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Trends in Bioremediation and Phytoremediation, 2010: 23-54 ISBN: 978-81-308-0424-8
Editors: Grażyna Płaza
3. Principles of bioremediation processes
Artin Hatzikioseyian
National Technical University of Athens (NTUA), School of Mining and Metallurgical Engineering, Division of
Metallurgy and Materials Engineering, Laboratory of Environmental Science and Engineering
Heroon Polytechniou 9, 15780 Zografou, Athens, Greece
Abstract. The field of bioremediation has experienced a dynamic evolution and
remarkable development over the past decades. The bioremediation of
contaminated former industrial sites or waste deposits became a well established
field of environmental biotechnology. This chapter gives an overview of
engineered bioremediation treatment options and focuses mainly on the biological
aspects of biostimulation, bioaugmentation, monitored natural attenuation
(MNA), and biotransformations of metals, metalloids and radionuclides in
contaminated sites. Some aspects of phytoremediation are also presented.
Introduction
Bioremediation is concerned with the biological restoration and rehabilitation of
historically contaminated sites and with the cleanup of areas contaminated in more recent
times, either accidentally or incidentally, as a result of the production, storage, transport,
and use of organic and inorganic chemicals [1, 2]. Bioremediation offers the possibility
of degrading, removing, altering, immobilising, or otherwise detoxifying various
chemicals from the environment through the action of bacteria [3-5], fungi [3, 4] and
plants [5-11]. Most of the advances in bioremediation have been realised through the
assistance of the scientific areas of microbiology, biochemistry, molecular biology,
analytical chemistry, chemical and environmental engineering, among others. These
different fields, each with its own individual approach, have actively contributed to the
development of bioremediation progress in recent years [12].
Correspondence/Reprint request: Dr. Artin Hatzikioseyian, National Technical University of Athens (NTUA), School
of Mining and Metallurgical Engineering, Division of Metallurgy and Materials Engineering, Laboratory of Environmental
Science and Engineering, Heroon Polytechniou 9, 15780 Zografou, Athens, Greece. E-mail: artin@metal.ntua.gr
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Artin Hatzikioseyian
The systematic use of biological processes in the design of engineered environmental
remediation systems is today well developed and documented [1, 4, 12-44].
Contemporary environmental biotechnology utilises techniques that employ usually
natural strains of microorganisms in order to mobilise and remove organic or immobilize
inorganic pollutants from the environment. The key advantage of bioremediation
processes as compared to other biological technologies is that they can employ enzymatic
metabolic pathways that have been evolved in nature over very long periods of time, thus
becoming very specific. The combination of such pathways can make possible the
degradation of a wide variety of even hazardous pollutants. Today, metabolic pathways
for the degradation of compounds previously considered as non degradable have been
identified [15]. Laboratory and field techniques for assessing the applicability of
biological processes for the degradation of the pollutants that are present at specific sites
have become available [25, 45, 46].
Historically the engineering aspects of environmental biotechnology were perceived
as an extension of the conventional biological wastewater treatment processes. The idea
was to bring together microorganisms and the waste organic materials under conditions
that would enable the microorganisms to utilise the organic molecules as a substrate. The
corresponding process efficiency could then be measured by the use of an easy lumped
parameter such as BOD5 or COD and occasionally by the use of compound specific
analyses. Experience however quickly led to the realisation that systems dealing with
environmental contamination through bioremediation were much more complex, and that
moving from the flask to the field was not a simple task.
Today, the experience accumulated over the last decades, has improved our
understanding in many aspects of this multidisciplinary technology. However, technical
and non technical issues associated with large scale implementation of bioremediation
technology still exist. The non technical level of uncertainty associated with the
technology is the result of combination of legal, regulatory, social and financial
considerations. Non technical issues can at times override the potential technical
feasibility for a specific application. The combined evaluation of the technical and non
technical issues is an important step towards the successful application of environmental
biotechnology in remediation.
The state of the art of bioremediation technology as well as examples of more or less
successful case studies are published in many books during the last 20 years [14, 16, 19,
25, 26, 29, 31, 35, 37, 39-41, 43-45, 47]. Many books are specialized in specific
applications of bioremediation such as bioremediation of MTBE [48], perchlorate [49],
oil spills – hydrocarbons contamination [29, 39, 50], chlorinated solvents [51] metal ions
[27, 29, 34, 52-54], radionuclides [38, 55], explosives [56] in various environments, e.g.
marine harbours [57], cold areas [58-60] etc. The financial aspects of bioremediation is
also very important and documented [23, 44]. Finally, a novel area such as
phytoremediation has gained attention in the recent years [5-10, 29].
Currently many websites are also providing researchers with advice tools,
technologies and methods for assessing and cleaning up hazardous waste sites. The
hazardous waste clean-up information (CLU-IN) website (http://clu-in.org/) provides
information about innovative treatment and site characterization technologies to the
hazardous waste remediation community. It describes programs, organizations, publications, and other tools for federal and state personnel, consulting engineers, technology
developers and vendors, remediation contractors, researchers, community groups, and
individual citizens. The site was developed by the US Environmental Protection Agency
(EPA) but is intended as a forum for all waste remediation stakeholders. The Federal
Principles of bioremediation processes
25
Remediation Technology Roundtable (FRTR) provides the remediation technologies
screening matrix and reference guide version 4.0, which serves as "yellow pages" of
remediation technologies. It is intended to be used to screen and evaluate candidate
cleanup technologies for contaminated installations and waste sites in order to assist
remedial project managers (RPMs) in selecting a remedial alternative. To reduce data
collection efforts and to focus the remedial evaluation steps, information on widely used
and presumptive remedies is provided [61]. The guide is available online from the
website (http://www.frtr.gov/matrix2/top_page.html). Some more web resources are
presented and discussed in the literature [62, 63].
It is beyond the scope of this chapter to cover every aspect of bioremediation in
details and to review any relevant results which have been obtained by the huge
community of researchers in science and industry in the last decades. However, a general
overview is attempted and a collection of interesting books and selected papers from the
literature is proposed for in depth study.
Contaminated soil remediation options
In terms of managing soil pollution, there are three generic approaches. The first
approach is to contain the contamination within the affected area. This option does not
reduce the total concentration of the pollutants in soil, but aims to manage exposure to the
pollutants, and hence environmental and health risks, through technologies that lower
pollutant mobility and bioavailability. The second approach is to treat the contaminated
mass in-situ, while the third approach is to remove the contaminated mass from its
surrounding environment and then treat it or dispose it ex-situ. The last two clean-up
technologies aim to reduce significantly the total concentration of the pollutants in soil to
a maximum allowable total concentration by applying passive or active bio/remediation
technologies [64].
