EXPLORATION OF BIOLOGICAL TREATMENT SYSTEMS FOR THE REMOVAL

EXPLORATION OF BIOLOGICAL TREATMENT SYSTEMS FOR THE REMOVAL
EXPLORATION OF BIOLOGICAL TREATMENT SYSTEMS FOR THE REMOVAL
OF PERSISTENT LANDFILL LEACHATE CONTAMINANTS AND
NANOPARTICLES
by
Francisco Gomez-Rivera
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
2011
2
THE UNIVERSITY OF ARIZONA
GRADUATE COLLEGE
As members of the Dissertation Committee, we certify that we have read the dissertation
prepared by Francisco Gomez-Rivera
entitled Exploration of Biological Treatment Systems for the Removal of Persistent
Landfill Leachate Contaminants and Nanoparticles
and recommended that it be accepted as fulfilling the dissertation requirement for the
Degree of Doctor of Philosophy
_________________________________________________Date: 11/24/10
James A. Field
_________________________________________________Date: 11/24/10
Maria Reyes Sierra Alvarez
_________________________________________________Date: 11/24/10
Raina M. Maier
_________________________________________________Date: 11/24/10
Robert G. Arnold
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
recommended that it be accepted as fulfilling the dissertation requirement.
_________________________________________________Date: 11/24/10
Dissertation Director: 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 head of the major department or the Dean of the Graduate College when in
his or her judgment the proposed use of the material is in the interests of scholarship. In
all other instances, however, permission must be obtained from the author.
SIGNED: Francisco Gomez-Rivera
4
ACKNOWLEDGEMENTS
I would like to take this opportunity to express my deepest gratitude and appreciation to
my advisors, Dr. James A. Field and Dr. Maria Reyes Sierra Alvarez, for their excellent
guidance, assistance, and invaluable support. My advisors provided me with the tools and
knowledge required to successfully complete my degree. I would like to thank as well to
my Committee members, Dr. Raina M. Maier and Dr. Robert G. Arnold, for their time
and expertise to improve my work.
I am heartily thankful to all the people that in any way assisted me and supported me
while pursuing my degree; especially to those who directly helped me in effectively
achieving my goals, including professors, technicians, administrative staff, and graduate
and undergraduate students.
I would like to thank to the Consejo Nacional de Ciencia y Tecnologia (CONACyT) for
providing the financial support. The labor that CONACyT does is remarkable, giving the
opportunity to numerous students to study an advanced degree in renown Universities.
I am especially grateful to Aida Tapia-Rodriguez. Her company and understanding is the
motor that keeps me going. Finally I would like to deeply thank my family. I would not
have gotten this far without their unconditional trust and encouragement.
5
TABLE OF CONTENTS
LIST OF FIGURES .......................................................................................................... 11
LIST OF TABLES ............................................................................................................ 16
ABSTRACT ...................................................................................................................... 17
OBJECTIVES ................................................................................................................... 19
CHAPTER 1: INTRODUCTION ..................................................................................... 21
1.1 Hydrocarbons ...................................................................................................... 21
1.1.1 Aromatic hydrocarbons ................................................................................. 22
1.1.2 Phenols .......................................................................................................... 25
1.1.3 Chlorinated hydrocarbons ............................................................................. 26
1.2 Hydrocarbons in landfill leachates ...................................................................... 28
1.2.1 Characterization of landfill leachates ............................................................ 28
1.2.2 Physicochemical technologies for treating landfill leachates........................ 29
1.2.3 Physicochemical processes for groundwater treatment................................. 30
1.2.4 Bioremediation. An alternative for groundwater and landfill leachate
treatment ................................................................................................................. 33
6
TABLE OF CONTENTS – Continued
1.3 Removal of emerging contaminants during wastewater treatment ..................... 35
1.3.1 Nanoparticles ................................................................................................. 36
1.3.2 Health effects and environmental fate of nanoparticles ................................ 39
1.3.3 Surface chemistry of nanoparticles ............................................................... 41
1.4 References ........................................................................................................... 46
CHAPTER 2: ANOXIC OXIDATION OF TOLUENE, BENZENE, M-XYLENE, AND
CIS-DCE IN THE PRESENCE OF BIOLOGICAL SLUDGE ........................................ 59
2.1 Abstract ............................................................................................................... 59
2.2 Introduction ......................................................................................................... 61
2.3 Material and Methods .......................................................................................... 66
2.3.1 Inocula sources .............................................................................................. 66
2.3.2 Microcosm preparation.................................................................................. 67
2.3.3 Respikes and transfers ................................................................................... 70
2.3.4 Analytical procedure ..................................................................................... 71
2.4 Results ................................................................................................................. 73
2.4.1 Toluene degradation under nitrate reducing conditions ................................ 73
2.4.2 Toluene degradation under chlorate reducing conditions ............................. 75
2.4.3 Transferring of denitrifying cultures ............................................................. 76
2.4.4 Transferring of methanogenic cultures.......................................................... 79
2.4.5 Nitrate consumption and nitrite formation linked to toluene degradation ... 81
2.4.6 m-Xylene degradation under nitrate reducing conditions ............................. 84
2.4.7 Benzene degradation under nitrate and chlorate reducing conditions ........... 85
2.4.8 Experiments summary ................................................................................... 87
7
TABLE OF CONTENTS – Continued
2.5 Discussion ........................................................................................................... 89
2.5.1 Toluene degradation ...................................................................................... 89
2.5.2 Transfer of microorganisms .......................................................................... 91
2.5.3 Nitrate consumption ...................................................................................... 93
2.5.4 m-Xylene ....................................................................................................... 94
2.5.5 Benzene ......................................................................................................... 95
2.5.6 cis-DCE ......................................................................................................... 96
2.6 Conclusions ......................................................................................................... 97
2.7 References ........................................................................................................... 98
CHAPTER 3: BIOLOGICAL TREATMENT OF A SYNTHETIC LANDFILL
LEACHATE IN A COMBINED ANAEROBIC – AEROBIC SYSTEM ..................... 106
3.1 Abstract ............................................................................................................. 106
3.2 Introduction ....................................................................................................... 108
3.3 Material and Methods ........................................................................................ 113
3.3.1 Anaerobic reactor ........................................................................................ 113
3.3.2 Aerobic reactor ............................................................................................ 115
3.3.3 Synthetic leachate ........................................................................................ 116
3.3.4 Analytical procedure ................................................................................... 117
3.4 Results ............................................................................................................... 120
3.4.1 UASB reactor .............................................................................................. 120
3.4.2 DHS reactor ................................................................................................. 128
8
TABLE OF CONTENTS – Continued
3.5 Discussion ......................................................................................................... 131
3.5.1 UASB reactor .............................................................................................. 131
3.5.2 DHS reactor ................................................................................................. 134
3.6 Conclusions ....................................................................................................... 136
3.7 References ......................................................................................................... 137
CHAPTER 4: FATE OF CERIUM OXIDE (CeO2) NANOPARTICLES DURING
MUNICIPAL WASTEWATER TREATMENT ............................................................ 146
4.1 Abstract ............................................................................................................. 146
4.2 Introduction ....................................................................................................... 147
4.3 Material and Methods ........................................................................................ 151
4.3.1 CeO2 stability in aqueous dispersions ......................................................... 151
4.3.2 Laboratory-scale secondary treatment......................................................... 153
4.3.3 Batch adsorption experiments ..................................................................... 156
4.3.4 Analytical procedure ................................................................................... 157
4.3.5 Electron microscopy .................................................................................... 158
4.3.6 Organic content of the wastewater and biomass concentration .................. 159
4.4 Results ............................................................................................................... 160
4.4.1 Stability of CeO2 NPs .................................................................................. 160
4.4.2 Removal of CeO2 NPs from synthetic wastewater during activated sludge
treatment ............................................................................................................... 165
4.4.3 Fate of CeO2 NPs during activated sludge treatment fed with municipal
wastewater ............................................................................................................ 166
9
TABLE OF CONTENTS – Continued
4.4.4 Contribution of biomass to CeO2 removal during activated sludge secondary
treatment ............................................................................................................... 171
4.4.5 COD removal during secondary treatment .................................................. 173
4.5 Discussion ......................................................................................................... 177
4.5.1 Fate of CeO2 during wastewater treatment.................................................. 177
4.5.2 Mechanisms for CeO2 removal ................................................................... 179
4.5.3 Impact of CeO2 on the wastewater treatment process ................................. 182
4.6 Conclusions ....................................................................................................... 183
4.7 References ......................................................................................................... 183
CHAPTER 5: REMOVAL OF Al2O3 NANOPARTICLES DURING ACTIVATED
SLUDGE SECONDARY TREATMENT ...................................................................... 189
5.1 Abstract ............................................................................................................. 189
5.2 Introduction ....................................................................................................... 191
5.3 Material and Methods ........................................................................................ 194
5.3.1 Stability of Al2O3 NPs in aqueous dispersions ............................................ 194
5.3.2 Laboratory-scale secondary treatment......................................................... 195
5.3.3 Analytical procedure ................................................................................... 198
5.3.4 Electron microscopy .................................................................................... 201
5.4 Results ............................................................................................................... 202
5.4.1 Stability of Al2O3 NPs ................................................................................. 202
5.4.2 Removal of Al3O2 NPs from real domestic wastewater during activated
sludge treatment ................................................................................................... 206
5.4.3 Fate of the organic content in the secondary treatment ............................... 210
10
TABLE OF CONTENTS – Continued
5.5 Discussion ......................................................................................................... 215
5.5.1 Fate of Al2O3 during wastewater treatment ................................................. 215
5.5.2 Mechanisms of Al2O3 removal .................................................................... 217
5.5.3 Impact of Al2O3 on the performance of the activated sludge treatment ...... 219
5.6 Conclusions ....................................................................................................... 220
5.7 References ......................................................................................................... 221
CONCLUSIONS............................................................................................................. 227
REFERENCES ............................................................................................................... 231
11
LIST OF FIGURES
FIGURE 1.1, Structural formula of a) benzene, b) toluene, c) p-xylene, d) phenol,
c) p-cresol, and f) cis-dichloroethylene ............................................................................ 24
FIGURE 1.2, Schematic diagram showing the electrical double layer at the surface of a
particle............................................................................................................................... 43
FIGURE 1.3, Schematic showing the dependence of the zetal-potential on the surface of
the particle on the pH of the media ................................................................................... 45
FIGURE 2.1, Toluene degradation under nitrate reducing conditions using Nedalco
sludge. ─♦─, Abiotic control; ─■─, No NO3-; ─▲─, NO3-; ─●─, Poisoned. Error bars
represent standard deviation of duplicate assays .............................................................. 74
FIGURE 2.2, Evidence of toluene degradation under methanogenic conditions using
Nedalco sludge. ─♦─, Abiotic control; ─■─, No ClO3-; ─▲─, ClO3-; ─●─, Poisoned.
Error bars represent standard deviation of duplicate assays ............................................. 77
FIGURE 2.3, Transfer of granular sludge containing toluene degrading microorganisms
(Nedalco sludge); intact vs crushed granules. ─♦─, Abiotic control; ─■─, No NO3(crushed); ─▲─, NO3- (crushed); ─Δ─, NO3- (intact); ─●─, Poisoned (crushed). Error
bars represent standard deviation of duplicate assays....................................................... 78
FIGURE 2.4, Transfer of toluene degrading microorganisms (Nedalco sludge. 10% of
total crushed granules previously adapted to methanogenic conditions). ─♦─, Abiotic
control; ─■─, No NO3-; ─▲─, NO3-; ─●─, Poisoned. Error bars represent standard
deviation of duplicate assays ............................................................................................ 80
FIGURE 2.5, Consumption of a) NO3- and formation of b) NO2- during the anaerobic
degradation of toluene. NO3- spiked on day 16. Nedalco sludge. ─♦─, Abiotic control;
─■─, No NO3-; ─▲─, NO3-; ─●─, Poisoned. Error bars represent standard deviation of
duplicate assays ................................................................................................................. 82
FIGURE 2.6, Anaerobic degradation of m-xylene under methanogenic conditions. Pond
sediments (Agua Caliente park). ─♦─, Abiotic control; ─■─, No NO3-; ─▲─, NO3-;
─●─, Poisoned. Error bars represent standard deviation of duplicate assays, except for
─■─ .................................................................................................................................. 85
12
LIST OF FIGURES - Continued
FIGURE 2.7, Benzene degradation under nitrate reducing conditions. Nedalco sludge.
─♦─, Abiotic control; ─■─, No NO3-; ─▲─, NO3-; ─●─, Poisoned. Error bars represent
standard deviation of duplicate assays, except for ─▲─ ................................................. 87
FIGURE 2.8, Benzene removal using chlorate as terminal electron acceptor. Anaerobic
digested sludge. ─♦─, Abiotic control; ─■─, No ClO3-; ─▲─, ClO3-; ─●─, Poisoned.
Error bars represent standard deviation of duplicate assays, except for ─▲─ ................ 88
FIGURE 3.1, Diagram of the upflow anaerobic sludge blanket reactor. (1) Influent
leachate, (2) sludge blanket, (3) gas effluent, (4) treated effluent .................................. 114
FIGURE 3.2, Schematic diagram of the upflow hanging sponge reactor used in the
experiments. (1) Pre-treated leachate influent, (2) humidified air influent, (3) shower, (4)
polyurethane sponge, (5) humidified air exit, (6) recirculation, (7) treated effluent ...... 116
FIGURE 3.3, Fate of COD during anaerobic treatment in the UASB reactor, where (■)
represents the influent and (●) the effluent ..................................................................... 122
FIGURE 3.4, CH4 generation linked to the anaerobic degradation of the organic
components in the synthetic leachate during anaerobic treatment in the UASB reactor
during periods I and II..................................................................................................... 123
FIGURE 3.5, VFA concentration in the influent (■) and effluent (●) of the UASB reactor
as function of operation time .......................................................................................... 124
FIGURE 3.6, Influent and (■) and effluent (●) concentration of a) phenol and b) p-cresol
in the UASB reactor as function of time ......................................................................... 127
FIGURE 3.7, Fate of NH4+ during anaerobic treatment in the UASB reactor as function
of time, where the influent is indicated by (■) and the effluent by (●) .......................... 128
FIGURE 3.8, Fate of a) NH4+-N, b) NO3--N and c) NO2--N in the DHS reactor where the
influent and effluent concentrations are indicated by (■) and (●), respectively ............. 130
FIGURE 4.1, Transmission electron microscope image of nano-sized ceria with an
average particle size 50 nm ............................................................................................. 153
13
LIST OF FIGURES - Continued
FIGURE 4.2, Schematic of the laboratory-scale secondary treatment system used.
(1) Aeration tank, (2) air, (3) settler, (4) effluent, (5) recycled sludge ........................... 155
FIGURE 4.3, a) Particle size distribution and b) zeta-potential of CeO2 in different
matrices at t=0 (■) and t=24 h (■ ) incubation in MQ = Milli-Q water (pH 3.11 and 7.06);
SWW = Synthetic wastewater (pH 7.10); WW = Real wastewater (pH 7.09) ............... 163
FIGURE 4.4, Ce concentration in the supernatant from different matrices after t =24 h
incubation. Control = Milli-Q water at t=0 (pH 3.11); MQ = Milli-Q water (pH 3.11 and
7.06); SWW = Synthetic wastewater (pH 7.10); WW = Real wastewater (pH 7.09)..... 164
FIGURE 4.5, Fate a) total and b) < 200 nm Ce in the influent (●) and (■) effluent
samples when feeding synthetic wastewater to the secondary treatment ....................... 168
FIGURE 4.6, Concentration of a) total unfiltered and b) filtered Ce in the (●) influent and
(■) effluent of activated sludge treatment operated with real domestic wastewater. Bold
markers in second panel indicate CeO2 < 200 nm, while empty markers refer to CeO2
< 25 nm ........................................................................................................................... 169
FIGURE 4.7, a) SEM image of a sample from the aeration tank after being fed with CeO2
for 2 months and b) EDS analysis of the sample. The white rectangle in the image
indicates the section of the sample where the EDS analysis was performed. Accelerating
voltage: 12.0 kV .............................................................................................................. 170
FIGURE 4.8, Sorption of CeO2 to biomass. Total unfiltered Ce was measured from the
supernatant of each assay after 20 h. Two pH values were considered for this test.
Treatments consisted of Milli-Q water (W) at pH 3.0 and 6.0, sludge biomass (B) at pH
3.0 and 5.9, and rinse sludge (S) at pH 2.7 and 6.0. The Milli-Q water treatment was
sampled at the beginning of the experiment (Water 0 h) to be used as a control ........... 172
FIGURE 4.9, Total COD removal in the secondary treatment system when using
synthetic wastewater, where (●) indicates the influent and (■) the effluent. Vertical line
indicates the addition of CeO2 to the influent wastewater .............................................. 174
14
LIST OF FIGURES - Continued
FIGURE 4.10, Fate of a) soluble and b) total COD in the aerobic activated sludge system
when feeding real wastewater, where (●) indicates the influent and (■) the effluent.
Feeding of CeO2 to the reactor is represented by the continuous vertical line. Dashed line
indicates fresh wastewater batches ................................................................................. 175
FIGURE 4.11, Acetate concentration in the influent, (●), and effluent, (■), of the
secondary treatment when feeding real wastewater. Continuous vertical line indicates
the addition of CeO2 to wastewater. Dashed vertical lines represent fresh wastewater
batches............................................................................................................................. 176
FIGURE 5.1, Schematic of the laboratory aerobic activated sludge treatment system.
(1) Aeration tank, (2) air, (3) settler, (4) effluent, (5) recycled sludge ........................... 196
FIGURE 5.2, a) Average particle size and b) zeta-potential of CeO2 in different matrices
at t=0 (□) and t=24 h (■) incubation. MQ = Milli-Q water (pH 3.10 and 7.01) and
WW = Real wastewater (pH 6.97) .................................................................................. 204
FIGURE 5.3, Al concentration in the supernatant from different matrices after t =24 h
incubation. Control = Milli-Q water at t=0 (pH 3.10); MQ = Milli-Q water (pH 3.10 and
7.01); WW = Real wastewater (pH 6.97). The intended Al concentration at t=0 was set
equal to 18 mg Al L-1 (34 mg Al2O3 L-1) ........................................................................ 206
FIGURE 5.4, Fate of a) total unfiltered and b) filtered Al in the (●) influent and (■)
effluent of activated sludge treatment. Bold markers in second panel indicate Al2O3 < 200
nm, while empty markers refer to Al2O3 < 25 nm. Horizontal line represents the average
Al concentration in wastewater; broken lines indicate the standard deviation ............... 208
FIGURE 5.5, SEM image of the original sludge use to seed the reactor (left) and sludge
exposed to Al2O3 NPs for 167 days (right). EDS spectrum of the sludge (bottom left) and
the exposed sludge (bottom right)................................................................................... 210
FIGURE 5.6a) Total and b) soluble COD concentration in the aerobic activated sludge
system, where (●) indicates the influent and (■) the effluent. Feeding of Al2O3 to the
reactor is represented by the continuous vertical line. Doted lines indicate fresh real
domestic wastewater batches .......................................................................................... 212
15
LIST OF FIGURES - Continued
FIGURE 5.7, Concentration of acetate in the influent (●) and effluent (■) of the
secondary activated sludge treatment. Continuous vertical lines specify the addition of
Al2O3 to wastewater. Doted vertical lines represent fresh wastewater batches .............. 214
16
LIST OF TABLES
TABLE 1.1, Maximum contaminant level (MCL) of benzene, cis-1,2- dichloroethylene,
phenol, toluene, and xylene .............................................................................................. 27
TABLE 1.2, Physical and chemical properties of benzene, cis-1,2-dichloroethylene,
phenol, p-cresol, toluene, and p-xylene ............................................................................ 32
TABLE 1.3, Average bond energies of some covalent bonds relevant to hydrocarbons . 34
TABLE 1.4, Current applications of certain nanomaterials and their forecasted
production ......................................................................................................................... 38
TABLE 2.1, Range of BTX and cis-DCE concentration found in landfill leachates ....... 62
TABLE 2.2, Description of treatments used in microcosm experiments ......................... 72
TABLE 2.3, Microcosm experiment results for BTX and cis-DCE anaerobic
degradation ........................................................................................................................ 76
TABLE 2.4, Electron-milliequivalent balance over the first 3-days from the transfer
experiment shown in Figure 5........................................................................................... 83
TABLE 2.5, Energy associated to toluene oxidation ........................................................ 90
TABLE 3.1, Concentration range of contaminants commonly found in landfill
leachates .......................................................................................................................... 110
TABLE 3.2, pH, organic loading rate (OLR), and volumetric CH4 production in the
UASB .............................................................................................................................. 121
TABLE 3.3, Nitrogen balance in the DHS reactor ......................................................... 131
TABLE 4.1, Operation parameters of the laboratory-scale activated sludge system fed
with synthetic and real wastewater ................................................................................. 155
TABLE 5.1, Operational parameters of the laboratory aerobic activated sludge treatment
in the absence and presence of Al2O3 ............................................................................. 198 17
ABSTRACT
The integrity of groundwater sources is constantly threatened by contaminant
plumes generated by accidental gasoline leakages and leachates escaping landfills. These
plumes are of concern due to their proven toxicity to living organisms. Aromatic and
chlorinated hydrocarbons, volatile fatty acids, phenols, and ammonia have been found in
these leachates. In addition, benzene, toluene, and xylenes (BTX) are major components
of gasoline. The lack of oxygen in groundwater makes anaerobic bioremediation desired
for the treatment of groundwater contaminated with BTX and chlorinated solvents. With
the objective of finding microorganisms capable of BTX and cis-dichloroethylene
(cis-DCE) degradation under anaerobic conditions for their use in permeable reactive
barriers, different inocula were tested in batch experiments. Toluene was rapidly
degraded by several inocula in the presence of alternative electron acceptors. Benzene
and m-xylene were eliminated by few of the inocula tested after incubation periods
ranging from 244 to 716 days. cis-DCE was highly recalcitrant as no degradation was
observed over 440 days. Biological processes have been successfully applied for the
treatment of landfill leachates as well. In an effort to provide an effective and economical
alternative, an anaerobic-aerobic system was evaluated using a synthetic media
simulating the organic and ammonia content of real leachates. The removal of the organic
content reached 98% in an upflow anaerobic sludge blanket reactor, and resulted in the
formation of methane. During the aerobic process, in an innovative down-flow sponge
18
reactor, ammonia was highly transformed to nitrite and nitrate. Complete nitrification
was eventually achieved.
The capacity of current wastewater treatment plants for removing nanoparticles
has been questioned during the last years. Nanoparticles have been incorporated into
numerous applications and their presence in wastewater seems to be inevitable. A
laboratory-scale secondary treatment system was set-in to study the behavior of cerium
and aluminum oxide nanoparticles during wastewater treatment. The nanoparticles were
highly removed, suggesting that secondary treatment is suitable for their elimination. The
removal of these nanoparticles was influenced by the pH and organic content of the
wastewater. Aluminum nanoparticles proved to be toxic; however the performance of the
system for eliminating the organic content was recovered over time.
19
OBJECTIVES
The aim of this research was to assess the capability of different biological
technologies for the removal of organic and inorganic contaminants commonly found in
gasoline plumes and landfill leachates and oxide nanoparticles present in municipal
wastewater. The particular objectives of this PhD dissertation are as follows:
I.
Evaluate the presence of microorganisms, in different inocula, capable of
degrading BTX and cis-DCE under anaerobic conditions using nitrate and
chlorate as alternative electron acceptors.
II.
Study the feasibility of obtaining pure cultures of BTX and cis-DCE
degrading organisms by serial transfers.
III.
Evaluate the performance of an integrated anaerobic-aerobic system for
the treatment of a synthetic media simulating a real landfill leachate.
IV.
Evaluate the partial nitrification of ammonia in the synthetic leachate to
nitrite to promote the elimination of nitrogen under aerobic conditions,
eliminating the implementation of a post-treatment process.
20
V.
Study the behavior of cerium and aluminum oxide nanoparticles during
activated sludge secondary treatment.
VI.
Elucidate the mechanisms that promote the removal, if any, of the
nanoparticles during treatment.
VII.
Investigate the effect of the nanoparticles over the performance of the
treatment system for removing the organic content of the wastewater.
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CHAPTER 1
INTRODUCTION
1.1
Hydrocarbons
Hydrocarbons are organic compounds composed entirely of carbon and hydrogen
and can be found as gases, liquids, or solids. They can have different shapes, including
straight and branched chains and cyclic arrangements. According to the carbon-carbon
(C-C) bonding features in the molecule, hydrocarbons can be classified into
single-bonded (alkanes), double-bonded (alkenes), triple-bonded (alkynes), and
mono- or polycyclic aromatics. Hydrocarbons are widely abundant in nature and are
formed by geochemical reactions or during the degradation of organic matter mediated by
living organisms [1]. Hydrocarbons can be synthetically produced as well, for example,
by the catalytic mediated reaction between hydrogen and carbon monoxide (CO) [2].
Hydrocarbons are the most important source of energy around the globe [3] and find
numerous industrial and commercial applications. It is believed that the use of
hydrocarbons as fuels is of major importance for the human development in main areas
such as the economic and social sectors [4]. The economic importance of hydrocarbons is
22
not only related to their use as fuels, but also to their utilization in the manufacturing of
plastics, dyes, paints, pesticides, among others. However, the dependence of the society
on hydrocarbons may pose a risk to the environment, as different studies have
demonstrated that anthropogenic activities are the main source of hydrocarbon
contamination in different ecosystems [5-8].
1.1.1
Aromatic hydrocarbons
Aromatic hydrocarbons comprise the second most abundant group of organic
compounds in nature [9]. They alternate single and double bonds between carbons in the
molecule and contain one or more benzene rings. Benzene, toluene, and (o-, m-, p-)
xylene (BTX) are volatile monoaromatic hydrocarbons of great importance to society
since they are used in diverse processes and commercial products. Figure 1.1 provides the
structural formula of each of the BTX compounds. They are found in petroleum and
petroleum derivatives and are extensively utilized for the manufacturing of plastics,
fibers, pesticides, paints, inks, dyes, and nylon. In 1997, the industrial production of these
compounds was estimated in the range of megatons per year (1 x 1012 grams) [10]. It is
not surprising that BTX compounds are recognized as one of the major causes of
environmental pollution [11], resulted from landfill leachates reaching groundwater,
leakage of underground storage tanks (USTs) and pipelines, spills at production wells,
refineries, and distribution terminals, and improper disposal and accidents during
transport [12]. BTX are so prevalent as environmental contaminants, that these chemicals
23
have been listed in the Comprehensive Environmental Response, Compensation, and
Liability Act (CERCLA) priority list. Concerns about exposure to BTX are due to their
known toxic effects [13] and their relatively high mobility, which allows them to travel
long distances. According to the Agency for Toxic Substances and Disease Registry
(ATSDR), long term exposure to toluene [14] and the three xylene isomers [15] can
result in damage of the central nervous system, the liver and/or the kidneys; whereas
benzene is a suspected carcinogen [16]. For this reason, the Environmental Protection
Agency (EPA) has set a maximum contaminant level (MCL) in potable water of 0.005, 1,
and 10 mg L-1, for benzene, toluene, and (o-, m-, p-) xylene, respectively (Table 1.1).
Although BTX are so widespread as pollutants in nature, other organic compounds of
concern as well.
24
a)
c)
b)
d)
f)
e)
Figure
F
1.1. Structural formula of a) benzenee, b) toluenne, c) p-xyylene, d) phhenol,
c)) p-cresol, an
nd f) cis-dich
hloroethylen
ne.
