., 2006],
Cenibacterum arsenoxidans [Muller et al., 2003] and Thermus thermophilus strain HB8
[Gihring et al., 2001]. The arsenite reductase (Aso) shows two sub-units, a larger 88-kDa
polypeptide containing the Mo-pterin, a HiPIP 3Fe-4S center, and a smaller 14-kDa
subunit with the Rieske 2Fe-2S center. The Aso enzyme is located on the outer surface of
49
the inner cell membrane [Anderson et al., 1992]. Oxidation of As(III) to As(V) hinders
the transport of As into microbial cells and reduces microbial toxicity.
1.4.3.3. Oxidation of As(III) to As(V) by chemolithotrophic arsenite oxidizers
In contrast to the HAOs, several As(III)-oxidizing bacteria have been isolated as
chemolithoautotrophs, which can use inorganic carbon to gain growth energy from the
oxidation of As(III). CAOs can utilize As(III) as electron donor to reduce either oxygen
[Ilyaletdinov and Abdrashitova 1981; Santini et al., 2000; Inskeep et al., 2007] or nitrate
[Oremland et al., 2002; Rhine et al., 2006; Hoeft et al., 2007]. The energy produced in
this metabolic process is used to fix carbon dioxide into the organic cellular material and
support cell growth. The arsenite oxidases of autotrophic As(III)-oxidizers are still not
fully understood and characterized, but they appear to be soluble periplasmic enzymes
[Santini et al., 2000]. Although the arsenite oxidases of known CAOs are similar to the
heterotrophic oxidases, the difference in activity and lower sequence homology suggests
that the aroAB genes may comprise a separate group [Santini and Vanden Hoven, 2004;
Mukhopahyay et al., 2002; Silver and Phung, 2005]. As(III) oxidases are heterodimeric
structures with two subunits containing molybdopterin and Rieske iron-sulfur domains
[Rhine et al., 2007; Santini et al., 2004; Stolz et al., 2006]. The full diversity and
distribution of As(III) oxidase genes in environmentally relevant bacteria is not presently
50
known [Silver and Phung, 2005]. Inskeep et al. [Inskeep et al., 2007] and Rhine et al.
[Rhine et al., 2007] recently designed primer sets to examine As(III) oxidase-like genes
(aroA, asoA and aoxB) in environmental samples. The results suggest that a variety of
aerobic As(III) oxidizers are widespread in As-contaminated environments.
The first CAO strain known, Pseudomonas arsenitoxidans, was isolated from a
gold-arsenic deposit (sulfidic gold ore) in the former Soviet Union [IIyaletdinov and
Abdrashitova, 1981]. P. arsenitoxidans has been proven to be an obligate autotroph and
can not oxidize As(III) in the presence of organic carbon. Nine novel species of As(III)
oxidizers were isolated from arsenopyrite from a gold mine located in Australia including
three HAOs (NT-4, NT-5and NT-10 which are all members of β-Proteobacteria) and six
CAOs (NT-2, NT-3, NT-6, NT-14, NT-25 and NT-26 are all members of αProteobacteria) [Santini et al., 2000]. Among these CAOs strains, the NT-26 strain can
grow by using only As(III) as an electron donor, oxygen as an electron acceptor, and
carbon dioxide or bicarbonate as carbon sources. NT-26 also grows with organic carbon
as its carbon source without any inhibition, but the growth rate is greater in the presence
of As(III) than in its absence. Both NT-26 and P. arsenitoxidans are aerobes and utilize
As(III) as the electron donor to oxygen via electron transport chains.
Although aerobic microbial oxidation of As(III) provides energy for rapid growth
of CAOs, this process is would not be functional in subsurface environments where
dissolved oxygen (DO) is absent. However, recently it has come to light that alternative
oxidants, such as nitrate [Oremland et al., 2002] and selenate [Fisher et al., 2008] can be
51
utilized by anaerobic microorganisms to gain energy from As(III). The first indications
that the anoxic oxidation of As(III) occurs in the environment comes from a field study
evaluating the speciation arsenic in relation to nitrate profiles in anoxic lake water
columns [Senn and Hemond, 2002]. At depths where nitrate was depleted, As(III) was
present, but at shallower depths where nitrate was present, As(V) was the dominant
species. Additionally, arsenic was attenuated by injection of nitrate into contaminated
groundwater in Bangladesh [Harvey et al., 2002] but this was potentially attributable to
anoxic biological oxidation of Fe(II) to Fe(III) oxyhydroxides, which in turn adsorbed the
arsenic.
An anoxic chemolithoautotrophic haloalkaliphilic arsenite oxidizing bacterium,
Alkalilimnicola ehrlichii strain MLHE-1, was isolated from an soda lake in California
(Mono Lake), which was capable of linking the oxidation of As(III) to As(V) to partial
denitrification of nitrate to nitrite as shown in equation 1 [Oremland et al., 2002; Hoeft et
al., 2007].
H3AsO3 + NO3- → 2H+ + HAsO42- + NO2-
[eq 1]
Subsequently, two anoxic chemolithoautotrophic strains were isolated from
arsenic polluted industrial soil [Rhine et al., 2006]. These included Azoarcus strain
DAO1 and strain DAO10 which was most closely related to Sinorhizobium. These two
strains metabolized high concentrations of As(III) (5 mM) and the metabolism was linked
to the complete denitrification of nitrate to dinitrogen (N2) gas (equation 2) as was
52
implied from the stoichiometry between As(V) formed and nitrate consumed as well as
the successful amplification of the nitrous oxide gene, nosZ.
5H3AsO3 + 2NO3- → 8H+ + 5HAsO42- + N2 + H2O
[eq 2]
1.5. Environmental significances of biogeochemical cycles of As associated with Fe
Biogeochemical processes affecting the mobility of arsenic by adsorption on the soil and
sediment have been a matter of considerable research interest in order to determine the
factors controlling the release of arsenic into groundwater [Oremland and Stolz, 2003].
Soil and sediments containing oxyhydroxy oxides such as those of aluminum and iron,
adsorb both As(III) and As(V) [Lin and Wu, 2001; Dixit and Hering, 2003] and
effectively attenuate the soluble As concentration from groundwater and surface water.
Compared to As(V), As(III) is less strongly adsorbed by some common-occurring soil
oxides such as aluminum oxyhydroxides [Lin and Wu, 2001; Hering and Dixit, 2005]. In
comparison, Iron oxides show strong adsorption capacity on both As(III) and As(V) in
circumneutral environments [Raven et al., 1998; Dixit and Hering, 2003].
Figure 1.4 illustrates major pathways of ecological importance carried out by
microorganisms resulting in the reduction and oxidation of As and iron. Under anaerobic
conditions, the respiratory reductions of As(V) [Stolz and Oremland, 1999; Oremland
and Stolz, 2003] and Fe(III) [Straub et al., 2001; Lovley et al., 2004] are ubiquitous
53
energy-yielding processes carried out by a diverse group of bacteria utilizing various
electron donating substrates. Release of arsenic in anaerobic conditions can also result
form two main mechanisms: the reduction of As(V) to As(III), which decrease the
sorption capacity on clay minerals containing mostly Al oxides with less iron content
[Zobrist et al., 2000]; and the dissimilatory reductive dissolution of iron oxyhydroxides
[Straub et. al., 2001], which are important sorbents for As(V) and As(III) [Dixit and
Hering, 2003; Anawar et al., 2006]. The latter is generally ascribed as the dominant
means of As displacement from the solid phase.
On the other hand in aerobic environments (Figure 1.4), As(III) is oxidized to
As(V) by a large variety of As(III) oxidizing microorganisms [Inskeep et al., 2007].
Additionally, Fe(II) is oxidized by both chemical and biological activity in the presence
of dissolved oxygen [Moses and Herman, 1989; Rentz et al., 2007]. The iron
oxyhydroxides formed from the reaction strongly adsorb the inorganic As species and
remove them from the liquid phase [Katsoylannis and Zouboulis, 2006]. For example,
freshly precipitated iron(III) oxides, which were formed by oxidation of Fe(II) in the
presence of oxygen, showed a much greater capacity to remove arsenic than adsorption to
the pre-formed Fe(III) particles [Lytle et al., 2005]. The combined results of the microbial
oxidation of As(III) and Fe(II) could enhance the removal of arsenic from groundwater,
decreasing a threat of As contamination in drinking water [Dixt and Hering, 2003;
Katsoylannis and Zouboulis, 2006].
However, the dissolved oxygen (DO) in water may be consumed by various
reducing compounds such as organic matter, sulfides and others in the subsurface
54
environments. These interferes prevent the oxidation of As(III) and Fe(II) by aerobic
microorganisms, finally hamper the As immobilization processes. Gaseous oxygen can be
injected into the anaerobic zone of a contaminated environment to stimulate
bioconversions, but this is costly and inefficient. Nitrate is an example of an ecologically
significant alternative oxidant that can support these anoxic oxidation processes. Nitrate
is very soluble and therefore can potentially occur in subsurface water at concentrations
far exceeding the electron accepting capacity of DO [Nolan et al., 1997]. Since anaerobic
nitrate-dependent ferrous iron oxidation was first observed, it has been well known that
Fe(II) oxidation was coupled to various nitrate reduce pairs including NO3-/ NO2-, NO3/N2 and NO3-/ NH4+ [Straub et al., 2004; Weber et al., 2006] (Figure 1.4). Anoxic
oxidation of As(III) linked to denitrification was also illustrated to widespread in various
arsenic contaminated environments [Oremland et al., 2002; Rhine et al., 2006] (Figure
1.4). Anaerobic microbes can be able to gain growth energy from the anoxic oxidation of
Fe(II) and As(III). The standard reduction potential (E°’) for the redox couple
As(V)/As(III) and Fe(III)/Fe(II) are 0.139 V [Madigan et al., 2006] and 0.77 V [Weber et
al., 2006]; while that for NO3-/N2 is 0.747 V, which equates to a standard change of
Gibb’s free energy (∆G°’) for As(III) oxidation linked to denitrification (eq. 3 and 4) of 117.3 kJ/mol As(III) and -59.9 kJ/mol Fe(II), respectively.
5H3AsO3 + 2NO3- → 5HAsO42- + N2 + 8H+ + H2O
[eq 3]
10Fe(II) + 2NO3- + 12H+ → 10Fe(III)solid + N2 + 6H2O
[eq 4]
55
Thus, the immobilization of arsenic in anoxic environment can be generally
rationalized by two mechanisms related to minerals containing Al or Fe oxides. The first
mechanism is due to the microbial anoxic oxidation of As(III) to As(V) (Figure 1.4),
which increase the adsorption of As on sediments with predominantly aluminum oxides
including clay minerals. It is well established that aluminum oxides have stronger affinity
on As(V) than As(III) under circumneutral conditions [Lin and Wu, 2001; Hering and
Dixit, 2005] Thus, the pre-oxidation of As(III) to As(V) have the potential to enhance the
removal of soluble arsenic on Al oxides.
The second mechanism is involved into the anoxic oxidation of soluble Fe(II) to
form solid–phase iron(hydr)oxides by anaerobic bacteria, and subsequent adsorption of
both As(III) and As(V) on the solid surface (Figure 1.4). The sorption of As(III) and
As(V) has been previously studied that iron oxides were able to attenuate or remediate
the soluble arsenic in groundwater and sediments [Katsoyiannis and Zouboulis, 2006;
Gimenez et al., 2007]. Biological oxidation of iron provided very efficient removal of
arsenic by adsorption of arsenic on biogenic iron oxides, when arsenic was
simultaneously present in the groundwater under neutral pH conditions [Katsoyiannis et
al., 2002; Katsoyiannis and Zouboulis, 2004]. Additionally, the formed schwertmannite
by the rapid oxidation of Fe(II) was found to effectively remove the As(V) released in the
acid mine drainage [Fukushi et al., 2004].
56
Arsenic Cycle
Iron Cycle
arsenite oxidation
Fe(II) Oxidation
O2
NO3-
O2
O
OH
HO
As
HO As OH
OH
OH
arsenate, As(V)
arsenite, As(III)
NO3-
As(V) As(V)
Fe(OH)3
Fe(II) + As(V)
As(V) As(V)
aqueous
solid
organic matter
organic matter
arsenate reduction
Fe(III) reduction
Figure 1.4 Biogeochemical redox cycles of arsenic and iron as well as their interaction.
The anoxic processes of As(III) and Fe(II) oxidation are the missing link to the As
biogeochemical cycle. They would account for the redox cycling of As and Fe in absence
of oxygen. This of As(III) to As(V) increase the As retention in soil and sediments
containing non-iron metal oxides. Furthermore, the oxidation of Fe(II) could generate
iron (hydr)oxides and immobilize As on the solid phases. Due to the low level of oxygen
in the saturated subsurface, nitrate could be utilized as an alternative electron acceptor
with advantages of high solubility, less reactive enabling it stimulate the conversions. The
microbial nitrate-dependent oxidations of Fe(II) and As(III) have the great potential to
enhance the immobilization of arsenic in the anoxic environments.
57
1.6 Objectives
The objective of this study is to evaluate the importance of anoxic oxidation of As(III) to
As(V) by anaerobic microorganisms such as chemolithotrophic denitrifying bacteria and
chlorate respiring bacteria in the biogeochemical cycle of arsenic. This study is also to
investigate the potential bioremediation strategy of arsenic by injecting nitrate or chlorate
into contaminated groundwater and surface water under anaerobic conditions. Three
types of biogeochemical scenarios will be distinguished:
i)
The anoxic microbial oxidation of As(III) linked to chemolithotrophic
denitrification or chlorate reductions in batch and continuous flow-through
anaerobic environments.
ii)
The role of denitrification on As(III) oxidation and arsenic mobility in anoxic
sediment column model with activated aluminum, independent of iron oxides.
iii)
Anoxic microbial oxidation of As(III) and Fe(II) linked to chemolithotrophic
denitrification for the immobilization of As in anoxic environments.
The central question addressed in this research is whether microbial anoxic
oxidation of As(III) is a ubiquitous processes in anaerobic environments. The dissertation
will explore if the microbial oxidation of As(III) linked to denitrification or chlorate
reduction occur at significant rates, as well as steady over prolong period in the
continuous bioreactor with bacteria immobilized in bio-film granules. Likewise, the
58
project will attempt to identify and characterize the diversity of microorganisms involved
and the nature of the chemolithotrophic reactions they carry out. Another pertinent
question is whether a bioremediation strategy by injecting NO3- is able to support the
microbial oxidation of Fe(II) and As(III) in the subsurface as a means to immobilize
arsenic. Two models were utilized to illustrate the As removal mechanisms: 1) sediment
columns packed with activated aluminum were utilized to demonstrate the role of nitrate
in supporting microbial As(III) oxidation and arsenic mobility in anoxic sediments
containing mostly non-iron oxides; 2) sand-packed columns were used to simulate natural
anaerobic groundwater and sediment system with co-occurring As(III) and Fe(II) in the
presence or absence of nitrate. Microbial oxidation by denitrifying microorganisms lead
to the formation of iron (hydroxides) which adsorbed As(V) formed from As(III)oxidation.
59
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66
CHAPTER 2
ANOXIC OXIDATION OF ARSENITE LINKED TO DENITRIFICATION IN
SLUDGES AND SEDIMENTS
2.1. Abstract
In this study, denitrification linked to the oxidation of arsenite (As(III)) to arsenate
(As(V)) was shown to be a widespread microbial activity in anaerobic sludge and
sediment samples that were not previously exposed to arsenic contamination. When
incubated with 0.5 mM As(III) and 10 mM NO3-, the anoxic oxidation of As(III)
commenced within a few days, achieving specific activities of up to 1.24 mmol As(V)
formed g-1 volatile suspended solids d-1 due to growth (doubling times of 0.74 to 1.4 d).
The anoxic oxidation of As(III) was partially to completely inhibited by 1.5 and 5.0 mM
As(III), respectively. Inhibition was minimized by adding As(III) adsorbed onto activated
alumina (AA). The oxidation of As(III) was shown to be linked to the complete
denitrification of NO3- to N2 by demonstrating a significantly enhanced production of N2
beyond the background endogenous production as a result of adding As(III)–AA to the
cultures. The N2 production corresponded closely the expected stoichiometry of the
reaction, 2.5 mol As(III) mol-1 N2-N. The oxidation of As(III) linked to the use of
commonly occurring nitrate as an electron acceptor may be an important missing link in
the biogeochemical cycling of arsenic.
Keywords: Chemolithotrophic, Arsenic, Arsenate, Nitrate, Anaerobic, Biotransformation
67
2.2. Introduction
The occurrence of arsenic in drinking water is a global public health concern.
Epidemiological studies have provided compelling evidence for a link between the
chronic exposure to arsenic in drinking water and several types of cancer and other
medical disorders [ATSDR, 2007]. The evidence has prompted various environmental
agencies to lower the maximum contaminant level in drinking water from 50 to 10 µg l-1
[USEPA, 2001; WHO, 1993]. The main anthropogenic sources of arsenic to the
environment are from liberation of arsenic by mining [Roussel, 2000], use of arsenical
pesticides [Bednar, 2002], and treatment of preserved wood with arsenic [Katz, 2005].
However, the natural release of arsenic to groundwater from the weathering of arsenic
bearing rocks and sediments is the most important source from a global public health
perspective [Smedley and Kinniburgh, 2002].
Microorganisms play a key role in the speciation and mobility of arsenic in the
environment [Oremland and Stolz, 2003]. In anaerobic environments, a biodiverse group
of dissimilatory arsenate reducing bacteria utilize arsenate (As(V)) as a terminal electron
acceptor and reduce it to arsenite (As(III)). As(V) reduction to As(III) is also microbially
catalyzed by the ars operon as a detoxification mechanism which pumps arsenic out of
the cell [Silver and Phung, 2005]. Compared to As(V), As(III) is less strongly adsorbed
by some common-occurring soil oxides such as aluminum oxyhydroxides [Hering and
Dixit, 2005]. Release of arsenic in anaerobic conditions can also result form the
dissimilatory reductive dissolution of iron oxyhydroxides [Straub et. al., 2001], which are
68
important sorbents for As(V) and As(III) [Dixit and Hering, 2003]. The combined result
of the microbial reduction of arsenic and iron is the enhanced mobility of arsenic in
groundwater, posing a threat of arsenic contamination in drinking water [Anawar et al.,
2006; Smedley and Kinniburgh, 2002]. On the other hand in aerobic environments,
As(III) is oxidized to As(V) by a large variety of heterotrophic and autotrophic As(III)
oxidizing bacteria [Inskeep et al., 2007]. Additionally, Fe(II) is oxidized by both
chemical and biological activity in the presence of dissolved oxygen [Moses and Herman,
1989; Rentz et al., 2007]. The iron oxyhydroxides formed from the reaction strongly
adsorb the inorganic arsenic species and remove them from the liquid phase
[Katsoylannis and Zouboulis, 2006]. Thus, oxidizing conditions provided in aerobic
environments lower the risk of arsenic mobility and groundwater contamination.
The maximum solubility of dissolved oxygen (DO) in water equilibrated with air
is approximately 9 mg l-1. The modest inventory of DO in water may be consumed by
organic matter, sulfides and other reducing compounds in the subsurface preventing the
oxidation of As(III) and Fe(II) by aerobic microorganisms. However, recently it has come
to light that alternative oxidants can be utilized by anaerobic microorganisms to gain
energy from As(III) [Oremland et al., 2002] and Fe(II) oxidation [Straub et al., 2001].
Nitrate is an example of an ecologically significant alternative oxidant that can support
these anoxic oxidation processes. Nitrate is very soluble and therefore can potentially
occur in subsurface water at concentrations far exceeding the electron accepting capacity
of DO. The first indications that the anoxic oxidation of As(III) occurs in the
environment comes from a field study evaluating the speciation arsenic in relation to
69
nitrate profiles in anoxic lake water columns [Senn and Hemond, 2002]. At depths where
nitrate was depleted, As(III) was present, but at shallower depths where nitrate was
present, As(V) was the dominant species. Additionally, arsenic was attenuated by
injection of nitrate into contaminated groundwater in Bangladesh [Harvey et al., 2002]
but this was potentially attributable to anoxic biological oxidation of Fe(II) to
oxyhydroxides, which in turn adsorbed the arsenic.
An anoxic chemolithoautotrophic haloalkaliphilic arsenite-oxidizing bacterium,
Alkalilimnicola ehrlichii strain MLHE-1, was isolated from an soda lake in California
(Mono Lake), which was capable of linking the oxidation of As(V) to As(III) to partial
denitrification of nitrate to nitrite as shown in equation 1 [Oremland et al., 2002; Hoeft et
al., 2007].
H3AsO3 + NO3- → 2H+ + HAsO42- + NO2-
[eq 1]
Subsequently, two anoxic chemolithoautotrophic strains were isolated from
arsenic polluted industrial soil [Rhine et al., 2006]. These included Azoarcus strain
DAO1 and strain DAO10 which was most closely related to Sinorhizobium. These two
strains metabolized high concentrations of As(III) (5 mM) and the metabolism was linked
to the complete denitrification of nitrate to dinitrogen (N2) gas (equation 2) as was
implied from the stoichiometry between As(V) formed and nitrate consumed as well as
the successful amplification of the nitrous oxide gene, nosZ.
70
5H3AsO3 + 2NO3- → 8H+ + 5HAsO42- + N2 + H2O
[eq 2]
In this study, we report on the anoxic biological oxidation of As(III) by
denitrifying microorganisms in sludges and sediments with no prior exposure to arsenic.
The study demonstrates that the biological capacity to link As(III) oxidation to
denitrification is widespread in anaerobic environmental samples. However, the activity
is sensitive to high concentrations of As(III). The production of N2 gas linked to As(III)
oxidation was demonstrated by direct measurements.
2.3. Material and Methods
2.3.1. Microorganisms
Sludge and sediment samples obtained from different locations were used as inocula in
the batch bioassays. Aerobic activated sludge (RAS) and anaerobically digested sewage
sludge (ADS) were obtained from a local municipal wastewater treatment plant (Ina
Road, Tucson, Arizona). Methanogenic granular sludge (biofilms pellets) samples were
obtained from industrial upward-flow anaerobic sludge blanket (UASB) treatment plants
treating recycled paper wastewater (EGS) (Industriewater, Eerbeek, The Netherlands)
and alcohol distillery wastewater (NGS) (Nedalco, Bergen op Zoom, The Netherlands).
Chemolithotrophic denitrifying granular sludge was obtained from a laboratory-scale
71
thiosulfate-oxidizing
denitrifying
enrichment
bioreactor
(TDE)
(Chemical
and
Environmental Engineering Department, University of Arizona). Duck pond sediments
were obtained at the Agua Caliente Park (DPS) (Tucson, Arizona). Additional sediments
were also collected from Pinal Creek (PCS) (Arizona) and from a Winogradsky column
(WCS) (Chemical and Environmental Engineering Department, University of Arizona)
originally inoculated with a mixture of cattle manure lagoon sludge mixed with creek
sediments obtained in Patagonia, Arizona. The total suspended solids’s (TSS) contents of
the sludge and sediment samples was 6.32 ± 0.27, 17.64 ± 1.12, 79.2 ± 0.7; 74.48 ± 4.01,
16.65 ± 0.21 and 1.63 ± 0.09% wet weight basis respectively; and the volatile suspended
solids’s (VSS) content was 5.96 ± 0.28, 12.91 ± 0.97, 0.72 ± 0.01, 0.67 ± 0.07, 10.03 ±
0.1 and 1.12 ± 0.07% wet weight basis for NGS, EGS, DPS, PCS, ADS and RAS,
respectively. The granular sludge samples were washed and sieved before use to remove
fines. The inocula were stored under nitrogen gas at 4°C.
2.3.2. Basal medium
The standard basal medium was prepared using ultra pure water (Milli-Q system;
Millipore) and contained the following compounds (mg l-1): NH4HCO3 (3.16); NaHCO3
(672); CaCl2 (10), MgSO4.7H2O (40); K2HPO4 (300); KH2PO4.2H2O (800); and 0.2 ml l-1
of a trace element solution containing (in mg l-1): FeC13.4 H20 (2000); CoCl2. 6 H20
72
(2000); MnCl2 4 H20 (500); AlCl3 6 H20 (90); CuCl2.2H20 (30); ZnCl2 (50); H3BO3 (50);
(NH4)6Mo7O24.4 H2O (50); Na2SeO3.5 H2O (100); NiCl2.6 H20 (50); EDTA (1,000);
resazurin (200); HCl 36% (1 ml).
2.3.3. Batch bioassay
Batch bioassays were performed in shaken flasks, which were incubated in a dark
climate-controlled room at 30±2˚C. Serum flasks (160 ml) were supplied with 120 ml of
a basal mineral medium (pH 7.0-7.2) containing bicarbonate as the only carbon source, as
described above. The medium was also supplemented with arsenite (As(III)) as electron
donor (concentration indicated in Tables and Figures of each experiment) and nitrate as
the electron acceptor, typically 10 mM, unless otherwise specified. Flasks for anoxic
assays were sealed with butyl rubber stoppers, and then the medium and headspace were
purged with N2/CO2 (80/20, v/v) for 20 min to exclude oxygen from the assay. Various
controls (e.g. abiotic controls, killed sludge controls, controls without electron acceptor,
etc) were applied based on the requirements of each experiment. Abiotic controls were
prepared without adding microbial inoculum. Killed sludge controls were prepared by
autoclaving the flasks added with inoculum for 20 min at 121ºC, the content was allowed
to cool down overnight, and the cycle was repeat two more times and then sealed
aseptically. Controls lacking As(III) were included to measure endogenous consumption
73
of nitrate and correct the net nitrate reduction linked to As(III) oxidation. All assays were
conducted in triplicate. Liquid samples were analyzed for the concentration of electron
acceptor and biotransformation products (NO3- and NO2-) and arsenic species (As(III)
and As(V)). Granular sludge and sediment inoculum was added to the assays at 10-12
wet weight g l-1 medium; RAS and ADS inoculum was added to the assays at 6% (v/v).
2.3.4. Batch assays to determine the terminal product of denitrification
End products of denitrification were measured by monitoring gaseous nitrogen species in
the headspace of bioassays flushed with He/CO2 (80/20, v/v) in lieu of N2/CO2 as
described above. Gaseous nitrogen in As(III) spiked samples (3.5 mM) was compared to
endogenous controls. Microbial reduction of NO3- coupled to As(III) oxidation was
assessed in shaken batch bioassays inoculated with NGS and DPS.
The NO3-
concentration utilized was 4.0 and 5.0 mM for the experiments with NGS and DPS,
respectively. Some samples were supplied 50 g l-1 of either activated aluminum (AA400G, Alcan Bauxite and Alumina) or TiO2 (Dow Chemicals) adsorbents to lower the
effective aqueous concentration of As(III) so as to minimize its toxicity. Headspace
samples were analyzed periodically for N2 and N2O with a pressure lock gas tight syringe
(1710RN, 100 µl (22s/2‫״‬/2), Hamilton Company) to confirm denitrification. Liquid
74
samples were analyzed for NO3- and NO2-. Flushed headspace controls incubated with
just water were monitored to ensure background levels of N2 were low.
2.3.5. Analytical methods
All liquid samples were centrifuged (10 min., 10,000 rpm) or membrane filtered (0.45
µm) immediately after sampling and stored in polypropylene vials. Samples for arsenic
analysis were then stored at -20°C. As(III) and As(V) were analyzed by ion
chromatography–inductively coupled plasma–mass spectroscopy (HPLC-ICP-MS). The
system consists of an HPLC (Agilent 1100) and an ICP-MS (Agilent 7500a) with a
Babington nebulizer as the detector. The operating parameters were as follows: Rf power
1500 watts, plasma gas flow 15 l min-1, carrier flow 1.2 l min-1, arsenic was measured at
75 m/z and terbium (IS) measured at m/z 159. The injection volume was 10 µl. The
detection limit for the various arsenic species was 0.1 µg l-1. The total concentration of
arsenic in liquid samples was determined using an ASX500 auto sampler (CETAC
Technologies, Omaha, NE) and an Agilent 7500a ICP-MS. The analytical system was
operated at an Rf power of 1500 W, a plasma gas flow of 151 min-1 and a carrier gas flow
of 1.21 min-1. The acquisition parameters used were as follows: arsenic measured at m/z
75; terbium (IS) measured at m/z 159; three points per peak; 1.5 s dwell time for As, 1.5 s
dwell time for Tb; number of repetitions =7.
75
The total arsenic content and arsenic speciation in solid matrices (i.e., activated
alumina, biomass) were measured following extraction of samples with 25 ml of NaOH
(2.0 M) in anaerobic tubes (under nitrogen) immersed in a shaking water bath at the
temperature of 90±2°C for 12 h. All liquid samples were adjusted to pH 6-9 with HNO3
(2.5 M) and membrane filtered (0.45 µm) immediately after sampling and stored in
polypropylene vials at -20°C.
All liquid samples were centrifuged or membrane filtered (0.45 µm) immediately
after sampling and stored in polypropylene vials at 4°C. Nitrate, nitrite and arsenate
(As(V)) were analyzed by suppressed conductivity ion chromatography using a Dionex
500 system (Sunnyvale, CA, USA) fitted with a Dionex IonPac AS11 analytical column
(4 mm x 250 mm) a AG16 guard column (4 mm x 40 mm). During each injection, the
eluent (KOH) used was 20 mM for 20 min. All liquid samples were centrifuged or
membrane filtered (0.45 µm) immediately after sampling and stored in polypropylene
vials. Samples for arsenic analysis were then stored at -20°C until the analysis was
performed in order to reduce changes in arsenic speciation. Other filtered samples were
stored at 4°C.
N2 and N2O were analyzed using a Hewlett Packard 5890 Series II gas
chromatograph fitted with a CarboxenTM 1010 Plot column (30 m x 0.32 mm) and a
thermal conductivity detector. The temperatures of the column, the injector port and the
detector were 220, 110 and 100°C, respectively. Helium was used as the carrier gas and
the injection volume was 100 µl.
76
Other analytical determinations (e.g., pH, TSS, VSS, etc.) were conducted according to
Standard Methods [APHA, 1999].
2.4. Results
2.4.1. Screening
Sediments and sludge samples from environments not known to be contaminated with
arsenic were incubated with As(III) (0.5 mM) either in the presence and absence of NO3(10 mM) and were incubated in the absence of elemental oxygen. An example time
course from these incubations is provided in Figure 1 for DPS and NGS. In treatments
incubated with NO3-, As(V) formation started after approximately 2 d and was
concomitant with the disappearance of As(III). After approximately 1 week the reaction
was complete. The reactions were dependent on the presence of NO3- as evidenced by the
lack of any significant conversion in incubations lacking NO3-. Likewise, the reactions
were also dependent on being inoculated, since no reaction occurred in bottles receiving
the medium with As(III) and NO3- but lacking sediment or sludge.
The results for all the sludge and sediments samples screened are shown in Table
1. Six out of the eight inocula tested displayed microbial activity towards the anoxic
oxidation of As(III). Five out of the six of the positive samples, showed a dependency on
the presence of nitrate for the anoxic oxidation of As(III). One of the positive samples,
PCS, had this activity both in the presence and absence of added NO3- which may have
77
been due to high levels of oxidized manganese species known to occur in that sediment
[Lind and Hem, 1993]. The average molar yield of nitrate linked As(III) oxidation was
0.943±0.040 mol As(V) formed mol-1 AS(III) consumed and the value for the PCS
sediment were similar.
0.6
A
As(III) (mM)
0.5
0.4
0.3
0.2
0.1
0.0
0
2
4
6
8
10
0.6
B
0.4
0.0
Ln(As (V)) (mM)
As(V) (mM)
0.5
0.3
0.2
y = 0.9323x - 5.0972
2
R = 0.9972
-1.0
-2.0
-3.0
y = 0.6089x - 4.0626
2
R = 0.9741
-4.0
0
0.1
1
2
3
Time (Days)
4
5
0.0
0
2
4
6
8
10
Time (Days)
Figure 2.1 The removal of As(III) (panel A) and formation of As(V) (panel B) by NGS
and DPS under denitrifying condition, NGS and DPS with NO3- (■) and (▲), NGS and
DPS without NO3- (□) and (), abiotic (○). Insert shows natural logarithm plot of data
during exponential growth phase.
78
Table 2.1 Summary of microbial As(III) oxidation under denitrifying conditions
As(V) formation ‡
Inoculum
†
Time†
Number
Sources
With NO3-
Without NO3-
(Days)
1
NGS
0.423±0.004
__
6
2
EGS
__
__
__
3
ADS
0.425±0.003
__
13
4
RAS
__
__
__
5
TDE
0.416±0.002
__
10
6
DPS
0.415±0.003
__
6
7
WCS
0.413±0.001
__
5
8
PCS
0.295±0.020
0.298±0.014
>14+
: Time to oxidize 80% of 0.5 mM As(III) to As(V)
‡
Conversion of As(III) to As(V): “Average ± STDE” for “conversion ≥ 80%”; “__ “ for
“conversion ≤ 5%”
As(V) formation =concentration of As(V) (at the end of the experiment, d10 or d14)
minus concentration of As(V) (at day 0)
+
: In PCS, at day 14 only 60% of As(III) was converted to As(V)
79
2.4.2. Kinetics
The experiments conducted with NGS, DPS and ADS were monitored with sufficient
sampling times to obtain kinetic data. The growth rates and maximum specific activity
normalized to volatile suspended solids (VSS) in the sludge or sediment sample are
summarized in Table 2. The growth rate, estimated from the slope of the natural
logarithm of As(V) formation versus time, indicated doubling times of around 1 d. The
maximum specific activities ranged from 0.16 to 1.2 mmol As(V) formed g-1 VSS d-1.
Table 2.2 Summary of kinetics† of microbial As(III) oxidation (0.5 mM) under denitrifying conditions
Highest specific activity
Inoculum sources
*
Doubling time (days)
(mmol As (V) g-1 VSSadded d-1)
†
*
NGS
0.744±0.027
1.243±0.045
ADS
1.041±0.028
0.160±0.004
DPS
1.145±0.106
0.609±0.060
: Estimated from As(V) formation data.
The coefficient of determination (R2) was 0.9972, 0.9804 and 0.9741 for the
ln(∆As(V)) versus time plots of NGS, ADS and DPS
80
2.4.3. Toxicity
The anoxic oxidation of As(III) was inhibited by incremented As(III) concentrations as
illustrated in Figure 2. Compared to the kinetic parameters observed at 0.5 mM As(III),
the maximum specific activities and growth rates were inhibited by approximately 50% at
1.24 mM As(III), and 90% at 3.5 mM As(III) in assays utilizing NGS as inoculum. For
the most part, no activity was observed at 5 mM As(III) or higher. Albeit, that inhibition
was severe, As(III) could be completely biologically transformed at 3.5 mM by anoxic
oxidation over long incubation times (Figure 3).
Rate of As(V) formation
(mmol/g VSS*d)
1.2
0.9
0.6
0.3
0.0
0.0
2.0
4.0
6.0
Initial As(III) concentration (mM)
Figure 2.2 Substrate toxicity as shown by a decrease in the rate of As(V) formation (♦)
from the oxidation of As(III) at different initial As(III) concentrations.
5
5
4
4
3
3
2
2
1
1
0
0
0
4
12
19
26
30
34
40
As(V) formation (mM)
As(III) elimination (mM)
81
48
Time (days)
Figure 2.3 Elimination of As(III) and formation of As(V) by NGS under denitrifying
condition, As(III) removal (■), As(V) formation (□), abiotic control for As(III) (●) and
As(V)(○).
2.4.4. Consumption of NO3- and formation of N2 linked to As(III) oxidation
In environmental samples of sludges and sediments, the measurement of NO3consumption and N2 formation linked to As(III) oxidation is a difficult task. The
inhibition of As(III) limits the concentration that can be used, such that the electron
equivalents provided by As(III) are relatively small compared to the endogenous
substrates in the environmental sample. Therefore, a distinction between denitrification
attributable to As(III) oxidation and that which is attributable to endogenous substrate
utilization is infeasible. A distinction can potentially be made if a higher supply of As(III)
82
is introduced without causing inhibition. In this study, this was achieved by using weak
adsorbents of As(III), activated alumina (AA) and TiO2 to slowly release As(III) over the
course of the bioassay. By including 50.0 g dwt l-1 of AA or TiO2, together with As(III),
supplied at a rate of 3.5 mmol l-1; the equilibrium concentration of As(III) was 0.65 and
0.63 mM, respectively. Based on the NO3-, N2O-, N2O and N2 data (Figures 4 and 5), the
adsorbents dramatically decreased the toxicity of As(III) to denitrification. When 3.5 mM
As(III) was introduced directly, the inhibited denitrification resulted in a major
accumulation of N2O (0.95±0.07 or 1.62±0.12 mmol N2O-N l-1 in different experiments)
instead of complete denitrification to N2, as well as a longer period of NO2- accumulation.
On the other hand, when sorbents were utilized, the NO3- consumption was more rapid,
there was a shorter period of NO2- accumulation and the denitrification was complete to
N2. The molar ratios of As(III) consumption and As(V) formation to the net NO3consumption and N2-N formation are provided in Table 3. The molar ratios of As(III)
removal to NO3- consumption and N2-N formation from the amended adsorbent by the
DPS and NGS cultures were close to the theoretical ratio of 2.5, which is expected for
As(III) oxidation linked to complete denitrification to N2.