Containment systems
Containment is necessary whenever contaminated materials are to be buried or left in
place at a site. The containment option for soil, sediment, bedrock and sludge includes
technologies such as cover systems, vertical barriers, horizontal barriers and hydraulic
control measures [55].
Cover systems: A covered system is composed by a multilayer construction placed
over the contaminated soil to reduce the harmful effects of the contaminants at the
surface, minimise water infiltration through the contaminants by rain, prevent upward
migration of groundwater by capillary rise, prevent airborne migration of the
contaminants, and where appropriate control gas migration (Figure 1). With such systems
contaminated soil and ground water can be physically isolated with low-permeability
barriers such as landfill caps, liners, and cutoff walls. The optimum combination of the
layers, in terms of composition, thickness, and sequence of materials is based on an
assessment of the physical and chemical properties of the entire system (e.g., chemical
resistance, physical resistance to climatic conditions, and ground conditions such as
cracking and channelling due to drought, freeze/thaw, settlement), construction aspects,
consideration of the reduction in environmental risk of the underlying contaminated land
and cost [55]. Construction difficulties and questions about long-term reliability have
limited the application of containment technologies in the past. Questions such as: (a)
How can methods for detecting defects in containment systems be improved? (b) How
26
Artin Hatzikioseyian
Figure 1. Schematic layout of covered system [61].
can the bottoms of vertical walls be effectively sealed? (c) What is the long-term
reliability of different materials used for containment? and (d) How significant is
diffusive transport of contaminants across barriers over long time scales? always arise
[65]. Technical details concerning the characteristics of geomembranes and the
construction of covered systems can be found in the literature [61, 64, 66, 67].
Vertical barriers: Vertical barriers are installed adjacent to contaminated ground (i)
to prevent the off-site lateral migration of contaminated groundwater, (ii) to divert clean
groundwater away from contaminated ground, and (iii) to reduce the extraction rates of
contaminated groundwater from hydraulic control measures. They can be used to funnel
groundwater to an in-ground treatment center (so called funnel and gate) and also be used
to cut-off the underground migration of gases. To be effective, vertical barriers are
normally tied into a natural low permeable layer at depth (e.g., a clay layer) or to an inground horizontal barrier. There are three common types of vertical barriers (i)
displacement systems (e.g., sheet piling, membrane walls), (ii) excavated barriers (e.g.,
shallow cut-off walls, slurry trench walls, secant walls), and (iii) injection barriers (e.g.,
chemical grouting, auger mixing, jet grouting).
Horizontal barriers: In-ground horizontal barriers are installed below the
contaminated ground to prevent vertical migration. They can be used in combination with
vertical barriers to isolate potentially mobile contamination. Horizontal barriers are
generally formed by injection of cement slurries at depth [55].
Hydraulic control measures: Hydraulic control measures are used to adjust the
groundwater flow around a contaminated area so that no further spread of contamination
takes place. This can involve preventing or reducing the contact of the groundwater with
the contaminated mass (e.g., lowering the water table), reducing, intercepting or
containing a plume of contaminated groundwater, supporting other remediation methods
such as in-ground barriers, or being part of groundwater remediation operations.
Hydraulic control measures are commonly carried out by pumping out groundwater from
a number of wells, or using diversion or interceptor trenches [55, 68].
Immobilisation, stabilisation and solidification of the contaminated area either in-situ
or ex-situ aim to reduce the mobility of the contaminants by the following actions:
Stabilisation: Forming chemically immobile compounds of the contaminant,
Solidification: Binding the soil together to form a monolithic block to prevent access
by external mobilising agents such as wind, rain, and groundwater. Contaminated
Principles of bioremediation processes
27
soil can be contained by solidifying it in place with chemical fixatives (e.g.
cementing agents including pozzolan-portland cement, lime-flyash pozzolan, and
asphalt) [65],
Vitrification: Melting and rapidly cooling the soil so that the contaminants are
immobilised and encapsulated into a glassy matrix. Vitrification may be carried out
in-situ or ex-situ. In-situ vitrification is performed by introducing an array of four
graphite electrodes into the soil and heating them electrically with powerful
generators to temperatures between 1600–2000 °C. At these temperatures the soil
melts and forms a glass block. Upon cooling, organic contaminants are pyrolysed and
reduced to gases during the melting process, while heavy metals remain enclosed in
the stabilised glass mass. This method has also been successfully used in treating
soils contaminated by radioactive materials. Vitrification may also be done ex-situ in
special appliances where contaminated soil would be molten in presence of
borosilicate and soda lime to form a solid glass block [43, 55].
In-situ treatment systems
In the case of in-situ techniques, soil and associated groundwater are treated in place
without excavation, [69]. Examples of in situ techniques include pump and treat,
percolation (flooding), bioventing, air sparging, bioslurping and permeable reactive
barriers, [23-25]. In-situ remediation ranges from partially closed (contained) systems to
completely open ones such as oil spills, (i.e. Exxon Valdez case [70]).
Pump-and-treat systems, which are applied to saturated-zone remediation the
removal of any contaminated water from the ground, treatment either at an on-site or offsite plant and return it back to the contaminated soil zone, (Figure 2) [64, 69, 71, 72].
This is one of the most traditional methods of remediation, but it is not the most effective
method for all contaminants because it is costly in investment - maintenance and very
time consuming [72]. However, the advantage of this method is that the contaminant is
actually removed entirely from this system if it is water soluble. For example in the cases
of methyl tertiary-butyl ether (MTBE), which is a fuel additive very soluble in water,
chlorinated solvents and hydrocarbons in groundwater, the pump and treat method
combined with carbon adsorption and air stripping is very effective [72, 73]. However,
for contaminants that bind very closely to the soil, such as polycyclic aromatic
hydrocarbon (PAHs), desorption of the contaminant from the soil to the groundwater is
very slow. When the groundwater is pumped out of the system, the contaminant present
in the aqueous phase is also removed, but a significant portion of the contaminant still
remains present in the ground. In order for pump and treat to be effective, it must be done
over a long period, in order to give the contaminant sufficient time to desorb from the
soil. However, this option becomes much more cost efficient and timely with the addition
of mobility agents such as cosolvents (e.g. alcohols), surfactants (e.g. amphiphilic
polyurethane nanoparticles, extracellular polymers, and cyclodetrins) and steam, that will
loosen the bond of the contaminant to the soil particles and increase the apparent
solubility of the contaminant [69, 74-76].