25
1.1.2
Phenols
Phenolic compounds, also called phenols, consist of one or more aromatic
molecules bonded to one or more hydroxyl groups (-OH). Although phenols are similar
to alcohols, they are not classified as such, since the –OH group is bonded to an aromatic
ring, what makes phenols weak acids [17]. Phenolic compounds are naturally produced
organic molecules essential for the growth, development, and defense of plants; phenols
are commercially employed as dyes and food additives [17], among other applications.
Phenol is the simplest of the phenolic compounds and consists of a single benzene ring
bonded to a single –OH group (Figure 1.1). It is commonly used as disinfectant and it is
employed in numerous synthetic chemical processes since phenol is the basic structural
unit for different organic compounds [18, 19]. Phenol production was estimated in 1.25
billion Kg in 1980 [19]. Other important phenolic compounds are cresols, which are
methylated phenols. Cresols can be obtained from coal tar and petroleum or from the
oxidation of toluene; and like phenol, cresols are employed in the synthesis of other
organic compounds, and are extensively used in the manufacture of pesticides and resins
[20]. The structural formula of p-cresol, or 4-methylphenol, is shown in Figure 1.1.
Releases of phenolic compounds are of concern since they are known to be toxic at high
concentrations and they are genotoxic at low concentrations [21]. Dissemination of
phenolic compounds into the environment is common due to their presence in countless
products and processes; they have been frequently identified in landfill leachates [22, 23]
posing a threat to groundwater.
26
1.1.3
Chlorinated hydrocarbons
Chlorinated hydrocarbons are organic molecules bonded to at least one chlorine
atom. These compounds are produced in large quantities since they are extensively used
as solvents for degreasing fats, oils, waxes, and resins [24]. cis-1,2- Dichloroethylene
(cis-DCE) is industrially used as solvent for polymers, waxes, and resins; extended
exposure to this contaminant results in damage to the central nervous system, and it is a
suspected carcinogen [25]. The structural formula of cis-DE is shown in Figure 1.1. In
order to reduce the risk of exposure to cis-DCE, the EPA has set a MCL for cis-DCE
equal to 0.07 mg L-1 (Table 1.1). cis-DCE can contaminate groundwater due to improper
disposal of chlorinated solvents and/or from landfill leachates escaping the landfill.
Although large amounts of cis-DCE are produced each year, its occurrence in
groundwater is primarily due to the microbial degradation of higher chlorinated solvents,
mainly tetrachloriethylene (PCE) and trichloroethylene (TCE) [26]. Due to the risk that
all these contaminants pose to human health and the environment, their removal from
groundwater and landfill leachates is a high priority.
27
Table 1.1. Maximum contaminant level (MCL) of benzene, cis-1,2- dichloroethylene,
phenol, toluene, and xylene.
MCL
(mg L-1)
Potential health
effects above the
MCL
Common sources of
contaminant in
drinking water
Benzene
0.005
Anemia; increased
risk of cancer
Discharge from factories;
leaching from gas storage
tanks and landfills
cis-1,2Dichloroethylene
0.07
Liver problems; risk
of cancer
Discharge from industrial
chemical factories
*Phenol
10
Gastrointestinal and
skin damage
Discharge from industrial
chemical factories
Toluene
1
Nervous system,
kidney, or liver
problems
Discharge from petroleum
factories
Xylene
10
Nervous system
damage
Discharge from petroleum
factories; discharge from
chemical factories
Contaminant
* Phenol is listed only in the national recommended water quality criteria. p-Cresol is not
classified by the EPA as a priority pollutant. Source: Environmental Protection Agency,
2011 [18, 27, 28].
28
1.2
Hydrocarbons in landfill leachates
Sanitary landfills are the preferred option for the disposal of household waste,
since they are the most economical alternative for handling the waste and allow its
decomposition under controlled conditions [29]. It is estimated that around 70% of the
household refuse is disposed off in landfills worldwide [30] and normally consists of
paper, food waste, vegetative matter, plastics, textiles and glass [31]. However, hazardous
waste can be found as well, due to illegally dumped industrial material and the presence
of toxic compounds in household products [32]. Landfills have been recognized to pose a
threat to the environment since they can emit pollutants that have the potential to
jeopardize the quality of the air [33, 34] and surface water and groundwater [35, 36]
surrounding the landfill.
1.2.1
Characterization of landfill leachates
Landfill leachates are formed as percolating water mixes with chemicals leached
from the buried waste due to the intrinsic physicochemical and microbial conditions in
landfills. Landfill leachates pose a threat to groundwater as numerous organic
contaminants have been identified in the leachates, including BTX, phenol, p-cresol, and
cis-DCE [37-41]. Baun et al [22] found 55 xenobiotic organic compounds and 10
degradation byproducts in leachates from 10 Danish landfills. In Japan, Yasuhara et al
[42] identified more than 100 organic compounds, including toxic chemicals such as
29
1,4-dioxane, phthalates, and bisphenol, in leachates from 11 different landfills. Paxeus
[23] detected more than 200 organic compounds in leachates from 3 landfills in Sweden.
The composition of the leachates can vary with the age of the landfill. Leachates
produced during the initial aerobic phase have a neutral pH and a high concentration of
complex compounds. Once oxygen is depleted, anaerobic conditions develop. Further
degradation of the high molecular weight organics results in a leachate of acidic pH and
high biological oxygen demand (BOD) content due to the presence of soluble organic
compounds, such as volatile fatty acids (VFA). Methanogenic conditions gradually
dominate over time as VFA and other compounds are degraded, which causes a decrease
of the BOD content of the leachate and an increase of the pH to circumneutral values [4345].
1.2.2
Physicochemical technologies for treating landfill leachates
Different technologies exist for the treatment of landfill leachates and can be
divided in three major groups, including (a) leachate recirculation, (b) physicochemical
methods, and (c) membranes [40, 45, 46]. Leachate recirculation was a preferred option
for treating landfill leachates and consists in the recirculation of the leachate to the top of
the landfill. Although this technique is efficient and relatively inexpensive, ammonia
(NH4+), an inorganic contaminant occurring at high concentrations in landfill leachates, is
not treated and it accumulates. Physicochemical processes, which include coagulationflocculation, adsorption, and advance oxidation, are used to remove suspended solids,
30
colloidal particles, and biological refractory compounds. Due to their high operation costs
they are normally used as a pre- or post-treatment. Finally, techniques employing
membranes, such as reverse osmosis (RO) and ultra-, micro-, and nanofiltration, are
mainly utilize for the polishing of treated effluents to avoid a rapid saturation of the
membranes.
1.2.3
Physicochemical processes for groundwater treatment
Landfill leachates can contaminate groundwater if landfills lack of contention
systems or if such systems fail. Hydrocarbons have shown to be a constant threat to
groundwater streams. Only in the US it is estimated that groundwater from 300,000 to
400,000 sites is contaminated with organic chemicals [47]. Different techniques have
been utilized for the removal of aromatic and chlorinated compounds in groundwater;
perhaps the most commonly used are air striping, carbon adsorption, and chemical
oxidation. The used of these technologies relies on the physical and chemical properties
of the contaminants, such as density, henry’s law constant, and water solubility. Table 1.2
provides information about these physicochemical characteristics for BTX, phenol, pcresol, and cis-DCE. Air stripping consists in exposing the contaminated groundwater to
a clean air supply, resulting in the release of the organic compounds into the gas phase
[48]. It can enhance the biodegradation of the pollutants by providing oxygen that may be
lacking on the site. Depending on the type and concentration of the organic contaminants,
the gas phase might require treatment before being discharge into the atmosphere [49].
31
Another important method for treating contaminated groundwater involves the use of
activated carbon (AC). Processes based on AC technologies are so widely used that are
normally employed as the primary treatment, preceding polishing processes [50]. AC can
be obtained from different raw materials, including wood, coconut shells, peats, and coals
[51], and it is characterized by a its high porosity and large surface area. When
groundwater is treated using AC, it is generally pumped out and passed through filters
containing granules of AC. The removal of the contaminant from the groundwater is
achieved by adsorption between the contaminant and the AC; which occurs when the
attractive forces between the organic molecules and the liquid are overcome by the
attractive forces at the carbon surface. The main factors affecting the adsorption of the
contaminant onto the AC are the pore size distribution and the surface chemistry of the
AC [50]. After time, the AC must be replaced as it reaches its maximum adsorption
capacity, increasing the operation costs. Finally, chemical oxidation is another viable
option for treating contaminated groundwater. During chemical oxidation, the organic
contaminants are degraded by the addition of oxidizing agents, including ozone and
hydrogen peroxide [52]. Higher degradation rates are achieved when advance oxidation
processes are used instead of chemical oxidation. During advance oxidation processes
hydroxyl radicals (OH-) are generated most commonly by cavitation (ultrasonic
irradiation), photocatalytic oxidation (ultraviolet radiation), or Fenton’s chemistry, to
oxidize the organic molecules [53]. Although oxidation processes have demonstrated to
be a promising alternative for the degradation of organic contaminants, operation costs
can be high due to the use of expensive reactants [54].
32
Table 1.2. Physical and chemical properties of benzene, cis-1,2-dichloroethylene, phenol,
p-cresol, toluene, and p-xylene.
Contaminant
Density
at 20○C
(Kg L-1)
Henry’s constant
at 25○C (atm m3
mol-1)
Solubility at
25○C (g L-1)
Octanol-water
partitioning
coefficient (Log Kow)
Benzene
0.8787
(15○C)
5.50 x 10-3
0.188 (%)
2.13
cis-1,2Dichloroethylene
1.284
4.08 x 10-3
3.50
1.86
Phenol
1.0545
3.00 x 107
82.4
1.46
p-Cresol
1.0341
7.92 x 101
21.5
1.94
Toluene
0.8669
5.94 x 10-3
0.535
2.72
p-Xylene
0.8611
6.90 x 10-3
0.162
3.15
Data obtained from the Agency for Toxic Substances and Disease Registry (ATSDR)
[14-16, 18, 20, 25].
33
1.2.4
Bioremediation. An alternative for groundwater and landfill leachate treatment
Physicochemical processes are not the only viable alternative for treating
contaminated groundwater and landfill leachates. Processes relying in bioremediation
technologies have gained importance during the last 30 years since it is suggested that
these technologies are safer, less costly, and less disruptive than some current
physicochemical processes [55]. Bioremediation is defined by the EPA as any process by
which microorganisms, or their enzymes, transform the structure of harmful chemicals
released into the environment. Many microorganisms, including bacteria and fungi, are
known to be capable of degrading a diverse range of hydrocarbons. It is believed that,
since hydrocarbons are ubiquitously present in the environment, microorganisms have
developed pathways to use these compounds as growth substrates [1, 56]. Moreover, it
has been suggested that one of the primary removal mechanisms of petroleum and other
organic contaminants involves the biodegradation of such compounds by natural
populations of microorganisms in the environment [57]. Many microorganisms are
known to be capable of degrading numerous natural and synthetic organic compounds in
the presence of oxygen. In this case, the initial attack requires the use of molecular
oxygen as co-substrate [58]. Hydrocarbons were thought to resist biodegradation in the
absence of oxygen since aromatic and aliphatic compounds are very stable, resulted from
the resonance energy in the aromatic ring and the inertness of C-C and H-C bonds in the
organic molecules [56]. The energy associated to these bonds is shown in Table 1.3.
However, it is now widely known that different organic contaminants can be degraded in
34
the presence of alternative electron acceptors [1, 3, 12, 58]. Anaerobic biodegradation of
hydrocarbons in the presence of alternative electron acceptors, including nitrate (NO3-),
ferric iron (Fe3+), sulfate (SO42-), manganese (Mn4+) and carbon dioxide (CO2), has been
demonstrated [3]. In contrast to the aerobic degradation, the initiation reactions under
anoxic conditions are widely diverse [58]. During aerobic degradation, the organic
contaminants are transformed mainly to CO2 and biomass sludge, whereas the main
products of the anaerobic degradation are methane (CH4) and CO2 [29]. Bioremediation
has been successfully applied in the cleanup of groundwater, soils, wastewater, among
others; it has even proved to be effective for large scale applications as demonstrated by
the remediation of the Exon oil spill in Alaska [59]. However, certain environmental,
biological, and physicochemical factors can impede bioremediation to occur. Thus, more
research is needed to make bioremediation processes a more feasible treatment
alternative.
Table 1.3. Average bond energies of some covalent bonds relevant to hydrocarbons.
Bond
Bonds order
Bond energy (KJ mol-1)
H–C
1
411
C–C
1
346
C=C
2
602
C=C
3
835
Data obtained from “Inorganic chemistry: principles of structure and reactivity” by J.E.
Huheey et al, 1993 [60].
35
1.3 Removal of emerging contaminants during wastewater treatment
Wastewater commonly contains pollutants, pathogenic microorganisms, and
nutrients that pose a risk to human health and environmental welfare if released
untreated. With the objective of removing these contaminants and hazardous
microorganisms,
wastewater
treatment
plants
(WWTPs)
have
implemented
physicochemical and biological processes to produce effluents that can be safely
discharged into the environment. Wastewater treatment has evolved over the years. The
elimination of colloidal and suspended material, biodegradable organics, and pathogenic
organisms based on aesthetic and environmental concerns was the primary goal of the
treatment from the 1900s until the 1980s; during the last 30 years the goal of WWTPs has
been the removal of constituents that may pose a threat to health and the environment in
the long-term [61]. However, as new emerging pollutants previously unexpected or
unrecognized to reach WWTPs are being found in wastewater, the capability of WWTPs
for eliminating these contaminants is being questioned. Nanoparticles (NPs) are one
important group of emerging pollutants. These particles are defined by the National
Nanotechnology Initiative (NNI) as materials with at least one dimension of 100 nm or
less and can be naturally formed (volcanic dust, colloids in fresh water, soil erosion) or
synthetically
engineered.
NPs
are
much
smaller
(500 – 5000 nm) and many viruses (10 – 300 nm).
in
diameter
than
bacteria
36
1.3.1
Nanoparticles
Engineered NPs (ENPs) can be manufactured by mechanical and chemical
processes and by gas-phase synthesis [62]. NPs characteristics such as size, shape,
chemical composition, and surface area and charge influence their complex colloid and
aggregation chemistry. The size of the NPs has a great impact over the charge
distribution within the particle and their surface area. The reactivity of the nanomaterials
is mainly affected by their surface structure which is defined by the shape and chemical
composition of the NP. The molecular bonding pattern of the material greatly changes as
its shape is modified. Finally, the chemical composition refers mainly to the crystallinity
of the NP and the presence of functional organic and/or inorganic groups on its surface.
All these properties make ENPs to have a dual behavior between the quantum effects of
atoms and bulk properties of large materials.
The special characteristics of ENPs have caused an exponential increase in the
utilization of these materials in countless processes and products. According to a 2001
report by the National Science Foundation (NSF), nanotechnology is estimated to reach a
$1 trillion dollar market by 2015 [63] and visualizes this field as a relevant technology
that could revolutionize the science and the industry. ENPs are now found in
stain-resistant clothing, cosmetics, food packing, drugs and many other applications.
Sunscreen production is one of the most important markets for ENPs, where titanium
dioxide (TiO2) and zinc oxide (ZnO) are regularly used. These nano-oxides are the most
common inorganic agents utilize in sunscreens for scattering purposes [64]. It has been
37
estimated that sunscreen products sales reached near a half billion dollars in the year
2005 [65]. ENPs can also be used for interesting applications in the field of
environmental engineering [66-68], resulting in cheaper and more effective remediation
techniques. The most common nanomaterial used to date in environmental applications
has been zero-valent iron (ZVI); however, many other nanomaterials are being
considered such as nano-zeolites, nano-oxides such as TiO2, and carbon nanotubes [69].
The most important characteristic that makes nanomaterials interesting for environmental
applications is their high surface area to volume ratio that could enhance their reactivity
with pollutants [70].
Semiconductor industrial processes have incorporated NPs in many operations.
The most important nanomaterials utilize in this market are the inorganic nano-oxides
including silica (SiO2), alumina (Al2O3), and ceria (CeO2), which are used in slurries used
for the chemical mechanical polishing (CMP) of wafers. The objective of the CMP is to
obtain wafers with smooth surfaces and free of defects by mechanical abrasion and
chemical reaction when in contact with the slurries. Mechanical grinding can achieve
planarization of the surface of the wafer; however it leaves a rough pattern. To eliminate
the damage on the wafer the chemical components of the slurries react with the wafer to
achieve the smooth surfaces. The slurry market for CMP applications was calculated to
reach $ 400 million in 2003 and accounted for the 74.2% of the worldwide NP market in
2005, including magnetic and optoelectric applications [71]. According to the
Semiconductor Industry Association sales jumped 11% from 2000 to 2005, due in large
part to CMP. The extensive and increasing use of engineered NPs has led to increasing
38
concern about their potential toxicological effects and environmental impact. Table 1.4
shows current uses of different nanomaterials and their estimated production.
Table 1.4. Current applications of certain nanomaterials and their forecasted production.
Materials/Devices
Application
Estimated
production from
2011 – 2020 (tonnes
per year)
Ceramics, catalysts
composites, coating,
powders, metals
Structural
104 - 105
Metal oxides (TiO2, ZnO,
Fe2O3)
Skincare products
103 or less
Nanoelectronics,
optoelectronic materials
Information and
communication technology
103 or more
Nanoencapsulates, quantum
dots
Biotechnology
10
Nanofiltration, membranes
Environmental
103 - 104
Nanoelectronic-mechanical
systems
Instruments, sensors,
characterization
102 - 103
Adapted from “Nanomaterials – the driving force” by M.J. Pitkethly, 2004 [62].
39
1.3.2
Health effects and environmental fate of nanoparticles
The use of nanotechnology has been associated to certain benefits, including
improvements in energy production, storage and efficiency as well as pollution and waste
minimization by implementing “green” technologies that will generate less undesirable
by-products and reducing consumption of resources [72]. Nanotechnology, together with
biotechnology, was thought to pose the potential to eliminate the concept of waste and
pollution by creating products and services at the molecular scale [73]. However it has
been argued that the same characteristics that provide NPs their unique behavior could
influence their toxicological effects to humans and the environment. In some cases, cell
membranes seem to be an ineffective barrier to certain NPs under 100 nm of diameter,
which can readily pass through them [74]. The inherent small size of NPs results in large
surface area, as previously discussed. This feature, which is concentration dependent, is
of concern because some of these NPs are very reactive, posing a hazard to organisms.
The capacity to generate reactive oxygen species (ROS) [75] is a clear example of the
potential risks associated to the large surface area and reactivity of these materials.
Awareness of the possible adverse effects of the production, manipulation, release, and
exposure to NPs has increased in the last years. The US federal government will invest
$480 million dollars between 2005 and 2011 to support research related to the potential
environmental, health, and safety (EHS) impacts of nanoparticles, as stated by the
National Nanotechnology Initiative (NNI). Despite the increasing concern about the
40
potential environmental impact and adverse health effects of NPs, nothing has been done
towards the regulation of such materials [76].
Engineered nanomaterials have modified surfaces that could threaten the
well-being of living organisms if exposed to these particles. The possible routes of
exposure to NPs in humans include inhalation and ingestion of ultrafine particles, as well
as direct contact. Particles of 50 nm, or smaller, behave more like gas molecules, thus
they can deposit anywhere in the respiratory track; moreover they can reach the liver or
the kidneys when ingested and can penetrate the skin rather easy [77]. Interactions of NPs
with other molecules affect their fate in the organisms since they can bind to proteins
forming complexes that are more mobile and that can penetrate tissue sites otherwise
inaccessible [78]. The intrinsic characteristics of NPs play a major role in their excretion
as well, normally via urine, feces or sweat; properties such as surface charge, size, and
chemical composition greatly determine their toxicity and whether or not these materials
are expulse out of the body and at what extent. In general, smaller particles are thought to
be more toxic than their larger counterparts since the surface area increases as the
diameter of the particles decreases. Documented adverse effects of NPs include oxidative
stress by formation of ROS, briefly discussed above. Under normal coupling conditions
in the mitochondrion ROS species are readily neutralized; however, as the concentration
of ROS increases the capacity of the cells are surpassed, causing inflammation as
observed in the lungs [79] during ambient and occupational exposure. Other effects are
particular to each NP species. For example, to mention a few, silver nanomaterials have
shown to cause lung cancer, nano TiO2 and ZnO may damage nucleic acids, and carbon
41
nanotubes may cause dermal toxicity [72]. A comprehensive risk assessment about the
use and exposure to NPs is not possible at this stage due to the lack of knowledge about
their fate in the organisms and the environment. Despite data about the toxicity of these
materials constantly emerge, most of his information is obtained from observational
studies and limited to species used for regulatory toxicology [80]. Information about NPs
in the environment is even more restricted. As recognized by the EPA [81], little is
known about the environmental fate of nanomaterials released directly or indirectly from
manufacturing and processing of nanomaterials, oil refining processes, chemical and
materials manufacturing processes, among others.
1.3.3
Surface chemistry of nanoparticles
Contrary to bulk materials, a large fraction of the atoms in NPs are exposed on the
surface [82]; therefore adsorption processes are thought to control their behavior in the
environment. In general, in aqueous solutions, the fate of NPs is determined by their
interactions with organic and inorganic chemicals, the naturally occurring biotic/abiotic
processes, and the stability of the dispersions they form [81]. In aquatic environments
NPs are rarely dissolved; instead they form colloidal dispersions [83] whose stability
greatly depend on pH and ionic strength (IS), since colloids are generally charged. When
particles are small and have negligible settling velocity, Brownian motion is considered
to be the most important collision mechanism. Brownian motion refers to the random
movement of particles suspended in a fluid and is characterized by (i) continuous motion
42
and (ii) independent gaussian increments [84]. Particles may collide or not into each other
depending on their concentration and affinity; after collision they may adhere depending
once again on their affinity and agitation of the media. The affinity between NPs is
defined by the zeta potential (ζ-potential) of the colloidal particles and represents the
magnitude of the electrostatic forces between the surface of the particle and the bulk
media, which can be repulsive or attractive.
Solid oxide particles dispersed in aqueous media are often electrically charged
due to an imbalance between the densities of adsorbed H+ and OH- and the presence of
ionic groups and/or the adsorption of metal hydroxo complexes on the surface of the
particle [85]. The ions adsorbed on the surface of the particle attract ions of opposite
charge, present in the aqueous phase, forming two parallel charge layers that surround the
particle called electrostatic diffuse double layer (EDL) and is shown in Figure 1.2. The
EDL can be either positive or negative, depending on the surface ligands of the particle.
The forces associated with the EDL are generally repulsive due to the electrostatic
surface charge of the materials and dominate as the distance between particles increases.
Attraction between neutral NPs is the result of London-var der Waals (LVW) forces that
dominate in the short range, as the separation of such NPs decreases. LVW interactions
are exhibited by nonpolar materials and are originated by the dipoles in the molecules. As
the absolute charge of the particles increases (positive or negative), the absolute value of
ζ-potential exhibited by the particles increases as well, resulting in stable dispersions. On
the other hand, as the particles loose charge and become neutral, the absolute ζ-potential
43
decreases (po
ositive or negative) and the dispersiion becomess unstable, w
which can leead to
Surface of the particle
th
he sedimentaation of the particles.
p
Figure
F
1.2 Schematic
S
diiagram show
wing the eleectrical doubble layer at the surfacee of a
particle. Adap
pted from Handy et al, 2008
2
[86].
The pH
p of the media
m
has an
a importannt influence over the ζζ-potential oof the
particles, affeecting their average parrticle size. A
Any changee in the pH of the meddia is
fo
ollowed by change
c
in th
he ζ-potentiaal, resulting in an increaase or decreaase of the suurface
ch
harge of thee NPs. At low
l
pH vallues, metalliic NPs becoome positivvely chargedd (+),
44
whereas at high pH the charge becomes negative (-). The pH at which the surface of the
NPs is neutral is called the isoelectric point (IP) (Figure 1.3) and particles are expected to
agglomerate. At pH values distant from the IP particles are stabilized. Different studies
have demonstrated the effect of pH over particle size and stability of NPs dispersions [8790]. Since the stability of the particles is function of the charge of the NPs, the ionic
strength (IS) of the dispersion plays an important role as well. The IS is an indicator of
the concentration of ions in solution and increases in the presence of salts, since they
normally dissociate into ions. Salts reduce the repulsion between the double layers of the
particles, making it possible for NPs to get very close to each other. In the case of
positively charged NPs, the surface charge layer is compressed, thus subsequent
collisions of the NPs will cause aggregation, and probably further sedimentation [80].
Studies have demonstrated the influence of salt concentration over particle size in
solution [91-93]. Amal et al [94] observed that the agglomeration velocity of haematite
NPs was directly dependent of the salt concentration. Thus, higher agglomeration
velocities were observed as the salt content in solution was increased.
45
Figure
F
1.3 Scchematic sho
owing the deependence oof the zetal-ppotential on tthe surface oof the
particle on the pH of the media.
m
nd particle size distrribution (PS
SD) can bbe modifiedd by
The stability an
en
ncapsulating
g the NPs in
n organic orr inorganic polyelectrollytes, or stabbilizers [95,, 96].
Polyelectroly
ytes dissociatte when in solution andd can imparrt surface chharge to the NPs,
in
ncreasing thee stability off the disperssion. The moost commonn stabilizers in the markeet are
Carbowax-20
C
0
M,
poly-N-vinyl-2-pyrrollidone,
poly-acrylicacid-b-polysttyrene-6,
mide,
poly-N-isoppropylacrylam
3-aminoprop
3
pyl-trimethooxysilane
((APS),
soodium
polyacrylate, dendrimers, and cubic silsesquioxxane [97]. Sttabilizers arre pH depenndent,
hus differen
nt stabilizerss can behav
ve distinctivvely at the ssame pH, eenhancing oor not
th
ag
gglomeration
n of the NPss [98]. In general, NP diispersions uttilized in thee industry coontain
sttabilizers and dispersantts that stabiliize the NPs at neutral pH
H values. Thhis allows N
NPs to
reemain disperrsed for extended periods of time aand travel loong distancees, increasinng the
46
likelihood to reach WWTPs. Most of the engineered nanomaterials are released into the
environment via sewage and industrial wastewater discharges [99]. Different computer
models have suggested the potential of many ENPs to reach municipal WWTPs [100,
101], which was confirmed by the finding of TiO2 NPs in effluents of municipal WWTPs
across the nation [102]. Not only industrial wastewater can contain significant
concentrations of ENPs. These materials can be found in municipal sewage due to their
increased use in numerous household products. The fate of ENPs during wastewater
treatment is limited and more research is needed to determine if current technologies are
appropriate for their removal.
1.4
References
1.
Widdel, F. and R. Rabus, Anaerobic biodegradation of saturated and aromatic
hydrocarbons. Current Opinion in Biotechnology, 2001. 12(3): p. 259-276.
2.
Underwood, A.J.V., Industrial synthesis of hydrocarbons from hydrogen and
carbon monoxide. Industrial and Engineering Chemistry, 1940. 32: p. 449-454.
3.
Chakraborty, R., Anaerobic degradation of monoaromatic hydrocarbons. Appl
Microbial Biotechnol, 2004. 64: p. 437-446.
4.
Hall, C., et al., Hydrocarbons and the evolution of human culture. Nature, 2003.
426(6964): p. 318-322.
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5.
Bicego, M.C., et al., Results from a 15-year study on hydrocarbon concentrations
in water and sediment from Admiralty Bay, King George Island, Antarctica.
Antarctic Science, 2009. 21(3): p. 209-220.
6.
Steinhauer, M.S. and P.D. Boehm, The composition and distribution of saturated
and aromatic-hydrocarbons in nearshore sediments, river sediments, and coastal
peat of the alaskan beafort sea - implications for detecting anthropogenic
hydrocarbon inputs. Marine Environmental Research, 1992. 33(4): p. 223-253.
