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ANAEROBIC BIODEGRADATION OF POLYCYCLIC AROMATIC HYDROCARBONS USlNG FERRIC IRON AS TERMINAL ELECTRON ACCEPTOR by Kevin A. Robertson A thesis submitted to the Department of Chemical Engineering in conformity with the requirements for the degrce of Master of Science (Engineering) Queen's University Kingston, Ontario, Canada December, 1 998 copyright O Kevin A. Robertson, 1998
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ANAEROBIC BIODEGRADATION O F POLYCYCLIC AROMATIC HYDROCARBONS USlNG FERRIC IRON AS TERMINAL ELECTRON ACCEPTOR
by
Kevin A. Robertson
A thesis submitted to the Department of Chemical Engineering in conformity with the requirements for the degrce of
Master of Science (Engineering)
Queen's University
Kingston, Ontario, Canada
December, 1 998
copyright O Kevin A. Robertson, 1998
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Abst ract
Although aerobic biodegradation has been successfully demonstrated, a large niche
exists whereby intrinsic bioremediation via ~ e + ' redudion would provide a
significant economic advantage. This work examined a novel and promising in situ
remed iation approach to address the widespread contamination of pol ycycl ic
aromatic hydrocarbons (PAHs) using ferric iron as the terminal electron acceptor.
Two consortia enriched from contaminated soil/sediment, QU 1 and QU2, exhibited
characteristics consistent with PAH degradation coupled to ferric iron reduction. In
subsequent treatability studies, QU2 was shown to be capable of appreciable PAH
mineralization of the low molecular weight compounds, such as naphthalene and
phenanthrene. The poor solubility, and hence bioavailabi l ity, of the higher
molecular weight compounds was probably the source of their recalcitrance, and s o
surfactants and cyclodextrin were investigated as aids to enhance their dissolution.
At concentrations above their critical micelle concentration (CMC), Brij35 and
Trition x-100 were both toxic to the microorganisms. Cyclodextrin did improve the
rate and extent of phenanthrene mineralization slightly, but at higher cyclodextrin
concentrations it appeared to be consumed as the preferen tial carbon source,
thereby in hi biting phenanthrene metabolism. The low solubility of ferric oxides is
often ratecontroll ing and so EDTA was used to chelate them. Although EDTA
dramatical l y increased fenic oxide solubil ization, it only en hanced Fe'' reduction
of monoaromatic hydrocarbons, and inhibited PAH (napthalene, phenanthrene, and
anthracene) degradation. To detennine the effect of in situ conditions on
naphthalene mineralization, the influence of various environmental factors was
examined. The results showed that within a typical subsurface range of temperature
(lO°C - 30°C), pH (6 - 8), and nutrient profile (nutrient rich, nutrient poor),
mineral ization of PAHs could proceed. The capacity of the microorgan isms for
hydrocarbon degradation coupled to iron reduction was a fundion of the bacterial
population. When inoculated with QU 1, toluene mineralization occurred with a
concomitant accumulation of ferrous iron. The mineralization of low and, to a
lesser degree, high molecular weight PAHs catalyzed by QU2 was accompanied by
signs consistent with the presence of an iron biogeochernical cycle. A mechanism
was proposed whereby PAH rnetabolism occurred through a transfer of electrons to
ferric oxide, resulting in the production of soluble Fe? Diffusion of the ferrous
species into the overlying aqueous layer may have been followed by its re-oxidation
to ferric iron which precipitated ont0 the water-sediment interface. The mechanism
responsible for the regeneration is unclear, however several theories are presented.
Co-Authorship
The research and writing of this thesis was performed by Mr. Kevin Robertson.
However, during the course of this project, several undergraduate students
contributed work which went towards their senior theses. These students
conducted experiments that were designed and supervised by Mr. Robertson, but
their individual perspectives have provided a unique and valuable addition to this
work. Their assistance cannot go without being recognized.
Mr. Trevor Bugg examined the effects of synthetic surfactants on the bioremediation
of PAHs. Figure 6.1 contains information related to work that he performed for his
senior thesis in Engineering Chemistry.
Mr. Nicolas Acay investigated the potential role of cyclodextrin in the remediation
of PAH-contaminated sites. Figures 6.3, to 6.6 summarize some of this work, also
for his senior thesis in Engineering Chemistry.
Ms. Caroline Seto tested the selective ability of EDTA to solubilize iron forms of
varying crystallinity. This data is included as Figure 7.2, and was part of her work
for her senior thesis in Engineering Chemistry.
Acknowledgements
I owe many thanks to the numerous people who have offered their support along
the way. I am especially grateful to rny supervisor, Dr. Juliana Ramsay, from whom
1 have grown in the field of biorernediation, and research, as well as in life. For her
guidance, as well as her confidence in my abilities I am appreciative. Also, to Dr.
Bruce Ramsay, who offered both insight and his home, I am grateful.
I am truiy indebted to MarieCiaire Aly-Hassan at Ecole Polytechnique in Montreal
for her many houn of dedication towards this project. She has made a valuable
contribution and her efforts don't go unheeded.
It has also been my pleasure to work alongside Janani Swamy, Todd Adamsson, and
James Smith who offered encouragement and great friendships. 1 wish them the
very best in their future endeavours.
Departmental administration and Steve Hodgson found ways to help in almost any
situation, and kept everything running smoothl y.
Lastly to my friends, family, and new wife who extended great support under any
circumstance, I give than ks.
Table of Contents
Abstract ....................... .... .......,..... ..................
......................................................... Co-Authorship ................... ... .... ... ............................................. ............................. Acknowledgements ............
............................................................................................... List of Tables
........................................................... List of Figures .................................
......................................................................................... Nomenclature
Chapter 1 .
Chapter 2 .
2.9
Chapter 3 .
3.1
.................. .... ................. Introduction ..., ....
........................................ .................... titerature Review .,,,
........................................................ PAHs as Priority Pohtants ........................... ........................ Entry into the Environment ..
Fate .......................................................................................... ................................................ ................. Remediation ......
......................................................................... B ioremediation 2.5.1 Mass Transfer Limitations .......................................... 2.5.2 Chemical Structure Dependence ............................... 2.5.3 Substrate Mixtures ....................................................
.......................................... ...... ................... Biostimulation .,., .... 2.6.1 PAH Bioavailability Enhancement ................................ 2.6.2 Nutrient Supplementation .......................... .. .............. 2.6.3 Redox Conditions .................... ,,, ..............................
Anaerobic Microbial Degradation ........................ .. ................ 2.7.1 NitrateReduction ........... ............................................ 2.7.2 SulphateReduction .................................................... 2.7.3 Methanogenesis ............................. .. .........................
lron Reduction ......................................................................... 2.8.1 Ligand-Stimulated lron Reduction .............................. 2.8.2 Compounds Degraded ..............................................
............... Summary ...................................................................
..................................... Materiais and Methods .................... ..,.,
.......................................................................... Introduction
3.2 Ferric Oxide Synthesis .................... .... ............................. 3.3 lron Analysis .........................................................................
.......................................................................... 3.4 14C Analysis
Chapter 4 . lnoculum Development ...................................... ................
......... ........ 4.1 Introduction ............................................................. . ..................*........... 4.2 En richment Methods .................... ... ...
4.2.1 Sources of lnoculum ................................................. 4.2.2 lnoculum Development ............................................ 4.2.3 Isolation .................................................................. 4.2.4 lnoculum Evaluation .................................................
4.3 Results and Discussion .............................................................. .............................................................................. 4.4 Conclusions
Chapter 5 . Substrate Ut il ization ...................... ..... ...................................
...................................... 5.1 introduction .................................... ......................................................................... 5.2 Experimental
5.3 Results and Discussion .......................................................... 5.3.1 Naphthalene ............................................................ 5.3.2 Phenanthrene ........................................................... 5.3.3 Toluene ................................................................... 5.3.4 PAH Family of Compounds .......................................
.......................................................................... 5.4 Concl usions
Chapter 6 . Enhancement of PAH Bioavailability .........................................
6.1 Introduction .............................................................................. ..................................... 6.2 Synthetic Surfactants ................... ...
6.2.1 Experimental ............................................................... 6.2.2 Results and Discussion ............................. ....,,.. .........
...................................... 6.3 Cyclodextrin ................................... .. 6.3.1 Experimental ...............................................................
6.3.1.1 PAH Solubilization in an Aqueous Solution ... 6.3.1.2 PAH Desorption from a Sand Matrix .............. 6.3.1.3 Mobilization of PAHs through a Packed
Column ................... .... ......................... ........................... 6.3.1.4 Biomineralization of PAHs
6.3 -2 Results and Discussion .................................... ............ ... 6.3.2.1 PAH Solubilization in an Aqueous Solution
.............. 6.3.2.2 PAH Desorption from a Sand Matrix 6.3.2.3 Mobilization of PAHs through a Packed
Column ...................................................... ........................... 6.3.2.4 Biomineralization of PAHs
......................... 6.4 Conclusions ................................. ,..............
Chapter 7 . iron Chelation .......................................... ........................
............................................. 7.1 Introduction .... ...... 7.2 Experimental ..................... .. .................................................
............................. 7.2.1 Effect of EOTA on Fe Dissolution .,
............. 7.2-2 Effect of EDTA on Phenanthrene Degradation 7.3 Resultsand Discussion ..........................................................
............................... 7.3.1 Effect of EDTA on Fe Dissolution .................... 7.3.2 EDTA-En hanced Hydrocarbon Oxidation
............................................... ....................... 7.4 Concl usions ....
Chapter 8 .
Chapter 9 .
9.4
Chapter 10 .
Environmental Factors .............................................................
.......................................................................... introduction ......................................................................... Experi mental
.......................................................... Resu l ts and Discussion ............................................................ 8.3.1 Temperature
.......................................................................... 8.3.2 pH 8.3.3 Ferric Oxide Crystal l inity, lnocul um Percentage,
Nutrients ................................................................. .............................................. ........................ Concl usions ..
.......................................................................... Redox State
................................................. ....................... Introduction ,,
............................. Experi mental .. ........................................ Results and Discussion .......................................................... 9.3.1 Oxygen ................................................................... 9.3.2 Nitrate ..................................................................... 9.3.3 Sulphate .................................................................. 9.3.4 lron .........................................................................
9.3.4.1 QU1 Consortium ......................................... 9.3.4.2 QU2 Consortium .........................................
.......................................................................... lmpi ications
Conclusions and Recommendations ......................................
References ....................................................................................................
Curriculum Vitae .................................................... ................................
List of Tables
2 . i Physiochemical properties of selected PAHs relating to their bioavailability .................... .... .................................................... 12
2.2 Theoretical free energy change associated with the mineralization of ..................................... PAHs coupled to various electron acceptors 30
.............................. 4.1 Growth medium and trace mineral compositions 40
4.2 Qualitative evaluation of plated rnicrobial species ............................ 43
4.3 PAH concentrations in QU2 consortium sediment ............................ 45
5.1 Experimental conditions used for spiking radio-labeled cornpounds .... 47
5.2 Cornparison of mineralization rates for organic substrates studied ....... 55
List of Figures
Calibration curve for ferrous and ferric iron analyses .........................
Schematic of chernostat used in inoculum developrnent ....................
Mineralization of [14C]naphthalene ............................. ..... ..............
................................................ Mineralization of [14Cjphenanthrene
Minerakation of ['4qtoi uene .........................................................
................................ Toxic effect of Brij35 and Triton x-100 on QU1
Structural representation of Pcyclodextrin .......................................
Solubilization effect of HPCD on selected PAHs ...............................
Cyclodextrin-enhanced desorption of pyrene from sand ....................
Elution profiles for phenanthrene from a soi1 column treated with ................ water and a solution of MCD ................................... ...
Elution profiles for pyrene from a soi1 column treated with water and a solution of MCD .........................................................................
............ Influence of HPCD on the biomineralization of phenanthrene
Three-dimensional illustration of Fe(0i-i. )(EDTA) complex .................
.................... EDTA-enhanced dissolution of ferrihydrite and hematite
............................................ Ligand-stimulateci iron oxide reduction
......................................... Effed of EDTA on hydrocarbon oxidation
............................ Inhibition of naphthalene mineralization by EDTA
Temperature-dependence of naphthalene mineralization ...................
Effect of periodic pH adjustment on phenanthrene mineralization ......
8.3 Effect of periodic pH adjustment on the mineralization of selected h yd rocarbon s ....................... ... ............................................... 98
8.4 Effect of semicontinuous pH adjustment on the mineralization of naphthalene .................................................................................. 100
9.1 Efficiency of nitrate sparging for the removal of dissolved oxygen ....... 106
9.2 Observations and proposed mechanism for biogeochemical iron cycling .................... .... ............................................................ 112
Nomenclature
BTEXs
CD
CMC
CMCD
DIRB
DO
EDTA
HPCD
MCD
MGP
MSM
NAPL
NTA
PAHs
SCD
SRB
TMS
Benzene, Toluene, Ethylbenzene, Xylenes
Cyclodextrin
Critical Micelle Concentration
Carboxymethyl-Bcyclodextrin
Dissirnilatory Iron-Reducing Bacteria
Dissolved Oxygen
Ethylenediaminetetraacetic Acid
Hydroxypropyl-hclodextrin
PMethyl Cyclodextrin
Manufactured Gas Plant
Minerai Salts Medium
Non-Aqueous Phase Liquid
N itrilotriacetic Acid
Polycyclic Arornatic Hydrocarbons
Sulfateci-bclodextrin
Sulfate-Reducing Bacteria
Trace Mineral Solution
Chapterl Introduction
Polycycl ic aromatic hydrocarbons (PAHs) are ubiqu itous, carci nogen ic
pollutants which penist in the environment, and yet techniques for their
remediation are limited. The few methods which do exist are mediocre at best,
reflecting the technological difficulties associated with the removal of these
hydrophobic contaminants. Treatment options have been predominantly restricted
to incineration, land filling, and other ex situ techniques, but are inherently
expensive. In situ approaches offer potentially large cost savings, and less risk of
exposure to remediation personnel.
The intrinsic fate of PAHs is principally govemed by biological processes,
which explains the growing interest in bioremediation for their removal. The aim of
in situ biodegradation is to capitalize on nature's reflex to contamination, and use
techniques to enhance it. Aerobic degradation is commonly attempted but is
dependent, in most subsurface environments, upon the input of an external oxygen
source. Such injection techniques are complicated by factors such as: biomass
clogging, uncontrolled volatilization of selected compounds, and low permeable
matrices. The added cost also depreciates the value of this technique.
As an alternative, anaerobic PAH biodegradation offen a method to avoid
these obstacles through the use of NO,; Soi2, Mn+4, Fe", or CO, as electron
accepter. The limited research which has been conduaed in this area has
documented successful PAH degradation under bath nitrate- and sulphate-reducing
conditions. Of the potential electron accepton however, Fe+' is perhaps the most
promising, considering its natural abundance and favourable thermodynamics
compared to oxygen. However, dissimilatory iron reduction has only been l inked
to the degradation of straight chain, and rnonoarornatic hydrocarbons, although
PAHs are theoretically amenable. The demonstration of femc iron as the terminal
electron acceptor for the in situ redudion of PAHs, and a further understanding of
the mechanisms involved were the focus of this work.
The scope of this work has been Iimited to an examination of the proposed
scherne, including the identification of factors which contribute significantly to its
performance. This research was organized as follows:
A review of current literature was done to provide a basis from which the
experirnents were designed (Chapter 2)
Using enrichment techniques, an inoculum source was developed to be used in
subsequent treatabi lity experiments (Chapter 4)
Microcosms were prepared to determine the ability of the isolated cultures to
mi neral ize a range of aromatic hydrocarbons (Chapter 5)
Surfactants and cyclodextrins were used in an attempt to overcome mass-
transfer limitations typicaily encountered with PAH degradation (Chapter 6)
Organic ligands were investigated for their ability to accelerate the reduction o i
poorly soluble ferric oxides associated with the resulting increase in their
bioavailability (Chapter 7)
Experirnents were designed to evaluate the influence of various in situ
conditions on PAH mineralization (Chapter 8)
Evidence compiled throughout the preceding studies is discussed regarding the
ability of the developed consortiums to couple the degradation of organics to
the redudion of various eiectron acceptors (Chapter 9)
Chapter 2 Literature Review
2.1 PAHs as Priority Pollutants
Over the past few years, increasing awareness has been focused on
environmental pollutants and their effects on human and ecological health.
Statistio justify such concem. Environmental contaminants such as pesticides,
heavy metals, and organic solvents have been linked to reproductive and
developmental abnormal ities l ike reduced fertil ity, low birthweight, and congen i ta1
malformations (Moel ler, 1 992). Furthermore, increasingly contaminants are being
linked to the explosion in the number of cancer cases across the country.
Polycyclic aromatic hydrocarbons (PAHs) have recently received attention,
mainly due to their potential toxic, mutagenic, and carcinogenic effects (Dipple et
ai., 1 990). Based on neoplastic, genotoxic, and popu lation-level effects observed in
aquatic biota at sites contaminated with PAHs across Canada, Environment Canada
and Health Canada concluded that PAHs are entering the environment in a quantity
or concentration or under conditions that may have harmful effects on the
environment (Environment Canada and Health Canada, 1 994). Of the five higher
molecular weight PAHs they studied, including benzo[a] pyrene,
benzo[b]fl uoranthene, benzo[i]fluoranthene, benzo[k]fl uoranthene, and
indeno[î ,2,3-cd]pyrene, al1 were classified as "toxic" and 'probably carcinogenic to
humans"
For these reasons, the United States Environmental Protection Agency
(USEPA) has included 16 PAH compounds in their list of priority pollutants (U.S.
GPO, 1978). Strict regulatory cleanup standards have been irnplernented as well as
a requirement to monitor industrial effluents, and aquatic and terrestrial ecosystems
(Keith and Telliard, 1979). These efforts have undoubtedly helped to reduce the
input of PAHs into the environment, but the damage has already been done.
Decades of mismanaged environmental practices have left thousands of sites
scattered across the country contaminated with PAHs.
2.2 Entry into the Environment
PAHs are released into the environment through natural as wel l as
anthropogen ic sources. Major natural contri butors incl ude: forest fires, volcan ic
act ivity, d iagenesis, and biosynthesis. Of these, forest fires represent the largest
source of PAH emission to the environment (in Canada), averaging 2000
tonnedyear (Environment Canada and Health Canada, 1 994). The release from this
source, however, is often sepaated by large gaps in both time and geographical
location, and therefore does not represent a continuai supply of PAHs to the
environment. Although programs have been established and awareness brought to
the importance of forest fire prevention, quite often they are inevitable. In view of
this, concerted efforts are directed towards reducing PAH emissions from
anthropogen ic sources.