83
2.0
N 2 -N (mmol/L liquid )
A
1.5
1.0
0.5
0.0
0
10
20
30
40
50
60
10
20
30
40
50
60
10
20
30
40
Time (days)
50
60
N 2 O-N (mm ol/L liquid )
2.0
B
1.5
1.0
0.5
0.0
Delta NO 3 -/NO 2 - conc.(mM)
0
2.0
C
1.5
1.0
0.5
0.0
0
Figure 2.4 Denitrification at a high initial As(III) aqueous concentration (3.5 mM). The
formation of N2 (panel A) and N2O (panel B) by NGS. Legend for panels A and B: Full treatment
with As(III) and NO3- (♦); endogenous control with only NO3- added (◊); and the abiotic control
with As(III) and NO3- (○). Panel C shows the consumption of NO3- and the temporal
accumulation of NO2-. Legend for panel C: NO3- consumption in the full treatment with As(III)
and NO3- (▲); NO3- consumption in the endogenous control with only NO3- added (♦); NO3consumption in the abiotic control with As(III) and NO3- (●); NO2- accumulation in the full
treatment with As(III) and NO3- (∆); NO2- accumulation in the endogenous control with only
NO3- added (◊);NO2- accumulation in the abiotic control with As(III) and NO3- (○).
84
2.0
N2-N (mmol/Lliquid)
A
1.5
1.0
0.5
0.0
0
10
20
30
40
50
60
10
20
30
40
50
60
2.0
Delat NO3-/NO2- conc.(mM)
C
1.5
1.0
0.5
0.0
0
Time (days)
Figure 2.5 Denitrification with As(III) (3.5mM) adsorbed to activated aluminum. The
formation of N2-N (panel A) by NGS. Legend for panel A: Full treatment with As(III)
and NO3- (▲); endogenous control with only NO3- added (∆); and the abiotic control
with As(III) and NO3- (○); panel B shows the consumption of NO3- and the temporal
accumulation of NO2-. Legend for panel B: NO3- consumption in the full treatment with
As(III) and NO3- (▲); NO3- consumption in the endogenous control with only NO3added (♦); NO3- consumption in the abiotic control with As(III) and NO3- (●); NO2accumulation in the full treatment with As(III) and NO3- (∆);NO2- accumulation in the
endogenous control with only NO3- added (◊); NO2- accumulation in the abiotic control
with As(III) and NO3- (○). No N2O-N was observed in any treatment.
85
Table 2.3 Summary of sorbed As(III) oxidation linked to denitrification
NGS
DPS
AA
Parameters
AA
TiO2
As(III)+NO3
-
NO3- only
As(III)+NO3
-
NO3- only
As(III)+NO3
-
NO3- only
As(III) fed (mM)
3.52±0.08
__
3.55±0.06
__
3.55±0.06
__
NO3- consumed (mM)
1.84±0.01
0.58±0.03
3.37±0.01
1.94±0.10
2.72±0.14
1.13±0.07
Corrected† NO3- consumed (mM)
1.27±0.04
__
1.43±0.10
__
1.59±0.14
__
N2-N formed (mmol/Lliquid)
1.84±0.05
0.47±0.01
3.21±0.03
1.60±0.17
2.53±0.13
0.98±0.09
Corrected N2-N formed (mmol/Lliquid)
1.37±0.06
__
1.60±0.15
__
1.55±0.19
__
Corrected N2-N formed/corrected NO3consumed
1.09±0.07
__
1.12±0.13
__
0.97±0.04
__
Mol As(III) fed/corrected mol NO3- consumed 2.78±0.12
__
2.48±0.14
__
2.24±0.21
__
__
2.23±0.23
__
2.32±0.31
__
Mol As(III) fed/corrected mol N2-N formed
†
2.57±0.07
: Corrected for the endogenous nitrate consumption or N2-N formation determined in the nitrate treatment lacking added
As(III).
86
2.5. Discussion
2.5.1. Evidence of bioconversion
In this study, the capacity of microorganisms to form diverse anaerobic samples of
sludges and sediments to utilize nitrate as an electron acceptor for the oxidation of As(III)
was demonstrated. The biological nature of the reaction is inferred from the lack of any
conversion in uninoculated samples, heat killed samples and the inhibition of the reaction
at high As(III) concentrations. Likewise, kinetic measurements indicate an exponential
increment in the conversion rate with time, which is consistent with microbial growth.
The biological conversion of As(III) to As(V) in the absence of O2, was
dependent on the presence of NO3-. No As(III) oxidation occurred in the inoculated
incubations without NO3- addition, except for one sediment sample known to contain
manganese oxides. The presence of As(III), supplemented in a non-inhibitory fashion,
greatly enhanced N2 production beyond the background endogenous production,
suggesting its role as an electron donor to the microbial reaction. The stoichiometry of
the As(III) conversion to As(V) in relation to NO3- conversion to N2 clearly suggests
As(III) oxidation was linked to the complete denitrification to N2 as indicated in equation
(2). The experimentally measured ratios approximated 2.5 mol arsenic metabolized mol-1
NO3- converted as is expected from the equation. While the microbial activity in this
study was demonstrated in pristine samples, there is at least one reported precedent from
arsenic contaminated soils for the enrichment and isolation of two denitrifying bacterial
87
strains capable of oxidizing As(III) at the expense of NO3- reduction [Rhine et al., 2005].
Likewise, a partial denitrifying bacterium capable of oxidizing As(III) was isolated form
a soda lake [Oremland et al., 2002].
2.5.2. The occurrence of anoxic As(III) oxidizing bacteria
Microorganisms capable of gaining energy from the oxidation of inorganic chemicals are
widespread in the environment, including anoxic environments. There are ample
examples of chemolithotrophic denitrification utilizing H2 [Lee and Rittmann, 2002],
Fe(II) [Straub et al., 2001], S0 [Sierra-Alvarez et al., 2007], H2S [Cardoso et al., 2006],
and U(IV) [Beller, 2005] as electron donors among others. Nitrate is also a common
alternative electron acceptor found in groundwater and lakes that could potentially
support these reactions [Nolan et al., 1997; Senn and Hermond, 2002]. Energy for
microbial growth can also be gained from the anoxic oxidation of As(III), The standard
reduction potential (E°’) for the redox couple As(V)/As(III) is 0.139 V [Madigan and
Martinko, 2006]; while that for NO3-/N2 is 0.747 V which equates to a standard change of
Gibb’s free energy (∆G°’) for As(III) oxidation linked to complete denitrification (eq.
(2)) of -117.3 kJ mol-1 As(III).
With respect to arsenic, the environmental samples used in this study can be
regarded as pristine samples. The spring that feeds the pond from where DPS was
collected has a measured arsenic concentration of 67 nM [RECON, 2002]. The effluent
from the wastewater treatment plant where ADS was collected has a measured arsenic
88
concentration below the detection limit of 52 nM [PAG, 2002]. NGS was collected from
wastewater originating from a distillery utilizing sugar beet molasses as a feedstock.
Sugar beet molasses is reported to have an arsenic concentration of 2.4 µmol kg-1 [Skrbic
and Durisic-Mladenovic, 2005], the actual concentration in the wastewater would be
lower due to dilution of the feedstock into the process water. TDE was an enrichment
culture fed with medium prepared with milliQ water to which no arsenic was added.
Nonetheless, microbial activity towards anoxic oxidation of As(III) developed rapidly in
these inoculum samples. The microbial community had a moderate growth rate with
doubling times ranging from 0.74 to 1.34 d. Once the conditions of As(III) and NO3- were
provided, responsible organisms grew and established a large enough population to
account for significant As(III) activity in the microcosms, ranging from 0.16 to 1.24
mmol As g-1 VSS d-1. The presence of the anoxic arsenite oxidizers in the anaerobic
samples lacking arsenic is probably due to the metabolic versatility of organisms
responsible for the reaction. As an example, one of the previously reported anoxic
As(III)-oxidizing isolates, Azoarcus sp. strain DAO1 [Rhine et al., 2006], is from a genus
well known for denitrification utilizing diverse substrates. The substrate spectrum
includes simple organic acids, amino acids, sugars, aromatic acids, phenolics and
aromatic hydrocarbons [Rabus, 2005; Reinhold-Hurek and Hurek, 2006; Rhine et al.,
2006]. Aside from these heterotrophic carbon sources, the CO2 fixing gene for ribulose1,5′-bisphosphate carboxylase/oxygenase was identified in Azoarcus sp. strain DAO1,
indicating the capability for autotrophic metabolism [Rhine et al., 2006].
89
2.5.3. As(III) substrate inhibition
The previous attempts to isolate anoxic As(III)-oxidizing bacteria have been from sites
containing high concentrations of arsenic. For example, the soda lake from where
Alkalilimnicola ehrlichii strain MLHE-1 was isolated contains 0.2 mM arsenic
[Oremland et al., 2002]. Thus investigators utilized high concentrations of arsenic for the
enrichment and isolation. The initial enrichments leading to the isolation of Azoarcus sp.
strain DAO1 and Alkalilimnicola ehrlichii strain MLHE-1 contained 5 mM As(III)
[Oremland et al., 2002; Rhine et al., 2006]. Such high concentrations of As(III) are highly
toxic to microorganisms [Stasinakis et al., 2003]. The 50% inhibitory concentrations of
As(III) to methanogenic activity are reported to be as low as 15 µM [Sierra-Alvarez et al.,
2004]. A concentration of 5 mM As(III) was lethal to the As(III) oxidizing microbial
communities described in this study, and concentrations from 1 to 3.5 mM caused partial
inhibition compared to kinetic rates observed at 0.5 mM. Therefore, the previous studies
selected for anoxic As(III)-oxidizing microorganisms with remarkable resistance to
As(III). A common resistance mechanism is afforded by the ars operon which pumps
arsenic out of cells [Silver Phung, 2005]. By inhibiting microorganisms with low As(III)
resistance, the previous studies potentially overlooked the biodiversity of anoxic As(III)
oxidation. The environmentally relevant concentrations in the subsurface are typically
well below the toxicity range to microorganisms [Cullen and Reimer, 1989; Smedley and
Kinniburgh, 2002], and therefore anoxic As(III) oxidizers that would thrive at low As(III)
90
concentrations are highly relevant to the speciation and mobility of arsenic in the
environment.
2.5.4. Sorbed As(III) as a substrate for denitrification
Sediments contain oxyhydroxy oxides such as those of aluminum, which adsorb As(III)
[Lin and Wu, 2001] and attenuate the effective soluble arsenic concentration. In the
presence of AA and TiO2, As(III) was adsorbed and the equilibrium soluble
concentrations was approximately six-fold lower than treatments without adsorbents,
lowering the inhibitory impact of As(III). The adsorbed As(III) was effectively utilized
as an electron donor for denitrification to N2 gas. In parallel incubations without
adsorbents, N2O gas accumulated instead due to the severe toxicity of non-attenuated
aqueous As(III) concentrations. In cultures with AA and TiO2, the quantity of N2 gas
produced corresponded to the complete oxidation of As(III), including the adsorbed
fraction. The complete oxidation was also evidenced by the recovery of As(V) when the
AA was extracted at the end of the incubation. These results clearly indicate that
adsorbed As(III) was bioavailable for oxidation by denitrifiers, most likely due to the
reestablishment of the equilibrium upon consumption of aqueous As(III). This is
consistent with previous observations that As(V) coprecipitated with aluminum
hydroxide was bioavailable for bacterial dissimilatory As(V) reduction to As(III) [Zobrist
et al., 2000]. Likewise biological reduction has been found to be the main reason for the
91
remobilization of As(V) adsorbed to AA and TiO2 when incubated with anaerobic mixed
cultures [Jing et al., 2008; Sierra-Alvarez et al., 2005].
2.5.5. Environmental significance
The findings presented here demonstrate a widespread capacity in the environment for
the anoxic microbial oxidation of As(III). The oxidation of As(III) linked to the use of
ubiquitous nitrate as an electron acceptor may be an important missing link in the
biogeochemical cycling of arsenic between two common inorganic species, As(III) and
As(V), where DO is absent. The concentrations of arsenic in the subsurface pore water
are attenuated by adsorption to metal oxyhydroxides. The aqueous concentrations
encountered are generally not inhibitory to microorganisms. Therefore the biodiversity of
anoxic As(III) oxidizers may be greater than the few arsenic resistant strains reported
previously. In this study, it was shown that tolerance to As(III) was not a prerequisite for
anoxic As(III) oxidation by anaerobic microbial communities.
2.6. Conclusions
(1) Microorganisms capable of linking anoxic As(III) oxidation to denitrification are
widespread in anaerobic sediments and sludges. The use of ubiquitous nitrate as an
92
electron acceptor may be an important missing link the biogeochemical cycling of arsenic
between two common inorganic species, As(III) and As(V), where DO is absent.
(2) The doubling times for growth of the anoxic As(III) oxidizers range from 0.74 to 1.34
d.
(3) The anoxic oxidation of As(III) linked to denitrification is inhibited by As(III) with 5
mM causing complete inhibition.
(4) Dinitrogen gas is the end product of denitrification linked to As(III) oxidation if
As(III) is not present at inhibiting concentrations, nitrous oxide accumulates a major
product at inhibitory As(III) concentrations
(5) As(III) weakly adsorbed to AA or TiO2 is available as an electron donor for
denitrification.
(6) The biodiversity of anoxic As(III) oxidizers is potentially greater than the few arsenic
resistant strains reported previously. In this study, it was shown that tolerance to As(III)
was not a prerequisite for anoxic As(III) oxidation by anaerobic microbial communities.
93
2.7. Acknowledgments
The work presented here was funded by a USGS, National Institute for Water Resources
104G grant (2005AZ114G), and by a grant of the NIEHS-supported Superfund Basic
Research Program (NIH 5 P42 ES004940). The authors are grateful to Ron Oremland for
reviewing the manuscript. The use of trade, product, or firm names in this report is for
descriptive purposes only and does not constitute endorsement by the U.S. Geological
Survey. Also arsenic analyses were performed by the Analytical Section of the Hazard
Identification Core (Superfund Basic Research Program grant NIEHS-04940).
94
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100
CHAPTER 3
MOLECULAR CHARACTERIZATION AND IN SITU QUANTIFICATION OF
ANOXIC ARSENITE OXIDIZING DENITRIFYING ENRICHMENT CULTURES
3.1. Abstract
To explore bacteria involved in the oxidation of arsenite (As(III)) under denitrifying
conditions, three enrichment cultures (ECs) and one mixed culture (MC) were
characterized that originated from anaerobic environmental samples. The oxidation of
As(III) (0.5 mM) was dependent on NO3- addition, and N2-formation was dependent on
As(III) addition. The ratio of N2-N formed to As(III) fed approximated the expected
stoichiometry of 2.5. A 16S rRNA gene clone library analysis revealed three predominant
phylotypes. The first, related to the genus Azoarcus from the division β-Proteobacteria,
was found in the three ECs. All of the cultures examined in this study contained clones
representing one of two phylotypes closely related to the genus Acidovorax within the
Comamonadaceae family of β-Proteobacteria. Fluorescent in situ hybridization (FISH)
confirmed that Azoarcus accounted for a large fraction of bacteria present in the ECs. The
Azoarcus clones had 96% sequence homology with Azoarcus sp. strain DAO1, an isolate
previously reported to oxidize As(III) with nitrate. FISH analysis also confirmed that
Comamonadaceae clones were present in all cultures. These results taken as a whole
101
suggest that bacteria within the genus Azoarcus and the family Comamonadaceae are
probably involved in the observed anoxic oxidation of As(III).
102
3.2. Introduction
Arsenic (As) contaminated drinking water poses a risk to millions of people around the
world (Smedley & Kinniburgh, 2002). Numerous studies have provided compelling
evidence linking As in drinking water with cancer and other medical disorders (ATSDR,
2007). Arsenic contamination in groundwater generally results from naturally occurring
As-bearing geological material rather than specific anthropogenic sources (Smedley &
Kinniburgh, 2002) (White, et al., 1997). Arsenic usually occurs as either arsenite (As(III),
H3AsO3) or arsenate (As(V), H2AsO4- and HAsO42-) in circumneutral aqueous
environments. The mobility of As in the environment is highly influenced by microbial
transformations, which affect As speciation (Oremland & Stolz, 2003). A large diversity
of anaerobic microorganisms have been discovered that reduce As(V) to As(III) utilizing
two biochemical systems. One system involves a cytoplasmic As(V) reductase (arsC) as
part of a specific As efflux mechanism in resistant strains (Silver & Phung, 2005, Lloyd
& Oremland, 2006). The other system involves a periplasmic dissimilatory As(V)
reductase (arrA) whereby microorganisms gain energy for growth utilizing As(V) as a
terminal electron acceptor (Malasarn, et al., 2004, Stolz, et al., 2006). The formation of
As(III) from the microbial reduction of As(V) increases the public health risk, since
As(III) is generally considered to be the more mobile and toxic form of As (Smedley &
Kinniburgh, 2002, Sierra-Alvarez, et al., 2004).
The oxidation of As(III) to As(V) has the potential to immobilize soluble As and
lower the risk of As contamination in groundwater. Microorganisms from physiologically
103
diverse groups, including both heterotrophs and autotrophs, can oxidize As(III) to As(V)
in the presence of elemental oxygen (O2) in various environments (Stolz, et al., 2006,
Inskeep, et al., 2007). The heterotrophic oxidation of As(III) is catalyzed by periplasmic
enzymes, most likely as a detoxification mechanism because energy for growth is not
derived from the reaction (Silver & Phung, 2005). However, some chemolithotrophic
As(III)-oxidizing bacteria can grow using the energy gained from the oxidation As(III)
(Santini, et al., 2000, Rhine, et al., 2007). As(III) oxidases are heterodimeric structures
with two subunits containing molybdopterin and Rieske iron-sulfur domains (Santini &
vanden Hoven, 2004, Stolz, et al., 2006, Rhine, et al., 2007). The full diversity and
distribution of As(III) oxidase genes in environmentally relevant bacteria is not presently
known (Silver & Phung, 2005). Inskeep et al. (Inskeep, et al., 2007) and Rhine et al.
(Rhine, et al., 2007) recently designed primer sets to examine As(III) oxidase-like genes
(aroA, asoA and aoxB) in environmental samples. The results suggest that a variety of
aerobic As(III) oxidizers are widespread in As-contaminated environments.
Compared with aerobic As(III)-oxidizing microorganisms, little is known about
anoxic As(III)-oxidizers. Although oxidation with O2 is more favorable based on
biochemical energetic considerations, alternative oxidants with lower reduction potentials
are feasible for the oxidation of As(III). The data of Senn and Hemond (Senn & Hemond,
2002) were one of the first indications that anoxic As(III) oxidation occurs in the
environment, based on the observation that seasonal changes in arsenic speciation were
related to nitrate in anoxic lake water. Recently it has come to light that nitrate can be
utilized by anaerobic microorganisms to gain energy from As(III) (Oremland, et al.,
104
2002). A facultative chemolithoautotrophic arsenite oxidizing bacterium, Alkalilimnicola
ehrlichi sp strain MLHE-1, capable of oxidizing As(III) to As(V) linked to partial
denitrification of nitrate to nitrite, was isolated from an arsenic-containing soda lake in
California (Mono Lake, CA) (Oremland, et al., 2002, Hoeft, et al., 2007). Subsequently,
two anaerobic chemolithoautotrophic As(III)-oxidizing denitrifying bacteria were
isolated from arsenic-contaminated soil and shown to oxidize As(III) with a
stoichiometry consistent with the complete denitrification of nitrate to dinitrogen (N2) gas
(Rhine, et al., 2006) as shown in equation 1. The occurrence of the nitrous oxide gene,
nosZ, was also demonstrated in the isolates (Rhine, et al., 2006). The two novel strains,
DAO1 and DAO10, are phylogenetically similar to Azoarcus and Sinorhizobium on the
basis of 16S rRNA sequences, respectively.
5H3AsO3 + 2NO3- → 5HAsO42- + N2 + 8H+ + H2O
[eq 1]
In this study, we report on bacteria involved in the anoxic oxidation of As(III)
linked to denitrification in three enrichment cultures (ECs) and one mixed culture (MC)
that were established from sludges and sediments with no prior exposure to As as well as
from a biofilm of an anoxic reactor exposed to nitrate and As(III). Complete
denitrification in the ECs and the MC was confirmed by direct measurement of N2
formation coupled with As(III) oxidation. The cultures were characterized by 16S rDNA
gene clone library and fluorescent in situ hydridization (FISH) analysis (Amann, et al.,
1995, Sanz & Kochling, 2007). The clone libraries were used to identify the predominant
105
phylotypes in the cultures. FISH probes were then utilized to quantify the predominant
phylotypes as fractions of the total cells. The results provide insights on the types of
bacteria responsible for the widespread occurrence of anoxic As(III)-oxidizing
denitrifying bacteria in the environment.
3.3. Materials and Methods
3.3.1. Enrichment cultures
The three ECs differed based on the source of the original inoculum. The inoculum for
enrichment culture 1 (EC1) was anaerobic effluent containing suspended biofilm from a
laboratory-scale bioreactor (Department of Chemical and Environment Engineering,
University of Arizona, Tucson, AZ) which coupled the anoxic oxidation of As(III) with
denitrification. The bioreactor was fed with 3.75 mM As(III) and 6.4 mM NO3- using a
hydraulic retention time of 1 day. The inoculum was obtained from a sulfoxidizingdenitrification reactor originally inoculated with methanogenic granular sludge (biofilm
pellets) obtained from an industrial upward-flow anaerobic sludge blanket (UASB)
treating recycle paper wastewater in Eerbeek, The Netherlands. The effluent sample was
collected after 498 days of operation. The inoculum for enrichment culture 2 (EC2) was a
methanogenic granular sludge obtained from a UASB treating alcohol distillery
wastewater (Nedalco, Bergen op Zoom, The Netherlands). The inoculum for enrichment
culture 3 (EC3) and for the mixed culture (MC) was derived from duck pond sediment
106
obtained at the Agua Caliente Park (Tucson, Arizona). The culture volumes during the
initial feeding were different from the routine volumes used for the ECs. The initial feed
was performed in 500 mL serum bottles, with 250 mL media. EC1 received 10 ml of
effluent (with resuspended biofilm), EC2 and EC3 received 5 g of wet sludge and
sediment, respectively. The MC can be considered a subset of EC3 because it was
inoculated with a colony cultured from a plate inoculated from the 5th transfer of EC3
(the colony used for inoculation was a single colony that upon later examination
contained a mixed culture). The plate contained the same medium used for the ECs with
25 g l-1 sterile noble agar (Difco Laboratories, Detroit, MI) and was incubated
anaerobically for approximately 4 weeks under an N2 atmosphere.
The three ECs and the MC were maintained in the absence of O2 in a basal
mineral medium amended with 0.5 mM As(III) as electron donor, 10 mM NO3- as
electron acceptor, and 8 mM HCO3- as the sole carbon source. The cultures were
incubated in 160 ml serum flasks with a liquid volume of 100 ml that were sealed with
butyl rubber stoppers. The medium and headspace was purged for 20 min with N2:CO2
(80:20, v/v). Cultures were serially transferred (5%, v/v) to fresh medium after
incubations of 2-3 weeks at 30ºC and confirmation of As(III) oxidation (measured by
As(V) formation). The enrichment process was continued for 25 transfers.
107
3.3.2. Medium composition
The standard As(III)/NO3--containing basal medium (A/N-BM) (pH 7.0-7.2) was
prepared using ultra pure water (Milli-Q system; Millipore) and contained (mg l-1):
K2HPO4 (638); KH2PO4·2H2O (1700); NH4Cl (593); MgCl2·6H2O (166); MgSO4·7H2O
(22); CaCl2 (22), and 2 ml l-1 of a trace element solution containing (mg l-1): FeC13·4 H2O
(2,000); CoCl2·6 H2O (2,000); MnCl2·4 H2O (500); AlCl3·6 H2O (90); CuCl2·2H2O (30);
ZnCl2 (50); H3BO3 (50); (NH4)6Mo7O24·4 H2O (50); Na2SeO3·5 H2O (100); NiCl2·6 H20
(50); EDTA (1,000); resazurin (200); HCl 36% (1 ml). The basal medium was amended
with 8 mM HCO3- (NaHCO3), 0.5 mM As(III) (NaAsO2) and 10 mM NO3- (KNO3). The
basal medium with NO3- was sterilized by autoclave, while NaHCO3 and As(III) were
sterilized using membrane filtration (0.2 µm).
3.3.3. Experimental incubations
Once the cultures were established, they were used to evaluate the time course of As(III)
oxidation under denitrifying conditions using the A/N-BM medium. Samples were
periodically removed for measurement of As(III), As(V) and NO3-. Various controls were
utilized based on the requirements of each experiment. Abiotic controls were prepared
without adding enrichment culture inoculum. Heat killed controls were prepared by
autoclaving the flasks added with inoculum for 20 min at 121ºC, on three consecutive
days. Controls lacking As(III) were included to measure endogenous consumption of
108
NO3-. All assays were conducted in triplicate. To avoid the contamination of carry-over
NO3- from old cultures to fresh medium, the cultures (5% volume of previous culture)
were centrifuged in sterilized eppendorf tubes at 10,000 rpm for 10 min, the pellets were
collected and washed into sterilized MiliQ water for two cycles. The pellets were
resuspended into same volume of sterilized MiliQ water and transferred to the
experiments.
In one experiment the impact of low concentrations of organic carbon
supplements were evaluated. In this case the A/N-BM medium was supplemented with
yeast extract (1.0 mg l-1) or pyruvate (1.7 mg l-1). The experiment was inoculated with
EC1 at 1% (v/v).
3.3.4. Most probable number (MPN)
MPN assays were performed in A/N-BM. Culture samples (1 mL) were taken after 23
transfers, were homogenized by vortexing, and were serially diluted in 10-fold
increments to 10-9 in sterile medium under anaerobic conditions. Each dilution was set up
with five replicate MPN tubes. The tubes were incubated in an orbital shaker (110 rpm) at
30°C under anaerobic conditions with N2/CO2 (80:20) in the headspace. After 3-4 weeks
incubation, samples from each tube were removed and analyzed for As(V). Conversion of
> 80% of the As(III) to As(V) was considered as a positive tube. Finally an MPN table
for five tubes (APHA, 1999) was utilized to enumerate the number of bacteria with
As(III)-oxidizing capability.
109
3.3.5. Batch assays to determine the terminal product of denitrification
End products of denitrification were measured by monitoring gaseous nitrogen species in
the headspace of bioassays flushed with He/CO2 (80:20, v:v) in lieu of N2/CO2 as
described above. Gaseous nitrogen measured in As(III) spiked samples (3.0 mM) was
compared to endogenous controls. Microbial reduction of NO3- coupled to As(III)
oxidation was assessed in shaken batch bioassays inoculated with EC1 and EC3. The
NO3- concentration utilized was 2.5 mM for the experiments. Samples were supplied 50 g
l-1 of activated aluminum (AA-400G, Alcan Bauxite and Alumina) adsorbents to lower
the effective aqueous concentration of As(III) so as to minimize its toxicity. As(III) was
supplied at 3.0 mM; however by using this procedure, the aqueous equilibrium
concentration of As(III) was 0.59 mM. The high concentration of As(III) supplied was
required to have enough electron equivalents to measure N2 production properly.
Headspace samples were analyzed periodically for N2 and N2O with a pressure lock gas
tight syringe (1710RN, 100 µl (22s/2‫״‬/2), Hamilton, Reno, Nevada USA) to confirm
denitrification. Liquid samples were analyzed for NO3- and NO2-. Flushed headspace
controls incubated with just water were monitored to measure background levels of N2.
The background level was low.
110
3.3.6. Analytical methods
Samples (1 ml) were taken from sealed anaerobic serum flasks by piercing the stoppers
using sterile 1.0 ml syringes with 16-gauge needles. All liquid samples were centrifuged
(10 min, 10,000 rpm) or membrane filtered (0.45 µm) immediately after sampling and
stored in polypropylene vials. Samples for As analysis were stored at -20°C. As(III) and
As(V) were analyzed by ion chromatography–inductively coupled plasma–mass
spectroscopy (IC-ICP-MS). The system consists of an IC (Agilent 1100) and an ICP-MS
(Agilent 7500a) with a Babington nebulizer as the detector. The operating parameters
were as follows: Rf power 1500 watts, plasma gas flow 15 l/min, carrier flow 1.2 l/min,
arsenic was measured at 75 m/z and terbium (IS) measured at m/z 159. The injection
volume was 10 µl. The detection limit for the various arsenic species was 0.1 µg l-1.
Nitrate and nitrite were analyzed by suppressed conductivity ion chromatography
using a Dionex 500 system (Sunnyvale, CA, USA) fitted with a Dionex IonPac AS11
analytical column (4 x 250 mm) a AG16 guard column (4 mm x 40 mm). During each
injection the eluent (KOH) used was 20 mM for 20 min. The same procedure also was
suitable as an alternative analysis of As(V).
N2 and N2O were analyzed using a Hewlett Packard 5890 Series II gas
chromatograph fitted with a CarboxenTM 1010 Plot column (30 m x 0.32 mm) and a
thermal conductivity detector. The temperature of the column, the injector port and the
detector were 220, 110 and 100 ºC, respectively. Helium was used as the carrier gas and
the injection volume was 100 µl.
111
Other analytical determinations (e.g., pH, TSS, VSS, etc.) were conducted
according to Standard Methods (APHA, 1999).
3.3.7. 16S rRNA gene clone libraries
Community genomic DNA was extracted from 5 ml samples taken from each EC and the
MC by a modification of the Current Protocols extraction method for Genomic DNA
from bacteria (Ausubel, et al., 1995). Briefly, cells were pelleted and resuspended in TE
buffer (10 mM Tris Cl, 1 mM EDTA, pH 8.0) and subjected to three freeze-thaw cycles
in liquid nitrogen and boiling water.
Cells were then lysed and the DNA purified
according to the protocol. Before the extraction all sterile tubes, caps, and solutions used
in the protocol were exposed to UV light for 10 min in a UVC-508 Ultraviolet
Crosslinker (Ultra-Lum, Claremont, CA) to remove any potential DNA contamination.
The extracted DNA was stored at -20°C. Extraction blanks were processed in parallel
throughout the full procedure as negative controls to evaluate potential DNA
contamination from reagents. The DNA was quantified with TBS-380 Mini-Fluorometer
(Tuner BioSystems, Sunnyvale, CA) Using Molecular Probes’ PicoGreen dsDNA
Quantitation Reagent (Molecular Probes, Inc., Eugene, Oregon).
The 16S rRNA gene was PCR amplified from community DNA extracts using
universal primers 27F and 1492R (Lane, 1991). Each 25-µl reaction contained 0.5 mM of
each primer, 0.2 mM of each dNTP, 1x buffer consisting of 10 mM Tris-HCl, 50 mM
KCl, 2.0 mM MgCl2 (pH 8.3), 5% dimethyl sulfoxide (DMSO), 0.5 U of Taq DNA
112
polymerase (Roche, Indianapolis, IN), and 2.5 µl of DNA extract. The DNA was
amplified with an initial denaturation step of 95°C for 5 min followed by 30 cycles of:
94°C for 1 min, 60°C for 1 min and 72°C for 1.15 min, with a final extension for 10 min
at 72°C in a Perkin Elmer GeneAmp PCR System 2400 (PerkinElmer Inc., Boston, MA).
The PCR products were visualized on a 1% GenePure LE agarose gel (Intermountain
Scientific Corp., Kaysville, UT) and then purified using a Quick PCR purification kit
(Qiagen, Chatsworth, CA).
The purified PCR products were cloned into plasmid vector pCR®2.1-TOPO®
using the TOPO TA cloning system (Invitrogen, Carlsbad, CA) according to the protocol
described by the manufacturer. The control TOPO® Cloning reaction was performed
using the reagent included in the kit in parallel with real samples. The plasmid DNA was
purified using PureLinkTM Quick Plasmid Miniprep Kit (Invitrogen, Carlsbad, CA)
according to the protocol described by the manufacturer. Purified plasmid DNA was
sequenced
using
four
primers
including
the
vector
primers
T7
(5’-
TAATACGACTCACTATAGGG-3’) and M13R (5’-AGGAAACAGCTATGACCATG3’) and universal internal primers 518F and 1070R (Lane, 1991). Nearly full-length 16S
rRNA gene sequences were obtained using an Applied Biosystems 3730xl DNA analyzer
(Applied Biosystems, Foster City, CA) (University of Arizona Research Laboratory
Genomic Analysis and Technology Core, Tucson, AZ).
The number of clones analyzed for each culture was determined using a
rarefaction curve to estimate the diversity (Analytic Rarefaction 1.3, UGA Stratigraphy
Lab, University of Georgia, Atlanta, GA). An exponential model, y = a × [1-exp(-b × x)],
113
was used based on the formulation of Tipper (Tipper, 1979) to fit the clone distribution
data. Sequences were aligned by using FAKII Fragment Assembly Kernel (FAKtory,
Biotechnology Computing Facility, University of Arizona, Tucson, AZ) and compared to
known sequences through the Basic Local Alignment Search Tool (BLAST)
(http://www.ncbi.nlm.nih.gov/BLAST/). The clones were clustered into phylotypes on
the basis of having sequence similarities of > 99. 5%. ARB program package (Ludwig, et
al., 2004) was used for the phylogenetic analysis. Sequence data were aligned and a tree,
including 16S rRNA gene sequences from reference bacterial strains (GenBank) and
unique phylotypes recovered from each of the three ECs and the MC, was constructed
using the Maximum Likelihood algorithm (Felsenstein, 1981) with a bacterial filter of the
fastDNAml program (Olsen, et al., 1994) in the ARB program package. Tree confidence
was tested using the Hasegawa method (Hasegawa, et al., 1985). The sequences obtained
in this work (of selected clones representing each phylotype obtained in each culture)
have been deposited in the GenBank database with accession numbers: EC1-1
(EU708507); EC2-2 (EU741793); MC-1 (EU741795); EC3-10 (EU708501); EC1-12
(EU708506); EC1-7 (EU708505); EC3-1 (EU708500); EC2-7 (EU741794); EC1-17
(EU708503); EC1-10 (EU708508); EC2-2 (EU741793); EC1-33 (EU708504); MC-9
(EU741796); and EC3-11 (EU708502).
114
3.3.8. FISH analysis
In order to quantify the bacterial composition of enrichment cultures, 10 ml of each
culture were fixed with formaldehyde (4% in phosphate saline buffer, PBS) for 4 h. After
that, the samples were washed with PBS and stored in a mixture of PBS:ethanol 50:50
v:v at -20°C until they were hybridized. 10 µl of each sample were placed on to multiwell slides. All samples were further dehydrated by immersion in 50, 80, and 100%
ethanol solutions for 3 min each time. FISH was applied according to protocols described
elsewhere ((Amann, et al., 1990, Amann, 1995)). Samples were hybridized with the
probes (Table S1) Azo644, specific for Azoarcus (Hess, et al., 1997); DEN220, which
hybridizes with the Comamonadaceae family (Ginige, et al., 2005) (to target members of
the genera Alicycliphilus and Diaphorobacter within the Acidovorax cluster); and the
universal probe for Bacteria domain Eub338 (Amann, et al., 1990). The probes were
purchased from Biomers (Ulm, Germany). The formamide concentration used for
hybridization (2 h at 46°C) was 20% (Azo644), 40% (DEN220) and 35% (Eub338) and
the NaCl for washing (15 min at 48°C) was 215 mM (Azo644), 46 mM (DEN220) and 70
mM (Eub338). After hybridization the samples were stained with 4’, 6’-diamin
phenylindol (DAPI, 1 mg ml-1) in order to determine the total cells present. Twenty
randomly selected fields (approximately 1000 DAPI-stained cells) were analyzed using a
Zeiss Axiovert 200 fluorescence inverted microscope.
115
3.4. Results
3.4.1. Anoxic As(III) conversion by enrichment cultures
The three ECs and the MC were incubated under anoxic conditions with As(III) either in
the presence or absence of NO3- (Figure 1). In treatments incubated with NO3-, As(V)
formation started after approximately 1 d and the reaction came to completion after 4-6 d.
The maximum reaction rate, calculated from the formation of As(V), was 221 ± 9, 221 ±
8, 175 ± 12, and 138 ± 1 µM d-1 for EC1, EC2, EC3, and MC, respectively. The reactions
were dependent on the presence of NO3- as evidenced by the lack of any significant
conversion in incubations without added NO3- (Figure 1). The reactions did not occur in
bottles receiving the medium with As(III) and NO3- but lacking inoculum (Figure 1) or
with heat-killed inoculum (data not shown).
116
0.6
As(V) (mM)
0.5
0.4
0.3
0.2
0.1
0.0
0
2
4
6
8
Time (Days)
Figure 3.1 Formation of As(V) by cultures under denitrifying condition. Incubated with
0.5 mM As(III) and 10 mM NO3-: EC1 (▲), EC2 (■), EC3 (●), and MC (♦) Control with
only 0.5 mM As(III) and no NO3-: EC1 (), EC2 (□), EC3 (○), and MC (◊) abiotic
control (X).
MPN analysis of the As(III)-oxidizing denitrifying populations performed after
the 23rd transfer revealed that 4.9×107, 2.2×108, 1.7×107, and 2.3×107 cells were present
per ml of EC1, EC2, EC3 and MC, respectively; corresponding to 3.3 × 1012 to 4.4 × 1014
cells produced per mol of As(V) formed.