7.
El Deeb, K.Z., et al., Distribution and sources of aliphatic and polycyclic
aromatic hydrocarbons in surface sediments, fish and bivalves of Abu Qir Bay
(Egyptian Mediterranean Sea). Bulletin of Environmental Contamination and
Toxicology, 2007. 78(5): p. 373-379.
8.
Bicego, M.C., et al., Assessment of contamination by polychlorinated biphenyls
and aliphatic and aromatic hydrocarbons in sediments of the Santos and Sao
Vicente Estuary System, Sao Paulo, Brazil. Marine Pollution Bulletin, 2006.
52(12): p. 1804-1816.
9.
Brackmann, R. and G. Fuchs, Enxymes of anaerobic metabolism of phenoliccompounds - 4-hydroxybenzoyl-CoA reductase (dehydroxylating) from a
denitrifying Pseudomonas species. European Journal of Biochemistry, 1993.
213(1): p. 563-571.
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59
CHAPTER 2
ANOXIC OXIDATION OF TOLUENE, BENZENE, M-XYLENE, AND CIS-DCE IN
THE PRESENCE OF BIOLOGICAL SLUDGE
2.1
Abstract
Benzene, toluene, xylene (BTX), and cis-1,2-dichloroethene (cis-DCE) are
common groundwater contaminants as they are used in many industrial processes and
household products. The low oxygen concentration found in groundwater makes aerobic
degradation less feasible. Technologies based on anaerobic processes have emerged as a
low cost alternative for the treatment of contaminated groundwater. These technologies
rely in the utilization of alternative electron acceptors, including nitrate (NO3-), sulfate
(SO42-), ferric iron (Fe3+), manganese (Mn4+) and chlorate (ClO3-), among others, to
promote the degradation of organic molecules. The objective of this work was to evaluate
the presence of microorganisms in a total of six different sludge and sediment samples
taken from diverse sources, capable of BTX and cis-DCE removal under anaerobic
conditions in the presence of NO3- or ClO3-. The final goal was to determine if such
microorganisms were ubiquitously present in the sludges and sediments tested in order to
consider their use for the anoxic oxidation of BTX and cis-DCE. For this purpose batch
experiments were set up in which individual BTX and cis-DCE compounds were tested
60
for their biodegradation in microcosms amended with a solution of NO3- or ClO3-.
Non-inoculated and poisoned controls were incubated in parallel to verify that removals
by live inocula were due to biological reactions. Degradation was monitored by
measuring the contaminant concentration in the headspace. Toluene was found to be the
easiest compound to be degraded, as it was readily removed under denitrifying conditions
in the majority of inocula tested and in some cases the removal was observed in less than
7 days. NO3- analysis during toluene degradation demonstrated that NO3--consumption
was linked to toluene degradation. Nitrite (NO2-) was observed as an intermediate during
denitrification. Toluene degradation was not observed under ClO3--reducing conditions;
whereas in one inoculum, slow degradation of toluene was observed under methanogenic
conditions. Benzene was removed in one duplicate after extended incubation in the
presence of NO3- and ClO3- by granular and liquid sludge after a lag time of 100’s of
days. m-Xylene degradation was only observed when using pond sediments under
methanogenic conditions after an initial lag phase of approximately 100 days. Finally,
cis-DCE was not degraded under any of the conditions investigated. Data suggests that
microorganisms suitable for anaerobic degradation of BTX and cis-DCE were not present
in the inocula tested.
61
2.2
Introduction
Benzene, toluene, (o-, m-, p-) xylene (BTX), and cis-DCE are common
contaminants that can reach groundwater mainly from leaking underground fuel storage
tanks and leachates escaping landfills. Groundwater contamination by cis-DCE occurs
not only by leachates containing this pollutant found in waxes, polymers and resins, but
also by the anaerobic degradation of highly chlorinated ethenes and ethanes. Table 2.1
summarizes the concentration ranges of these contaminants in landfill leachates reported
in the literature. Additionally, BTX compounds constitute an important fraction of fuels,
such as gasoline and diesel, and oil. Approximately 480,000 leaking fuel storage tanks
are estimated to occur in the US [1]. These contaminants are known to cause adverse
effects to human health. Damage to central nervous system, the liver, and kidneys has
been associated to long exposure to toluene [2] and (o-, m-, p-) xylene [3]. Likewise,
benzene and cis-DCE are suspected carcinogens [4, 5].
62
Table 2.1. Range of BTX and cis-DCE concentration found in landfill leachates.
Pollutant
Benzene
Toluene
Xylenes
cis-1,2Dichloroethylene
Conc. (µg/L)
Reference
0.2 - 1,630
[6]
2.3 - 38.9
[7]
0.65 - 3,800
[8]
3,800
[9]
1 - 123,000
[6]
1.9 – 241
[7]
0.01 - 41,000
[8]
41,000
[9]
0.01 – 1.29
[10]
0.8 - 3,500
[6]
2 - 2,220
[7]
4 - 170,000
[8]
0.03 - 0.208
[10]
1.4 – 470
[6]
1.4 -60
[8]
Bioremediation provides an efficient and cost effective alternative for the
treatment of groundwater contaminated by landfill leachates and gasoline plumes [11].
Natural attenuation (NA) relies on physical, chemical, and biological processes occurring
with no human intervention, and relies in the presence of organisms capable of
transforming the contaminants affecting the site and the occurrence of electron acceptors
along the path of the contaminated plume [12]. When the contamination cannot be
63
removed by NA, engineered bioremediation becomes an interesting option. In situ
treatment is a cost-effective innovative technology that allows remediation without
removing the contamination from one place to another. Furthermore, it is less destructive
than many other alternatives. Another popular emerging technology is the utilization of
permeable reactive barriers (PRBs). PRBs are built below ground by digging a narrow
trench, which intercepts the plume, filled with reactive materials. Degradation is
promoted by filling the barriers with either electron acceptors or donors, nutrients or any
other component lacking in the site, which are slowly released over time. This technology
achieves treatment and containment of the plume at relatively low cost [13]. In fact,
PRBs are extensively used for the treatment of chlorinated hydrocarbons [14].
Degradation of organic compounds can occur in the presence and absence of
oxygen, and ample evidence exists about the biological transformation of BTX under
aerobic conditions [15-17]. Aromatic hydrocarbons are known to be aerobically degraded
by both prokaryotes and eukaryotes by the incorporation of molecular oxygen into the
aromatic molecule by oxygenase enzymes. While eukaryotes generally cometabolize
these compounds, prokaryotes use BTX as carbon source and obtain energy from their
degradation. In non-ligninolytic fungi, an oxygen atom is incorporated into the aromatic
ring via monooxygenase resulting in the formation of an arene oxide which is further
oxidized to a dihydrodiol [18]. Bacteria, on the other hand, can perform either a
monooxygenase or dioxygenase attack. In unsubstituted aromatics the ring is
hydroxylated by diooxygenase enzymes forming a cis-dihydrodiol, which is converted to
cathechol by dehydrogenase enzymes [19]. The cathechol formed can be cleaved in the
64
ortho- or metha- position to ultimately produce low molecular weight organics that are
readily oxidized in the Krebs cycle [20]. Substituted aromatics can be subjected to a
monooxygenase attack, where the methyl or ethyl substituents can undergo oxidation
prior the activation of the aromatic ring to produce carboxylic acids or substituted
pyrocatechols that are further transformed into simpler organic molecules [21]. Aerobic
processes are normally faster than those relaying on alternative electron acceptors;
however biodegradation of BTX is very sensitive to the oxygen concentration in the
groundwater, which is generally limited or absent due to microbial respiration [22].
Utilization of hydrogen peroxide and oxygen releasing compounds (ORCs) has been
adopted to overcome this problem; however certain drawbacks to their use have been
identified. Hydrogen peroxide can cause complete microbial inhibition at concentrations
of 200 mg L-1 or higher [23] and ORCs, such as calcium and magnesium peroxide,
generate one mol of oxygen per two moles of the ORCs.
Anaerobic degradation for treating groundwater then arises as a viable option.
Although degradation rates achieved under anaerobic conditions are lower than those
attained in the presence of oxygen, different electron acceptors can be used for the
removal of aromatic compounds. Degradation can occur under NO3-, SO42-, Fe3+, (Mn4+),
and methanogenic reducing conditions [24, 25]. The overall degradation pathway
involves the transformation of the aromatic molecule, resulting in the formation of a few
central intermediates. Benzoyl-coenzyme A (benzoyl-CoA) is the most common aromatic
intermediate formed and is further transformed to non-aromatic compounds via
hydrolysis. Subsequently reactions allow the formation of acetyl-coenzyme A
65
(acetyl-CoA) which is readily oxidized in the Krebs cycle. Bacteria belonging to the
Azoracus, Geobacter, Pseudomonas and Thaurea species are known to successfully
metabolize aromatic compounds; however complete mineraliztion is not always achieved,
especially in the case of benzene [26-29]. Degradation of toluene is enhanced by the
addition of fumarate to the methyl group, forming benzylsuccinate which is converted to
benzoyl-CoA by a series of reactions [30]. Likewise, fumarate is incorporated to one of
the methyl groups in xylene, resulting in the formation of methylbenzylsuccinate that is
further transformed to benzoyl-CoA [28]. Benzene has shown to be the most recalcitrant
of the BTX compounds under anaerobic conditions, thus the degradation pathway is not
well understood. It has been suggested that benzene can be transformed to phenol,
benzoate or toluene as intermediates, which are then transformed to benzoyl-CoA [31].
Chlorinated solvents, such as tetrachloroethene (PCE), trichloroethene (TCE),
cis-DCE, and vinyl chloride (VC), can be degraded in the absence of oxygen by
sequential reductive dechlorination reactions resulting in the formation of ethane and/or
ethane [32]. Under these conditions microorganisms transfer the electrons from organic
molecules or hydrogen to the chlorinated solvents, which are used as terminal electron
acceptors [33]. The rate of the dechlorination reactions decreases as the number of
chlorine atoms in the molecules decreases, which might explain the accumulation of
cis-DCE and VC in anaerobic contaminated sites [34]. However, complete dechlorination
has been achieved under methanogenic and SO42- reducing conditions [35]. Since
cis-DCE can be highly recalcitrant to anaerobic reductive dehalogenation, in this work a
different approach for its degradation was assessed in the absence of oxygen. The
66
removal of cis-DCE was investigated in anaerobic batch experiments, where it was
supplied as substrate. Only a few reports exist of the capability of certain microorganisms
to oxidize cis-DCE and VC under manganese and VC under iron reducing conditions to
CO2 [36].
The scope of this work was to evaluate the capability of microorganisms in
different inocula for degrading of benzene, toluene, m-xylene, and cis-DCE, using NO3and ClO3-as terminal electron acceptors, to consider their use in PRBs treating
contaminated groundwater. NO3- and ClO3- are convenient for their use in PRBs due to
their high solubility. Moreover, reduction of ClO3- results in the release of oxygen, which
could enhance further degradation of the contaminants.
2.3
Materials and Methods
2.3.1
Inocula sources
Six different inocula were used for the BTX and cis-DCE biodegradation
experiments. Granular methanogenic sludge from industrial upflow anaerobic sludge
blanket (UASB) wastewater treatment plants included: Nedalco (distillery, Bergen Op
Zoom, Holland), Aviko (starch processing, Steederen, Holland), and Mahou (brewery,
Guadalajara, Spain). Anaerobic digested sludge (ADS) and returned activated sludge
(RAS) used in some of the microcosms were obtained from a municipal wastewater plant,
67
Ina Road facility located in Tucson, Arizona. Finally, sediments employed in the
experiments were taken from a local pond (Agua Caliente park, Tucson, Arizona). The
selection of the inocula was based on their exposure to a wide diversity of organic
compounds, which increases the likelihood of finding BTX and cis-DCE degrading
bacteria. The sludge was stored in the refrigerator at 4°C.
2.3.2 Microcosm preparation
Biodegradation experiments were performed to determine the presence of
microorganisms capable of toluene, benzene, m-xylene or cis-DCE degradation. Each
experiment consisted of a full treatment (live inocula, alternative electron acceptor, and
contaminant), an abiotic control (no inoculum), a killed inoculum (poisoned inoculum)
and a methanogenic control (no alternative electron acceptor). Each treatment and
corresponding controls were run in duplicate. At first, the biotic control consisted of heat
killed inocula. For this purpose the sludge was autoclaved for three consecutive days for
30 min at 120°C. However, this protocol released chemical with surfactant properties into
the liquid interfering with the headspace measurements of the BTX and cis-DCE. To
avoid false readings from the disturbed partitioning of the substrates, killed controls were
then arranged by using poison. The killed control contained a final concentration of
0.50 g L-1 and 0.25 g L-1 of sodium azide (NaN3) and mercury sulfate (HgSO4),
respectively.
68
Microcosms were set in 160 mL (Wheaton, Millville, NJ, USA) glass bottles of
which 50 mL was occupied by the liquid phase and the rest, 110 mL, was comprised by
the headspace. According to the treatment, the liquid volume consisted of mineral media,
sludge, electron acceptor, poison, and substrate (Table 2.2). The mineral media contained
the following chemicals (g L-1): NH4Cl (0.028), KCl (0.027), K2HPO4 (0.017),
CaCl2 2H2O (0.001), MgCl2 6H2O (0.015), NaHCO3 (5.00), yeast extract (0.002), and
trace elements (0.1 mL L-1). The trace elements solution consisted of (g L-1):
FeCl3 4H2O (2.00), CoCl2 6H2O (2.00), MnCl2 4H2O (0.50), AlCl3 6H2O (0.09),
CuCl2 2H2O (0.03), ZnCl2 (0.05), H3BO3 (0.05), (NH4)6Mo7O2 4H2O (0.05), NaSeO3
5H2O (0.10), NiCl2 6H2O (0.05), EDTA (1.00), rezasurine (0.20) and HCl 36%
(1 mL L-1). The pH of the medium was adjusted between 6.8 to 7.2 before adding the
sodium bicarbonate.
The concentration of the granular sludge and pond sediments was set at 1.5 g and
10 g volatile suspended solids per liter (g VSS L-1), respectively; while for the more
diluted sludges such as RAS and ADS the concentration was set at 10% (v/v) with
respect to the liquid volume. A stock solution of electron acceptors (KNO3 and NaClO3)
was prepared based on the stoichiometric electron equivalence (e-eq) relationship
between the target substrate (BTX or cis-DCE) and the electron acceptor. After dilution
with the other liquid constituents of the microcosms, NO3- and ClO3- were in a 2.5 to 5
fold excess to ensure a sufficient concentration of electron acceptor to sustain the
degradation of the pollutants.
69
After filling the microcosms, the glass bottles were sealed using Teflon faced
butyl septa (Supelco, San Louis, MO) and aluminum crimps (Fisher Scientific,
Pittsburgh, PA). A mixed gas containing 80% N2 and 20% CO2 was flushed through the
treatments for 6 min to achieve anaerobic conditions. Benzene, toluene and cis-DCE were
introduced into treatments via concentrated stock solution, which was diluted in the
microcosms to achieve a designed total concentration ranging from 100 to 120 mg L-1,
depending on the substrate used. Only m-xylene was supplemented directly into
treatments to a designed total concentration of 120 mg L-1 with a 10 µL gas lock
chromatographic syringe (Hamilton, Reno, Nevada) due to its low aqueous solubility,
making the preparation of stock solution unfeasible. The total concentration of the
substrates in the treatments was calculated by applying Henry’s Law; which is a
thermodynamic relationship of the partitioning of a compound between the gas and liquid
phase. Thus, by knowing the concentration of the contaminant in the either the headspace
or liquid phase, it is possible to determine its total concentration. Treatments were
incubated at 30°C in the dark in a constant temperature room.
Different stoppers were tested to find the appropriate septa to be used in the
microcosm experiments in order to avoid the escaping of the aromatic compounds. For
this purpose 160 mL glass bottles, in duplicate, containing a toluene solution at a total
concentration ranging from 5 to 100 mg L-1, in 50 mL of liquid, were sealed with rubber,
Viton, and Teflon faced butyl stoppers (Supelco, San Louis, MO) and monitored
periodically. Treatments were incubated in the dark at 30 °C. After 110 days, bottles
sealed with Teflon faced butyl septa showed negligible variation in the total
70
concentration of toluene, whereas a significant decreased in toluene was observed in all
the treatments sealed with rubber and Viton stoppers (data not shown). These results
indicated that Teflon faced butyl stoppers were appropriate for their use in the
microcosms experiments.
2.3.3
Respikes and transfers
Fresh substrate was added to treatments that showed removal using a
chromatographic syringe in an effort to restore the initial concentration. The respiking of
the sealed bottles was performed in a fumehood. Microorganisms capable of degradation
were transferred to clean glass bottles containing a fresh liquid phase previously flushed
with N2/CO2 (80% / 20%) and sealed. Transfers were performed in an anaerobic chamber
(COY Laboratory Products, Grass Lake, MI) transferring 10% of total granules or 10% of
the liquid volume according to the characteristics of the sludge. After decanting the liquid
phase of the microcosms subjected to be transferred, the remaining granules were sucked
up with a 10 mL syringe and crushed. Then, new substrate was added in the fumehood.
71
2.3.4 Analytical procedure
Substrate degradation was monitored by taking 100 µL from the headspace with a
gas-tight syringe. BTX samples were injected into a gas chromatograph HP5890A Series
II (Agilent Technologies, Palo Alto, CA) equipped with a flame ionization detector (FID)
and a 30 m x 0.32 mm GS-GASPRO (J&W Scientific, Palo Alto, CA) column. The
carrier gas was helium at a flow rate of 32.45 mL min-1, while the split flow was
maintained at 12 mL min-1. Both inlet and detector temperature was set at 250°C and the
column temperature was 195°C. cis-DCE samples were analyzed in a gas chromatograph
HP5890E Series II (Agilent Technologies, Palo Alto, CA) equipped with an electron
capture detector (ECD) and a 30 m x 0.32 mm column GS-GASPRO (J & W Scientific,
Palo Alto, CA). The flow through the column was 18 mL min-1 and the split flow was
17.6 mL min-1. The temperature of the column was set at 195°C, while the temperature of
the inlet and the detector was 180°C and 275°C, respectively. The syringe was cleaned
after each injection using a syringe cleaner (Hamilton, Reno, NV) to avoid accumulation
of previous samples. The headspace concentration of the BTX and cis-DCE in the
treatments was determined by using standards of known concentrations. Standards were
prepared in 160 mL glass bottles and the headspace to water ratio was maintained the
same as that set for the treatments. The required concentration was achieved by adding a
known volume of a stock solution and amended with Milli-Q water to reach 50 mL of
liquid. The total concentration in the treatments was calculated using the Henry’s Law
constant (H). The H values were obtained at 25°C [2-5] and corrected to for 30°C. The
72
dimensionless H for benzene, toluene, m-xylene and cis-DCE used were 0.279, 0.284,
0.391 and 0.196, respectively.
NO3-, NO2-, and ClO3- were measured by ion chromatography using a
Dionex (Sunnyvale, CA) chromatograph equipped with an anion self-regenerating
suppressor (ASRS ULTRA II) and an Ionpac AS18 (4 x 250 mm) column. The eluent
was KOH (25 mM) at a rate of 1 mL min-1. Concentration in treatments was determined
by using external standards previously prepared.
Table 2.2. Description of treatments used in microcosm experiments
*
Electron
acceptor
Treatment
Medium
Substrate
Abiotic control
X
X
Killed control
X
X
X
X
Full treatment
X
X
X
X
Methanogenic*
X
X
X
no e-acceptor control
Sludge
NaN3
HgSO4
X
X
X
73
2.4
Results
2.4.1
Toluene degradation under nitrate reducing conditions
A series of screening experiments were set in using different inoculum sources to
determine the presence of toluene degrading bacteria. The expectation was that the
biodegradation would be stimulated by using NO3- as a terminal electron acceptor.
Figure 2.1 shows an example of the time course of toluene degradation by one of the
inocula (Nedalco sludge) used in the screening. The toluene concentration decreased with
time in the presence of inoculum and the electron acceptor. The treatment lacking NO3displayed very low activity. Toluene concentration remained constant in controls (abiotic
and killed) containing NO3-; indicating that toluene degradation was the result of a
biological reaction in the presence of NO3-. In the full treatment, toluene was removed in
6 days. Results indicate that Nedalco inoculum contained bacteria capable of toluene
degradation under denitrifying conditions. Some other inocula tested catalyzed toluene
removal as well, as shown in Table 2.3. To determine if the activity could be sustained,
the full treatments were respike with toluene.
In the microcosms where toluene degradation was observed, pure toluene was
spiked into the assays in order to confirm degradation and maintain toluene degrading
activity. As shown in Figure 2.1 toluene spiked into the full treatment of Nedalco sludge
was removed in about 10 days. The controls and methanogenic treatment showed no
decrease in toluene concentration over 18 days, including the 10 days following the
74
respike in the full treatment. The results demonstrate that the microbial activity could be
sustained beyond the five feedings. Figure 2.1 also shows that the total concentration of
toluene spiked was 5 times higher than the intended value (100 mg L-1). However, after a
few days, the concentration decreased to the intended value; suggesting that the fresh
substrate spiked needed a few days to achieve equilibrium with respect to partitioning
between headspace, water and free product toluene (the latter eventually dissolving).
Figure 2.1. Toluene degradation under nitrate reducing conditions using Nedalco sludge.
─♦─, Abiotic control; ─■─, No NO3-; ─▲─, NO3-; ─●─, Poisoned. Error bars represent
standard deviation of duplicate assays.
75
2.4.2
Toluene degradation under chlorate reducing conditions
A screening experiment was also conducted to determine if toluene degradation
under ClO3- reducing conditions could be observed with any of the several sources of
inoculum tested. An example of the screening results is shown in Figure 2.2 with Nedalco
sludge as an inoculum source. The figure shows that after 130 days negligible
degradation was observed in the full treatment containing ClO3- as an electron acceptor.
On the other hand, toluene was completely removed under methanogenic conditions after
a lag phase of 40 days. Since no degradation was observed in the killed and abiotic
controls, the observed reaction is considered to be due to biological activity under
methanogenic conditions.
On day 87 and again on day 123, pure toluene was spiked into the methanogenic
treatment, to an intended total toluene concentration of 180 mg L-1, in order to determine
if the microbial activity could be sustained. After each spike, toluene was allowed to
equilibrate and subsequently was removed in time periods of 10 to 20 days. The results
suggest that microbial activity in the methanogenic treatment was maintained after
several feedings.
76
Table 2.3 Microcosm experiments results for BTX and cis-DCE anaerobic degradation.
Inocula
Toluene
m-Xylene
Benzene
cis-DCE
CH4
NO3-
ClO3-
CH4
NO3-
ClO3-
CH4
NO3-
ClO3-
CH4
NO3-
ClO3-
Nedalco
+
+++
-
-
-
-
-
+/-**
-
NT
NT
NT
Aviko
-
++
NT
-
-
NT
-
-
NT
NT
NT
NT
Mahou
-
NT
-
-
NT
-
-
NT
+/-*
NT
NT
NT
PS
-
++
NT
+/-
-
NT
NT
NT
NT
-
-
-
ADS
-
-
-
-
-
-
-
-
+/-*
-
-
-
RAS
-
NT
-
-
-
-
-
NT
-
-
-
-
+++ loss before 6 days; ++ loss before 13 days (86% removal); + loss before 73 days;
+/- loss in one duplicate after 244 d, on the other after 153 day (88.69%); +/-* loss in one
duplicate after 413 d; +/-** loss in one duplicate after 716 d; - No degradation after
extended incubation; NT not tested.
2.4.3
Transferring of denitrifying cultures
The supernatant from treatments that showed complete toluene removal under
NO3- reducing conditions was transferred (10% v/v) in an effort to enrich the microbial
community. After several unsuccessful attempts it was decided that instead, intact and
crushed granules should be transferred (10% of total granules in treatments). Figure 2.3
shows that no lag time was observed for complete toluene removal under nitrate reducing
conditions when intact and crushed granules were used. The activity in the full treatments
could be sustained.
77
Figure 2.2. Evidence of toluene degradation under methanogenic conditions using
Nedalco sludge. ─♦─, Abiotic control; ─■─, No ClO3-; ─▲─, ClO3-; ─●─, Poisoned.
Error bars represent standard deviation of duplicate assays.
After the full treatments were respiked and the spike equilibrated, toluene was
removed. The rates however were different in the treatments containing intact granules
compared to crushed granules, corresponding to 4,117.23 and 7,477.65 mg g VSS-1·d-1,
respectively; in the time period from day 9 to day 16. Furthermore, complete removal
was only achieved with crushed granules by day 30 when the experiment was terminated.
The toluene removal reaction seems to be biological since no degradation was observed
in both controls, killed and abiotic. These results indicate that bacteria responsible of
78
toluene removal were embedded in the granule sludge. Figure 2.3 shows that toluene was
removed initially in the methanogenic treatments when using crushed granules; however
this activity was not sustained since no further degradation was observed in about 20 days
after the fresh toluene was respiked.
Figure 2.3. Transfer of granular sludge containing toluene degrading microorganisms
(Nedalco sludge); intact vs crushed granules. ─♦─, Abiotic control; ─■─, No NO3(crushed); ─▲─, NO3- (crushed); ─Δ─, NO3- (intact); ─●─, Poisoned (crushed). Error
bars represent standard deviation of duplicate assays.
79
2.4.4 Transferring of methanogenic cultures
Nedalco granules containing bacteria capable of toluene degradation under
methanogenic conditions were crushed and transferred (10% of total granules in
treatments) into fresh media in an effort to enrich them. Growth conditions with and
without NO3- as electron acceptor were compared. Toluene was rapidly degraded in the
treatment containing NO3-, as can be seen in Figure 2.4. After the initial spike at the
beginning of the experiment, all the toluene was already removed after the first
measurement on day 3. In order to determine if activity could be sustained pure toluene
was respiked into the cultures to an intended total concentration of 200 mg L-1. The
microbial activity was clearly sustained since in the nitrate reducing treatments, toluene
was removed.
However, the degradation after each respike took longer than the previous one,
indicating that no enrichment of toluene degraders was occurring. Removal was also
observed in the methanogenic treatment; however the activity was significantly slower
compared with the full (NO3--amended) treatment. The initial toluene spike at the
beginning of the experiment was not completely degraded after 40 days, when the
experiment was ended. The results indicate that degradation is biological, since abiotic
and killed controls containing NO3- presented negligible activity. NO3- appears to be an
important electron acceptor for the biological degradation of toluene, since its presence
dramatically increased the microbial activity towards toluene removal.
80
Figure 2.4. Transfer of toluene degrading microorganisms (Nedalco sludge. 10% of total
crushed granules previously adapted to methanogenic conditions). ─♦─, Abiotic control;
─■─, No NO3-; ─▲─, NO3-; ─●─, Poisoned. Error bars represent standard deviation of
duplicate assays.
81
2.4.5
Nitrate consumption and nitrite formation linked to toluene degradation
The liquid phase of the controls and treatments shown in Figure 2.4, were
sampled to determine the fate of both nitrate and nitrite. Figure 2.5a shows that NO3concentration decreased with time in the full treatment where toluene was rapidly
degraded. As expected, no NO3- was detected in the methanogenic treatment since that
treatment received no added NO3-. The NO3- concentration in the abiotic and poisoned
controls showed very little variation and no evidence of consumption was observed. On
day 16, when toluene and NO3- were respiked, it can be observed that the NO3concentration continued to drop. NO3- consumption was associated with the formation of
NO2- (Figure 2.5b). An analysis was made to compare the consumption of toluene with
the consumption of NO3- and formation of NO2- over the first 3-days of the experiment,
based on the concentration of total electron milliequivalents (e-meq) transformed.