A wide variety of such origins exist, of which the prirnary ones are: by-
products from incornplete fossil fuel combustion; residues of coal processing; wood
preservation using creosote; residential wood heating; leakage of underground
storage tanks and pipelines; and spills at production wells, refineries and
distribution terrninals. Combustion of fossil fuels, burning of refuse, and coke ovens
contribute more than 50% of the PAH emissions across the US., with as rnuch as
another 35% coming from vehicle exhaust (Eng, 1985). The same study found that
an estimateci i 1 billion gallons of coal tar was produced in the United States
between 1 820 and 1 950 at coai gassification plants, a significant proportion of
which was discarded as waste. The nature of the sources themselves dictates higher
and more prevalent contamination in areas of denser population. Urban areas
exhibit 10-1 00 times higher PAH concentrations than those found in rural or less
populated areas (Harvey, 1997). In accordance with this is a much higher potential
for a human health epidemic. In many cases, drinking water supplies are in danger
of becoming contaminated if that is not already the case.
Consequently, PAH contamination has become a ubiquitous problem in air
(Daisey et al., 1979), soi1 (Bossert and Bartha, 1984), and aquatic (Andelman and
Snodgrass, 1 974) environments. In aquatic sediments, PAHs have become one of
the most common organic pollutants (Hites et al., 1980), the implications of which
are only now being discovered. For instance, sediments in the New YorWNew
jersey harbour have accumulateci tu the point where dredging is necessary to keep
the shipping channels open. The harbour is in danger of closing and losing huge
arnounts of revenue to other ports of entry. Dredging has already begun but, once
brought to the surface, the dredge spoiis are classifieci as hazardous and therefore
proper treatment/disposal must be considered.
Once PAHs enter the environment the problern becomes much more
corn pl icated and difficult to remedy, especial ly if the contam inants have becorne
'weathered" (Burford et al., 1993). Because of their hydrophobicity and strong
capacity to sorb ont0 soi1 organ ic matter (Karickhoff et al., 1 979; Subba-Rao and
Alexander, 1982), PAHs tend to accumulate, and bioconcentrate (Bulman et al.,
1988; South et al., 1983). Except for some of the lighter compounds, PAHs are
relatively non-volatile, contributing to their persistence within the environment.
Although, due to these characteristics, tar plumes do not exhibit significant
mobility, their dissolution is sufficient to contaminate nearby groundwater and
potential drinking supplies (Luthy et al., 1994).
2.4 Remediation
Characterization of PAH-contaminated areas has progressed, however only
few attempts have been made towards remediating these sites (Luthy et al., t 994).
Although few proven remediation technologies exist, a great deal of work has been
done to develop innovative remediation approaches (CRI, 1987; EPRI, 1990; EPRI,
1991). Common practice has been to extrapolate from technologies used for
remediating petroleum hydrocarbons and light-fraction organic contaminants to
restore sites polluted with PAHs (hg, 1985). In situ and ex situ technologies have
been studied, such as: biorernediation, chernical oxidation (Sims and Bass, 19841,
incineration, soi1 vapour extraction, soi1 leaching, solidification/stabilization (EPRI,
1 991 ), air sparging, low and high temperature thermal desorption (CRI, 1 988; CRI
1989; CRI, 1 990), soi1 washing (EPRI, 1992), and solvent/chemical extraction (Luthy
et al., 1992; Villaurne, t 991 ). Of these, biorernediation, soi1 vapour extraction, air
sparging, and low temperature thermal desorption are appl id most frequentl y in
Ontario (Sibul, 1 996).
However many unforeseen technical difficulties have surfaced due to large
discrepancies in the physiochemical properties of petroleum hydrocarbons and
PAHs. These techniques are generally effective for volatile andor water-solu ble
contaminants only, of which PAHs are generally neither (Rao and Loehr, 1 992).
Most attempts, therefore, are limited to removal of source materials and free
product, as wel l as l i miting off-site migration through groundwater pump and treat
strategies. It is not surprising that by far the most commonly employed approach for
restoring these sites has been to excavate and dispose the soils at licensed landfills
(Sibul, 1996). Along with its ease and timing of cleanup, and cost competitiveness,
excavation and landfilling elirninates the need for long term monitoring. In
addition, due to the Iack of effective technologies for the cleanup of hydrophobic,
non-volatile contaminants, excavation and landfilling is frequently the only
available option. Despite these advantages, this standard practice may be short-
lived. Landfill space is dwindling which inevitably means increased disposal costs.
The creation of new landfill sites, which was once regarded as a tax incentive for
participating cornmunities, is now encountering opposition due to a greater
awareness of the in herent human and ecological suscepti bil ity.
2.5 Bioremediation
Due to unprecedented levels of PAH contamination virtually evevhere
(Jones, 1988; Yland, 1986), the global estimate of the cost of restoration is
incalculable. The burden therefore lies upon the development of innovative
rernediation technologies. Although many have been proposed, their application is
usually cost-prohibitive. Over a seven year span, between 1982 and 1 989, costs of
rernediation have increased more than ten-fold (Abelson, 1992). This is due in part
to more stringent regulations, higher disposal fees, and the realization that standard
techniques are ineffective. Economics is therefore fundamental in determining a
technology's viability. Because PAHs are naturally removed predominantly through
biological mechanisms (Cerniglia, 1993; Park et al., 1990), technologies which
could predia, control, and accelerate this process would potentially offer many
advantages over conventional physiochemical techniques. As a result,
bioremediation research has grown as well as its implementation in the field
(Finnerty, 1 994).
However, the notion of indigenous PAH-degrading bacteria is not new.
Cores of mud have shown that these bacteria date back to at least 1 800, presumabl y
surviving on PAHs released frorn forest fires. Ever since then, in the rapidly
developing technological age, anthropogenic PAH contamination has dramatically
increased the global levels. Correspondingly the population of P~Hdegrading
bacteria has increased six times (Eng, 1985). The comprehension of this
evolutionary response was perhaps an event that inspired its research as a
potentially powerful tool for active site remediation. The acceptance of
bioremediation was strengthened following its successful application in the 1989
Exxon Valdez oil spill, where the bacterially mediated removal of hydrocarbons and
monoaromatic contaminants was rapid.
PAHs, however, were found to be much more of a complex problem than
straight chain and monoaromatic hydrocarbons. Similarly to physiochemical
approaches, traditional biological treatment strategies for PAH-contaminated sites
found Iimited success. Reflecting this, is the observed progression towards the
utilization of ex situ technologies such as: slurry reactors (ENSR, 1991; French
Lim ited Task Croup, 1 988), landfarming, and composting. But, excavation, capital
expenditures, and operating budgets, as uniquely required for ex situ techniques,
increase the overall cost immensely. An in situ technology that could reduce PAH
concentrations to below standard would offer many advantages.
Such demonstrations in the field have been limited, but a considerable
amount of experirnentation has been performed at lab scale (Hughes et al., 1997).
Unfortunately too often researchers have tried to apply labotatory results, where
systems are well controlled and more simplified (e.g. aqueous microcosms, pure
cultures, single substrates, ideal mixing), to predid field applications and seen large
discrepancies (Cemiglia, 1993). Field remediation demands consideration of matrix
effects, environ mental influences, contaminant characteristics, redox conditions,
and site geology. Consequently, each site that is contaminated with PAHs must be
treated on an individual basis. With further experience, inferences can be made.
At this point, however, feasibility studies fint must be done on a laboratory or pilot
scale for each new site. All attempts should be made to create conditions as close
to the field as possible. Increasingly, factors which influence degradation are
becoming known and studied so that the process may be optimized.
2.5.1 Mass Transfer Limitations
Bidegradation of PAHs is often modeled using concentrationdependent
rate equations, such as Monod or first-order rate expressions (Choshal et al., 1 996;
Larson, 1 980; Ramaswami et al., 1994; Scow et al., 1 986). An inherent
requirement of these relationships, however, is that the contaminant be available to
the degrading species. Physiochemical properties of the contaminant, Iike water
sol ubi l ity, and adsorptive capacity, and matrix characteristics, such as soi 1 type,
degree of contact (Burford et al., 1993; Hatzinger and Alexander, 1995), moisture
content, and percent hurnic material (Karickhoff, 1 980), determine the degree to
which the PAHs are bioavailable.
Mass transfer limitations such as these have been directly correlated with
observed rates of degradation. Volkering et ai. (1 992) found that faster growth rates
of a Pseudomonas sp. on naphihalene could be achieved by decreasing the size of
the naphthalene crystals. Bioavailability is one of the most influential factors
affecting the rate of PAH degradation in many contaminated environments
(Cern igl ia, 1 993).
2.5.2 Chemical Structure Dependence
A property that correlates well with the bioavailability of PAHs is their
chernical structure. Physiochemical properties relating to compound degradability
are listed for selected PAHs in Table 2.1. An increase in Log Kow represents a
Table 2.1 Physiochemical properties of selected PAHs relevant to bioavai labil ity.
Water Vapour
Compound Chemical Molecular L , ~ K ~ , Solubility Pressure @
Structure Weight @ 25°C 25°C (mg/L) (mPa)
Naphthalene
Acenaphthene
Phenanthrene
Anthracene
Pyrene
FI uoranthene
Benzo[apyrene
greater hydrophobicity. Similarly, vapour pressure is an indication of a compound's
degree of volatilization. in addition to water solubility, the net result of these
parameters provides insight into the extent of bioavailability. If PAHs are grouped
according to similarities in chernical structure, a trend in these properties becomes
apparent. This is a good explanation for the generat !y accepted rule that high
molecular weight compounds (4 and 5 rings) are more recalcitrant compared to
lower molecular weight PAHs (2 and 3 rings) (Hughes et al., 1997; Shuttleworth
and Cern igl ia, 1 995). An important parameter, therefore, for remediating PAH-
contaminated sites is the proportion of 4 and 5-ring to 2- and bring contaminants.
With the majority of sites contaminated with a complex mixture of PAHs, possibly
as well as monoaromatic and straight-chain hydrocarbons, remediation methods
rnust be capable of simultaneously degrading multiple substrates.
2.5.3 Substrate Mixtures
Multiple substrate degradation requires consideration of not just each
individual component, but the interactions among them. A mixed compound
systern may result in inhibition, cometabol isrn, augmentation, or no effect.
Laboratory studies using defined mixtures of PAHs (Bauer and Capone, 1988;
Bouchez et al., 1995; Stringfellow and Aitken, 1995; Tiehm and Fritzsche, 1 995),
and those which use contaminated sediments from field sites (French Limited Task
Croup, 1988; Lewis, 1993) have shown combinations of these effects. Of these,
inhibition has been the most common (Hughes et al., 1997).
The aqueous solubility of individual compounds, and common enzymatic
pathways have been proposed as possible causes for the observed inhibition
patterns in some PAH mixtures (Bouchez et al., 1995; Stringfellow and Aitken,
1995; Tiehm and Fritzsche, 1995). In several of these studies more soluble PAHs
inhibited the degradation of the less soluble ones. One of the most soluble PAHs,
naphthalene, is also among the most toxic compounds in the water-soluble fraction
of petroleum (Heitkamp et al., 1987). This must be taken into consideration when
interpreting results and planning remediation strategies.
Inoculation of mixed, as opposed to pure, cultures has shown the abil ity to
mitigate observed inhibition effects (Bouchez et al., 1995; Trzesicka-Mlynarz and
Ward, 1 995). Symbiotic relationships are ofien seen with consortiums, providing a
greater tolerance to toxic products in some cases. Results have been contradidory,
however, and the reason for this could lie with an understanding of the solubility
and partitioning of individual components.
Cometabolism has also been obsetved in PAH mixtures, and has been
defineci as 'the transformation of a non-growth substrate in the obligate presence of
a growth substrate or another transformable compound' (Dalton and Stirling, 1982).
Bouchez et al. (1 995) found that a Pseudomonas sp. was incapable of degrading
fluoranthene when supplied as the sole carbon source, but in the presence of
phenanthrene it was degraded cometabol ical l y. I t was proposed that cometabol ism
might be an important mechanism for larger, more recalcitrant PAHs when mixed
with more readily degradable, smaller ones. lnsights such as these are significant
towards stimulating biological activity, an area of research that needs further work.
2.6 Biostimulation
The first step towards understanding the needs of stimulating biological
processes is to identify the factors that are rate controlling. Several such parameters
that are recurrent in the liteature include bioavailability, and the presence of
nutrients and suitable electron accepton.
2.6.1 PAH Bioavailability Enhancement
The most common of al1 atternpts to increase the bioavailability of carbon
su bstrates has been the use of non-ionic surfactants (Shuttleworth and Cernigl ia,
1 995) and CO-solvents. Feasibil ity of co-solvent use for site remediation is l i mited
because significant sol ubil ization is not achieved until its vol urne fraction is greater
than ten percent. Surfactants have been proposed to enhance degradation rates by
decreasing capillary forces within the sediment rnatrix (Bury and Miller, 1 993) or by
apparently solubilizing the contaminants when present above their critical micelle
concentration (CMC) (Edwards et al., 1991). Release of PAHs sorbed to the soi1
matrix has been observed and attribut4 to the increased concentration gradient
established at the soil-water interface (Liu et al., 1 991), and increases in diffusivity
associated with swelling of the soi1 organic matrix (Yeom et al., 1 996). Surfactants
applied to the biodegradation of PAHs have resulted in confliding reports. In the
presence of non-ionic surfactants, some researchen have observed inhibition
(Aronstein et al., 1991; Laha and Luthy, 1991; Tiehm, 1994), whereas othen have
seen enhancement of PAH degradation at below (Aronstein and Alexander, 1993;
Aronstein et al., 1 991 ), or above their CMC (Bury and Miller, 1 993; Guerin and
Jones, 1988).
In addition to the questionable effectiveness, surfactan t-mediated
remediation strategies are compkated by further disadvantages. Cations present
within the contaminated formation may result in the precipitation or sorption of
many surfactants, thereby increasing the amount required uafvert and Heath, 1991;
Palmer and Fish, 1992). Also, some surfactants tend to form high-viscosity
emulsions that are difficult to remove (Palmer and Fish, 1 992). So, addition of a
synthetic chemical, where complete recoveiy is questionable, to an already
contaminated environment appears only to trade one contaminant for another.
The emergence of surfactant technology in the field of bioremediation has
spawned interest in the area of microbiallyderived surface-active agents; or
biosurfactants. The complex structure and resulting physiochemical properiies of
biosurfactants results in comparable or superior qualities to many of their synthetic
counterparts. Additionally, biosurfactants are less toxic (Lang and Wagner, 1 993),
offer potentially greater specificity, generally exhibit lower CMC values, and offer a
more environmental ly friendly alternative. Biodegradation of biosurfactants occurs
more rapidly and to completeness in contrast to the environmental persistence of
some synthetic surfactants (Finnerty, 1 994). They have been tested predom inantly
on crude oil biodegradation (Muller-Hurtig et al., 1993), and enhanced oil recovery
systems (Jack, 1 991 ; Finnerty, 1 992). Biosurfactants offer great potential for
accelerating the remediation of PAH-contaminated sites (Mueller et al., 1992). but
its economic feasibility remains to be deterrnined (Mulligan and Gibbs, 1 993).
A relatively new method of enhancing the bioavailability of organic
contaminants is through the use of cyclodextrin, a cyclic oligosaccharide produced
from the enzyrnatic degradation of starch by bacteria. Cyclodextrin has been shown
to have the ability to increase the apparent aqueous solubility of low-polarity
organic contaminants (Wang and Brusseau, 1993; Wang and Brusseau, 1995) by
forming water-sol uble inclusion complexes (Bender and Komiyama, 1 978). These
studies have dernonstrated that a 1 Ob solution of hydroxypropyl-hclodextrin
(HPCD) can increase the aqueous solubility of naphthalene and anthracene by 5.8
and 29.1 times respectively. Desorption and transport of contaminants in soi1 is an
additional property observed of cyclodextrins (Brusseau et al., 1994). Cyclodextrin
has the advantages of low reactivity with soils, resistance to poreexclusion
phenomena, insensitivity to pH and ionic strength effects, a non-synthetic origin,
and non-toxicity. HPCD also proved capable of removing aged contaminants just
as effectiveiy as from recentl y contaminated soils. Modifications to cyclodextrin-
mediated remediation techniques such as mixtures of different cyclodextrins and
addition of selected alcohols have both resulted in enhanced activity. A great deal
more work needs to be done to explore areas such as these so that in situ
remediation can be optirnized.
For any of these technologies to be feasible they must meet the following
criteria: eohance the bioavailability of a wide range of PAHs; be technically
possible at ful l-scale; uninhibitory; cost effective; and have a low environmental
impact.
2.6.2 Nutrient Supplementation
Microorganisrns have a basic metabolic requirement for some nutrients such
as nitrogen and phosphorus, among others. Rapid microbial activity following
contamination in a closed system may deplete nutrient sources to a point where
degradation stops, or slows considerably. Studies have investigated the effect of
nutrient addition in such systems and have found positive results (Churchill et al.,
1 995; Heitkamp and Cerniglia, 1 989; Manilal and Alexander, 1 991 ; Venosa et al.,
1992). Caution must be exercised when applying these results to other sites,
however. The nutrient profile will differ between sites, not to mention spatially
within the same site, as well as throughout the degradation process. Monitoring
nutrient levels and compensating where necessav is the best approach in each
case.
2.6.3 Redox Conditions
The microbial degradation of contaminants is accomplished through a
complex array of redox reactions. Through the transfer of electrons, contaminants
are mineralized, producing end products of carbon dioxide and water, or form
dead-end products. However, in order for the reaction to proceed, an appropriate
electron acceptor must be present in a suitable form to complete the cycle.
Microbial processes are aerobic if they incorporate oxygen as the terminal
electron acceptor, and anaerobic if other elements serve the role. Aerobic
microbial degradation dom inates, reflecting favourable thermodynamic conditions
and the abundance of oxygen in many environments. Nevertheless, at the onset of
contamination, redox conditions can quickly change in response to rapid oxygen
consumption by aerobic m icroorganisms (McFarland and Sims, 1 99 1 ; Mi helcic and
Luthy, 1 988b). This, combined with a slow rate of atmospheric recharge for
subsurface and groundwater geological formations, and the poor solubil ity of
oxygen, can create a ratelimiting obstacle for biorernediation. Ecosystems which
typically become anoxic include: soils with poor drainage, stagnant water,
municipal landfilis, sewage treatment digesters, industrial plants that produce
methane from organic waste, and sediments of the oceans and other natural water
bodies (Evans and Fuchs, 1988). The accessibility of suitable electron acceptors can
be equal ly important as substrate bioavailabil ity.
injection of oxygen into the subsurface is commonly done to stimulate
aerobic microbial metabolism. This approach lends itself to several technical
problems. One problern that is commonly encountered is the production of large
amounts of biomass (Holliger and Zehnder, 1 996). Consequently in situ treatment
systems often become clogged and therefore ineffective. High aeration rates
normally used in these operations could result in the volatil ization of a considerable
portion of some compounds, and in turn create an air pollution problem.