Some of the more conventional modifications of groundwater pump-and-treat
remediation and aquifer restoration strategies from petroleum hydrocarbons include
bioenhanced degradation [72]. Bioenhanced biodegradation can be done in situ and
involves the introduction of nutrients and air or peroxide to enhance the natural biologic
processes. Aboveground treatment can include a variety of strategies including chemical
precipitation, evaporation, reverse osmosis, and ion exchange to remove metals, and carbon
28
Artin Hatzikioseyian
Figure 2. A simplified pump and treat scheme [77]. Treated water is usually returned back in the
ground water.
absorption, air stripping, ultraviolet oxidation and use of bioreactors to remove organics.
Common dissolved hydrocarbon constituents treated by these methods are the aromatic
fuel hydrocarbons such as benzene, toluene, ethylbenzene, xylenes and fuel additives
such as lead and methyl tertiary-butyl ether (MTBE) [72].
Percolation consists of applying water, containing nutrients and possibly a microbial
inoculum, to the surface of a contaminated area and allowing it to filter into the soil and
mix with the ground water [69].
Bioventing is the process of injecting air (i.e., aeration) to an unsaturated soil zone
through the installation of a well(s) connected to associated pumps and blowers which
draw a vacuum on the soil [17, 58, 63, 69]. This process combines an increased oxygen
supply with vapour extraction. A vacuum is applied at some depth in the contaminated
soil which draws air down into the soil from holes drilled around the site and sweeps out
any volatile organic compounds. The development and application of venting and
bioventing for in situ removal of petroleum from soil have been shown to remediate
approximately 800 kg of hydrocarbons by venting, and approximately 572 kg by
biodegradation [18, 78].
Air sparging or biosparging involves the injection of air into the saturated zone of a
contaminated soil, at low flow rates (<5 m3/h per point) [17, 58, 69]. This is used to
increase the biological activity in soil and to promote aerobic biodegradation by
increasing the O2 supply via sparging of air or oxygen into the soil (Figure 3). In some
instances air injections are replaced by pure oxygen to increase the degradation rates.
However, in view of the high cost of this treatment in addition to the limitations in the
amount of dissolved oxygen available for microorganisms, hydrogen peroxide (H2O2) is
introduced as an alternative, and it is used on a number of sites to supply more oxygen
[18]. In biosparging volatilization is typically less than that of the standard air sparging
system [58].
Bioslurping is an in situ technology that combines vacuum-enhanced free-product
recovery with bioventing of subsurface soils to simultaneously remediate petroleum
Principles of bioremediation processes
29
Figure 3. Biosparging system used with soil vapor extraction [63].
Figure 4. Bioslurping technology [63].
hydrocarbon-contaminated groundwater and soils (Figure 4). Vacuum-enhanced recovery
utilizes negative pressure to create a partial vacuum that extracts free product and water
from the subsurface. The technology is portable and uses a single pump to extract free
product, groundwater, and soil gas from multiple wells. Groundwater and soil gas may
require treatment before being discharged. Bioslurping is used at petroleum spill sites and
has proven most effective in fine-to-medium textured soils or fractured rock in areas with
a low water table [44, 72].
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Artin Hatzikioseyian
Ex-situ treatment systems
In the case of ex-situ techniques, soil and groundwater are removed from their
original locations for treatment. Examples of ex-situ techniques include land farming,
irrigation, compost piles, engineered biopiles, and ex-situ slurry techniques [69].
Land farming is land treatment of soil for degradation or transformation of
contaminants by a combination of volatilization and biodegradation by indigenous
microorganisms. A common practice is to place the soil as a shallow layer within a
bermed and lined treatment cell (or biocell), occasionally amend the soil with nutrients
and water to stimulate biodegradation, and regularly till (aerate) the soil to mix and aid
contaminant volatilization [17, 58, 69]. A land farming bioremediation case study for a
site contaminated with hydrocarbons (e.g. PAHs, BTEX ) is presented in the literature [15].
Compost piles consist of soil supplemented with composting material (i.e. wood
chips, straw, manure, rice hulls, etc.) to improve its physical handling properties and its
water- and air-holding capacities. Although compost piles are exposed to the atmosphere,
the interior is often anaerobic due to the oxygen demand of the contaminants and
amendments. Thus, air is drawn through the compost (by vacuum, although aerated piles
have been used to enhance the drainage) to supply O2 to the soil for promoting aerobic
degradation of organic material and remove evaporated water. Compost piles are
subjected to intermittent mixing using specially manufactured equipment that is capable
of turning the pile over onto itself. Temperatures can increase to 60–70°C due to the
exothermic nature of biodegradation, and mixing, aeration, and moisture addition help
dissipate excess heat that could be inhibitory to biodegradation [15, 69]. One advantage
of composting is that it is more effective than other solid-phase treatment systems for
soils and sludges contaminated with viscous substances such as coal tar, creosote, or
petroleum production and still bottoms. Soil treatment using composting systems is
limited to biodegradable chemicals. The technology cannot treat metals and most other
inorganic chemicals. Additionally, the technology cannot readily biodegrade halogenated
chemicals. However the composting system effectively remediates soils that are heavily
contaminated and cannot be treated by in-situ methods as well as wastes containing
hazardous volatile constituents untreatable by land farming methods [44]. An example
case study of composting for bioremediation of a site contaminated with explosives (e.g
TNT, RDX) is presented in the literature [15].
Biopile is a remediation technique that involves placing contaminated soils into piles
or cells above ground, and stimulating aerobic or anaerobic microbial activity within soils
through controlled aeration and/or addition of minerals and nutrients (Figure 5.). Air is
supplied to the biopile via a pipe-and-pump system, which either forces air into the pile
(positive pressure) or draws air through the pile (negative pressure). Forcing air into the
pile helps maintain constant temperature and aerobic conditions, while drawing air out of
the pile can create anaerobic conditions. Although composting systems require large
amounts of nutrients and bulking agents, fewer additives are needed for biopiles. Biopiles
are normally operated at lower temperatures since less organic material is added. Biopiles
have some potential limitations. For example, certain chemicals such as polychlorinated
biphenyls (PCBs) and other hydrocarbons are resistant to biodegradation. In addition,
high concentrations of toxic metals, such as lead, copper, and mercury, may limit biopiles
treatment [44, 58, 63, 69].