82
Figure 2.5. Consumption of a) NO3- and formation of b) NO2- during the anaerobic
degradation of toluene. NO3- spiked on day 16. Nedalco sludge. ─♦─, Abiotic control;
─■─, No NO3-; ─▲─, NO3-; ─●─, Poisoned. Error bars represent standard deviation of
duplicate assays.
83
Data obtained from samples taken during this period showed that toluene was
consumed from an initial total concentration of 165.79 mg L-1 (1.80 mM) to levels below
the detection limit (0.25 mg L-1), which corresponded to 64.80 e-meq L-1.
Simultaneously, 263.55 mg NO3- L-1 (2.25 mM) were consumed and 92.19 mg NO2- L-1
(2.00 mM) were generated. As consequence, 4.00 e-meq L-1 were involved in the
formation of NO2- linked to NO3-reduction during toluene degradation. The remaining
moles of NO3- consumed, were assumed to be fully reduced to N2, corresponding to
11.25 e-meq L-1. This analysis is shown in Table 2.4 and indicates that only 23.5% of
electron equivalents in the toluene removed over the first three days could be accounted
for by metabolism of NOx-, assuming the toluene would have been converted to CO2 by
the denitrifying bacteria.
Table 2.4. Electron-milliequivalent balance over the first 3-days from the transfer
experiment shown in Figure 5.
Redox couples
mM of NO3transformed
Toluene to CO2
electron m-eq / L *
Total e- meq / L
64.80
64.80
NO3- to N2
2.25
11.25
NO3- to NO2-
2.00
4.00
15.25
*Milliequivalents of toluene converted to CO2 (ideally) = 64.80 e-meq L-1.
Milliequivalents of NO3- converted to N2 = 11.25 e-meq L-1. Milliequivalents of NO3converted to NO2- = 4.00 e-meq L-1. Analysis performed from day 0 to 3.
84
2.4.6
m-Xylene degradation under nitrate reducing conditions
A series of experiments were conducted to determine the presence of m-xylene
degrading bacteria in different inocula. Figure 2.6 shows that no degradation occurred in
the NO3- containing treatment inoculated with pond sediment. This was also true for all
the other inocula tested. On the other hand, complete removal of m-xylene was achieved
under methanogenic conditions but only with pond sediment inoculum (Figure 2.6).
Degradation commenced after a 100 days of lag phase. Since there was negligible activity
in the controls the degradation in the methanogenic treatment was most likely due to a
biotic reaction.
In order to determine if the microbial activity could be sustained, a single respike
of pure m-xylene was added to the methanogenic treatment on day 485. Figure 2.6 shows
that after the respike, m-xylene was degraded. The rate of degradation was greater than
during the initial feeding. During the removal of the m-xylene respiked into the
methanogenic treatment, there was little removal in the controls and in the NO3treatment. The spiking data demonstrate that the microbial activity was enhanced after
exposure to the compound.
85
Figure 2.6. Anaerobic degradation of m-xylene under methanogenic conditions. Pond
sediments (Agua Caliente park). ─♦─, Abiotic control; ─■─, No NO3-; ─▲─, NO3-;
─●─, Poisoned. Error bars represent standard deviation of duplicate assays, except for
─■─.
2.4.7
Benzene degradation under nitrate and chlorate reducing conditions
A series of screening experiments were set up for the degradation of benzene
under anaerobic conditions using different inocula and electron acceptors. For most
inocula no benzene degradation was observed after extended incubation periods of up to
2 years. However, benzene removal was observed in one duplicate, under nitrate reducing
86
conditions after 420 days with the Nedalco inoculum (Figure 2.7). The lack of
degradation in the methanogenic treatment suggests that NO3- was critical as an electron
acceptor for the oxidation of benzene. Since negligible activity was observed in the
abiotic and poisoned controls, the removal of benzene in the NO3- treatment was
biological. The data indicate the presence of benzene degrading bacteria in some of the
inocula tested.
However, a long time period was required for the population of benzene
degraders to become enriched. In two of the inocula tested (ADS and Mahou), benzene
removal was observed in one duplicate each when ClO3- was used as electron acceptor.
Figure 2.8 shows the results with anaerobic digested sludge. After a lag phase of 111
days, the benzene concentration in the full treatment decreased considerably. The
degradation reaction was microbially mediated since negligible benzene removal was
observed in the controls. The presence of chlorate was a limiting factor since no
degradation was observed in the methanogenic treatment after more than 400 day.
87
Figure 2.7. Benzene degradation under nitrate reducing conditions. Nedalco sludge.
─♦─, Abiotic control; ─■─, No NO3-; ─▲─, NO3-; ─●─, Poisoned. Error bars represent
standard deviation of duplicate assays, except for ─▲─.
2.4.8 Experiments summary
Table 2.3 provides an overview of the biodegradation experiments performed and
which combinations of inoculum and electron acceptor resulted in biodegradation. Of all
the compounds tested, toluene was the easiest to degrade. It was rapidly removed in
treatments containing NO3- for the majority of the inocula tested. On the other hand,
methanogenic degradation was observed after extended incubation in only one inoculum.
The use of ClO3- as electron acceptor did not help with the degradation of this
contaminant. m-Xylene proved to be recalcitrant under NO3- and ClO3- reducing
88
condition since no removal was observed. Nonetheless, m-xylene was degraded under
methanogenic conditions after a long incubation period but only with pond sediments as
inoculum. The trend with benzene was the opposite of that observed for m-xylene, since
no degradation was observed under methanogenic conditions. Instead removal linked to
NO3- reduction was achieved in one of the inocula tested; and removal linked to ClO3reduction was observed in two inocula. However very extended incubation periods were
needed to observe benzene degradation. cis-DCE was tested in RAS, ADS, and pond
sediments and no degradation was observed after extended incubation when using NO3or ClO3-. There was also no evidence of cis-DCE removal with the methanogenic
treatments.
Figure 2.8. Benzene removal using chlorate as terminal electron acceptor. Anaerobic
digested sludge. ─♦─, Abiotic control; ─■─, No ClO3-; ─▲─, ClO3-; ─●─, Poisoned.
Error bars represent standard deviation of duplicate assays, except for ─▲─.
89
2.5
Discussion
2.5.1 Toluene degradation
In this study toluene, m-xylene, and benzene were shown to be degraded by some
anaerobic inocula tested under denitrifying conditions and, in some cases, under
methanogenic conditions. The degradation of these BTX compounds only occurred in the
presence of unadapted live inocula. There was no removal in killed controls or in
non-noculated controls, indicating that degradation was microbially mediated. Thus,
some of the inocula tested contained microorganisms capable of BTX degradation under
anaerobic conditions.
Toluene was removed under methanogenic conditions after 40 days when using
Nedalco sludge. Treatments containing the same inocula were capable of completely
removing toluene in 6 days in the presence of NO3-. The data demonstrate that although
toluene degradation occurs in the absence of external electron acceptor after extended
incubation, the presence of an exogenous electron acceptor stimulated the anoxic
degradation. The fact that NO3- caused an increase of the degradation rate of toluene
compared to methanogenic conditions may be due to the higher energy yield of
denitrification compared to methanogenesis. Table 2.5 shows the calculated standard
delta Gibbs energy (ΔG0’) for different redox reactions. This behavior was also observed
by Phelps and Young [37] in polluted sediments from a freshwater site.
90
Table 2.5. Energy associated to toluene oxidation.
Reaction
ΔG0 (KJ mol-1)
C7H8 + 7.2 NO3- + 0.2 H+
7 HCO3- + 3.6 N2 + 0.6 H2O
-3,547
C7H8 + 18 NO3- + 3 H2O
7 HCO3- + 18 NO2- + 7 H+
-2,456
C7H8 + 12 NO2- + 5 H+
7 HCO3- + 6 N2 + 3 H2O
-4,283
C7H8 + 7.5 H2O
2.5 HCO3- + 4.5 CH4 + 2.5 H+
-144
The data for computing the ΔG0 values of the reactions was taken from Reittmann and
McCarty [38] and Thauer et al [39].
Nonetheless, toluene degrading microorganisms were not found in all the inocula
tested. Only half of the inocula samples used in the experiments were capable of toluene
degradation which is consistent with previous reports that toluene can be recalcitrant
under anaerobic conditions in pristine soils [37]. Nonetheless, there are several reports of
toluene removal by unpolluted sediments under nitrate reducing conditions [40, 41], as
was demonstrated in this study. While toluene oxidation coupled to denitrification has
been widely studied, the use of ClO3- as electron acceptor has not been evaluated as
thoroughly. Extended incubation of the unadapted inocula did not result in an accelerated
degradation of toluene linked to ClO3- reduction. Therefore microorganisms capable of
utilizing chlorate for BTX degradation were not found in the pristine inocula used in this
study. Logan and Wu [42] demonstrated, in column experiments inoculated with ADS,
91
that ClO3- can enhance toluene biodegradation; however enrichments were not obtained
when using ClO3- as the sole electron acceptor.
Treatments that showed degradation were spiked with fresh substrate to evaluate
if the microbial activity could be sustained. After injection of new substrate into the
bottles, the initial measured concentration was much higher than the intended
concentration. A few days later the concentrations approached the intended values. This
behavior is attributed to a certain amount of time required for the toluene to equilibrate in
concentration between the liquid and headspace. The concentrations measured were
based on headspace analysis. After repeated spiking, the norm was an increase in time
required for the removal of the fresh substrate after each addition, suggesting loss of
required growth factors over time or toluene toxicity.
2.5.2
Transfer of microorganisms
The first attempts to transfer bacteria capable of toluene oxidation consisted in the
transfer of microorganisms suspended in the supernatant to fresh media from treatments
that showed degradation. However, this approach did not result in degradation of toluene.
One possibility for the lack of degradation in the new medium is that toluene degradation
may be due to cometabolism, and co-substrates present in the sludge granules were lost
by transferring only the supernatant. Cometabolic oxidation of toluene was observed by
Rabus and Heider [43] by ethylbenzene adapted bacteria. The lack of degradation could
also be attributed to the absence of endogenous organic matter, otherwise present in the
92
granules, which may have helped increase toluene degradation by generating fumurate
during organic matter decomposition. Fumurate is an essential reactant for the initiation
of anaerobic toluene degradation [44] and it is known to enhance the degradation. The
results observed could indicate as well that microorganisms capable of toluene removal
might be attached to granules. To evaluate this concept, both intact and crushed granules
were transferred into fresh media.
Toluene removal was observed in both cases; however the degradation rate of
crushed granules was higher than that showed by intact granules. accounting possible
explanation for this behavior is that crushing the granules might increase the mass
transfer of toluene. Microorganisms capable of toluene oxidation might be clustered
inside the granules. If so, the difference of the degradation rate could be the result of
faster diffusion of toluene to smaller broken up granules. Despite the fact that the first
transfer worked in both cases, it was not possible to enrich the cultures for toluene
degradation. This could indicate that organic matter present in the granules is not the sole
factor needed to sustain toluene degradation. Clearly other factors are important, such as
undefined growth factors. Likewise, it is possible that continued exposure to toluene
could be toxic to the responsible microorganisms, which is known to have inhibitory
effects to certain organisms [45, 46].
93
2.5.3
Nitrate consumption
A complete analysis for toluene, NO3-, and NO2- was performed for one the
transfer experiments. No NO2- was fed into the assay; however toluene removal linked to
NO3- reduction resulted in formation of NO2-. Its formation coincided with NO3consumption indicating incomplete denitrification. NO2- accumulated for a few days until
it reached a maximum value. Accumulation of NO2- from toluene denitrification has been
observed previously [47]. The following days the NO2- concentration decreased
presumably due to complete denitrification to N2. Evans et al [48] observed NO2reduction during toluene denitrification when the NO3- concentration was nearly
depleted. In this study an electron equivalent balance was performed for toluene
oxidation and NO3- reduction from day 0 to day 3. The analysis indicated that 23.5% of
the electron equivalents in toluene were accounted for by NO3- conversion considering
complete oxidation of toluene to CO2. This low correlation could indicate partial
oxidation of toluene in the treatments as observed by Evans et al [49] who found
formation of two recalcitrant byproducts, benzylsuccinic acid and benzylfumaric acid,
during toluene oxidation linked to denitrification when using strain T1, an anaerobic
microorganism capable of toluene oxidation under denitrifying conditions. Another
possibility would be the use of toluene or its partially oxidized products by methanogens.
Lastly, it should be noted that a greater proportion of the electron equivalents in toluene
would be accounted for by NO3- reduction, if dissimilatory nitrate reduction to
94
ammonium (DNRA) had taken place. DNRA is known to occur in methanogenic sludge
[50].
2.5.4
m-Xylene
In this study, removal of m-xylene was only observed under methanogenic
conditions after an extended incubation time when using non-polluted pond sediments
from Agua Caliente Park. There was no degradation in treatments containing neither
NO3- nor ClO3-, indicating that probably bacteria capable of utilizing these compounds as
electron acceptors for the oxidation of m-xylene might have not been present in the
sediments. It is not clear what factors enhance bioremediation of m-xylene under
anaerobic conditions. Although natural attenuation of m-xylene in contaminated sites has
been reported [51], anaerobic degradation is not observed in many sediments, sludge, or
water samples from both contaminated and pristine sediments. Biotransformation is most
likely limited to a few enrichment cultures [52] or bacterial strains [53]. Nontheless,
degradation of this contaminant using uncontaminated sediments has been reported by
Rabus and Widdel [54] who were able to isolate the denitrifying strain mXyN1, that
grows only in the presence of m-xylene, from fresh mud.
95
2.5.5
Benzene
Benzene is the most recalcitrant of the BTX compounds under anaerobic
conditions; the lack of functional groups in the molecule makes it extremely stable which
could explain why there was no degradation under methanogenic conditions after
extended incubations. However, the use of alternative electron acceptors promoted the
removal of this contaminant. In this study, treatments containing NO3- and ClO3enhanced the bioremediation after more than 400 days in the fastest treatment
(Figure 2.8). Data obtained from batch experiments suggest that microorganisms capable
of degradation were found in sludge from municipal and industrial treatment plants as
well as natural pond sediments. In general, removal of benzene in anaerobic treatments
requires long incubation periods, even when utilizing hydrocarbon contaminated
sediments as inoculum. Kazumi et al [55] reported benzene degradation ranging from 180
to 590 days under different redox conditions. Rapid degradation was observed by Botton
and Parsons [56] under iron reducing conditions; around 10 µM where removed in about
70 days when using an enrichment culture derived from a contaminated site.
96
2.5.6
cis-DCE
The low oxidation state of the carbon atoms in lower chlorinated aliphatic
compounds makes them recalcitrant under anaerobic environments. The most common
approach to biologically degrade these contaminants is via reductive dehalogenation
where chlorine atoms are replaced for hydrogen to ideally form ethene. Chlorinated
ethenes have been removed from contaminated sediments under anaerobic conditions.
Lendvay et al [57] compared the effect of bioaugmentation versus biostimulation.
Sediments amended with Dehalococcoides did initiated degradation without a lag phase;
meanwhile it took three months to start observing degradation in sediments amended with
lactate and nutrients but not bioaugmented. Dechlorination can also occur in pristine
sediments as demonstrated by He at al. [58] who were able to isolate a strain capable of
reductively transforming tetracholoethene to ethene, denominated Dehalococcoides sp.
strain FL2 from river sediments not contaminated with chlorinated solvents. However,
while microbial reduction rates of PCE to cis-DCE are high, transformation of cis-DCE
does not always occurs, and when it does, it is slower.
Some microorganisms can oxidize chlorinated hydrocarbons to carbon dioxide.
Chapelle and Bradley [59] documented natural attenuation of chlorinated solvents in
contaminated groundwater. They observed reduction of PCE and TCE to cis-DCE, and
VC in the SO42- reducing conditions zone, which were further oxidized to carbon dioxide
and chloride in the iron reducing zone. In this project the anaerobic degradation of
cis-DCE was also studied; the approach employed attempted to anoxically oxidize the
97
contaminant so that it would serve as a carbon and energy source to promote growth of
microorganisms. For this purpose, sludge from municipal wastewater treatment plants
and pond sediments were used in microcosm experiments. No degradation was observed
after extended incubation times in any of the treatments, which included methanogenic,
nitrate reducing, and chlorate reducing conditions.
2.6
Conclusions
Microorganisms capable of toluene oxidation under nitrate and methanogenic
reducing conditions were found in many of the inocula tested; however no removal was
observed when ClO3- was used as electron acceptor. m-Xylene and cis-DCE proved to be
recalcitrant under anoxic environments; bacteria capable of gaining energy from the
reduction of NO3- and ClO3- linked to the oxidation of these contaminants were not found
in the inocula studied. However, m-xylene was removed under methanogenic reducing
conditions. Benzene, on the other hand, was oxidized in the presence of NO3- and ClO3-,
yet the oxidation was observed only in one of the duplicates in all the cases.
Despite having observed the anoxic oxidation of BTX, microorganisms capable of
performing this task were not present in all of uncontaminated inocula tested. Toluene,
which was the fastest contaminant to be degraded, was only easily removed by three
different sludges. Of the inocula tested, none was suitable for application in PRBs to
support the anoxic oxidation of a mixture of aromatic and chlorinated contaminants.
98
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106
CHAPTER 3
BIOLOGICAL TREATMENT OF A SYNTHETIC LANDFILL LEACHATE IN A
COMBINED ANAEROBIC – AEROBIC SYSTEM
3.1
Abstract
Landfilling is the preferred alternative worldwide for disposing household waste
due to its simplicity and low cost, compared to other options. However, landfills pose a
risk to surrounding water bodies due to the formation of leachates that can contaminate
surface water and groundwater. Landfill leachates can contain numerous organic and
inorganic contaminants since they are formed as percolating water mixes with chemicals
leached out from the buried waste. The objective of this work was to evaluate the
performance of a system, consisting of an upflow anaerobic sludge blanket (UASB)
reactor and a down flow hanging sponge (DHS) column, for the treatment of a synthetic
landfill leachate with a chemical oxygen demand (COD) and ammonia (NH4+) content
comparable to those found in real leachates. The organic contaminants in the synthetic
leachate, comprised by acetic, propionic, butyric and valeric acid, phenol, and p-cresol,
were successfully removed during the anaerobic treatment in the UASB reactor. The
hydraulic retention time (HRT) in the reactor was decreased after 31 days from 48 to 24
h, which resulted in an increase of the organic loading rate (OLR) from an average of
107
1.04 + 0.46 to 2.3 + 0.56 g COD-CH4 L-1 d-1. This change did not seem to affect the
removal of the pollutants since the efficiency of the reactor remained constant. Mass
balance calculations showed good correlation between the COD removed and the amount
of methane (CH4) gas formed in the reactor. COD removal increased with time, reaching
up to 98% elimination. No NH4+ removal was observed in the UASB reactor. The aerobic
treatment was conducted in a DHS column, started 41 days after the UASB. High
nitrification rates were observed during the operation of the reactor. The average loading
rate of NH4+ to the DHS, which operated at an HRT of 1 day, was
180.9 + 41.6 mg NH4+-N L-1 d-1. During the last 79 days of operation, an average 82% of
the NH4+ entering the reactor was converted to nitrate (NO3-). Nitrite (NO2-) was formed
during the oxidation of NH4+ as it was frequently detected in the effluent, suggesting that
partial nitrification could be achieved by controlling the aeration rate. The UASB-DHS
system proved to be adequate for the removal of organic contaminants and nitrification of
the synthetic leachate.
108
3.2
Introduction
The increment of waste generation is strongly linked to the progress of the
modern lifestyle. Nowadays, Americans generate about three times more solid waste than
in the 1960’s and the most common practice is to discard it by disposal into municipal
solid waste (MSW) landfills. Of the 251 million tons of refuse generated in 2006,
approximately 12.5% was burned and 55% landfilled, and the rest, 32.5%, was recycled
[1]. Landfilling is so widely used that it is estimated that around the world 70% of
household waste is disposed of in landfills [2]. Despite the extensive use of landfills, they
have been recognized to pose a high threat to air quality [3] as well as surface- and
groundwater quality [4]. Water contamination has been associated with the generation of
leachates, which result from the mixing of percolating water with the chemicals present
in the landfilled waste. Although engineered landfills are utilized to minimize this risk,
they can be subject to failure. High density polyethylene and geotextile liners can become
deteriorated by the high temperatures reached inside the landfill [5]. Moreover, old
landfills were built without liners and collection systems [6], increasing the likelihood of
adverse environmental impacts to surrounding waters. Different studies have
demonstrated the negative effects of landfill leachates to groundwater [7-9]. Leachates
are generally high in pH and rich in COD and NH4+. They are known to cause
eutrophication and saprobibation to receiving waters [10] and they proved to be toxic to
certain living organisms [11, 12].
109
The composition of the leachates depends on the characteristics of the landfill,
composition of the buried waste, and the environmental conditions of the location [13].
Contaminants in leachates can be grouped into four major categories, consisting in
dissolved organic matter (DOM), inorganic macrocomponents, heavy metals, and
xenobiotic organic compounds [14]. Table 3.1 shows the concentration ranges of
pollutants commonly found in landfill leachates. Volatile fatty acids (VFA) are normally
present at the highest concentration among organic compounds, as they are generated
during the decomposition of proteins, lipids, and carbohydrates [15]. Other organic
pollutants commonly present in leachates are low molecular weight (LMW) alcohols and
amines, phenols, aliphatic and aromatic hydrocarbons and volatile esters [16].
Occurrence of phenols in leachates results from their presence in certain household and
hazardous products and from the degradation of the lining [17]. Toxic chemicals from
plasticizers, pharmaceuticals, solvents, and oils have been identified in landfill leachates
as well [18]. Different studies have proved the toxicity of leachates to certain test
organisms, including aerobic bacteria (Salmonella typhimurium and Photobacterium
phosphorium), zooplankton (Daphnia magna and Daphnia pulex), green algae
(Selenastrum capricornutum) and rainbow trout [11, 12, 19]. Given the negative effects
of these contaminants on public health and the environment, treatment of landfill
leachates is an important matter. Many techniques have been applied for the treatment of
the leachates, including aerobic and anaerobic biological processes, adsorption onto
activated
carbon,
flocculation-coagulation,
recirculation,
ozonation,
and
new
technologies such as reverse osmosis and filtration [20, 21]. In general, biological
110
processes have shown to be effective in removing organic compounds and nitrogen (N) in
young leachates at a low cost [22]. Compared to anaerobic processes, aerobic treatment
requires more energy and generates more waste sludge [23]. During anaerobic digestion
complex molecules are converted to carbon dioxide (CO2), methane (CH4), and NH4+
after multiple intermediate steps [24].
Table 3.1. Concentration range of contaminants commonly found in landfill leachates.
Group
Contaminant
Concentration
range (mg/L)
Reference
Biological Oxygen
Demand (BOD5)
20 - 570,000
100 - 90,000
3 - 25,000
[6]
[25]
[22]
Chemical Oxygen
Demand (COD)
140 - 152,000
150 - 100,000
556 - 70,900
1,700
[6]
[25]
[22]
[26]
50 - 2,200
[6]
1 - 1,500
0.2 - 13,000
85
[25]
[22]
[26]
Phenol
0.6 – 1,200
0.6 – 2.2
30 – 17,000
[6]
[27]
[5]
Cresols
1 – 2,100
0.12 – 60.9
2,100
[6]
[27]
[5]
Organic matter
Inorganic
components
Ammonium-N
Organic components
(µg L-1)
111
Short chain VFAs such as propionate, butyrate, and valerate are converted by
acetogenic bacteria to acetate, hydrogen gas (H2), and CO2 depending on the parent
compound. A low partial pressure of H2 is essential for the acetogenic process [28].
Finally CH4 is produced from acetate by acetoclastic bacteria and from CO2 and H2 by
reductive methanogenesis [29]. Evidence exist that most of the CH4 generated during
anaerobic digestion is produced from acetate reduction [30]. Phenol can be degraded
anaerobically as well, and its biodegradation by a methanogenic consortia was reported
many years ago [31]. Complete mineralization of phenol and p-cresol was observed in
sewage sludge and their transformation resulted in generation of CH4 [32]. Benzoate and
acetate were detected as intermediate products during phenol degradation which were
further transformed to CH4 and CO2 [33].
The utilization of anaerobic processes for wastewater treatment has gained
importance due to the development of the upflow anaerobic sludge blanket (UASB)
reactor, which allows treatment of industrial and domestic wastewater at low cost [34].
The system relies in a high sludge concentration in the reactor that results in capability
for treating high organic loadings [35]. Moreover, the sludge biomass has a high
settleability preventing loss of biomass due to washout and produced biogas that can be
used to generate energy. Due to their simplicity, efficiency and cost, full scale UASBs
have been successfully implemented in Mexico, Colombia, Brazil and Uruguay, among
others [36]. Alternatively, the utilization of the UASB reactor for treating landfill
leachate has been investigated. Kennedy [37] studied the efficiency of batch and
continuous UASB reactors. Although performance was similar at low to medium OLRs,
112
continuous UASB performed better at high OLRs than the sequencing batch treatment.
The performance of the reactors for COD removal varied between 71 and 92% depending
on the OLR and the concentration of the leachate. UASB reactors treating industrial and
domestic wastewater are normally operated at 30 to 35oC in order to achieve high
efficiencies. A pilot UASB treating a municipal landfill leachate showed the dependence
of the efficiency of the reactor with temperature; where the COD removal reported
between 18 to 23oC was 65-75% COD whereas that at 13-14oC only 50-55% [25].
Although effluents from UASB reactors contain low organic concentrations, their N
content remains fairly constant during treatment and rarely comply with discharging
standards [38], and require pos-treatment.
Methanogenic conditions are not appropriate for removing NH4+. For this
purpose, aerobic processes are desired after anaerobic degradation. Activated sludge,
trickling filters or rotating biological contactors are commonly used as post-treatment
alternatives [22]. The downflow hanging sponge (DHS) reactor has recently emerged as a
simple and low cost option for aerobic post-treatment. The DHS reactor was designed to
treat UASB effluents from municipal wastewater treatment plants (WWTPs) and consists
of polyurethane rectangles as packing material where bacteria can attach and grow [39].
Machdar et al [40] installed a DHS reactor fed with treated wastewater from a UASB in a
sewage treatment site in Japan. The combined system, with an overall HRT of 8 h,
achieved 81-84% total COD removal; however the nitrification was not as efficient
reaching only a maximum NH4+-N elimination of 61%. A similar arrangement system
was tested in a municipal wastewater treatment plant in Japan, achieving higher quality of
113
the treated effluent [41]. The average total COD removal was equal to 90% at an overall
HRT of 10.7 h, whereas up to 86% elimination of NH4+ was achieved in the DHS reactor.
Moreover, Agrawal et al [42] documented up to 84% N removal in a DHS connected to
the effluent of an UASB treating raw sewage after post-denitrification. In this case, N
elimination resulted from the combined activity of nitrification and denitrification.
The objective of this work was to study the COD and NH4+ removal achieved by
a system of laboratory-scale reactors consisting of an UASB followed by a DHS fed with
a synthetic media rich in VFA, NH4+, and phenolic compounds. The role of a DHS
reactor to polish the COD of the UASB effluent and nitrify NH4+-N was also evaluated.
3.3
Materials and Methods
Two laboratoty-scale reactors, one anaerobic and one aerobic, were set up in
series to study the removal of COD and NH4+ from a synthetic landfill leachate. Different
analyses were performed to evaluate the treatability of the leachate and the efficiency of
both reactors.