Additional ly, the presence of reduced alternative electron acceptors rnay
dramatically increase the required amount of oxygen by preferential oxidation of
these elements over the organic contaminants (Barcelona and Holrn, 1991 ).
Technical issues, such as matrix permeability and remoteness, may make in situ,
aerobic bidegradation infeasible or at the very least costly.
2.7 Anaerobic Microbial Degradation
Rather than atternpting to maintain subsurface aerobic conditions, alternative
electron acceptors either present naturally or injected could be u s 4 in an anaerobic
rernediation scheme. The practicality of this approach is dependent upon the
achievable rate and extent of degradation. In observance of the fact that oxygen is
used both as a terminal electron accepter, and as a reactant in the ring-cleavage
process, anaerobic biodegradation of aromatics was met with much skepticism.
This perception changed, however, after a 1934 report was published clairning that
benzene could be microbially degraded in the absence of molecular oxygen (Tarvin
and Buswell, 1 934). lnterest into anaerobic biodegradation of aromatic compounds
has been growing ever since. However, compared to aerobic metabolism, very
little work has been done to elucidate the rnechanisms. The majority of work has
focussed on nitrate, sulphate, and CO, reduction.
2.7.1 Nitrate Reduction
As long ago as the early eighties, researchen studied the feasibility of
pumping a nitrate solution into the subsurface to treat a site contaminated with
hydrocarbons (Holliger and Zehnder, 1996). Since then nitrate has been used to
successfully degrade aromatics such as benzene (Major et al., 1988), toluene (Major
et al., 1988; Zeyer et al., 1986; Kuhn et al., 1988; Frazer et al., 1993; Flyvbjerg et
al., 1993), xylene (Major et al., 1988; Kuhn et al., 1 988; Kuhn et al., i 985;
Hutchins, 1991 ; Hutchins et al., 1 Wl), ethylbenzene (Hutchins, 1991 ; Hutchins et
al., 1 991 ; Bail et al., 1 991 ), benzoic acid (Taylor et al., 1 970; Dolfing et al., 1 990),
phenol (Flyvbjerg et al., 1993; Bakker, 1977), and cresol (Flyvbjerg et al., 1 993;
Rudolphi et al., 1 991).
PAH degradation under denitrifying conditions has not been thoroughly
examined. Bouwer and McCarty (1 983) investigated naphthatene degradation in
primary sewage effluent under denitrifying conditions, but discovered no significant
activity over an 1 1-week period. This further strengthened the belief that un-
substitut4 PAH compounds were microbially recalcitrant in the absence of
molecular oxygen (Cernigl ia, 1 992). However a few years iater Al-Bas hi r et al.
(1 990) were able to achieve 9046 mineralization of naphthalene over a 50-day
incubation period under denitrifying conditions. After a lag period of approximately
1 8 days, rnineralization proceeded alrnost linearly followed by a gradua1
logarithmic decrease. When naphthalene was supplied at a concentration of 50
ppm, the rate of rnineralization was 60% less than when it was added at 200, or
500 ppm. They attribut4 this to different rate-controlling factors in each case.
Because naphthalene at 50 ppm is close to its solubility a smaller amount was
available in the aqueous phase due to sorption ont0 solid material. When present
at 200 or 500 ppm, naphthalene reached saturation in the water phase, and
therefore its bioavailability no longer acted as the initial rate-controlling step. After
the first 50 ppm were mineralized, the subsequent rate decreased indicating that the
rate of degradation was quicker than the rate of desorption. At this point the rate-
controlling factor became substrate availabil ity. The maximum degradation rate
achieved over the entire incubation period was 1.8 ppdday, the same magnitude
as those achieved under aerobic conditions (Bauer and Capone, 1988).
Mi helcic and Luthy (1 988a) also studied PAH degradation under den itrifying
conditions and had positive results. They were able to degrade naphthalene,
initially supplied at 7 ppm, to nondetectable levels over a 45day period with an
observed 1 M a y accl imation period. They were also successful in degrading
acenaphthene from 0.4 ppm to nondetectable levels in 40 days with a lag period of
15 to 20 days.
2.7.2 Sulphate Reduction
Coastal marine sediments typicall y con tain molecular oxygen and nitrate
only in a thin surficial layer, which in most cases would be capable of degrading
only a small fraction of the total organic matter present (Canfield et al., 1993). If
bidegradation could be achieved only through the use of molecular oxygen or
nitrate as terminal electron acceptors, this would have sign ificant ramifications for
contaminant fate and remediation options. As discussed, in situ remediation could
be attempted by continuous injections of either oxygen or nitrate, but would
in herentl y increase the cost of treatment. Natu rai attenuation offers the most cost-
effective approach to remediation, but without a sufficient concentration of electron
acceptor, inadequate amounts of poll utant would be removed. If, however, another
potential terminal electron acceptor could be identified which was in natural
abundance and was capable of completing the redox reaaion, intrinsic remediation
might be a promising course of action. Attention was directed towards sulphate to
fiIl this role.
Sulphate reduction combined with the oxidation of aromatics has been
studied less extensive1 y than for nitrate. Under su1 phate-reducing conditions
complete mineralkation of benzene has been shown (Lovley et al., 1995; Edwards
and Grbic-Galic, 1992). In addition, toluene (Flyvbjerg et al., 1993; Haag et al.,
1991; Edwards et al., 1992; Rabus et al., 1993), xylene (Flyvbjerg et al., 1993; Haag
et al., 1 991 ; Edwards et al., 1992), phenol (Flyvbjerg et al., 1 993; Bak and Widdel,
1986; Gibson and Suflita, 1986), and cresol (Flyvbjerg et al., 1993; Bak and
Widdel, 1986; Suflita et al., 1989) have been shown to be degraded under sulphat*
reducing conditions.
PAHs were long thought to be recalcitrant when present within sulphate-
reducing conditions. This belief was even held by research groups which had
demonstrated the successful degradation of PAHs under other anaerobic conditions
(Al-Bashir et al., 1990; Leduc et al., 1992; Mihelcic and Luthy, 1988a; Mihelcic and
Luthy, 1988b). A recent study proved othenvise. Coates et al. (1 996a) were able to
mineralize both phenanthrene and naphthalene under strictly anaerobic conditions.
Sulphate was inferred as the terminal electron acceptor due to several factors: 1)
strict anaerobic conditions were rnaintained in a reducing environment, such that
even trace amounts of O, would be rapidly consumed; 2) results of ion
chromatography revealed that nitrate was present below detectable levels (< 200
nM) and that sulphate was available at a sufficient concentration (1 0 mM) to a d as
the terminal electron acceptor in the oxidation of the spiked PAHs; and 3 ) when 20
mM molybdate, a specific inhibitor of sulphatduction, was added to a sediment
slurry, rnineralization of phenanthrene stopped irnrnediately.
Sediments from two distinct sites were used in this study for inoculation.
One sediment was taken from a site which had been heavily contaminated with
PAHs (33 mgkg) over an extended period, while the other sediment contained
PAHs at levels much lower (4 m&$ Mineralization of phenanthrene and
naphthalene was immediate and rapid in the microcosm with the highly
contaminated sediment, whereas a lag period followed by a slow rate of
mineralization was observed using the other sediment. It was proposeci that long-
term exposure to PAHs under sulphatereducing conditions may be necessary to
establish a significant PAHdegrading population, and that this could be the source
of failure in previous experimental studies.
2.7.3 Methanogenesis
Detailed information regarding aromatic hydrocarbon degradation using CO,
as the terminal electron acceptor is not as available as for either nitrate or sulphate.
Nevertheless, oxidation of benzene (Vogel and GrbicGalic, 1986), and toluene
(Edwards and GrbicGalic, 1994; Vogel and GrbicGalic, 1986; Grbic-Calic and
Vogel, 1987), has been achieved. Although available electron acceptors for
methanogenesis is usually not a problem, the reaaion is only slightly exergonic and
degradation, therefore, can be slow. More thermodynamically favourable electron
accepton usually intervene. PAH degradation using CO, as the terminal electron
acceptot has not knowingly been accomplished to date.
2.8 lron Reduction
Ferric iron reduction is therrnodynamically comparable to aerobic
degradation, and of all the theoretically possible electron acceptors, it is the single
most abundant, with greater than 90% of the total oxidative capacity in many cases
(Fredrickson and Gorby, 1 996; Lovley et al., 1 994; Ponnamperiuma, 1 972). In
many North American glacial lakes, Fe can make up several percentage points of
the dry weight in sediments (Neaison and Myers, 1992). This offen considerable
potential for intrinsidin situ anaerobic bioremediation with substantially less
operational expenses than alternative techniques.
Many species have been identifid which have shown to be capable of
dissi mi latory iron reduction including: Geobacter metallireducens (Lovley et al.,
1 993), Shewanella putrefaciens (Myers and Nealson, 1 988) Th iobacillus
thiooxidans (Brock and Gustafson, 1 976; Kino and Usami, 1 982), Thiobacillus
ferrooxidans (Sugio et al., 1 985), Bacillus circulans flroshanov, 1 968; Troshanov,
1969), Bacillus polymyxa (Munch and Ottow, 1982; Munch and Ottow, 1983),
Clostridîum butyricum (Munch and Ottow, 1983), Vibrio sp. Uones et al., 1983),
Sulfolobus acidocaldarius (Brock and Gustafson, 1 976; Kino and Usami, 1 982), and
Pseudomonas sp. (Obuekwe et al., i 981; Obuekwe and Westlake, 1982a;
Obuekwe and Westlake, 1 982b), illustrating the broad phylogenetic divenity of
dissirnilatory iron-reducing bacteria (DIRB).
lntrinsic remediation of organic contaminants coupled to ferric iron
reduction has been observed (Lovley et al., 1989a; Lyngkilde and Christensen,
1992). Rates of degradation, however, have been strongly correlateci with the
enzymatic redudion of ferric rninerals (Fredrickson and Gorby, 1996). Those
minerals that are poorly crystal l ine, such as ferri hydrite, and hence offer greater
surface area, are reduced more quickly than rninerals that are highly ordered, such
as hematite (Munch and Ottow, 1 982; DeCastro and Ehrlich, 1 970). When Fe" is
supplied at a high enough concentration, present as the metastable intermediate
ferrihydrite, dissimilatory redudion was no longer dependent upon the kinetics of
iron redudion but rather was governeci by microbial physiology (Arnold et al.,
1988). When the iron was supplied as hematite, reductive dissolution was
contingent upon the reactive surface area of the mineral (Arnold et al., 1988).
Early experiments even suggested that DIRB are capable of reducing only
amorphous and poorly crystalline iron oxides in natural sediments and that highl y
ordered rninerals such as hematite and goethite are, in cornparison, recalcitrant
(Phillips et al., 1 993). This poses a problern since the majority of iron in oxic
terrestrial subsurface sedirnents exists as ferric hydr(oxide) minerals (Fredrickson
and Corby, 1996) and consists of an array of mineralogies. Therefore the ratio of
amorphous to crystal 1 ine m inerals would potentiai ly determ ine whether the site
would be arnenable to rernediation via in situ, dissirnilatory iron reduction.
More recently, however, with increasing interest k i n g paid to DlRB for
rerned iation purposes, researches have isolated cultures which have been able to
reduce even highly crystalline minerals such as hematite, goethite, and even the
mixed valence iron oxide mineral magnetite (Arnold et al., 1988; Roden and
Zachara, 1996; Kostka and Nealson, 1 995). A possible explanation for the
conflicting evidence has been postulated. A5 with sulphate, long-terrn exposure to
ironieducing conditions can potentially select for cultures that have the ability to
reduce more crystalline minerals (Heron and Christensen, 1995). Whereas DlRB in
newly contaminated sites may only exhibit the abil ity to d u c e poorly crystalline
forms of iron oxide (Lovley and Phil lips, 1 986b).
Still, even with the demonstrated ability of some DlRB to reduce a broad
range of iron oxide minerals (Ehrlich, 1981), the rates of reaction are strongly
dependent upon mineral stability. With an approximate solubility of IO*" M
(Stumm and Morgan, 1 981), the limited bioavailability of iron poses the question of
its potentiai as a terminai electron acceptor.
2.8.1 Ligand-Stimulated lron Reduaion
Lovley et al. (1 994) demonstrated a possible means of overcorning this
problem. They found that the addition of chelators, organic ligands that form strong
complexes with metal species, dramatically enhanced the rate of reductive
dissolution. They attributed this to either increased bioavailability of the Fe"
andor the complexation/solubilization of adsorbed ~ e ' ~ . Other research groups
have observeci this phenornenon as well (Arnold et al., 1988). Degradation of
toluene and benzene, with added nitrilotriacetic acid (NTA), by DlRB was
accelerated to the point where it was comparable to tates in oxic sediments, and
fifty times faster than under denitrifying conditions (Lovley et al., 1994). In this
same study, ethylenediaminetetraacetic acid (EDTA), another chelator, has been
shown to stimulate degradation at least as well as NTA, but each reacts differently
with various mineals. Another group found that citrate, a tridentate ligand, resulted
in a threefold increase in the rate of reductive dissolution (Jones et al., 1983).
Use of such techniques in the field is unlikely as large amounts of synthetic
chelator would be added to the subsurface, further contaminating the site.
However studying synthetic chelators provides useful information for predicting the
possible effects of m icrobial l ysynthesized chelators (siderophores). Some species
have the ability to produce siderophores when under iron-lirniting conditions
(Haselwandter, 1 995; Drechsel et al., i 995). The siderophores a a simi larl y to
synthetic chelators in that they form strong associations with iron compounds and as
a result enhance their accessibility to microorganisrns for reduction. At the same
ti me, because they are produced natural l y, siderophores are environ mental l y
friendly and an acceptable form of treatment.
2.8.2 Compounds Degraded
DlRB have been found capable of degrading many different straight chain
and aromatic hydrocarbons, including benzene, benzoate, benzylalcohol,
benzaldehyde, phydroxybenzaldehyde, phydroxybenzoate, benzoic acid, toi uene,
phydroxybenzylalcohol, pcresol, ethanol, phenol, lactate, acetate, yeast extract,
butyrate, and propionate (Lovley et al., 1989a; Lovley and Lonergan, 1990; Lovley
and Phillips, l986b; Lovley and Phillips, 1988; Semple and Westlake, 1987).
Complete oxidation of some compounds to carbon dioxide has been observed
(Loviey et al., 1989a).
No dissimi latory iron-reducing culture or consortium has been identified
which has shown the ability to degrade PAHs. The only known reference towards
such work was done by Coates et al. (1 996b), who were unable to show
degradation of phenanthrene when tested against five different strains. The
complexities of each redox component combine to present a unique and
challenging remediation process. However the successful oxidation of PAHs under
nitrate- and sulphate-reducing conditions, and the demonstrated reduction of ferric
oxides coupled to monoaromatic hydrocarbon degradation is encouraging.
These convictions are further strengthened through thermodynamic calculations,
confirming the theoretical feasibil ity of this reaction fiable 2.2). A negative free
energy change associated with each of the electron acceptors indicates a potentially
favourable redox reaction. The magnitude of the free energy change is a measure of
the energy released, and must be sufficient to satisfy microbial maintenance and
growth requirements. Sirnilar values for 0, and Fe') suggest comparable
energetics.
A hierarchy of electron acceptor use exists, whereby the sequen tial reduction
of electron accepton woufd follow ~ n + ' , 4, NO, Fe'I, Mn4, FeOOH, SO,'*,
and CO,, as predicted by thermodynamics. The importance of increasing the
Table 2.2 Free energy change associated with naphthalene and pyrene mineralization using various electron acceptors. (Source: McFarland and Sims, 1 99 1 1
AG0 (25OC, pH 7.0)
Electron Acceptor Naphthalene Anthracene Phenanthrene Pyrene
Non-Metai
~ n + ~ Fe+3
M n 4 FeOOH
soluble fraction of iron is once again justified. The intrinsic generation of
successive redox zones mirroring this profile has been documented (Achtnich et al.,
1995; Lyngkilde and Christensen, 1992; Nealson and Myers, 1992). The
development of conditions suitable for iron reduction results in a redox zone that
can be a significant means of organic carbon oxidation. This has considerable
implications for the natural attenuation of PAHs in the subsurface, an area of
research that has not been fuliy explored. This thesis discusses the investigations
that were done to determine the feasibility of iron redudion for the degradation of
PAHs, and the factors that influence this process.
2.9 Summary
A review of the literature has revealed the following points:
The health effects and ubiquity of PAHs make them worthy contaminants to
further study
PAHs persist in the environment and there is a lack of adequate technologies for
their remediation
Bioremediation offers a cost-effective alternative, but factors such as
bioavailabil ity, chemical structure, nutrient profile, substrate mixtures, and the
redox environment must be considered
Enhancement of PAH bioavailability has been accomplished through
amendments with synthetic and biological surfactants, and cyclodextrins
Poorly permeable subsurface ecosystems becorne anaerobic soon afier
contamination
Maintenance of aerobic conditions through injection of oxygenating compounds
is common, but often leads tu technical difficulties
Anaerobic m icroorganisms can couple the oxidation of man y cornpounds,
including PAHs, to alternative electron acceptors such as nitrate and su1 phate
Ferric iron is a potential eiedron acceptor due to its abundance in nature and
favourable therrnodynamics
Ferric oxide insolubility and crystallinity is a significant factor in hindering its
use as a terminal electron acceptor
Acceieration of iron reduction has been achieved by the addition of organic ,
ligands which solubilize the oxide minerals
Although monoarornatic hyd rocarbons have been degraded when coupled io
iron reduction, this has not been demonstrated in the literature for PAH
degradation.
Chapter 3 Materials and Methods
3.1 Introduction
The work presented in this thesis is structured as a compilation of several
chapters. The materials and methods unique to each are described therein.
However, some techniques were commonly employed, the details of which are
included in this chapter.