Ex situ slurry techniques involve the creation and maintenance of a soil-water slurry
as the bioremediation medium. The slurry can be maintained either in a bioreactor or in a
pond or lined lagoon [29, 69]. Adequate mixing and aeration are key design requirements
Principles of bioremediation processes
31
Figure 5. Typical biopile system [63].
for slurry systems. Decomposition of organic contaminants takes place usually via
aerobic processes. Nutrients and, perhaps, inoculum may be added to the slurry [69]. A
significant benefit in the use of slurry biodegradation is the enhanced rate of contaminant
degradation, a direct result of improved contact between the microorganisms and
hazardous compounds. The agitation of contaminants in the water phase provides high
degree of solubilisation of compounds and greater homogeneity. More details for slurry
biodegradation technology can be found in the literature [16, 29].
Ex-situ solid applications involve the addition of water, nutrients and sometimes
addition of cultured indigenous microbes or inocula. They are often conducted on lined
pads to ensure that there is no contamination of the underlying soil [69].
Both in-situ and ex-situ techniques are capable of saturated and unsaturated zone
remediation, although restriction exists depending on the exact system used. However,
many factors can influence the effectiveness of each technique [69]. Ex situ techniques
allow more opportunities to control or engineer conditions for remediation. Although
some sites may be more easily controlled with ex-situ configurations, others are more
effective with in-situ treatment. For example, many sites are located in industrial and
commercial areas, making excavation extremely difficult. In addition if the contamination
is deep in the subsurface, excavation becomes too expensive. As a result of these physical
barriers, the required excavation efforts may make ex-situ biotreatment impracticable.
Other factors could also have an impact on the type of treatment. At a typical site, the
contamination is basically trapped below the surface. Exposing the contamination to the
open environment through excavation can result in potential health and safety risks. In
addition, the public’s perception of the excavation of contaminants could be negative,
depending on the situation. All of these conditions clearly favour in-situ biotreatment.
Nonetheless, the key is to carefully consider the parameters involved with each site
before evaluating which technique to use [74].
Factors affecting bioremediation
The most important principle of bioremediation is that microorganisms (mainly
bacteria or fungi) can be used to destroy hazardous contaminants or transform them to
less harmful forms. Thus, bioremediation of contaminants is an application of the
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Artin Hatzikioseyian
microbial metabolic activity. Microorganisms, through their enzymatic pathways, act as
biocatalysts and facilitate the progress of biochemical reactions that degrade the targeted
contaminants. As a result, bioremediation techniques are only applicable in environments
that can sustain life. The microorganisms act against the contaminants only when they
have access to a variety of materials-compounds to help them generate energy and
nutrients to build more cells. In very few cases the natural conditions at the contaminated
site provide all the essential materials in large enough quantities that bioremediation can
occur without human intervention - a process called intrinsic bioremediation. More often,
bioremediation requires the construction of engineered systems to supply microbestimulating materials - a process called engineered bioremediation. Engineered
bioremediation relies on accelerating the desired biodegradation reactions by encouraging
the growth of more organisms, as well as by optimizing the environment in which the
organisms must carry out the detoxification reactions [47].
The metabolic characteristics of the microorganisms in association with the
physicochemical properties of the targeted contaminants determine whether a specific
microorganism - contaminant interaction is possible. The actual successful interaction
between the two, however, depends on the environmental conditions of the site of the
interaction. Specific constrains should therefore be fulfilled for a successful
bioremediation attempt. These constrains encompass the microbial, chemical and
environmental characteristics of the targeted site.
Microbial constrains
A successful bioremediation effort relies on the utilisation of the appropriate
microorganisms [25, 79]. Such microbial populations can in theory be consortia of
naturally existing species or genetically engineered microorganisms. Most applications
rely on the use of naturally existing microbial populations which often are not well
characterised. That is to say the microbial populations are effective in their desired
application but the complete characterisation of the population is not well known. This
knowledge gap is not necessarily the result of a scientific inability but rather of the
continuous dynamic adaptation of the microbial species to their environments. An
example of this ability of microbial populations to adapt to the presence of man-made
chemicals comes from the field of medicine, where the rapid adaptation of pathogenic
organisms and their resulting immunity to specific classes of antibiotics as a result of the
excessive use of these antibiotics has been well documented.
These adaptational mechanisms advance through selection processes in which variant
species with a specific survival advantage for the given environment take over and
survive successfully. The survival advantage often relies on the ability of an organism to
metabolise as substrate organic molecules (pollutants) existing in a given site.
Contemporary microbiological techniques allow the identification of such
transconjugants that originate from a background microbial population confirming that
such processes are active in bioremediation practice [70].
Horizontal transfer of catabolic plasmids among different species existing within a
site may also result into species that possess enhanced catabolic or resistance potential
[80]. Such plasmid containing bacteria have been isolated from polluted sites [81-84].
Species that can through such plasmid transfer catabolise as sole carbon source hazardous
xenobiotic compounds (as for example 3-chlorobenzoate) have been reported [85, 86].
The ultimate impact, however, of such plasmid transfer processes on the field application
potential of bioremediation will have to pass through the previously described path of
Principles of bioremediation processes
33
natural selection. A newly acquired metabolic advantage will be assessed, through the
mechanism of natural selection, and may allow the ultimate successful establishment of a
transconjugant species in a contaminated site.
Genetically modified microorganisms (GMOs) have often been presented as offering
a major potential advantage for bioremediation. The development of recombinant DNA
and other genetic engineering technologies, in the late 1970s, was believed that could be
widely applied for environmentally-beneficial purposes, including the clean-up of
contaminated soil and water [80, 87-89]. The continuously growing knowledge on
catabolic pathways and critical enzymes provides the basis for the rational genetic design
of new and improved enzymes and pathways for the development of more effective
processes. Many researchers had expected that genetically modified organisms having
novel biochemical traits or enzymatic activities would quickly find broad applicability in
bioremediation of hazardous chemicals from the environment [20, 30]. However the
practical impact of GMOs is likely to remain low for many key reasons. Public,
regulatory, economic and technical issues associated with the release of genetically
engineered, or recombinant, microbial species into an open environment usually arise.