3.3.1 Anaerobic reactor
The anaerobic UASB reactor, having a total and working volume of 420 and
401 mL, respectively, was made of glass (Figure 3.1). The maximum diameter and height
were 52 mm and 338 mm, respectively, with the effluent port located 209 mm above the
114
base. An opening located at the top of the reactor was used for collecting the gases. The
reactor was packed with methanogenic granular sludge obtained from a distillery
wastewater treatment plant located at Nedalco distillery in Bergen Op Zoom, Holland.
The total mass of sludge, which was determined by its volatile suspended solids (VSS)
content, was equal to 121.8 g and once packed occupied less than 1/3 of the reactor’s
working volume. The UASB was operated with a flow rate corresponding to a 48 h HRT
for the first 31 days of operation, at which time the HRT was decreased to 24 h. The OLR
increased from 1.04 + 0.46 to 2.30 + 0.56 g COD-CH4 L-1 d-1.
Figure 3.1. Diagram of the upflow anaerobic sludge blanket reactor. (1) Influent
leachate, (2) sludge blanket, (3) gas effluent, (4) treated effluent.
115
3.3.2 Aerobic reactor
The aerobic reactor was started up 41 days after the starting of the UASB. It
consisted of a DHS column made of acrylic, having a total and working volume of 350
and 300 mL, respectively; with a height of 268 mm and a 50.8 mm diameter (Figure 3.2).
The influent was evenly distributed by a shower, located at the top of the reactor, made of
the same material as the reactor. Polyurethane sponge was cut in small rectangles of
1 x 0.5 cm and packed into the reactor until completely filling the working volume.
Aerobic returned activated sludge (RAS) from a local wastewater treatment plant (Roger
Road. Tucson, AZ) was used for seeding the reactor at a 10% (v/v) ratio. For this purpose
a sample of 100 mL of sludge, prior homogenization, was placed in a clean graduated
cylinder and allowed to sediment for 1 h. The supernatant liquid was decanted and the
sponge rectangles were mixed with 30 mL of the remaining precipitated material. A
down-flow stream of humidified air, at a constant flow rate of 5 mL s-1, was supplied
from a port located 2 cm from the top. The air was pumped by a commercial pump
(Aqua Culture, Walmart) through a clean Erlenmeyer beaker containing Milli-Q water.
Another port located at the bottom of the reactor was used for effluent recirculation at a
flow set 6 times higher than the influent. The retention time of the reactor was maintained
at 24 h.
116
Figure 3.2. Schematic diagram of the upflow hanging sponge reactor used in the
experiments. (1) Pre-treated leachate influent, (2) humidified air influent, (3) shower, (4)
polyurethane sponge, (5) humidified air exit, (6) recirculation, (7) treated effluent.
3.3.3. Synthetic leachate
The synthetic media resembled the COD and NH4+ concentrations normally found
in real landfill leachates; it contained as well macro- and micro-components needed for
bacterial growth. The organic components of the leachate were (g L-1): yeast extract
(0.020), acetic acid (0.585), propionic acid (0.217), butyric acid (0.500), valeric
acid (0.164), and phenol and p-cresol which concentration was increase from an average
0.03 to 0.90 g L-1 after 103 days of first being added into the leachate. The inorganic
117
fraction of the media consisted of (g L-1): CaCl2 2H2O (0.010), KH2PO4 (0.037),
MgSO4 7H2O (0.015), MgCl2 6H2O (0.078), NH4Cl (0.700), NaHCO3 (2.00), and trace
elements (1 mL L-1). The trace elements solution contained (g L-1): FeCl3 4H2O (2.00),
CoCl2 6H2O (2.00), MnCl2 4H2O (0.500), AlCl3 6H2O (0.090), CuCl2 2H2O (0.030),
ZnCl2 (0.050), H3BO3 (0.050), (NH4)6Mo7O2 4H2O (0.050), NaSeO3 5H2O (0.100),
NiCl2 6H2O (0.050), EDTA (1.00), rezasurine (0.200) and HCl 36% (1 mL L-1).
Chemicals were purchased from Sigma-Aldrich (St. Louis, MO). The pH of the synthetic
leachate was adjusted between 6.8 to 7.2 before adding the sodium bicarbonate. The
media was flushed with N2 gas for 30 min and then connected to a N2 bag to ensure
anaerobic conditions. The influent was kept at 4ºC for no longer than 4 days to avoid
degradation of the organic components.
3.3.4 Analytical procedure
Samples from three different locations were taken to monitor the efficiency of the
reactors. The influent was sampled directly from the leachate container and the effluent
of both reactors was sampled from the effluent lines.
In order to monitor the concentrations of the VFAs, phenol, and p-cresol, 1.5 mL
of each sample were centrifuged at 10,000 rpm for 15 min. Then, 1 mL of the supernatant
was transferred into 1.6 mL glass vials (Fisher Scientific, Pittsburgh, PA) and amended
with 10 μL formic acid (Sigma-Aldrich, St. Louis, MO). A volume of 0.5 μL of the
prepared samples were injected, with help of an autosampler, into a gas chromatograph
118
(Agilent Technologies 7890A, Palo Alto, CA) equipped with a flame ionization detector
(FID) and a 30 m x 0.53 mm column (Restek Stabilwax-DA, Bellefonte, PA). The carrier
gas was helium at a flow rate of 115 mL min-1 with a split ratio equal to 6:1. The inlet
temperature was set at 250ºC while the detector temperature was 275ºC. Standards
containing known concentrations of VFA, phenol, and p-cresol, amended with 10 μL
formic acid, were measured together with the samples.
The soluble COD fraction was determined by colorimetric analysis. 20 mL Kimax
test tubes (Fisher Scientific, Pittsburgh, PA), in duplicate, were filled with 2.5 mL of
sample previously passed through 200 nm filters (Fisher Scientific, Pittsburgh, PA) and
amended with 1.5 mL of digestion solution and 3 mL of concentrated sulfuric acid. The
digestion solution contained (g L-1): H2Cr2O7 (10.22), Hg2SO4 (33.30), and
H2SO4 (167 mL L-1). The sulfuric acid solution was prepared by dissolving 10.25 g of
Ag2SO4 in 1 L H2SO4. Tubes then were caped and placed in the oven for 2 h at 150ºC.
Chemicals used for COD measurements were purchased from Fisher Scientific
(Pittsburgh, PA). Two tubes containing only 2.5 mL of Milli-Q water, used as blanks,
were prepared along with the samples. Once at room temperature, the blanks and samples
were analyzed in a DU 530spectrophotometer (Beckman-Coulter, Fullerton, CA) at a
wavelength of 600 nm.
NH4+ was measured using a DC 218-NH4 NH4+ probe (Mettler Toledo,
Columbus, OH). For this purpose 2 mL of sample, prior centrifugation at 10,000 rpm for
15 min, were spiked with 40 μL of an ionic strength adjustor (ISA) buffer solution
(Mettler Toledo, Columbus, OH). The concentration read by the probe was based on a
119
calibration set before the analysis, where the standards were prepared following the same
procedure as that for the samples. The pH was measure using an Orion pH probe (Fisher
Scientific, Pittsburgh, PA). The probe was calibrated using standards (Thermo Scientific,
Waltham, MA) before measuring the samples.
NO3- and NO2- were measured by ion chromatography using a Dionex IC-300
system (Sunnyvale, CA) chromatograph equipped with an anion self-regenerating
suppressor (ASRS ULTRA II) and an Ionpac AS18 (4 x 250 mm) column. The eluent
concentration was 25 mM of KOH at a rate of 1 mL min-1. The instrument was calibrated
by using external standards.
CH4 production in the anaerobic reactor was monitored by liquid displacement.
The exhaust open was connected with plastic tubing to a 2% NaOH solution dyed with
methylene blue (Fisher Scientific, Pittsburgh, PA). The displaced liquid was captured in a
plastic container and weighted to determine the CH4 formation after several mathematical
calculations.
Finally, the flow of the anaerobic reactor was determined by measuring the
weight loss in the leachate container every delta time. The aerobic reactor flow was
monitored by measuring the effluent container every delta time.
120
3.4
Results
A UASB reactor packed with methanogenic granular sludge was set up to
evaluate the feasibility of removing COD of a synthetic landfill leachate composed of a
mixture of VFA and phenols. A DHS column was connected in series to the UASB to
treat the NH4+-rich effluent. The fate of the organic and inorganic components was
examined to evaluate the efficiency of both reactors.
3.4.1
UASB reactor
Initially, the UASB reactor was operated at a HRT of 48 h (period I), which was
decreased to 24 h on day 31 (period II). As a consequence, the OLR was increased from
an average of 1.04 + 0.46 to 2.3 + 0.56 g COD-CH4 L-1 d-1 as shown in Table 3.2,
together with the influent pH of the synthetic leachate and the volumetric CH4
production.
The COD removal indicted by difference in influent and effluent samples
demonstrated that the UASB was highly efficient for removing the organic content from
the media, achieving an average elimination above 88% and 98% during period I and II,
respectively. Figure 3.3 shows the concentration of COD measured in the samples. The
average COD during period I was 2.54 + 0.17 g COD L-1 in the influent and
0.30 + 0.10 g COD L-1 in the effluent; whereas for the second period an average of
2.43 + 0.43 and 0.05 + 0.01 g COD L-1 were measured in the influent and effluent of the
121
reactor, respectively. The decrease of the HRT caused an slight increase of COD in the
treated effluent as observed in Figure 3.3. However, the COD spike in the effluent was
negligible and microorganisms in the reactor adapted to the higher OLR in approximately
15 days. Moreover, extended operation time resulted in higher COD removal efficiencies,
as higher removal was achieved after day 73.
Generation of CH4 in the UASB was expected as it is the final product of the
methanogenic degradation of organic compounds, and the OLR as function of CH4
formed in the reactor is presented in Figure 3.4. The average volumetric CH4 production
obtained from CH4 measurements were higher than those obtained from direct COD
readings (Table 3.2), indicating effective conversion of organics in the media, as
previously discussed, and possible degradation of organic content in the sludge itself.
CH4 production increased as a more organic-rich leachate was fed to the reactor after day
124. Further analysis of individual contaminants were followed as well.
Table 3.2. pH, organic loading rate (OLR), and volumetric CH4 production in the UASB.
Period
HRT (h)
pH In
pH Out
OLR
(g COD / L d)
Volumetric CH4
production (g
CH4-COD / L d)
I
48
8.23 + 0.08
7.81 + 0.18
1.02 + 0.46
1.11 + 0.22
II
24
8.43 + 0.20
8.16 + 0.24
2.17 + 0.56
2.47 + 0.49
122
Figure 3.3. Fate of COD during anaerobic treatment in the UASB reactor, where (■)
represents the influent and (●) the effluent.
Influent and effluent samples were used to measure the fate of the VFA mixture,
phenol, and p-cresol. Figure 3.5 shows that regardless of the fluctuations of VFAs
concentration in the influent, the content in the UASB effluent was low. Moreover, it can
be observed that the effluent VFA measurements followed the same trend observed with
the COD measurements, where higher removal was documented after day 76. The
increase in the OLR from day 31 until the end of the experiment did not seem to have any
effect over the efficiency of the UASB reactor for removing the organic acids.
123
Figure 3.4. CH4 generation linked to the anaerobic degradation of the organic
components in the synthetic leachate during anaerobic treatment in the UASB reactor
during periods I and II.
The average VFA concentration in the influent during the operation of the reactor
was
2.16
+
0.45
g
VFA-COD
L-1,
whereas
that
of
the
effluent
was
0.88 + 0.22 g VFA-COD L-1. Thus, the overall average elimination of VFA reached 95%.
Analysis of each individual VFA (data not shown), indicated that mainly acetic acid was
present in the effluent. The synthetic leachate had an average measured concentration of
acetate of 633 + 97.2 mg L-1 and the average level detected in the effluent was
58.9 + 14.1 mg L-1, which represents approximately 90% elimination. Propionic, butyric,
124
and valeric acid were intermittently found in the treated effluent throughout the operation
of the reactor. Concentration of these acids in effluent samples containing these
compounds was higher in samples collected during the first weeks of reactor operation,
showing that microbial adaptation was important for the elimination of such
contaminants.
Figure 3.5. VFA concentration in the influent (■) and effluent (●) of the UASB reactor
as function of operation time.
125
Phenol and p-cresol were amended, via stock solution, to the synthetic leachate
after 20 days of operation. The concentration of both contaminants was then doubled
after day 124. Figure 3.6a shows that the bacteria in the sludge bed readily became
enriched with degraders of phenol since high removal efficiencies were observed
immediately after this compound was introduced into feed. It can be observed as well that
the reactor performance increased with time. Doubling the COD OLR, which occurred at
day 31, did not seem to have a significant effect over elimination of phenol as the effluent
concentration, 11.5 + 2.21 mg L-1, remained fairly constant from day 25 to day 37
achieving approximately 74% removal. Phenol removal gradually increased during the
following 30 days reaching up to 92% elimination, and from day 69 its concentration in
the treated UASB effluent was below the detection limit, which was found to be
0.25 mg L-1. More importantly, increasing the concentration of phenol from
41.2 + 6.47 to 90.6 + 8.42 mg L-1 on day 124, did not affect the removal efficiency of the
rector and microorganisms were capable of acclimatizing to the higher concentration.
The fate of p-cresol is shown in Figure 3.6b. Adaptation of bacteria to this contaminant
resulted slightly more difficult than to phenol. During the first 17 days after addition of pcresol, the effluent concentration steadily increased until reaching a maximum value, 34.6
+ 0.24 mg L-1, at day 37. After increasing the COD OLR, the overall p-cresol removal
from day 33 to day 40 was greatly impacted, resulting in an average elimination of only
57%. However, during the next following 28 days, the efficiency gradually increased and
during this period the removal efficiency averaged 90%. From day 73 until the end of the
experiment the concentration of p-cresol in the effluent samples was below the detection
126
limit of 0.5 mg L-1. As observed with phenol, increasing the average p-cresol
concentration in the leachate from 39.6 + 8.94 to 85.8 + 6.52 mg L-1 did not impact the
performance of the UASB reactor.
The anaerobic treatment proved to be suitable for treating the organic content of
the synthetic leachate. However, as shown in Figure 3.7, NH4+ remained in the treated
effluent, since methanogenic processes are not appropriate for nitrification. In order to
achieve NH4+ removal, the effluent of the UASB was sent to an aerobic DHS reactor. The
DHS was started on day 41 of the UASB operation. The HRT was set to 24 h and the
effluent of this reactor was recirculated back to the top at a rate 6 times higher than the
influent. Samples were taken from the influent and effluent to monitor the fate of NH4+,
NO3-, and NO2-, as well as the COD. A 11-day period was required for nitrifying
microorganisms to enrich to a point where they could remove most of the NH4+ in the
pre-treated leachate, as shown in Figure 3.8a. During the following 40 days the effluent
concentration remained quite low. The influent and effluent average concentration was
equal to 165.8 + 50.4 and 13.7 + 11.7 mg NH4+-N L-1, respectively, achieving an average
removal of approximately 91%. However, the efficiency of the reactor decreased
considerably from day 105 to day 122. There were no apparent reasons for this shift as
the operational conditions, pH, and OLR did not change during this period. High NH4+
elimination was observed again from day 130 until the end of the experiment, reaching an
average of approximately 96% removal, as the average influent concentration,
154.8 + 20.8 mg NH4+-N L-1, was decreased to 5.08 + 1.08 mg NH4+-N L-1. Nitrification
in the reactor was monitored by measuring NO3- and NO2-.
127
Figure 3.6. Influent and (■) and effluent (●) concentration of a) phenol and b) p-cresol in
the UASB reactor as function of time.
128
Figure 3.7. Fate of NH4+ during anaerobic treatment in the UASB reactor as function of
time, where the influent is indicated by (■) and the effluent by (●).
3.4.2
DHS reactor
Figures 3.8b and 3.8c show that NO3- and NO2- were present only in the DHS
effluent and not in the UASB effluent (fed to DHS). For the first 40 days, low NO3content was observed in the effluent, and the average concentration was
11.3 + 4.88 mg NO3--N L-1. From day 90, NO3- formation started to increase until
reaching a maximum during the last 30 days of operation of the reactor, with an average
129
concentration of 145.0 + 5.70 mg NO3--N L-1. NO2- showed the opposite behavior as
observed in Figure 3.8c. High concentrations of NO2- were detected in the samples during
the first 30 days; then NO2- content gradually decreased until day 134. After this point,
NO2- was intermittently detected in the treated effluent at an average concentration equal
to 1.45 + 0.32 mg NO2--N L-1. These data suggest that NO2- was the primary product of
NH4+oxidation in the reactor; however there was a shift in the products of nitrification,
with NO3- becoming the dominant the species produced over time. Table 3.3 shows the N
balance for the DHS reactor. The table is divided in three periods of time based on the
NH4+ removal efficiency of the reactor. The 11-day lag phase was not taken into account
for this analysis. During the first period, from day 55 to day 98, NO2- was the main
product detected of the nitrification. From day 105 to day 124, the main product shifted
to NO3-. In the last period, from day 130 to day 209, the efficiency of the reactor
remained fairly constant and NO3- was present at higher concentrations than NO2-. The
overall mass balance indicated N loses during treatment as the content was always higher
in the influent than in the treated effluent. This could have resulted from analytical errors
or the presence of other byproducts, not measured in the samples. A small fraction of this
difference could be attributed to the formation of anaerobic niches within the packed
sponges which could have enhanced the complete denitrification of NO3- to nitrogen gas
(N2). The average COD concentration in the influent samples was found to be
0.05 + 0.01 g COD L-1 which could have resulted in the formation of up to 8.75 mg N2-N
L-1, assuming 100% efficiency of the denitrification process.
130
Figure 3.8. Fate of a) NH4+-N, b) NO3--N and c) NO2--N in the DHS reactor where the
influent and effluent concentrations are indicated by (■) and (●), respectively.
131
Table 3.3. Nitrogen balance in the DHS reactor.
Day
55 – 98
105 - 124
130 – 209
N total In (mmol)
172.0
120.2
327.5
N total Out (mmol)
135.7
97.7
292.0
Removal of NH4+ (%)
89.8 + 1.7
36.6 + 1.3
96.5 + 0.1
Conversion to NO3- (%)
7.6 + 2.3
14.1 + 1.1
82.9 + 8.8
Conversion to NO2- (%)
77.4 + 7.9
20.1 + 1.2
4.14 + 1.2
3.5
Discussion
3.5.1
UASB reactor
The methanogenic reactor exhibited a high COD removal efficiency throughout
the experiment despite the decrease in the HRT at day 31. Most of the organic load in the
influent was converted to CH4 gas, as indicated by CH4 measurements and corroborated
by mass balances (Table 3.2). Formation of CH4 would have resulted from the conversion
of VFA, phenol, and p-cresol to CH4 by the combined activity of acetogenic and
methanogenic microorganisms [43]. Conversion of phenol, and p-cresol, also present in
132
the synthetic leachate, would have depended on phenolic compound degrading bacteria
converting the phenols to VFA and H2 [44]. A small fraction of the CH4 formed in the
reactor could have resulted from the endogenous decay of sludge biomass. The COD data
show that the organic contaminants in the leachate were efficiently removed. The
capability of the UASB for removing the pollutants added to the synthetic leachate could
be explained by the wide diversity of microorganisms present in anaerobic methanogenic
sludges [45-47]. The increase of the OLR, from 1.04 +0.46 to 2.3 + 0.56 g COD L-1 d-1,
did not seem to have an effect in the performance of the reactor since the average COD
removal during the experiment was up to 95%, which demonstrates the capability of the
UASB for treating the synthetic media. High COD elimination in laboratory-scale UASB
reactors treating raw landfill leachate has been observed at similar ORLs,
2.1 and 2.3 g COD L-1 d-1 [48, 49] and at ORLs as high as 2.3 g COD L-1 d-1 [50].
The fate of the VFAs was monitored to obtain information about the behavior of
each individual fatty acid during treatment. Figure 3.5 shows that, after an adaptation
period, the concentration in the effluent decreased to very low values. In fact, some data
points were below the detection limit of the analysis. The overall average VFA removal
achieved in the UASB was 95% and the individual analysis of each of the VFA added to
the synthetic leachate (not shown) demonstrated that most of the VFA content in the
leachate after treatment corresponded to acetic acid. The presence of acetate could be the
result of an inefficient degradation of the organic acid or a partial fermentation of the
organic components of the leachate since acetate has been identified as a fermentation
byproduct of low molecular weight organic molecules, including higher fatty acids,
133
fumarate, succinate, glycerol, lactate among others [51]. Propionic acid was
intermittently observed in the samples which could indicate partial removal of the
propionic acid or an incomplete acetogenesis of certain VFA in the synthetic media.
However, its average concentration in the treated media was lower than 5 mg L-1. Finally,
valeric and butyric acid were below the detection limit in the effluent of the reactor.
The data obtained show that microorganisms in the reactor were capable of
rapidly adapting to the presence of the phenolic compounds. Moreover, increasing the
average phenol concentration in the influent from 41.2 +6.47 to 90.6 + 8.43 mg L-1 and
the p-cresol content from 39.6 + 8.94 to 85.8 + 6.52 mg L-1 did not impact the removal of
such contaminants. In general, the degradability of phenol and p-cresol could be linked to
the presence of the hydroxyl group in the molecule, as observed by Battersby [52]..
Degradation of phenolics contaminants by digested sludge has been previously
documented in batch experiments [32, 53-55]. Removal under continuous feeding
conditions has been observed as well. High elimination of phenol and p-cresol in UASB
reactors has been documented [56-58]. Data shows that after introducing the phenol and
p-cresol into the treatment, bacteria enriched faster to the presence of phenol, as higher
removal efficiencies were achieved for this compound after the first days. Interestingly,
neither of these compounds were detected above the detection limit in the treated effluent
after day 73, indicating a possible use of p-cresol as co-substrate after the feeding of these
contaminants.
134
Lastly, the NH4+ concentration was monitored in influent and effluent of the
UASB (Figure 3.7). The content of NH4+ in the samples remained fairly constant,
indicating there is no mechanism of anaerobic removal of NH4+ in the absence of O2 or
NOx-. The minor fluctuations observed in the NH4+ content of the treated media could
have resulted from the endogenous decay of biomass, which results in releasing of NH4+,
or from the uptake of NH4+ due to cell yield, which would have caused a small increase
in the effluent. The effluents of UASB reactors have been recommended to undergo posttreatment to achieve good nutrient removal [59].
3.5.2
DHS reactor
The effluent of the UASB was low in COD but rich in NH4+, thus it was treated in
the DHS reactor with the main objective of oxidizing the high concentration of NH4+
present in the anaerobically pre-treated synthetic leachate. The data show that nitrifying
microorganisms enriched rapidly since NH4+removal occurred quickly without much of a
lag period before achieving excellent removal efficiency. The overall average elimination
of NH4+ was up to 96% and the main by-products of the oxidation were NO2- and NO3-.
These results coincide with those obtained by Tandukara et al [60], which demonstrated
high NH4+ oxidation in a similar reactor. A N mass balance performed with the data
obtained from the influent and effluent samples, demonstrated that after adaptation most
of the NH4+ was partially oxidized to NO2- for a period of 43 days, indicating that NH4+
oxidizing microorganisms dominated during the first days of operation of the reactor. The
135
conversion of NH4+ decreased considerably in the following 26 days, but after day 130
high removal was obtained again. At this point NO3- was found at high concentrations in
the effluent, whereas NO2- was hardly detected. The decrease in the performance of the
DHS reactor at the middle of the experiment could have resulted from a switching in the
bacterial communities that dominated in the reactor. Nitrosomonas and Nitrobacter
species had been identified in sponges used in DHS reactors used for NH4+ removal [61].
Uemura et al [62] observed a succession of microorganisms related to Nitrosomonas and
Nitrobacter in a DHS treating wastewater rich in nitrogen.
The activity of the nitrite-oxidizing bacteria increased with time, indicating the
dependence of these organisms to ammonia-oxidizing bacteria. Evidence exists that
nitrite-oxidizing organisms grow around ammonium-oxidizers in biofilms formed from
wastewater [63]. To achieve complete nitrification, the NO2- formed from the oxidation
of NH4+ must be further oxidized to NO3-. However, this is not always achieved and NO2can accumulate. Partial nitrification to NO2- can result in the removal of N at reduced
operational cost, since less O2 would be required in the process. The presence of NH4+
and NO2- at low O2 concentrations could enhance the anaerobic ammonium oxidation
(ANAMMOX), which involves the utilization of NO2- as electron acceptor by
microorganisms to promote the conversion of NH4+ to N2 gas [64]. ANAMMOX results
in N elimination without a denitrification post-treatment. DHS reactors could attain N
removal by controlling the air flow to promote the formation of NO2-, instead of NO3-.
Chuang et al [65] achieved partial nitrification by controlling the O2 concentration a DHS
reactor; however, a slight change in the supply of O2 promoted complete nitrification.
136
Due to the extremely low organic carbon present in the influent media, carbon needed by
the microorganisms must have been fixed from the bicarbonate in the leachate.
3.6
Conclusions
The anaerobic-aerobic system used in this study was suitable for treating a
syntheting landfill leachate containing a high COD concentration, in the form of VFAs,
phenol, and p-cresol, and high NH4+concentration. The organic loading feeding the
UASB was efficiently eliminated from the beginning of the experiment, and was
recovered in the form of CH4 gas. The low concentrations of VFAs detected in the
effluent corresponded mainly to acetate, probably resulting as a fermentation byproduct.
However, the values were so low that removal exceeded 95%. Phenol and p-cresol
proved to be degradable by methanogenic sludge and no toxic effects were observed.
Microorganisms were capable to rapidly enrich to oxidize these pollutants enabling high
efficiency from the start. Additionally, NH4+ was effectively removed in the DHS reactor.
Its operation was divided in two periods, one dominated by ammonia oxidizers and
another dominated by both ammonia and nitrite oxidizer microorganisms; however the
efficiency during the whole operation time was high.
137
3.7
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146
CHAPTER 4
FATE OF CERIUM OXIDE (CeO2) NANOPARTICLES DURING WASTEWATER
TREATMENT
4.1
Abstract
Nanoparticles (NPs) are materials with at least one dimension of 100 nm, as
established by the National Nanotechnology Initiative (NNI). Nano-sized inorganic
oxides such as silica (SiO2), ceria (CeO2), and alumina (Al2O3) are used in many
industrial processes, including catalysis, polymers, coatings and semiconductor
manufacturing. These three inorganic oxides are on the Organization for Economic Cooperation and Development (OECD) list of priority nanomaterials for immediate testing.
Little is known about the behavior of these contaminants in the environment and
information on the stability of NPs in wastewaters is very scarce. The objective of this
research was to investigate the fate of CeO2 NPs during municipal wastewater treatment
and elucidate the main mechanisms contributing to their removal. The study was
conducted in a laboratory-scale activated sludge system (A/S) fed with municipal
wastewater collected from a local treatment facility and a nano-sized CeO2 dispersion
(average primary particle size = 50 nm). Continuous bioreactor experiments were also
conducted with a well-defined synthetic wastewater to study the contribution of specific
wastewater components to NP (de)stabilization. The organic matter removal efficiency
147
was determined by monitoring the total as well as the soluble chemical oxygen demand
(COD) in the influent and effluent. The effect of pH and organic matter on the
aggregation, and possible sedimentation, of the particles was evaluated in batch
experiments by dispersing the NPs in different matrices. Finally, the contribution of the
biomass for the removal of CeO2 during treatment was investigated in batch treatments in
the presence and absence of biomass.
4.2
Introduction
The NNI defines NPs as materials with at least one dimension of 100 nm or less.