EC1 was tested for its ability to utilize low concentrations of organic carbon
sources for growth under denitrifying conditions. The treatments amended with yeast
extract and pyruvate (1.0 and 1.7 mg l-1, respectively) showed no significant difference in
117
the rate or extent of As(III) oxidation compared with the autotrophic oxidation conditions
without any organic carbon added (data not shown).
Experiments performed to determine whether NO3- consumption and N2 formation
is linked to As(III) oxidation are shown in Fig. 2. In EC cultures amended with both
As(III) and NO3-, N2 was formed as an end product (Figure 2A), NO3- was consumed
(Figure 2B), and NO2- was transiently observed with its appearance and disappearance
correlating closely with the consumption of NO3- and formation of N2 respectively
(Figure 2C). In the absence of As(III) and in the abiotic controls, no consumption of NO3or appearance of either NO2- or N2 was observed. The molar ratios of As(III)
consumption to the NO3- consumption and N2-N formation (corrected for background
endogenous denitrification) are provided in Table 1. The molar ratios of As(III) fed to
NO3- consumption and N2-N formation from the adsorbent amended EC1 and EC3
enrichment cultures were close to the theoretical ratio of 2.5 expected for As(III)
oxidation linked to complete denitrification to N2 (equation 1).
118
N2-N (mmol/Lliquid)
1.6
1.2
0.8
0.4
0
0
10
20
30
40
50
60
70
30
40
50
60
70
50
60
70
NO3- consumption (mM)
1.6
1.2
0.8
0.4
0
0
10
20
0
10
20
NO2- concentration (mM)
1.6
1.2
0.8
0.4
0
30
40
Time (days)
Figure 3.2 The influence of As(III) on the formation of N2-N (Panel A), consumption of
NO3- (Panel B) and accumulation of NO2- (Panel C) by EC1 (squares) and EC3
(triangles) cultures under denitrifying conditions. Legend: full treatment with 3 mM
As(III) and 2.5 mM NO3- (■ and ▲), endogenous control with only 2.5 mM NO3- (□ and
∆), and abiotic control (●). All assays contained activated alumina (50 g l-1).
119
Table 3.1 Summary of denitrification linked to As(III) oxidation when adsorbed onto acitivated alumina (50 g l-1)
EC1
Parameters
As(III)+NO3-
NO3- only
As(III)+NO3-
NO3- only
3.10±0.02
__
3.10±0.02
__
NO3 consumed (mM)
1.41±0.07
0.18±0.01
1.56±0.23
0.12±0.07
Corrected† NO3- consumed (mM)
1.23±0.07
__
1.44±0.18
__
N2-N formed (mmol/Lliquid)
1.24±0.06
0.02±0.01
1.18±0.03
0.01±0.02
Corrected N2-N formed (mmol/Lliquid)
1.21±0.06
__
1.17±0.02
__
Corrected N2-N formed/corrected NO3- consumed
0.99±0.06
__
1.12±0.13
__
Mol As(III) fed/corrected mol NO3 consumed
2.53±0.17
__
2.17±0.24
__
Mol As(III) fed/corrected mol N2-N formed
2.56±0.14
__
2.64±0.03
__
As(III) fed (mM)
-
-
†
EC3
: Corrected for the endogenous nitrate consumption or N2-N formation determined in the nitrate treatment lacking added As(III).
120
3.4.2. Community composition of enrichment cultures
The composition and structure of the microbial community in the ECs and MC were
analyzed by preparing a 16S rRNA gene clone library for each. Rarefaction analysis for
each clone library (Figure S1) suggested that 46, 18, 21 and 20 clones were sufficient for
capturing the community composition in the clone libraries corresponding to EC1, EC2,
EC3, and MC, respectively. A total of eight unique phylotypes were obtained from the
four cultures. Of the eight unique phylotypes recovered, 6, 3, 3, and 2, were found in
EC1, EC2, EC3, and MC respectively (Figure 3). Note that the 2 phylotypes found in the
MC were also recovered from EC3 which was expected since the colony used to
inoculate the MC was from the EC3 (Figure 4).
The 105 clones analyzed in this study fell into four phylogenetic divisions, αProteobacteria β-Proteobacteria, γ-Proteobacteria, and Flavobacteria accounting for
1.9, 84.8, 9.5, and 3.8% of all the clones from the four cultures, respectively. The
phylogenetic relationships among the phylotypes recovered are shown in Figure 4. The βProteobacteria accounted for a large majority of all the clones and four of the eight
phylotypes recovered were most closely related to members of this group (Azoarcus,
Alcyliphilus, Diaphorobacter and unclassified Burkholderiales). One phylotype belonged
to the α-Proteobacteria and was most closely related to Rhizobiaceae. Two phylotypes
belonged to the γ-Proteobacteria and were most closely related to Stenotrophomonas and
Dokdonella, and the remaining phylotype was affiliated with Flavobacteria and was most
closely related to Kaistella.
121
Figure 3.3 Phylogenetic distributions in the four cultures. Diagrams show the relative
abundance of 16S rRNA phylotypes of clones from each culture with a total of 8 unique
phylotyptes found in the study. Out of these 8 unique phylotyptes, 6, 3, 3, and 2 were
found in EC1, EC2, EC3, and the MC, respectively.
122
Figure 3.4 Phylogenetic relationships of the EC and MC clones recovered in this study.
Maximum likelihood analysis was used to generate the tree from 16S rRNA gene sequences of
bacterial strains and unique phylotypes represented by clones recovered from the four cultures
labeled with EC#-# (the first # refers to the enrichment culture, the second # refers to the clone
number), or MC-# (where the # refers to the clone number). Genbank accession numbers are
indicated in parentheses. Confidence values for tree topology (1,000 replicates) are given for
nodes with ≥ 50% support.
123
Three out of the eight phylotypes identified in the ECs and MC dominated in
terms of clones analyzed. One of these phylotypes (including clones EC1-7, EC3-1,
EC2-7) was 100% likely to belong to the genus Azoarcus (according to the RDP
Classifier (Cole, et al., 2005)) and was found in the three ECs (Figure 3 and 4). The
remaining two dominant phylotypes were related to the genera Alcycliphilus (clones
EC1-1, EC2-2; 98%; RDP Classifier) or Diaphorobacter (clones MC-1, 99%; EC3-10,
96%; RDP classifier) which are both closely related to Acidovorax belonging to the
Comamonadaceae. All of the ECs and the MC contained one of the two phylotypes
related to the Comamonadaceae (Figure 3 and 4). The ECs originating from biofilm and
sludge (EC1 and EC2) contained the phylotype related to the Alcyliphilus genus of this
cluster; whereas those originating from pond sediment contained the phylotype most
closely related to Diaphorobacter.
The remaining five phylotypes were present as small fractions of the clones
examined. The phylotype most closely related to Dokdonella (clone EC1-10; 100%;
RDP Classifier) accounted for 11 to 17% of the clones in the the ECs originating from
biofilms and sludge (EC1 and EC2). The phylotype most closely related to Kaistella
(clones EC3-11, MC-9; 100%; RDP Classifier) accounted for about 10% of the pond
sediment-derived cultures (EC3 and MC). The phylotypes most closely related to
Stenotrophomonas (clone EC1-17; 100%; RDP Classifier), Rhizobiaceae (clone EC1-33;
100%; RDP Classifier), and unclassified Burkholderiales (clone EC1-12; 94%; RDP
Classifier), were from the EC started with the biofilm incolum (EC1), and each accounted
for about 4% of all the clones in that EC.
124
FISH analysis showed that the cells detected by the Eubacteria probe, Eub338,
ranged from 75 to 80% of the total DAPI-stained cells in the different ECs and MC
(Figure 5). The Azoarus-specific probe, Azo644, hybridized with 38 to 72% of DAPIstained cells in EC1, EC2 and EC3. As expected, there was no hybridization of Azo644
in the MC which did not contain Azoarcus-like clones. The Comamonadaceae-specific
DEN220 probe (used to target Alicycliphilus and Diaphorobacter) detected a smaller
proportion of the total cells ranging from 4 to 20% of total DAPI-stained cells in the four
cultures.
Figure 3.5 FISH analysis of the EC and MC communities using universal (Eub338) and
group-specific (Azo644 and DEN220) probes. Results are presented as percentages
relative to DAPI counts (set at 100%).
125
It should be noted that the FISH results and clone library results are not identical.
For example in EC1, the clone library results indicate that the Azoarcus and
Comamonadaceae clones are approximately equal in number. In contrast, the FISH
results indicate that the Azoarcus is strongly dominant. It should be recognized that both
of these techniques have associated biases which likely explain this discrepancy.
However, these techniques, used in combination are in agreement as to the presence or
absence of Azoarcus and Comamonadaceae in the ECs and MC. Further, these
techniques provide strong evidence that in addition to Azoarcus members of the
Comamonadaceae are capable of carrying out As(III) oxidation under anoxic conditions.
Numbers of unique clones
7
6
5
4
3
2
1
0
0
10
20
30
40
50
Numbers of clones tested
Figure 3.6 Evaluation of representative clones obtained from four ECs by rarefaction
analysis: EC1 (♦), EC2 (●), EC3 (▲) and EC4 (■).
126
Table 3.2 Groups-specific oligonucleotide probes used in this study
Target organisms
Probe
Probe sequence (5’-3’)
Formamide†
(%)
NaCl‡
mM)
Reference
Bacteria
Eub338
GCTGCCTCCCGTAGGAGT
35
70
Amann et
al., 1990a
Azoarcus
Azo644
GCCGTACTCTAGCCGTGC
20
215
Hess et al.,
1997
Comamonadaceae
Den220
GGCCGCTCCGTCCGC
40
46
Ginige et
al., 2005
†
: Percentage of formamide (%v/v) in the hybridization buffer.
‡
: Concentration of sodium chloride in the wash buffer.
3.5. Discussion
3.5.1. Evidence of As(III) oxidation linked to denitrification
Multiple lines of evidence demonstrate that the bacterial communities in the ECs and MC
evaluated in this study linked the anoxic oxidation of As(III) to denitrification. First, the
formation of As(V) from As(III) was dependent on the presence of NO3- and inoculum.
No oxidation occurred in controls lacking either inoculum or NO3-. Second, the
production of N2 was linked to the addition of As(III) to the cultures and the yield of N2
corresponded to the expected stoichiometry from the electron equivalents in As(III).
Third, bacteria appeared to benefit from the process as evidenced by the large number of
transfers (25×) performed in this study. Likewise, the MPN assay indicated high cell
127
densities of microorganisms were produced with up to 4 × 1014 cells mol-1 As(V) formed.
The enrichment cultures were sustained without addition of organic C, and small
additions of organic C (yeast extract or pyruvate) did not stimulate growth. These
findings indicate the cultures are chemolithoautotrophic, obtaining energy from As(III)
and most likely obtaining carbon from added HCO3-/CO2.
3.5.2. The As(III) oxidizing bacterial community – Azoarcus and Comamonadaceae
One phylotype belonging to the β-Proteobacteria and related to the genus Azoarcus
constituted a large fraction of the clones analyzed in all the ECs studied. FISH analysis
supported the dominance of this genus, 38 to 72% of the cells in these cultures were
detected with the Azoarcus probe.
The Azoarcus phylotype was closely related to
Azoarcus strains PbN1 (98%) and EbN1 (97%) (Rabus & Widdel, 1995) (Figure 4).
These strains are known for their ability to degrade propyl- and ethylbenzene (Szaleniec,
et al., 2007). In general, the Azoarcus genus is known for the ability to degrade aromatic
substrates under anaerobic denitrifying conditions (Reinhold-Hurek & Hurek, 2006).
Azoarcus strain EbN1 and a number of closely related aromatic-degrading strains within
the Azoarcus/Thauera cluster are being proposed as a new genus, “Aromatoleum”. In
fact, EbN1 is known as Aromatoleum aromaticum, a metabolically versatile
representative of the group capable of degrading a wide variety of aromatic and nonaromatic compounds under both anoxic and aerobic conditions (Szaleniec, et al., 2007,
Wohlbrand, et al., 2007).
128
The Azoarcus clones from this study had 96% similarity to Azoarcus strain
DAO1, which is an isolate from an arsenic-contaminated site known to link As(III)
oxidation to denitrification (Rhine, et al., 2006). Therefore, there is a precedent within the
genus Azoarcus for the anoxic oxidation of As(III). Azoarcus strain DAO1 does not
contain a gene that hybridizes with degenerate primers designed from known sequences
of arsenite oxidases (Rhine, et al., 2007). Nonetheless, it is interesting to note that the key
enzyme produced by “Aromatoleum aromaticum” strain EbN1,
ethylbenzene
dehydrogenase (Johnson, et al., 2001, Kloer, et al., 2006, Szaleniec, et al., 2007) are
related to arsenite oxidases (Inskeep, et al., 2007, Rhine, et al., 2007) since both contain
molybdate cofactors and are members of the DMSO family.
The genome of “Aromatoleum aromaticum” strain EbN1 (97% similarity with
EC1-7 and EC3-1) has been sequenced (Rabus, 2005, Rabus, et al., 2005). The genome
contains putative genes for arsenic resistance in the ars operon family (arsC, arsD, arsR,
and arsA) (Rabus, et al., 2005). The ars operon could potentially explain the ability of the
Azoarcus strains in the ECs to be viable in the presence of 0.5 mM arsenic (in some
experiments 1.0 mM). The genome also contains a complete set of denitrification genes
(e.g. NarG, NirS, NorC, and NosZ) (Rabus, 2005) which could account for the observed
production of N2 from NO3-.
All of the cultures examined in this study (ECs and MC) contained clones
representing one of two phylotypes related to a cluster of genera (Acidovorax,
Alicycliphilus and Diaphorobacter) within the family Comamonadaceae which also
belongs to the β-Proteobacteria (Figure 4). The fraction of clones belonging to this
129
group ranged from 29 to 90% depending on the culture. FISH analysis with the DEN220
probe confirmed the presence of Comamonadaceae but suggested they constitute a
smaller fraction of the community (Figure 5). One Comamonadaceae cluster phylotype
recovered from sludge-derived ECs (EC1 and EC2) was related to the genus
Alicycliphilus. This clones have 99% similarity to an uncultured clone from activated
sludge, Alicycliphilus strain R-24611. These clones also have 99% similarity to several
other Acidovorax isolates. An example is an H2-consuming nitrate reducing isolate
Acidovorax strain Ic31 from a hydrogenotrophic denitrification batch reactor (Vasiliadou,
et al., 2006), indicating a precedent for autotrophic denitrification in closely related
species. Another example is Acidovorax avenae strain C1 which was isolated from a low
O2 phenol-degrading nitrate-reducing culture (Baek, et al., 2003).
The second Comamonadaceae cluster phylotype was obtained from the pond
sediment-derived cultures (EC3 and MC) and was related to the genus Diaphorobacter.
A number of published isolates belonging to species Diaphorobacter nitroreducens have
99% similarity with clones EC3-10 and MC-1 recovered in this study. These are known
for their ability to degrade polyhydroxyalkanoate bacterial energy storage polymer,
poly(3-hydroxybutyrate-co-3-hydroxyvalerate) (PHBV), under denitrifying conditions
(Khan & Hiraishi, 2001, Khan, et al., 2002, Khan & Hiraishi, 2002 ). Interestingly, the
genus Azoarcus is known for the production of poly-β-hydroxybutyrate (PHB) as energy
storage polymers, which accumulate in cells as granules (Reinhold-Hurek & Hurek,
2006, Serafim, et al., 2006). Thus, the PHB produced by Azoarcus could potentially
benefit the Comamonadaceae members of the community related to Diaphorobacter by
130
serving as a source of carbon and energy suggesting one possible interaction between two
important members of the EC communities.
Acidovorax sp. JS42, which is a well known aerobic nitroaromatic degrading
bacterium (Lessner, et al., 2003), also has 99% similarity with clones EC3-10 and MC-1.
The genome of Acidovorax sp. JS42 has been sequenced and the genome contains ars
operon genes (arsA, arsR and arsC) as well as a gene for an arsenical-resistance protein
which could explain the ability of these microorganisms to tolerate As. The genome also
contains genes involved in denitrification (nitrate reductase, nitrite reductase, nitric oxide
reductase large subunit, and nitrous oxide reductase) as well as one gene for an esterase
for PHB depolymerization.
There is no precedent for anoxic arsenite oxidation in the Comamonadaceae
cluster. In the MC, Diaphorobacter was dominant which indicates that it is likely
involved in using As(III) as an energy source. Recently, an aerobic As(III) oxidizing
strain (GW2) was isolated in Chinese sediments that belongs to Acidovorax sp. (Fan et
al., 2008), and it was shown to possess an arsenic oxidase gene. Strains in the
Comamonadaceae cluster are also known for their ability to link Fe(II) oxidation to
dentrification (Straub, et al., 2004).
3.5.3. The As(III) oxidizing bacterial community – less dominant members
A phylotype belonging to the γ-Proteobacteria was recovered from sludge-derived ECs
(EC1 and EC2). This phylotype has 97% similarity to Dokdonella fugitive, an isolate
131
obtained from potting soil (Cunha, et al., 2006) (Figure 4). A second phylotype belonging
to the γ-Proteobacteria was recovered only from EC1. This phylotype (clone EC1-17)
has 99% similarity to Stenotrophomonas sp. YC-1 (Figure 4), which was isolated from an
chlorpyrifos
manufacturing
wastewater
treatment
plant
on
the
basis
of
organophosphorous pesticide degradation (Yang, et al., 2006). Two additional isolates
which have 99% similarity to EC1-17 include Stenotrophomonas acidaminiphila sp.
nov., which is a strictly aerobic bacterium isolated from an upflow anaerobic sludge
blanket reactor (Assih, et al., 2002), and Stenotrophomonas sp. BO from a
microaerophilic denitrification reactor fed methane (Costa, et al., 2000). The latter isolate
can carry out denitrification with acetate to predominantly N2O and to a lesser extent, N2.
Stenotrophomonas nitritireducens L2 which has a 98% similarity to EC1-17 was isolated
from a biofilter treating ammonia (Finkmann, et al., 2000), and the strain was able to
reduce NO2- to N2O but could not reduce NO3-.
The phylotype belonging to the Flavobacteria was recovered only from the pond
sediment-derived cultures (EC3 and MC). This phylotype contains 2 clones (EC3-11 and
MC-9), that have 99% similarity with Kaistella koreensis strain Chj707, which was
isolated from a fresh water stream in Korea and it is known to reduce nitrate (Kim, et al.,
2004).
A phylotype belonging to the α-Proteobacteria was recovered from EC1. This
phylotype (EC1-33) has 99% similarity to Amorphomonas oryzae strain B26 in the
family Rhizobiaceae (Fig. 4). Amorphomonas oryzae is a free-living nitrogen-fixing
bacterium isolated from rice roots. There are also closely related Rhizobium-
132
Sinorhizobium strains that are known for their ability for aerobic chemolithoautotrophic
oxidation of arsenite such as strains NT-2, NT-3 and NT-4, which have 96% similarity
with EC1-33 (Santini, et al., 2002). Other bacteria that are related included aerobic
As(III)-oxidizing strains, Rhizobium NT-26 (Santini, et al., 2000) Sinorhizobium sp.
GW3 (EF550173) and Agrobacterium GW4 (EF550174). The latter two strains were
isolated from sediments naturally enriched with arsenic (Fan, et al., 2008). There is also
a precedent in the “Rhizobium” cluster for anoxic oxidation of As(III). Sinorhizobium
strain DAO10 (99.8% similarity with EC1-33) is reported to be As(III)-oxidizing
denitrifying bacterium (Rhine, et al., 2006). DAO10 was shown to contain an arsenite
oxidase like gene with 73.6% amino acid similarity with arsenite oxidase (AroA) from
NT-26 (Rhine, et al., 2007).
The last phylotype recovered belongs to the β-Proteobacteria and was recovered
from EC1. The clone EC1-12 has 97 to 98% similarity to uncultured bacterial clones
(AY548933 and AY548944) from the order Burkholderiales (Fig. 4) that were obtained
from anaerobic ammonium oxidizing (Anammox) microbial communities. The family
Oxalobacteraceae
within
the
Burkholderiales
contains
at
least
one
species,
Herminiimonas arsenicoxydans strain ULPAs1T (AY728038), which is known for
aerobic As(III) oxidation and contains arsenite oxidase genes (Muller, et al., 2006).
133
3.6. Conclusions
Clone library and FISH analysis suggest that phylotypes representing Azoarcus and
Comamonadaceae are the main As(III) oxidizers in the microbial communities of the
three ECs studied. For Azoarcus, this conclusion is supported by the fact that a related
isolate from this genus (DAO1) can link As(III) oxidation to denitrification (Rhine, et al.,
2006). Further, the closely related isolate EbN1 contains arsenic resistance genes as well
as a full set of denitrification genes. For phylotypes related to the Comamonadaceae
cluster, this is the first report of anoxic As(III) oxidation. The fact that As(III) was
oxidized in the MC (which contained no Azoarcus) implies that the phylotype related to
Diaphorobacter is responsible for anoxic As(III) oxidation in the MC. This conclusion is
based on the finding that in the MC this phylotype was represented by the majority of the
clones, and that the FISH probe for the Comamonadaceae cluster demonstrated an
enrichment of the cells belonging to this phylogenetic group compared to the ECs. The
occurrence of aerobic As(III) oxidizers in the Comamonadaceae cluster (Fan, et al.,
2008), together with the occurrence of arsenic resistance- and denitrification genes in the
genome of the closely related Acidovorax sp. strain JS42, support this potential role.
3.7. Acknowledgements
The work presented here was funded by a grant 2005AZ114G from the USGS National
Institute for Water Resources 104G and by grant 2 P42 ES04940-11 from the National
134
Institute of Environmental Health Sciences Superfund Basic Research Program, NIH.
Funding was also obtained from the Spanish Ministerio de Educacion y Ciencia including
grant CTM2006-04131/TECNO to J.L.S. and grant SAB2006-0087 to J.A.F. as well as a
pre-doctoral fellowship from the Comunidad Autonoma de Madrid to N.F. Also arsenic
analyses were performed by the Analytical Section of the Hazard Identification Core
(Superfund Basic Research Program grant NIEHS-04940).
135
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142
CHAPTER 4
ANOXIC OXIDATION OF ARSENITE LINKED TO CHEMOLITHOTROPHIC
DENITRIFICATION IN CONTINUOUS BIOREACTORS
4.1. Abstract
In this study, the anoxic oxidation of As(III) linked to chemolithotrophic denitrification
was shown to be feasible in continuous flow bioreactors. Biological oxidation of As(III)
was stable over prolonged periods of operation ranging up to 3 years in the continuous
denitrifying bioreactor with responsible bacteria immobilized in biofilm granules. The
microbial population adapted to high influent concentrations of As(III) up to 5.0 mM.
When 7.6 mM As(III) was fed, the As(III) oxidation process was severely inhibited until
the As(III) concentration was lowered again to 3.5 mM. When the biofilm granules were
incubated in a bioassay with 3.5 mM As(III) and 2.5 mM NO3-, the oxidation of As(III)
was shown to be linked to the complete denitrification of NO3- to N2 by demonstrating a
significantly enhanced production of N2 beyond the background endogenous production.
The N2 production corresponded closely to the expected stoichiometry of the reaction, 2.5
mol As(III) mol-1 N2-N.
Keywords: Chemolithotrophic, Arsenite, Denitrification, Anaerobic, Continuous bioreactor
143
4.2. Introduction
Arsenic (As) contamination in drinking water resources is a global health problem
affecting numerous countries, including India, Bangladesh, Chile, Mexico, Argentina and
China [Smedley and Kinniburgh, 2002]. Naturally occurring As-bearing minerals,
industrial, agricultural and mining activities are important sources of arsenic in the
environment [Oremland and Stolz, 2003; Sierra-Alvarez et al., 2004]. However, the
natural release of arsenic to groundwater from the weathering of arsenic bearing rocks
and sediments is the most important source [Smedley and Kinniburgh, 2002]. The
predominant forms of As are inorganic arsenate (As(V)) and arsenite (As(III)) in
circumneutral aqueous environments. As(III) is substantially more toxic than As(V)
[Abdullaev et al., 2001; Sierra-Alvarez et al., 2004], and tends to be more weakly bound
to the surface of clay and aluminum hydroxides [Manning and Goldberg, 1997; Lin and
Wu., 2001]. The microbial conversions between As species have a significant impact on
the mobility of As in the environment [Oremland and Stolz, 2005].
Diverse types of bacterial arsenate reductases have been recognized to reduce
As(V) to As(III). There are two main biochemical systems responsible for As(V)
reduction. One system is a detoxification mechanism, whereby As(V) is reduced by an
arsenate reductase (ArsC) and As(III) is exported from the cell by As(III)-specific efflux
pump (arsB) [Lloyd and Oremland, 2006; Silver and Phung, 2005]. The other system is
involved in anaerobic respiration where As(V) is utilized as a terminal electron acceptor
to support growth with various electron donors. As(V) is reduced by a periplasmic
144
dissimilatory As(V) reductase consisting of a larger molybdopterin subunit (arrA), and a
smaller [Fe-S] center protein (arrB) [Krafft and Macy, 1998; Malasarn et al., 2004; Silver
and Phung, 2005]. The reduction of As(V) is hypothesized to be a mechanism of As
mobilization in subsurface environments because the As(III) formed is more weakly
bound to sediments. If oxidizing conditions can be restored, As can be immobilized by
the formation of As(V) which is better adsorbed on soil metal hydroxides like those of
aluminum.
As(III)-oxidizing microorganisms are widespread in various aquatic, soil and
sediments [Stolz et al., 2006; Inskeep et al., 2007; Rhine et al., 2007]. As(III) oxidases
are heterodimeric structures with two subunits containing molybdopterin and Rieske ironsulfur [Rhine et al., 2007; Santini, et al., 2004; Stolz et al., 2006]. Heterotrophic and
chemoautotrophic microorganisms have been reported that are able to oxidize As(III)
under aerobic [Inskeep et al., 2007; Rhine, et al., 2007] and anaerobic conditions [Rhine
et al., 2006; Sun et al., 2008]. The oxidation of As(III) by heterotrophic microorganisms
is generally considered to be a detoxification mechanism as the microbes do not gain
energy from the reaction [Silver and Phung, 2005]. However, several chemolithotrophic
microorganisms are able to utilize As(III) as an electron donor coupled to CO2 fixation
for cell growth under aerobic conditions [Santini et al., 2002; Rhine et al., 2008; Inskeep
et al., 2007], conditions of partial nitrate reduction to nitrite [Oremland et al., 2002; Hoeft
et al., 2007], and complete denitrifying conditions [Rhine et al., 2006; Sun et al., 2008].
Several bioreactor technologies have been used to investigate the biological oxidation of
As(III) in aerobic environments [Simeonova et al., 2005; Battaglia-Brunet et al., 2006;
145
Seith and Jekel, 1997; Lievremont et al., 2003]. For example, an up-flow fixed bed
bioreactor was initially inoculated with the sterile pozzolana-attached As(III)-oxidizing
bacterial consortium (CAsO1), and then continuously fed with an autotrophic medium
(modified CSM medium) containing 1.33 mM As(III), the results showed that around
90% of the As(III) was oxidized to As(V) during the 60 days operation [Michel et al.,
2007].
Although aerobic microbial oxidation of As(III) occurs readily and rapidly, this
process would be prevented in subsurface aqueous environments where dissolved oxygen
(DO) is absent. Nitrate can considered as an alternative oxidant with advantages over
elemental oxygen due to its high solubility and lower reactivity which together enable it
to be better dispersed in the saturated subsurface [Nolan et al., 1997]. Chemolithotrophic
denitrifying bioreactors have been proposed for groundwater treatment utilizing various
inorganic electron donors such as elemental S (So) [Sierra-Alvarez et al., 2007], sulfide
(H2S) [Kleerebezem and Mendez, 2002], hydrogen (H2) [Lee and Ritaman, 2002;
Rezania et al., 2007; Smith et al., 2005], [Fe(II)EDTA]2- [Kumaraswamy et al., 2006;
Maas et al., 2006].
The objectives of this research were to explore whether As(III) can be efficiently
oxidized with addition of nitrate in a continuous denitrifying bioreactor under anoxic
conditions over prolonged periods of operation. The specific goals of this study was to
demonstrate the dependence of the anoxic oxidation of As(III) on the presence of nitrate,
determine if the stoichiometry of the reaction is due to the complete denitrification and
acclimate the microbial population to high toxic influent concentration of As(III).
146
Two bioreactor studies were conducted. In the first study, two bioreactors were
seeded with chemolithotrophic sulfoxidizing denitrifying sludge and were operated in the
parallel with and without nitrate in the medium to confirm the dependence of the reaction
on nitrate. A second study utilized a larger bioreactor seeded with anaerobic granular
sludge from a methanogenic reactor to determine if the biomass could be enriched for
chemolithotrophic As(III)-oxidation linked to denitrification. The dependence of
denitrification on As(III) was evaluated by interrupting the As(III) feed during specific
periods.
4.3. Material and Methods
4.3.1. Microorganisms
The two bioreactors (R1 and R2) of the first study were inoculated with
chemolithotrophic sulfoxidizing denitrifying granular sludge, which was obtained from a
laboratory-scale
thiosulfate-oxidizing
denitrifying
enrichment
bioreactor
(TDE)
(Chemical and Environmental Engineering Department, University of Arizona). The
larger bioreactor column (R3) of the second study was inoculated with methanogenic
granular sludge obtained from a full-scale upward flow anaerobic sludge blanked
(UASB) treating alcohol distillery wastewater (NGS) (Nedalco, Bergemop Zoom, The
Netherlands). The total suspended solids’s (TSS) contents of the sludge TDE and NGS
was 8.06±0.13 and 6.32±0.27% on a wet weight basis, repectively; and the volatile
147
suspended solids’s (VSS) content was 5.76±0.18 and 5.96±0.28% on a wet weight basis,
respectively. The granular sludge samples were washed and sieved before use to remove
fines. The inocula were stored under nitrogen gas at 4°C.
4.3.2. Basal medium
The standard basal medium for continuous column prepared using ultra pure water (MilliQ system; Millipore) and contained the following compounds (mg l-1): NH4HCO3 (3.16);
NaHCO3 (672); CaCl2 (10), MgSO4.7H2O (40); K2HPO4 (300); KH2PO4.2H2O (800);
Yeast extract (0.3) and 0.2 ml l-1 of a trace element solution containing (in mg l-1):
FeC13.4 H20 (2,000); CoCl2. 6 H20 (2000); MnCl2 4 H20 (500); AlCl3 6 H20 (90);
CuCl2.2H20 (30); ZnCl2 (50); H3BO3 (50); (NH4)6Mo7O24.4 H2O (50); Na2SeO3.5 H2O
(100); NiCl2.6 H20 (50); EDTA (1000); resazurin (200); HCl 36% (1 ml).
4.3.3. Continuous columns
Two laboratory-scale flow-through columns of the first study (420 ml), R1 and R2, were
operated in parallel (Figure 1A) under anoxic conditions. Both reactors were seeded with
12.44 g VSS l-1 inoculum at the beginning of the operation. One column (R1) was
148
continuously fed with As(III) and nitrate; the other column (R2) was a control reactor
without nitrate that was operated to verify there were no other factors contributing to
As(III) oxidation (e.g. contamination of O2). R1 and R2 were supplied with a basal
medium containing As(III) (0.5, 2.5 and 3.75 mM, depending on the period of operation)
as electron donor, nitrate (2.5 and 6.4 mM) as electron acceptor, and bicarbonate as major
carbon source. The feed of the control-reactor (R2) was prepared similarly but nitrate
addition was omitted. The columns were operated for a period of 608 days and the
operation was divided into five periods, which could be distinguished based on the
concentration of arsenic as shown in the Table 1.
Table 4.1 Operational periods for the two continuous columns of the first continuous experiment
Nitrate Reactor (R1)
Period
Days
Control Reactor (R2)
Influent
Influent
HRT (d)
*
HRT (d)
As(III) (mM)
NO3- (mM)
As(III) (mM)
Avg.
Std
Avg.
Std
Avg.
Std
Avg.
Std
Avg.
Std
I
0-153*
_
_
_
_
_
_
_
_
_
_
II
154-309
1.06
0.13
2.89
0.40
6.61
0.65
1.06
0.09
2.79
0.37
III
310-511
1.04
0.14
3.75
0.32
6.48
0.49
1.05
0.28
3.59
0.77
IV
512-540
1.08
0.04
0†
0†
6.82
0.26
1.08
0.04
0†
0†
V
541-608
1.10
0.03
3.63
0.22
6.25
0.32
1.09
0.03
3.52
0.31
: Cultivation period, reactor was fed 0.5 mM As(III) and 2.5 mM NO3- with an approximate HRT of 1 d.
The results were monitored only on an occasional basis. †: As(III) was not included in the feed of medium.
149
The larger column (R3) of the second experiment was a 2-l bench-scale upwardflow anaerobic sludge bed (UASB) reactor (Figure 1B). The goal was to enrich the
nitrate-dependent As(III)-oxidizing microorganisms linked to chemolithoautotrophic
denitrification from a methanogenic consortium. This reactor was inoculated with 27 g
VSS l-1 of granular methanogenic anaerobic sludge. The reactor was fed with basal
mineral medium and As(III) (concentrations vary according to operational periods as
shown in Table 2) as the sole energy source, nitrate (6.57±0.46 mM) as the sole electron
acceptor and NaHCO3 as the major carbon source.
The average hydraulic retention time of the columns in both experiments was
approximately 1 d. All columns were placed in a climate controlled room at 30±2°C and
covered by aluminum foil to avoid growth of phototrophic bacteria. The pH of the
influent was adjusted to 7.2 with NaOH or HCl, as required. The influent was maintained
at all times under an N2 atmosphere to minimize dissolved oxygen from entering the
medium which could result in the unwanted aerobic oxidation of As(III). Influent and
effluent samples were prepared immediately for analysis to minimize possible changes in
arsenic speciation upon exposure to the atmosphere. The pH value was determined
immediately after sampling. Samples for analysis of arsenic speciation, nitrate and nitrite
were centrifuged (10,000 rpm. 10 min) or membrane filtered (0.45 µm) prior to dilution.
The samples for analysis of arsenic speciation were stored at -20°C, and nitrate and nitrite
samples were stored at 4 °C.
150
Table 4.2 Operational periods for the UASB reactor of the second continuous experiment
Influent
HRT (days)
Period
†
Days
As(III) (mM)
NO3- (mM)
Avg.
Std
Avg.
Std
Avg.
Std
I
0-158
1.05
0.15
0.40
0.05
6.52
0.26
II
160-223
1.11
0.16
1.60
0.15
6.50
0.19
III
225-284
1.03
0.22
2.47
0.30
6.73
0.37
IV
286-347
1.06
0.22
3.78
0.60
6.69
0.39
V
349-403
1.05
0.21
0†
0†
6.45
0.29
VI
405-475
1.07
0.18
3.46
0.37
6.52
0.36
VII
477-566
1.02
0.16
5.24
0.64
6.74
0.58
VIII
568-676
1.06
0.20
7.63
0.55
6.34
0.41
IX
678-725
1.14
0.31
8.08
1.40
6.35
0.16
X
727-798
1.12
0.22
4.04
0.20
6.29
0.55
XI
800-927
1.09
0.14
0†
0†
6.76
0.22
XII
929-954
1.02
0.09
0.98
0.14
7.15
0.28
XIII
956-1017
0.97
0.24
2.20
0.45
6.87
0.21
XIV
1019-1045
1.09
0.07
3.66
0.09
7.09
0.12
: As(III) was not included in the feed of medium.
151
Figure 4.1. Panel A, anaerobic bioreactor R1 and R2 for the first continuous experiment;
panel B, schematic diagram of the UASB bioreactor R3 for the second continuous
experiment.
152
4.3.4. Batch bioassay
Batch bioassays were performed in shaken flasks, which were incubated in a dark
climate-controlled room at 30±2˚C. Serum flasks (160 ml) were supplied with 120 ml of
a basal mineral medium (pH 7.0-7.2): NH4HCO3 (3.16); NaHCO3 (2,150); CaCl2 (10),
MgCl2 (83), MgSO4.7H2O (10); K2HPO4 (320); KH2PO4.2H2O (850); and 0.2 ml l-1 of a
trace element solution as described before. The medium was also supplemented with
arsenite (As(III)) as electron donor (concentration indicated in Tables and Figures of each
experiment) and nitrate as the electron acceptor, typically 10 mM, unless otherwise
specified. Flasks for anoxic assays were sealed with butyl rubber stoppers, and then the
medium and headspace were purged with N2/CO2 (80/20, v/v) for 20 min to exclude
oxygen from the assay. Various controls (e.g. abiotic controls, controls without nitrate,
controls without As(III) etc.) were applied based on the requirements of each experiment.
Abiotic controls were prepared by excluding the addition of microbial inoculum. Controls
lacking As(III) were included to measure endogenous consumption of nitrate so that the
net nitrate reduction linked to As(III) oxidation could be estimated. Controls without
NO3- were used to estimate if there were other factors contributing to As(III) oxidation.
All assays were conducted in triplicate. Samples were taken from sealed anaerobic serum
flasks by piercing the stoppers using 1.0 ml of sterile syringes with 16-gauge needles.
Liquid samples were analyzed for the concent of electron acceptor (NO3-),
biotransformation product (NO2-) and inorganic arsenic species (As(III) and As(V)).