3.2 Ferric Oxide Synthesis
Two forms of ferric oxides were used: ferrihydrite (5Fq0,-9H,O), and
hernatite (a-Fe,O,). Fenihydrite was synthesized using a modified formula given by
Lovley and Phillips (1 986a). A solution of FeCI, in distilled water was neutralized
with NaOH and diluted to provide the desired final concentration of Fe(lli). No
attempts were made to remove the chloride, upon determining that it was not
inhibitory to the inoculums at the concentration provided. The resulting colloidal
suspension was continuously mixed during its transfer to microcosms to ensure an
equal distribution of Fe(lll). To form hematite, ferrihydrite was produced as
described and filtered to recover the ferric oxide precipitate, which was
subsequently dried in an oven at 80°C ovemight then ground into fine particles
using a mortar and pestle (Fisher Scientific). The procedure used to synthesize
hematite was adapted from Lovley and Phil l ips (1 986a).
lron Analysis
The phenanthroline colourimetric procedure (American Water Works
Association, 1992) was used to quantify both ferric and ferrous iron in aqueous and
dualphase (sediment-water) samples. For analyses performed on aqueous systerns,
a i .5 rnL sample was centrifuged in a microcentrifuge (louan MR14.11) for 2
minutes at 9,000 rpm to remove any suspended solids. The supernatant was used
in the subsequent colourirnetric reaction.
lron was quantified in slurry microcosms by fint obtaining a homogeneous
sample of approxirnately 0.5 g using a syringe with a bevelled needle through the
reactors' septum. The actual sample weight was rneasured in tared, 7-rnL
scintillation vials (Fisher Scientific) on an analytical balance, accurate to 0.1 mg. 5
mL of a 5 M HCI solution was added to each via! to extract the majority of Fe
fractions from the sediment over a period of 21 days (Heron et al., 1994). The
digest4 contents were centrifuged for 5 minutes at 9,000 rpm, and the supernatant
used for analysis. Due to the propensity for iron to adsorb, al l vials were treated
with 1 N HCI prior to their use to minimire false high readings.
The reagents were prepared according to the specifications given in Standard
Methods (American Water Works Association, 1992). A calculated volume of
sample was used to provide no more than 200 pg of Fe for total iron analysis. To
this sample, the following were added: 1 O mL ammonium acetate buffer solution
(250 g NH,C,H,O, in 150 mL distilled water and 700 mL glacial acetic acid), 4 mL
phenanthroline solution (0.1 96 solution of 1 ,l O-phenanthroline monohydrate in
distilled water), 1 mL hydroxylarnine solution (1 0% NH,OH-HCI in distilled water),
and 2 mL concentrated HCI. The contents were brought to a final v ~ h ~ m e of 50
with distil led water, mixed thoroughly, and al lowed tu react for 1 5 minutes for
maximum colour development. The absorbance was rneasured
spectrophotometrically at 5 10 nm (Philips PU8720), and compared against a
calibration curve (Figure 3.1). The instability of ferrous iron in an oxidizing
environrnent requires that special attention be paid towards its analysis. Although
an acidic environrnent helps to stabil ize the ferric-ferrous ratio, samples should be
analyzed immediately fol lowing collection to min imize the effects.
Determination of ferrous iron was accomplished by mixing a volume of
sample, not exceeding 50 pg of total Fe, with i O mL ammonium acetate buffer
solution, 20 mL phenanthroline solution, 2 mL concentrated HCI, and distilled
water to a final volume of 50 rnL. Maximum colour development was reached after
5 minutes of incubation time. Exposure of samples to light will result in the
photoredudion of ferric iron and hence overestimate the ~ e + ~ concentration. To
minimize these effects, amber bottles or flasks wrapped in aluminum foi1 were used
during experiments, and samples were stored during extraction/cornplexation in the
absence of light. Similarly to total iron analysis, the absorbance was measured at
51 0 nm and compared against a calibration curve to give the ferrous iron
concentration. Ferric iron was determined as the difference between total and
ferrous iron.
Figure 3.1 lron calibration curve for ferrous and total iron analyses.
3.4 l4C Analysis
Many mineralization experiments were used to determine the ability of
different inoculums to degrade hydrocarbon contaminants. Although the conditions
varied between microcosms, the procedures used for 14C anal ysis rernained the
same. 125-mL serum botties were used as microcosms in al1 mineralization studies.
Within each bottle was a test tube containing 2 mL KOH solution to trap the 14C0,
evolved. The carbon substrates were added to the microcosms predominantly as
non-radiolabeled cornpounds and their associated radioisotope at approximately
100,000 DPM. In sarnpling, the entire volume of KOH solution was replaced with
a fresh, equal volume. The recovered KOH was quenched in a scintillation cocktail
(Optiphase 'Hi Safe" 3, Fisher Scientific, Loughborough Leics, England) in 20-mL
scintillation vials (Fisher Scientific) and measured for I4C on a Wallac 1409
scintillation counter. The results were given in units of DPM and converted to
percent mineralization.
Chapter 4 lnoculum Development
4.1 Introduction
Pure cultures of PAHdegrading (Bouchez et al., 1995; Heitkamp and
Cemiglia, 1 988; Heitkarnp and Cemiglia, 1 989; Stringfellow and Aitken, 1 995;
Tiehm and Fritzsche, 1995; Walter et al., 1991; Ye et al., 1996), and dissimilatory
iron-reducing (Coates et al., 1995; Lovley et al., 1993; Myen and Nealson, 1 988;
Roden and Lovley, 1993) bacteria have been obtained. However, no pure or mixed
culture has been shown to couple the oxidation of PAHs with ferric iron reduction.
The objective of this section was to enrich for an inoculum which could accomplish
this goal, and subsequently be used in future treatability studies. Several
enrichment steps were utilized for this purpose.
The development of a pure culture was not necessary for this work, and
therefore not the focus of this Chapter. Moreover, a consortium can have many
advantages over single strains. Such a mixed microbial population would provide a
system more representative of a natural ecosystem. Laboratory results could then
more easily be extrapolated to model the expected behaviour in the field. It has
also been indicated that in natural systems manganese- and iron-reducen may
function optimally in mixed populations (Jones et al., 1983). These researchers
found that more Fe could be solubilized by a mixed culture than from the pure
cultures which were derived from it.
4.2 En richment Methods
4.2.1 Sources of lnoculurn
Several soi1 and sediment samples were used to initially incubate the batch
and chemostat enrichment experiments. These were taken from the anaerobic zone
of environments contaminated with toxic organics. This approach is a tool which
allows for the collection of a large number and wide variety of microorganisms
which possess the desired characteristics (Hunter-Cevera et al., 1 986). Samples
were stored at 4OC to sustain the cultures' viability, and without headspace to
maintain anaerobicity.
4.2.2 lnoculum Development
lnoculum development was performed using batch and continuous
fermentation techniques. Shake flask enrichments were done by combining 50 mg
yeast extract, 0.5 g contaminated soil, and 100 mL of mineral salts medium (Table
4.1 ) in 125-ml glas, amber, boston round bottles closed with a screw cap top. The
growth medium also contained 1.4 g/L Fe as ferrihydrite which served as the major
source of terminal electron accepter, and 20 mgR naphthalene, 50 mg/L toluene, or
50 mg/L acetone which senred as a source of carbon and energy. The medium and
headspace were purged for 15 minutes with high purity nitrogen to displace oxygen
prior to being sealed. The slurries were mixed on a gyrotory shaker (New
Brunswick) at 180 rpm and incubated at room temperature (2S°C f 3).
Table 4.1 Constituents in the A: mineral salts medium; B: Trace Mineral Solution (This) (Modified from Lovley and Phillips, 1 986a).
lngredient Conc. (g/L) lngredient Conc. (g/L) - - -
NaHCO, 2.5 CaCI,-2H20 O. 1 KCI 0.1 N H4CI 1.5 NaH, PO, 0.6 NaCl O. 1
NaMo0,-2H20 0.037 N ICI2-6H20 0.024 EDTA 1 .S MgS04*7H20 3.0 MnS0,-4H,O 0.6 NaCl 1 .O
MgCl,-6H,O O. 1 FeS0,-7H20 0.1 MgS04-7H20 0.1 CoS0,-7H20 0.1 8 MnS04-H20 0.0043 CaC1,-2H20 0.1 N aMo0,-2H,O 0.00 1 ZnSO, 0.1 TMS 10 mL CUSO,.~ H 2 0 0.0 1
AIK(SOJ2-1 2H20 0.01 8
W O , 0.01
A multistaging procedure was adopted to ensure that substrates and nutrients
were not limiting, and to maintain the culture in physiological state of growth. This
involved the transfer, every 3-4 weeks, of a 1096 (vol/vol) inoculum to fresh growth
medium. After several serial dilutions the inoculum became essentially free of any
solid material.
A chemostat (Figure 4.1 ) was also designed to enrich for a PAHdegrading
population using Fe'3 as the terminal electron acceptor. The same mineral salts
medium with 0.7 g/L Fe as ferrihydrite, chelated with 3.6 g/L
ethylenediaminetetraacetic acid (EDTA) was added to a 7-L feed resewoir. To the
chemostat the following were added: 50 g of contaminated soil, and naphthalene
Figure 4.1 Chemostat used for inoculum development. A: highpurity N, gas cylinder; B: medium reservoir; C: sterile air filter; D: magnetic stirrer; E: magnetic stirring bar; F: peristaltic pump; C: bioreactor vessel; H: level control outlet fine; 1: waste reservoir.
supplied at approxirnately 30% above its solubility. The volume in the reactor was
brought to 400 rnL with medium and mixed in batch mode for two days to allow a
culture to become established before operation as a chernostat. Throughout the
incubation, the contents were continously sparged with a low strearn of purified N,
at room temperature (25OC f 3).
When the system was changed to continuous operation, a feed flow rate of
approximately 40 mVhr was used, producing a dilution rate of O. 1 hi'.
Periodically, naphthalene crystals were added to rnaintain its dissolved
concentration at its solubility. The reactor was covered with aluminurn foi1 to
prevent naphthalene photodegradation, and light-induced iron redudion.
The development of an inoculurn source from sediment collected at the site
of a former gassification plant in Kingston was done by creating a dual-phase slurry
in a four-liter tissue culture roller bottle. The contents were brought to a final
volume of 3.5 L, and consisted of the defined minerai salts medium, 1.0 g/L Fe as
ferrihydrite, 20 ppm each of naphthalene, phenanthrene, and anthracene, and
sediment at a final concentration of 20% (wt/vol). The sluny was purged with high-
purity N, prior to k i n g capped and continuously mixed on a Modular Cell
Production Roi ler Apparatus (Wheaton Instruments) at 1 00°' motor speed. The
growth medium, reactors, or bottles used in inoculum development were not
sterilized prior to use.
4.2.3 Isolation
A first attempt at isolating bacterial colonies from the liquid growth medium
was accomplished using agar plates. The agar medium consisted of mineral salts
medium with 1.5 g/L Fe as ferrihydrite, 2.0 g/L EDTA, 20 ppm naphthalene, and 15
g/L agar. After autoclaving and allowed to cool briefly, the solution was dispensed
into petri dishes and cooled until sol idified. The agar plates were streaked with a
loopful of liquid culture using a serial dilution technique to obtain well-isolated
colonies. Plates were incubated upside down, at 30°C, in an anaerobic jar using a
GasPak system (BB L Microbiology Systems).
4.2.4 lnoculum Evaluation
The streak plates were observed for microbial growth over several weeks of
incubation. Table 4.2 provides a qualitative comparison for the plated strains, and
includes the substrate(s) that were initially used to enrich them in Iiquid culture.
Each culture in Table 4.2 was derived from a petroleurn-contaminated soi1 in Sarnia,
Ontario, and was evaluated for growth afier approximately one month of
incubation.
Table 4.2 Evaluation of anaerobic plated cultures grown on 20 ppm naphthalene initially isolated on: A - Acetone; N - Naphthalene; T - Toluene. Note: N N would denote isolation on acetone followed by naphthalene, prior to plating. A negative (-) growth indicates no observable growth. Positive growth given on a relative sale from ( + ) to (+ + + +) which indicates poor to ven/ good observable growth, respectively.
Su bstrate lsolate lsolated on Growth (+/O)
4.3 Results and Discussion
The enrichment and ensuing isolation of cultures from contaminated soi1
samples has lead to the development of four inocula which had good growth on
agar plates containing naphthalene. These were: K-2, K-6, K-25, and K-26. These
strains were subjected to a quantitative screening (discussed in Chapter 6) in which
K-25 (later renamed QU 1 consortium) was selected for the remaining experiments.
QU1 consortium was originally developed from a soi1 sample collected from
a petroleum-contaminated site in Sarnia, Ontario. The sample was taken a few feet
below the surface in an attempt to obtain a high population of anaerobic
microorganisms. The inoculum grew on rnonoaromatic hydrocarbons and reduced
ferric iron. Re-activation of the species from liquid culture following several months
without transfer to fresh medium reflects the documented ability of anaerobic
biomass to become fully operational shortly after an extended period of dormancy
(Anderson et al., 1982).
An additional inoculum, derived from the dual-phase sediment slurry and
without isolation on plates, was also chosen for further work. This inoculum
source, labeled QU2 consortium, would be used in its slurry form to inoculate
microcosrns. QU2 was obtained from sediment collected under approximately 10
feet of water in Lake Ontario off the shore of Kingston. The area i s the site of a
former gassification plant that operated from 1848 to 1957. Soil and bedrock
analyses performed in 1988 showed high levels of coal tar contamination and its
migration over several blocks (Kingston This Week, 1 997 (September 3)).
A homogenous, sediment sample from the former gassification site was
analyzed for PAH composition at an extemal laboratory (Novamann Maxxarn
Analytique, Quebec). The results are summarized in Table 4.3. In addition to
PAHs, high background levels of ferrous iron (1.4 mg/L) were also detected.
Table 4.3 PAH concentrations in QU2 consortium sediment.
Compound Concentration
(mglkg dry sediment)
Naphthalene Acenaphthylene Acenaphthene Fluorene Phenanthrene Anthracene Fluoranthene Pyrene Chrysene Benzo(b + j + k)fluoranthene Benzo(a) pyrene Indeno(lf2,kd)pyrene
Total PAHs (21 analyzed) 255.3
4.4 Conclusions
Several cultures were enriched from a variety of environmental samples and
screened for their potential to degrade PAHs when ferric iron was provided as a
terminal electron acceptor. Based on this criteria, two inocula, labeled QU 1 and
QU2, were chosen to be used in succeeding experiments.
l ntroduction
For almost 100 years, combustible gas was manufactured from coke, coal,
and oil at 1 O00 to 2000 sites across the United States (Luthy et al., 1994). Poor
waste management practices has led to the contamination of soil, sediment, and
groundwater at many of these sites. Process residuals including ammonia liquors,
a h , oils, tars, and sludges remain at these locations and serve as a continual supply
to groundwater contamination. Al though tan are compris4 of hundreds of
compounds, the primary constituents include 2-, 3 , 4-, and Eringed PAHs and
BTEXs (Hughes et al., 1997). Consequently contaminants exhibiting a wide range
of physiochernical properties can be present (Table 2.1). The low sotubility, high
adsorption capacity, and low volatility associated with high molecular weight PAHs
has been correlated with their recalcitrance. The degradation rates of these
compounds often dictates the time required to remediate these sites.
Previous laboratory work with pure cultures has demonstrated the diversity
that exists within the PAHdegrading microbial community. Some species have the
ability to metabolize a broad spectrum of PAHs (Boldrin et al., 1993; Muelier et al.,
1991; Tiehm and F ritzsche, 1995; Weissenfels et ai., 1990), whereas others are
effective only on lower molecular weight compounds (Bouchez et al., 1995;
Cemiglia, 1 992; Cernigl ia, 1 993; Heitkamp and Cern igl ia, 1 987). The need exists
to be able to biodegrade a wide array of contarninants from within a complex
mixture and matrix.
5.2 Experimental
Duplicate biotic microcosms and an abiotic control were set up to study the
mineralization of each of the following six PAHs and one monoaromatic
hydrocarbon: naphthalene, phenanthrene, anthracene, fi uoranthene, pyrene,
benzo[a]pyrene, and toluene. PAHs were separately provided as both ["Cl-
radioisotopes dissolved in either tol uene or methanol, and non-radiolabeled crystals
to a final concentration of 20 pprn. Non-radioactive and ['4CJtoluene are liquid at
room temperature and were added as such to a concentration of 20 ppm. Table 5.1
summarizes the experirnental set-up for the radioactive components used.
Heterogeneous treatment slurries were prepared in 1 25-mL glass, serum
bottles sealed with a butyl rubber stopper and aluminurn crimp. A final volume of
Table 5.1 Experimental design used for mineralization studies. Some compounds were added with a uniformly labelled (UL) carbon backbone, whereas others were restricted to selected locations as indicated in the second column.
Volume Activity Radioisotope 14C Positioning Solven t (PL) (dpm)
Naphthalene UL Methanol 10 92,000 Phenanthrene 9 Met hanol 1 100,000 Anthracene UL Tol uene 1 86,000
Pyrene 4,5,9,1 0 Toluene 0.2 99,000 Fluoranthene 3 Methanol 1 770,000
Benzo[aJpyrene 7 Toluene 8 85,000 Toluene UL None 0.5 108,000
100 mL was achieved with: 50 mL defined growth medium (Table 4.1), 10 mL
ferrihydrite solution which provided a stoichiometric excess of Fe (100 mg), 30 mL
distilleci water, and 10 mL of QU2 consortium (10% inoculum). The contents were
purged with N, to remove oxygen, and incubated in the dark without mixing. A
final 0.26/0 (wt/voi) solution of sodium azide was added to sterilize slurries for use as
abiotic controls. I4CO, was collected and analyzed as described in Chapter 3.
5.3 Results and Discussion
5.3.1 Naphthalene
The aqueous naphthalene concentration at former manufactured gas plant
(MGP) sites i s often used as a worst-case indicator of groundwater contamination
due to its relatively high solubility. Of the compounds that exist within the water-
soluble fraction of petroleum, naphthalene is one of the most toxic (Heitkamp et al.,
1987). Fortunately, naphthalene is one of the easier PAH compounds to
microbially break down (Hughes et al., 1997; Shunleworth and Cerniglia, 1995).
Because of its known potential toxicity, mineralization experiments with
naphthalene were performed at concentrations both below (20 ppm) and above its
solubility (50 ppm). The results of these experiments are shown in Figure 5.1.
Abiotic loss of naphthalene, over a period of five months, only arnounted to
5.5 of its initial concentration. Although reasonably low, naphthalene
evaporation was over 16 times faster than for phenanthrene, which was the next
most volatile.
O 20 ppm 50 ppm
A Abiotic Control
O 20 40 60 80 100 120 1 40 1 60
Time (Days)
Figure 5.1 Minerakation of [l4qnaphthalene in duplicate anaerobic sedi ment slurry microcosms.