In this size range, particles have physicochemical characteristics that differ from their
bigger counterparts. Probably the most important feature of NPs is their large specific
surface area and increased quantum effects [1]. As a consequence of these unique
properties, NPs are being incorporated into numerous products and industrial processes.
Cosmetics, health care products, food packing materials, clothing, tires, and many other
consumer goods contain NPs. In 2006, 212 nano-enable consumer products were
identified in a survey; three years later the number exceeded 1,000 and the trend
continues to increase [2]. Despite the ever increasing commercial and industrial
applications of NPs, little is known about their possible toxic effects and fate in the
environment.
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A significant fraction of engineered NPs can be expected to reach municipal
wastewater treatment plants (MWWTPs). This was recently demonstrated based on a
model that considered the fate of these materials once released from consumer products
or industrial processes [3]. Evidence of the occurrence of NPs in treatment plants
includes the finding of silver sulfide NPs in the final stage sludge from a full scale
MWWTP [4]. The question to be asked is whether or not MWWTPs are suitable for
removing these materials during treatment. Current MWWTPs were not designed to treat
these emerging contaminants [5] and the detection of non-manufactured NPs, i.e.
biogenic colloids, in treated wastewater could suggest the incapability of MWWTPs for
removing such materials [6]. Additionally, titanium dioxide (TiO2) NPs were detected in
the treated effluent from several MWWTPs located in Arizona, California, Colorado,
Iowa, Maryland, and New York [7]. The characteristics of the wastewater will have a
great effect on the behavior of the NPs in the treatment plant. Many NPs do not dissolve
in water; rather they form thermodynamically unstable colloidal dispersions whose
stability depends on the pH and ionic strength (IS) of the wastewater.
Sorption onto primary and secondary sludge and entrapment into microbial
floccules have been suggested as additional mechanisms of NP removal in MWWTP [8].
Extracellular polymeric substances (EPS) are the most important constituents of
microbial floccules and they can behave as ligands, enhancing the attraction of charged
particles [9]. However, [10] in one study reduced aggregation of TiO2, ZnO, and CeO2
NPs in the presence of naturally occurring organic matter (OM) was observed, and this
finding was independent of the pH [11]. Certain NPs can bind to OM in the wastewater
149
[12, 13] due to their overall negative charge from the presence of carboxylic and phenolic
functional groups.
Cerium (Ce) is the most abundant of the rare earths elements at a concentration of
60 parts per million in the Earth’s crust [14]. It is a very strong oxidizing agent that
becomes stabilized when bonded to oxygen [15]. The adverse effects of CeO2 are still not
well known. CeO2 NPs showed limited toxicity to human mesothelioma and rodent
fibroblast cell lines when incubated for 6 days at a concentration of 30 mg CeO2 L-1 [16].
They caused 50% inhibition after a three day incubation period with such cells; yet
activity remained fairly constant during the next following 3 days. Evidence of
significant cytotoxicity and oxidative stress to human lung cancer cells by 20 nm CeO2
NPs was documented by Lin et al [17]. Oxidative species were generated at all
concentrations tested, ranging from 3.5 to 23.3 µg CeO2 mL-1, causing lipid peroxidation
and cell membrane damage. However, Schubert et al [18] found that CeO2 NPs could act
as antioxidants, protecting nerve cells from oxidative stress. Antioxidant effects could
result from special crystal defect structure. Hence, more research is needed to understand
the toxicity of CeO2 NPs to living organisms.
Exposure to Ce may occur mainly from cerium oxide (CeO2) containing products,
which include glasses, ceramics, televisions tubes, semiconductors, gasoline, etc.
Incorporation of CeO2 NPs in diesel engines can reduce particulate matter emissions [19].
CeO2 NPs are used as an oxygen donating catalyst for the oxidation of carbon monoxide
(CO) and hydrocarbons [20]. In the semiconductor industry CeO2 NPs, together with
aluminum oxide (Al2O3) and silicon dioxide (SiO2) NPs, are employed for the
150
chemical-mechanical polishing (CMP) of wafers. Smooth wafer surfaces free of defect
are achieved by CMP processes [21]. In 2003 the slurry market for CMP was estimated to
be $400 million [22], and the annual average growth rate of the CMP global market
during 2008 was 14.2% [23]. CeO2 NPs are expected to be present in municipal and
industrial wastewater since most of the engineered NPs are released to sewer systems
[24].
Only a few studies exist related to the fate of NPs during wastewater treatment.
Removal of bare and coated SiO2 NPs in a laboratory-scale primary treatment was
studied using unscreened and screened real domestic wastewater [25]. Coagulation and
further sedimentation was only observed in Tween-coated particles. Limbach et al
reported up to 94% removal of CeO2 NPs during simulated secondary treatment when
dispersing the material in synthetic wastewater [26]. Moreover NPs proved to be highly
stabilized in the synthetic media. Elimination of TiO2 NPs in sequencing batch reactor
simulating secondary treatment was studied by Kiser et al [7]. An average elimination of
88% was observed when treating synthetic wastewater containing the NPs. Although
these results provide a general understanding about the behavior of NPs during treatment,
their short duration and the utilization of synthetic wastewater, instead of real domestic
wastewater, might not represent accurately the conditions NPs would encounter under
real conditions.
151
The scope of this work was to assess the behavior of CeO2 NPs during secondary
treatment in a bench-scale aerated activated sludge system over long term experiments
with real domestic wastewater. CeO2 was selected for this project due its low background
concentration in wastewater, which facilitates its monitoring without interference of
natural or anthropogenic occurring CeO2. Stability of CeO2 NPs in water and wastewater,
as well as their sorption to biomass, was studied to insight on the dominant mechanisms
of CeO2 removal.
4.3
Materials and Methods
The NPs were obtained as a fine powder with an average particle size (APS) of
50 nm in diameter according to the manufacturer (Sigma-Aldrich, St Louis, MO).
However a transmission electron microscope image of the powder (Figure 4.1) shows
particles as small as 35 nm in diameter.
4.3.1 CeO2 stability in aqueous dispersions
The impact of pH and wastewater composition on the stability of CeO2 NP
dispersions was tested based on measuring particle aggregation and sedimentation. A
two-step process was used to prepare the CeO2 stock suspension. A concentrated stock
suspension containing 4.07 g Ce L-1 (5.00 g CeO2 L-1) was prepared by dispersing the
152
NPs in 1 mM HCl and applying sonication for 10 min at 70% intensity to achieve a
uniform dispersion (Daigger GEX130, 130W, Cole-Parmer Instruments, Vernon Hills,
IL). Then, the stock was diluted in 1 mM HCl to 0.81 g Ce L-1 (1.00 g CeO2 L-1) and a
pH of 3.10 + 0.20. The diluted stock suspensions were diluted 11-fold to an intended
concentration of 74.0 mg Ce L-1 (90.9 mg CeO2 L-1) with real domestic wastewater or
synthetic wastewater adjusting the pH to 7.10, which was the original pH of the
wastewater, as well as 11-fold dilution with Milli-Q water at pH 3.11 or 7.06 in 15 mL
falcon tubes (BD Biosciences, Bedford, MA) in duplicate. Both real and synthetic
wastewater samples were passed through 25 nm filters (Millipore.Billerica, MA) to avoid
interference of suspended materials. A 1 mL sample was taken from the initial 11 mL
content from the treatment containing Milli-Q water at pH 3.11 after vortex mixing the
aqueous dispersions to determine average particle size (APS) and zeta-potential at the
beginning of the experiment. Then the tubes were incubated for 24 h without any
mechanical mixing. After incubation, 1 mL sample was taken from the upper 15% of
supernatant of each assay to measure CeO2 recovery. The Ce content in the supernatant
was analyzed after digesting the samples.
153
Figure 4.1. Transmission electron microscope image of nano-sized ceria with an average
particle size 50 nm.
4.3.2
Laboratory-scale secondary treatment
The bench-scale aerobic activated sludge treatment system, constructed with clear
acrylic plastic, consisted of an aeration tank connected to a settler with a working volume
of 1.19 and 0.66 L, respectively (Figure 4.2). The laboratory-scale secondary treatment
system operated at 27.5 ºC and was supplied with wastewater at a steady volumetric flow
rate set to 2.77 + 0.09 L d-1 corresponding to a hydraulic retention time (HRT) of
10.3 + 0.32 h in the aeration tank. The average flow leaving the settler was
2.66 + 0.21 L d-1 equal to an HRT of 6.50 + 0.92 h (Table 4.1). The flow rate was
154
measured by collecting a defined volume of influent and effluent during a determined
time. The aeration tank was seeded with returned activated sludge (RAS) at a volatile
suspended solids (VSS) concentration of 3.50 g VSS L-1. Two air pumps
(Aqua, Walmart) provided air at a rate of 430 L d-1 and mixing for the aeration tank and
recirculation of the settled sludge. The laboratory-scale secondary treatment system was
operated in two different experiments. In the first experiment the reactor system was fed
with synthetic wastewater for 28 days (experiment I). CeO2 NPs were introduced into the
influent after 13 days of operation and they were supplied until the end of the experiment,
13 days more. The synthetic wastewater consisted of (g L-1): peptone (0.220), meat
extract (0.150), urea (0.010), K2HPO4 (0.008), NaHCO3 (0.400). The pH was adjusted to
7.00 + 0.10 before adding the sodium bicarbonate. The average pH of the wastewater
during the operation of the reactor was 7.68 + 0.15 and the total chemical oxygen demand
(COD) content was equal to 371.7 + 33.03 mg COD L-1. In the second experiment the
reactor was fed with real domestic wastewater for 70 days (experiment II). In this case
the feeding of the NPs lasted for 57 days, as they were added at day 7. Real domestic
wastewater was collected after primary sedimentation from Roger Road, a local
MWWTP located in Tucson, AZ. The average wastewater pH was found to be
7.71 + 0.14. The wastewater contained an average COD concentration of
247.4 + 50.39 mg COD L-1. The CeO2 NPs were introduced into the system via diluted
stock dispersion at a concentration equal to 0.81 g Ce L-1 (1 g CeO2 L-1). The stock was
continuously stirred to avoid agglomeration of the NPs and combined with wastewater at
a volumetric ratio of 1:10 at the entrance of the aeration tank.
155
2
4
1
3
5
Figure 4.2. Schematic of the laboratory-scale secondary treatment system used.
(1) Aeration tank, (2) air, (3) settler, (4) effluent, (5) recycled sludge.
Table 4.1. Operation parameters of the laboratory-scale activated sludge system fed with
synthetic and real wastewater.
Parameter
Synthetic wastewater
Real wastewater
Operation time (days)
27
65
Θ in the aeration tank (h)
10.58 (+ 0.54)
10.11 (+ 0.96)
Average pH In
7.68 (+ 0.14)
7.41 (+ 0.14)
Average pH Out
7.05 (+ 0.63)
7.05 (+ 0.42)
Mixed liquor volatile
suspended solids
concentration (g/L)
3.13 (+ 0.74)
3.23 (+ 0.24)
156
Samples were taken from the influent and effluent of the settler to follow the
removal of the organic content of the wastewater by measuring the total and soluble COD
and acetate concentration. The pH of the samples was monitored as well as it is an
important variable that can affect the microbial activity in the secondary treatment. To
maintain a fairly constant biomass concentration in the reactor 90 mL of sludge were
taken from the aeration tank every 3 days, which were replaced with wastewater.
Biomass concentration was determined by measuring the VSS content of the samples.
The fate of CeO2 NPs during treatment was followed by measuring the total and filtered
Ce concentration in the samples, including the diluted stock dispersion. Scanning electron
microscopy (SEM) and energy dispersive spectroscopy (EDS) analysis were performed to
confirm the presence and absence of Ce in selected samples.
4.3.3
Batch adsorption experiments
The contribution of the sludge biomass to the removal of NPs was studied in a
sorption experiment. For this purpose, treatments were set up at pH 3.0 and 6.0 in
duplicated 50-mL falcon tubes. The treatments consisted of MilliQ-water alone,
MilliQ-water with washed sludge at a VSS concentration of 3.50 g VSS L-1, as well as
the supernatant from the MilliQ-water with washed sludge. The sludge was washed four
times by mixing it with Milli-Q water and centrifuging the mixture afterwards at
4000 rpm for 30 min. The intended Ce concentration added to the treatments was set to
81.40 mg Ce L-1 (100.0 mg CeO2 L-1). The treatment containing of Milli-Q water was
157
sampled right after diluting the stock to establish the initial dispersed Ce concentration.
The tubes were continuously mixed for 15.23 h and then incubated for 4.83 h under
stationary conditions to emulate the secondary treatment (10.3 and 6.50 h HRT in
aeration basin and settler, respectively). Finally the supernatant from each treatment was
sampled and digested to measure the total Ce concentration remaining in suspension.
CeO2 sorbed onto sludge was estimated by comparing the Ce content in the supernatant
of each treatment.
4.3.4 Analytical procedure
Total Ce concentration was measured in the samples taken from the stability test,
the laboratory-scale secondary treatment, and sludge biomass sorption experiment. For
this purpose samples were digested in an automated microwave system (MDS2100, CEM
Corp., Matthews, NC) in a solution containing 8 mL of HNO3 (69%) and 2 mL of H2O2
(30%) for 30 min at 70 PSI. After digestion, samples were diluted in Milli-Q water and
analyzed in an inductively coupled plasma – optical emission spectrometer (ICP-OES,
Optima 2100 DV, Perkin Elmer, Waltham, MA) at 413 nm wavelength. Filtered samples
were passed through 200 nm (Whatman. Piscataway, NJ) and 25 nm (Millipore.Billerica,
MA) filters and then amended with a drop of HNO3 concentrated to preserve the samples.
The resulting filtrate was diluted with 4% HNO3 (v/v) before analysis in an ICP-OES.
Particle size (PS) and zeta-potential were measured by light scattering using a
Zeta Sizer Nano ZS (Malvern, Inc., Westborough, MA). For this purpose a capillary
158
vessel was filled with 1 mL sample, carefully removing any air bubble trapped in the
capillary. The temperature of the analysis was 25oC and Milli-Q water was used as
dispersant. The refractive index of CeO2 was set equal to 1.828.
The pH of the samples and stock dispersions were measured using a pH probe
(Orion-Thermo Scientific, Cincinnati, OH) calibrated before measurements with certified
standards (Orion-Thermo Scientific, Cincinnati, OH). Ce and acetate determinations were
quantified by analyzing standards of known concentrations together with the samples.
4.3.5
Electron microscopy
Preparation and analysis of the samples taken from the aeration tank after being
exposed to CeO2 was performed at the University spectroscopy and imaging facilities
(USIF) at the University of Arizona. The samples were provided to USIF in aqueous
solution and without addition of any preservative. Formaldehyde was added as a fixative
at 50% (v/v) ratio 4 h before the preparation of the samples. Later the specimens were
vacuum dried and hand broken into small pieces forming a powder-like material, which
was placed and glued in a metallic base. Finally, the samples were coated with gold (Au)
and palladium (Pd). The coating serves as a protective barrier from the electron beam at
which the samples are subjected when using these techniques. Electron dispersive
spectroscopy (EDS) was used to prove the presence of Ce in the samples.
159
4.3.6
Organic content of the wastewater and biomass concentration
The total COD concentration in the liquid samples was measured by mixing
2.5 mL of sample with 1.5 mL of digestion solution and 3.5 mL of a sulfuric acid
(H2SO4) solution. The digestion solution contained (g L-1): H2Cr2O7 (10.27),
Hg2SO4 (33.30), and H2SO4 (167 mL L-1). The H2SO4 solution was prepared by
dissolving10.25 g of Ag2SO4 in 1 L of concentrated acid. The resulting mixture was
digested for 2 h at 150 ºC. Lastly, the absorbance of the digested samples was analyzed in
a spectrophotometer (Beckman-Coulter, Fullerton, CA) at a 600 nm wavelength. The
soluble COD fraction was determined following the same procedure after centrifuging
the samples for 10 min at 10,000 rpm.
The fate of acetate in the system was monitored as well. After centrifugation of
the influent and effluent samples, 1 mL of the supernatant was transferred into 1.5 mL
glass vials and amended with 0.5 µL of concentrated formic acid. The acetate content of
the samples was measured in a 7890A Agilent Technologies gas chromatograph (GC)
(Agilent Technologies, Palo Alto, CA) equipped with a flame ionization detector (FID)
and a 30 m x 0.53 mm Stabilwax-DA column (Restek, Bellefonte, PA). The temperature
at the inlet was set to 250ºC and 275ºC for the detector. The initial temperature in the
oven was gradually increased from 90ºC to 140ºC at a rate of 10ºC per minute. Samples
were injected using an autosampler applying a 6:1 split ratio. Ultra high purity helium
was used as carrier gas at a flow rate of 16 mL min-1.
160
Biomass content in the reactor was determined by measuring the VSS
concentration of the samples. For this purpose 20 mL of the sample were filtered through
450 nm glass fiber filters (Whatman. Piscataway, NJ) applying vacuum. Filters
containing the retentate were placed in aluminum vessels (Fisher Scientific, Pittsburgh,
PA) and dried overnight at 105oC. Finally, the dried material was ashed in a muffle
furnace at 550oC for 4 h.
4.4
Results
4.4.1
Stability of CeO2 NPs
The stability of CeO2 NPs was studied in aqueous dispersions at different pH
values and organic composition. Figure 4.3a shows the APS of the NP dispersions in the
assays before and after 24 h incubation without mixing. Initial samples were taken from
the bulk after diluting the NP stock in the aqueous treatments. After incubation, samples
were taken from the overlying supernatant. At pH 3.11, NPs were stable as demonstrated
by the small average particle size of the suspension, which remained constant over time.
The average hydrodynamic diameter of the particles measured at the beginning and end
of the test was 147.7 + 2.17 nm. Destabilization was observed at circumneutral pH.
Aggregation of NPs in Milli-Q water at pH 7.06 was evident.
161
The APS measured after dilution of the stock, 782.7 + 263.2 nm, was up to five
times greater than that obtained at pH 3.11. The diameter of the particles that remained
suspended in the supernatant after incubation was smaller than those present in the bulk
at the beginning of the experiment. Surprisingly, the effect of pH on CeO2 NPs was
overcome when diluted in synthetic wastewater. In the presence of the synthetic
wastewater, the particles behaved similarly as those in Milli-Q water at low pH. The
suspension was very stable since the small average hydrodynamic particle size of
157.8 + 2.62 nm was sustained over time. In contrast, the presence of real wastewater
caused the NPs to readily aggregate and the measured APS increased considerably to the
micron range. The average diameter of the particles exposed to real wastewater was
2,457 + 211.7 nm.
The zeta-potential exhibited by the particles in the different treatments is
presented in Figure 4.3b. The data obtained show that at higher absolute values (either
positive or negative), dispersions become more stable. The average zeta-potential over
the 24 h period in the treatment consisting of Milli-Q water at pH 3.11 was equal to
45.8 + 2.88 mV, whereas that of the synthetic wastewater was -32.6 + 0.30 mV.
Agglomeration was enhanced as the absolute value of the zeta-potential of the
particles decreased, as demonstrated by the increased aggregation observed in the real
domestic wastewater and Milli-Q at pH 7.06 treatments, which showed an average zetapotential closer to zero. Results indicated that the charge of the particles play a major role
over stability. Nevertheless, other factors may affect the agglomeration of the particles as
162
the APS was up to four times higher in real domestic wastewater compared to the Milli-Q
water treatment at pH 7.06.
The residual CeO2 NPs was monitored by measuring the Ce in the supernatant of
each assay after incubation (Figure4.4). The designed total Ce concentration in the
treatments was 74.0 mg Ce L-1 (90.9 mg CeO2 L-1). The treatment consisting of Milli-Q
water at pH 3.11 was sampled from the bulk at the beginning of the experiment and
digested in order to determine the maximum Ce content that could be expected in the
supernatant of each treatment, which was found to be 71.8 + 0.53 mg Ce L-1
(88.2 CeO2 L-1). This measurement was set as the control. The same treatment was
sampled 24 h later and found that NPs remained in suspension over time as indicated by
the Ce content in the supernatant, 74.0 + 8.82 mg Ce L-1 (90.9 + 10.8 mg CeO2 L-1),
which matched the one obtained for the controls (Figure 4.4). Neutral pH values
promoted the sedimentation of the particles suspended in clean water. The Ce
concentration in the supernatant of Milli-Q water at pH 7.06 after incubation decreased
significantly when compared to the Milli-Q treatment at pH 3.11. The average Ce
concentration in the supernatant was 25.0 + 6.14 mg Ce L-1 (30.7 + 7.54 mg CeO2 L-1),
indicating that the majority of the particles, approximately 65.5%, settled during the test.
However the effect of the pH was overcome to some extent when CeO2 NPs were diluted
in synthetic wastewater at pH 7.10 as was observed by the high content of Ce in the
samples. The residual Ce concentration in that treatment was 90.2% based on measuring
65.8 + 2.16 mg Ce L-1 (80.8 + 2.65 mg CeO2 L-1) in the supernatant.
163
Figure 4.3. a) Particle size distribution and b) zeta-potential of CeO2 in different matrices
at t=0 (■) and t=24 h (■) incubation in MQ = Milli-Q water (pH 3.11 and 7.06);
SWW = Synthetic wastewater (pH 7.10); WW = Real wastewater (pH 7.09).
164
In the case of real wastewater most of the aggregates formed readily sedimented
rather than stay in suspension. Only a small fraction from the total Ce initial
concentration, 5.70%, did not sediment, as the concentration of Ce measured in the
samples was equal to an average 4.15 + 3.38 mg Ce L-1 (5.10 + 4.15 mg CeO2 L-1). In
general, the residual Ce concentration in the supernatant followed the behavior observed
in terms of particle size for each treatment. Suspensions with smaller aggregate sizes
resulted in higher stability of the Ce concentration during incubation.
Figure 4.4 Ce concentration in the supernatant from different matrices after t =24 h
incubation. Control = Milli-Q water at t=0 (pH 3.11); MQ = Milli-Q water (pH 3.11 and
7.06); SWW = Synthetic wastewater (pH 7.10); WW = Real wastewater (pH 7.09).
165
4.4.2
Removal of CeO2 NPs from synthetic wastewater during activated sludge
treatment
The fate of CeO2 NPs during secondary treatment was studied using a
laboratory-scale activated sludge system. The reactor was fed with synthetic wastewater
for 13 days without adding CeO2 to allow stabilization of the system. Afterwards NPs
were continuously amended to the influent wastewater to achieve a designed CeO2 NP
concentration of 90.9 mg L-1 (74.0 mg Ce L-1). NPs were fed for the following 13 days,
from day 13 to day 26. Figure 4.5a shows the fate of total Ce in the reactor. The total
average Ce concentration measured in the influent was equal to 69.9 + 20.3 mg Ce L-1
whereas the average concentration detected in the effluent was 3.94 + 1.73 mg Ce L-1.
The average removal of the total Ce during secondary treatment was 94.4%, indicating
that the activated sludge treatment was effective in removing the majority of the CeO2
supplied to the system. The efficiency of the treatment for removing small particles was
studied by filtering the samples using 200 nm filters. Figure 4.5b illustrates that only a
small fraction of the suspended CeO2 particles entering the reactor were smaller than
200 nm. Their average concentration in the influent was 0.64 + 0.40 mg Ce L-1 (0.78 +
0.49 mg CeO2 L-1), representing less than 1.0% of the CeO2 entering the reactor. After
treatment, the content of these particles in the effluent was found to be 0.13 + 0.08 mg Ce
L-1, which represents an overall removal of 78.9% of Ce in particles
< 200 nm.
Although the < 200 nm size fraction of CeO2 particles were significantly removed in the
bench-scale secondary treatment system, an appreciable fraction of around 21.1% did
166
escape the treatment. The synthetic wastewater may represent the COD concentration
encountered in real domestic wastewater; however it does not properly represent the
complex OM and other suspended solids that could potentially affect the behavior of NPs
in the activated sludge treatment, as was observed during the stability test.
4.4.3 Fate of CeO2 NPs during activated sludge treatment fed with municipal
wastewater
In order to assess the fate of the CeO2 NPs in the secondary treatment with
complex OM, the reactor was stopped, seeded with fresh sludge, and fed with real
domestic wastewater for 65 days. The designed concentration of CeO2 NPs in the influent
wastewater was set to 90.9 mg CeO2 L-1 (74.0 mg Ce L-1) as when done feeding synthetic
wastewater. Figure 4.6a shows that the influent contained an average total concentration
of 54.9 + 12.8 mg Ce L-1 (67.5 + 15.7 mg CeO2 L-1). The discrepancy between the
designed and measured Ce content could be attributed to sedimentation of CeO2 in the
influent line. Most of the particles that reached the secondary treatment were removed
from the wastewater during treatment. The average concentration of total Ce leaving the
settler was 1.83 + 1.42 mg Ce L-1, achieving an average removal of approximately
96.6%, providing evidence that the activated sludge system was suitable for treating real
wastewater containing CeO2. The fate of the particles in size fractions smaller than
200 nm and 25 nm was monitored by serial filtrations, to study the efficiency of the
treatment for eliminating the small aggregates and primary particles. The fraction of the
167
total CeO2 particles passing through the 200 nm membrane filter in the influent was
14.4%. Figure 4.6b shows their behavior in the treatment. The average removal achieved
was equal to 98.6% based on the influent and effluent concentrations of < 200 nm CeO2
size fraction which corresponded to 7.94 + 5.94 and 0.12 + 0.11 mg Ce L-1, respectively.
Most of these particles were actually smaller than 25 nm as they accounted for up to
94.2% of the < 200 nm CeO2. Hence, the content of < 25 nm CeO2 NPs in the samples
was very similar to that observed for the < 200 nm particles (Figure 4.6b). The average
removal achieved for these NPs was 92.3%, which shows that the secondary treatment
system was highly efficient for eliminating such small materials. A sludge sample taken
from the aeration tank after being exposed to CeO2 for 56 days was imaged using SEM
(Figure 4.7a). The image shows the presence of bacteria, protozoa, and extracellular
products. The presence of Ce in the sludge sample was confirmed by EDS analysis
(Figure 4.7b) of the white square marked in Figure 4.7a. The spectrum also shows high
counts of oxygen and carbon which could be attributable to the presence of microbial
biomass in the sludge.
168
Figure 4.5. Fate a) total and b) < 200 nm Ce in the influent (●) and (■) effluent samples
when feeding synthetic wastewater to the secondary treatment.
169
Figure 4.6. Concentration of a) total unfiltered and b) filtered Ce in the (●) influent and
(■) effluent of activated sludge treatment operated with real domestic wastewater. Bold
markers in second panel indicate CeO2 < 200 nm, while empty markers refer to CeO2
< 25 nm.
170
Figure 4.7. a) SEM image of a sample from the aeration tank after being fed with CeO2
for 2 months and b) EDS analysis of the sample. The white rectangle in the image
indicates the section of the sample where the EDS analysis was performed. Accelerating
voltage: 12.0 kV.
171
4.4.4
Contribution of biomass to CeO2 removal during activated sludge secondary
treatment
The role of sludge biomass on the removal of CeO2 NPs was studied, at low and
neutral pH, in batch experiments that resembled the mixing and settling conditions of the
laboratory-scale secondary treatment (Figure 4.8). Samples taken from the treatments
with Milli-Q water at pH 3.05 and 6.02 after dispersing the CeO2 stock were used as
controls to determine the maximum Ce content expected in the supernatant of the
treatments at the end of the experiment. Ce concentration from these samples was found
to be 87.6 + 0.11 mg Ce L-1 at pH 3.05, and 81.6 + 2.97 mg Ce L-1 at pH 6.02 which was
close to the designed concentration of 81.4 mg Ce L-1. The results in the figure indicate
that NPs remained stable over time at low pH and nearly half sedimented at
circumneutral pH when dispersed in Milli-Q water, which coincided with the behavior
observed in the stability test. The presence of washed sludge biomass overcame the
stabilizing effects of pH, as contact with CeO2 NPs resulted in large and significant
decrease in Ce concentration in the supernatant. A similar residual Ce content in the
supernatant was measured in this treatment at pH 3.03 (2.50 + 0.66 mg Ce L-1) and
at pH 5.9 (4.93 + 1.11 mg Ce L-1). This may be due to sorption onto biomass; however a
large decrease in Ce content can also be caused by the destabilizing effect of soluble
organics released from the sludge on the nano-dispersion, as observed in the treatments
containing only sludge supernatant. The contribution of the sludge supernatant to the
agglomeration of the particles also increased with pH, as shown in Figure 4.8.