Many site owners, consultants and regulators are more comfortable choosing
technologies and methods with which they are familiar, have a long track record of
success and thus a greater predictability. Legislative reasons are associated with the strict
control on the release of such organisms into the environment. There is significant
concern about the long term survival of genetically engineered species into a natural
environment where they would have to compete with the naturally existing consortia that
had ample time to adapt to the prevailing environmental conditions. Thus, difficulties in
obtaining permission to use genetically engineered microorganisms from government
regulatory agencies as well as public controversies have made companies reluctant to
develop bioremediation strategies based on GMOs [20, 30]. Finally, their use is
considered costly. Technically speaking, it seems more plausible to use GMOs in ex-situ
bioremediation treatment schemes in bioreactors, designed for use with defined soil
slurries or water streams in tightly controlled environments. Not only does this limit the
widespread release of the GMOs into the environment and avoids the problem of
competition with indigenous microflora, but also allows the microorganism to be
maintained at controlled temperatures and other growth conditions, to be used with
relatively well-defined waste streams containing one or a small number of specific
contaminants.
The application of the genetically engineered microorganisms in industrial scale
bioremediation is not yet prominent. Until today genetically engineered microorganisms
have not been used in commercial site remediation projects, with few only exceptions
[90]. Most bioaugmentation projects have used naturally-occurring bacteria for which
obtaining regulatory approval is relatively easy. However, recently transgenic plants
begin to find applicability in commercial phytoremediation projects.
Chemical constrains
Bioavailability of contaminants
In order for the pollutants to be amenable to biological degradation they must be
bioavailable [91]. Bioavailability is related to the physical state of the contaminant and
the possibility of efficient contact between the microorganism and the contaminant. This
contact is best when the microorganism-contaminant interface is maximised. Regarding
physical state, microorganisms generally assimilate pollutants from the liquid phase and
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Artin Hatzikioseyian
cannot effectively degrade a pollutant until it desorbs from aquifer solids, diffuses out of
nanopores, or dissolves from nonaqueous phase liquids (NAPLs) into the bulk solution.
In such cases, the rate of biodegradation can be controlled by the diffusion, desorption, or
dissolution rates. Polar, water soluble contaminants are more easily bioavailable. The
increase of the contaminant - microorganisms contact surface for hydrophobic
contaminants may require the addition of surface active agents. Knowledge of
partitioning and rates of transfer of a chemical between its disolved-sorbed-volatile states
becomes important in defining its bioavailability. Bioavailability comprises the effects of
all the physical and chemical parameters that eventually dictate the potential for the
microbial utilisation of a compound and thus its biodegradation potential [15, 33, 46, 91].
Biodegradability of contaminants
The success of any bioremediation project depends mainly on the chemical structure
of the organic molecules present in the contaminated site [25, 46]. Some structural
features of organic compounds that are not common in nature, called ‘‘xenophores’’ (e.g.,
substitutions of H with Cl, NO2, CN, and SO3 groups), make such molecules difficult to
be metabolized by microorganisms. Thus, contaminants that contain such xenophores
tend to be recalcitrant to microbial degradation, [14, 25, 33, 46]. Table 1. presents the
experienced biodegradability potential of different target organic molecules.
Numerous mechanisms and pathways have been elucidated for the biodegradation of
a wide variety of organic compounds [25]. All metabolic reactions are mediated by
enzymes. These belong to the groups of oxidoreductases, hydrolases, lyases,
transferases, isomerases and ligases. Many of the oxygenase enzymes that attack
aromatic hydrocarbons have a remarkably wide degradation capacity due to their non
specific substrate affinity. For example, toluene dioxygenase is capable of degrading
more than 100 different compounds, including TCE, nitrobenzene, and chlorobenzene.
Other examples are esterases, which break down ester bonds by the addition of water;
depolymerases, which hydrolyze polymers; dehalogenases, which remove halogen atoms
Table 1. Biodegradability of various compounds [37, 63].
Simple hydrocarbons, C1–C15
Very easy
Alcohols, phenols, amines
Very easy
Acids, esters, amides
Very easy
Hydrocarbons, C12–C20
Moderately easy
Ethers, monochlorinated hydrocarbons
Moderately easy
Halogenated and non-halogenated volatile organic
compounds (VOCs)
Moderately easy
Halogenated and non-halogenated semi-volatile organic
compounds (SVOCs)
Moderately easy
Hydrocarbons, greater than C20
Moderately difficult
Multichlorinated hydrocarbons
Moderately difficult
Polycyclic aromatic hydrocarbons (PAHs), Polychlorinated
biphenyls (PCBs),
Pesticides and herbicides
Very difficult
Principles of bioremediation processes
35
such as chlorine and replace them with —OH groups; decarboxylases which remove CO2
groups (i.e., decarboxylation), hydratases which add water to alkenes converting them
into secondary alcohols; glutathione S-transferase which transfers the thiol group to
chlorinated compounds with concomitant dechlorination; racemases which catalyze Land D-amino acid interconversions and finally CoA-ligase, which adds -S-CoA to fatty
acids during beta-oxidation. For details of different biotransformation pathways the
reader is referred to the literature [15, 22, 25, 46, 92].
Other contaminant properties
Contaminant properties are critical to contaminant-soil interactions, contaminant
mobility and to the ability of treatment technologies to remove, destroy or immobilize
contaminants. Important contaminant properties include: Solubility in water, dielectric
constant, diffusion coefficient, molecular weight, vapour pressure, density and aqueous
solution chemistry [40].
Nutrients
Most in-situ bioremediation methods practiced today rely on the stimulation of
indigenous microbial populations at the site of contamination, by addition of appropriate
nutrients, principally carbon, oxygen, nitrogen and phosphorus, and by maintaining
optimum conditions of pH, moisture and other factors, to trigger increased growth and
activity of indigenous biodegradative microorganisms [17].
Nitrogen and phosphorus requirements are often estimated by calculating a carbon to
nitrogen to phosphorus ratio C/N/P close to 100/(10 to 5)/1. Many authors report
optimum experimental results C/N/P ~70/3/0.6, [93], 8/1/0.07, [94], 1/11/2 [95] for crude
oil bioremediation of different origin. Fertilizers such as paraffinized urea and
octylophosphate in C/N/P 100/10/1 respectively have been suggested for optimal growth.