6 iomineral ization of naphthalene was highl y dependent upon its
concentration. When suppl ied at 20 ppm, naphthalene was immediatel y
mineral ized without a lag period. Initial rates of mineral ization (calculated over the
first week of incubation) suggest thaï biological activity controlled degradation rates.
Following the initial penod of apid growth, rnineralization decreased in a
logarithmic-like pattern until, at approximately ten week of incubation, the rate of
degradation became linear, probably due to mas transfer limitation of naphthalene
adsorbed onto the sediment matrix. This pattern was also evident in desorption and
biodegradation experiments perforrned by Al-Bashir et al. (1 990). They attributed
linear growth to the strong sorption of naphthalene ont0 the soi1 matrix, and the
subsequent slow rate of dissolution into the aqueous phase where it became
microbially available. They found that substrate-dependent kinetics was dominant
once the aqueous phase naphthalene concentration dropped below the saturation
1 eve! .
The experiment was terminated after approximately five months, accounting
for almost 55% naphthalene mineralization. It was not determined what percentage
of the radioactive carbon was incorporated into biomass or deadend metabolic
products. The gradval development of an orange hue within the overlying aqueous
layer in settled microcosms was observed. This may be an indication of the
accumulation of the deadend product, 1.2-naphthaquinone (Auger et al., 1995)~ or
the precipitation of ferric oxides (see Chapter 9). Quinones emit a characteristic
orange colour within an aqueous solution, and 1 ,tnaphthaquinone is known to
resist fiirther transformation,
Experiments done with 50 ppm naphthalene showed no mineralization,
compared to the abiotic control, over the entire five month duration. This
characteristic toxicity of naphthalene was also observed by Bouchez et al. (1 995),
Muelier et al. (1 990). and Weissenfels et al. (1 991 ). Bouchez et al. (1 995) were
able to induce growth when naphthalene was supplied in the vapour phase, but not
as crystals. Nevertheless, cultures which are capable of degrading much higher
concentrations of naphthalene have been identified (Al-Bashir et al., 1 990). This
further i Il ustrates the potential divenity between ecosystems and the importance of
site-specific treatability studies.
5.3.2 Phenanthrene
Mineralization of phenanthrene occurred in sediment slurry microcosms
over a five-rnonth period (Figure 5.2). Abiotic phenanthrene loss was linear, and at
a much lower rate than for naphthalene (0.3% over 100 days). The duplicate biotic
microcosms both showed good reproducibility (average standard deviation of 0.44
in un its of percent mineral ization) and imrnediate mineral ization without a lag. The
initial rate of mineralization was only 40% as fast as that for naphthalene, and
amounted to almost seven percent phenanthrene mineralization at the cornpletion
of the experiment.
If it is assumed that the initial rate-limiting step is biokinetic in origin, this is
easii y explained. Mineralization of phenanthrene, a 3-ring PAH, would require a
more complex metabolic pathway than for naphthalene, which has only two rings.
Consequently, it is reasonable that degradation wouid occur at a slower rate,
especially if ring cleavage is the rate-limiting step.
If dissolution, on the other hand, is the initial rate-controlling factor, slower
mineral ization can be due to differences in aqueous solubility and sorption.
Naphthalene is about 30 times more soluble than phenanthrene, and less Iikely to
sorb onto sedirnent organic matter (see Table 2.1). Similarly to naphthalene, rates
of phenanthrene minerakation decreased with time and became linear after ten
weeks of incubation.
O 20 40 60 80 100 120 1 40 160
Time (Days)
Figure 5.2 Mineralization of [I4qphenanthrene in duplicate anaerobic sediment slurry microcosms.
5.3.3 Toluene
Considering the frequent presence of BTEXs at former MGP sites, the QU2
consortium was evaluated for its ability to degrade toluene (Figure 5.3). Abiotic
losses were comparable to naphthalene, relating to its high volatility and solubility.
O Active Microcosm
A Abiotic Control
O 20 40 60 80 1 00 120 140 1 60
Time (Days)
Figure 5.3 Mineralization of ["qtoluene in dupl icate anaerobic sediment slurry m icrocosrns.
No biomineralization was observed until about three weeks of incubation, when the
rate of '%O, production in the active microcosms increased relative to the abiotic
control. A linear mineralization rate was achieved after six weeks and continued
üntil almost four months at which point the rate decreased slightly. The total
disappearance of toluene by the end of the experimental period amounted to only
seven percent, one-third of which was due to biological activity.
An explanation for this may be associated with the source of the inoculum.
The sedirnent used to inoculate each microcosm was taken from a site with
prolonged exposure to rnoderate levels of PAHs (Table 4.2). There was probably a
seledive pressure that enriched for microbial cornmunities capable of PAH
degradation. Two inferences may be made. First i s that the enzyme systern used by
QU2 for the degradation of PAHs is incapable of breaking down toluene. The
second conclusion that can be drawn is that the nurnber of toluenedegrading cells
must be relatively low initially, and is responsible for the observed acclimation
period. This i s reasonable since the indigenous microorganisms would most likely
have depleted the supply of monoaromatic hydrocarbons much earlier, resulting in
predominantly PAHdegraders.
5.3.4 PAH Family of Compounds
Although only naphthalene, phenanthrene, and toluene were discussed in
detail, al1 of the previously mentioned seven organic compounds (naphthalene,
phenanthrene, anthracene, fi uoranthene, pyrene, benzo[a]pyrene, and toi uene)
were studied. These compounds were chosen to provide a range of contaminants
with varying physiochemical properties, from monoaromatic hydrocarbons to 2-, 3-,
4-, and 5-ring PAHs. The ability of the mixed microbial consortium QU2 to
degrade these constituents is an important indicator for the potential feasibi lity of in
situ remediation ai the former MGP site.
Table 5.2 provides a sumrnary of the biodegradation results for the
contaminants examined. In general, the magnitude of abiotic loss for each
compound corresponded with the trend in their solubility and volatility (Table 2.1 ).
The initial biotic rate of mineralization was deterrnined as the average 14C evolved
over the first week of incubation.
Table 5.2 Summary of mineralization rates for organic compounds studied.
Rate of Mineralization (ndday)
Initial Conc. Average '10 Minera! ized
Compound (pprn) ~b io t i c Initial Biotic Linear Biotic (at Day 147)
Toluene 20 Naphthalene 20
50 Phenanthrene 20 Anthracene 20 Fluoranthene 20 Pyrene 20 Benzo(a]pyrene 20
Naphthalene showed the highest rate of biomineralization, initial ly, and
when control led by diffusional limitations. This is consistent with other reports
comparing the biodegradation of various PAH compounds (Bauer and Capone,
1988; Herbes and Schwall, 1978; Park et al., 1990). Phenanthrene was also
mineralized at high rates throughout the duration of the experiment. Its relative
ease of biodegradation was also shown by Weissenfels et al. (1 990) and Park et al.
(1 990).
Anthracene, pyrene, and possibly benzo[a]pyrene exhibited indications of
initial biological mineralization without any accl imation period, but rates quickly
declined to leveis comparable with their respective abiotic controls. It is likely that
the poor bioavailability of these PAHs is the cause of their limited mineralization.
The minimal solubility and strong sorption of the higher molecular weight PAHs
will result in very low concentrations in the dissolved phase. Biological activity will
depend upon the desorption and dissolution of these compounds. It has been
suggested that at the extremely slow rates of dissolution for these organic
constituents, the total mass of degradable carbon could drop below the critical level
needed to maintain a viable population of PAHdegrading bacteria (Brubaker and
Stroo, 1992).
The initial mineralization of naphthalene, phenanthrene, anthracene, pyrene,
and benzo[a]pyrene immediately fol lowing inoculation impl ies that either a pure
culture within the QU2 consortium contains an oxygenase system common to each
of those compounds, or multiple species are present within the inoculum with
unique enzyme systems. Further study is required to answer this question.
Fluoranthene was the only compound which showed no mineralization at
any point. Although cultures have been identified which can utilize fluoranthene as
sole carbon source (Keck et al., 1989; Wiessenfels et al., 1990; Muelier et al.,
1991), Bouchez et al. (1 995) found that a pure culture of a Pseudomonas sp was
unable to degrade fluoranthane alone but could cometabolically metabol ire it in
the presence of phenanthrene. This could also be true of QU2, and hence
responsible for its inabiliw to metabolize fluoranthene as the sole carbon and
energy source.
5.4 Conclusions
The ability of QU2 to degrade a diverse group of individual organic
compounds was evaluated. Naphthalene and phenanthrene were susceptible to
degradation, and toluene following a short lag period. The higher molecular weight
PAHs seemed to be amenable to degradation but at much slower rates probably due
to lower solubility and higher adsorption of these compounds.
Chapter 6 En hancement of PAH Bioavailability
6.1 Introduction
In general, the farnily of PAH compounds demonstrate an inverse
proportionality between their degree of bioavailability and their molecular weight.
A consequence of this is that high molecular weight compounds may not provide
sufficient carbon to sustain microbial growth. This is reflected in their observed
recalcitrance. Low molecular weight compounds, although susceptible to microbial
degradation, are frequently removed at a rate which is govemed by dissolution
kinetics (Cerniglia, 1993). Techniques to accelerate degradation by increasing the
aqueous substrate concentration has, as a result, become a growing area of
research. Many innovative approaches have been taken, but only few have found
application at full-scale. The effects of surfactants, such as Triton x-100
(C,H,,C,H,O(CH,CH,O),,H), and Brij35 (C,,H,,O(CH,CH,O),,H), and cyclodextrin
on bioremediation were investigated in this report.
6.2 Synthetic Surfactants
Surfactant-amendment is the most frequently applied technology to enhance
the bioavailability of hydrophobic contaminants (Shuttleworth and Cerniglia, 1995).
The dual acîion of a surfactant's hydrophobic and hydrophilic moieties serves to
mediate between the immiscible aqueous and organic phases Viehm, 1 994). The
resulting solubil ization or emulsification improves the rate of mass transfer of
contaminant5 from the solid to the liquid phase. Yeom et al. (1 996) explained that
this was due to an increase in the concentration gradient at the soil-water interface,
and an increase in diffusivity of PAHs due to the swelling of the soi1 organic matrix
resulting in contaminant mobilization and concentration in the water phase.
The ability of surfactants to solubilize and mobilize PAHs has been
demonstrated (Auger et al., 1995; Tiehm, 1994; Tiehm and Fritzsche, 1995).
However, their value in bioremediation is uncertain due to contradictory reports in
the literature. The objective of the work presented in this chapter was to establish
whether the selected surfactants (i) affected microbial growth and (ii) could enhance
PAH degradation.
6.2.1 Experimental
Batch experiments were conducted, in du pl icate, in 1 25-m L screw-capped,
glass, amber bottles fitted with teflon,silicone septa and filled to capacity. The
growth medium contained, per liter, 0.5 g yeast extract, 0.9 g Fe as hematite, 0.9 g
EDTA, 50 mg phenanthrene, plus Brij35 at either 2.2 g or 4.4 g or Triton x-100 at
1.5 rnL or 3.0 m l in defined anaerobic mineral salts medium (Table 4.1). The high
and low concentrations for each surfactant were chosen based on their theoretical
ability to solubilize 50 and 100 ppm phenanthrene, respectively. The contents of
the bottles were sparged with extra-high purity nitrogen gas for 15 minutes, then
inoculated with an anaerobic culture of QU 1 consortium. Abiotic controls were
prepared with the addition of 1.1 g/L mercuric sulphate. Each microcosm was
incubated at room temperature and mixed on a Modular Cell Production ROI fer
Apparatus (Wheaton Instruments) at 100% motor speed.
A homogeneous, EmL, slurry sample was obtained using a syringe with a
bevelled needle. Solid hematite was allowed to settle briefly (approximately 1 5
seconds), and turbidity of the aqueous phase rneasured at 600 nm on a UV-Vis
scanning spectrophotometer (Phil ips PU8720) as an indication of microbial growth.
6.2.2 Resu lts and Discussion
Anaerobic microcosms were set up with concentrations of Brij35 or Triton X-
100, from 18 to 36 times and 1 i to 22 times their CMC values respectively, to
determine whether these surfactants affected the growth of the inoculated rnicrobial
species. The turbidity of the abiotic control, except for a small gain initially,
remained relatively constant over the entire threeweek trial. Following a two-week
acclimation period, exponential growth was indicated in du pl icate microcosms
without surfactant amendment. There was no growth of QU1 with phenanthrene
solubilized by either Brij35 or Triton X-100 at either concentration used.
To differentiate between inhibition caused by a higher aqueous
phenanthrene concentration and surfactant toxicity, the experiment was repeated
with toluene (50 ppm). Over a four-week sampling period, biotic duplicates
showed good reproducibility with an average standard deviation in absorbance
units of 0.009. For clarity, only the average values are plotted in Figure 6.1.
Although the turbidity in surfactant-amended microcosms varied slightly throughout
6 no surfactant 1.51 ml/L Triton x-1 O0
O 2.98 mUL Triton x-1 O 0
A 2.2 g/L Brij35
A 4.4 gk 8rij35
M Abiotic Contol CTriton x-100)
O Abiotic Contol (Brij35)
Figure 6.1 Average turbidity (absorbante at 600 nm) showing toxic effect of Brij35 and Triton x-100, at high and low levels, on the degradation of 50 ppm toluene inoculated with QU1 consortium.
the course of the experiment, no significant change was observed when compared
to abiotic controls. In contrast, the biotic samples without surfactant increased in
turbidity by a factor of four to five.
The results cleariy demonstrate the toxic effm of these two surfactants at the
concentrations tested on the QU1 consortium. These findings coincide with those
of Laha and Luthy (1 991,1992), Aronstein et al. (1 991), and Tiehm (1 994), among
others, who observed inhibition of microbial growth in the presence of synthetic
surfactants above their CMCs. However, unlike Laha and Luthy's work, inhibition
occurred even when supplied with toluene at a non-toxic concentration, indicating
that the toxicity was associated with the surfactant itself, rather than the elevated
aqueous PAH concentration.
It is possible that at a lower surfactant concentration, toxicity would not have
been observed. However the effective enhancement of bioavailability is
questionable at these levels. It has been proposeci that displacement of non-
aqueous phase liquids (NAPLs), such as PAHs, would require a significant reduction
in the NAPL-water interfacial tension resulting in the need for large doses of
surfactant (Abriola et al., 1993; Pennell et al., 1993). Further quantities of
sudactant, in excess of this arnount, would be required to cornpensate for the
propensity of surfactants to sorb ont0 the soi1 rnatrix.
To further investigate the growth that was detected on phenanthrene without
surfactant, the experiment was repeated. In addition tu QU 1 consortium, three
other uncharacterized, plated inocula (K-2, K-6, K-26) were studied for their ability
to degrade phenanthrene (50 ppm) and naphthalene (32 ppm) in anaerobic
microcosms. In the absence of surfactant, these microcosms were set up, and
followed by monitoring turbidity and soluble PAH concentration.
Except for the QU 1 consortium, none of the active rnicrocosms showed any
growth or change in PAH concentration through the experimental period compared
to the abiotic control. Afier a ho-wedc lag period, QU1 consortium showed an
apparent exponential increase in turbidity. Concurrent with the onset of growth
was the development of foam at the air-water interface when shaken, and elevated
soluble PAH concentrations. Compared to the abiotic control and the other biotic
consortiums, the soluble naphthalene and phenanthrene concentrations were four
and five times higher, respectively with Q U I . These observations are consistent
with biosurfactant production (Finnerty, 1 994).
B iosurfactants are microbial l y-synthesized surfactants wh ic h possess many of
the same qualities as their synthetic counterparts but are biodegradable, potentially
less toxic, exhibit lower CMC values, and therefore are more environmentally-
acceptable. Biosurfactant production is often seen in some microorganisms when
carbon bioavailability is limited (Finnerty, 1 994).
6.3 Cyclodextrin
Cyclodextrin, a cycl ic oligosaccharide (Brusseau et al., 1 997), has recently
been found to be a potentially effective aid in contaminant remediation. Its
properties of low soi l reactivity, biodegradabi l ity, and non-toxicity (Wang and
Brusseau, 1993) make cyclodextrïn amendment a strong candidate as a
bioavailabilityenhancing technique. Some cyclodextrins are equally as effective at
removing weathered contaminants as from recentl y contam inated soi ls (Brusseau et
al., 1997). Three different cycloûextrin homologues (a, P, and y) exist which
possess varying propeities. &clodextrin is the least expensive, but exhibits poor
water solubility (Wang and Brusseau, 1993). Chernical modification of this element
to form any of the derivatives hydroxypropyl-~clodextrin (HPCD), su lfated-P
cyclodextrin (SCD), pmethyl cyclodextrin (MCD), and carboxymethyl-p
cyclodextrin (CMCD) enhances its water solubility. HPCD is particularly water-
soluble.
The cyclodextrin molecule can be represented by a toroid (Figure 6.2) where
O R '
O$: 8- - @" Figure 6.2 Schematic of Bcyclodextrin. A: chemical structure; B: three- dimensional rnolecular model; C: topology. (Source: Manunza et al., 1998)
hydroxyl groups are armnged along its exterior but are absent from its cavity
(Manunza et al., 1998). Accordingly, this configuration results in properties of
hydrophilicity on its surface and hydrophobicity intemally. Such a compound is
ideal for solubilizing low-polarity organic cornpounds such as PAHs (Wang and
Brusseau, 1995). Detailed information regarding cyclodextrin as a solubilizing
agent is limited. Data that does exist is primarily restricted to physiochemical
response (Brusseau et al., 1994; Brusseau et al., 1997; Wang and Brusseau, 1 993).
The objectives of the studies in this chapter were (i) to examine the
solubi l ization of different cyclodextrin homologues on various PA&, (i i) determ ine
their capacity to desorb analytes from a solid matrix, (iii) evaluate the transportation
of the solubilized PAHs through a porous medium, and (iv) study the effect of
cyclodextrin on the bioremediation of PAHs.
6.3.1 Experirnental
6.3.1 .1 PAH Solubilization in an Aqueous Solution
individual PAHs (naphthalene, acenaphthene, phenanthrene, anthracene,
pyrene, or fluoranthene) were added to 20-mL scintillation vials at final
concentrations above their solubilities (1-1 0 mg). Appropriate volumes of water and
an aliquot of a stock solution of SCD, MCD, or HPCD were added to give final CD
concentrations in the range 1-5% (wtlvol). Equilibration was achieved by mixing
the vials on a platform shaker for 48 houn. &nL aliquots were then withdrawn and
centrifuged at 7000 rpm for 1 5 min. PAH concentrations were determined, in the
presence and absence of CD, by UV spectrophotometry by monitoring the
corresponding maximum absorbante wavelength in the region 200 to 500 nn.