153
4.3.5. Batch assays to determine the terminal product of denitrification
End products of denitrification were measured by monitoring gaseous nitrogen species in
the headspace of bioassays flushed with He/CO2 (80/20, v/v) in lieu of N2/CO2 as
described above. Microbial reduction of NO3- coupled to As(III) oxidation was assessed
in shaken batch bioassays inoculated with highly enriched nitrate-dependent As(III)oxidizing bacteria from bio-film of nitrate bioreactor (R1). The NO3- concentration
utilized was 2.5 mM for the experiments. Gaseous nitrogen was measured in As(III)
spiked assays (3.5 mM) and compared to that measured in endogenous controls lacking
the As(III) spike. Headspace samples were analyzed periodically for N2 and N2O with a
pressure lock gas tight syringe (1710RN, 100 µl (22s/2‫״‬/2), Hamilton Company, Reno,
Nevada USA) to confirm denitrification. Liquid samples were analyzed for NO3- and
NO2-. Flushed headspace controls incubated with just water were monitored to ensure
background levels of N2 were low.
154
4.3.6. Analytical methods
As(III) and As(V) species were analyzed by high performance liquid chromatography–
inductively coupled plasma–mass spectroscopy (HPLC-ICP-MS) by the University of
Arizona Superfund Basic Research Program Hazard Identification Core. The system
consists of an HPLC (Agilent 1100) and an ICP-MS (Agilent 7500a) with a Babington
nebulizer as the detector. The operating parameters were as follows: Rf power 1,500
watts, plasma gas flow 15 l min-1, carrier flow 1.2 l min-1, arsenic was measured at 75
m/z and terbium (IS) measured at m/z 159. The injection volume was 10 µl. The
detection limit for the various arsenic species was 0.1 µg l-1. The total concentration of
arsenic in liquid samples was determined by direct injection into the ICP-MS using an
ASX500 auto sampler (CETAC Technologies, Omaha, NE). The analytical system was
operated at an Rf power of 1,500 W, a plasma gas flow of 151 min-1 and a carrier gas
flow of 1.21 min-1. The acquisition parameters used were as follows: arsenic measured at
m/z 75; terbium (IS) measured at m/z 159; three points per peak; 1.5 s dwell time for As,
1.5 s dwell time for Tb; number of repetitions =7.
Nitrate, nitrite and arsenate (As(V)) were analyzed by suppressed conductivity ion
chromatography using a Dionex 500 system (Sunnyvale, CA, USA) fitted with a Dionex
IonPac AS11 analytical column (4mm x 250 mm) and a AG16 guard column (4 mm x 40
mm). During each injection the eluent (20 mM KOH) was used for 20 min.
155
N2 and N2O were analyzed using a Hewlett Packard 5890 Series II gas
chromatograph fitted with a CarboxenTM 1010 Plot column (30 m x 0.32 mm) and a
thermal conductivity detector. The temperatures of the column, the injector port and the
detector were 220, 110 and 100°C, respectively. Helium was used as the carrier gas and
the injection volume was 100 µl.
Ammonia was analyzed spectrophotometrically at a wavelength of 510 mm. The
analytical determination was conducted according to Standard Methods (APHA 1999).
Other analytical determinations (e.g. pH, TSS, VSS, etc.) were conducted according to
Standard Methods [APHA, 1999].
4.4. Results
4.4.1. Nitrate-dependent oxidation of As(III) to As(V) in continuous bioreactor under
anoxic conditions
Two laboratory continuous bioreactors was operated for a period of 608 days and the
summary of performance data for the two columns is presented in Table 3 for the total
duration of this study. The microbial oxidation of soluble As(III) and the concomitant
formation of soluble As(V) in the first experiment with bioreactors R1 and R2 as a
function of operation time are illustrated in Figure 2. The removal efficiency of As(III) in
column R1 averaged 93.8±7.6% for the combined results of period II, III and V. The
156
Figure 4.2 Concentration of As(III) and As(V) in the continuous bioreactor R1 and R2 as a function of time: As(III) (panel A
and C), ()- Influent, (▲)- Effluent; As(V) (panel B and D), (○) Influent, (●) Effluent.
157
conversion of As(III) removed to As(V) formed in column R1 was more than 90.0%
during the same periods. In comparison, the removal of As(III) and the formation of
As(V) were negligible in the control column R2 lacking nitrate. The results indicated that
the microbial oxidation of As(III) in R1 could be attributed to the addition of nitrate, and
its use as an electron acceptor for As(III) oxidation. Likewise the absence of
contaminating O2 in the influent was verified by the lack of any significant As(IIII)
oxidation in R2.
In period I (Day 0-153), the concentration of As(III) in the influent of both
reactors was 0.5 mM. The initial period was utilized for the cultivation of enrichment in
the sludge capable of linking arsenite oxidation to denitrification. After the day 153, the
influent concentration of As(III) was raised to 2.9 mM.
In period II (Day 154-309) and III (Day 310-511), the influent concentration of
As(III) was kept average of 2.9 and 3.75 mM, respectively, while the concentration of
nitrate was constant at 6.4 mM. Starting from day 154, the dominant arsenic species in
the effluent of R1 was As(V), indicating the occurrence of microbial As(III) oxidation
under denitrifying conditions. The formation of As(V) in the effluent corresponded with
the almost stoichiometric removal of As(III) in the influent, indicating that As(V) was the
main product of the conversion. In contrast, negligible oxidation of As(III) in R2 was
observed in the absence of nitrate. Figure 3 A and B illustrates the oxidation efficiency of
As(III) to As(V) in both columns during the period II and III, respectively. As(III) was
efficiently oxidized in the nitrate-amended column (R1) with removal efficiencies of
87.6±8.8 and 92.7±10.6%, respectively, compared with nearly no oxidation in the control
158
column (R2). The volumetric loadings of As(III) for column R1 was 179±35 and
259±53% mg As Lr-1 d-1 for period II and III, respectively, where Lr refers to the empty
bed volume of the reactors.
In period IV (Day 512-540), no As(III) was fed to either column. The endogenous
consumption of nitrate by organic matter present in the bio-film was estimated and used
to correct the ratio of As(III) removal to NO3- consumption in the R1 column. Results of
the consumption of nitrate are presented by Figure 4 including periods when no As(III)
was added to determine the endogenous consumption of NO3-. The effluent concentration
of NO3- started to increase, thus the NO3- consumption decreased when As(III) was
omitted from the feed in the influent. When As(III) was fed again, the gradual increase in
the As(III) oxidation to As(V) was reflected by a gradual decrease of NO3- concentration
in the effluent and increase for the NO3- consumptions.
In period V (Day 541-608), feeding of As(III) to both columns was reestablished
at 3.63 mM. After 28 d of operation without As(III), the anaerobic denitrifying microbes
in the R1 column readily oxidized As(III) to As(V). The reactor became fully stable after
a recovery period of 3 weeks. During this 3 week, recovery at the start of period V, a
gradual increase in As(V) formation (Fig 3B) was paralleled by a gradual decrease in
effluent NO3- concentration. The arsenic removal efficiency of 98.1% was similar to the
previous efficiency in period III of 94.9%. The As(III) removal efficiencies for the R1
were very stable with values of 88%, and the ratio of As(V) formed to As(III) removed
was more than 90%.
159
4.0
A
Arsenic (mM)
3.0
2.0
1.0
0.0
4.0
B
Arsenic (mM)
3.0
2.0
1.0
0.0
4.0
C
Arsenic (mM)
3.0
2.0
1.0
0.0
R1 Infl
R1 Effl
R2 Infl
R2 Effl
Figure 4.3 Comparison of the period average As(III) and As(V) concentrations in the
influent and effluent of bioreactors R1 and R2 during: Period II (Panel A, day 154-309),
III (Panel B, day 310-511) and V (Panel C, day 541-608). As(III): solid block; As(V):
empty block.
160
4.4.2. Nitrate reduction coupled to oxidation of As(III) to As(V)
Figure 4 shows the nitrate consumption in column R1 as a function of bioreactor
operation time.
The nitrate concentration supplied (6.4 mM) was in excess of the
concentration required for stoichiometric conversion of As(III) to As(V). The
stoichiometric requirement is 1.0 and 1.5 mM NO3- for 2.5 and 3.75 mM As(III),
respectively. In period IV (Day 512-540), the endogenous consumption of nitrate in the
absence of As(III) was measured as 1.18±0.31 mM. The molar ratio of As(V) formed
compared to NO3- consumed involved in the coupled reactions was calculated from the
formation of As(V) (As(V)) and the corrected NO3- (NO3-) consumption (NO3consumption corrected for the endogenous nitrate consumption in period IV). For the
periods II, III and V, the calculated molar ratios of As(V): NO3- are presented in Table
3. These ratios are very close to the theoretical stoichiometry ratio of 2.5 for As(III)
oxidation linked to complete denitrification of NO3- to N2 as shown in equation 1. The
production of nitrite (NO2-) and ammonium (NH4+), two possible products from the
microbial degradation of nitrate, were not detected in the effluent of R1 and R2
throughout the experiment. These findings indicate that nitrate was completely denitrified
to the benign end product dinitrogen gas (N2).
5H3AsO3 + 2NO3- → 8H+ + 5HAsO42- + N2 + H2O
[Eq. 1]
161
Period II
Period III
Period IV Period V
500
400
300
200
-
-1
NO3 concentration (mg L )
600
100
0
150
200
250
300
350 400 450
Time (days)
500
550
600
Figure 4.4 The influent and effluent concentrations of NO3- determined in the continuous
bioreactor R1 as a function of time: (◊) - Influent, (♦) - Effluent. Note: Phase IV is the
endogenous period without feeding of As(III).
4.4.3. Kinetics of chemolithotrophic As(III) oxidizers in the sludge in the column R1
To study the specific activity of granular sludge taken from bioreactor R1, a batch assay
was conducted with basal medium amended with 0.5 and 1.5 mM As(III) as the electron
donor, 5 mM NO3- as the electron acceptor. The time course of the elimination of As(III)
and the formation of As(V) are illustrated in Figure 5 A and B. The maximum specific
162
activities normalized to volatile suspended solids (VSS) are 0.98±0.04 and 0.73±0.02
mmol As(V) formed g-1 VSS d-1 for 0.5 and 1.5 mM As(III), respectively.
0.6
0.5
0.5
0.4
0.4
0.3
0.3
0.2
0.2
0.1
0.1
0.0
0.0
1
2
3
4
5
6
2.4
7
8
9
As(V) concentration (mM)
0.6
0
As(III) concentration (mM)
0.7
A
10
B
2.4
2.0
2.0
1.6
1.6
1.2
1.2
0.8
0.8
0.4
0.4
0.0
0.0
As(V) concentration (mM)
As(III) concentration (mM)
0.7
0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16
Time (days)
Figure 4.5 Elimination of As(III) and formation of As(V) by R1 biofilm at 0.5 (Panel A)
and 1.5 mM (Panel B) As(III) under denitrifying conditions, As(III) (●) and As(V) (○) by
the abiotic treatment, As(III) (■) and As(V) (□) supplied with As(III) and NO3-; As(III)
(▲) and As(V) () supplied with As(III), but lacking NO3-. The inoculum concentration
was 0.3 g VSS l-1.
163
To better understand the growth kinetics, batch assays were established with a
very low inoculum concentration of 0.05 g VSS l-1. The experiment contained a basal
medium amended with 1.0 mM As(III) as the electron donor, 2.5 mM NO3- as the
electron acceptor. The results (Figure 6A) demonstrated complete oxidation of As(III) to
As(V) within 48 d. The formation of As(V) corresponded to an almost stoichiometric
elimination of As(III). The As(III) removal and As(V) formation was 0.94±0.01 and
0.92±0.12 mM, respectively. Figure 6B showed the time course of the N2-N formation as
the end product of nitrate reduction in all the treatments. The actual measured NO3--N
reduction and N2-N production (corrected for the endogenous nitrate consumption and N2
production) was 0.37±0.03 and 0.36±0.07 mmol lliquid-1 for NO3--N reduction and N2-N
production, respectively, which close to 0.38 and 0.37 mmol lliquid-1 as expected linked to
anoxic oxidation of As(III). These results support that energy for growth was obtained
from the autotrophic oxidation of As(III) coupled to complete denitrification of NO3- of
N2. The specific activity of denitrification calculated on the N2-N formation curve was
1.46±0.20 mmol N2-N g VSS-1 d-1, which is close the activity of As(V) formation was
1.01±0.24 mmol As(V) formed g VSS-1 d-1. The growth rate of the population of As(III)oxidizing bacteria with an initial As(III) concentration of 1 mM was estimated from the
ln(∆As(V) and ln(∆N2) versus time insert graphs (Figures 6A and 6B). The doubling
times were estimated to be 9.7 to 13.5 d.
164
Figure 4.6 Panel A, formation of As(V) under denitrifying conditions, As(V) (●) by the
abiotic treatment, As(V) (■) by the R1 bio-film supplied with As(III) and NO3-; As(V)
(▲) by the R1 biofilm supplied with As(III), but lacking NO3-; panel B, formation of N2
in all the treatments link As(III) oxidation to denitrification, Abiotic control (○); R1 biofilm supplemented with As(III) and nitrate (■); with As(III) (♦); with nitrate (▲).
165
4.4.4. Terminal products of autotrophic denitrification linked to As(III) oxidation to
As(V)
Continuous flow bioreactor studies suggested that microorganisms in R1 column had the
ability to oxidize As(III) (3.75 mM) to As(V) at high concentrations of As(III) in the
influent. The measured molar ratio of As(III) removal or As(V) formation compared to
the corrected NO3- consumption (Table 3) was very close to stoichiometric ratio of 2.5
expected for complete denitrification. Therefore N2 was the expected end product of the
chemolithotrophic denitrification process. The N2 gas production could not be carefully
monitored in the R1 column. Therefore, a batch experiment was set up with bio-film
obtained from R1 in order to confirm that N2 gas was the end product of nitrate reduction.
The results shown in Figure 7 demonstrate the time course of As(III) removal and
As(V) formation in the presence of nitrate, as well as in control treatments without
nitrate, abiotic control treatments and an additional treatment amended with acetylene
(C2H2) in the headspace to inhibit the final step of denitrification and favor accumulation
of N2O. The C2H2 was injected on d 15 to avoid an initial inhibition of the reaction. In all
treatments, the formation of As(V) corresponded to an almost stoichiometric elimination
of As(III). The results indicated that the enriched microbial consortium could readily
oxidize all As(III) to As(V) within 60 days in the presence of nitrate as electron acceptor
without any lag phase. Before C2H2 was added, there was already 31.7% of As(III)
oxidized to As(V). By the end of the experiment, only 52.02±13.9% of the initial As(III)
166
in the treatments with C2H2 was oxidized to As(V), compared with 99.80±0.24%
conversion in the treatment lacking C2H2.
Table 4.3 Summary of As(III) oxidation linked to denitrification in bioreactor R1
Parameters
Period
Units
II
As(III) volumetric load
III
mg As lr-1d-1 178.9±34.8 259.2±52.8
IV
V
—
242.9±14.8
As(III) removal efficiency
%
88.38±7.2 92.71±10.6
—
98.58±1.8
As(III) removed
mM
2.52±0.38 3.56±0.57
—
3.55±0.22
As(V) formed
mM
2.39±0.31 3.32±0.48
—
3.33±0.27
mol mol-1 0.90±0.15 0.95±0.15
—
0.94±0.06
As(V) formed /As(III) removed
NO3- removed
mM
2.36±0.53 2.63±0.57 1.18±0.31 2.56±0.41
Corrected NO3- removed
mM
1.24±0.49 1.45±0.57
—
1.38±0.41
As(III) removed/corrected* NO3- removed
mol mol-1 2.49±0.98 2.75±0.90
—
2.83±0.74
As(V) formed/corrected* NO3- removed
mol mol-1 2.32±0.95 2.50±0.15
—
2.65±0.71
*
Corrected for endogenous nitrate consumption measured in period IV
4.0
4.0
3.0
3.0
2.0
2.0
1.0
1.0
0.0
0.0
0
8
15
21
28 35 42
Time (days)
49
56
62
As(V) concentration (mM)
As(III) concentration (mM)
167
69
Figure 4.7 Elimination of As(III) and formation of As(V) under denitrifying conditions,
As(III) (●) and As(V) (○) by the abiotic treatment, As(III) (■) and As(V) (□) by the R1
sludge supplied with with As(III) and NO3-, but lacking C2H2; As(III) (▲) and As(V) ()
by the R1 sludge supplied with As(III), NO3- and C2H2). The C2H2 was injected into the
headspace of bottles on day 14.
The possible products from nitrate reduction, including NO2-, NO, N2O, N2 and
NH4+, were also monitored to determine the fate of nitrate in the anoxic oxidation of
As(III). Figure 8 A and 8B demonstrates that N2 was the only end product in the
treatment without C2H2. In the treatment with C2H2, N2O had accumulated and no further
reduction was observed. The results indicate direct evidence that N2 was formed from
nitrate reduction linked to oxidation of As(III) to As(V). In the treatment without C2H2,
the formation of N2-N corresponded to 100.30±12.01% of the measured net removal of
NO3--N, also confirming that N2 was the only product of the conversion of NO3- due to
As(III). The As and N mass balances determined in this experiment are shown in Table 4.
168
The molar ratio of As: NO3- involved in the reaction was calculated from the As(V)
formed (As(V)) to the NO3- (NO3-) consumption corrected for NO3- removed in the
endogenous control and N2-N (N2-N) formation (corrected for the endogenous N2
formed). The calculated ratios of As(V) to NO3- and N2-N were 2.41±0.05 and
2.42±0.25, respectively, which are very close to the theoretical stoichiometric ratio of 2.5
(Eq. 1) for As(III) oxidation linked to complete denitrification. Accumulations of NO2-,
NO, or NH4+ were not detected in any of the treatments. These results confirm that the
bio-film in R1 is capable of complete denitrification to the benign N2 end product when
utilizing As(III) as the electron donor.
169
170
Figure 4.8 Formation of N2 (Panel A) and N2O (Panel B) in all the treatments under
denitrifying conditions, Abiotic control (○); R1 sludge supplemented with As(III)/nitrate
and no C2H2 (■), with As(III)/nitrate and C2H2 (♦); with nitrate and no C2H2 (▲); with
nitrate and C2H2 (X).
171
4.4.5. Growth with various electron acceptors
Bio-film from granular sludge in R1 column were tested for their ability to link As(III)
oxidation to the various electron acceptors. Oxygen, nitrate, nitrite, chlorate and
perchlorate were chosen as different possible electron acceptors for this enriched biofilm.
The R1 biofilm, incubated with 8 mM HCO3-, 0.5 mM As(III) was able to completely
oxidize As(III) to As(V) with either NO3- or O2 within 5 d. NO2-, ClO3- and ClO4- did not
support As(III) oxidation under anaerobic conditions (data not shown).
4.4.6. Anoxic oxidation of As(III) to As(V) linked to denitrification in continuous benchscale UASB bioreactor
A second continuous experiment using a larger volume bioreactor (R3) was initiated with
a methanogenic granular sludge with no prior exposure to As(III). R3 was operated for a
period of 1,045 days and the operation was divided into fourteen periods, which were
distinguished based on the concentration of As and the inclusion of NO3- as shown in the
Table 2. The volumetric loadings of As(III) ranged from 72-1443 mg As Lr-1 d-1 for all
the experimental periods, where Lr refers to the empty bed volume of the reactors. The
nitrate concentration supplied (6.4 mM) was in excess of the concentration required for
stoichiometric conversion of As(III) to As(V). The anoxic oxidation of As(III) and the
concomitant formation of As(V) in reactor as a function of time are illustrated in Figures
172
9 A and B, respectively. Table 5 summarizes the speciation of the identified soluble
arsenic compounds in the reactor for each period of operation.
The removal of As(III) from the influent was very stable averaging 96.5±6.6% in
all experimental periods with 5 mM As(III) or less. The dominant arsenic species in the
effluent was As(V) (Figure 9B), indicating the occurrence of microbial As(III) oxidation
under denitrifying conditions. The formation of As(V) in the effluent corresponded with
the almost stoichiometric removal of As(III) in the influent, indicating that As(V) was the
main product of the conversion. The conversion of As(III) removed to As(V) formed was
average at 94.34±16.6% during the same periods.
In period V (Day 349-403) and XI (Day 800-927), the endogenous consumption
of nitrate was measured as 0.79±0.18 and 0.30±0.24 mM respectively, when no As(III)
was added in the influent. Figure 10 shows the nitrate consumption as a function of
bioreactor operation time including two endogenous consumption periods. The effluent
concentration of NO3- started to increase, thus the NO3- consumption decreased when
As(III) was omitted from the feed in the influent. The endogenous NO3- consumption was
almost zero in the last As(III) feed interruption indicating biomass stabilized in the
bioreactor.
For the period of VI and XII when As(III) was fed again, the anaerobic
denitrifying microbes readily oxidized As(III) to As(V). The reactor became fully stable
after a recovery period of less than 2 weeks. During the recovery process, a gradual
173
increase in the As(III) oxidation to As(V) (Figure 9) co-occurred with a gradual increase
of the NO3- consumptions.
The denitrifying bacteria in the bioreactor was able to tolerate up to 5 mM As(III)
in the Influent. When 7.6 mM or more As(III) was applied on the periods of VIII and IX,
there was a steady decline in the As(III) conversion efficiency that was remedied by
lowering the influent As(III) concentration back to 3.75 mM. It took 2-3 weeks for the
conversion efficiency to recover. And a gradual increase in As(V) formation (Figure 9B)
was paralleled by a gradual decrease in effluent NO3- concentration (Figure 10).
For all the experimental periods, the calculated molar ratios of As(V): NO3- are
presented in Table 5. The molar ratio of As(III) removed and As(V) formed compared to
nitrate consumed (corrected for background removal by the endogenous substrate) was
found to average 2.62±0.24 and 2.71±0.21, respectively for all experimental periods with
5 mM As(III) or less. These values are very close the theoretical stoichiometry of 2.5 for
complete denitrification as shown in Eq. 1. The production of nitrite (NO2-) and
ammonium (NH4+), two possible products from the microbial degradation of nitrate, were
not detected in the effluent of R3 throughout the experiment. These findings indicate that
nitrate was completely denitrified to the benign end product dinitrogen gas (N2).
174
Figure 4.9 Influent and effluent concentrations of As(III) (Panel A) and As(V) (Panel B)
during the operation of a 2-L bench-scale up-flow anaerobic sludge blanket reactor:
As(III): ()- Influent and (▲)- Effluent; As(V): (○)- Influent and (●)- Effluent.
175
176
4.5. Discussion
4.5.1. Microbial anoxic oxidation of As(III) linked to chemolithotrophic denitrification in
the continuous bioreactors
Biological denitrification has great potential for the remediation of groundwater and
drinking water [Kapoor and Viraraghavan, 1997; Rittmann and McCarty, 2001].
However, the full-scale applications have been hampered due to the concerns about
bacteria contamination and biofouling, when organic matters were used as electron
donors. Interest for chemolithotrophic denitrification can overcome the disadvantages of
heterotrophic dentrification by using inorganic electron donors. For example hydrogen
coupled to denitrification has been used in small-scale or pilot-scale bioreactors to
develop a cost effective system for treating nitrate-contaminated drinking-water supplies
[Smith et al., 2005; Lee and Rittamann, 2002]; a chemolithotrophic denitrifying
bioreactor packed-bed with S0:limestone granules (1:1, v/v) was utilized to explore the
feasibility to treat nitrate-contaminated groundwater [Sierra-Alvarez et al., 2007].
Therefore chemolithotrophic denitrifying bioreactors have the potential to utilize As(III)
as electron donor. The resulting As(V) formed would be expected to be more strongly
adsorbed by various metal oxides such as those of aluminum.
177
The anoxic oxidation of As(III) by the continuous bioreactors in this study
showed significant dependence on the presence of nitrate in the feed. The bioreactor R1
had a very high efficiency of As(III) removal in the presence of nitrate whereas in R2
without nitrate, the removal efficiency was negligible. Additional evidence for the
dependence on nitrate was observed in the batch experiments. When the biofilm granules
were inoculated with As(III), there was no activity in the batch assays lacking nitrate,
compared with the complete biological oxidation of As(III) in the presence of nitrate.
On the other hand, denitrifying microorganisms utilized the As(III) as the main
electron donor to reduce the nitrate. During the continuous column studies of R1 and R3,
the observed biological nitrate consumptions in periods fed with As(III) were much
higher than the endogenous consumption of nitrate in periods when NO3- was omitted
from the feed. Moreover, the NO3- consumption apparently decreased in the periods
without feeding of As(III), and increased again when As(III) was added again. The net
difference in NO3- consumption in periods with and without As(III) is referred to as the
“endogenous” consumption of NO3- due to native biomass as electron donor. To calculate
the nitrate consumption linked to As(III) oxidation, the total nitrate consumption was
“corrected” by subtracting, the endogenous nitrate consumption. The molar ratio of
As(III) removed (or As(V) formed) to the corrected NO3- consumption was
approximately equal to that of the theoretically expected value of 2.5 for complete
denitrification. After feed interruptions ranging up to 4 months, it took only 2-3 weeks
for the denitrifying bacteria in the nitrate bioreactor (R1 and R3) to recover the ability to
oxidize As(III) to As(V). Furthermore, the batch bioassays demonstrated that nitrate
178
consumption and N2 production were significantly higher in the biological treatments
containing As(III) than endogenous controls without As(III). The results from this study
provide compelling evidence that a chemolithotrophic denitrification process was
responsible for the microbial oxidation of As(III) to As(V).
Nitrate-dependent As(III)-oxidizing bacteria have been isolated from various
environments with historical arsenic contamination, for examples: Alkalilimnicola
ehrlichii sp. strain MLHE-1 from an arsenic-rich, alkaline hypersaline soda lake in
California (Mono Lake) [Hoeft et al., 2007], and strain DAO1 and strain DAO10 isolated
from arsenic polluted industrial soil [Rhine et al., 2006]. Beside the evidence of
microorganisms from arsenic-contaminated sites, As(III)-oxidzing denitrifying bacteria
were also indentified in mixed cultures established from sludges and sediments with no
prior exposure to As [Sun et al., 2008].
4.5.2. End production of denitrification linked to As(III) oxidation in the continuous
bioreactors
In this study, multiple lines of evidence point to the fact that microorganisms in the
biofilm of R1 and R3 reactor linked the anoxic oxidation of As(III) to complete
denitrification as opposed to partial denitrification to nitrite or dissimilatory nitrate
reduction to ammonia. Firstly, this was confirmed by the lack of any accumulation of
nitrite. Secondly, the stoichiometric ratio of As(III) oxidation to NO3-/NO2-, NO3-/N2 and
NO3-/NH4+ would be 1, 2.5 and 4, respectively; and the observed ratio in this study was
179
consistently, approximating 2.5 (Tables 3 and 5), corresponding to complete
denitrification to N2. Thirdly, the batch study with biofilm granules from R1 confirmed
the production of N2 due to the oxidation of As(III) and the yield of N2 corresponded to
the expected stoichiometry from the electron equivalents in As(III). The measured N2
was not an artifact of atmospheric N2, because controls had negligible background N2,
and inhibiting the reaction with C2H2 generated N2O which has no potential for artifact
interferences.
The first chemolithoautotrophic As(III)-oxidizing denitrifying bacteria isolated,
Alkalilimnicola ehrlichii sp. strain MLHE-1, only partially denitrifies nitrate to nitrite
[Oremland et al., 2002; Hoeft et al., 2007]. It has already been demonstrated that
Azoarcus strain DAO1 and Sinorhizobium strain DAO10 have the ability to completely
denitrify nitrate to N2 when oxidizing As(III), based on the stoichiometry of the reaction
and the amplification of the nitrous oxide reductase gene, nosZ [Rhine et al., 2006]. In
our previous study, the As(III) adsorbed on activated alumina was effectively utilized as
an electron donor for the denitrification to N2 gas. In parallel incubations N2O gas
accumulated without adsorbents due to the severe toxicity of non-attenuated aqueous
As(III) concentrations [Sun et al., 2008]. A 16S rRNA gene clone library characterization
of enrichment cultures indicated that the predominant phylotypes were from the genus
Azoarcus and the family Comamonadaceae [Sun et al., 2008]. These results confirm
beyond doubt that chemolithotrophic anoxic oxidation of As(III) was successfully linked
to the complete denitrification process, in which nitrate was reduced to the benign
gaseous product, N2.
180
4.5.3. As(III) substrate inhibition on anoxic oxidation of As(III) in continuous bioreactors
The toxicity of arsenic compounds to bacteria is of great environmental concern. But the
mode of toxicity depends on the chemical form of arsenic. The toxicity of As(V) is due
mainly to its similar structure and properties with phosphate. As(V) can replace
phosphate and inhibit oxidative phosphorylation, which is the main energy-generation
system for living organisms [Oremland and Stolze, 2003; Silver Phung, 2005]. As(III) is
more toxic than As(V) according to various studies [Abdullaev et al., 2001; SierraAlvarez et al., 2004], because it can enter the cell through aqua-glycerolporins, bind to
thiols or vicinal sulfhydryl groups and inactivate/denature the normal functions of many
proteins and enzymes, even finally causing cell damage [Mukhopahyay et al., 2002].
Although arsenic is toxic to many bacteria, some bacteria are resistant to arsenic
due to an efflux system mediated by the plasmid- or chromosomally-encoded ars operon
[Silver and Phung, 2005]. Some microorganisms also can evolve transformation
pathways to arsenic reduction, oxidation and methylation [Oremland and Stolz, 2003;
Qin et al., 2006; Inskeep et al., 2007].
This study demonstrates that the anoxic oxidation of As(III) linked to
denitrification can be sustained over prolonged periods of up to 3 years in continuous
bioreactors with bacteria immobilized in biofilm granules. The microbial population in
the continuous bioreactors (R1 and R2) acclimated to high concentration of As(III) up to
3.75 mM in the influent. The R3 reactor was operated within the concentration
181
increments of As(III) in the influent from 0.5 to 7.6 mM. R3 reactor tolerated As(III) up
to influent concentration of 5.0 mM, However, when the concentration of As(III) in the
influent was increased to 7.6 mM, inhibition of As(III) oxidation became evident and the
efficiency of As(III) conversion dramatically decreased compared to time periods fed
with lower concentrations of As(III). Such high concentrations of As(III) are highly toxic
to microorganisms [Stasinakis et al., 2003]. The 50% inhibitory concentrations of As(III)
to methanogenic activity are reported to be as low as 15 µM [Sierra-Alvarez et al., 2004].
Nine arsenic resistant strains exhibited an effective 50%-inhibiting concentration (EC50)
in the range of 0.27-9.73 mM [Takeuchi et al., 2007]. In our previous batch study, mixed
consortia from sludge and sediments without prior exposure to As demonstrated the
capacity to oxidize As(III) up to 3.5 mM linked to denitrification, but at that
concentration the rate was severely inhibited, compared to the rates observed at 0.5 mM
[Sun et al., 2008]. The denitrification was inhibited as reflected by the N2O accumulation
linked to oxidation of As(III) at 3.5 mM [Sun et al., 2008]. However, in this study, the
highly enriched sludge from R1 reactor could link readily the oxidation of 3.5 mM
As(III) to complete denitrification of nitrate to N2 without toxic inhibition.
4.6. Conclusions
(1)
Anoxic oxidation of As(III) linked to denitrification was feasible over prolonged
periods of operation in the continuous denitrifying bioreactor with bacteria
immobilized in bio-film granules.
182
(2)
During this continuous experiment, the bioreactor (R1) fed with nitrate (R1)
demonstrated a reliable high conversion efficiency with more than 90% the
influent As(III) oxidized to effluent As(V). In contrast, negligible oxidation of
As(III) was observed in the control bioreactor (R2) lacking nitrate.
(3)
In this study, the anoxic oxidation of As(III) was dependent on the presence of
nitrate; and vice versa, nitrate denitrification to N2 was dependent on As(III) as
the electron donor.
(4)
Dinitrogen gas is the end product of denitrification linked to As(III) oxidation.
(5)
The microorganisms in the continuous bioreactor could acclimate to the toxicity
of high concentrations of As and tolerated up to 5 mM As(III) in the influent.
4.7. Acknowledgments
The work presented here was funded by a USGS, National Institute for Water Resources
104G grant (2005AZ114G), and by a grant of the NIEHS-supported Superfund Basic
Research Program (NIH ES-04940). Also arsenic analyses were performed by the
Analytical Section of the Hazard Identification Core (Superfund Basic Research Program
grant NIEHS-04940).
183
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CHAPTER 5
THE ROLE OF DENITRIFICATION ON ARSENITE OXIDATION AND ARSENIC
MOBILITY IN ANOXIC SEDIMENT COLUMN MODEL WITH ACTIVATED
ALUMINA
5.1. Abstract
Arsenite (As(III)) is the predominant arsenic (As) species in reduced anaerobic
environments. As(III) is less strongly adsorbed than As(V) at circumneutral pH
conditions by common non-iron metal oxides in sediments such as those of aluminum.
Therefore, oxidation of As(III) to As(V) could contribute to an improved immobilization
of As by sediments and thus help mitigate arsenic contamination in groundwater.
Microbial oxidation of As(III) is known to readily occur in aerobic conditions, however,
aeration of groundwater in anaerobic subsurface zones maybe important due to the poor
solubility of oxygen and its high chemical reactivity with reduced compounds. Nitrate,
which commonly occurs in groundwater, can be considered as an alternative electron
acceptor with a higher solubility and lower chemical reactivity.
In this study, two sediment columns packed with activated alumina (AA) and
inoculated with a chemolithotrophic denitrifying sludge were utilized to demonstrate the
role of denitrification on As(III) oxidation and As mobility in anoxic environments
containing non-iron oxides. During the 250 days operation of the continuous experiment,
more than 75% of the cumulative influent As had adsorbed in the column fed with As(III)
190
together with nitrate. This was considerably more than the 50% retained in the control
column lacking nitrate. The solid-phase As extraction results revealed that As(V) was the
dominant As species retained on AA surfaces in the presence of nitrate; whereas As(III)
was predominant from in the absence of nitrate. This study validates a bioremediation
strategy based on the anoxic microbial nitrate-dependent oxidation of As(III) to more
readily adsorbed As(V) as a means to enhance the immobilization of As on non-iron
metal oxide particles in the subsurface environments.
Keywords: Arsenite, Arsenate, Nitrate, Oxidation, Activated Alumina, Adsorption
191
5.2. Introduction
Arsenic (As) is generally found as a contaminant in soil, sediments and water systems
[Smedley and Kinniburgh, 2002]. The occurrence of elevated As in groundwater or
surface water has been attributed to natural biogeochemical reactions such as weathering
and dissolution of As-containing minerals, which promote its release into groundwater
[Smedley and Kinniburgh, 2002]. The toxicity of arsenic is an important issue of current
global public health. Increasing evidences indicates that As contributes to non-cancer and
cancer diseases worldwide [WHO, 2001; ATSDR, 2007]. The U.S. Environmental
Protection Agency (EPA) has enacted a stricter drinking water standard for arsenic that
lowers the maximum contaminant level (MCL) from 50 ug l-1 to 10 ug l-1 [USEPA,
2001].
In natural environments, arsenic generally occurs as either arsenate (As(V))
encountering under oxidizing conditions, or arsenite (As(III)) predominating in reduced
anaerobic environments. At circumneutral pH values, As(V) occurs as deprotonated
oxyanions of arsenic acid (H2AsO4- and HAsO42-); As(III) exists as non-ionic form of
H3AsO3 [Cullen et al., 1989]. Of the two commonly occurring species, As(III) is
generally considered more mobile and toxic [Oremland and Stolz, 2003].
The main mechanism of immobilizing As is its sorption from the aqueous phase
onto the solid phase [Dixit and Hering, 2003]. Clay minerals are common constituents of
the subsurface environment, which act as adsorbents. As(III) and As(V) are adsorbed on
the surface of various clay minerals, which are largely consist of alternating layers of
192
silica oxide and Al oxide [Lin and Puls, 2000; Wang and Mulligan, 2006]. It is well
established that As(III) is less strongly bound compared to As(V) onto clay and minerals
with aluminum oxides with low content of iron [Lin and Wu, 2001; Hering and Dixit,
2005]. Activated alumina (AA) shows a 2-fold higher affinity for As(V) than As(III) at
pH 7 [Ghosh and Yuan, 1987], and kaolinite and montmorillonite were studied as
sorbents of As(III) and As(V) from landfill leachate and exhibited higher affinities for
As(V) than As(III) [Frost and Griffin, 1977]. Surface analyses demonstrate that As(V) is
adsorbed predominantly onto aluminum oxides by strong inner sphere complexes. In
contrast, the non-ionized form of As(III) is bound to alumina by the same inner-sphere
and weaker outer-sphere adsorption sites [Arai et al., 2001]. The outer-sphere complexes
were dominant in the pH range of 5.5-8, which decrease the binding strength of As(III)
on alumina under circumneutral conditions.
Therefore, oxidation of As(III) to As(V) could contribute to improve As sorption
onto clay mineral and other aluminum (hydr)oxides, thus alleviating arsenic
contamination in groundwater. The microbial oxidation of As(III) under aerobic
conditions is well known and has been reviewed in several references [Oremland and
Stolz, 2003; Inskeep et al., 2007]. However, aeration of groundwater in anaerobic
subsurface zones maybe important due to the poor solubility of oxygen and its high
chemical reactivity with reduced compounds. Recently evidence is growing that
microbial oxidation of As(III) can also occur under anoxic conditions in the presence of
nitrate [Oremland et al., 2002; Sun et al., 2008] or selenate [Fisher and Hollibaugh,
2008]. Nitrate may have a better potential to control As mobility in anaerobic
193
environments compared to oxygen, because of its higher solubility and lower chemical
reactivity. In our previous study, microbial nitrate-dependent oxidation of As(III) was
shown to be feasible over prolonged periods of operation in a continuous bioreactor with
bacteria immobilized in biofilm granules [Chapter IV]. Some studies have demonstrated
an improved removal of As in the activated alumina packed columns with an chemical
pre-oxidation of As(III) [Bissen and Frimmel, 2003].