172
Figure 4.8. Sorption of CeO2 to biomass. Total unfiltered Ce was measured from the
supernatant of each assay after 20 h. Two pH values were considered for this test.
Treatments consisted of Milli-Q water at pH 3.0 and 6.0, sludge biomass at pH 3.0 and
5.9, and rinse sludge at pH 2.7 and 6.0. The Milli-Q water treatment was sampled at the
beginning of the experiment (Water 0 h) to be used as a control.
173
4.4.5
COD removal during secondary treatment
The performance of the activated sludge treatment was monitored with COD
measurements of the influent and effluent. In both experiments the secondary treatment
was first operated by feeding wastewater without addition of NPs. This step was done to
obtain baseline information about the average COD elimination in absence of CeO2.
Figure 4.9 shows the fate of total COD concentration when using synthetic wastewater.
The average COD removal before the feeding of CeO2 was 92.2%. The efficiency was
maintained as NPs were introduced into the treatment when the average COD removal
was 88.2%. The performance of the activated sludge treatment, when utilizing real
domestic wastewater, is presented in Figure 4.10. The soluble fraction of the total COD
in the influent wastewater was found to be 64.6%. The total and soluble COD elimination
achieved before addition of CeO2 NPs was 81.1 and 83.5%, respectively. After feeding
the CeO2 NPs there was a slight decrease in both soluble and total COD removal. The
efficiency observed during the feeding of CeO2 NPs was 65.4 and 65.9% for the soluble
fraction and total COD, respectively. The efficiency drop off after adding CeO2 was not
due to a deterioration in effluent quality but rather it was more related to a general
decrease in the average influent COD.
The fate of acetate during treatment was monitored when feeding real domestic
wastewater (Figure 4.11). Although the influent concentrations fluctuated considerably,
acetate was consistently highly removed in the activated sludge system. There was no
noticeable effect of the presence of CeO2 to acetate removal as its concentration in the
174
effluent was below the detection limit during the first 16 days of operation. Although
acetate was detected in samples taken from day 17 until the end of the experiment, its
average concentration during this period was only 2.02 + 1.60 mg L-1 which represents an
average net removal of 93.4%.
Figure 4.9. Total COD removal in the secondary treatment system when using synthetic
wastewater, where (●) indicates the influent and (■) the effluent. Vertical line indicates
the addition of CeO2 to the influent wastewater.
175
Figure 4.10. Fate of a) soluble and b) total COD in the aerobic activated sludge system
when feeding real wastewater, where (●) indicates the influent and (■) the effluent.
Feeding of CeO2 to the reactor is represented by the continuous vertical line. Dashed line
indicates fresh wastewater batches.
176
Figure 4.11. Acetate concentration in the influent, (●), and effluent, (■), of the secondary
treatment when feeding real wastewater. Continuous vertical line indicates the addition of
CeO2 to wastewater. Dashed vertical lines represent fresh wastewater batches.
177
4.5
Discussion
4.5.1
Fate of CeO2 during wastewater treatment
In the present study, the laboratory-scale activated sludge treatment proved to be
efficient for removing CeO2 dispersed in synthetic and real domestic wastewater as
demonstrated by the high elimination achieved. When operating the secondary treatment
system with real domestic wastewater only an average 3.4% of the total Ce in the influent
was detected in the effluent of the clarifier. A similar behavior was observed when using
synthetic wastewater, where the total Ce removal achieved was higher than 94.0%. Data
obtained from Ce content in filtered samples revealed that elimination of CeO2 < 200 nm
varied significantly between experiments, achieving a higher removal in the experiment
using real domestic wastewater. Approximately 98.6% of CeO2 < 200 nm was eliminated
when NPs were dispersed in real domestic wastewater. Moreover, most of these particles
were < 25 nm and their elimination exceeded 92.0%. On the other hand, more than 20%
of the < 200 nm particles were not removed during treatment when the reactor was fed
with synthetic wastewater. The organic and/or inorganic content in the real domestic
wastewater could have enhanced the further elimination of CeO2 < 200 nm.
178
Previous studies exist investigating the fate of CeO2 [26] and TiO2 [7] NPs
suspended in synthetic wastewater during simulated secondary treatment. TiO2 NPs were
more persistent than CeO2 NPs, as the overall average removal achieved was equal to
88.0 and up to 94.0% for Ti and Ce, respectively. Moreover, Ce was highly eliminated
from the influent wastewater even in the presence of biodegradable and
non-biodegradable surfactants. Although these results provide an insight about the
treatability of wastewater containing metal-oxide NPs, they could fail in accurately
representing the behavior of the particles during treatment. Removal of Ce in the present
study when using synthetic wastewater was similar to that observed by Limbach et al
[26]. However, adding the NPs in real domestic wastewater resulted in higher
elimination. The short duration of the previous studies did not allow the determination if
the removal efficiencies were sustained with time, which was confirmed for the case of
CeO2 from the data obtained in this project.
The overall removal efficiency was greater for bulk Ce than < 200 nm or < 25 nm
fraction Ce when using either synthetic or real domestic wastewater. These data suggest
that elimination mainly occurred by sedimentation of agglomerated larger particles,
which could have partitioned onto the sludge present in the reactor. The SEM images
obtained from sludge exposed to Ce for extended time indicated mineral-like structures
accumulated next to bacteria and protozoa. The occurrence of Ce in the sludge was
confirmed by the EDS analysis. Evidence exists suggesting that granulated Ce could
sorbed onto sludge [26]. Additionally, biosorption of ionic Ce onto aerobic granules has
been demonstrated to be technically feasible [27].
179
4.5.2
Mechanisms for CeO2 removal
CeO2 was subjected to pH change as NPs were taken from the diluted stock
dispersion (pH 3.10) to being mixed with wastewater (above pH 7.40). Based on colloidal
chemistry, the expectation was an enhanced agglomeration of the particles as a result of
the pH change, since the isoelectric point (IP) of the Ce is found between 6.75 and 7.90
[28]. Readings of the zeta-potential at the surface of the CeO2 were higher when particles
were dispersed in Milli-Q water at pH 3.11 than at pH 7.06. Particles become neutral as
the pH of the dispersion approaches the IP, promoting agglomeration as the zeta-potential
decreases. On the other hand, at pH values away from the IP the zeta-potential increases
as result of the increased charge, causing particles to repeal each other. Data obtained
from the stability test confirmed the expectations. The APS of CeO2 agglomerates in
Milli-Q water at circumneutral pH, and zeta-potential close to zero, were up to four times
higher than that observed at pH 3.11. Even at acidic pH the hydrodynamic diameter of
the particles was bigger than the primary particles. This observation indicates that even at
low pH some aggregation is taking place. Agglomeration of commercially available NPs
is common and despite the fact that aggregates can be broken by ultrasonication, it is not
always possible to obtain the original small material [29]. Even though there was some
agglomeration at pH 3.11, NPs largely remained in suspension over time as confirmed by
the residual Ce concentration in the supernatant after incubation, which was close to that
measured in the control.
180
The pH of the dispersion does have an important impact over aggregation of the
particles, as previously discussed. However, this effect can be overcome by the presence
of organic and/or inorganic compounds. Synthetic wastewater dispersed CeO2 NPs better
than Mili-Q water at pH 7.06. The presence of OM has shown to induce dispersion
stability of NPs in aquatic environments as NPs are bound to acidic functional groups,
providing steric and electrostatic stability [30, 31]. Evidence exists proving that NPs can
bind to proteins as well. Gold NPs were stabilized in the presence of mixed peptide and
polyethylene glycol monolayers; and that NPs stability increased with increasing polymer
lenght [32]. CeO2 NPs are known to bind to some proteins, even when surfaces have the
same charge [33, 34]. Additionally, it has been demonstrated that certain inorganic
chemicals can stabilize some NP dispersions. For example, magnetite NPs in the presence
of phosphate (PO43-) have been effectively stabilized when dispersed in different media
[35, 36]. Hence, the dispersing impact of the synthetic wastewater was not surprising,
since the main organic components were peptone and meat extract, and PO43- was an
important inorganic constituent of the synthetic media as well. The impact of real
domestic wastewater was drastically different than that of the synthetic wastewater.
Adding CeO2 in real domestic wastewater resulted in large particle size and greater
sedimentation compared to the treatment consisting of Milli-Q water at pH 7.06.
Instability was promoted by the constituents of the real domestic wastewater, as its pH
was equal to 7.09. It has been suggested that NPs are likely to aggregate in MWWTPs
due to the organic content in the sewage [37]. Aggregation of colloidal particles in fresh
181
water has been shown to be enhanced by the presence of naturally occurring biopolymers,
which overcome the stabilizing effect of fulvic acids [38].
During secondary treatment with synthetic wastewater NPs were highly removed.
These results appear to contrast the expectations generated from the stability test, where
CeO2 was stabilized over time when dispersed in the synthetic media. However, the main
difference between the activated sludge treatment and the stability test resides in that
polypeptides and amino acids, from the meat extract and peptone, are degraded during
the biological treatment. Hydrolysis is an important step for the biological treatment of
municipal wastewater, and different enzymes have been identified serving this purpose,
as bacteria can only metabolize simple molecules [39, 40]. In this study, the overall
removal efficiency of CeO2 was greater when the particles were added to real domestic
wastewater compared to synthetic wastewater. This difference might be attributable to the
type of organic matter and/or inorganic constituents that promoted stable dispersions in
synthetic wastewater, albeit that it was subjected to degradation, and coagulation in real
domestic wastewater.
The presence of biomass might enhance further removal of NPs as they become
sorbed onto the sludge. Different NPs have been found to undergo biosorption onto
activated sludge suggesting that this might be the major removal mechanisms of NPs
from wastewater [41]. In the present study, sludge contributed significantly to the
removal of CeO2 in suspension at pH 3.00 and 6.02. However, it was observed that the
soluble constituents of the sludge, dissolved during incubation, contributed by themselves
to the destabilization of the nanodispersions in the absence of sludge. The extent of extra
182
Ce loss due to sedimentation from the supernatant of the treatment accounts in large part
for the amount lost with the sludge present. Therefore, there is no strong evidence for
sorption of CeO2 onto biomass.
4.5.3
Impact of CeO2 on the wastewater treatment process
The addition of CeO2 into the influent wastewater did not cause disruption of the
COD removal activity of the activated sludge in both experiments neither when feeding
synthetic wastewater nor when using real domestic wastewater. For the latter case, small
losses in COD removal were observed in the period after introducing the CeO2 NPs
compared to the values obtained before addition of the particles. This difference was
mostly due to a lower COD content of the influent real domestic wastewater in the period
after addition of CeO2. Moreover, in a preliminary study performed in our research
group, toxicity experiments of CeO2 over acetate consuming activity of activated sludge
have demonstrated that concentrations of up to 1000 mg CeO2 L-1 had no noticeable
impact on acetate consumption. Although the toxicity of many NPs to microorganisms
has been explored [42, 43], information is limited assessing the impact of metal oxide
NPs over biological activity in wastewater treatment processes.
183
4.6
Conclusions
Nanodispersions of CeO2 are highly removed during activated sludge treatment
regardless of the complexity of the organic content of the wastewater. Elimination
occurred mainly by the destabilization of the particles dispersed in the wastewater,
promoted by a shift in pH and, in the case of real domestic wastewater, by enhanced
agglomerating effects of its constituents. The pH impacts the zeta-potential of the double
layer surrounding the particles, which decreases the surface charge, resulting in
agglomeration of the neutral particles. Aggregation could also occur from adhesion of
NPs into proteins cause their destabilization, and hence, the formation of bigger particles.
4.7
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189
CHAPTER 5
REMOVAL OF Al2O3 NANOPARTICLES DURING ACTIVATED SLUDGE
SECONDARY TREATMENT
5.1
Abstract
Aluminum oxide (Al2O3) is widely used in the fabrication of ceramics, catalysts,
and abrasives, among others, and it is also extensively utilized in the semiconductor
industry for the chemical-mechanical polishing (CMP) of wafers. The stability of
nanosized oxides in sewage is likely to be altered due to the significant changes in the
water chemistry, which might result in particle agglomeration and eventually even
sedimentation. In addition to particle agglomeration and gravitational settling,
interactions of nanoparticles (NPs) with microorganisms involved in biological
wastewater treatment might be an additional mechanism contributing to their removal.
The main goal of this work was to investigate the behavior of Al2O3 NPs during activated
sludge secondary treatment. The pH and organic content of the dispersion were
investigated as possible mechanisms for the removal of the particles. Finally, the effect of
Al2O3 on the performance of the reactor for removing the chemical oxygen demand
(COD) of the wastewater was evaluated. For this purpose a laboratory-scale secondary
treatment was continuously fed with real domestic wastewater containing Al2O3
190
nanoparticles. The average concentration of total Al measured in the samples was
18.10 + 3.88 mg Al L-1 (34.2 + 7.29 mg Al2O3 L-1) in the influent and
2.10 + 0.91 mg Al L-1 (3.94 + 1.69 mg Al2O3 L-1) in the effluent. This information
suggests that the treatment system was suitable for the removal of Al2O3 NPs. The
efficiency of the system for eliminating Al2O3 could be higher than that reported in this
work since Al was found in the real domestic wastewater used to feed the reactor at an
average concentration of 1.04 + 0.68 mg Al L-1. The elimination of particles with
hydrodynamic diameter smaller than 200 nm and 25 nm was evaluated as well. However,
there was no evidence of removal of these small particles since the average concentration
of Al < 200 nm and < 25 nm measured in the samples was similar to that measured in the
wastewater. The addition of Al2O3 to the wastewater greatly affected the efficiency of the
treatment system for removing the total and soluble COD. The average elimination of
total COD reached up to 92% before addition of the NPs and decreased down to 23% in
the presence of Al2O3. However, COD removal efficiency increased with time to an
average of 89%. Batch experiments demonstrated that pH and organic content play a
major role on the aggregation and sedimentation of the particles, enhancing their removal
during treatment. NPs became readily destabilized in Milli-Q water at pH 7.01 and the
initial average particle size (APS) (518.0 + 74.2 nm) measured was up to 2.4 higher than
the APS of NPs dispersed in Milli-Q water at pH 3.10 (214.5 + 8.66 nm). The initial APS
of NPs dispersed in real domestic wastewater at pH 6.97 was equal to 1.03 + 0.26 µm,
suggesting that, not only the pH, but also the presence of organic matter promotes the
destabilization and agglomeration of the particles.
191
5.2
Introduction
Engineered NPs are materials with at least one dimension of 100 nanometers (nm)
or less. They are used for numerous applications in the manufacturing of electronics,
cosmetics, pharmaceuticals, and biomedicine, among others [1]. Due to their small size
range, the fraction of atoms exposed at the NP surfaces is significant, contributing to
modifying the physical, chemical, electronic and atomic properties of the particles
compared to their bulk particle counterparts[2]. Therefore, NPs can act as an effective
bridge between the bulk and the molecular properties of the materials [3]. NPs provide
enhanced characteristics to products, that otherwise would not be possible or cost
effective, resulting in an increased utilization of these materials in the last years. There
were more than a thousand nano-enable consumer products identified in 2009, produced
by 50 companies distributed in 20 countries [4]. Increased concern about the presence of
NPs in consumer products and the environment results from the possible toxic effects of
nano-sized materials to living organisms [5-7].
Nano-metal oxides are the nanomaterials (NMs) group with most commercial
applications to date [8]. Al2O3, also called alumina, together with other metallic NMs
accounted for up to 83% of the NP market in 2005 [9]. Alumina is commonly used in the
manufacturing of abrasives, catalysts, refractories and electrical insulators [10].
Moreover, Al2O3 is an important ceramic material that is extensively employed for
electrical and biomedical applications [11]. Alumina NPs have numerous applications in
the electronic, optical, and magnetic industry [12]. For microelectronic purposes, it is
192
used during the CMP of wafers, which is an important process to achieve even surfaces
free of defects [13]. For this purpose, slurries containing metal-oxide NPs, including
Al2O3, surfactants and additives at controlled pH, are applied to the wafers [14]. The
CMP slurry market was estimated at $400 million dollars in 2003 [15]. Although Al2O3
NPs are of great commercial importance, the adverse effects to human health of such
materials it is still not fully explored.
The very same properties that make NPs interesting for their application in
products and processes raise concern about the adverse effects of such materials on
biological systems [16], resulted from their small size and high surface area [17].
Toxicity of Al2O3 seems to increase as particle size decreases. A study comparing microand nano-sized Al2O3 showed that NPs posed higher toxicity to certain bacteria than their
bulk counterparts [18]. Additionally, Al2O3 NPs caused considerable damage to human
brain microvascular endothelial cells [19], and they might be as well responsible for the
breakdown of the blood-brain barrier [20]. However, Al2O3 NPs toxicity seems to depend
on the target cell or organisms, as relative no adverse effects were observed when HT22
cells, taken from rodent nervous system, were exposed to Al2O3 [21].
Although unintentional and controlled releases/discharges of engineered NPs to
the environment become inevitable as their use increases in household products and
industrial processes, the fate of such materials once in the environment is largely
unknown. Modeling studies have suggested that a considerable fraction of these NPs can
find their way to the municipal wastewater treatment plants (MWWTPs) [22, 23].
Nevertheless, there is not much information about the behavior of metal-oxide NPs
193
during wastewater treatment. The detection of titanium dioxide (TiO2) NPs in the effluent
of several treatment plants [24] might suggest that MWWTPs are not suitable for treating
wastewater containing these emerging contaminants.
Different mechanisms have been proposed for the removal of NPs from
wastewater during treatment, including sorption and entrapment onto biomass [25]. A
few reports exist suggesting that the removal efficiency of the treatment for eliminating
the NPs in wastewater will depend on the physicochemical characteristics of the particles
and the inorganic and organic components of the wastewater. For example Limbach et al
[26] observed high elimination of CeO2 NPs dispersed in synthetic media, where the
average efficiency in the simulated secondary treatment was 94%. Moreover, Jarvie et al
[27] studied the sedimentation of surfactant-coated and uncoated SiO2 NPs in a
bench-scale primary treatment using screened and unscreened domestic wastewater. Only
Tween-coated NPs were effectively removed during primary treatment.
In this work the fate of Al2O3 NPs during laboratory-scale activated sludge
treatment was investigated. Results obtained in this study provide an important insight
about the behavior of such NPs during wastewater treatment, as no previous reports have
been published. In contrast with former studies, NPs were introduced into real domestic
wastewater, and the secondary treatment was operated over extended time to simulate the
operation of large scale system.
194
5.3
Materials and Methods
The Al2O3 NPs were obtained from Sigma-Aldrich (Sigma-Aldrich, St Louis,
MO) as a fine powder, and according to the manufacturer the average particle size was
50 nm.
5.3.1 Stability of Al2O3 NPs in aqueous dispersions
The Al2O3 NPs were suspended in Milli-Q water and real domestic wastewater
with the objective to study the impact of pH and inorganic/organic components over the
stability and aggregation of the particles. For this purpose an Al2O3 stock dispersion
containing 200.0 mg Al L-1 (377.9 mg Al2O3 L-1) was prepared in 1 mM HCl by applying
sonication (Daigger GEX130, 130W, Cole-Parmer Instruments, Vernon Hills, IL) for 15
min at 70% intensity. Then, 1 mL of the stock was diluted in 10 mL Milli-Q water at pH
3.10 and 7.01 as well as in 10 mL of real domestic wastewater at pH 6.97. Treatments
were prepared in duplicate in falcon tubes (BD Biosciences, Bedford, MA), and in the
case of wastewater, it was centrifuged for 30 min at 4000 rpm and passed through 25 nm
membrane filters (Millipore, Billerica, MA) to remove any suspended solid that could
interfere with the analysis. A 1 mL sample was taken after vortex mixing the treatments
for 30 s, which allowed the determination of the initial zeta-potential and particle size
distribution (PSD). Subsequently, treatments were incubated for 24 h without mechanical
mixing, allowing the sedimentation of the aggregates. Finally, the upper 15% of the
195
supernatant formed in each treatment was sampled and digested to measure the remaining
Al2O3 NPs in suspension. The treatment consisting of Milli-Q water at pH 3.10 was
vigorously vortex mixed after taking the supernatant sample with the purpose of
determining the total concentration of Al in the treatments.
5.3.2
Laboratory-scale secondary treatment
The fate of the Al2O3 NPs was studied in a laboratory-scale activated sludge
treatment consisting of an 1.19 L aeration tank and a 0.66 L settler, all constructed in
acrylic plastic (Figure 5.1). The real domestic wastewater was collected after primary
clarification from Roger Road, a local WWTP located in Tucson, AZ, in a weekly basis
and was kept under refrigeration to avoid the decomposition of the organic constituents
of the water. The reactor was seeded with returned activated sludge (RAS) obtained from
the same plant. In order to avoid any interference of pre-existing Al in the sludge, it was
washed 7 times with Milli-Q water. The biomass concentration, measured as volatile
suspended solids (VSS), in the aeration tank was set equal to 2.55 g VSS L-1. The Al2O3
NPs were introduced into the secondary treatment system via a stock dispersion. The
Al2O3 stock preparation was a two-step process. First a concentrated stock was prepared
by dispersing the NPs in 1 mM HCl using sonication at 70% intensity for 15 min to an Al
concentration of 2.00 g Al L-1 (3.78 g Al2O3 L-1). Then the stock was further diluted in
1 mM HCl to 200 mg Al L-1 (378 mg Al2O3 L-1). The diluted stock was kept under
constant mixing to avoid aggregation of the NPs. The diluted stock was mixed with the
196
wastewater
w
in
n the entran
nce of the aeration
a
tankk to achievee an Al conncentration in the
in
nfluent of 18
8.2 mg Al L-1 (34.2 mg Al
A 2O3 L-1).
The average
a
pH and
a chemicaal oxygen deemand (COD
D) of the inffluent wastew
water
measured
m
after dilution with
w the NP
Ps was 7.26 + 0.22 andd 313.4 + 888.7 mg COD
D L-1,
reespectively. The averag
ge flow ratee was 2.99 + 0.15 L d-1, which ccorresponds to a
hy
ydraulic reteention time (HRT)
(
equal to 9.49 + 00.12 h (Tablle 5.1). The recirculationn of a
frraction of th
he settle slud
dge and air and mixingg was achievved by usinng two air puumps
(A
Aqua, Walm
mart) providin
ng a flow ratte of 400 L d -1.
2
4
1
3
5
Figure
F
5.1. Schematic of the labo
oratory aerobbic activateed sludge trreatment syystem.
(1
1) Aeration tank,
t
(2) air, (3) settler, (4)
( effluent, (5) recycledd sludge.
197
The fate of the NPs and the performance of the reactor were monitored by taking
samples from the diluted stock, the influent wastewater containing the NPs, and the
clarified effluent. The COD and acetate concentration were measured in the influent and
effluent samples. The pH was monitored in the influent and effluent samples as well. In
order to maintain a relatively constant biomass concentration in the reactor, 90 mL of
sludge were taken from the aeration tank and replaced with wastewater (without Al2O3
NPs) every other day. Since the aluminum has an inherent alkalinity, the pH of the NP
stock was measured every 2 days and adjusted adding 0.03 mL of concentrated HNO3 to
a 1 L stock. Keeping the low pH values avoids agglomeration of the particles. Finally the
behavior of the Al2O3 NPs was monitored by measuring the total concentration in the
samples, as well as the concentration of Al2O3 that passed through a 200 nm and 25 nm
membrane filters.
Images from the original sludge used to seed the reactor and samples of sludge
exposed to Al2O3 NPs taken from the aeration tanked were obtained using scanning
electron microscopy (SEM). The presence of Al was determined by energy dispersive
spectroscopy (EDS).
198
Table 5.1. Operational parameters of the laboratory aerobic activated sludge treatment in
the absence and presence of Al2O3 NPs.
Operation time (days)
Real wastewater (No
Al2O3 added)
24
Real wastewater (Al2O3
added)
167
Θ in the aeration tank (h)
9.41 + 0.77
9.58 + 0.43
Average pH In
7.37 + 0.09
7.24 + 0.25
Average pH Out
6.57 + 0.27
6.76 + 0.26
1.88 + 0.46
2.03 + 0.52
Parameter
Mixed liquor volatile
suspended solids
concentration (g/L)
5.3.3 Analytical procedure
The total and soluble COD concentrations in liquid samples were measured
during secondary treatment. The soluble fraction was obtained by centrifuging the
samples for 10 min at 11,000 rpm. The COD was determined by mixing 2.5 mL of
sample, either soluble or total, with 1.5 mL digestion solution and 3.5 mL sulfuric acid
concentrated (H2SO4). The digestion solution consisted of (g L-1): H2Cr2O7 (10.22),
Hg2SO4 (33.30), and H2SO4 (167 mL L-1). The H2SO4 concentrated contained
10.25 g Ag2SO4 L-1. Digestion of the resulting mixture occurred at 150ºC for 2 h. The
199
COD content was then obtained by measuring the absorbance of the digested samples
using a spectrophotometer (Beckman-Coulter, Fullerton, CA) at a wavelength of 600 nm.
The fate of acetate, a simple organic acid, was followed in the activated sludge
treatment system as well. For this purpose 1.5 mL of sample from the influent and
effluent were centrifuged for 15 min at 10,000 rpm. Then, 1 mL of the formed
supernatant was transferred into a 1.5 mL glass vial and amended with 0.5 µL of
concentrated formic acid. Finally, acetate was measured using a gas chromatograph (GC)
(7890A, Agilent Technologies, Palo Alto, CA) equipped with a flame ionization detector,
and a 30 m x 0.53 mm Stabilwax-DA column (Restek, Bellefonte, PA). The carrier gas
used was ultra-high purity helium at a flow rate of 16 mL min-1. The liquid samples were
injected into the GC by an autosampler applying a 6:1 (v/v) split ratio. The inlet and
detector temperatures were equal to 250ºC and 275ºC, respectively.
As previously mentioned, the biomass content in the aeration tank was determined
by measuring the VSS content. Twenty milliliters of sample were vacuumed filtered
using 45 nm glass fiber filters (Whatman. Piscataway, NJ). The filters with the retentate
were placed in aluminum vessels (Fisher Scientific, Pittsburgh, PA) and dried overnight
at 105oC. Then, the samples were ashed in a muffle furnace at 550oC for 4 h.
Subsequently, containers were allowed to rich ambient temperature in a dissecator and
weighted.
200
Total Al was measured in the stability test and in the samples taken from the
activated sludge treatment samples, including the diluted stock. A volume of 5 mL of
sample was mixed with 5 mL of HCl 6.75 M (Fisher Scientific, Pittsburgh, PA) and
digested in a Mars 5 microwave digestion oven (MDS2100, CEM Corp., Matthews, NC)
at 71 PSI for 30 min. In the case of the diluted stock only, 1 mL was used for the
digestion due to its high Al concentration; nevertheless it was amended with 4 mL MilliQ water. After digestion the effluent samples were diluted with 10 mL Milli-Q water, and
the diluted stock and influent digested samples were diluted with 25 mL Milli-Q water.