6.3.1.2 PAH Desorption from a Sand Matrix
Pristine sand, pre-washed with distil led water and dried in an oven at 500°C,
was used to create a slurn/ with 1 5 mL distilled water and 1 00 pL of stock pyrene
solution to give a final concentration (0.07 PM) which was below its saturation
limit. Cyclodextrin was added at different concentrations to produce a range
between 2.7 and 25 mM. A control without cyclodextrin and another without sand
were also set up and run in parallel. Aqueous pyrene concentration was
determ ined with a fi beroptic instrument with fluorescence detedion currentl y under
development in the laboratory of Dr. S. Brown at Queen's University. A linear
response was obtained for the concentration range tested. The initial aqueous
pyrene concentration was determined for each batch, following a brief period of
mixing. Mixing of the slurry continued for an additional 24 hours using a magnetic
stirrer. An aliquot was taken and centrifuged at 7000 rpm for 15 minutes and the
liquid analyzed for soluble pyrene concentration. Further mixing for 24 hours was
followed by a final pyrene analysis. Ottawa sand was used in al l experirnents.
6.3.1 -3 Mobilization of PAHs throunh a Packed Column
Two columns (50 mL burets) were packed, from bottom to top, with a thin
layer of glass wool, 10 cm of pristine sand, 5 g of sand impregnated with 0.5 mg
pyrene and 1.2 mg phenanthrene in ethanol, and an additional 10 cm of pristine
sand. The columns were wetted with water and a peristaltic feed was attached at
the top. Distilled water was introduced into one of the columns and a 4% (wtlvol)
solution of MCD in the other. Each solution was applied at an approximate
flowrate of 5 mumin. Aqueous sarnples were periodically taken from each outlet
and tested for pyrene and phenanthrene concentration using fiberoptic fluorescence
detection.
6.3.1.4 Biornineral ization of PAHs
Du pl icate microcosms were prepared in 1 25-rnL glas boules sealed with
butyl rubber stoppers and crimped to form an airtight seal. Each microcosm
contained: 83 mg Fe as ferrihydrite, and 40 rnL Lake Ontario water collected at the
former MGP site where the QU2 consortium was collected. A final volume of 80
rnL was achieved with distilled water and a 10% (dv) inoculation with QU2
consortium. HPCD was added to give a final concentration of 50, 500, or 5000
mg/L. A control, without cyclodextrin, was included as well. Phenanthrene was
introduced to each microcosm as solid crystals and a radiolabeled spike (89,000
dpm) to provide a final concentration of 20 ppm. An abiotic microcosm was
sterilized with sodium azide at a final concentration 0.2% (wt/vol). Microcosms
were stored in the dark, at room temperature, without mixing. Each bottle
contained a KOH trap that was sampled periodically and analyzed for 14C0,
according to the methods outlined in Chapter 3.
6.3.2 Results and Discussion
6.3.2.1 PAH Solubilization in an Aaueous Solution
Three cyclodextrin derivatives were studied for their ability to enhance the
solubilization of six representative PAHs. Of these, HPCD displayed the strongest
solubilizing power, followed by MCD and SCD. Only the solubilization effect of
HPCD on PAHs is shown (Figure 6.3). A iinear relationship was observed between
cyclodextrin concentration and the apparent aqueous solute concentration for al1
compounds tested. This behaviour was expected since cyclodextrin forrns 1 :i
inclusion complexes when solubil izing compounds (Wang and Brusseau, 1 993).
Pyrene and phenanthrene were solubilized the greatest, with enhancement
factors, at 5'6 (wt/vol) HPCD, of approximately 50 and 42, respectively. The
remaining four PAHs exhibited increased solubil ization by factors of no less than 4
and no more than 13 times. The varying solvating strengths for the analytes tested
has been explained by di fferences in stereoselective interactions and h ydrogen
bond formation (Wang and Brusseau, 1 993).
However, the enhancernent factors compare poorly with similar studies done
with HPCD, which showed naphthalene and anthracene dissolution 7 and i i tirnes
greater, respective1 y (Wang and Brusseau, 1 993).
6.3.2.2 PAH Desoption frorn a Sand Matrix
The adsorption/desorption kinetics of pyrene in sand slurries was examined
to determine the effectiveness of cyclodextrin addition in maintain ing the
Naphthalene A Acenaphthene
A Phenanthme O Anthracene
Fluoranthene
1 2 3 4
Weight O/a HPCD
Figure 6.3 Sol ubilization-enhancement effect of HPCD on seleaed PAHs, where Sf io is the aqueousphase concentration of the d u t e with HPCD relative to its natural solubility in the absence of HPCD.
concentration of hydrophobic contaminants in the dissolved phase. ~ f t e r a lag
period, dissolved pyrene was rapidly depleted in the cyclodextrin-free control in
response to its adsorption ont0 the sand matrix or the surface of the glass column
(Figure 6.4). After 48 hours, soluble pyrene was reduced to almost half of its initial
+ no amendment
++ 15 mM MCD (no sand)
-rlr 2-7 rnM MCD -C- 5.3 mM MCD
-C- 25 mM MCD
O 10 20 30 40 50
Time (Hours)
Figure 6.4 Desorption capacity of MCD for pyrene on pristine Ottawa sand.
value. It is anticipated that higher soil-to-water ratios would have a significant
impact on substrate availability for microbial attack.
The physiochemical response of cyclodextrin on soluble pyrene was
evaluated by observing the effect of a 15 m M MCD solution in the absence of sand.
In the first 24 hours, the concentration of soluble pyrene decreased by 1796. The
next 24 hours, however, resulted in a rebound of soluble pyrene to 97% of its initial
value. No proven theory exists to explain this observation, but severai possibi l ities
are offered. Pyrene may adsorb ont0 the surface of the glass apparatus, followed by
complexation with cyclodextrin and re-dissolution. Another possibility is that the
initial complexation reaction is rapid, but its soiubilization is rate-limiting.
Cyclodextrin mixed with sand resulted in a consistent decrease in soluble
pyrene, over the first 24 hours, of 24%, regardless of MCD concentration. The
additional disappearance, over cyclodextrin alone, is due to adsorption onto the
sand rnatrix. After 48 hours of mixing, the average soluble pyrene concentration
returned to more than 94O/0 of its original concentration. Higher cyclodextrin
concentrations resulted in a marginal increase in desorption. Without cyclodextrin,
soluble pyrene concentrations never increased following adsorption. This illustrates
a common difficulty encountered when attempting to treat areas contaminated with
higher molecular weight PAHs.
6.3.2.3 Mobilization of PAHs from a Sand-Packed Column
A legitimate concern of in situ soi1 flushing with cyclodextrin is the poteniial
for contaminant spreading by leaching PAHs, and polluting nearby groundwater.
Enhanced mobil ization was simulated by applying a steady flow of a 4% (wt/vol)
MCD solution through a porous matrix impregnated with phenanthrene and pyrene.
These results were compared to distilleci water applied at the same flow rate.
Figures 6.5 and 6.6 show the breakthrough curves for phenanthrene and pyrene,
respectively through sand-packed columns.
O 1 2 3 4 5 6 7 8 9 10
Pore Vdurnes
Figure 6.5 Elution profiles for phenanthrene with water and 4% (wtfvol) MCD as desorption solutions.
Ten pore volumes of water removed on ly 4% of the phenanthrene frorn the
sand matrix. Although probably not a hurnan health hazard, this concentration may
a 4 Wto? MCD
A Water
O 2 4 6 8 10
Pore Vdumes
Figure 6.6 Elution profiles for pyrene with water and 4% (wthol) MCD as desorption solutions.
have adverse effects on the aquatic or terrestrial ecosystems (Environment Canada
and Health Canada, 1 994). For the same volume of 4% (wtlvol) MCD solution,
approximately 1 5% of the adsorbed or insoluble phenanthrene was removed. A
plateau has been reached whereby no further phenanthrene will be desorbed.
Much higher desorption was observed by Brusseau et al. (1 997) where almost
i 00% of the phenanthrene was eluted.
Only 0.75% of pyrene was recovered in ten pore volumes of water. Acute
toxic effects are not of concem at these concentrations, however the
bioaccumulation of pyrene may present chronic problems. Pyrene desorption
increased markedly to 90% when a comparable volume of 4% (wt/vol) MCD was
passed through the column.
6.3.2.4 Biomineralization of PAHs
The solubilization and improved mobilization of low-polarity organic
cornpounds using cyclodextrin has been repeatedly shown in literature (Brusseau et
al., 1994; Brusseau et al., 1997; Wang and Brusseau, 1993). Never has this
technology been studied for the purpose of enhancing the bioavailability of PAHs in
sediments to accelerate their degradation. For the first time, this was examined.
Samples were taken, on average, weekly for a period of eleven weeks and analyzed
for percent phenanthrene minerai ization (Figure 6.7). Over this tirne the abiotic
control showed minimal activity, accounting for phenanthrene volatil ization andor
incomplete sterilization. Minerakation in each of the biotic microcosms occurred
without lag, immediately followed by a decline in activity. lmrnediately following
inoculation, there appears to be a negligible difference in the rate of metabolisrn for
each of the active systems. This changed following one week of incubation, when
the 0.5% (wt/vol) HPCD microcosm showed an increase in reaction rate above the
others. The 400,40, 4, and O mg HPCD microcosms have instantaneous
no amendment O
O 4mg HPCD - R4û mg HPCO
U 400 mg HPCD A Abiotic Control
O 10 20 3 0 40 50 60 70 80
Time (Days)
Figure 6.7 Effect of HPCD concentration on ("Cjphenanthrene mineralization in anaerobic slurry microcosms inoculated with 2% (w/v) Kingston sediment slurry. Data points represent the average of dupl icate microcosms.
mineralization rates at this time of 0.23, 0.14, 0.1 4, and 0.1 2 ppdday,
respectively. The delay in enhancement can be explained by the tirne required for
the HPCD to complex with phenanthrene and for the ceIl density to increase.
Unexpectedly, from weeks three to six, both du pl icate m icrocosms with 400
mg HPCD had gradually decreasing phenanthrene mineralization activity. The rate
of mineralization at week six is only 2.4% of that at week one. Since both of the
duplicates displayed this behaviour, the results are likely representative. Because
degradation proceeded successfully for several weeks prior to this event, HPCD
toxicity is not a viable explanation. The accumulation of toxic by-products can also
be ru led-out since the other HPCD-amended microcosms continued to flourish
upon reaching and surpassing the equivalent total mineralization. The most
probable answer is that the phenanthrene-degrading microbial commun ity began to
consume HPCD as a preferential carbon source over phenanthrene. Because the
400-mg HPCD microcosms contain 1 0, and 100 times the quantity of the other
HPCD-amended systems, a greater selection pressure was created for alternative
carbon consumption. This behaviour was not observed in either of the other
HPCD-amended slurries during the eleven-week period in which mineralization
was monitored. It is conceivable that at lower HPCD concentrations this metabolic
transformation might require a longer incubation period.
After eleven weeks of incubation, the HPCD control had mineral ized 30%
phenanthrene, while the 4 and 40-mg HPCD-amended slurries mineralized 38O/',
and 3 6 O / 0 , respectively. Due to the nature of the inclusion complexes, each
molecule of contaminant solubilized requires one cyclodextrin molecule. Without
the possibility of recycling, this technique may be cost prohibitive. Further
experimentation i s necessary to predict whether the possibi lity of recycl ing H PCD,
and hence reducing associated costs, exists. Also, a wider range of PAHs should be
studied, individually and in mixtures. Only then can it be determined if and when
HPCD-amendment is advantageous.
6.4 Conclusions
Conventional synthetic surfactants as well as new bidegradable agents were
examineci for their potential to enhance the bioavailability of PAHs. At
concentrations above their CMC, Bri j35 and Triton X-100 were both toxic to QU 1,
and hence were elirninated from further studies. It was found that QU1 was
capable of producing a biosurfaaant, probably in response to carbon limitation, and
consequently increased the concentration of soluble naphthalene and
phenanthrene.
Several cyclodextrin derivatives were tested on PAHs for their ability to
enhance solubilization, effect desorption from a sand matrix, accelerate
contaminant mobi lization through a sand-packed colurnn, and increase the rate and
extent of biotic mineralization. For most PAHs tested, a high concentration of
cyclodextrin was required to provide a sign ificant en hancement of its sol u bi l ization
and desorption. However, results of the mineralization study suggested that at high
concentrations the QU2 consortium developed a tendency to preferentially
consume CD as carbon source instead of the contaminant.
Cyclodextrin-amendment did have a positive influence on phenanthrene and
pyrene bioavailability at low concentrations. Further work needs to be perforrned
to detemine whether cyclodextrin could be used to enhance biodegradation of
higher molecular weight PAHs.
Chapter 7 lron Chelation
7.1 Introduction
lncreasing the soluble fraction of PAHs is only half of the challenge
associated with their anaerobic biodegradation with ferric iron as terminal electron
acceptor. Although at the onset of anoxic conditions Fe(l Il) represents greater than
902 of the oxidative capacity in many environments (F redrickson and Gorby,
1996), at a neutral pH, iron is present as relatively insoluble fenic oxides with a
water solubil in/ of approximately 1 O-" M (Chiswell and Zaw, 1 989; Fox, 1 988;
Murray, 1979; Schwertmann, 1988; Stumm and Morgan, 1983). At such low
concentrations, its intrinsic enzymatic redudion in biologically-mediated reactions
is limited (Anderson and Morel, 1982; Arnold et al., 1985; Fredrickson and Gorby,
1996; Lovley et al., 1994).
A variety of Fe (Ill)-oxides and oxyhydroxides results from soi1 and sediment
biogeochemical processes (Fine and Singer, 1989). Typical minerals include
ferrihydrite, lepidocrocite, maghemite, magnetite, hematite, and goethite. The
reduaive dissolution of Fe(lll) is apparently dependent upon the mineral stabil ity,
with crystalline species being more recalcitrant (Arnold et al., 1988; Arnold et al.,
1986). The transition from amorphic to structured oxides with age, and therefore
depth of sediment, is reflected in the persistance of ferric species in deep aquifers.
(Coey et al., 1974; Frouelich et al., 1979; Sakata, 1985; Verdouw and Dekken,
1 980; Walker, 1 984; Phillips et al., 1 993). This has important implications for the
natural attenuation of organic contaminants in these environments.
Difierences in the strength and number of bonds to be broken and steric
hinderences of the various mineral identities have been offered as explanations for
the observeci selective dissolution (Borggaard, 1990). The redox readion is further
complicated by multiple phases and requires interaction between the cell and solid
ferric oxide. Possible mechanisms proposed include: solubilization of the ferric
oxide, direct physical contact on the surface followed by electron transfer, or
transport of the iron oxide into the cell as a solid (Neaison and Myers, 1992).
An alternative approach is the use of organic ligands to chelate and
solubil ize the ferric oxides. It was found that when Fe(lll) is presented in the
dissolved phase, dissimilatory iron redudion can be very fast (Arnold et al., 1988;
Lovley et al., 1 994; Lovley and Phillips, 1 988). This resulted in accelerated rates of
degradation of compounds such as toluene and benzene when coupled to
dissirnilatory iron reduction (Lovley et al., 1994). Selected chelators with high
affinities for Fe(lll) such as citrate, nitrilotriacetic acid (NTA), humic acids,
ethylenediaminetetraacetic acid (EDTA), and biologically-produced ligands
(siderophores) have been examined, and have shown the ability to solubi 1 ize Fe(lll)-
oxides (Jones et al., 1 983; Arnold et al., 1986; Lovley et al., 1994; Haselwandter,
1995; Drechsei et al., 1995).
Chelate ligands act by binding at more than one site to the metal atom of the
ferric oxide and create a more stable complex which exhibits a higher water
soiubility over the metal oxide alone. Figure 7.1 illustrates the pentagonal
bipyramidal chemical structure formed when EDTA binds with a representative
ferric hydroxide.
Figure 7.1 Three-dimensional chemical structure of Fe(OH,)(EDTA). (Source: McArd ie, 1 98 1 )
Experirnents were desigiied to investigate and evaluate the performance of
EDTA on the physiochemical chelation of ferric oxides, and its ability to accelerate
the reduction of Fe(lll) associated with monoaromatic hydrocarbon oxidation. The
results of this work are compared to literature data. This study was also the first to
examine whether EDTA-arnended microcosms with PAWs as the sole carbon source
could be applied to aid in the remediation of these contaminants.
7.2 Experirnental
Multiple experiments were prepared to determine the effectiveness of iron
chelation to aid in PAH degradation.
7.2.7 Effect of EDTA on Fe Dissolution
A concentration range of EDTA (O - 1 mole ratio EDTA:Fe in increments of
0.2) was mixed with 0.5 g/L Fe as ferrihydrite or hernatite and diluted with distilled
water to 100 mL in 250-mL erlenmeyer flasks. The contents were mixed at 180 rpm
on a platform shaker (New Brunswick Instruments) for 48 hours. The soluble Fe(ll)
concentration was then determined.
7.2.2 Effect of EDTA on Phenanthrene Degradation
Within 125-mL amber glas bottles with silicondteflon septa, the following
ingredients were added per liter: 0.5 g yeast extract, 1.6 g Fe (as FeCI,-6H,O), 50
mg phenanthrene, 100 mL inoculum, and EDTA added to provide a range of mole
ratios to total iron of O to 1 in incrernents of 0.1. The contents were neutralized
with 1 M NaOH and brought to 125 mL with defined growth medium (Table 4.1).
All of the bottles were inoculated with QU1 consortium. Mercuric sulfate was
added to one microcosm at a final concentration of 1 Oh to create an abiotic control.
The contents were continuously rnixed on a Modular Cell Production Roller
Apparatus (Wheaton Instruments) at 100% motor speed and incubated at room
temperature (2S°C f 3). Samples were taken weekly, centrifuged at 9000 rpm for 1
minute, and the supernatant analyzed for soluble Fe(ll).
The experirnent was repeated with toluene, naphthalene, phenanthrene, and
anthracene, each supplied at 50 ppm, without yeast extract. EDTA was added to
each microcosm (including one without any hydrocarbon) at a mole ratio of 0.5. A
parallel experiment with '"[C]toluene was run to evaluate its mineralization in the
presence of EDTA.
A second microcosm was prepared with 80 mg EDTA, naphthalene (20
ppm), 1 .O g Fe as ferrihydrite, 50 mL defined growth medium (Table 4.11,
inoculated with 10 mL QU, and diluted with distilled water to 100 mL.