The objective of this study was to evaluate the potential of injecting nitrate into
anaerobic groundwater as a potential bioremediation method to promote the anoxic
oxidation of As(III) by denitrifying bacteria. We set out to test the hypothesis using two
sediment columns packed with AA and fed with As(III) either in the presence or absence
of nitrate. The columns were used to simulate anaerobic sediments containing mostly
non-iron oxides. The specific goals were to (1) evaluate the adsorption isotherms of
As(III) and As(V) on the AA at neutral pH conditions; (2) investigate the potential of
biological anoxic oxidation of As(III) linked to nitrate reduction followed by As(V)
adsorption on AA to remediate groundwater contaminated with As.
5.3. Materials and methods
5.3.1. Microorganisms
A chemolithotrophic As(III)-oxidizing denitrifying granular biofilm was used to
194
inoculate the two sediment columns packed with AA. The inoculum was obtained from a
2-l bench-scale upward flow anaerobic sludge bed As(III)-oxidizing denitrifying
bioreactor after 466 days of operation. The bioreactor was fed with As(III) (3.46±0.37
mM) as electron donor, NO3– (6.47±0.31 mM) as electron acceptor, and NaHCO3 (8.0
mM) as major carbon source with a hydraulic retention time (HRT) of 1.12±0.21 d. The
total suspended solids’s (TSS) and volatile suspended solids’s (VSS) content of the
sludge was 6.08 ± 0.18% and 5.76 ± 0.23%, respectively, on a wet weight basis. The
granular biofilm was washed and sieved before use to remove fines. The sludge was
stored under nitrogen gas at 4ºC for one week.
5.3.2. Basal medium
The standard basal medium was prepared using ultra pure water (Milli-Q system;
Millipore) and contained the following compounds (mg l-1): NH4HCO3 (3.16); NaHCO3
(100); CaCl2 (10), MgSO4.7H2O (10); K2HPO4 (204); KH2PO4.2H2O (204); and 0.02 ml
l-1 of a trace element solution containing (in mg l-1): FeC13.4 H20 (2,000); CoCl2. 6 H20
(2,000); MnCl2 4 H20 (500); AlCl3 • 6 H20 (90); CuCl2.2H20 (30); ZnCl2 (50); H3BO3
(50); (NH4)6Mo7O24 • 4 H2O (50); Na2SeO3.5 H2O (100); NiCl2.6 H20 (50); EDTA
(1,000); resazurin (200); HCl 36% (1 ml). The pH of the basal medium was adjusted to
7.2 using concentrated NaOH or HCl, as needed.
195
5.3.3. Adsorption isotherms of As(III) and As(V) on activated alumina (AA)
Sorption isotherm experiments with activated alumina (regenerable AA-400G, Alkan
Chemicals, Cleveland, Ohio) were conducted in triplicate using glass flasks (160 ml)
supplemented with the sorbent (5.0 g dry weight) and 100 ml of a pH-7.2 sodium
bicarbonate buffer (33.3 mM) spiked with As(III) (as AsNaO2) or As(V) (as Na2HAsO4·7
H2O). The initial adsorbate concentrations were 0.6, 1.0, 2.0, 3.0, 4.0, 5.0, 7.0, 9.0, 15.0,
20.0, 35.0, 45.0, and 60.0 mg l-1 for As(III). The concentration tested for As(V) were the
same except that an As concentration of 80.0 mg l-1 was used instead of 60.0 mg l-1.
Control flasks lacking sorbent were run in parallel to correct for possible removal of
arsenic by other mechanisms than adsorption. Flasks were shaken under a N2/CO2
(80/20) atmosphere in an orbital shaker (150 rpm) for 2 d at 30±2°C. Removal of As(III)
and As(V) from solution was determined following centrifugation and filtration (0.45
µm) of the samples to remove insoluble matter.
Langmuir and Freundlich isotherms were used to describe the adsorption
equilibrium between adsorbed (q) As(III) and As(V) on the AA (mg As g-1 dwt AA) and
the equilibrium concentration (Ce) of As(III) and As(V) in solution (mg l-1). The
Langmuir isotherm is defined by:
Cs =
a ⋅ b ⋅ Ce
1+ b ⋅ Ce
[1]
196
Where Cs is the concentration of the solute in the solid phase (mg As g-1 AA), Ce is the
equilibrium concentration of the solute in solution (mg As l-1), a and b are Langmuir
adsorption constants; a represents the maximum achievable surface concentration of the
solute, and b is the equilibrium constant for the sorption reaction.
The Freundlich equation is an empirical relationship describing the sorption of
solutes from a liquid to a solid surface. The Freundlich equation is defined by:
C s = K F ⋅ Ce
1/ n
[2]
Where KF [(mg As g-1 AA)(mg As l-1)-1/n] is the Freundlich adsorption constant or
capacity factor and n is the Freundlich exponent which provides a measure for the
sorption intensity. The KF value represents the loading in mg adsorbate per gram of
sorbent at an aqueous equilibrium concentration of 1.0 mg l-1 of the compound.
The values of a and b of Langmuir isotherm were calculated from the slope and the
intercept of the linear plot of 1 q-1 versus 1 Ce-1, and the values of KF and 1 n-1 of
Freundilich isotherm were calculated from the slope and the intercept of the linear plot
between log q and log Ce.
5.3.4. Attenuation of As(III) under chemolithotrophic denitrifying conditions in sediment
columns packed with AA
Anoxic As(III) oxidation and immobilization of As(V) formed under denitrifying
conditions was investigated in two glass columns (each 420 ml) continuously fed with the
197
synthetic basal medium. The columns were placed in a climate controlled room at
30±2°C and covered with aluminum foil to avoid growth of phototrophic
microorganisms. Each reactor was packed with 350 dry weight of AA and inoculated
with 2.94 g VSS l-1 anaerobic biofilm obtained from the chemolithoautotrophic As(III)oxidizing denitrifying bioreactor. The full-treatment reactor (S1) was the biologically
active column inoculated with the chemolithotrophic As(III)-oxidizing denitrifying
biofilm and fed with basal medium, As(III) (6.7 uM) and nitrate (2.5 mM NO3-). The
control reactor (S2) was inoculated and fed As(III) in the same fashion as S1, but lacked
nitrate. Both reactors were supplied with bicarbonate (100 mg l-1) as the major carbon
source, except that 18.8 mg l-1 acetate was added to support the microbial consumption of
any possible traces of dissolved oxygen that could potentially enter from the medium and
cause unwanted aerobic oxidation of As(III). The influent of both reactors was
maintained at all times under a N2 atmosphere supplied via a gas bag (SKC-West Inc,
Fullerton, CA) to minimize any exposure to O2.
The columns were operated with an empty bed hydraulic retention time averaging
24 h (from day 0-30), and 12 h for the remainder of the experiment. Fresh liquid samples
were collected periodically from the influent and effluent lines and prepared immediately
for analysis to minimize possible changes in arsenic speciation upon exposure to the
atmosphere. The pH value was determined immediately after sampling. Samples for
analysis of arsenic speciation, total arsenic, nitrate and nitrite were centrifuged (10,000
rpm, 10 min) or membrane filtered (0.45 µm) prior to dilution.
198
5.3.5. Arsenic extraction
The total arsenic content and arsenic speciation in solid matrices (i.e., activated alumina,
biomass) were measured following extraction of the samples with 25 ml of NaOH (2.0
M) in anaerobic tubes (30.0 ml) under nitrogen gas immersed in a shaking water bath at a
temperature of 90±2°C for 12 h. After the extraction, the liquid samples were
immediately adjusted to pH lower than 1.0 with HNO3 (2.5 M) to avoid oxidation at
alkaline conditions when the samples first became exposed to air. Subsequently, the
samples were adjusted to a final pH of 6.0-6.5 with 2.0 M NaOH and membrane filtered
(0.45 µm) and stored in polypropylene vials at -20°C. A simple assay was set up to
confirm the preservation of As speciation during the alkaline extraction and sample
preservation processes. The results shown (Table S1) demonstrated that the extraction
process on solid-phase AA did not change the As speciation.
199
Table S1. Alkaline extraction protocol for solid-phase As on the AA
Adsorbed As
Extracted As
(ug As/g AA)
(ug As/g AA)
As recovery (%)
Initial As speciation
AA-As(III)
Speciation
As(III)
As(V)
As(III)
As(V)
98.7±7.5
ND
96.5±8.2
ND
AA-As(V)
Total As
†
*
recovery†
ND
109.5±0.6
ND
104.2±6.2
recovery*
97.7±1.09
95.9±8.4
Total As
Speciation
Sum of As Extracted As
99.4±1.0
95.2±1.68
102.0±7.0
97.9±1.1
: The sum of As recovery was calculated from the adsorbed sum(As(III) and As(V))/extracted sum(As(III)+As(V))*100.
: The extracted As recovery was calculated from the extracted sum(As(III) and As(V))/extracted total As*100.
200
5.3.6. Analytical methods
As(III) and As(V) speciation was analyzed by ion chromatography–inductively coupled
plasma–mass spectroscopy (HPLC-ICP-MS). The system consists of an HPLC (Agilent
1100) and an ICP-MS (Agilent 7500a) with a Babington nebulizer as the detector. The
operating parameters were as follows: Rf power 1500 watts, plasma gas flow 15 l min-1,
carrier flow 1.2 l min-1, arsenic was measured at 75 m/z and terbium (IS) measured at m/z
159. The injection volume was 10 µl. The detection limit for the various arsenic species
was 0.1 µg l-1. The total concentration of arsenic in liquid samples was determined by
direct injection into ICP-MS using an ASX500 auto sampler (CETAC Technologies,
Omaha, NE). The analytical system was operated at an Rf power of 1500 W, a plasma
gas flow of 151 min-1 and a carrier gas flow of 1.21 min-1. The acquisition parameters
used were as follows: arsenic measured at m/z 75; terbium (IS) measured at m/z 159;
three points per peak; 1.5 s dwell time for As, 1.5 s dwell time for Tb; number of
repetitions =7.
Nitrate, nitrite and arsenate (As(V)) were analyzed by suppressed conductivity ion
chromatography using a Dionex 500 system (Sunnyvale, CA, USA) fitted with a Dionex
IonPac AS11 analytical column (4 mm x 250 mm) and a AG16 guard column (4 mm x 40
mm). During each injection the eluent (20 mM KOH) was used for 20 min. Other
analytical determinations (e.g., pH, TSS, VSS, etc.) were conducted according to
Standard Methods [APHA, 1999].
201
5.4. Results
5.4.1. Adsorption isotherms of As(III) and As(V) on the AA
Sorptions of both As(III) and As(V) on AA are shown in Figure 1. The constants of
regression equations and parameters of the adsorption isotherm models for As(III) and
As(V) are presented in Table 1. For the adsorption of As in the concentration range of
0.6-60 As(III) or 0.6-80 As(V) mg l-1, the experimental data could only be reliably fit
with Freundlich model yielding R2 values > 0.98. In comparison, the Langmuir fits were
unreliable yielding R2 values of < 0.80 and < 0.53, respectively for As(III) and As(V).
When considering a lower initial concentration range of 0.6 to 7.0 mg l-1, the sorption
both fit well to the Freundlich model. The AA had a significantly higher adsorption
capacity for As(V) compared to As(III) as is illustrated in Figure 1, with lower adsorption
of As(III) compared to As(V) at any given concentration of Ce. The difference in
adsorption is also reflected by a 5-10 fold lower value of KF for As(III) than As(V),
which means the maximum coverage. At the MCL (Maximum Contamination Level) of
As of 10 ug l-1 for drinking water, the sorption of As(III) and As(V) on AA was estimated
form the data to be 0.8 µg As g-1 AA and 8.9 µg As g-1 AA, respectively; based on the
Freundlich isotherms.
202
Table 5.1 Freundlich and Langmuir isotherm model parameters for the adsorption of As(III) and As(V) on activated alumina.
Arsenic
As(III)
†
pH
7.2
Reference
Freundlich isotherms
Langmuir isotherms
Q=KF*Ce (1/n)
Cs=a*b*Ce/(1+b*Ce)
Initial concentration
ranges (mg L-1)
KF [(mg As g-1 AA)(mg As l-1)-1/n]
1/n
R2
a (mg g-1)
b (l mg-1)
R2
0.6-5.0
0.096
1.026
0.993
-0.715
-0.123
0.164
0.6-60.0
0.102
1.097
0.988
-1.755
-0.050
0.531
0.79-4.9
0.052
0.65
0.998
3.48
0.59
0.989
0-0.8 (Ce)†
0.14
0.32
0.97
0.15
8.28
—
0.6-7.0
0.987
1.023
0.985
7.474
0.123
0.003
0.6-80.0
0.537
0.830
0.984
2.211
0.384
0.800
This study
As(III)
6.9
Lin, 2001
As(III)
7.6 Sarvinder, 2004
As(V)
7.2
This study
As(V)
7.2
Lin, 2001
2.85-11.5
0.36
0.27
0.980
9.93
8.25
0.996
As(V)
7.0
Jeong, 2007
0-0.6 (Ce)†
—
—
—
0.14
9.76
0.87
As(V)
7.0 Anderson, 1976
0-120
—
—
—
37.31
2.37
—
As(V)
6.55
0-80 (Ce)†
0.13
0.24
0.99
9.0
32.82
0.97
Jang, 2002
: Equilibrium concentration of adsorption isotherms experiment.
—: Means data not available.
203
1.6
Cs (mg As/g AA)
0.3
Cs (mg As/g AA)
1.2
0.8
0.2
0.1
0.0
0.0
0.2
0.4
0.6
0.8
Ce (mg/l)
0.4
0.0
0.0
2.0
4.0
Ce (mg/l)
6.0
8.0
Figure 5.1 Adsorption isotherms of As(V) (▲) and As(III) (■) on AA. Initial As
concentration range for: A 0.6-60 and 0.6-80.0 mg l-1 for As(III) and As(V), respectively;
Insert: 0.6-5.0 and 0.6-7.0 mg l-1 for As(III) and As(V), respectively. Fitted Freundlich
isotherm adsorption model, As(V) (—) and As(III) (----).
5.4.2. Attenuation of As(III) in AA packed bed columns
A simple experimental model system was utilized in which AA was placed into anaerobic
columns and exposed to continuous feeding of As(III) with basal medium in the presence
or absence of nitrate. During the first 50 day period, the pH of the effluent was high, in
the range of 9.5-10.0 as shown in Figure 2, due to the release of the alkalinity from the
AA. After lowering the concentration of bicarbonate and phosphate, as well as continued
204
washing of the AA, the pH stabilized around 7 after day 50. The cumulative release of
total arsenic from the two AA packed bed columns fed with As(III) with NO3- (S1) and
without NO3- (S2) is illustrated in Figure 3. Within this initial period, the concentration of
total As in the effluent remained as high as 75% and 50% of the influent concentration
for column S1 and S2, respectively; then dropped to lower concentrations when the pH
decreased to 7.2. And there were 65% of arsenic in the influent retained in S1 column,
compared with 86% in S2 column within the period of day 0 to 50.
11
10
pH
9
8
7
6
0
50
100
150
200
250
Time (days)
Figure 5.2 Effluent pH of two AA packed columns fed with a mineral medium
containing 6.67 µM As(III). Column S1 (fed with 2.5 mM nitrate), (●); Column S2
(without nitrate): (○). The vertical line on day 50 indicates the time point after which the
effluent pH stabilized in the circumneutral range, when prior to day 50 the pH ranged
from 9.5 to 10. The vertical arrow indicates the time point when HRT decreased to 0.5 d.
205
700
Total As (ug L-1)
600
500
400
300
200
100
0
0
50
100
150
Time (days)
200
250
Figure 5.3 Removal of total arsenic in two activated alumina packed columns fed with a
mineral medium containing 6.67 µM As(III). Column S1 (fed with 2.5 mM nitrate): (■)
influent, (▲) effluent; Column S2 (without nitrate): (□) influent, () effluent. The vertical
line on day 50 indicates the time point after which the effluent pH fell within the
circumneutral range.
From day 50 onwards, when circumneutral pH conditions prevailed, the results
indicate the release of total As was greater in S2 compared to S1 in accordance with the
expectation that anoxic oxidation would lead to As(V) formation, which in turn would be
adsorbed more efficiently. The total As in the effluent of column S2 started to increase on
day 60, and continued to increase until almost 100% of As(III) in the influent broke
through on day 200. In the same time period from day 50-200, only around 40% the
arsenic entering the column S1 was released to the effluent. On day 250, 70% and 100%
of the As as total arsenic in the influent was recovered in the effluents of columns S1 and
206
S2, respectively, when the operation of two columns was terminated. The result implies
that As had not yet broken through completely by the end of the experiment.
The daily concentration of arsenic species (As(III) and As(V)) in the influent and
effluent of the two columns is illustrated in Figure 4 for S1 and S2. The results illustrate
that As(III) was completely removed in column S1 from day 50 to 100, then the arsenic
started to slowly pass through with the dominant species as As(V) through the end of the
experiment. In column S2, As(III) broke through completely on day 200 with As(III)
occurring as the prevalent species in the effluent. The cumulative release of soluble
identified arsenic species was several folds to many folds greater in S2 compared to S1.
The results indicates that oxidation of As(III) to As(V) in the column SF1 with nitrate
greatly decreased the As mobility.
Figure 5 illustrates the average speciation of arsenic in the influent and effluent of
column S1 and S2 during the period from day 50 to 100 and 202-250. The graph
demonstrates that the arsenic in the influent and the effluent of S2 was predominately
composed of As(III); whereas the effluent of column S1 contained only As(V). There was
incomplete recovery of As(V) in the S1 effluent due to its continued adsorption by AA.
In contrast, the influent and effluent of column S2 were dominated by As(III). At the end
of the experiment, As(III) in the effluent of S2 was similar to the influent As(III) since
the As(III) adsorption capacity of the column was completely exhausted.
207
600
A
-1
500
As(V) Concentration (ug L )
-1
As(III) Concentration (ug L )
600
400
300
200
100
400
300
200
100
0
0
0
50
100
150
200
250
0
600
50
100
150
200
250
100
150
Time (days)
200
250
600
C
500
As(V) Concentration (ug L -1)
As(III) Concentration (ug L -1 )
B
500
400
300
200
100
D
500
400
300
200
100
0
0
0
50
100
150
Time (days)
200
250
0
50
Figure 5.4 Concentrations of As(III) and As(V) in the influent and effluent of column S1 (Panels A and B, respectively) and
column S2 (Panels C and D, respectively) as a function of time: As(III) concentrations: (○) influent; (●) effluent. As(V)
concentrations: influent (∆); effluent (▲).
208
700
A
-1
Arsenic (ug L )
600
500
400
300
200
100
0
700
B
-1
Arsenic (ug L )
600
500
400
300
200
100
0
S1 Infl
S1 Effl
S2 Infl
S2 Effl
Figure 5.5 Speciation of arsenic in the influent and effluent of activated alumina packed
columns supplied with 6.67 µM As(III) and 2.5 mM nitrate (column S1), or only 6.67 µM
As(III) (column S2): Panel A: days 50-100; Panel B: days 202-250. As(III): (solid block);
As(V): (empty block).
209
5.4.3. Residual arsenic in activated alumina and sludge
At the end of the column experiment, the AA was extracted to determine the As adsorbed
on the column packing. The extracted As from the column was compared with the
quantity of As retained in the column as estimated from the differences between the
influent and effluent As (some of As species and total As) to make the As mass balances.
The recovery of arsenic calculated as the ratio of As extracted to As retained for both
columns is shown on Table 2. The adsorbed As concentration profile of S1 and S2 over
the height of the reactors are indicated in Figure 6. There was clearly a greater level of
adsorption at the base of S1 compared to the top of the reactor. In S2, the As sorption was
lower than S1 and there was no obvious profile. The profiles als show that the speciation
of As was consistent over the reactor height, being predominate As(V) in S1 and As(III)
in S2. The results indicate that As(V) adsorption on AA was kinetically limited in S1, but
As(III) adsorption on AA was not kinetically limited in S2. The results are consistent
with a much greater adsorption capacity of AA for As (V) than As(III) as was expected
from the isotherms. The arsenic adsorbed on the AA and the cumulative identified
species mobilized from the column effluents indicate that S1 was predominately
composed of As(V) on both liquid and solid phase confirming the occurrence of
microbial As(III) oxidation by chemolithoautotrophic denitrifiers; whereas, the effluent
of S2 contained only As(III) and with arsenic immobilized in the AA which was also
predominantly in the As(III) form.
210
Table 5.2 Mass balance of arsenic at the end of the experiment.
Parameter
Sum of species of As
As retained in the column
As extracted
Recovery†
influent-effluent (mg)
from column packing (mg)
(%)
S1
69.22
67.05
96.9
S2
48.25
40.54
84.0
S1
65.01
63.65
97.9
S2
50.37
39.42
78.3
Column
Total As
*
†
*
S1: Column fed with nitrate (2.5 mM) and As(III) (6.67 µM). S2: Column fed with As(III) only (6.67 µM).
: The recovery of arsenic calculated as the ratio of As extracted to As retained
211
25-50
0-25
Height (cm)
50-75
A
0
0.1
0.2
0.3
0.4
0.5
0.6
Cs (mg As/g AA)
25-50
0-25
Height (cm)
50-75
B
0
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
Cs (mg As/g AA)
Figure 5.6 The profile of sorbed arsenic in two activated alumina packed columns at the
end of the continuous experiment. Panel A: Column S1 fed with 6.67 µM As(III) and 2.5
mM. Panel B: Column S2 fed with 6.67 µM As(III) but lacking nitrate. Total As: (cross
hatched block); As(V): (empty block); As(III): (solid block).
212
5.5. Discussion
5.5.1. Bioremediation of arsenic by the addition of nitrate
As(III) is more prevalent in reducing environments, due to microbial catalyzed reduction
of As(V) to As(III). The formation of As(III) enhances the release of As sorbed on
aluminum oxides into the aqueous phase [Oremland and Stolz, 2003]. Zobrist et al.
[2000] demonstrated that As(V) reduction by strain Sulforospirillum barnesii can lead to
As mobilization from As(V) co-precipitated with aluminum hydroxide. A previous
studies illustrated that leachates generated from landfills result in the biologically
catalyzed mobilization of arsenic from As(V)-laden drinking water residuals due to
As(V) reduction [Sierra-Alvarez et al., 2005].
Taken together the results indicate, the reduction of As(V) causes mobilization
from aluminum (hydr)oxide solid phases. Thus, the question arise if the oxidation of
As(III) would result in its immobilization on the aluminum phase. In the continuous study
presented here, the addition of NO3- supported the anoxic oxidation of As(III) to As(V)
and its subsequent adsorption on AA. In the control column S2 without NO3- addition,
there was complete break through of As(III) by day 200. On the other hand, complete
break through in column S1 with nitrate never occurred, even after 250 days of operation.
Within the circumneutral period of operation from day 50-200 and 200-250, only 15.8
and 59.6% respectively of all the As fed had passed through column S1, compared with
38.1 and 95.4% in column S2. Therefore, the addition of nitrate played an important role
213
in improving the immobilization as evidenced by the considerably lowered removal of
arsenic in the nitrate-free control column S2. These findings are consistent with the
adsorption behaviors of flow through columns of AA comparing As(III) with As(V),
where an earlier break through of As(III) was witnessed [Clifford, et al., 1990].
5.5.2. Microbial nitrate-dependent oxidation of As(III) to As(V)
The main prerequisite for the enhanced As immobilization in this study is that microbial
nitrate-dependent oxidation of As(III) to As(V). During the whole operations, the As
speciation in the effluent of column S1 was dominated as As(V), which indicated the
occurrence of the As(III) oxidation in the presence of nitrate. The fact that the
denitrifying activity in column S1 played an important role in the As(III) oxidation
removal process was evidenced by the considerably less removal of As(III) in nitrate-free
control column S2. Furthermore, the solid-phase As extraction results showed that AA in
the column S1 contained 1.5-fold higher As than column SF2. Furthermore, the As
speciation of the extract of the solid phase AA at the end of the experiment revealed that
As(V) was the dominant arsenic species retained in the solid phases in the presence of
nitrate in column S1. On the other hand, As(III) was the predominant species in the solid
phase extract of column S2 without nitrate. The mass of As(V) retained on the AA
theoretically close to the As(III) removed in column S1. Microbial anoxic oxidation of
214
As(III) to As(V) by nitrate was the mechanism to enhance the arsenic immobilization on
AA.
Previous studies have illustrated that As(III) oxidizing denitrifying bacteria are
widespread in nature, including arsenic-contaminated lakes [Oremland et al., 2002; Senn
and Hemond, 2002] and soil [Rhine et al., 2006], sluges and sediments unknown
exposure to arsenic [Sun et al., 2008]. Indeed, two novel As(III) oxidizing denitrifying
strains, DAO1 and DAO10, are phylogenetically similar to Azoarcus and Sinorhizobium
on the basis of 16S rRNA sequences, respectively. Furthermore, our previous study
demonstrate that three phylotypes representing Azoarcus, Comamonadaceae and
Diaphorobacter are the main denitrifying As(III) oxidizers suspected to be implicated in
the microbial communities of three enrichment cultures and one mixed cultures studied
[Chapter III].
5.5.3. Adsorption of As(III) and As(V) on AA
Adsorption isotherms of both As(III) and As(V) in this study illustrate that As(V) is more
strongly adsorbed by AA compared to As(III) at circumneutral conditions, which is
consistent with previous findings in literature reports [Lin and Wu, 2001; Goldberg et al.,
2002; Arai et al., 2001]. Previous reports indicate that both Freundlich and Langmuir
could fit well the adsorption of As(III) and As(V) on AA, and the constants of nonlinear
regression equations and parameters of the adsorption isotherms for As(III) and As(V)
215
are summarized in Table S2 from this study and literatures. The nonlinear regression
coefficients indicated that Freundlich isotherm successfully described the partition
behavior between water and AA surface for As(III) and As(V) in this study. The KF
values of As(V) are much higher than those of As(III) in this study and literatures, which
confirm again that AA has a higher affinity for As(V) than As(III).
In the column study from day 0-50, the effluent pH was high in the range 9.5-10.
During the initial period, the sorption of As on AA in column S1 was lower than column
S2, even though the As speciation of S1 was dominated by As(V) compared As(III)
which was prevalent in S2. Previous studies have shown that the adsorption of As(III)
and As(V) on AA is a highly pH-dependent process [Lin and Wu, 2001; Xu et al., 1991].
The As(V) species predominant in aqueous solution as H3AsO4, H2AsO4-, HAsO42- and
AsO43-. The pKa values for those species are 2.3, 6.8 and 11.6. As(III) exists as H3AsO3,
H2AsO3- and HAsO32- in aqueous solution with pKa values of 9.2 and 13.1 [Welch et al.,
2000]. When the effluent pH value was around 9.5-10, the anionic HAsO42- was the main
As(V) species. In contrast, As(III) occurs as mixture of the non-ionic H3AsO3 and the
deprotonated H2AsO3- forms. Since the point of zero charge (pzc) for different types of
alumina is around 8.4-9.1 [Lin and Wu, 2001], the surface of AA would be negatively
charged when the pH is higher than pzc. Therefore the adsorption of the anionic form of
As(V) would be suppressed due to strong repulsion from the negatively charged surface
of AA. In contrast, the less ionic As(III) species would have a better adsorption than
As(V) because of weaker repulsion at high pH.
216
Later, when the pH in the effluent of columns stabilized and maintained in the
circumneutral interval, column S1 demonstrated a much stronger and efficient As
removal than column S2. Under circumneutral conditions, the surface of AA is positively
charge, which is favorable for the sorption of anions due to coulumbic attraction. The
negatively ionic forms As(V) (as H2AsO4- and HAsO42-), would thus have a stronger
interaction (specific binding) with the AA surface and have a higher adsorption. In
contrast, As(III) would solely be present in its non-ionic form (H3AsO3), which only has
weak van der waal force between the solute and the alumina surface. This could explain
why the As(III) adsorption on AA had broken through much earlier in column S2 than
column S1 under neutral pH conditions. This also proves that oxidation of As(III) to
As(V) would promote the As removal from groundwater in circumnuetral environments.
5.5.4. Implications
Biogeochemical processes affecting the adsorption As on soil and sediment minerals
have been a matter of considerable research interest in order to determine the factors
controlling the release of As into groundwater under anoxic environment [Oremland and
Stolz, 2005]. This work used AA as model to determine the impact of microbial
conversion of the As speciation on the mobility of As in the presence of non-iron metal
(hydr)oxides. As(III) is the prevalent As species detected in the anaerobic environments,
which is attributed to the reduction As(V) to As(III) by anaerobic microorganisms. This
conversion increases the mobility of As in soil and sediments containing aluminum
217
oxides with less iron content. To counteract the mobilization due to anaerobic condition,
aeration can be considered. However, raising the dissolved oxygen concentration in the
saturated subsurface water may be impractical, due to its limited solubility and high
chemical reactivity. However, nitrate, which has a higher solubility and lower chemical
reactivity, could be utilized as an alternative electron acceptor to promote the microbial
oxidation of As(III) in the saturated subsurface. This study validates a bioremediation
strategy based on anoxic microbial nitrate-dependent oxidation of As(III) to As(V) which
is more readily adsorbed onto aluminum (hydr)oxides in sediments. The net result would
be an enhanced immobilization of As in anoxic environments.
5.6. Conclusions
(1) AA was demonstrated to have higher adsorption capacity for As(V) than As(III)
under circumneutral pH conditions.
(2) Biological nitrate-dependent oxidation of As(III) to As(V) enhanced the
adsorption of arsenic on activated alumina.
(3) The concentrations of arsenic in the groundwater and subsurface water could be
attenuated by adsorption to aluminum oxides, attributed to anoxic oxidation of
As(III).
218
5.7. Acknowledgments
The work presented here was funded by a USGS, National Institute for Water Resources
104G grant (2005AZ114G), and by a grant of the NIEHS-supported Superfund Basic
Research Program (NIH ES-04940). The use of trade, product, or firm names in this
report is for descriptive purposes only and does not constitute endorsement by the U.S.
Geological Survey. Also arsenic analyses were performed by the Analytical Section of
the Hazard Identification Core (Superfund Basic Research Program grant NIEHS-04940).
219
5.8. References
(1)
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(2)
Arai, Y., Elzinga, E.J. and Sparks, D.L. (2001) X-ray adsorption spectroscopic
investigation of arsenite and arsenate adsorption at the aluminum oxide-water
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(3)
ATSDR (2007) Toxicological profile for arsenic, p. 500, Agency for Toxic
Substances and Disease Registry, U.S. Department of Health and Human
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(4)
Bissen, M. and Frimmel, F.H. Arsenic – a review. Part II: oxidation of arsenic
and its removal in water treatment. Acta hydrochim. hydrobiol. 31(2), 97–107.
(5)
Clifford, D.A. (1990) Ion exchange and inorganic adsorption. In Water Quality
and Treatment, 4th edn., ed F. W. Pontius. American Water Works Association,
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(6)
Cullen, W.R. and Reimer, K.J. (1989) Arsenic speciation in the environment.
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Dixit, S. and Hering, J.G. (2003) Comparison of arsenic(V) and arsenic(III)
sorption onto iron oxide minerals: Implications for arsenic mobility. Environ.
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Fisher, J.C. and Hollibaugh, J.T. (2008) Selenate-dependent anaerobic arsenite
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(9)
Frost, R.R. and GRIFFIN, R.A. (1977) Effect of pH on adsorption of arsenic
and selenium from landfill leachate by clay-minerals. Soil Science Society of
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(10)
Jiang, J.Q. (2001) Removing arsenic from groundwater for the developing
world – a review. Water Sci. Technol. 44, 89–98.
(11)
Hering, J.G. and Dixit, S. (2005) Contrasting sorption behavior of aresnic(III)
and arsenic(V) in suspensions of iron and aluminum oxyhydroxides. In: O'Day,
P.A., Vlassopoulos, D., Meng, X. and. Benning, L. (Eds.). Advances in Arsenic
Research. ACS Symposium Series No. 915; American Chemical Society,
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(12)
Inskeep, W.P., Macur, R.E., Hamamura, N., Warelow, T.P., Ward, S.A. and
Santini, J.M. (2007) Detection, diversity and expression of aerobic bacterial
arsenite oxidase genes. Environ. Microbiol. 9(4), 934-943.
(13)
Ghosh, M.M. and Yuan, J.R. (1987) Adsorption of inorganic arsenic and
organoarsenicals on hydrous oxides. Environmental Progress 6(3), 150-157.
(14)
Gimenez, J., Martinez, M., Pablo, J. D., Rovira, M. and Duro L. (2007) Arsenic
sorption on natural hematite, magnetite and goethite. J. Hazard. Mat. 141, 575580.
(15)
Goldberg, S. (2002). Competitive adsorption of arsenate and arsenite on oxides
and clay minerals. SOIL SCI. SOC. AM. J., 66(2), 413-421.
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Lin, Z. and Puls, R.W. (1999) Adsorption, desorption and oxidation of arsenic
affected by clay minerals and aging process. Environment Geology 39(7), 753759.
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(17)
Lin, T.F. and Wu, J.K. (2001) Adsorption of arsenite and arsenate within
activated alumina grains: Equilibrium and kinetics. Water Res. 35(8), 20492057.
(18)
Nolan, B.T., Ruddy, B.C., Hitt, K.J. and Helsel, D.R. (1997) Risk of nitrate in
groundwaters of the United States - A national perspective. Environ. Sci.
Technol. 31(8), 2229-2236.
(19)
Oremland, R.S., Hoeft, S.E., Santini, J.A., Bano, N., Hollibaugh, R.A. and
Hollibaugh, J.T. (2002) Anaerobic oxidation of arsenite in Mono Lake water
and by facultative, arseniteoxidizing chemoautotroph, strain MLHE-1. Appl.
Environ. Microbiol. 68(10), 4795-4802.
(20)
Oremland, R.S. and Stolz, J.F. (2003) The ecology of arsenic. Science
300(5621), 939-944.
(21)
Oremland, R.S. and Stolz, J.F. (2005) Arsenic, microbes and contaminated
aquifers. Trends Microbiol. 13, 45-49.
(22)
Rhine, E.D., Phelps, C.D. and Young, L.Y. (2006) Anaerobic arsenite oxidation
by novel denitrifying isolates. Environ. Microbiol. 8(5), 899-908.
(23)
Senn, D.B. and Hemond, H.F. (2002) Nitrate controls on iron and arsenic in an
urban lake. Science 296(5577), 2373-2376.
(24)
Sierra-Alvarez, R., Field, J.A., Cortinas, I., Feijoo, G., Moreira, M.T., Kopplin,
M. and Gandolfi, A.J. (2005) Anaerobic microbial mobilization and
biotransformation of arsenate adsorbed onto activated alumina. Water Res.
39(1), 199-209.
(25)
Smedley, P.L. and Kinniburgh, D.G. (2002) A review of the source, behaviour
and distribution of arsenic in natural waters. Appl. Geochem. 17(5), 517-568.
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(26)
Sun, W., Sierra-Alvarez, R., and Field, J.A. (2008) Anoxic oxidation of arsenite
linked to denitrification in sludges and sediments. Water Research 42, 45694577.
(27)
Sun, W., Sierra-Alvarez, R., Fernandez, N., Sanz, J.L., Amils, R., Legatzki, A.,
Maier, R.M. and Field, J.A. (2008) Molecular Characterization and In Situ
Quantification of Anoxic Arsenite Oxidizing Denitrifying Enrichment Cultures.
FEMS Microbiology Ecology. (under review)
(28)
USEPA (2001) National Primary Drinking Water Regulations: Arsenic and
Clarifications to Compliance and New Source Contaminants Monitoring.
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(29)
Wang, S. and Mulligan, C.N. (2006) Natural attenuation processes for
remediation of arsenic contaminated soils and groundwater. Journal of
Hazardous Materials B138, 459–470.
(30)
Welch, A., Ryker, S., Helsel, D., Hamilton, P. 2001. Arsenic in groundwater of
the United States: An overview. Water Well J.2, 30–33.