Digested samples were placed in 15 mL falcon tubes and analyzed in an inductively
coupled plasma – optical emission spectrometer (ICP-OES, Optima 2100 DV, Perkin
Elmer, Waltham, MA) at a wavelength of 396 nm. The fate of Al < 200 nm and < 25 nm
was monitored in the samples from the secondary treatment. For this purpose samples
were passed, using a 3 mL plastic syringe, through 200 nm (Whatman, Piscataway, NJ)
and 25 nm (Millipore, Billerica, MA) membrane filters. After filtration 0.015 mL HNO3,
were added to preserve the samples. Finally, 2.5 mL of the filtered sample were diluted
with 2.5 mL 4% HNO3 (v/v) and directly measured in an ICP-OES.
The APS and zeta-potential were measured in a Zeta Sizer Nano ZS (Malvern,
Inc., Westborough, MA) by light scattering caused by the NPs. The equipment was set at
a refractive index and absorption value equal to 1.828 and 0.01, respectively, using water
as dispersant at a 25oC. This operation was accomplished by filling a DTS1060C
capillary vessel with 1 mL of sample, carefully to avoid air bubbles that could interfere
with the analysis.
201
A variable that was measured in all the samples was the pH. It was determined
using a probe (Orion-Thermo Scientific, Cincinnati, OH) calibrated before measurements
with certified standards at pH 4.0, 7.0, and 10.0 (Orion-Thermo Scientific. Cincinnati,
OH).
5.3.4
Electron microscopy
The SEM and EDS analysis of the sludge samples were executed in the
University spectroscopy and imaging facilities (USIF) at the University of Arizona. The
SEM the samples were fixed using formaldehyde in a 50% (v/v) ratio 4 h before the
preparation of the samples. Then, samples were vacuumed dried and manually broken
until forming a fine powder, which was mounted and glued into a plastic base. The
specimen was finally coated with a thin layer of gold (Au) and palladium (Pd) to protect
the sample from the electron beam. EDS measurements were done by directing an x-ray
beam into the sample and measuring the energy reflected and measured by the detector.
SEM and EDS analysis were performed in a S-4800 UHR field emission scanning
microscope (Hitachi, Tokyo, Japan) at 5 kV.
202
5.4
Results
5.4.1 Stability of Al3O2 NPs
In this study, the stability of Al2O3 NPs dispersions in Milli-Q water, at pH 3.10
and 7.01, as well as in wastewater at its natural pH (6.97), was investigated. The study
was conducted to get a better understanding about the main factors that promote the
aggregation (if any) of the particles. Figure 5.2a shows the APS of the NPs at the
beginning of the experiment and the supernatant after 24 h incubation without mechanical
mixing. The average hydrodynamic diameter of the Al2O3 NPs was bigger than the 50 nm
primary particle size reported by the manufacturer. At low pH, the APS was
214.5 + 8.66 nm. The APS of the low pH dispersion remained constant over time and
only changed marginally during incubation. When the NPs were suspended in Milli-Q
water at neutral pH the APS right after dispersion was 518.0 + 74.2 nm and increased to
938.6 + 173.9 nm after 24 h of incubation. The results suggest that aggregation is
considerably enhanced at circumneutral pH values.
Significant agglomeration of Al2O3 was observed in the presence of real domestic
wastewater. At the beginning of the experiment the APS in the treatment containing
wastewater was 1.03 + 0.26 µm and the APS increased to 1.46 + 0.09 µm over the 24 h
incubation period. The data show that Al2O3 NPs in wastewater readily aggregate to form
particles in the micron-size range. Moreover, when comparing these results to those
obtained for the Milli-Q water at neutral pH treatment, it is evident that certain
203
constituents of the wastewater promote agglomeration of the Al2O3 NPs. The zetapotential values of the samples were measured as well and these are shown in
Figure 5.5b. The average zeta-potential for the treatment consisting of Milli-Q water at
pH 3.10 did not vary considerably over time and it was equal to 37.8 + 0.21 mV during
the experiment. As observed with the size measurements, the zeta-potential for the
Milli-Q water at pH 7.01 treatment varied with time, decreasing from an initial value of
22.4 + 2.51 mV to only 4.73 + 2.49 mV after 24 h. Comparing these two Milli-Q water
treatments, the data suggest that low absolute values of zeta-potential promote
agglomeration of the particles. An exception was observed in the treatment containing
real domestic wastewater. Although the zeta-potential values at the beginning
(-23.3 + 1.61 mV) and at the end (-22.3 + 0.82 mV) of the experiment were comparable
to those measured in the Milli-Q water pH 7.01 treatment, the APS of the Al2O3 in
wastewater was considerably higher.
204
Figure
F
5.2. a)
a Average particle size and
a b) zeta-ppotential of CeO2 in diff
fferent matricces at
t=
=0 (□) and t=24 h (■)) incubation
n. MQ = M
Milli-Q wateer (pH 3.100 and 7.01)) and
WW
W = Real wastewater
w
(pH
( 6.97).
205
The NP dispersion stability was also evaluated by monitoring the residual Al2O3
NPs remaining in suspension 24 h after incubation (Figure 5.3). To determine the
maximum Al content expected in the supernatant of the treatments, the treatment at pH
3.10 was sampled after diluting the NPs and used as a control, giving a concentration
equal to 17.1 + 0.84 mg Al L-1 (32.3 mg Al2O3 L-1), which was close to the designed Al
concentration set at 18 mg Al L-1 (34 mg Al2O3 L-1). Although the APS did not increase
considerably in the treatment at pH 3.10 and remained constant over time, surprisingly
around 30% of the particles sedimented during the experiment based on total Al
concentration measurements. The residual Al concentrations in the treatments consisting
of Millli-Q water at pH 7.01 and real wastewater were 5.71 + 2.66 (10.8 + 5.02) and
6.86 + 1.39 mg Al L-1 (12.9 + 2.62 mg Al2O3 L-1), respectively, which represent an
average drop off in the Al concentration of 66% for the treatment consisting of Milli-Q
water at pH 7.01 and 59% for the real wastewater treatment, compared to the control
(Milli-Q water at pH 3.10). Even though the final APS in the treatment containing
wastewater was more than 1.5 higher, the remaining Al content in both treatments was
very similar.
206
Figure
F
5.3. Al
A concentrration in thee supernatannt from diffe
ferent matricces after t =
=24 h
in
ncubation. Control
C
= Miilli-Q water at t=0 (pH 33.10); MQ = Milli-Q waater (pH 3.100 and
7.01); WW = Real wasteewater (pH 6.97).
6
The iintended Al concentratioon at t=0 waas set
-1
-1
eq
qual to 18 mg
m Al L (34
4 mg Al2O3 L ).
5.4.2 Removal of Al3O2 NPs from real domesstic wastewaater during activated slludge
trreatment
b
of Al
A 2O3 NPs suspended iin real wasteewater was investigatedd in a
The behavior
laaboratory-scale activated sludge trreatment sysstem. Initiallly the secoondary treattment
sy
ystem was fed
f with reaal domestic wastewater. After 24 ddays, NPs w
were continuously
in
ntroduced in
nto the systeem via a stocck dispersioon which waas diluted w
with real dom
mestic
wastewater
w
to
o achieve a design conccentration off 18.2 mg A
Al L-1 (34.2 m
mg Al2O3 L-1) in
207
the influent. The fate of total Al in the system is presented in Figure 5.4a. The average
concentration of total Al measured in the influent was 18.10 + 3.88 mg Al L-1
(34.2 + 7.29 mg Al2O3 L-1), whereas in the influent was found to be
2.10 + 0.91 mg Al L-1 (3.94 + 1.69 mg Al2O3 L-1). The Al concentration in the
wastewater was measured before starting the experiment to determine its natural Al
background concentration. The Al content of these samples was below the detection limit
of the ICP-OES (<1 part per billion). However, an average concentration of
1.04 + 0.68 mg Al L-1 was found in the wastewater collected from the MWWTP during
the operation of the reactor, which is considerably higher than the preliminary
measurements. After stopping the secondary treatment, additional measurements revealed
no detectable Al in the wastewater. This could indicate that Al was added intermittently
to the wastewater during treatment in the MWWTP. The natural soluble Al content in the
wastewater is indicated in Figure 5.4a by a horizontal line, and the upper and lower
broken lines represent the standard deviation.
The fate of the < 200 nm and < 25 nm particles was studied by serial filtrations
and is shown if Figure 5.4b. The real domestic wastewater was filtered through a 200 nm
membrane filter and its Al content was measured, which is indicated by the horizontal
lines. As it can be observed, the Al concentration in the influent and effluent samples
taken from the secondary treatment was masked by the Al < 200 nm occurring in the real
domestic wastewater. Results indicate no elimination of Al < 200 nm and Al < 25 nm
during treatment since the Al content measured in the samples was similar to that
observed in the real domestic wastewater.
208
Figure
F
5.4. Fate
F of a) totaal unfiltered and b) filterred Al in thee (●) influennt and (■) efffluent
of activated sludge treatm
ment. Bold markers
m
in se cond panel iindicate Al2O3 < 200 nm
m,
while
w
empty markers
m
refeer to Al2O3 < 25 nm. Horrizontal linee represents tthe average A
Al
co
oncentration
n in wastewaater; broken lines
l
indicatte the standaard deviationn.
209
A sample of the original sludge used to seed the reactor was imaged and EDS
analyzed together with a sample taken from aeration tank at the end of the experiment,
exposed to Al2O3 for 167 days. Figure 5.5 shows the image of both samples and their
EDS spectrum. Although the SEM shows a similar composition of the samples,
consisting of bacteria, filamentous microorganisms, and extra-cellular material, the Al
counts greatly varied. The presence of a small amount of Al in the original sludge could
suggest the utilization of Al in the MWWTP, and might explain its presence in the
wastewater used to feed the laboratory-scale secondary treatment. Even though the
number of counts does not represent the concentration of an element in a sample per se,
higher counts indicate higher amounts of the target element under study. The Al counts in
the sample of sludge exposed to Al2O3 were up to 27 times higher than those in the
original sludge, suggesting sedimentation of the particles added to the wastewater and
possible entrapment or sorption onto the biomass.
210
Figure
F
5.5. SEM
S
image of the origiinal sludge uuse to seed the reactor (left) and slludge
ex
xposed to Al2O3 NPs forr 167 days (rright). EDS spectrum off the sludge (bottom leftt) and
th
he exposed sludge
s
(botto
om right).
5.4.3
Fate of
o the organiic content in
n the secondaary treatmennt
The capability of the activated sludge treeatment for rremoving thee organic coontent
frrom the wasstewater wass evaluated by
b measurinng the COD and acetate concentratiion in
th
he influent and
a effluentt samples. The
T secondarry treatmentt system opperated durinng 24
days without addition of the Al2O3 NPs
N in orderr to allow accclimatizatioon of the bioomass
to
o the new environmenta
e
al condition
ns. During thhis period tthe average removal off total
211
COD was found to be 49%, as the average influent concentration was
247.2 + 42.35 mg COD L-1 and 126.3 + 32.60 mg COD L-1 in the effluent (Figure 5.6a).
The performance of the reactor for removing total COD gradually increased with time
and from day 25 to day 38 the average total COD removal was 87%. The total COD
content in the influent and effluent samples was 270.4 + 65.54 mg COD L-1 and 34.29 +
10.16 mg COD L-1, respectively. However, the efficiency of the treatment system
unexpectedly decreased considerably after day 40 and a significant and steady increase in
the COD effluent concentration was observed until day 81. Finally, during the last 91
days of operation of the reactor, from day 98 to day 190, the COD content in the samples
remained rather low and fairly constant. The average COD concentration in the influent
and effluent was 331.9 + 103.3 mg COD L-1 and 44.9 + 24.7 mg COD L-1 respectively,
resulting in an average elimination of 86%. The COD data suggest that the period of low
removal efficiency could have been caused by the presence of Al2O3 in the reactor since
the COD content in the influent wastewater showed little variation,. The elimination of
the soluble COD fraction followed a similar trend as that one observed in the removal of
total COD (Figure 5.6b). The performance of the reactor considerably decreased after 40
days of operation. The efficiency of the treatment system for eliminating soluble COD
was more affected than for removing total COD. After 74 days, when the soluble COD
content in the effluent decreased and stabilized, the average soluble COD concentration
in the influent was equal to 131.4 + 60.46 mg COD L-1 and that of the effluent was
30.40 + 13.62 mg COD L-1. From day 74 until the end of the experiment on day 190 the
average soluble COD removal was 77%.
212
Figure
F
5.6. a)
a Total and
d b) soluble COD conceentration in the aerobic activated slludge
sy
ystem, wherre (●) indicaates the inflluent and (■
■) the efflueent. Feedingg of Al2O3 tto the
reeactor is rep
presented by
y the contin
nuous verticcal line. Doted lines inndicate freshh real
domestic wasstewater batcches.
213
The acetate content in the influent and effluent samples was measured and is
presented in Figure 5.7. Acetate degradation is vital to maintain a circumneutral pH in the
reactor, since microbial activity is decreased at low pH values. The figure shows despite
the fact that acetate concentration in the influent wastewater varied significantly, it was
highly removed without any lag phase and its elimination increased with time. The
average acetate concentration in the influent samples was found to be 29.7 + 17.6 mg
acetate L-1 which reflects the great variability in the wastewater. Acetate content in the
effluent was 0.93 + 0.52 mg acetate L-1, resulting in an average elimination of 96%. More
importantly, the behavior observed in the removal of total and soluble COD was not
detected in the case of acetate.
214
Figure
F
5.7. Concentratio
C
n of acetate in the influeent (●) and effluent (■) of the seconndary
acctivated slud
dge treatment. Continuo
ous verticall lines speciify the addiition of Al2O3 to
wastewater.
w
Doted
D
verticaal lines repreesent fresh w
wastewater bbatches.
215
5.5
Discussion
5.5.1
Fate of Al2O3 during wastewater treatment
The data obtained in this work suggest that significant removal of Al2O3 can be
achieved during activated sludge secondary treatment. The content of total Al was
decreased from an average of 18.10 + 3.88 mg Al L-1 (34.2 + 7.29 mg Al2O3 L-1) in the
influent to only 2.10 + 0.91 mg Al L-1 (3.94 + 1.69 mg Al2O3 L-1) in the treated effluent.
However, the presence of Al in the real domestic wastewater used to feed the reactor does
not allow determining the real capacity of the treatment system for removing Al2O3 NPs.
The content of Al measured in the wastewater during the operation of the reactor was
very close to that measured for Al < 200 nm and Al < 25 nm, suggesting no effective
removal of the very fine particles. The presence of Al in the wastewater suggests its
utilization in the local MWWTP where the wastewater batches were taken from. Alum
(Al2(SO4)3 nH2O) is commonly used during wastewater treatment for coagulation
purposes [28] and might explain the presence of Al in the wastewater. Apparent removal
of soluble Al from the real domestic wastewater seemed to occurred during treatment, as
the Al content for certain effluent samples was lower than that measured in the
wastewater.
216
Different studies in laboratory-scale reactors have proved that activated sludge
processes are highly efficient for the elimination of metal-oxides NPs. Limbach et al [26]
observed an average removal of 94% CeO2 NPs and Kiser et al [24] documented
efficiencies of up to 88% when studying elimination of TiO2 NPs. However, NPs were
introduced into synthetic wastewater, instead of real domestic wastewater and the
reactors were operated only for short periods of time, providing limited information about
the behavior of the particles in a MWWTP over the long term.
The high elimination of total Al2O3 from the influent wastewater suggests the
formation of large particles that sedimented onto the biomass in the aeration tank. The
SEM image obtained from the original sludge used to seed the reactor and a sample taken
from the aeration tank when the secondary treatment was stopped reveals a similar
biological composition. An EDS analysis of the samples exposed the presence of Al in
both cases. The presence of Al in the original sludge strengthens the possibility of its
utilization at the MWWTP. Nevertheless, the number of counts was relatively small
compared to the sample taken from the reactor at the end of the experiment. Additionally,
these observations indicate that Al accumulated in the reactor and that biomass might
play an important role in the removal of Al2O3 from wastewater. Evidence exists that
silver sulfide NPs can bind to biological sludge during treatment in MWWTPs [29]. The
sorption capacity of the biomass depends on the physical and chemical characteristics of
the particles as observed by Kiser et al [30], who reported 97% removal of silver NPs but
only 13% of fullerol NPs when exposed to wastewater biomass...
217
5.5.2 Mechanisms of Al2O3 removal
As Al2O3 is taken from the continuously stirred diluted stock dispersion to the
influent real domestic wastewater, particles are subjected to a pH change from an average
3.6 up to 7.2. Such a significant change of pH was expected to cause agglomeration of
the particles since the isoelectric point (IP) of the Al2O3 has been estimated to be at pH
7.9 [31]. At the IP, particles become neutral and the repulsion between them decreases.
The zeta-potential of NPs added to Milli-Q water at pH 3.10 remained constant over a
24 h period and was equal to 37.8 (+ 0.21) mV. When NPs were suspended in Milli-Q
water at pH 7.01 the zeta-potential was equal to 4.73 (+ 2.49) mV after 24 h. These
results show that the absolute value of the zeta-potential is lower at high pH, and are
comparable to observations by Gosh et al [31] who measured decreasing zeta-potential
values on the surface of Al2O3 NPs with increasing pH. Moreover, data suggest that the
zeta-potential can decrease with time when the pH of the dispersion is close to the IP.
The zeta-potential represents the electric potential between the surface of the particle, and
the bulk medium and is affected by the presence of ions in the dispersion. Charged
particles repeal each other, whereas when neutralized are attracted by van der Waals
forces, resulting in the formation of bigger particles. This was confirmed by the APS
measured from the supernatant formed after incubation of the treatments consisting of
Milli-Q water at pH 3.10 and 7.01. The average size obtained from the treatment at pH
7.01 was up to 3.4 times higher than that at pH 3.10. The data obtained indicate that the
pH of the dispersion influence the zeta-potential of the particles, and hence their size
218
distribution. Moreover, it was observed that the absolute value of the zeta-potential
decreased with time, resulting in further agglomeration. This behavior is not uncommon
as increased size over time for certain NPs has been previously documented [32]. Even
though particles were highly charged at pH 3.10 some degree of agglomeration was
observed since the size of the Al2O3 NPs claimed by the manufacturer was 50 nm and the
average hydrodynamic size measured after sonication was 214.5 nm. Different NPs,
including TiO2, Fe2O3, ZnO, and NiO, agglomerated when dispersed in water and cannot
be broken down into primary particles after extended ultrasonication [33, 34]. Despite the
zeta-potential and the APS remained constant during the experiment for the treatment at
pH 3.10, around 30% of the particles settled over time. This observation suggests the
formation of even larger agglomerates that did not remain in suspension.
The agglomeration of the Al2O3 NPs was influenced not only by the pH but also
by certain constituents of the wastewater, since Al2O3 NPs diluted in real domestic
wastewater at pH 6.97 showed a considerably higher APS than that at pH 7.01 in Milli-Q
water. Al2O3 is known to bind to humic aicd and other organic compounds [35, 36], and
this capacity can be increased by the presence of certain bridging compounds, including
polymerin and sodium dodecylsulfate (SDS) [37, 38]. Different studies have reported the
stabilization of certain NPs in the presence of organic matter and found it to be
independent of the pH of the dispersion [31, 39, 40]. However evidence exist that fulvic
acids and other organic polymers can destabilize colloids present in water [41, 42]. The
Al content in the supernatant, after incubation, from the treatment containing wastewater
was similar to that observed in Milli-Q water at pH 7.01, indicating that even larger
219
particles can remain in suspension by their interaction with simple organic molecules in
the wastewater. The high removal efficiency achieved during secondary treatment
corroborated the findings from the stability test, where the pH of the dispersion and the
organic matter in the wastewater were identified as the principal mechanisms for the
removal of the particles.
5.5.3
Impact of Al2O3 on the performance of the activated sludge treatment
The introduction of Al2O3 into the activated sludge secondary treatment system
disrupted the overall performance of the reactor for removing the organic content from
the wastewater. Microorganisms were affected by the presence of Al as demonstrated by
the low total and soluble COD removal measured approximately 14 days after the feeding
of Al2O3 NPs. The degradation of the soluble COD fraction in the reactor was more
affected than the removal of total COD by the presence of Al, where at the most critical
point removal decreased down to 17%. It is known that Al can be toxic to
microorganisms. Toxicity can be caused by the binding of Al to cell walls, enzymes,
substrates, DNA, and ATP, by the complexation of Al to nutrients, and by a decrease of
the pH of the medium [43, 44] Guida et al [45] reported growth inhibition when studying
the effects of aluminium nitrate to Escherichia coli and found it to be dependent of pH.
Illmer et al [46] observed severe inhibition of nitrogen mineralization by microorganisms
in soils exposed to Al
The COD removal efficiency steadily increased with time and
during the last 92 days of operation of the reactor remained constant at an average
220
elimination of 76%, a clear indication that microorganisms adapted to Al2O3. More
importantly, accumulation of the Al particles did not cause further inhibition. Different
studies have revealed that certain microorganisms can tolerate the presence of Al.
Pseudomonas fluorescens was able to grow in in mineral media containing up to
50 mmol Al L-1 [47]. Fischer et al [48] observed increased resistance of a
chemolithoautotrophic acidophilic bacteria, Acidiphilium cryptum, after precultivation in
1
mmol
L-1
Al2(SO4)3
and
subsequent
transfer
to
a
medium
containing
300 mmol L-1 Al2(SO4)3. The continuous high removal of acetate from the wastewater
revealed that microbial activity for acetate degradation was not affected by the presence
of Al2O3, suggesting that the partial inhibition caused
by Al affected only the
degradation of higher molecular weight organics.
5.6
Conclusions
The fate of Al2O3 NPs was studied during activated sludge secondary treatments.
The high elimination of Al was observed, suggests that Al2O3 in wastewater can be
effectively removed in MWWTPs. The pH of the wastewater played a major role for the
elimination of the Al, since it caused a decreased of the zeta-potential of the Al2O3
particles.. The destabilization of the nano-suspension caused by the pH of the wastewater
resulted in agglomeration of the Al2O3, which enhanced the removal of Al. The
efficiency of the treatments system for eliminating COD was affected by the presence of
Al. Microorganisms responsible for the degradation of COD were partially inhibited.
221
However, the microbial community in the reactor adapted and recovered to the presence
of Al over time.
5.7
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CONCLUSIONS
Organic and inorganic contaminants present in gasoline plumes and landfill
leachates are known to be persistent in nature and toxic to living organisms. In this
dissertation different biological techniques were explored for their potential use for the
treatment of landfill leachates and contaminated groundwater.
Different inocula were used to study the anaerobic degradation of toluene,
benzene, m-xylene, and cis-DCE using nitrate and chlorate as alternative electron
acceptors. Toluene degradation occurred under nitrate and methanogenic reducing
conditions; however higher energy yield was obtained from the reduction of nitrate, since
toluene removal occurred at higher rate in the presence of nitrate. In this case, toluene
degradation was linked to denitrification as nitrite was formed in significant amounts.
Although various respikes and transfers were performed in treatments that exhibited
toluene removal, enrichment cultures were not obtained. The fact that toluene removal
stopped after just a few transfers suggests that degradation occurred via cometabolism.
Microorganisms capable of toluene removal using chlorate as electron acceptor were not
observed.
m-Xylene can be degraded under methanogenic conditions after extended
incubation. Moreover, microorganisms can adapt to the presence of m-xylene over time,
as demonstrated by the fast degradation of the fresh m-xylene respiked to treatments that
showed m-xylene removal. Pure cultures could be obtained after serial transfers.
228
Surprisingly microorganisms responsible for the degradation of m-xylene using
nitrate or chlorate as electron acceptors were not found in the inocula tested. Benzene can
be degraded in the absence of oxygen when nitrate and chlorate are used as electron
acceptors. Nevertheless, this is not a common process since removal was observed after
extended incubation only in a few experiments. Microorganisms responsible for benzene
degradation under methanogenic conditions were not found in the inocula used in the
experiments.
cis-DCE
proved
to
be
recalcitrant
in
anaerobic
environments.
Microorganisms capable of using cis-DCE as carbon source were not present in the
inocula tested.
The treatability of a synthetic landfill leachate containing volatile fatty acids,
phenol, p-cresol and ammonia was investigated in an anaerobic – aerobic system
configuration. The high performance of the reactors demonstrates that the system can be
used for the treatment of landfill leachates. The upflow anaerobic sludge blanket (UASB)
is suitable for removing the organic content of the leachate, as indicated by the high
elimination of the chemical oxygen demand. During anaerobic treatment in the UASB,
organics were converted to methane gas. Microorganisms in the reactor can rapidly adapt
to phenol and p-cresol since these contaminants were removed with no lag phase.
Volatile fatty acids are highly degraded in the UASB and their removal increases with
time. The persistence of ammonia in the leachate after treatment indicates that anaerobic
processes are not suitable for its removal.
229
The aerobic downflow hanging sponge (DHS) column proved to be suitable for
the nitrification of ammonia, as demonstrated by the high removal achieved during
treatment. The extremely low organic content of the influent suggests that the consortia
of microorganisms responsible for nitrification are autotrophs. Oxidation of ammonia in
the DHS is divided in two major processes: 1) oxidation of ammonia to nitrite and 2)
oxidation of nitrite to nitrate. Ammonia oxidizing bacteria, responsible for partial
nitrification to nitrite, dominate during the first days of operation; however ammonia
oxidizers and nitrite oxidizers dominate over time, demonstrated by the complete
nitrification of ammonia achieved during extended operation of the reactor.
The fate of different oxide nanoparticles during domestic wastewater was studied
in this dissertation. The adverse effects of oxide nanoparticles are not well understood;
however evidence suggests that these materials pose a risk to human health. To date, it is
still unknown if the current wastewater treatment processes are suitable for the removal
of nanoparticles, or if these materials escape in the treated effluents.
The behavior of cerium oxide nanoparticles was investigated during activated
sludge secondary treatment of synthetic and real domestic wastewater. Data shows that
the secondary treatment is adequate for the elimination of the particles in the influent
wastewater. Efficiency was higher when using real wastewater, suggesting that
inorganic/organic components lacking in the synthetic wastewater enhance further
removal of particles. Particles agglomerate when diluted in synthetic and real wastewater
during treatment since less than 15%, at best, of the total ceria showed to be smaller than
230
200 nm. Performance of the secondary treatment system for removing the small particles
(<200 and <25 nm) is not as high as that observed for total ceria.
Removal of ceria is dependent of the pH and organic content of the wastewater.
Circumneutral pH values enhance agglomeration of the particles, resulting in the
formation of unstable aggregates. Depending on the kind of organic content that is in the
wastewater, particles can be stabilized or destabilize. In the synthetic wastewater, peptone
seems to stabilize the particles, whereas the complex organic matter found in real
wastewater causes the aggregation of the particles. When particles agglomerate tend to
sediment. The detection of ceria in the sample of sludge taken from the aeration tank at
the end of the experiment suggests accumulation of the material over time. The presence
of ceria in the reactor did not cause microbial inhibition. The chemical oxygen demand
and acetate were continuously removed during treatment.
In a similar experiment aluminum oxide was added to real domestic wastewater
and its removal investigated during secondary treatment. Aluminum oxide is highly
removed during treatment. The pH and organic matter enhance the agglomeration and
sedimentation of the particles, as observed for ceria. Aluminum oxide causes partial
inhibition to microorganisms responsible for the removal of the organic content in the
wastewater. Adaptation occurs over time since efficiency gradually increased, despite
accumulation of the material in the reactor. Acetate degradation was not inhibited by the
presence of aluminum oxide.
231
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