Dibble and Bartha [96] suggest ratio of C/N/P 800/13/1, illustrating that the nutrient
requirement is specific to oil-in-water mixtures and needs individual consideration for
any case. Suggested C/N values for composting are between 30-40 [97]. A detailed
excellent review for nitrogen and phosphorous requirements for bioremediation as well as
the detrimental effects of excess nutrients can be found in the literature [98].
Nutrients typically are delivered by controlling ground water flow using injection
wells or burred perforated pipes (infiltration gallery). In the most common configuration,
ground water withdrawn from production wells downgradient from the biostimulation
zone is amended with the nutrients required for biostimulation, treated if necessary to
remove contaminants, and reintroduced to the aquifer upgradient of the biostimulation
zone using the injection wells or infiltration galleries. Water from an external source is
required if the flow of withdrawn water is insufficient to control the subsurface flow or if
it is infeasible to reinject the withdrawn ground water. The rate of nutrient delivery to the
biostimulation zone, therefore, is often limited by the solubility of the nutrients in water
and the reinjection flow rate [47].
Oxygen, air, hydrogen peroxide
In the majority of applications, bioremediation is an oxidation process. During
oxidation of the contaminants, microorganisms extract energy via electron transfer.
Electrons are removed from the contaminant and transferred to a terminal electron
acceptor which, during aerobic biodegradation, is oxygen. During decomposition of the
36
Artin Hatzikioseyian
organic substrate, oxygen concentrations in the subsurface may become depleted [29].
The major kinetic limitation on aerobic bioremediation is often the availability of oxygen
due to the low solubility of oxygen in water. This is more intense in the cases of organic
molecules with high oxygen demand such as petroleum hydrocarbons. Air, oxygen, or
other oxygen sources (e.g., hydrogen peroxide, ozone) may need to be added to the
infiltration water to promote aerobic biodegradation. Air sparging of water can supply
8 mg/L dissolved oxygen, sparging with pure oxygen can deliver 40 mg/L, while
application of hydrogen peroxide can provide more than 100 mg/L oxygen. Therefore,
while air sparging is the simplest and most common oxygen delivery technique, the use
of oxygen or hydrogen peroxide may speed the bioremediation process and decrease
the pumping required. However, in some cases the increased cost and potential
explosion hazard associated with pure oxygen supply may limit the applicability of
direct oxygen use [47]. On the other hand, application of hydrogen peroxide to in-situ
bioremediation is limited by its toxicity to microorganisms, its potential for causing
aquifer plugging due to the highly reactive nature of hydrogen peroxide resulting in
chemical oxidations of organic and inorganic compounds, producing precipitates [99],
and the formation of oxygen bubbles which decreases aquifer permeability [16, 100,
101]. The problems associated by the use of alternative oxygen sources are discussed in
the literature [29].
Alternative electron acceptors
In the absence of molecular oxygen, anaerobic microorganisms use other forms of
combined oxygen. For example, denitrifying bacteria use nitrate (NO3-), nitrite (NO2-), or
nitrous oxide (N2O); dissimilatory metal-reducing bacteria use manganese or ferric iron
oxides (e.g., MnO2, Fe(OH)3, or FeOO-); sulfate-reducing bacteria use sulfate (SO42- );
and methanogens use carbon dioxide (CO2) or bicarbonate (HCO3-) as electron acceptors
[15, 102]. In cases where oxygen is progressively depleted, electron acceptors are
generally used up in a set sequence determined by the appropriate redox potentials of the
oxidation reactions under consideration [15, 103]. Thermodynamic concepts imply the
following sequence of electron acceptor utilization:
O2 → NO3− → Mn 4+ → Fe3+ → SO42− → HCO3−
The implication of this thermodynamic analysis is that when the electron acceptor
demand is relatively high (e.g., near the source zone), microbial degradation would
sequentially deplete the available oxygen, then nitrate, manganese, ferric iron, and sulfate
before methanogenesis becomes predominant. Thermodynamic considerations also imply
that heterotrophic microorganisms capable of deriving the maximum amount of energy
per unit of carbon oxidized would have a competitive advantage over other species, and
their respiration mode would become dominant until their specific electron acceptor is
used up [15].
Metal ions
Although some metals are essential in trace quantities for microbial growth, heavily
contaminated sites with high concentrations of metal ions in contaminated soil or water
usually inhibit the metabolic activity of the cells, thus affecting directly any
bioremediation process [74].
Principles of bioremediation processes
37
Toxic compounds
High aqueous phase concentrations of some contaminants can create toxic effects to
microorganisms, even if the same chemicals are readily degraded at lower concentrations.
Toxicity prevents or slows down microbial metabolic activity and often prevents the
growth of new biomass needed to stimulate rapid contaminant removal. The degree and
mechanisms of toxicity vary with specific toxicants, their concentration, and the exposed
microorganisms. Some organic compounds are toxic to targeted life forms such as insects
and plants and may also be toxic to microbes. These compounds include herbicides,
pesticides, rodenticides, fungicides, and insecticides. In addition, some classes of
inorganic compounds such as cyanides and azides are toxic to many microbes; however,
these compounds may be degraded following a period of microbial adaption [74].
Biogeochemical parameters
Measurements of various biogeochemical parameters such as dissolved oxygen
(DO), redox potential, CO2, and other parameters such as NH4+, NO3-, NO2-, SO42-, S2and Fe2+ will give an indication of the existing (natural or intrinsic) microbial metabolic
activity at the site [37].
Environmental constrains
Temperature
Microbial metabolism is substantially affected by temperature [104]. Most
microorganisms grow well in the range of 10 to 38°C. Technically it is extremely
difficult to control the temperature of in-situ processes, and the temperature of ex-situ
processes can only be moderately influenced, sometimes with great expense. Fortunately,
although temperatures within the top 10 m of the subsurface may fluctuate seasonally,
subsurface temperatures down to 100 m typically remain within 1° to 2°C of the mean
annual surface temperature suggesting that bioremediation within the subsurface would
occur more quickly in temperate climates [105].
pH
The pH range within which most bioremediation processes operate most efficiently is
approximately 5.5 to 8. It is no coincidence that this is also the optimum pH range for
many heterotrophic bacteria, the major microorganisms in most bioremediation
technologies. The optimum pH range for a particular situation, however, is site-specific.
The pH is influenced by a complex relationship between organisms, contaminant
chemistry, and physical and chemical properties of the local environment. Additionally,
as biological processes proceed in the contaminated media, the pH may shift and
therefore must be monitored regularly. The pH can be adjusted to the desired range by the
addition of acidic or basic substances (i.e., mineral acids or limestone, respectively).