(31)
Xu, H., Allard, B. and Grimvall, A. (1991) Effects of acidification and natural
organic materials on the mobility of arsenic in the environment. Water, Air and
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(32)
Zobrist, J., Dowdle, P.R., Davis, J.A. and Oremland, R.S. (2000) Mobilization
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223
CHAPTER 6
ANOXIC OXIDATION OF As(III) AND Fe(II) LINKED TO CHEMOLITHOTROPHIC
DENITRIFICATION FOR THE IMMOBILIZATION OF As IN ANOXIC
ENVIRONMENTS
6.1. Abstract
Microbial reduction of Arsenate (As(V)) and arsenic-bearing iron(hydr)oxides are
suggested to be the dominant mechanisms of arsenic (As) mobilization in subsurface
environments. Oxygen can be introduced into the anaerobic zone by injection of gaseous
O2 to prevent or reverse the reactions, but O2 is poorly soluble and reactive and thus
difficult to distribute in the subsurface. Nitrate can be considered as an alternative oxidant
having the advantage of being more soluble, less reactive and thus more readily dispersed
in the saturated subsurface, including groundwater and sediments. In this study, a
bioremediation strategy was explored by injecting NO3- to support the anoxic oxidation
of Fe(II) and As(III) in the subsurface as a means to immobilize As in the form of As(V)
adsorbed on biogenic iron(III) hydroxides. Continuous flow sand-packed columns were
used to simulate a natural anaerobic groundwater and sediment system with co-occurring
As(III) and Fe(II) in the presence or absence of nitrate. During the long term operation of
250 days, the removal of Fe(II) and As(III) in column SF1 in the presence of nitrate was
10-fold higher than column SF2 lacking nitrate. Solid-phase extraction also revealed that
the amount of iron and As retained on the sand surface of column SF1 was much higher
224
than those on column SF2. The recovered solid-phase iron and As was close to the iron
and As removed during columns operation. The dominant speciation of iron and As was
Fe(III) and As(V) in column SF1, compared with Fe(II) and As(III) in column SF2.
Furthermore, mineralogy analysis by XRD and XPS suggests that the iron(III) oxide and
As(V) content in the active biological column amended with nitrate, provides
confirmation that anoxic oxidation of Fe(II) and As(III) can be coupled to denitrification.
Microbial oxidation of As(III) and Fe(II) linked to denitrification resulted in the
enhancement of immobilization of As in anaerobic environments.
Keywords: Arsenite, Ferrous iron, Nitrate, Anaerobic, Iron (hydr)oxides, Adsorption
225
6.2. Introduction
Although arsenic has a relatively low abundance in the earth’s crust, it is generally found
as a contaminant in soil and water systems due to various anthropogenic activities, such
as mining, discharge of industrial waste and agriculture, as well as from natural
biogeochemical reactions [Oremland and Stolz, 2003; Sierra-Alverez et al., 2006].
Arsenic is a known human carcinogen [WHO, 1993; ATSDR, 2007], and its
contamination of drinking water sources is presently a worldwide concern [Smedley and
Kinniburgh, 2002].
The predominant oxidation states of As found in surface water and groundwater
are arsenite (As(III), H3AsO3) and arsenate (As(V), H2AsO4- and HAsO42-). In natural
soil and sediments, iron (hydr)oxides have strong sorption for both As(III) and As(V) in
circumneutral environments [Raven et al., 1998; Dixit and Hering, 2003]. However,
As(V) generally binds more strongly to clays and aluminum hydroxides than As(III) [Lin
and Wu, 2001; Aria et al., 2001]. In anaerobic environments, microorganisms play an
important role in the mobilization of adsorbed arsenic [Oremland et al., 2005; Stolz et al,
2006]. Dissimilatory reductive dissolution of iron oxyhydroxides can lead to release of
both adsorbed As(III) and As(V) into the aqueous phase in anaerobic environments
[Herbel and Fendorf, 2002; Oremland and Stolz, 2005]. The microbial reduction of
As(V) also increases the mobility of As in environments due to the lower sorption
strength of non-iron metal oxides such as aluminum (hydr)oxide for As(III) [Zobrist et
al., 2000; Cummings et al., 1999]. Thus, in As contaminated sites, normal microbial
226
reduction of As(V) and Fe(III) could enhance the mobility of arsenic under anaerobic
conditions, posing a threat of arsenic contamination in drinking water [Anawar et al.,
2006; Smedley and Kinniburgh, 2002].
One strategy to suppress As reduction is to promote microbial re-oxidation of
As(III) and Fe(II). The oxidation of Fe(II) could result in the formation of iron
hydroxides that strongly adsorb both As(III) and As(V) and remove them from the
aqueous phase [Dixt and Hering, 2003]. Also the formation of As(V) is well-known to
improve their binding to clay and minerals containing mostly Al oxides [Lin and Wu,
2001; Frost and Griffin, 1977; Goldberg, 2002].
Fe(II) is subject to both spontaneous chemical oxidation [Appelo et al., 1999;
Lytle et al., 2004; Park and Dempsey, 2005] and microbial catalyzed oxidation [Emerson,
2000; Sobolev and Roden et al., 2001; James and Ferris, 2004] in the presence of
dissolved O2 (DO) to form iron(III) hydroxides at circumneutral conditions. Additionally,
it has been well established that As(III) is oxidized to As(V) by a large diversity of
microorganisms present in arsenic-impacted environments in aerobic conditions [Rhine et
al., 2006; Inskeep et al., 2007]. In fact, oxidation of Fe(II) and As(III) by aeration has
been utilized to remediate arsenic contaminated groundwater [Katsoylannis and
Zouboulis, 2005; Battaglia-Brunet et al., 2006]. However, the oxygen has limited
aqueous solubility and may be significantly consumed by organic matter, sulfides and
other reducing compounds in the groundwater. These limitations suggest exploration of
other routes for oxidation of As(III) and Fe(II) e.g., the use of alternative oxidants.
227
Recently, evidence indicates that biological nitrate-dependent As(III) oxidation
[Oremland et al., 2002; Rhine et al., 2006; Sun et al., 2008] as well as Fe(II) oxidation
[Straub et al., 2001; Weber et al., 2006] is catalyzed by several anoxic microorganisms in
the absence of O2. Anoxic biological oxidation of As(III) by denitrifying microorganisms
were shown to occur by a three strains of bacteria isolate in lakes and sediments with
historical arsenic contamination or naturally high levels [Oremland et al., 2002; Rhine et
al., 2006], and sludges and sediments with no prior exposure to arsenic [Sun et al., 2008].
The nitrate-dependent oxidation of As(III) was also demonstrated to be feasible and
stable for up to 3.5 years operation in the continuous bioreactor system with bacteria
immobilized in biofilms [Sun et al., 2008, Chapter IV]. Microbial oxidation of both
soluble and insoluble Fe(II) coupled to nitrate reduction has been demonstrated in various
freshwater and saline environmental systems at neutral pH [Straub et al., 2004; Weber et
al., 2001; Weber et al., 2006]. The biological oxidation of Fe(II) presents mechanism for
the formation of insoluble Fe(III) oxide minerals in anoxic soils and sediments, such as
ferrihydrite and other forms of iron oxides, which have been used as best available
technologies to remove As from drinking water [USEPA, 2001; Bissen and Frimmel,
2003].
The objectives of this study are to evaluate the addition of NO3- to support the
anoxic oxidation of Fe(II) and As(III) as a means to immobilize arsenic in the form of
As(V) adsorbed on biogenic iron(III) hydroxides. This study was performed in a
continuous flow column packed with sand. The goal of the study was to demonstrate the
potential of utilizing nitrate injection as a long term bioremediation approach for
228
removing arsenic from anoxic natural water system contaminated with co-occurring
Fe(II) and As(III). The specific goals were to 1) evaluate the sorption of As(III) and
As(V) on nitrate-dependent biogenic iron(III) oxide coated sands (ICS) at neutral pH; 2)
investigate the potential of biological anoxic oxidation of As(III) and Fe(II) linked to
nitrate reduction, and followed by As(V) adsorption on freshly formed ICS to remediate
groundwater contaminated with As; 3) characterize the solid-phase nitrate-dependent iron
mineral associated with sorbed As with respect to its chemical and mineralogical
composition.
6.3. Materials and methods
6.3.1. Microorganisms
The chemolithotrophic As(III)-oxidizing denitrifying granular sludge used to inoculate
the two sediment columns was obtained from a bench-scale As(III)-oxidizing denitrifying
bioreactor (2 l) after 828 days of operation. The original inoculum of the bioreactor was
methanogenic granular sludge obtained from full-scale upward flow anaerobic sludge
blanked (UASB) treating alcohol distillery wastewater (NGS) (Nedalco, Bergem op
Zoom, The Netherlands). The bioreactor was fed with As(III) (4.04 ± 0.20 mM) as
electron donor, NO3– (6.29 ± 0.55 mM) as electron acceptor, and NaHCO3 (8.0 mM) as
major carbon source together with 0.3 mg l-1 yeast extract. The hydraulic retention time
229
(HRT) was 1.12 ± 0.22 days. The total suspended solids’ (TSS) contents and volatile
suspended solids’ (VSS) content of the sludge recovered from the bioreactor were 6.08 ±
0.18% and 5.76 ± 0.23% wet weight basis, respectively. The granular sludge samples
were washed and sieved before use to remove fines. The sludge was stored under
nitrogen at 4ºC.
6.3.2. Basal medium
The standard basal medium was prepared using ultra pure water (Milli-Q system;
Millipore) and contained the following compounds (mg l-1): NH4HCO3 (3.16); NaHCO3
(100); CaCl2 (10), MgSO4.7H2O (10); K2HPO4 (204); KH2PO4.2H2O (204); and 0.02 ml
l-1 of a trace element solution containing (in mg l-1): FeC13.4 H20 (2,000); CoCl2. 6 H20
(2,000); MnCl2 4 H20 (500); AlCl3 6 H20 (90); CuCl2.2H20 (30); ZnCl2 (50); H3BO3 (50);
(NH4)6Mo7O24.4 H2O (50); Na2SeO3.5 H2O (100); NiCl2.6 H20 (50); EDTA (1,000);
resazurin (200); HCl 36% (1 ml). The pH of the influent was adjusted to 7.2 with NaOH
or HCl, as required.
230
6.3.3. Formation of biogenic iron(III) oxides coated sand (ICS) linked to denitrification
The experiment was conducted in triplicate using glass flasks (1,100 mL) supplied with
sand as supporting material (70 g dry weight) and 1,000 ml of pH-7.2 basal medium
amended with 15.0 mM Fe(II) supplied as FeCl2 and 10.0 mM NO3– supplied as KNO3–.
The bottles were inoculated 10% (w/v) with NGS. Flasks were sealed with butyl rubber
stoppers, and then the medium and headspace were purged with N2/CO2 (80/20, v/v) for
20 min to exclude oxygen from the assay, then were shaken under a N2/CO2 (80/20)
atmosphere in an orbital shaker (150 rpm) at 30 ± 2°C. The iron species (Fe(II) and
Fe(III)) were monitored every two days. After all the Fe(II) was oxidized to Fe(III), the
pH of the solution was adjusted to pH-12.0 with NaOH under anaerobic conditions,
which was intened to enhance the Fe(III) precipitation and adsorption on the surface of
the sand to generate ICS [Chang et al., 2006]. The sand slurry was maintained in water
bath at 80°C. Water vapors were continuously removed by flushing under N2 until
approximately 10% of water was remained in the suspension. After then the sand was
dried under N2 overnight. To remove the salts on the sand, dry sand was rinsed with
distilled three times and then dried again under N2. The dry weight content of final ICS
product was 93.63 ± 0.58%. The iron content of ICS on a dry weight basis was 11.9 ± 1.3
mg Fe g-1 dwt ICS with an average of 99.7 ± 0.2% of the iron as Fe(III).
231
6.3.4. Sorption of As(III) and As(V) on biogenic iron oxide coated sands (ICS)
Sorption experiments with biogenic ICS were conducted in triplicate using glass flasks
(160 ml) supplemented with the sorbent (2.05 g dry weight) and 100 ml of a pH-7.2
sodium bicarbonate buffer (33.3 mM) spiked with As(III) (as AsNaO2) or As(V) (as
Na2HAsO4·7 H2O). The initial adsorbate concentrations were 0.6, 1.0, 2.0, 3.0, 4.0, 5.0,
7.0, 9.0, 15.0, 20.0, 35.0, 45.0, and 60.0 mg l-1 for As(III). The concentration tested for
As(V) were the same except that an As concentration of 80.0 mg l-1 was used instead of
60.0 mg l-1. Control flasks lacking sorbent were run in parallel to correct for possible
removal of arsenic by other mechanisms than adsorption. Flasks were shaken under a
N2/CO2 (80/20) atmosphere in an orbital shaker (150 rpm) for 2 d at 30 ± 2°C. Removal
of As(III) and As(V) from solution was determined following centrifugation and filtration
(0.45 µm) of the samples to remove insoluble matter.
The experimental data were fitted to the Langmuir and Freundlich isotherm
models. The Langmuir isotherm is defined by:
Cs =
a ⋅ b ⋅ Ce
1+ b ⋅ Ce
[eq. 1]
Where Cs is the concentration of the solute in the solid phase (mg As g-1 Fe), Ce is the
equilibrium concentration of the solute in solution (mg As l-1), a and b are Langmuir
232
adsorption constants; a represents the maximum achievable surface concentration of the
solute, and b is the equilibrium constant for the sorption reaction.
The Freundlich equation is an empirical relationship describing the sorption of
solutes from a liquid to a solid surface. The Freundlich equation is defined by:
C s = K F ⋅ Ce
1/ n
[eq. 2]
Where KF [(mg As g-1 Fe)(mg As l-1)-1/n] is the Freundlich adsorption constant or capacity
factor and n is the Freundlich exponent which provides a measure for the sorption
intensity. The KF value represents the loading in mg adsorbate per gram of sorbent at an
aqueous equilibrium concentration of 1.0 mg l-1 of the compound.
6.3.5. Attenuation of As(III) on freshly biogenic iron(III) oxide under chemolithotrophic
denitrifying conditions in sediment columns packed with sand
Anoxic As(III) and Fe(II) oxidation under denitrifying conditions was investigated in two
glass columns (each 420 ml) continuously fed with synthetic basal medium. The columns
were placed in a climate controlled room at 30 ± 2°C and covered with aluminum foil to
avoid growth of phototrophic microorganisms. Each reactor was packed with 600 dry
weight of white quartz sand (SiO2; -50+70 mesh) and inoculated with 2.94 g VSS l-1
anaerobic sludge obtained from the a laboratory-scale As(III)-oxidizing denitrifying
233
bioreactor. The treatment reactor (SF1) was biologically active column inoculated with
chemolithotrophic As(III)-oxidizing denitrifying bacteria and fed with basal medium,
As(III) (0.5 mg l-1) and Fe(II) (20 mg l-1 Fe supplied as FeCl2) as electron donating
substrates, nitrate (35 mg l-1 NO3- supplied as KNO3-) and bicarbonate as the major
carbon source, except that 18.8 mg l-1 acetate was added to consume traces of DO that
could potentially enter from medium and cause oxidation of As(III). The control reactor
(SF2) was the same as SF1, but lacked nitrate in the medium. The influent of both
reactors was maintained at all times under a N2 atmosphere supplied via a gas bag (SKCWest Inc, Fullerton, CA) to minimize any exposure to O2.
The columns were operated with hydraulic retention times averaging 24 h for the
whole experiment. Fresh liquid samples were collected periodically from the influent and
effluent lines and prepared immediately for analysis to minimize possible changes in
arsenic speciation upon exposure to the atmosphere. The pH value was determined
immediately after sampling.
6.3.6. Batch assay
The biological activity of Fe(II) oxidation linked to denitrification in the effluent of SF1
column was monitored by measuring the conversion of iron speciation in batch assay.
Batch bioassay was performed in shaken flasks, which were incubated in a dark climate-
234
controlled room at 30 ± 2˚C. Serum flasks (160 ml) were supplied with 120 ml of a basal
mineral medium (pH 7.0-7.2): NH4Cl (300); NaHCO3 (2150); CaCl2 (10), MgCl2 (83),
MgSO4.7H2O (10); K2HPO4 (320); KH2PO4.2H2O (850); and 0.02 ml l-1 of a trace
element solution as described before. The bioassay was also supplemented with Fe(II)
(2.0 mM) as electron donor, nitrate (5 mM) as the electron acceptor and inoculated with
effluent of SF1 column (5% by volumn). Flasks for anoxic assays were sealed with butyl
rubber stoppers, and then the medium and headspace were purged with N2/CO2 (80/20,
v/v) for 20 min to exclude oxygen from the assay. Various controls (e.g. abiotic controls,
controls without nitrate) were applied based on the requirements of each experiment.
Abiotic controls were prepared by excluding the addition of inoculum. Controls without
NO3- were used to estimate if there were other factors contributing to As(III) oxidation.
All assays were conducted in triplicate. Samples were analyzed for the concentration of
Fe(II) and Fe(III).
6.3.7. Iron and arsenic extraction from ICS
Dry samples collected for analysis of Fe(II) and total Fe on ICS were extracted in 6.0 N
HCl under a N2 atmosphere in 30-mL anaerobic glass tubes with butyl rubber stoppers.
All tubes were incubated at 30±2°C in a shaking water bath (100 rpm) for overnight. The
235
Fe(II) and total Fe in extracts were analyzed using phenanthroline method as described
later.
To avoid the artificial interference of Cl-Cl on the As analysis, dry samples
collected for analysis of As speciation and total As on ICS were extracted in 6.0 N HNO3
under a N2 atmosphere in 30-mL anaerobic glass tubes with butyl rubber stoppers. All
tubes were incubated at 30±2°C in a shaking water bath (100 rpm) for overnight. The
tubes were allowed to settle, samples of the supernatants were collected and they were
adjusted immediately to pH 6.0-6.5 with NaOH (8.0 N). Subsequently, the samples were
membrane filtered (0.45 µm) and stored in polypropylene vials at -20°C till analysis of
total As content and As speciation. The total As, As(III) and As(V) in extracts were
quantified using ICP-MS and HPLC-ICP-MS as described later. A simple assay was set
up to confirm the preservation of As speciation during the extraction process. The results
shown (Table S1) demonstrated that the HNO3 was the best mineral acid to conduct
solid-phase As extraction from ICS, because it extracted most efficiently, and also did not
change speciation.
6.3.8. Mineral analysis
In preparation for X-ray diffraction (XRD) analysis, solid samples were dried under N2
overnight in anaerobic tube. The mineralogy of the pristine original sand, sand from
236
control column SF2 and ICS from biological column SF1 with As adsorption was
characterized by X-ray diffraction (Scintag XRD 2000 PTS, X-ray Diffractometer,
University of Spectroscopy and Imaging Facility, University of Arizona, Tucson, AZ). Xray diffraction spectroscopy (XRD) was used to determine the chemical compositions of
the coated metallic species. The operational parameters were as followed: copper Kα; λ=
1.5444 A; voltage= 30 kV; current intensity= 20 mA. The samples were scanned from
10º to 70º 2θ and the scanning rate was fixed at 2.0º min-1.
6.3.9. Scanning electron microscope/ Energy dispersive spectroscopy (SEM/EDS)
analysis
Samples for field emission SEM/EDS (Hitachi-S-4800 Type II scanning electron
microscope with Thermo-NORAN NSS energy dispersive spectroscopy, University of
Spectroscopy and Imaging Facility, University of Arizona, Tucson, AZ) were used to
observe the surface properties of the coated layer. The samples were coated with thin
platinum film in order to avoid the influence of charge effect during the operation of
SEM. Elemental micro-probe and elemental distribution mapping techniques were used
for analyzing the elemental constitution of solid samples. The operational conditions of
EDS analysis were scanning energy from 0 to 15 keV with an elapse time of 100 s.
237
6.3.10. X-ray Photoelectron Spectroscopy (XPS) analysis
Surface characterization was carried out with monochromatic X-ray photoelectron
spectroscopy (XPS) to examine the oxidation states and chemical structures of the iron
and As on the sand in both biological (SF1) and control columns (SF2). XPS studies were
conducted with a Kratos 165 Ultra X-ray photoelectron spectrometer (Laboratory for
Electron Spectroscopy and Surface Analysis, University of Arizona, Tucson, AZ)
equipped with a high sensitivity Al monochromatic Kα source at 1486.6 eV. For all the
data presented here, the analyzed spot size was 300 × 700 microns, with analyzer pass
energy of 20 eV.
6.3.11. Analytical methods
As(III) and As(V) speciation was analyzed by ion chromatography–inductively coupled
plasma–mass spectroscopy (HPLC-ICP-MS). The system consists of an HPLC (Agilent
1100) and an ICP-MS (Agilent 7500a) with a Babington nebulizer as the detector. The
operating parameters were as follows: Rf power 1500 watts, plasma gas flow 15 l min-1,
carrier flow 1.2 l min-1, As was measured at 75 m/z and terbium (IS) measured at m/z
159. The injection volume was 10 µl. The detection limit for the various As species was
238
0.1 µg l-1. The total concentration of As in liquid samples was determined by direct
injection into ICP-MS using an ASX500 auto sampler (CETAC Technologies, Omaha,
NE). The analytical system was operated at an Rf power of 1,500 W, a plasma gas flow
of 151 min-1 and a carrier gas flow of 1.21 min-1. The acquisition parameters used were as
follows: As measured at m/z 75; terbium (IS) measured at m/z 159; three points per peak;
1.5 s dwell time for As, 1.5 s dwell time for Tb; number of repetitions =7.
Nitrate, nitrite and arsenate (As(V)) were analyzed by suppressed conductivity ion
chromatography using a Dionex 500 system (Sunnyvale, CA, USA) fitted with a Dionex
IonPac AS11 analytical column (4 mm x 250 mm) a AG16 guard column (4 mm x 40
mm). During each injection the eluent (20 mM KOH) was used for 20 min.
Fe(II) was quantified by the colorimetric method using 5-ortho-phenantroline,
using an UV-visible spectrophotometer (Agilent 8453, Palo Alto, CA, USA). Total Fe
was obtained by reducing Fe(III) with a 1% hydroquinone solution (in pH 4.5 0.05M
acetate buffer) [APHA, 1999]. Other analytical determinations (e.g., pH, TSS, VSS, etc.)
were conducted according to Standard Methods [APHA, 1999].
239
6.4. Results
6.4.1. Adsorption isotherms of As(III) and As(V) on the ICS
Adsorption isotherms of both As(III) and As(V) on ICS are shown in Figure 1,
illustrating that As(III) is less strongly adsorbed by nitrate-dependent ICS compared to
As(V). The regression equations and parameters of the adsorption isotherm models for
As(III) and As(V) are presented in Table 1. For the adsorption of As(III) at the
concentration range of 0.6 to 60 mg l-1, As(III) experimental data was only fitted
satisfactorily by Freundlich isotherms (R2 > 0.88); in comparison, As(V) ranging from
0.6 to 80 mg l-1 was well described by Langmuir (R2 > 0.91) and Freundlich isotherm (R2
> 0.88).
The results in Figure 1 demonstrated that lower adsorption of As(III) than As(V) at
any given concentration of Ce. Compared with As(III), the value KF meaning the
maximum coverage was 18-784 fold higher for As(V) adsorbed by ICS, especially at
concentration of lower than 5.0 mg l-1 in Freundlich adsorption model. The value of (1/n)
was always in the range of 0 to 1 for both As(III) and As(V), suggesting favorable
adsorption conditions. The value of a, representing the theoretical monolayer adsorption
capacity on ICS in Langmuir equation, was 2-4 time higher for the As(V) than As(III).
For the adsorption energy on ICS indicated by the equilibrium constant b, the As(V) also
showed stronger binding than As(III). The maximum adsorptions of As(III) and As(V)
240
onto the ICS (qmax, denoted as mg g−1) was estimated from measurements as 212.7 and
77.5 mg As g−1 ICS, respectively. At the MCL level of 10 ug l-1 for As in drinking water,
the adsorption of As(III) and As(V) on AA was estimated form the data to be 0.5 and
34.4 mg As g-1 ICS, respectively.
200
Cs (mg As/g Fe(III))
160
120
80
40
0
0.0
1.0
2.0
3.0
4.0
5.0
6.0
Ce (mg/l)
Figure 6.1 Adsorption isotherms of As(V) (♦) and As(III) (■) on biogenic ICS.
Equilibrium As concentration (Ce) range for 0-6.0 mg l-1 for As(III) and As(V).
Simulated isotherm adsorption models, Freundlich (—) and Langmuir (----).
241
242
6.4.2. Microbial nitrate-dependent oxidation of Fe(II) and subsequent precipitation of
Fe(III) oxides formed on sand surface to form ICS in sediment columns
The cumulative release of soluble species of iron (Fe(II) and Fe(III)) is illustrated in
Figure 2 for both columns SF1 and SF2. The results show the removal of iron was 10fold greater in column SF1 compared to SF2, which is in accordance with the expectation
that Fe(II) was oxidized and retained as iron (hydr)oxides in column SF1 in the presence
of nitrate. As the column experiment progressed the first 20-30 days, the removal of iron
improved and achieved equilibration in column SF1. Initially during the first 2 weeks, the
removal was 50.2 (± 23.1)% after day 12 the removal average 91.33 (± 6.91)%. In the
column lacking nitrate (SF2), the removal of iron averaged only 14.14 (± 8.30)% from
day 2 onwards.
243
25
-1
Fe concentration (mg L )
A
20
15
10
5
0
0
40
80
120
160
200
240
80
120
Time (days)
160
200
240
25
Fe concentration (mg L-1)
B
20
15
10
5
0
0
40
Figure 6.2 Concentrations of Fe(II) and Fe(III) in the influent and effluent of biological
column SF1 (Panel A) and control column SF2 (Panel B) as a function of time: Fe(II)
concentrations: (▲) influent; (∆) effluent. Fe(III) concentrations: influent (●); effluent
(○).
244
Figure 3 illustrates the average speciation of iron in the influent and effluent of
reactors SF1 and SF2 during steady state period of operation after day 12. In SF1, the
soluble Fe(III) concentration is low in both influent and effluent, thus the Fe(III) formed
can most likely be rationalized by accumulation of iron (hydr)oxides inside the column;
However in column SF2, the influent and effluent were dominated by Fe(II), and with
more than 94.39 (± 4.43)% of Fe(II) entering SF2 passing through the column. The
results shown in Figure 4 demonstrated that the effluent of column SF1 have the
biological capacity of linking Fe(II) oxidation to denitrification. There was no obvious
oxidation in the abiotic control. 82% of Fe(II) was removed in the treatment with nitrate
after 10 days, compared with negligible Fe(II) elimination in control treatment lacking
nitrate. The formation of soluble Fe(III) did not match the removal of soluble of Fe(II)
due to the precipitation of Fe(III) from the liquid phase.
Iron concentration (mg L-1)
25
20
15
10
5
0
SF1 Influent
SF1 Effluent
SF2 Influent
SF2 Effluent
Figure 6.3 Iron speciation in the influent and effluent of sand packed columns supplied
with 6.67 µM As(III), 36 µM Fe(II) and 2.5 mM nitrate (column SFF1), or only 6.67 µM
As(III) and 36 µM Fe(II) (column SF2): Fe(II) (solid block) and Fe(III) (empty block).
245
150
Fe(II) concentration (mgl-1)
A
120
90
60
30
0
0
5
10
15
20
25
B
-1
Fe(III) concentration (mg l )
150
120
90
60
30
0
0
5
10
15
20
25
Time (days)
Figure 6.4 The removal of soluble Fe(II) and the formation of soluble Fe(III) by effluent
of column SF1: Fe(II) (panel A) and Fe(III) (panel B). Abiotic control with Fe(II) and
NO3- (○), effluent with Fe(II) and NO3- (■) and effluent with Fe(II) without NO3- (▲).
246
6.4.3. Microbial nitrate-dependent oxidation of As(III) and subsequent adsorption of
As(V) on ICS in sediment columns
The cumulative release of total As from the two columns is illustrated in Figure 5. The
result shows the release of As (measured as total As) was greater in SF2 compared to SF1
in accordance with the expected adsorption of As on the iron (hydr)oxides formed from
Fe(II) oxidation linked to nitrate reduction. During the first 30 days of operation, the total
As concentration passing through column SF1 started to gradually decrease from day 16
to 30. After 30 days, the iron (hydr)oxides were stably retained in the column with a
steady state with a concentration of total As averaging at 10.60 (±9.58) ug l-1, which
corresponded 47% removal efficiency. The results are consistent with adsorption of As
on freshly precipitated iron (hydr)oxides on sand surface. When the experiment was
terminated, 97.2 (±3.3)% of the initial As(III) added to the columns was adsorbed inside
the columns SF1 during the operation from day 50 to 225, whereas, less than 10.2
(±6.1)% the As entering the column SF2 passed through the column SF2 and was
adsorbed on the sand surface.
247
-1
Total Arsenic concentration (ug L )
800
600
400
200
0
0
40
80
120
Time (days)
160
200
240
Figure 6.5 Removal of total As in two sand packed columns fed with a mineral medium
containing 6.67 µM As(III). Column SF1 (fed with 36 µM Fe(II) and 2.5 mM nitrate):
(▲) influent, (∆) effluent; Column SF2 (fed with 36 µM Fe(II) without nitrate): (●)
influent, (○) effluent.
The daily concentration of soluble species of As (As(III) and As(V)) released
from the two columns is illustrated in Figure 6 for SF1 and SF2, respectively. The results
illustrate that As(III) was 99.7% eliminated from the column SF1 during the period of 50225 days, and it was not recovered as soluble As(V) in the effluent. Therefore As was
most likely adsorbed inside the column. The results shown in Figure 7, As(V) was the
dominant species of the small fraction of As that was discharged in the effluent, which
account for 89.5% of As speciation in effluent. These suggests that As(III) oxidation to
As(V) may have been taken place in column SF1 fed with nitrate. However, As(III)
removal was marginally 9.7% in the absence of nitrate in column SF2. The results
248
suggest that no adsorption had occurred, because there was consistent with low content of
retention of iron in the column. And it is also demonstrated (Figure 6) that As(III) was
the prevalent form of As discharge with SF2 effluent (99.8% of As speciation). These
findings indicate that little oxidation of As(III) had occurred.
249
800
As concentration (ug L -1)
A
600
400
200
0
0
40
80
120
160
200
240
800
As concentration (ug L -1)
B
600
400
200
0
0
40
80
120
160
Time (days)
200
240
Figure 6.6 Concentrations of As(III) and As(V) in the influent and effluent of column
SF1 (Panel A) and column SF2 (Panel B) as a function of time: As(III) concentrations:
(▲) influent; (∆) effluent. As(V) concentrations: influent (●); effluent (○).
250
-1
Arsenic concentration (ug L )
800
600
400
200
0
SF1 Influent
SF1 Effluent
SF2 Influent
SF2 Effluent
Figure 6.7 Arsenic speciation in the influent and effluent of sand packed columns
starting from at 30 when steady status was achieved. Column was supplied with 6.67 µM
As(III), 36 µM Fe(II) and 2.5 mM nitrate (column SF1), or only 6.67 µM As(III) and 36
µM Fe(II) (column SF2): As(III): (solid block); As(V): (empty block).
6.4.4. Residual iron and arsenic in the sand at the end of reactor operation
At the end of the column experiment, iron and As adsorbed on the ICS were extracted to
determine the solid-phase iron and As retained in the column packing. The extracted As
and iron from the column are compared with the quantity of As and iron retained in the
column as estimated from the differences between the influent and effluent As (As
251
species and total As) and iron (total Fe) to make the mass balances. The recovery
calculated as the ratio of extracted to retained for both columns is shown on Table 2.
The total iron extracted on the sand is 116.4% and 115.9% match up the total iron
adsorbed in column SF1 and SF2, respectively. The results indicate that the iron removal
could be accounted for by extractable iron in the sand bed. The result in Figure 8 showed
that Fe(III) was predominately solid-phase iron species, and there is clearly a greater
level of precipitation of iron hydroxides on the sand at the base in column SF1, which is
up-flow through column. In contrast, there was no profile in SF2 with less iron(II)
uniformly immobilized at all heights inside the column. The results indicate that Fe(III)
oxides precipitated more readily than Fe(II) on the surface of sand.
The extracted As at the end of the experiment is accounted for 109.9% and 97.5%
of the As retained in column SF1 and SF2 respectively shown in Table 2. And there is
clearly a greater level of As adsorption at the base as shown in Figure 9. In comparison,
only small amount of As(III) was adsorbed and evenly distributed in column SF2. The
extracted As speciation (Table 3) showed that As(V) was dominant solid As speciation in
column SF1, however the majority of As were As(III) in column SF2. This proved that
As(III) was oxidized to As(V) in column SF1 with nitrate, but not in column SF2 lacking
nitrate.
252
Table 6.2 Mass balance of arsenic and iron at the end of the experiment.
Parameter
Sum of species of As
Retained in the column
Extracted
Recovery†
Influent-Effluent (mg)
from column packing (mg)
(%)
SF1
45.16
49.65
109.9
SF2
6.4
6.24
97.5
SF1
43.27
49.61
114.7
SF2
4.87
6.59
142.7
SF1
1290.3
1501.5
116.4
SF2
181.4
210.3
115.9
Column
Total As
Total iron
*
*
: SF1: Column fed with nitrate (2.5 mM), Fe(II) (0.36 mM) and As(III) (6.67 µM). SF2: Column fed with Fe(II) (0.36 mM)
and As(III) (6.67 µM) only.
†
: Recovery was calculated from the mass ratio of extracted to retained.
253
Table 6.3 Mass of arsenic and iron on the ICS of the two continuous sediment columns at different depth intervals.
Column*
SF1
SF2
Profile#
Mass of ICS profile
Arsenic recovered from ICS
Iron recovered from ICS
(gr)
(mg)
(mg)
As(III)
As(V)
Total As
Total Fe
Top
211.73
0.013±0.004
1.272±0.167
1.329±0.156
49.545±4.235
Middle
195.06
0.027±0.004
4.843±0.543
5.003±0.573
140.830±6.437
Bottom
196.70
0.946±0.196
41.563±1.560
43.282±1.353
1311.073±65.394
Top
202.15
1.149±0.255
0.039±0.002
1.416±0.494
47.707±7.277
Middle
203.66
1.762±0.096
0.106±0.019
1.871±0.083
58.247±17.718
Bottom
197.60
2.974±0.419
0.210±0.034
3.302±0.472
104.33±14.622
*
SF1: Column fed with nitrate (2.5 mM), Fe(II) (0.36 mM) and As(III) (6.67 µM). SF2: Column fed with Fe(II) (0.36 mM)
and As(III) (6.67 µM) only.
#
Profile: The top, middle and bottom section have the same volume in the up-flow through column.
254
25-50
0-25
Height (cm)
50-75
A
0
1
2
3
4
5
25-50
50-75
B
0-25
Height (cm)
6
0
1
2
3
4
Cs (mg Fe/g ICS)
5
6
Figure 6.8 The profile of sorbed iron on the ICS in two sands packed up-flow columns at
the end of the continuous experiment. Panel A: Column SF1 fed with with 6.67 µM
As(III), 36 µM Fe(II) and 2.5 mM nitrate (column SF1), or panel B: only 6.67 µM As(III)
and 36 µM Fe(II) (column SF2). Fe(II): (solid block); Fe(III): (empty block).
255
25-50
0-25
Height (cm)
50-75
A
0
50
100
150
200
25-50
50-75
B
0-25
Height (cm)
250
0
50
100
150
Cs (ug As/g ICS)
200
250
Figure 6.9 The profile of sorbed As on the ICS in two sands packed columns at the end
of the continuous experiment. Panel A: Column SF1 fed with with 6.67 µM As(III), 36
µM Fe(II) and 2.5 mM nitrate (column SF1), or panel B: only 6.67 µM As(III) and 36 µM
Fe(II) (column SF2). As(III): (solid block); As(V): (empty block).
256
6.4.5. Mineralogy of iron and arsenic on the surface of sand
The SEM was used to observe the surface characteristics of the sand before and after
column studies. The original sand (Figure 10 A and B) showed very ordered silica
crystals at the surface. The natural sand had a relatively uniform and smooth surface, with
small cracks, micro pores or light roughness found on the sand surface. The image
obtained for ICS from SF1 (Figure 10 C) had significantly rougher surfaces and more
presence of porous structures than the original sand, which was composed of fresh
precipitates and bacteria. The precipitates were observed as mixture of amorphous and
crystalline form, which may be responsible for the high surface area and surface active
sites. The bacteria are rod and roughly 1-2.5 nm immobilized on the surface. And the
original sand surfaces are not visible. In comparison, the sand of control column SF2
(Figure 10 E) demonstrated the similar morphology with original sand surface, which
was still visible. The surface was pocked with occasional precipitates, but much less than
ICS in SF1. And there was less evidence of bacteria present on the surface. The EDS
analysis (Figure 10 D and F) revealed that much less iron and no As were adsorbed on
the sand surface of the control column SF2.
257
A
B
C
D
E
F
Figure 6.10 SEM-EDS for column profile: original sand: SEM (panel A) and EDS (panel
B); SF1 SEM (panel C) and EDS (panel D); SF2 SEM (panel E) and EDS (panel F).
258
The XRD analysis (data not shown) of precipitates recovered the end of the SF1
experiment revealed the presence of SiO2 (from original sand) and a mixture of Fe(III)
oxides dominated by crystalline hematite suggesting Fe(III) oxides formation due to
nitrate-dependent oxidation of soluble Fe(II); on the other hand, the spectra demonstrated
that SF2 surface was predominantly composed of SiO2 and a few Fe(II) minerals such as
siderite (FeCO3).
The XPS results are shown in Figure 11 comparing samples of original sand, SF1
and SF2 with As (As(III) and As(V)) and Iron (Hematite) standard spectra. The spectra of
the original sand surface were still largely recognized in the sample from the SF2 control
column lacking nitrate. The oxidation state of As cannot be clearly indentified on the
surface of SF2 sand due to very low content. The surface of SF1 sand clearly contained
Fe(III) with a spectrum characteristic of hematite. The iron spectra of SF1 matched the
hematite standard samples with same binding energy. The SF2 spectrum did not have any
signature regarding hematite. The As spectra indicated that the presence of As(V) in
column SF1 as evidenced by a binding energy which matched closely As2O5 standard.