7.3 Results and Discussion
The effects of EDTA to enhance solubilization of Fe(lll) and to stimulate
mono and polyaromatic hydrocarbon degradation was investigated.
7.3.1 Effect of EDTA on Fe Dissolution
The ability of EDTA to solubilize ferric oxides of different crystallinity is
shown in Figure 7.2. Chelation of ferrihydrite resulted in a linear increase in
soluble Fe(lll) concentration over the range of EDTA studied. Hematite dissolution
was linear until an EDTA:Fe ratio of 0.3, after which it exhibited signs consistent
with saturation. Complexation of the metal species was dependent upon the
mineral identity. Dissolution of ferrihydrite was 15% greater than for hematite at
the same EDTA:Fe mole ratio of 1 .O. This trend agrees with Borggaard (1 982) who
found that EDTA extracts amorphous Fe-oxides better than crystalline forms.
7.3.2 E DTA-En hanced Hydrocarbon Oxidation
A range of EDTA to iron ratios were used in microcosms to determine the
83
Ferrihydrite
O O .2 O .4 0.6 0.8 1
Mole Ratio E DTkFe
Figure 7.2 EDTA-en hanced dissolution of ferri hydrite and hematite.
optimum concentration for acceleration of ferric oxide dissolution coupled to
phenanthrene degradation. Monitoring of p H and accumulation of soluble ferrous
iron was performed periodically over the 33day experiment. Figure 7.3 shows the
change in Fe+2 over this duration, as well as the average pH, which was found to
remain relatively constant throughout the experiment.
Mole Ratio EDTA:Fe
Figure 7.3 Stimulation of ferrihydrite redudion by EDTA.
The first observation that can be noticed is that iron reduction is greatly
enhanced in the presence of EDTA. The greatest extent of reduction was obtained
at a mole ratio of EDTA to iron of 0.4. When cornpared to no EDTA, this
represented an enhancement of over 21 times. An equivalent increase in the iron
reduction rate was observed by Arnold et al. (1 986) for the oxidation of lactate with
a 1 : 1 equimolar concentration of NTA to total iron. EDTA stimulation has been
found to be similar with NTA, but was slightly dependent upon the mineral being
chelated (Arnold et al., 1988; Lovley et al., 1994). The absence of ferrous iron
production in sterilized microcosms proves that the observed reductions were the
result of biological activity.
The accumulation of small quantities of ferrous iron in the microcosm
without EDTA is possibly explained by observations made shortly following
inoculation. The appearance, in this microcosm, of a black precipitate in
conjunction with the production of a rotten-egg odour is characteristic of iron
sulfide generation (Lovley, 1991; Stumm and Morgan, 1981). The insoluble iron
sulphides arise from the biological production of H,S, due to sulphatereducing
bacteria (SRB), which subsequently causes the abiotic reduction of ferric oxides
(Chiorse, 1 988).
The determining factor for which inorganic compound is reduced appears to
be the degree of bioavailability of the iron oxide. Others have also found that the
addition of Fe(lll) oxides to sediment in which sulfate reduction or methanogenesis
was the predominant electron-accepting process resulted in a 50-100% inhibition of
these processes, and depended upon the sediments and type of Fe(lll) added (Ell is-
Evans and Lemon, 1989; King, I W O ; Lovley and Phillips, 1987). From these
results, it was established that the QU1 consortium has the dual ability to reduce
both sulphate, or iron when presented as a soluble species.
The sharp decline in iron-reducing capacity above an EDTA to iron mole
ratio of 0.4 is most Iikely due to a drop of approximately two pH units. It is
hypothesized that the buffering capacity of the medium was exceeded due to the
increased EDTA concentration. At a pH below five, it is reasonable to expect a
decline in microbial activity. The pH effect was examined independently and the
results are discussed in more detail in Chapter 8.
Although at fint, it would appear that phenanthrene was being degraded,
microcosms with EDTA but lacking phenanthrene showed a similar response in
Fe(ll) accumulation. The identification of the preferred carbon compound was
inconclusive due to another potential carbon source in yeast extrad. The
experiment was repeated without yeast extract and at a constant EDTA mole ratio of
0.5 with phenanthrene, toluene, anthracene, or naphthalene supplied at 50 ppn?
each. Figure 7.4 illustrates the change in soluble Fe(ll1 with time. Positive iron
redudion in the presence of EDTA alone demonstrates the susceptibility of this
compound to biodegradation by QU1 consortium. When other chelate ligands
were studied, NTA was found to resist degradation in the presence of toluene
(Lovley et al., 1994), but citratedegradation was prevalent (Lovley, 1991).
The growth curve for toluene was different enough from EDTA alone to
warrant a closer look. Of al1 of the organic compounds tested, toluene was the only
one which resulted in a prolonged lag period. Only after approximately two weeks
there was an increase in ferrous iron over its initial concentration. It was thought
that during this period the population of toluenedegrading microorganisms within
the QU1 consortium was increasing. This would mean that toluene would be the
preferred energy-yielding substrate over EDTA which is in agreement with Lovley et
al. (1 994) for NTA. Later mineralization experiments with QU 1 consortium and
toiuene with EDTA confirmed this.
EDTA
A Toluene
Naphthalene
x Phenanthrene
O Anthracene
A A a *
O 5 10 15 20 2 5 3 O 35 40
Tirne (days)
Figure 7.4 Ferric iron reduction attributed to the oxidation of seleaed carbon compounds amended with EDTA.
In the microcosrn with naphthalene, no ferric iron was reduced. This
suggests that at the concentration supplied (50 ppm), naphthalene was toxic to the
88
QUI consortium. This is in agreement with the toxic effect of higher
concentrations of naphthalene on QU2 (Section 5.3).
The results for phenanthrene and anthracene show similar growth curves to
that of EDTA alone. It was thought that perhaps EDTA was being consumed in
these microcosms rather than the PAH. Minera1 ization experiments were designed
to substantiate these cfaims. No mineral ization occurred in experirnents inoculated
with QU 1 when phenanthrene or anthracene were provided as the sole carbon
source.
Mineralization experiments conducted with QU2 and naphthalene at non-
toxic levels (20 ppm) confirmeci inhibition of PAH degradation in the presence of
EDTA (Figure 7.5). This represents additional evidence supporting the assertion that
EDTA was being consumed instead of phenanthrene and anthracene in the previous
study (Figure 7.4).
7.4 Conclusions
The reduction of Fe(lI 1) is often limited by the low solubility of natura .I ferric
oxides, and this can control the kinetics of hydrocarbon degradation. Previous
investigators have found that organic ligands which cornplex metal species can
dramatically enhance their dissolution and therefore the rate at which organic
compounds can be anaerobically degraded. This work focussed on the feasibilty of
using chelaton to accelerate the in situ remediation of hydrocarbon contaminants
coupled to dissimilatory iron reduction. It was found that EDTA greatly increased
O no ammendment
0 80 mg EDTA
A Abiotic Control
Figure 7.5 Inhibition of naphthalene mineralization by EDTA.
the soluble fraction of Fe(l l l), and that its efficacy was dependent upon the rnineral.
More ferric oxides were solubilized by EDTA when present as an amorphous versus
a crystalline form.
Ferric oxide reduction was improved when amended with EDTA, but above
a mole ratio of 0.5 it exceeded the buffering capacity of the medium, resulting in a
large drop in pH. The observed decline in microbial activity was found to result
from the acidic conditions in the growth medium. Enhanced hydrocarbon
degradation in the presence of soluble iron occurred for toluene, but not for any of
the PAHs studied. Results indicated that in the presence of PAHs, both the QU 1
and QU2 consortiums preferentially degradeci EDTA as carbon source. The
difference in the water solubilities between toluene and the PAHs tested is expected
to be responsible for the observed substrate preferences. Hence synthetic chelators
(EDTA) were omitted from further studies.
When presented as soluble Fe(lll), QU1 consortium was able to reduce iron.
However under iron-limiting conditions sulfate reduction was evident.
Chapter 8 Environmental Factors
8.1 Introduction
In designing an in situ bioremediation scheme the kinetic dependence of
contaminant degradation on environmental conditions must be considered. Such
factors include temperature, pH, geochemical conditions, culture density, and
nutrients. Especially in the Canadian environment, these conditions can Vary
considerably. One example is the potentially significant difference in the nutrient
profile between marine and fresh water environments. Also, the seasonal variation
in temperature may impact both the survival and activity of indigenous
rnicroorganisms. The interpretation of data to be used for modelling must include a
contribution from environmental factors to provide accurate predictions.
8.2 Experimental
To evaluate the effects of temperature, pH, ferric oxide crystal linity,
inoculum percentage, and nutrient profile on the mineralization of PAHs, 30
microcosms were prepared. One abiotic, and duplicate biotic microcosms were set
up for each of the parameters tested, and an identical set of microcosrns were run in
paral le1 to accommodate simultaneous mineral ization, and cornplementary iron
analyses. A solution composed of 20 ppm naphthalene, 1 .O g/L Fe as ferrihydrite,
50 mL anaerobic culture medium, inoculated with 10 mL QU2 consortium and
brought to 100 mL with distilled water constitueci the common growth slurry.
Microcosms were incubated at room temperature with no pH control except that
provided by the buffering capacity of the medium. An independent assessrnent of
each environmental parameter was performed by manipulating one corn ponent and
keeping the othen constant to create the five experimental systems.
The following parameters were evaluated: 1 ) Temperature - (1 0, 20, and
30°C); 2) pH - bi-weekly adjustrnents to 4.0, 6.0, or 8.0 f 0.25 or periodic spiking
were conducted with a minimal volume of NaOH or HCI; 3) Ferric lron
Crystallinity - the presence or absence of ferrihydrite or hematite at equivalent Fe
concentrations; 4) lnoculum Size - one, two and three times the inoculum volume
were used to study the influence of cell density on naphthalene degradation; 5)
Nutrients - to evaluate the nutritional requirements of the QU2 consortium, the
anaerobic culture medium was replaced with lake water.
8.3 Results and Discussion
8.3.1 Temperature
To examine the influence that temperature might have on the natural
attenuation of PAHs, microcosms were prepared and incubated at 10, 20, and 30°C.
A cornparison of the initial rates of degradation reveals an apparent temperature-
dependent teaction (Figure 8.1). At 1 O°C, mineralization proceeded, but only after
a threeday acclimation period and only half as fast as in microcosms at 30°C.
However, the extent of temperaturedependence diminishes after one week's
incubation. At the completion of the two-and-a-half month experiment, 4 7 O / 0 of the
O 20 deg C
30 deg C
O 10 20 30 40 50 60 70 80
Tirne (Days)
Figure 8.1 Temperaturedependence of naphthalene mineralkation.
original naphthalene was recovered as 14C0,, which represents al most 75 */O as
much as in the 30°C microcosm. The variability was low between duplicates, with
an average variation of only 3%.
Microbial adivity at 20°C and 30°C was immediate, with degradation rates
differing by only 2O0h, in favour of the warmer incubation temperature. At the
94
completion of the test, 57% and 64% of the naphthalene was mineralized at 20°C
and 30°C, respectively; a difference of only 1 0%. Once more, the dupl icate
microcosms were in good agreement, with variabilities at 20°C and 30°C of 5Of0 and
2 O/& respect ive1 y.
The obvious conclusion that can be drawn from this data is that although
metabolic activity is faster at a warmer temperature, the QU2 consortium has the
ability to adapt to more inclernent temperatures and therefore maintain reasonably
high degradation rates. In cornparison, the iron-reducing isulate, GS-15 has a
temperature optimum between 30-35OC, with no detectable activity at greater t han
50°C or less than 1 O°C (Lovley and Phil l ips, 1 988). Converse1 y, the accumulation
of Fe(ll) over a 2-month period was observed at an in situ temperature of only 9°C
(Lovley et al., 1989a). The transition from biokinetic to rnass-transfer control as
evidenced in prior substrate bioavailability experiments may be partially responsible
for the modest long-terrn influence of temperature on m ineral kat ion. The
implications of this is that for in situ remediation approaches with this consortium,
moderate fluctuations in temperature will only have a slight impact on the rate of
degradation.
The effect of pH on the mineralkation of PAHs was investigated in two
ways; using either intermittent or continuous pH adjustment. When pH was
allowed to fluctuate, it increased from an average initial value of 7.9 to 8.3 over the
fint two weeks, consistent with an accumulation of 14C02 from the mineralization
of phenanthrene (Figure 8.2). Elevated levels of pH was also observed in other
cases where iron-redudion was the predominant redox process (Aller et al., 1 986;
Lovley, 1990).
Afterward, the pH remained relatively constant and the rate of substrate
pH adjusted
no pH adjustment
% Mineralization
60 80 1 00
Time (Days)
Figure 8.2 pH spiking and its effea on the mineralization of phenanthrene.
nietabolism declined. Microbial inhibition was thought to arise when a basic
environment was created within the sluny microcosm. Al-Bashir et al. (1 990) also
witnessed a reduction in the rate of PAH mineralization coinciding with a rise in
pH, and attributed the former to the latter. Consequently, at six and nine weeks of
incubation, the pH in one of the duplicate rnicrocosms was adjusted with 1 N HCI
and not in the other. With a decrease in pH to 5.7, there was an accompanying
five-fold increase in the rate of mineral ization. lmmediately foi lowing the decrease
in pH,it rebounded sharpiy until day 63, when once again the microcosm was
acidified. The pH dropped to 5.2, but rather than increasing once again,
mineralization remained unchanged. The large fluctuations in pH over a short
period may have been lethal to the bacterial population. pH adjustments were
similarly done to the abiotic control, which showed no change in abiotic
naphthaiene loss. Intermittent pH adjustments were similarly done with alternative
carbon sources (Figure 8.3).
An explanation for the observed occurrence could be related to the
enhanced enzymatic reduction of ferric oxides when provided in a form accessible
to the degrading consortium Stumm and Wieland (1 991) found that pH affected the
surface charge of iron oxides, with more acidic conditions favouring increased
dissolution. If iron reduction represented the rate-l imiting step of the redox process,
augmentation of organics oxidation would occur following an increase in the
soluble iron. Arnold et al. (1 986) stated that when provided in the form of colloidal
iron hydroxides, the concentration of soluble Fe" was solely a function of pH.
Another possibility is that the metabolic end products of PAH degradation,
such as, formate, oxalate, citrate, pynivate, or other organic acids, or of sulphate
reduction (H,S ), abiotically reduced ferric iron. It has been dernonstrateci in the
literature that under anaerobic conditions, the abiotic redudion of Fe by microbial
Figure 8.3 Cornparison of mineralization rates prior to, and following pH adjustrnent for various carbon compounds. Note: Mineraiization rates for toluene, naphthalene, and anthracene have been scated to fit.
metabol ites is en hanced in acidic environments (Chiorse, 1 988; Jauregui and
Reisenauer, 1982), but not at circumneutral pH (Lovley et al., 1989b). This would
explain the increase in mineralization with HCI addition. The lack of response to
pH adjustment in the abiotic control, where rnetabolic by-products would be
absent, is consistent with this theory. Furtherrrtore, the lack of enhancement during
the second pH adjustment would imply that: 1) the capacity of the metabolites to
nonenzymatically reduce iroo has been exhausted, and 2) at a pH of around 5.0,
microbial activity is inhibited.
Experiments in which pH was continuously monitored and adjusted to
produce a quasi steady-state pH at 4, 6, or 8 resulted in similar kinetics of
naphthalene mineralization at pH 6 and 8 (Figure 8.4). This is in agreement with
Arnold et al. (1 986) who found modest variation in the mineralization rates of
lactate over a pH span from 6.5 to 7.5.
At a pH of 4, however, no biological rnineralization was detected. It is not
unreasonable to expect that the acidic environment would be too harsh for the
consortium's survival. The discrepancy between the results found for the
experiments with intermittent and continuous pH adjustment is consistent with the
findings of previous work. Arnold et al. (1 988) and Kostka and Nealson (1 995) both
found that dissimilatory reduction of ferric iron occur in the pH range of 5.0 - 6.0.
For example, Fe(lil)-redudion of rnagnetite was found to be thermodynamically
favourable only within this region (Kostka and Nealson, 1995). Below a pH of 5.0,
the conditions were too acidic to support microbial adivity.
O 1 O 20 30 40 50 60 70 80
Time (Days)
Figure 8.4 Influence of pH on naphthalene mineralization.
8.3.3 Ferric Oxide Crystallinity, lnoculum Percentage, Nutrients
In addition to temperature and pH, the significance of iron oxide
cn/stall inity, percent inoculum, and nutrient amendment on naphthalene
degradation were investigated. Slurry microcosms amended with ferrhydrite or
hematite were compared for their effect on the mineralization of naphthalene.
1 O0
When cornpared to a control without amendment, no notable difference was
observed. B a d on the abundance of work which has demonstrated a strong
correlation between iron oxide crystallinity and reduction, these results were
unexpected. Upon closer examination of the inoculum source however, it was
determined that the 10°h inoculum contained sufficient ferric iron for the
stoichiometric mineralization of 20 ppm naphthalene. A large excess of ferrihydrite
was added to the medium used for the culture development and subsequently
transferred with the consortium upon inoculation. Also, if the rate4 imiting factor is
carbon substrate bioavailability as presumed from the results shown in Chapter 5,
the effect of iron crystal linity/dissolution would not be observable.
An experiment was undertaken to investigate the response of naphthalene
mineralization to inoculum percentage. Slurry inoculations of 10, 20, and 30°h
resulted in sirnilar mineralization kinetic profiles. lnoculum size would result in
changes in cell density and solids content for substrate adsorption, each opposing
interactions. To address the first component, the absence of a lag period, even at
1 0% inoculation, indicates a sufficient initial population of naphthalenedegraders,
which would account for its lack of influence. Even with low cell numbers, as
Arnold et al. found (1 9881, the magnitude of iron reduction can be significant.
Additional sediment also represents an increase in surface area and organic
content which funaion as adsorption sites for naphthalene, thus limiting its
availability for microbial degradation. However because naphthalene was supplied
at levels below its solubility, minimal sediment was sufficient to establish a
soikwater partition equil ibrium. Further additions of sediment would not be
expected to alter this ratio.
Laboratory studies were conducted to model the impact of nutrient
amendment on in situ ?AH bioremediation. Naphthalene mineralization in
microcosrns with anaerobic culture medium was no different than when incubated
in lake water. Basic microbial maintenance and growth have minimal nutrient
requirements which are apparently k i n g met through the inoculum source. The
conclusion that can be drawn from this is that nutrients supplied from both the
defined medium (Table 4.1) within the inoculum and in the sediment itself are
adequate for the mineralization of at least 20 ppm naphthalene. Anderson et al.