However changes in soil pH will influence dissolution or precipitation of soil metals and
may increase the mobility of hazardous materials. Therefore, the soil buffering capacity
should be evaluated prior to application of amendments [29].
The effect of pH on permeability of soils and sediments is not fully understood but it
seems that soil pH has also significant effect. Soils have a negative permanent charge and
a pH-dependent variable charge. Therefore, pH affects soil dispersion and its permeability.
38
Artin Hatzikioseyian
A typical volcanic ash soil has a large amount of pH-dependent charge. Its saturated
hydraulic conductivity decreases under low and high pH conditions. When the
predominant anion is sulphate, hydraulic conductivity does not decrease even at low pH.
However, the saturated hydraulic conductivity of soils with montmorillonite and kaolinite
at pH 9 is smaller than that at pH 6 [21].
Moisture content - water activity
Moisture is a very important variable relative to bioremediation. Moisture content of
soil affects the bioavailability of contaminants, the transfer of gases, the effective toxicity
level of contaminants, the movement and growth stage of microorganisms, and species
distribution. During bioremediation, if the water content is too high, it will be difficult for
atmospheric oxygen to penetrate the soil, and this can be a factor of limiting growth
efficiency and determine the types of organisms that can flourish. Various workers in the
field have reported that the water content of the soil should be between 20 and 80%. In
cases where no extra source of oxygen is being provided (for example, bioremediation of
surface contamination), 20% moisture may be adequate; however, if a continuous
recirculation system (pipe networks) is being used for deeper contamination, 80% water
content would be more appropriate [74].
Soil moisture is frequently measured as a gravimetric percentage or reported as field
capacity. Evaluating moisture by these methods provides little information on the “water
availability” for microbial metabolism. Water availability is defined by biologists in
terms of a parameter called water activity (aw). In simple terms, water activity is the ratio
of the system’s vapour pressure to that of pure water (at the same temperature) [37, 74].
Redox potential
The redox potential of the soil (oxidation-reduction potential, Eh) is directly related to
the concentration of O2 in the gas and liquid phases. The O2 concentration is a function of
the rate of gas exchange with the atmosphere, and the rate of respiration by soil
microorganisms and plant roots. Respiration may deplete O2, lowering the redox potential
and creating anaerobic (i.e., reducing) conditions. These conditions will restrict aerobic
reactions and may promote anaerobic processes such as denitrification, sulfate reduction,
and fermentation. Many reduced forms of polyvalent metal cations are more soluble (and
thus more mobile) than their oxidized forms. Well-aerated soils have an Eh of about 0.8 to
0.4 V; moderately reduced soils are about 0.4 to 0.1 V; reduced soils measure about 0.1 to 0.1 V; and highly reduced soils are about 0.1 to -0.3 V. Redox potentials are difficult to be
measured in the soil or groundwater and are not widely used in the field [29].
Mass transfer characteristics
Mass transport characteristics are used to calculate potential rates of movement of
liquids or gases through soil and include: Soil texture, unsaturated hydraulic conductivity,
dispersivity, moisture content vs. soil moisture tension, bulk density, porosity, hydraulic
conductivity and infiltration rate [40, 106-108].
Site hydrogeologic characteristics
Hydrogeologic factors for consideration include aquifer type, hydraulic conductivity,
hydrogeologic gradient, permeability, recharge capability, depth to groundwater, moisture
Principles of bioremediation processes
39
content/field capacity, thickness of the saturated zone, homogeneity, depth to
contamination, extent of contamination, and plume stability. These are only some
parameters that should be factored into the design of any bioremediation system [37, 40, 72,
106-108].
Biostimulation
Biostimulation is the modification of the environment by adding nutrients, such as
nitrogen and phosphorus, as well as oxygen and other electron acceptors to stimulate the
rate of biological degradation of contaminants by indigenous microorganisms. This
alternative is also chosen when a natural microbial population exists at the site which has
the potential to degrade the chemicals, but is actually lacking oxygen, nitrogen or other
nutrients to degrade them. The missing component(s) can then be introduced into the
system and the degradative activity of the microbial community can be induced. Most
bioremediation systems employ some form of biostimulation [37, 58].
Commonly used water soluble nutrients include mineral salts (e.g., KNO3, NaNO3,
Ca(NO3)2, NH4NO3, (NH4)2SO4, K2HPO4, (NH4)2HPO4, MgNH4PO4,), anhydrous
ammonia (NH3), urea (NH2)2CO and many commercial inorganic fertilizers (e.g., the
23:2 N:P garden fertilizer used in the Exxon Valdez case). There is some disagreement
on the efficacy of various inorganic nitrogen salts for bioremediation. It is generally
not agreeable which nitrogen source ammonium or nitrate is the most preferable. Water
soluble nutrients are usually applied in the field through the spraying of nutrient
solutions or spreading of dry granules. This approach has been effective in enhancing
oil biodegradation in many field trials [109]. Compared to other types of nutrients,
water soluble nutrients are more readily available and easier to manipulate. Another
advantage of this type of nutrient over organic fertilizers is that the use of inorganic
nutrients eliminates the possible competition of carbon sources. However, water
soluble nutrients also have several potential disadvantages. First, they are more likely
to be washed away by the actions of tides and waves when applied in seawater, and
second, inorganic nutrients, ammonia in particular, should be added carefully to avoid
reaching toxic levels [17].
Slow released or control released granular nutrient sources have been also used for
bioremediation of oil spill contaminated areas, providing continuous release of nutrients
over extended time period [110]. Slow release fertilizers are normally in solid forms that
consist of inorganic nutrients coated with hydrophobic materials like paraffin or
vegetable oils (e.g MaxBac® [44]). Other examples of slow released nutrients are sulfurcoated urea and Osmocote®, slowly soluble materials (such as metal ammonium
phosphates), or materials that must be microbially mineralized to release nitrogen (e.g.
organic fertilizers and urea formaldehydes). To maximize effectiveness, controlledrelease nutrients should be released at a rate equivalent to microbial demand. However,
the main disadvantage of these materials is that the nutrient release rate cannot be easily
controlled in the field [109, 111-113].