The combined results suggest a significant increment of iron(III) oxides and
As(V) contents in the precipitates coating the sand in active biological column in the
presence of nitrate providing an additional confirmation for the anoxic oxidation of Fe(II)
and As(III) coupled to denitrification.
259
Figure 6.11 XPS for original sand, SF1 and SF2 column profile: As standard (panel A),
As samples (panel B), and Iron (panel C).
260
6.5. Discussion
6.5.1. Bioremediation of As in the presence of NO3-
In this study, two continuous flow sand-packed columns were operated to simulate a
natural anaerobic groundwater-sediment system with co-occurring As(III) and Fe(II) in
the presence or absence of nitrate. The results provide evidence that the microbial anoxic
nitrate-dependent oxidation of Fe(II) and As(III) enhanced the adsorption of As on the
freshly formed solid-phase iron(III) oxides coating sand surfaces. During the operation
period from day 30 to 225, 94.4 (±4.4) and 89.8 (±6.1)% respectively of the influent
Fe(II) and As(III) entering the column were retained in the column with added nitrate
(SF1), compared with only 8.9 (±7.8) and 3.6 (±3.3)% in the control column lacking
nitrate (SF2). The extraction results showed that sand in SF1 contained 10-fold higher
quantities iron and As than SF2. The biological denitrification activity of the column SF1
played an important role in the immobilization process as evidenced by the considerably
less removal of soluble Fe(II) and As(III) in nitrate-free control column.
6.5.2. The mechanisms of immobilization of As on iron (hydr)oxides in anoxic
environments
The end products of anaerobic microbial oxidation of Fe(II) have significant influences
on soil and sediment mineralogy and geochemistry through the formation of a broad
261
variety of environmentally relevant Fe(III) oxide-containing minerals, which could
immobilize heavy metals and metalloids contaminants in natural environments via
adsorption, co-precipitation [Finneran et al., 2002; Lack et al., 2002; Senn and Hemond,
2002]. Based on our observations and findings from literatures, the microbial nitratedependent oxidation of Fe(II) and As(III) could potentially enhance the precipitation of
As in subsurface environments. To better understand the processes of As immobilization
in anoxic environment, two possible mechanisms can be generally rationalized related to
Fe(II) and As(III) oxidation.
The first mechanism is that the removal of As(III) may occur through a two-step
process: iron (hydr)oxideds formation due to nitrate-dependent Fe(II) oxidation, and
subsequent adsorption of As(III). In this study, more than 91% of Fe(II) was adsorbed in
the SF1 column in the presence of nitrate. The solid-phase extraction revealed that iron
was retained on the sand surface as oxidation state of Fe(III), which was close to the
soluble Fe(II) eliminated. And the EDS results showed that a higher iron content on the
sand surface of column SF1 than column SF2. The mineralogy analysis of XRD and
XPS illustrated that the mixture of iron(III) oxides dominated by crystalline hematite
suggesting Fe(III) oxides formation due to nitrate-dependent oxidation of soluble Fe(II).
The microbial nitrate-dependent Fe(II) oxidation to Fe(III) oxides is a well
established process among prokaryotes in diverse ecosystems including activated sewage
sludge, anoxic aquifer sediments, and marine sediments [Senko et al., 2005; Weber et al.,
2006]. The effluent from the SF1 column demonstrated the microbial activity of anoxic
oxidation of Fe(II) linked to denitrification. The biological nature of the reaction is
262
inferred from the lack of any conversion in uninoculated samples and nitrate free
samples. Literature findings have identified nitrate-dependent biogenic formation of
minerals such as ferric oxyhydroxide, goethite, hematite and magnetite in anoxic
environments [Benz et al., 1998; Straub et al., 1998; Chaudhuri et al., 2001; Lack et al.,
2002; Weber et al., 2006].
Previous studies have illustrated that injecting dissolved O2 into groundwater with
high dissolved Fe(II) concentrations leads to formation of iron oxides, thereby, increasing
removal of arseinc [Appelo and deVet, 2003; Welch et al., 2008]. Katsoyiannis et al.,
[2002; Katsoyiannis and Zouboulis, 2004] have indicated that the removal of As(III)
following aerobic microbial and abiotic Fe(II) oxidation was very efficient, reaching
percentage efficiencies over 90%. It has also been found that during the biological
removal of iron, a significant simultaneous increase of As removal was also observed
[Katsoyiannis and Zouboulis, 2005]. Roberts et al. [Roberts et al., 2004] demonstrated
that addition of Fe(II) instead of the usually pre-formed Fe(III) (hydr)oxides showed
advantage of higher As removal efficiency, because iron(III) (hydr)oxides freshly formed
from the oxidation of Fe(II) by dissolved oxygen had higher sorption capacities.
The microbial oxidation of Fe(II) results in the formation of an iron (hydr)oxide
adsorbing matrix, removing As as has been observed in the field study in Bangladesh,
where As was attenuated by injection of nitrate into contaminated groundwater [Harvey
et al., 2002]. Several studies have effectively utilized different iron oxides coated
materials such as iron-oxide coated sand (IOCS) [Thirunavukkarasu et al., 2003; Gupta et
263
al., 2005], iron oxide coated cement (IOCC) [Kundu et al., 2006, 2007], iron impregnated
quartz sand (IIS) [Vaishya et al., 2003] and iron and manganese amended activated
alumina (IMAA) [Dhiman and Chaudhuri, 2007] to remove As from an aqueous solution.
Both As(III) and As(V) show high affinity, but different adsorption behaviors on
iron oxides in soil and subsurface environments [Fukuoka et al., 2006]. The sorption of
As(III) and As(V) has been studied previously, actually, onto various types of iron oxides
such as amorphous hydrous ferric oxide (FeOOH), poorly crystalline hydrous ferrihydrite
[Raven et al., 1998; Goldberg et al., 2001], hematite (Fe2O3) [Jeong et al., 2007; Gimenez
et al., 2007], goethite (α-FeOOH) [Dixit and Hering, 2003] and magnetite (Fe3O4)
[Gimenez et al., 2007] (Table S2 for As(III); S3 for As(V)). Most studies indicate that
both As(III) and As(V) show similar adsorption capacity on iron(hydr)oxides at
circumneutral conditions. For example, in the natural environments with the pH range 69, As(III) is sorbed to a similar extent as As(V) on hydrous ferric oxides (HFO) and
goethite [Dixit and Hering, 2003].
Unlike the studies mentioned before, the adsorption isotherms of As(III) and
As(V) in this study demonstrated that As(III) adsorption was much less efficiently than
As(V) on nitrate-dependent biogenic ICS at neutral pH. In this case, As(V) is present in
negatively ionic forms; however As(III) is dominant by non-ionic forms. As(V), as
anionic species, would thus have stronger interaction (specific binding) with positive
surface of iron oxides and have higher uptake than As(III). The findings from Bowell
demonstrated that the sorption capacity of As(V) is much higher than As(III) on natural
goethite and magnetite [Bowell et al., 1994]. Chang [Chang et al., 2008] found that the
264
adsorption rate of As(V) onto ICS was greater than that of As(III), and ICS showed a
greater adsorption capacity for the removal of As(V) than As(III).
The second mechanism is that nitrate-dependent oxidation of As(III) to As(V)
could enhance the adsorption of As on the iron (hydr)oxides, which shows higher affinity
on As(V) than As(III) in this study and literatures [Bowell et al., 1994; Chang et al.,
2008].
During the long term operation of continuous columns, the removal of As(III) was
10-fold higher adsorbed in column SF1 in the presence of nitrate than column SF2
without nitrate. The extraction results revealed that the As (mass sum of As(III) and
As(V)) recovered from sand corresponded to the As (mass sum of As(III) and As(V))
removed in both column SF1 and SF2. Although As(III) was prevalent in the influent in
column SF1, the As retained on the sand surface was dominant in the form of As(V).
Furthermore, the results obtained by XPS show a strong indication that As(V) was the
dominant speciation of solid-phase As retained on the ICS of column SF1. In contrast, the
As extracted from column SF2 was As(III), and the effluent was also As(III), suggesting
no oxidation of As(III) in column SF2 without nitrate.
In our previous study we report on diverse anaerobic microorganisms were
involved in the anoxic oxidation of As(III) linked to denitrification in from sludges and
sediments with no prior exposure to As [Sun et al., 2008]. The biological nature of the
reaction is inferred from the lack of any conversion in uninoculated samples, heat killed
samples and samples lacking nitrate.
265
Although most studies indicate that As(III) and As(V) show strong retention on
iron oxides, As(V) binds to ICS more extensively than As(III) in this study. Hsu [Hsu et
al., 2008] also found that the adsorption of As(V) in anion form was more favorable than
As(III) onto the ICS surface at pH 7.0. As(III) was contrarily shown to be more mobile
under flow conditions than As(V) [Gulens et al., 1979]. Moreover, Herbel et al. [2006]
demonstrated that As(III) desorption was much easier and extensive than As(V) in
continuous column packed with ferrihydrite, although As(III) showed a 2-fold higher
adsorption capacity on As(III) than As(V). In other words, these findings indicate that the
anoxic oxidation of As(III) was attributable to the presence of nitrate, which enhance the
adsorption and retention of As on ICS at pH neutral environments.
6.5.3. Implications
Iron (hydr)oxides strongly adsorb both As(V) and As(III) and, thus offer significant
control of the dissolved concentration of As in natural environments. Anaerobic
microbial reduction and dissolution of iron(hydr)oxides, as well as dissimilitary reduction
of As(V) to As(III) are major mechanisms of mobilizing As in soil and sediments. Fe(II)
and As(III) commonly co-occur in contaminated groundwater and surface water under
anaerobic conditions. On the other hand, the oxidation of Fe(II) and As(III) could be an
important bioremediation strategy to generate iron(hydr)oxides and immobilize As(V) on
the solid phases. Although dissolved oxygen can readily oxidize Fe(II) and As(III)
abiotically and biologically, respectively, it is difficult to diffuse dissolved oxygen into
266
anoxic zone of submerged subsurface due to its low solubility and high reactivity.
However, nitrate could be utilized as an alternative electron acceptor with advantages of
high solubility, and less reactive enabling it to disperses in the saturated subsurface. The
study presented here validates that microbial nitrate-dependent oxidation of Fe(II) and
As(III) enhance the immobilization of As in the anoxic environments.
6.6. Acknowledgments
The work presented here was funded by a USGS, National Institute for Water Resources
104G grant (2005AZ114G), and by a grant of the NIEHS-supported Superfund Basic
Research Program (NIH ES-04940). The use of trade, product, or firm names in this
report is for descriptive purposes only and does not constitute endorsement by the U.S.
Geological Survey. Also arsenic analyses were performed by the Analytical Section of
the Hazard Identification Core (Superfund Basic Research Program grant NIEHS-04940).
267
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276
CHAPTER 7
NOVEL STUDY OF ANOXIC OXIDATION OF As(III) TO As(V) LINKED TO
CHLORATE REDUCTION
7.1. Abstract
Microorganisms play a key role in the speciation and mobility of arsenic in the
environment. In this study, the anoxic oxidation of arsenite (As(III)) linked to
(per)chlorate reduction was shown to be a feasible phenomenon in anaerobic sludge
samples that were not previously exposed to arsenic contamination. When incubated with
0.5 mM As(III) and 3.0 mM ClO3-, the anoxic oxidation of As(III) occurred within a few
days, achieving specific activities of up to 1.24 mmol As(V) formed g-1 volatile
suspended solids d-1. No activity was observed in samples lacking inoculum or with heatkilled sludge, and in controls lacking ClO3-. The anoxic oxidation of As(III) was inhibited
50% at 0.29 mM As(III) and 85% at 1.2 mM As(III) compared to activity of lowest tested
concentration of 0.005 mM. The oxidation of As(III) was shown to be linked to the
complete reduction of ClO3- to Cl-. The Cl- production corresponded closely to the
amount expected from the theoretical stoichiometry of the reaction, which is 3.0 mol
As(III) oxidized mol-1 Cl- formed. Additionally, microbial anoxic oxidation of As(III)
was stable over prolonged periods of operation ranging up to 550 days in the continuous
bioreactor with bacteria immobilized in bio-film granules. The oxidation of As(III) linked
277
to the use of chlorate as an electron acceptor could be a potential bioremediation
technology by converting As(III) to less toxic and less mobile As(V).
Keywords:
Arsenite,
Oxidation,
Chlorate,
Anaerobic,
Enrichments, Continuous bioreactor, Bioremediation.
Autotrophic,
Heterotrophic,
278
7.2. Introduction
The contamination of drinking water with arsenic is an global public health issue. As is
human carcinogenic compound [WHO, 1993; ATSDR, 2007], which poses a risk to
millions of people around the world [Smedley and Kinniburgh, 2002]. In certain regions
in the world, As is found at relatively high concentrations in groundwater and sediments
as a consequence of the natural dissolution and weathering of naturally occurring Asbearing geological material [Smedley and Kinniburgh, 2002].
The most common oxidation states of As found in natural circumneutral pH
aqueous environments generally occur as either arsenite (As(III), H3AsO3) or arsenate
(As(V), H2AsO4- and HAsO42-). Clay minerals and iron oxides, the most common
constituents of the subsurface environments, can adsorb As and immobilize it in the solid
phase [Goldberg, 2002]. The relative binding affinity of As(III) and As(V) depends on
the dominant part what constitute in sediment and soil. Thus transformations between
two species may govern the mobility of As [Dixit and Hering, 2003]. Although As and Fe
transformations may be influenced by abiotic reactions, microbial processes play critical
roles in controlling the fate and transformation of As and Fe in the subsurface system
[Senn and Hermond, 2002; Oremland et al., 2005; Kocar et al., 2006; Herbel and
Fendorf, 2006]. As(V) binds to aluminum oxides more extensively than As(III) under
circumneutral conditions [Lin and Wu, 2001; Goldberg, 2002; Hering and Dixit, 2005].
Thus, microbial reduction of As(V) to As(III) will decrease the adsorption capacity of As.
Both As(III) and As(V) are strongly adsorbed on iron oxides [Dixit and Hering, 2003].
279
Dissimilatory reductive dissolution of iron oxyhydroxides is common place in anaerobic
environments [Straub et al., 2001]. The resulting loss in iron oxides sorbent mass should
be expected to induce the release As and enhance the mobility of As in groundwater and
surface water [Oremland et al., 2005; Anawar et al., 2006].
On the other hand, aerobic oxidation of Fe(II) and As(III) are capable of
generating Fe(III) oxyhydroxides for adsorption of As and As(V) formed increase the
removal efficiency on minerals [Katsoylannis and Zouboulis, 2006]. However, the
dissolved oxygen (DO) in water maybe consumed by organic matter, sulfides, and other
reducing compounds in the subsurface and become anoxic even strict anaerobic
conditions. Recently, several studies have demonstrated that nitrate-dependent As(III)
oxidation can be catalyzed by anaerobic microorganisms to gain energy from As(III)
oxidation [Sun et al., 2008]. As(III)-oxidizing denitrifying bacteria were reported to be
widespread in various environments including As-contaminated lakes and soil [Oremland
et al., 2002; Rhine et al., 2006], or As pristine sediments and sludges samples [Sun et al.,
2008]. As(III) oxidation linked to complete denitrification was demonstrated in three
enrichment cultures. Evidence for the complete denitrification was demonstrated by
direct measurement of conversion to N2. Clone library characterization of the enrichment
cultures indicates that the predominant phylotypes were from the genus Azoarcus and the
family Comamonadaceae and Comamonadaceae [Chapter III].
Beside nitrate, chlorate (ClO3- ) is also considered as a possible alternative oxidant
with the potential to bioremediate contaminated plumes [Coats et al., 1999; Chakraborty
et al., 2005], although this speices is not a naturally abundant compound and lacks
280
geological importance. (Per)chlorate (perchlorate or chlorate) has been recognized as a
widespread contaminant in groundwater and surface water [Bruce et al., 1999; Urbansky,
2000]. Perchlorate (ClO4-) is commonly used as a terminal electron acceptor by anaerobic
bacteria; as a result, it is completely degraded to the benign end product, chloride (Cl-).
Microbial reduction of perchlorate proceeds via a three-step process of: ClO4- → ClO3- →
ClO2- → O2+Cl- [Rikken et al., 1996; Steinberg et al., 2005]. Reduction of perchlorate to
chlorate, which is the limiting step, and chlorite is catalyzed by respiratory (per)chlorate
reductases [Bender et al., 2005; Logan et al., 2001]. Disproportionation of chlorite into
Cl- and O2 is catalyzed by chlorite dismutase, which is the fastest step, and the oxygen
produced is immediately consumed for energy of cell synthesis [Logan et al., 2001; Xu et
al., 2004]. In anoxic conditions, microbial oxidation of Fe(II) to form ferrihydrite or
amorphous ferric oxyhydroxides has been demonstrated utilizing nitrate, perchlorate and
chlorate as electron acceptor [Weber et al., 2006]. Such an oxidation could potentially
improve the removal of As(III) from the aqueous phase since Fe(III) (hydr)oxides are
strong adsorbents of inorganic As [Raven et al., 1998; Dixit and Hering, 2003] .
The main objective of this study is to explore the potentials of chlorate to serve as
an electron acceptor for the microbial oxidation of As(III) by anaerobic bacteria. The
theoretical stoichiometry of the reaction is presented below:
ClO3- + 3H3AsO3 → Cl- + 3HAsO42- + 6H+
(eq 1)
Based on bioenergetic considerations, the reaction is feasible as indicated by the
highly exergonic standard change in Gibb’s free energy (∆G0’= -249.6 kJ mol-1 As(III)).
281
In this study, we report on the anoxic biological oxidation of As(III) by chlorate respiring
microorganisms in sludges with no prior exposure to As. The study demonstrates that the
biological capacity for (per)chlorate linked to As(III) oxidation with chlorate reduction is
widespread in anaerobic environmental samples. However, the activity is sensitive to
high concentrations of As(III).
7.3. Material and Methods
7.3.1. Microorganisms
Sludge and sediment samples from different locations were used as inocula in the batch
bioassays. Aerobic activated sludge (RAS) and anaerobically digested sewage sludge
(ADS) was obtained from a local municipal wastewater treatment plant (Ina Road,
Tucson, Arizona). Methanogenic granular sludge (biofilm pellets) samples were obtained
from industrial upward-flow anaerobic sludge blanket (UASB) treatment plants treating
recycled paper wastewater (EGS) (Industriewater, Eerbeek, The Netherlands) and alcohol
distillery wastewater (NGS) (Nedalco, Bergen op Zoom, The Netherlands). Two
chemolithotrophic As(III)-oxidizing denitrifying granular sludge were obtained from two
laboratory-scale As(III)-oxidizing denitrifying bioreactor biofilm, AODBB (2 l)
described previous [chapter IV]. The total suspended solids’s (TSS) contents of the
sludge and sediment samples were 6.32±0.27, 17.64±1.12, 16.65±0.21, 7.10±0.54,
6.08±0.18 and 1.63±0.09% wet weight basis; and the volatile suspended solids’s (VSS)
282
content were 5.96±0.28, 12.91±0.97, 10.03±0.10, 4.78±0.08, 5.76±0.23 and 1.12±0.07%
wet weight basis for NGS, EGS, ADS, AODBB, NRS and RAS, respectively. The
granular sludge samples were washed and sieved before use to remove fines. The inocula
were stored under nitrogen gas at 4°C.
7.3.2. Basal medium
The standard basal medium was prepared using ultra pure water (Milli-Q system;
Millipore) and contained the following compounds (mg l-1): NH4HCO3 (3.16); NaHCO3
(672); CaCl2 (10), MgSO4.7H2O (40); K2HPO4 (300); KH2PO4.2H2O (800); and 0.2 ml l-1
of a trace element solution containing (in mg l-1): FeC13.4 H20 (2,000); CoCl2. 6 H20
(2,000); MnCl2 4 H20 (500); AlCl3 6 H20 (90); CuCl2.2H20 (30); ZnCl2 (50); H3BO3 (50);
(NH4)6Mo7O24.4 H2O (50); Na2SeO3.5 H2O (100); NiCl2.6 H20 (50); EDTA (1,000);
resazurin (200); HCl 36% (1 ml).
7.3.3. Batch bioassay
Batch bioassays were performed in shaken flasks, which were incubated in a dark
climate-controlled room at 30±2˚C. Serum flasks (160 ml) were supplied with 120 ml of
a basal mineral medium (pH 7.0-7.2) containing bicarbonate as the only C source, as
described above. The medium was also supplemented with As(III) as electron donor
(concentration indicated in Tables and Figures of each experiment) and chlorate as the
283
electron acceptor, typically 3.0 mM, unless otherwise specified. Flasks for anoxic assays
were sealed with butyl rubber stoppers, and then the medium and headspace were purged
with N2/CO2 (80/20, v/v) for 20 min to exclude oxygen from the assay. Various controls
(e.g. abiotic controls, killed sludge controls, controls without electron acceptor, and
controls without As(III)) were applied based on the requirements of each experiment.
Abiotic controls were prepared without adding microbial inoculum. Killed sludge
controls were prepared by autoclaving the flasks added with inoculum for 20 min at
121ºC, the content was allowed to cool down overnight, and the cycle was repeat two
more times and then sealed aseptically. Controls without electron acceptor was included
to prove no oxidation of As(III) by potential contamination of oxygen. Controls lacking
As(III) were included to measure endogenous consumption of chlorate and correct the net
chlorate reduction linked to As(III) oxidation. All assays were conducted in triplicate.
Liquid samples were analyzed for the concentration of electron acceptor and
biotransformation products (ClO3-, ClO2-, and Cl-) and As species (As(III) and As(V)).
Granular sludge was added to the assays at 10-12 g wet weight l-1 medium; liquid sludge
such as RAS and ADS inoculum was added to the assays at 6% (v/v).
7.3.4. Enrichment cultures
Under restricted anaerobic conditions, enrichment cultures were established by adding
granular sludge inoculum to the assays at 20 g wet weight l-1 medium. Liquid sludge
284
inoculum was inoculated to the assays at 50 ml l-1 medium (5%). The medium was basal
medium amended with 0.5 mM As(III) (supplied as NaAsO2) as electron donor, 3.0 mM
ClO3- (supplied as NaClO3) as electron acceptor and 8 mM HCO3- (NaHCO3) as major
carbon sources. The basal medium with NO3- was sterilized with an autoclave at 121 ºC
three cycles as described before the killed sludge controls, while NaHCO3 and As(III)
were sterilized using membrane filters (0.2 µm). Cultures for anoxic assays were
incubated in 160 ml sealed with butyl rubber stoppers, and then the medium and
headspace were purged with N2/CO2 (80/20, v/v) for 20 min to exclude oxygen from the
assay. Once oxidation of the added As(III) had occurred, a 5% (v/v) dilution of the active
cultures were made into fresh medium and incubated as previously described. After the
active enrichment cultures were serially transferred to the point where there was no more
sludge, aliquots of the active cultures were transferred to fresh liquid medium and
maintain alive for identification and characterization. The autotrophic cultures were
enriched with bicarbonate as sole carbon sources; the heterotrophic enrichments were
incubated with yeast extract (1.0 mg l-1) as the addition carbon sources besides
bicarbonate (8.0 mM). Autotrophic enrichment chlorate culture (ECC1 and ECC3), and
heterotrophic enrichment chlorate cultures (ECC2 and ECC4) were originally derived
from ADS and RAS, respectively.
Once the cultures had been cultivated, they were tested for their ability to grow on
As(III) under chlorate-respiring conditions in the same liquid medium. Growth was
monitored routinely by measuring As(III) oxidation and ClO3- reduction. The basal
medium previously described was used for routine maintenance and characterization
285
experiments. To avoid the contamination of carrying ClO3- from old cultures to fresh
medium, the cultures (5% volume of previous culture) were centrifuge in sterilized
eppendorfs at 10,000 rpm for 10 min. The pellets were collected and washed into
sterilized MiliQ water. Next the centrifugation and wash step were repeated, and finally
the pellets were resuspended into same volume of sterilized MiliQ water and transferred
to designed experiments. Cultures were serially transferred to fresh medium after
incubations of 2-3 weeks at 30ºC and confirmation of As(III) oxidation (measured by
As(V) formation). The enrichment process was continued for 23 transfers.
7.3.5. Most Probable Number (MPN)
MPN assays were performed in A/N-BM. Culture samples (1 ml) were taken after 23
transfers, were homogenized by vortexing, and were serially diluted in 10-fold
increments to 10-9 in sterile medium under anaerobic conditions. Each dilution was set up
with five replicate MPN tubes. The tubes were incubated in an orbital shaker (110 rpm) at
30°C under anaerobic conditions with N2/CO2 (80:20) in the headspace. After 3-4 weeks
incubation, samples from each tube were removed and analyzed for As(V). Conversion of
> 80% of the As(III) to As(V) was considered as a positive tube. Finally, an MPN table
for five tubes (APHA, 1999) was utilized to enumerate the number of bacteria with
As(III)-oxidizing capability.
286
7.3.6. Continuous column
In the continuous bioreactor study, a larger column was operated, which was 2-l benchscale upward-flow anaerobic sludge bed (UASB) reactor. This reactor was inoculated
with 33 g VSS l-1 of ADS. The reactor was fed with basal mineral medium and As(III)
(concentrations vary according to operational periods as shown in Table 5) as the sole
energy source, chlorate (5.47±0.44 and 3.34±0.29 mM for day 0-104 and 105-550,
respectively) as the sole electron acceptor and NaHCO3 as the major carbon source. The
goals were to explore whether As(III) can be efficiently oxidized with addition of
chlorate in continuous bioreactor under anoxic conditions over prolonged periods of
operation; to enrich the As(III)-oxidizing microorganisms linked to chemolithotrophic
chlorate reduction.
The average hydraulic retention time of the columns in both experiments was
approximately 1 d. The bioreactor was placed in a climate controlled room at 30±2°C and
covered by aluminum foil to avoid growth of phototrophic bacteria. The pH of the
influent was adjusted to 7.2 with NaOH or HCl, as required. The influent was maintained
at all times under an N2 atmosphere to minimize DO from entering the medium which
could result in the unwanted aerobic oxidation of As(III). Influent and effluent samples
were prepared immediately for analysis to minimize possible changes in As speciation
upon exposure to the atmosphere. The pH value was determined immediately after
sampling.
287
7.3.7. Analytical methods
As(III) and As(V) speciation was analyzed by ion chromatography–inductively coupled
plasma–mass spectroscopy (HPLC-ICP-MS). The system consists of an HPLC (Agilent
1100) and an ICP-MS (Agilent 7500a) with a Babington nebulizer as the detector. The
operating parameters were as follows: Rf power 1,500 watts, plasma gas flow 15 l min-1,
carrier flow 1.2 l min-1, As was measured at 75 m/z and terbium (IS) measured at m/z
159. The injection volume was 10 µl. The detection limit for the various As species was
0.1 µg l-1. The total concentration of As in liquid samples was determined by direct
injection into ICP-MS using an ASX500 auto sampler (CETAC Technologies, Omaha,
NE). The analytical system was operated at an Rf power of 1500 W, a plasma gas flow of
151 min-1 and a carrier gas flow of 1.21 min-1. The acquisition parameters used were as
follows: As measured at m/z 75; terbium (IS) measured at m/z 159; three points per peak;
1.5 s dwell time for As, 1.5 s dwell time for Tb; number of repetitions = 7.
Chlorate, chloride and arsenate (As(V)) were analyzed by suppressed conductivity
ion chromatography using a Dionex 500 system (Sunnyvale, CA, USA) fitted with a
Dionex IonPac AS11 analytical column (4 mm x 250 mm) and a AG16 guard column (4
mm x 40 mm). During each injection the eluent (20 mM KOH) was used for 20 min.
Other analytical determinations (e.g., pH, TSS, VSS, etc.) were conducted
according to Standard Methods [APHA, 1999].
288
7.4. Results
7.4.1. Screening
Sludge samples from sources known to be pristine with respect to As were incubated with
As(III) (0.5 mM) either in the presence or absence of ClO3- (3 mM) under strict anaerobic
conditions. Four out of the five inocula tested, including RAS, ADS, NGS, EGS and
AODBB1, displayed microbial activity towards the anoxic oxidation of As(III). The time
course from these incubations is provided in Figure 1 for ADS, RAS and NGS. In
treatments incubated with ClO3-, As(V) formation started after approximately 3 d and
was concomitant with the disappearance of As(III). After approximately 27 d, the
reaction was complete. The reactions were dependent on the presence of ClO3- as
evidenced by the lack of any significant conversion in incubations lacking ClO3-.
Likewise, the reactions were also dependent on being inoculated, since no reaction
occurred in bottles receiving the medium with As(III) and ClO3- but lacking sludge
inoculum or with heat-killed inoculum (data not shown).
289
0.60
As(III) Concentration(mM)
A
0.45
0.30
0.15
0.00
0
10
20
30
40
50
0.60
As(V) Concentration (mM)
B
0.45
0.30
0.15
0.00
0
10
20
30
Time (days)
40
50
Figure 7.1 The removal of As(III) (panel A) and formation of As(V) (panel B) by ADS,
RAS and NGS in the presence or absence of chlorate under anoxic condition, ADS (▲),
RAS (■) and NGS (♦) with ClO3-; ADS (), RAS (□) and NGS (◊) without ClO3-; and
abiotic (○).
290
The results of anoxic oxidation of As(III) for all the sludge samples and oxidants
screened are shown in Table 1. All the four positive samples including RAS, ADS, NGS,
and AODBB showed a dependency on the presence of chlorate for the anoxic oxidation
of As(III). The average molar yield of chlorate linked As(III) oxidation was 1.04±0.11
mol As(V) formed per mol As(III) consumed. Perchlorate was also used as an electronacceptor to oxidize As(III) under anoxic conditions with two of the inocula (table 1) but
incubations were longer taking up to 48 days for the conversion to come to complete with
the average yield at 1.01±0.14 by the RAS and ADS inoculum.
7.4.2. Kinetics
The bioassays with ClO3- as electron acceptor that were inoculated with NGS, ADS,
RAS and AODBB were sampled sufficiently to obtain kinetic data. The doubling time
and maximum activity in the sludge samples are summarized in Table 2. The activity,
estimated from the slope of the natural logarithm of As(V) formation versus time,
indicated doubling times of 2.5-5.5 d. The maximum activities ranged from 0.13 to 0.39
mM As(V) VSS d-1.
291
Table 7.1 Summary of oxidation of As(III) (0.5 mM) to As(V) linked to chlorate (3.0 mM) or perchlorate (2.5 mM).
As(V) formation (mM) ‡
Inoculum
Non-inoculated
ADS
RAS
NGS
EGS
AODBB
†
‡
Time†
With ClO3-
Without ClO3-
With ClO4-
Without ClO4-
__
__
__
__
0.454±0.024
__
0.448±0.018
__
0.486±0.036
__
0.408±0.064
0.497±0.009
__
__
0.476±0.026
Heat-kill
with ClO3- or
ClO4-
(Days)
ClO3-
ClO4-
__
__
__
27
27
__
__
15
15
__
__
__
15
__
__
__
__
__
__
__
__
__
__
__
__
: Time to oxidize 80% of 0.5 mM As(III) to As(V)
Conversion of As(III) to As(V): “Average ± STDE” for “conversion ≥ 80%”; “__ “ for “conversion ≤ 5%”. As(V) formation
=concentration of As(V) (at the end of the experiment, d 48) minus concentration of As(V) (at day 0).
292
Table 7.2 Summary of kinetics† of microbial As(III) oxidation (0.5 mM) under chlorate (3.0 mM) respiring conditions
Highest specific activity†
†
‡
Inoculum sources
Doubling time (days) ‡
NGS
3.34±1.40
0.22±0.02
ADS
2.62±1.25
0.39±0.18
RAS
5.02±0.98
0.30±0.05
AODBB
5.45±0.74
0.13±0.02
(mM As (V) d-1)
: Estimated from As(V) formation data.
: The coefficient of determination (R2) was 0.92, 0.96, 0.91 and 0.95 for the ln(∆As(V)) versus time plots of NGS, ADS, RAS
and AODBB.
293
7.4.3. MPN assays for enrichment cultures
MPN analysis of the As(III)-oxidizing, chlorate-respiring populations were performed
after the 13th and 23rd transfer of enrichment cultures. The results (Table S1) revealed that
2.2×107, 2.4×107, 0.33×107, and 1.7×107 cells were present per ml of ECC1, ECC2,
ECC3 and ECC4 respectively after 23rd transfer; corresponding to 6.5 × 1012 to 4.4 × 1013
cells produced per mol of As(V) formed.
Table S1 MPN estimations of cell density and cell yield of four enrichment cultures
Sample Source
Cell density†
(cells ml )
(cells ml )
(cells mol-1 As(V)
formed)
ECC1
ADS autotrophic
0.17×107
2.2×107
4.38×1013
ECC2
ADS heterotrophic
0.09×107
2.4×107
4.39×1013
ECC3
RAS autotrophic
0.11×107
0.33×107
0.65×1013
ECC4
RAS heterotrophic
0.07×107
1.7×107
3.39×1013
-1
†
: MPN assays were conducted at 13 time transfer on July, 2007
‡
: MPN assays were conducted at 23 time transfer on July, 2008
Cell density‡
Cell Yield
Cultures
-1
294
7.4.4. Rates of anoxic As(III) oxidation linked to autotrophic and heterotrophic chlorate
reduction
Two enrichment cultures (ECC1 and ECC3) were tested for their ability to utilize
inorganic and organic carbon sources for growth under anoxic conditions. The
autotrophic ECC1 and heterotrophic ECC3 were incubated under anoxic conditions with
As(III) both in the presence and absence of ClO3-. The time course from these
incubations is provided in Figure 2 (autotrophic ECC1) and Figure 3 (heterotrophic
ECC3). In all treatments, the formation of As(V) corresponded to the stoichiometric
elimination of As(III). In ECC1, As(V) formation started after approximately 1 d and the
reaction came to completion after 28-30 d both in the presence and absence of yeast
extract (YE, 1 mg l-1),. The maximum reaction rate, calculated from the formation of
As(V) (Table 3), was 0.13±0.03 and 0.20±0.08 mM As(V) formed d-1 for treatment with
and without YE, respectively. The results indicated that the autotrophic ECC1 could
oxidize As(III) in the presence of ClO3- without YE, although the addition of YE had a
modest stimulation of the activity. In ECC3, it took only 12-15 d to complete the
oxidation of As(III) to As(V) when YE or pyruvate (Pyr, 1.7 mg l -1) was added. In
comparison, ECC3 without any organic carbon addition did not convert As(III) to As(V).
The results suggested that the heterotrophic ECC3 has a strict requirement for organic
carbon. The highest reaction rate (Table 3) was 0.34±0.08 and 0.25±0.02 mM As(V)
formed d-1 for treatment with YE and Pyr, respectively. Both were higher than those for
ECC1.
295
1.0
As(III) Concentration (mM)
A
0.8
0.6
0.4
0.2
0.0
0
10
20
30
40
1.0
As(V) Concentration (mM)
B
0.8
0.6
0.4
0.2
0.0
0
10
20
Time (days)
30
40
Figure 7.2 The removal of As(III) (panel A) and formation of As(V) (panel B) by ADS
autotrophic enrichment (ECC1)in the presence of chlorate under anoxic condition.
Legend for panels A and B: Full treatment with As(III) + ClO3-, without YE (▲) and
with YE (■); control with only As(III) added, As(III) without YE (∆) and with YE (□),
and the abiotic control with As(III) and NO3- (○).
296
1.2
As(III) Concentration (mM)
A
0.9
0.6
0.3
0.0
0
5
10
15
20
25
30
1.2
As(V) Concentration (mM)
B
0.9
0.6
0.3
0.0
0
5
10
15
Time (days)
20
25
30
Figure 7.3 The removal of As(III) (panel A) and formation of As(V) (panel B) by ADS
heterotrophic enrichment (ECC3) in the presence of chlorate under anoxic condition.
Legend for panels A and B: Full treatment with As(III) + ClO3-, without YE (♦), with YE
(▲) and with Pyr (■); control with only As(III) added (●), and the abiotic control with
As(III) and ClO3- (○).
297
Table 7.3 kinetics† of microbial As(III) oxidation (1.0 mM) under chlorate (1.5 mM) respiring conditions by ECC1
Parameters
ADS autotrophic(ECC1)
ADS heterotrophic (ECC3)
ADS without OC*
ADS with YE
ADS without OC
ADS with YE‡
ADS with Pyr‡
Growth rate† (mM d-1)
0.13±0.03
0.20±0.08
__
0.34±0.08
0.25±0.02
Doubling time (days)
5.53±1.36
3.78±1.48
__
2.11±0.54
2.77±0.23
R2
0.93
0.99
__
0.95
0.92
†
: Estimated from As(V) formation data for the ln(∆As(V)) versus time plots.
‡
: The yeast extract (YE) and pyruvate (Pyr) were added at the concentration of 1.0 and 1.7 mg l-1, respectively.
*: OC=organic carbon.
298
7.4.5. End products of autotrophic and heterotrophic chlorate reduction linked to anoxic
As(III) oxidation.