(1 9821 stated a low nutrient requirement as one of the advantages of anaerobic over
aerobic treatment systems. A higher metabolizable organic content might result in
nutrient-l imitation and require amendment for degradation to continue.
8.4 Conclusions
The influence of the following environmental factors on the feasibility of in
situ PAH bioremediation was assessed: temperature, pH, indigenous microbial
density, iron crystallinity, and nutrients. The results found that under the conditions
studied, naphthalene m ineral ization is relatively independent of each of these
factors.
The implication of this work is that intrinsic degradation of PAHs would be
effective under these environmental conditions. Despite this, the principal
goveming force for natural attenuation is the redox environment. Swn after the
introduction of organic compounds into subsurface formations anoxic conditions
devefop. Both the natural presence of alternative electron acceptors and the ability
of indigenous microorganisms to reduce them will determine the continuing value
of intrinsic bioremediation. Accordingly studies were designed to test the
applicability of anaerobic carbon oxidation coupled to alternative redox systems,
specifically iron reduction. The results are discussed in the next chapter.
Chapter 9 Redox State
9.1 Introduction
The existence of microbial species with the capacity to couple the
degradation of organ ic contaminants to the reduction of al ternative electron
accepton has considerable ecological importance. Considering that most soil-water
systems become anoxic soon alter contamination with organics (Evans and Fuchs,
1988; Mihelcic and Luthy, 1988b3, the metabolic fate of these compounds depends
on the availability of suitable alternative electron acceptors, and microorganisms
that can use them. O,, NO,; SO;', Fe+), and CO, have been demonstrated as
effective electron acceptors.
The degradation of PAHs has been shown to occur under aerobic (Cerniglia,
1993; Parks et al., 1990; Voikering et ai., 1992), nitrate reducing (Al-Bashir et al.,
1 990; Mihelcic and Luthy, 1988a), and sulphate reducing (Coates et al., 1996a)
conditions, but never in iron-reducing or methanogenic environments. The
concentrations of these potential electron accepton were monitored during
previous experiments (Chapters 5 through 8) to assess the ability of the QU1 and
QU2 consortia to couple their reduction to the degradation of aromatic
hydrocarbons, especially PAHs. Because iron reduction was the focus of this work,
the concentrations of ferric and ferrous iron were foilowed more thoroughly than for
nitrate and su1 phate, which were determineci for the mineral ization experiments
inoculated with QU2 only.
The objective of this chapter was to identify which electron acceptor(s)
wadwere responsible for the degradation of the aromatic hydrocarbons, particularly
PAHs, during earlier experiments.
9.2 Experimental
Ferric and ferrous iron concentrations were determined in each experiment
as previously described (Chapter 3). Slurry samples from the mineralkation
experirnents were filtered through 0.45 Fm, polycarbonate filters, and the fiitrate
analyzed for nitrate, nitrite, phosphate, sulphate, and chloride concentrations by ion
ch romatograp hy.
9.3 Results and Discussion
An evaluation of each potential electron acceptor was done to determine the
active redox system in each microcosm, and the capabilities of QU1 and QU2 for
utilking potential inorganic electron acceptors.
9.3.1 Oxygen
Since the objective of this work was to examine PAH degradation under
anaerobic conditions, steps were taken to remove O, from the medium and
headspace in vials used for inoculum development, and the ensuing microcosms.
High purity nitrogen was bubbled through the media at a rate and duration
dependent upon the volume of liquid present. The efficiency of this process for the
1 OS
removal of 0, was gauged on the change in dissolved oxygen and redox potential
prior to and following sparging. Figure 9.1 illustrates the dissolved oxygen (DO)
concentration with time in a typical sparging routine. The rapid redudion in
2 3
Time (Minutes)
Figure 9.1 Efficiency of nitrate sparging for the removal of dissolved oxygen.
oxygen concentration over the fint 2 minutes is a good indication that the
microcosms prepared in previous studies were anoxic following sparging. A
reduction in the DO was also accompanied by a decrease in the redox potential
from + 178 mV to + 73 mV. Additional measures that were taken to produce
1 O6
anaerobic conditions included: minimizing headspace where possible, and sealing
vials to create a closed system.
If it is assumed that the oxygen in the gas phase of microcosms was
completely removed during sparging, but that the displacement of dissolved 0, was
ineffecîive, only 1 5% mineralization of naphthalene would be theoretically
possible. On the contrary, the sparging process was shown to be very effective at
removing dissolved oxygen (93% reduction) and, as shown in Chapter 5, as much
as 55% naphthalene was mineralized. It is therefore very probable that aerobic
degradation was not the predominant elearon accepting process.
9.3.2 Nitrate
Two pieces of evidence can be used to defend the argument that nitrate
reduction was not responsible for the observed mineral ization of PAHs. The first is
that nitrate was not provided in any form in the defined mineral salts medium
(Table 4.1 ), nor was it present at an appreciable concentration in the distilled water
(0.004 mM) or lake water (0.01 mM) used to prepare the microcosms. The only
remaining source of nitrate is in the sediment used for inoculation. Although
analyses showed that the sediment contributed enough nitrate to rnineralize al1 of
the added carbon, its concentration never decreased with incubation time nor did
nitrite increase in samples taken from the microcosms.
9.3.3 Sulphate
Unlike nitrate, sulphate was provided in the mineral salts medium, in varying
forms, at a concentration of approxirnately 0.4 mM. Its use as an electron acceptor
was found to Vary depending on the inuculurn used. During experiments to
determine the effect of EDTA on ferric iron redudion coupled to phenanthrene
degradation (Chapter 7)) several signs consistent with sulphate redudion were
apparent. The microcosms, inoculated with QUI, contained a range of EDTA at
mole ratios to iron of between O and 1 (Figure 7.3). In the presence of EDTA, as
much as 425 &mL Fe" accurnulated over the 33day incubation period. However
no ferric iron reduction occurred when it was not present in a bioavailable form.
Instead, a rottenegg odour and a black precipitate were generated, both
characteristics of hydrogen sulfide production. H,S is produced as a by-product of
sulphate reduction, catalyzed by sulphate reducing bacteria (SRB). Once formed,
sulphide can chemically reduce Fe+) and complex with it to form insoluble iron
sulphides (Jauregui and Reisenauer, 1 982; Lovley and Phillips, 1 986a; Pyzik and
Sommer, 1 98 1 ).
Two conclusions can be drawn from these results. First, since sulphate
reduction requires anaerobic conditions (G hiorse, 1 9881, purging techniques must
have been successful at removing the majority of gaseous and dissolved oxygen,
which supports the assumption that PAH degradation was not aerobic. Also,
because EDTA was not added to this microcosm, phenanthrene was the sole carbon
substrate. Indications of sulphate reduction would imply that QU1 is able to
degrade PAHs (phenanthrene) with sulphate as the terminal efectron acceptor.
However, su1 phate reduction is improbable in microcosms inoculated with
QU2 consortium for two reasons. The concentration of sulphate was monitored in
microcosms used for mineralization studies, and showed no decrease over the
duration of the experiment. The strongest evidence, however, cornes from the
results of spiking an active microcosrn with sodium molybdate, a specific inhibitor
of sulphate redudion (Oremland and Capone, 1988). If sulphate reduction was the
predominant redox process, the addition of sodium molybdate would inhibit PAH
oxidation. However, no variation in the rate of mineralization resulted in this case,
another indication that there was no sulphate reduction.
9.3.4 lron
Because the aim of this work was to demonstrate that PAHs could be
degraded with iron as the electron acceptor, ferric iron was provided in abundance
in each experiment, and other potential electron acceptors minimized where
possible. However, although the iron was added at a stoichiometric excess, due to
its poor bioavailability, it could still be limiting. To address this concern, studies to
examine the influence of iron oxide crystallinity and l igand-amendment on iron
reduction were conducted. The results of these experiments and others showed that
iron reduction was contingent upon the consortium used for inoculation.
9.3.4.1 OU 1 Consortium
The ability of QUI to use ferric iron as the terminal electron acceptor was
clearly established following experirnents with EDTA (Chapter 7). In the absence of
soluble Fe(lll), iron reduction, was negligible (Figure 7.3). However when chelated
with EDTA, considerable accumulation of ferrous iron was observed. It was
subsequently discovered that EDTA was apparently consurned as the preferential
carbon source rather than the PAHs. This contrasted the results for toluene which,
through mineralization experirnents, was confirmed to be degraded with
concomitant ferric iron redudion (Figure 7.4).
The discrepancy in preferential carbon consumption between monoaromatic
and poiycyclic aromatic hydrocarbons in EDTA-amended microcosms presumably
illustrates the dependence of QU 1 on the bioavailability of the potential substrates.
Since toluene is much more soluble than the PAHs studied, i t is probably
rnetabolized preferentially to EDTA. Due to the limited bioavailabi lity of PAHs,
however, QU 1 can fulfil 1 its rnetabolic requirements more easi ly through the
degradation of EDTA.
9.3.4.2 OU2 Consortium
Prelirninary experiments demonstrated the inhibitory effect of EDTA on
naphthalene mineralization (Figure 7.5), and hence eliminated its use in al1
subsequent investigations. lron analyses were done on samples obtained from the
rnicrocosms used in the mineralization experiments to determine if it was being
reduced. Although there was some variability in the concentrations of ferrous iron,
110
no observable trend could be established. This would seem to indicate that iron
was not being used as the terminal eledron accepter. However, a cornpelling
observation, made in the 4 4 bottle used to develop the QU2 consortium, indicates
otherwise.
After several weeks of mixing, the slurry used to inoculate the microcosms in
the mineralization experirnents was allowed to settle. Shortly thereafter, the
aqueous layer was observed to develop a hue which progressed from yellow to
orange. As well, the slow accumulation of an orange coloured precipitate at the
sediment-water interface was evident. Of al1 of the possible ferric or ferrous
complexes, lepidocrocite (y-FeOOH) is characterized as having the closest colour to
the observed precipitate (Heron et al., 1 994; Schwertmann and Taylor, 1 989).
Schwertmann and Taylor (1 989) went further to say that soi1 lepidocrocites
generally have the same morphology as the oxidation products of FeCI, solutions at
ambient conditions. Since FeCI, was hydrolyzed directly in the medium to form
colloidal ferrihydrite, sufficiently high concentrations of CI- would be available to
complex with any Fe(ll) produced. A similarly produced abiotic control, sterilized
with a 0.296 solution of sodium azide, did not display this behaviour. A
measurernent of the redox potential at the top of the water column and within the
surficial sediment produced Eh values of + 50 mV, and between -6 and -1 6 mV,
respectively. Based on these observations, and the results of the previous nitrate
and sulphate analyses, a biogeochemical iron cycling phenomenon was proposed
(Figure 9.2).
PAH QU2 Consortium
Figure 9.2 lron biogeochernical cycling. A: development of orange hue in aqueous layer and orange precipitate settled to sediment-water interface; B: abiotic control for corn parison; C: schematic representation of proposed biogeochemicai iron cycle (Adapted from: Nealson and Myers, 1992).
Because fenihydrite is insoluble at neutral pH and has a propensity to adsorb
ont0 solid matrices, it would primarily reside on the surface of the sediment.
Li kewise, PAHs and many microorganisms adsorb ont0 sol id surfaces. The
proposeci mechanism links the rnineralization of the PAHs to the transfer of
electrons to fenihydrite in the sediment phase, producing ferrous iron, CO,, and
perhaps some deadend products. The solubility of Fe(l1) is much higher than ferric
iron, and hence it diffuses into the overlying water body (Ponnarnperuma, 1972).
The insoluble ferric oxides are regenerated in the aqueous phase, and precipitates
ont0 the sediment-water interface. This could explain the absence of Fe(ll)
accumulation in the iron analyses, and account for the observed colour change in
the aqueous phase and orange precipitate formation at the surface of the sediment.
The mystery that remains is the mechanism responsible for the regeneration
of Fe(l il). Although there are references towards biogeochernical iron cycl ing
(Nealson and Myers, 1992; Lovley, 1991 ; Schwertmann and Taylor, 1989),
oxidation occurs as a result of oxygen andor bacteria (e.g. Thiobacillus
ferrooxidans). However, it has been assumed that oxygen has been effectively
removed from the system. If this assumption is false, an anoxic zone in the
sediment and an oxic zone in the water column could exist. Fe(1l) i s
thermodynamically unstable in an aerobic environment (Heron et al., 1994, and
would regenerate ferric oxide.
Another possibility is that an abiotic chemical reaction in the aqueous phase
is responsible for the regeneration. Species which have been shown to
nonenzymatical l y oxidize ferrous iron incl ude: manganese (Nealson and Myers,
113
1 9921, nitrite (krensen, 1 982), and chromium. However, these species are either
not soluble, or not anticipated to be at significant concentrations in the slurry.
Known processes which abiotically convert ferrous to ferric iron lends support for
other similar, but as yet undiscovered, mechanisms which could play a role in the
proposed biogeochemical iron cycle.
9.4 Implications
Both QU 1 and QU2 were capable of utilizing ferric oxides for the
degradation of hydrocarbon contaminants. The QU1 consortium was only able to
reduce ferric iron when it was supplied in a bioavailable form, and as such could
degrade the more soluble monoaromatic hydrocarbons only and not the PAHs.
On the other hand, QU2 substantially mineralized some PAHs with the
apparent reduction of ferric oxides. Also, the possible regeneration of Fe(lll)
provides a scenario in which iron would not become limiting. The absence of
synthetic chelators demonstrates the competency of QU2 to utilize the insoluble
ferric oxides as terminal electron accepter. The characteristics of this indigenous
population would be beneficial for the intrinsic remediation of low molecular
weight PAHs.
Chapter 10 Conclusions and Recommendations
A novel method for the intrinsic remediation of PAHs coupled to ferric iron
reduction was examined. This technique brings together the complexities inherent
with each part, and thus requires an appreciation for the multidirnensional
approach required for its investigation.
Enrichment techniques were used to develop two indigenous mixed cultures,
QU1 and QU2, which demonstrated PAH degradation and iron reduction. No
effort was directed towards the undentanding of their taxonomie classification,
genetics, or microbial interactions of these consortia, and these studies should be
incorporated into future work. Although it has been established that mixed
populations of Fe reducers rnay provide an advantage over pure cultures in natural
ecosystems, isolates would allow for an easier interpretation of the resulting
kinetics. This step should be incorporated into the next phase of work, and include
a corn parison between the mixed versus pure CU ltures ta determ ine their relative
val ue.
Mineralization studies to identify the aromatic hydrocarbons amenable to
degradation by QU2 revealed that the concentration of lower molecular weight
PAHs, and monoaromatic hydrocarbons following a short acclirnation period, were
significantly reduced. Linear mineralization rates for naphthalene and
phenanthrene after approximately ten weeks of incubation implies that degradation
was mass-transfer l imited. The accumulation of sinal l amounts of '%O,
immediately following inoculation, provided an indication that higher rnolecular
weight PAHs could be mineral ized, but that their lirnited bioavailability after
adsorption ont0 the sediment was insufficient to maintain the basic microbial
metabolic requirements of QU2.
The inabiiity of QU2 to mineralize fluoranthene, and the knowledge that
some species can do so cornetabolically, brought to Iight the importance of
substrate mixtures. Due to the Iimited scope of this work this was not investigated
further, bot should be in subsequent studies.
Synthetic surfactants, Brij35 and Triton x-100, used at concentrations above
their critical micelle concentration (CM0 were inhibitory to QUI. Surfactant
amendment could be further evaluated at concentratisns both below and at the
CMC.
After two weeks of incubation, an increase in soluble naphthalene and
phenanthrene with QU1 compared to an abiotic control and biotic microcosms
inoculated with other inocula indicated that QU 1 could manufacture a
biosurfactant, presumably in response to limiting substrate bioavailability. A greater
understanding of this process might enable its use to enhance PAH degradation.
Experirnents designed to test the physiochemical response of various
cyclodextrin derivatives revealed a considerable enhancement in the solubil ization,
desorption, and rnobilization of a range of PAHs. Additionally, it was shown for the
first time that cyclodextrin could provide a modest benefit to the anaerobic
biodegradation of phenanthrene. At higher concentrations, however, the ultimate
inhibition of PAH mineralization indicated preferential consumption of
hydroxypropyl~cldextr in (HPCD). The substantial improvement in the
solubi l ization, desorption, and mobil ization of high molecular weight PAHs
amended with cyclodextrin offers encouragement towards the potential of
cyclodextrins to greatly enhance their degradation. If this technique is to be
considered further, the possi bi 1 ity of recycl ing cyclodextrin must be addressed to
maintain its cost-effectiveness.
Because the enzymatic reduction of fenic oxides can be rate-limiting in
many systems, chelation with organic ligands was examined to increase their
bioavailability. The benefits of chelation were found to depend both on the mineral
crystal l in ity and the hydrocarbon compound being degraded. EDTA was SI ightl y
more effective at solubilizing ferrihydrite than the more crystalline ferric oxide,
hematite. EDTA also resulted in faster rates of biologically rnediated iron reduction
supplied with toluene, but was consumed preferentially to PAHs.
The influence of in situ environmental conditions was considered by varying
temperature, pH, iron oxide amendments, inoculation percentage, and nutrien t
supplements. The results showed that naphthalene mineralization was relatively
independent of in situ conditions within the range tested.
An assessrnent of the ability of QU 1 and QU2 to utilize various inorganic
electron acceptors concluded that hydrocarbon degradation was most likely
coupled to ferric iron reduction. QU 1 could metabolize toluene with concomitant
Fe(lll) reduction, but only when the fenic oxides were bioavailable. PAHs were not
degraded, due most likely to the preferential consumption of EDTA.
Observations made in slurry systems with QU2 were consistent with the
presence of a biogeochemical iron cycle coupled to PAH mineralization. The
proposed mechanism involved the degadation of PAHs coupled to ferric oxide
reduction, with the corresponding diffusion of Fe(ll) into the aqueous phase where it
was oxidized, regenerating an insoluble ferric oxide that precipitated ont0 the
sediment-water interface. The mechan ism responsible for the regeneration of Fe(l Il)
is unknown, the identification of which could be the focus of subsequent work.
The implications of these findings are significant and indicative of potentially
considerable intrinsic remediation of PAHs under iron-reducing conditions. The
results of this work should be incorporated into the design of field investigations to
substantiate the above daims. Additional evidence from the field which supports
the laboratory data will reinforce the industrial significance and potential of this
novel remedial approach.
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