Tthe possible products of chlorate respiration, including ClO2- and Cl-, were also
monitored to determine the end product of chlorate coupled to anoxic oxidation of
As(III). Figure 4 demonstrates that Cl- was the only end product in the treatment
inoculated with ECC1. And there was no ClO2- observed during the experiment (data not
shown). The formation of Cl- corresponded to 98.08±17.90% of the measured removal of
ClO3-, which also indicating that Cl- was the only product of the chlorate reduction. The
molar ratios of As(III) consumption and As(V) formation to the corrected ClO3consumption (corrected for endogenous ClO3- consumption) and corrected Cl- formation
(corrected for endogenous Cl- formation) are provided in Table 4. The molar ratio of
As:Cl involved in the reaction was calculated from the As(V) formed to corrected ClO3consumption and corrected Cl- formation . The molar ratios of As(V) formation to ClO3consumption and Cl- formation were 2.91±0.48 and 2.98±0.05, respectively, which were
close to the theoretical ratio of 3.0 expected for As(III) oxidation linked to complete
chlorate reduction to Cl- as can be seen in eq.1. The ECC3 also coupled to oxidation of
As(III) to the reduction of chlorate, forming chloride as sole end product (Table 4).
299
0.5
ClO3- consumption (mM)
A
0.4
0.3
0.2
0.1
0.0
0
10
20
30
40
0.5
B
0.3
0.2
-
Cl formation (mM)
0.4
0.1
0.0
0
10
20
Time (day)
30
40
Figure 7.4 Chlorate reduction at a initial aqueous As(III) concentration (1.0 mM). The
consumption of ClO3- (panel A) and the formation of Cl-1 (panel B) by ECC1 under
anoxic conditions. Legend for panels A and B: Full treatment with As(III) and ClO3-,
without YE (▲) and with YE (■); endogenous control with only ClO3- added, without
YE (∆) and with YE (□); and the abiotic control with As(III) and NO3- (○).
300
301
7.4.6. As(III) oxidation by autotrophic As(III) oxidizing chlorate reducing bacteria
utilizing oxygen as sole electron acceptor
Since O2 is formed during the reduction of chlorate, an experiment was conducted to
evaluate if the autotrophic and heterotrophic enrichment cultures could utilize oxygen as
an electron acceptor directly to oxidize As(III) to As(V) in the absence of chlorate. An
example of the time course from these incubations is provided in Figure 5 for ECC1
cultured under autotrophic conditions. The results showed that ECC1 has the ability to
effectively oxidize As(III) to As(V) only in the presence of lower oxygen concentration.
Reactions were only evidenced if the O2 was supplied at 5.0 mmol lliquid-1 or less. The
reaction was completely inhibited when O2 was supplied at 12.5 and 25 mmol lliquid-1. The
O2-dependent As(III) oxidation activity of ECC1 shown in Figure 6 ranged from 0.17 to
0.58 mM As(V) formed d-1, with highest activity of 0.58 mM As(V) formed d-1 at a
optimizing O2 concentration of 2.5 mmol lliquid-1.
302
1.2
As(III) Concentration (mM)
A
0.9
0.6
0.3
0.0
0
5
10
15
20
25
30
35
1.2
As(V) Concentration (mM)
B
0.9
0.6
0.3
0.0
0
5
10
15
20
Time (days)
25
30
35
Figure 7.5 The removal of As(III) (panel A) and formation of As(V) (panel B) by ECC1
in the presence of elemental oxygen. Legend for panels A and B: Full treatment with
As(III) + O2 with YE: O2 concentration at 0.25 mmol Lliquid-1 (▲), 1.25 mmol Lliquid-1 (●),
2.5 mmol Lliquid-1 (♦), 5.0 mmol Lliquid-1 (∆), 12.5 mmol Lliquid-1 (◊) and 25.0 mmol Lliquid-1
(■); control with only As(III) and YE added, (○); abiotic control with As(III) and O2
(1.25 mmol Lliquid-1), (□).
303
As(V) formation rate (mM d-1)
0.6
0.5
0.4
0.3
0.2
0.1
0
0
5
10
15
20
25
O2 concentration (mmol Lliquid-1)
Figure 7.6 Activity of A(III) oxidation in the presence of O2 by ECC1. The activity was
shown as the rate of As(V) formation (♦) in the presence of different O2 concentrations.
The two lines represent the rate of As(III) oxidation without YE (—) and with YE (----)
linked to chlorate reduction by ECC1.
7.4.7. Microbial oxidation of As(III) to As(V) utilizing chlorate as electron acceptor in
bench-scale continuous bioreactor under anoxic conditions
To determine if the process of As(III) oxidation linked to chlorate respiration could be
sustained in a continuous bioreactor, using a 2 L upward-flow anaerobic sludge bed
(UASB) bioreactor was initiated with ADS as inoculum. The UASB operated for a period
of 550 days and the operation was divided into seven periods, which were distinguished
based on the concentration of As and the inclusion of ClO3- as shown in the Table 5
(endogenous means no addition of any electron donor). The microbial oxidation of
304
As(III) and the concomitant formation of As(V) in the bioreactor as a function of
operation time are illustrated in Figure 7. The removal efficiency of As(III) in column
averaged 95.90±4.25% when considering data of an period when As(III) was supplied in
the feed. Also the conversion rate of influent As(III) to effluent As(V) averaged
98.26±3.98% during the same periods of consideration.
In period I (day 0-88), II (day 88-211) III (day 211-278) and IV (day 278-301),
the influent concentration of As(III) was gradually increased from 0.5 to 1.0, 1.75 and 2.0
mM, respectively, while the concentration of chlorate was fed at 5.0 mM (day 0-91) and
3.0 mM (day 92-550). Starting from day 20, the dominant As species in the effluent of
bioreactor was As(V), indicating the occurrence of microbial As(III) oxidation linked to
chlorate respiration under anoxic conditions. The formation of As(V) in the effluent
corresponded to an almost stoichiometric removal of As(III) in the influent, indicating
that As(V) was the main product of the conversion. Figure 8 illustrates the oxidation
efficiency of As(III) to As(V) in bioreactor during the period I, II III and IV, respectively.
As(III) was efficiently oxidized in the chlorate-amended column with removal efficiency
at 91.0±17.6, 95.3±9.1, 98.2±8.9 and 98.5±1.6%, respectively.
305
I
As(III) Concentration (mM)
2.4
II
III
IV
V
VI VII
VIII
1.8
1.2
0.6
0.0
0
100
I
200
II
300
III
IV
400
V
500
VI VII
600
VIII
As(V) Concentration (mM)
2.4
1.8
1.2
0.6
0.0
0
100
200
300
Time (days)
400
500
600
Figure 7.7 The removal of As(III) and the formation of As(V) in the continuous
bioreactor linking the anoxic oxidation of As(III) to chlorate respiration as a function of
time: (●) Influent, (○) Effluent.
306
As concentration (mM)
2.5
2.0
As(III) influent
As(III) effluent
As(V) influent
As(V) effluent
1.5
1.0
0.5
0.0
I
II
III
IV
VI
VII
VII
Operation period
Figure 7.8 The influent and effluent concentrations of As(III) and As(V) for each period
of the total operation in the continuous bioreactor link the anoxic oxidation of As(III) to
chlorate respiration.
In period V (day 301-423), no As(III) was fed to the column. The endogenous
consumption of chlorate and production of Cl- by organic matter naturally present in the
bio-film of the bioreactor was estimated and used to correct the ClO3- consumption and
Cl- production. The corrected ClO3- consumption was attributed to As(III) oxidation, and
used to calculate the ratios of As(III) removal and As(V) formation to ClO3- consumption
or similar ratios made with Cl- production. Results of the endogenous consumption of
chlorate and endogenous production of Cl- are presented by Figure 9.
In period VI (day 423-453), VII (day 453-500) and VIII (day 500-550), the
feeding of As(III) (0.5, 1.0 and 2.0 mM) to bioreactor was gradually reestablished in an
incremental fashion. After 112 d operation without As(III), the anaerobic microbes in the
307
bioreactor readily converted As(III) to As(V) indicating full recovery of population after
the feed interruption. The bioreactor became fully stable after a recovery period of 1
week, an the As(III) removal efficiency of 95.3±1.3% was reestablished during period of
day 432-453, which was similar to the average of periods I to IV of 93.3±4.6%. Table 5
presents a summary of the performance data for the total duration of this study.
308
Table 7.5 Results summary of operation periods for the UASB reactor
Period Days
†
‡
As(III)
As(III)
As(V)
loading rate
removed
formed
(mmol Lr-1 d-1
(mM)
(mM)
As(III)
ClO3
-
removal efficiency (%) consumed (mM)
-
Cl formed
(mM)
Cl- formed
/ClO3- consumed
(mol/mol)
I
0-88
33.8±6.0
0.45±0.08 0.40±0.08
92.69±6.94
0.40±0.09
0.38±0.05
1.00±0.26
II
88-211
81.8±15.0
1.09±0.20 1.04±0.19
97.31±2.91
0.66±0.11
0.56±0.06
0.87±0.18
III
211-278
126.0±10.5
1.68±0.14 1.64±0.11
97.97±1.24
0.83±0.17
0.72±0.07
0.91±0.21
IV
278-301
139.5±3.0
1.86±0.04 1.84±0.03
96.75±0.94
0.76±0.20
0.84±0.06
1.18±0.39
V
301-423
__
__
0.26±0.28†
0.23±0.15†
1.16±0.99
VI
423-453
35.2±1.5
0.47±0.02 0.47±0.02
95.28±1.31
0.42±0.03
0.38±0.02
0.90±0.09
VII 453-500
67.5±11.3
0.90±0.15 0.87±0.13
95.05±0.83
0.58±0.10
0.52±0.04
0.91±0.06
VIII 500-550
144.0±11.3
1.92±0.15 1.95±0.05
92.84±3.59
0.90±0.11
0.85±0.07
0.95±0.07
__
__
: The endogenous consumption of chlorate.
: Corrected for endogenous chlorate consumption measured in period fed with chlorate in the absence of As(III).
309
7.4.8. Chlorate respiration coupled to anoxic oxidation of As(III) to As(V) in bioreactor
Chlorate was fed to bioreactor with the concentrations of 5.0 mM (day 0-104) and 3.0
mM (day 105-550). Because chlorate in the influent was much higher than the chlorate
consumed within day 0-105, Figure 9 shows the chlorate consumption and chloride
production in bioreactor only after day 105 when the chlorate consumption is more
readily visible in the graph. The chlorate concentration supplied (5.0 and 3.0 mM) was in
excess of the concentration required for the stoichiometric conversion of As(III) to As(V)
(0.17, 0.33 and 0.67 mM ClO3- would be the theoretical requirement for 0.5, 1.75 and 2.0
mM As(III), respectively). In period V (Day 301-423), the endogenous consumption of
chlorate was measured as 0.26±0.28 mM. The formation of chloride was 0.23±0.15 with
the recovery ratio at 1.16(±0.99) mol Cl- mol-1 ClO3-. The molar ratio of As(V) formed
compared to ClO3- consumed and Cl- formed involved in the coupled reactions was
calculated from the formation of As(V) (As(V)) and the corrected ClO3- (ClO3-)
consumption and Cl- (Cl-) formation (corrected for the endogenous chlorate
consumption and chloride formation) as shown in Figure 10. For the total operation, the
calculated average molar ratios of As: Cl were 3.06±0.08 and 3.01±0.08 as As(III) or
As(V) :ClO3-, or 3.17±0.15 and 3.03±0.15 as As(III) or As(V) :Cl-, respectively.
These ratios are very close to the theoretical stoichiometry ratio of 3.0 for As(III)
oxidation linked to complete chlorate respiration (eq. (1)). Chlorite (ClO2-), one
intermediate product from the microbial degradation of chlorate, was not detected in the
310
effluent of reactor throughout the experiment. These findings indicate that chlorate was
completely reduced to the benign end product chloride (Cl-).
Chlorate Concentration (mM)
V
IV
VI
VII
VIII
4.0
3.5
3.0
2.5
2.0
100
200
I
1.2
Chloride Concentration (mM)
III
II
4.5
300
II
III
400
IV
500
V
VI VII
600
VIII
0.9
0.6
0.3
0.0
0
100
200
300
Time (days)
400
500
600
Figure 7.9 The removal of ClO3- and the formation of Cl- in the continuous bioreactor
link the anoxic oxidation of As(III) to chlorate respiration as a function of time: (●)
Influent, (○) Effluent.
311
5
Molar ratio
4
3
2
1
0
I
II
III
IV
V
VI
VII
VIII
Operation period
Figure 7.10 The summary of molar ratios of As(III) removed compared to ClO3consumed and Cl- formed, and As(V) formed compared to ClO3- consumed and Clformed in the continuous bioreactor link the anoxic oxidation of As(III) to chlorate
respiration as a function of time. Legends: As(III)/ ClO3- (○), As(III)/ Cl- (□), As(V)/
ClO3- (∆) and As(V)/ Cl- (◊). Theoretical ratio if ClO3- reduced to Cl- and O2 reduced to
H2O linked to As(III) oxidation (—); Theoretical ratio if O2 reduced to H2O linked to
As(III) oxidation (----).
7.4.9. As(III) substrates inhibition to As(III)-oxidizing chlorate-reducing bacteria
An experiment inoculated with As(III)-oxidizing chlorate-respiring bacteria (AOCRB) in
sludge sampled from bioreactor (on day 530) was monitored at various initial As(III)
concentrations to study inhibition of As(III) oxidation by As(III). Sufficient sampling
times were used to evaluate the activity of As(III) oxidation so as to obtain kinetic data
for the lowest As(III) concentration tested of 0.0005 mM. The specific activities
312
measured are shown in Figure 11. The anoxic oxidation of As(III) was inhibited by
incremental As(III) concentrations as illustrated by the decrease in the specific activity.
Compared to the kinetic parameters observed at 0.005 mM As(III), the maximum specific
activities were inhibited by 50% at 0.29 mM and approximately 97% at 5.0 mM As(III).
Albeit, that inhibition was severe, As(III) could be completely biologically transformed
by anoxic oxidation even in the presence of 10 mM, but after a prolonged lag phase of 25
days. The activity inhibition by As(III) at 10 mM was 98% compared to the specific
activity at 0.005 mM As(III).
Rate of As(V) formation
-1
(mM d )
Rate of As(V) formation
-1
(mM d )
5
4
3
2
5
4
3
2
1
0
0
1
0.05
0.1
0.15
0.2
0.25
Initial As(III) concentration (mM)
0
0
2
4
6
8
10
Initial As(III) concentration (mM)
Figure 7.11 Activity and substrate toxicity as shown by a decrease in the rate of As(V)
formation (♦) from the oxidation of As(III) by AOCRB from bioreactor at different initial
As(III) concentrations.
313
7.5. Discussion
7.5.1. Evidence of biological oxidation of As(III) to As(V) linked to chlorate reduction
In this study, the capacity of microorganisms from diverse anaerobic samples of sludge to
utilize chlorate as an electron acceptor for the oxidation of As(III) was illustrated for the
first time. The biological nature of the reaction is inferred from the lack of any
conversion in non-inoculated samples and heat killed samples. In addition, both the
autotrophic and heterotrophic enrichment cultures also linked the anoxic oxidation of
As(III) to chlorate respiration. The bacteria appeared to benefit from the process as
evidenced by the culture being to be sustained after a large number of transfers (23×) in
this study. Likewise, the MPN assays of four enrichments cultures indicate high cell
densities of microorganisms were produced with up to 4.39×1013 cells mol-1 As(V)
formed. Furthermore, the As(III), as the major substrates, showed inhibitory effects on
the activity of AOCRB, which is contrary to the expectation of a chemical reaction and
thus further supporting a bio-catalyzed reaction mechanism.
The biological conversion of As(III) to As(V) in the absence of O2, was
dependent on the presence of ClO3-. No As(III) oxidation occurred in the inoculated
incubations lacking ClO3-. The concentration of As(V) formed in the presence of ClO3corresponded with the concentration of As(III) removed, indicating that As(V) was the
main product of the conversion and there was no significant formation of other As
metabolites. The presence of As(III) greatly enhanced ClO3- consumption beyond the
314
background endogenous consumption, suggesting its role as an electron donor to the
microbial reaction. The production of Cl- in the incubations was linked to the addition of
As(III) to the cultures and the yield of Cl- corresponded to the expected stoichiometry
from the electron equivalents in As(III). In parallel the production of Cl- was nearly
100% recovered from ClO3- consumption without any accumulation of other chlorine
containing products. The stoichiometry of the As(III) conversion to As(V) in relation to
ClO3- conversion to Cl- clearly suggests As(III) oxidation was linked to the complete
chlorate reduction to Cl- as indicated in eq. (1). The experimentally measured ratios
approximated 3.0 mol As metabolized per mol ClO3- converted as is expected from the
eq. (1). These findings were proved again in the continuous bioreactor study. During the
whole operation, the molar ratios of As(III) removal or As(V) formation to ClO3consumption and Cl- formation were close to the theoretical ratio of 3.0 with various
As(III) concentration in the influent. If the As(III) oxidation was linked to the oxygen
generated from ClO3- reduction, the ratio of As(III) removed to ClO3- consumption would
have been approximately 2.0. Thus, these results confirm that oxidation of As(III) to
As(V) was successfully linked to the complete chlorate reduction process, in which
chlorate was reduced to benign products Cl-.
There are reported precedents from As contaminated sites for the enrichment and
isolation of nitrate-dependent [Oremland et al., 2002; Rhine et al., 2006] or selenatedependent [Fisher and Hollibaugh, 2008] As(III)-oxidizing bacterial capable of utilizing
nitrate or selenate as electron acceptor in anaerobic environments. Our previous study
reported that diverse anaerobic microorganisms were involved in the anoxic oxidation of
315
As(III) linked to denitrification in from sludges and sediments with no prior exposure to
As [Sun et al., 2008]. As far as we know, this is the first time that chlorate has been
reported toserve as an electron acceptor for the biological oxidation of As(III) to As(V)
under anoxic conditions. The microbial activity in this study was demonstrated in pristine
samples without exposure to As contamination.
7.5.2. The occurrence of As(III) oxidizing bacteria utilizing chlorate as electron acceptor
in anoxic environments
Perchlorate is mainly used as an oxidant in aerospace and defense industries that led to its
introduction into the environment [Cheremisinoff et al., 2001]. Chlorate entering the
environment ascribed to the production of chlorine dioxide and hypochlorite [Van Ginkel
et al., 1995; Rikken et al., 1996]. Microorganisms with the capacity of gaining energy
from the oxidation of inorganic compounds are widespread in the anoxic environments.
(Per)chlorate is a common contaminants found in groundwater and lakes that could
potentially support these reactions [Motzer et al., 2001]. It has been well known that H2
[Zhang et al., 2002], Fe0 [Son et al., 2006; Yu et al., 2007], Fe(II) [Achenbach et al.,
2001; Bruce et al., 1999], H2S [Achenbach et al., 2001], and S0 [Ju et al., 2007] served as
electron donors for perchlorate respiring bacteria (PRB); H2S and H2 [Van Ginkel et al.,
1995], as well as Fe0 [Son et al., 2003] and Fe(II) [Bruce et al., 1999] are also utilized by
316
chlorate respiring bacteria (CRB). Energy for microbial growth can also be gained from
the anoxic oxidation of As(III).
With respect to As, the environmental samples used as inoculum in this study can
be regarded as pristine samples [Sun et al., 2008]. Nonetheless, the microbial
communities in the inoculated samples demonstrated moderate biological activities of
As(III) oxidation with doubling times ranging from 2.62 to 5.45 d. Under the conditions
of co-occurring As(III) and ClO3-, responsible organisms grew and established a large
enough population to account for significant As(III) activity in the microcosms, ranging
from 0.13 to 0.39 mM As(V) formed d-1. The occurrence of the anoxic As(III) oxidizers
in the As pristine anaerobic samples is probably due to the metabolic versatility of PRB
[Bender et al., 2005; Coats et al., 1999] or CRB [Wolterink et al., 2003; Steinberg et al.,
2005], which are well known for utilizing diverse substrates.
7.5.3. The characteristic of anoxic As(III) oxidizing chlorate-respiring bacteria
The As(III) oxidizer in the ECC1 are chemolithotrophic, obtaining energy from As(III)
and most likely obtaining carbon from added HCO3-/CO2. The autotrophic ECC1 was
sustained without addition of OC, but small additions of OC (YE) did stimulate activity
slightly although with marginal statistical significance. It seems there are some
experiments when YE had no effect and others where it did in Table 3. The difference is
not statistically significant. And heterotrophic ECC3 was also maintained from the
317
incubation amended with YE. The As(III) oxidizing bacteria in the ECC2 showed strict
obligated heterotrophy evidenced by no activity for the incubation without organic
carbon.
Oxygen is produced as an intermediate during chlorate reduction. However, this
oxygen is rapidly consumed for cell respiration so that it does not accumulate in solution
to high levels [Van Ginke et al., 1996; Coats et al., 1998]]. Previous studies have
illustrated that biological oxidation of As(III) was very fast and efficient in the presence
of oxygen [Santini, et al., 2000]. Our results showed that supplied with low O2
concentrations (0.25 to 5.0 mmol lliquid-1) as sole electron acceptor, AOCRB from
autotrophic ECC1 could utilize the low levels of O2 to oxidize As(III) to As(V). In
contrast, exposure to higher O2 levels (12.5 to 25.0 mmol lliquid-1) caused inhibition on the
activity of the aerobic As(III) oxidation by AOCRB enriched. Thus, it is possible that the
AOCRB are microaerophilic microorganisms with the ability to oxidize As(III) as
electron donor in the presence of chlorate or oxygen. Wallace et al [1998] already
reported that HAP-1, a PRB strain, was a microaerobic organism, although no further
details on oxygen tolerance were provided. It has been well known that high
concentrations of dissolved oxygen inhibit the activity of (per)chlorate reduction by PRB
or CRB [Song et al., 2004]. Chlorate reductase was also sensitive to elemental oxygen
exposure [Wolterink et al., 2003]. Although chlorite dismutase activity is not inhibited by
oxygen, expression of this enzyme is not observed in CRB under aerobic conditions even
in the presence of chlorate [Bender et al., 2005].
318
7.5.4. Anoxic oxidation of As(III) linked to chemolithotrophic chlorate reduction in
continuous bench-scale UASB bioreactor
Several bioreactor technologies have been utilized for drinking water treatment and
bioremediation of perchlorate contaminated groundwater and surface water [Xu et al.,
2003]. Due to the concerns about the bacteria growth and disinfection byproducts in the
heterotrophic treatment process there has been some interest for exploring,
chemolithotrophic (per)chlorate reducing bioreactor that rely on inorganic chemicals as
electron donors instead of organics. Biological reduction of perchlorate was achieved
with excellent removal efficiency (> 99%) by autotrophic microorganisms attached to
zero valent iron in flow-through columns [Yu et al., 2007]. Highly effective removal of
perchlorate from ground water has been demonstrated with autotrophic hydrogenutilizing bacteria in a packed-bed bioreactor and membrane hollow fiber reactor [Giblin
et al., 2002; Logan and LaPoint, 2002]. Thus, the chemolithotrophic chlorate-respiring
bioreactors have the potential to utilize As(III) as electron donor and enhance the As
removal forming a more strongly absorbable species, As(V).
This study demonstrated that AOCRB immobilized in bio-films of a continuous
bioreactor effectively linked As(III) oxidation with chlorate reduction under anoxic
conditions. The concentration of As(III) was incremented slowly from 0.5 to 2.0 mM, the
bioreactor in the presence of chlorate showed a reliable As(III) removal efficiency
average 96%. Indeed, the bioconversion of As(III) to As(V) was also as high as 97%
319
during the whole operation. Furthermore, after a four monthly As(III) feeding
interruption, AOCRB in the bioreactor to recover the ability to oxidize As(III) to As(V).
Based on the mass balance of As and chloride speciation in Table 5, the results from this
study confirm that chemolithotrophic chlorate reduction to benign product Cl-1 was
responsible for the As(III) oxidation to As(V).
7.5.5. As(III) substrates inhibition on chemolithotrophic As(III)-oxidizing chloratereducing bacteria
It is well known that As is toxic compound to most microorganisms. The 50% inhibitory
concentration of As(III) for AOCRB described in this study was measured at 0.29 mM.
And a concentration of 1.2 mM As(III) started showing 85% inhibition on AOCRB,
compared to the activity observed at 0.005 mM As(III). As(III) at 3.5 mM demonstrated
the severe inhibition on the As(III) oxidation rates linked to denitrification, compared to
the rates observed at 0.5 mM [Sun et al., 2008]. The 50% inhibitory concentrations of
As(III) to methanogenic activity are reported to be as low as 15 µM [Sierra-Alvarez et al.,
2004]. As(III), entering the cell through aqua-glycerolporins, and binding to thiols or
vicinal sulfhydryl groups, can inactivate/denature the normal functions of many proteins
and enzymes, and cause cell damage [Mukhopahyay et al., 2002].
Despite the high level of inhibition, AOCRB in this study have the potentials to
oxidize As(III) at high concentrations up to the highest tested of 10.0 mM. The As(III)
was metabolized at a slow rate. In our previous study, mixed consortia from sludge and
320
sediments demonstrated the capacity to oxidize As(III) up to 3.5 mM linked to
denitrification [Sun et al., 2008]. The microbial population adapted to high influent
concentrations of As(III) up to 5.0 mM in the continuous denitrifying bioreactor with
responsible bacteria immobilized in bio-film granules [Sun et al., 2008, IV]. The previous
attempts to isolate anoxic As(III)-oxidizing bacteria have been from sites containing high
concentrations of As. For example, the soda lake from where Alkalilimnicola ehrlichii
strain MLHE-1 was isolated contains 0.2 mM As [Oremland et al., 2002]. The initial
enrichments leading to the isolation of Azoarcus sp. strain DAO1 and Alkalilimnicola
ehrlichii strain MLHE-1 contained 5 mM As(III) [Oremland et al., 2002; Rhine et al.,
2006]. The As-resistant bacteria have evolved several detoxicification mechanisms using
the plasmid- or chromosomally-encoded ars operon [Silver and Phung, 2005].
Is the toxic inhibition of As(III) on microorganisms an serious issue in
environment? Probably it applied at the anthropogenic contaminated sites with high
concentrated As. But maybe not for the natural environment with As contamination. The
environmental As concentrations in the subsurface are typically well below the toxicity
range to microorganisms [Cullen et al., 1989; Smedley and Kinniburgh, 2002], and the
soluble As concentration also could reach to the equilibrium with the solid phase
sorbents. Therefore, microorganisms would thrive at low As(III) concentrations in wide
environments and influence the speciation of As in the environment.
321
7.5.6. Bioremediation potential
Chlorate has been proposed as an alternative electron acceptor for the bioremediation of
hydrocarbon contaminated plume [Logan and Wu, 2002]. We have shown that it is also
used by anoxic As(III) oxidizer. Thus, chlorate injection into groundwater could be
considered as a means of remediation of As. Change in speciation from As(III) to As(V)
improves immobilization of As on clay minerals containing aluminum, non-iron oxides
[Dixit and Hering, 2003]. Chlorate also has potential to oxidize aqueous reduced Fe(II)
forming iron (hydr)oxides, such as ferrihydrite [Weber, et al., 2006], which could act as
sorbents for As [Raven et al., 1998; Dixit and Hering, 2003].
7.6. Acknowledgments
The work presented here was funded by a USGS, National Institute for Water Resources
104G grant (2005AZ114G), and by a grant of the NIEHS-supported Superfund Basic
Research Program (NIH ES-04940). The use of trade, product, or firm names in this
report is for descriptive purposes only and does not constitute endorsement by the U.S.
Geological Survey. Also arsenic analyses were performed by the Analytical Section of
the Hazard Identification Core (Superfund Basic Research Program grant NIEHS-04940).
322
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CHAPTER 8
CONCLUSIONS
8.1. Microbial oxidation of As(III) to As(V) in anoxic environments
8.1.1. Anoxic oxidation of As(III) to As(V) is linked to chemolithotrophic denitrification
in sludges and sediments.
a) Microorganisms capable of linking anoxic As(III) oxidation to denitrification are
widespread in anaerobic sediments and sludges, including samples not previously
exposed to As contamination.
b) The doubling times for growth of the anoxic As(III) oxidizers range from 0.74 to
1.34 d.
c) The experimentally measured ratios approximated 2.5 mol arsenic metabolized
mol-1 NO3- converted indicates that complete denitrification of nitrate to N2 was
linked to As(III) oxidation.
d) The anoxic oxidation of As(III) linked to denitrification is inhibited by As(III)
with 5 mM causing complete inhibition. However, As(III) weakly adsorbed to AA
or TiO2 is available as an electron donor for denitrification while the equilibrium
concentrations are low enough to minimize toxicity.
330
e) Dinitrogen gas is the end product of denitrification linked to As(III) oxidation if
As(III) is not present at inhibiting concentrations. Nitrous oxide accumulates as a
major product at inhibitory As(III) concentrations.
f) The biodiversity of anoxic As(III) oxidizers is potentially greater than the few
arsenic resistant strains reported previously. In this study, it was shown that
tolerance to As(III) was not a prerequisite for anoxic As(III) oxidation by
anaerobic microbial communities.
g) The use of ubiquitous nitrate as an electron acceptor may be an important missing
link the biogeochemical cycling of arsenic between two common inorganic
species, As(III) and As(V), where dissolved oxygen (DO) is absent.
8.1.2. Molecular characterization and in situ quantification of anoxic arsenite oxidizing
denitrifying enrichment cultures
a) Clone library and FISH analysis suggest that phylotypes in the genus Azoarcus
and in the family Comamonadaceae are the main As(III) oxidizers in the
microbial communities of the three enrichment cultures derived from sludge,
sediment and biofilms inocula.
b) The involvement of Azoarcus t is supported by the fact that a related isolate from
this genus (DAO1) can link As(III) oxidation to denitrification (Rhine, et al.,
331
2006). Furthermore, a closely related isolate EbN1 contains arsenic resistance
genes as well as a full set of denitrification genes.
c) This study provides the first clues for the possible involvement of members of the
Comamonadaceae family in anoxic As(III) oxidation.
d) The fact that As(III) was oxidized in the MC (which contained no Azoarcus)
implies that the dominate phylotype remaining (related to Diaphorobacter in the
Comamonadaceae familty) may be responsible for anoxic As(III) oxidation in the
MC. The occurrence of an aerobic As(III) oxidizers in the Comamonadaceae
familty (Fan, et al., 2008), together with the occurrence of arsenic resistance- and
denitrification genes in the genome of the closely related Acidovorax sp. strain
JS42, supports this potential role.
8.1.3. Anoxic oxidation of arsenite linked to chemolithotrophic denitrification in
continuous bioreactors
a) This study demonstrated for the first time that continuous denitrifying bioreactors
with bacteria immobilized in biofilm was able to readily and effectively link
As(III) oxidation in the presence of nitrate over prolonged periods of operation;
however, negligible oxidation of As(III) was observed in a control bioreactor
lacking nitrate.
332
b) As(III) feeding interruptions in the denitrifying bioreactors (R1 and R3)
demonstrate that nitrate removal was highly dependent on the presence of As(III).
c) Batch and continuous studies confirm beyond doubt that complete denitrification
of nitrate to benign product of N2 was successfully linked to oxidation of As(III)
to As(V) under anoxic conditions.
d) The microorganisms in the continuous bench-scale upward-flow anaerobic sludge
bed (UASB) bioreactor (R3) acclimated to higher concentrations of As and
tolerated up to 5 mM As(III) in the bioreactor influent.
e) For laboratory-scale denitrifying bioreactor (R1) in the presence of nitrate, the
volumetric loadings of As(III) ranged from 179 to 259 mg As Lr-1 d-1 during the
whole operation, where Lr refers to the empty bed volume of the reactors. The
removal efficiency of As(III) was average at 93.77±7.55% throughout the length
of the experiment. For the whole operation, the calculated molar ratios of As(V):
NO3- was average at 2.49±0.17. For the bench-scale UASB bioreactor (R3), the
removal of As(III) from the influent was very stable averaging 96.45±6.59% for
the whole experiment. The molar ratios of As(III) removed and As(V) formed
compared to nitrate consumed (corrected for background removal by the
endogenous substrate) was found to average 2.62±0.24 and 2.71±0.21,
respectively for the whole operation.
333
8.1.4. Novel Study of the Anoxic Oxidation of As(III) to As(V) linked to Chlorate
Reduction
a) In this study, the biological oxidation of As(III) to As(V) linked to chlorate
reduction was demonstrated for the first time. The findings presented here
demonstrate a widespread capacity for this reaction by diverse microorganism in
the environment. And chlorate was completely reduced to benign end product Cl-.
b) The As(III)-oxidizing chlorate-respiring bacteria (AOCRB) were enriched by
either autotrophic enrichment, which showed modest activity stimulation with the
small additions of organic carbon, or heterotrophic enrichments with strict
obligate heterotrophy.
c) The maximum specific activities of AOCRB ranged from 0.13 to 0.39 mmol
As(V) formed g-1 VSS d-1.
d) The AOCRB demonstrated the ability to oxidize As(III) utilizing oxygen as sole
electron acceptor.
e) The As(III) oxidation activity of AOCRB in the presence of oxygen ranged from
0.17 to 0.58 mM As(V) formed d-1, optimizing O2 at a concentration of 2.5 mmol
lliquid-1; however, completely inhibited when incubated with O2 at the
concentration of 12.5 mmol lliquid-1.
334
f) A 2-Liter bench-scale upward-flow anaerobic sludge bed (UASB) reactor was
operated with inoculum of an anaerobically digested sewage sludge. The
bioreactor was fed with basal mineral medium and As(III) as the main energy
source, chlorate as the sole electron acceptor and NaHCO3 as the major carbon
source. Anoxic oxidation of As(III) linked to chemolithotrophic chlorate
reduction was shown to be effective with more than 95% As(III) removal in this
continuous bioreactors over 550 days operation.
g) Injecting chlorate into anaerobic As plume as an electron acceptor may be a
potential bioremediation technology by converting As(III) to As(V) in subsurface
environments. As(V) is considered to be more strongly adsorbed than As(III) on
certain common metal oxides such as those of aluminum
8.2. The role of denitrification on arsenite oxidation and arsenic mobility in anoxic
dediment column model with activated aluminum
a) Activated aluminum (AA) exhibited higher adsorption capacity for As(V) than
As(III) under circumneutral conditions.
b) Two sediment columns packed with AA and inoculated with chemolithotrophic
denitrifying sludge were operated in a climate controlled dark room at 30±2°C.
The full-treatment reactor (S1) was the biologically active column inoculated with
335
the chemolithotrophic As(III)-oxidizing denitrifying biofilm and fed with basal
medium, As(III) (0.5 mg l-1) and nitrate (35 mg l-1 as NO3--N). The control reactor
(S2) was inoculated and fed As(III) same as S1, but lacked nitrate. Bicarbonate
(100 mg l-1) was supplied as the major carbon source, except that 18.8 mg l-1
acetate was added to consume DO entering from the medium.
c) Biological nitrate-dependent oxidation of As(III) to As(V) enhanced the removal
of arsenic on AA.
d) During the 250 days operation of the continuous experiment, only 25% of all As
in the influent had broken through the column S1 fed with As(III) together with
nitrate. This was considerably less than the 50% release in the control column S2
lacking nitrate.
e) The solid-phase As extraction results revealed that As(V) was the dominant
arsenic species retained on AA surfaces in the presence of nitrate; whereas As(III)
was predominant from in the absence of nitrate.
f) The concentrations of arsenic in the groundwater and subsurface water could be
attenuated by adsorption to Al oxides, attributed to anoxic oxidation of As(III)
linked to denitrification.
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8.3. Anoxic oxidation of As(III) and Fe(II) linked to chemolithotrophic
denitrification for the immobilization of As in anoxic environments.
a) In this study, nitrate-dependent biogenic iron(III) oxides coated sands (ICS)
showed stronger affinity on As(V) than As(III) at neutral pH conditions.
b) Continuous flow sand-packed columns were inoculated with 2.94 g VSS l-1
anaerobic sludge obtained from the a bench-scale As(III)-oxidizing denitrifying
bioreactor R3. The full-treatment reactor (SF1) was biologically active column
fed with basal medium, As(III) (0.5 mg l-1) and Fe(II) (20 mg l-1 Fe supplied as
FeCl2) as electron donating substrates, nitrate (35 mg l-1 NO3- supplied as KNO3-)
and bicarbonate as the major carbon source, except that 18.8 mg l-1 acetate was
added to consume traces of DO that could potentially enter from medium and
cause oxidation of As(III). The control reactor (SF2) was the same as SF1, but
lacked nitrate in the medium.
c) As(III) and Fe(II) were efficiently removed in the biological column SF1 in the
presence of nitrate; in contrast, limited removal of As(III) and Fe(II) occurred in
the control column SF2 lacking nitrate.
d) Denitrification was shown to be responsible for the oxidation of As(III) and
Fe(II); oxidized forms of iron and arsenic were demonstrated by extraction of
sand,
and
confirmed
by
spectrometric
techniques,
Energy
Dispersive
337
Spectroscopy (EDS), X-ray diffraction spectroscopy (XRD) and Energy
Dispersive Spectroscopy (XPS) . The data all indicate that Fe(III) (hydr)oxides
formed by denitrification effectively coated the sand and the resulting ICS
adsorbed As(V) in the biological column.
e) Thus study validates a bioremediation strategy based on injecting NO3- to support
the anoxic oxidation of co-occurring Fe(II) and As(III) in the subsurface. As(V) is
attenuated by becoming adsorbed on biogenic iron(III) hydroxides coating sand
particles and both products are formed from the anoxic microbial oxidation of
As(III) and Fe(II), respectively.
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