Another approach to overcome the problem of water soluble nutrients being rapidly
washed out is to utilize oleophilic organic nutrients (e.g. Inipol EAP22® - Elf Atochem of
North America, Inc. - an oleophilic urea-based fertilizer) [44, 58, 114, 115]. Since oil
biodegradation mainly occurs at the oil-water interface and since oleophilic fertilizers are
able to adhere to oil and provide nutrients at the oil-water interface, enhanced
biodegradation should result without the need to increase nutrient concentrations in the
bulk pore water [114]. According to the vendor, Inipol EAP-22 has several advantages:
40
Artin Hatzikioseyian
(a) Optimizes ratio between carbon, nitrogen, and phosphorus, (b) releases nutrients over
time, (c) inhibits the formation of water-in-oil emulsions and (d) is completely
biodegradable [44].
Some commercial proprietary biostimulation agents are branded by the names
FyreZymeTM, (Ecology Technologies International, Inc.) and HCZyme, (CCharbon
Consultants). Both are a combination of bacterial growth enhancing agents, extracellular
enzymes, and surfactants. The bacterial growth enhancers increase natural biological
processes by stimulating a logarithmic growth phase of indigenous microbes in soil and
water, while extracellular enzymes initiate the oxidation that degrades petroleum-based
contaminants, and surfactants help desorb the petroleum bound to soil particles.
FyreZymeTM also contains a biodegradable compound that adds oxygen to the soil,
thereby facilitating hydrocarbon degradation [44].
Biostimulation has been successfully applied in marine oil spills [18, 39, 58, 116]
and polycyclic aromatic hydrocarbon (PAH)-contaminated soils [117]. A review of oil
spill biodegradation literature is given by Swannell, Lee & McDonagh [109].
Bioaugmentation
Bioaugmentation is the soil amendment with non-indigenous allochthonous
microorganisms or cultivated indigenous species or engineered microbes via inoculation,
or use of bioproducts, (e.g. enzymes i.e. lipases, proteases, cellulases etc.), to enhance the
biodegradation of contaminants [58]. The introduced microorganisms augment, but do
not replace, the resident microbial population. Usually bacteria with necessary catalytic
activities and other required characteristics are injected directly into the polluted site
usually together with nutrients. Bioaugmentation can be necessary also in cases where
bacteria with the required catalytic activity although present at the site degrade the
pollutants incompletely and/or at a very slow rate.
Successful microbial inoculation requires a range of factors: (1) the population must be
capable of surviving and growing in the new environment; (2) the microorganisms must
retain their degradative abilities under the new conditions; (3) the organisms must come in
contact with the contaminants; and (4) the electron donors/acceptors and nutrients necessary
for microbial growth and contaminant degradation must be made available to the population
[118]. Once the microorganisms are injected into the aquifer, there must be some
mechanism for dispersing them throughout the biostimulation zone before they attach to the
solid matrix and carry out the degradation reaction of interest. Cell transport within porous
media is highly dependent on the characteristics of both the solid media and the microbial
cells. Experiments have shown that the conditions that best promote microbial transport in
porous media include (in order of their importance) highly permeable media, ground water
of low ionic strength, and small-diameter cells [119]. Unfortunately the microorganisms
that are efficient in the laboratory conditions do not cope well in the real world. Under
natural conditions, laboratory strains face unfavourable nutritional and physicochemical
conditions. They have to compete with already established indigenous communities and
they have to withstand a variety of predators. Moreover, there is often a mismatch between
the normal habitat of the introduced species and the ecological conditions in which they are
placed. Finally, when biostimulation and bioaugmentation are used simultaneously, it is a
common finding that the added nutrients favour mostly indigenous populations so that they
overgrow the introduced species.
The experience obtained from Exxon Valdez disaster at the coast of Alaska in 1989
in conjunction with research following other oil spills, demonstrated that in the case of
Principles of bioremediation processes
41
the Exxon Valdez clean-up, biostimulation of indigenous microbial consortia through the
use of nitrogen and other fertilizers is more effective than bioaugmentation through
seeding with exogenous organisms [112, 120, 121]. It seems that in most cases
biostimulation with nutrients is often more effective concerning biodegradation rates than
bioaugmentation with inoculums [3, 58, 116].
Most microbial inoculants or additives sold for use in bioaugmentation approaches
have historically been blends or consortia of microorganisms, purportedly tailored for the
types of compounds found in the target waste stream. Some commercial proprietary
bioaugmentation products are branded by the names M-1000TM, (Micro-Bac
International, Inc), Bac-Terra, (Microbe Technology Corporation), ABR (Augmented
Bioreclamation) Microbial Blends, (SSybron Chemicals, Inc.), WST Bioblends, (Waste
Stream Technology, Inc.), PetroKlenz, (Aqualogy BioRemedics), GT-1000, (Bio-Genesis
Technologies). Information for these products is limited and originates mainly from the
production/application companies. For example, Bac-Terra includes natural organic
matter with a blend of microbial consortia capable of working in both aerobic and
anaerobic environments, including psychrophilic, mesophilic, thermophilic, bacteria
cultures for use at temperatures ranging from 28 to 240°C. Bac-Terra requires soil
moisture greater than 20%. PetroKlenz is a dry powder containing specific cultured
facultative anaerobes, naturally occurring microbes that were originally derived from soil
and have been preserved through advanced drying techniques. The various strains are
grown individually in pure culture and compounded together with powdered wetting
agents, buffering agents, and other synergists that allow the organisms to readily adapt to
the treatment environment. The organisms have been carefully matched to complement
each other for the effective biodegradation of hydrocarbons. GT-1000 is a synergistic
group of non-pathogenic microorganisms. According to the vendors most of these
products have been used successfully in bench, pilot or full scale levels to demonstrate
the ability to degrade pollutants such as benzene, ethylene, toluene, and xylene (BETX),
volatile aromatic hydrocarbons, oil and grease, coal tars, phenolic compounds, and
chlorinated organic solvents [44]. Other microbial bioaugmentation products are cited in
the literature [20].
Up-to-date, there has been little convincing evidence for successful in situ
remediation of aquifers resulting from introduced microbial populations. The limitation
of distributing the exogenous microbial cultures in the subsurface and the question of
long-term survivability of these lab-grown cultures under field conditions also discourage
bioaugmentation. Bioaugmentation may play a prominent role in bioremediation when
the release of genetically engineered organisms will be permitted [17, 20, 30, 37, 47].
Monitored natural attenuation
Natural attenuation, often called intrinsic remediation, intrinsic bioremediation,
bioattenuation, or monitored natural attenuation (MNA), consists of unassisted and unenhanced physical, chemical, and biological processes (e.g. biodegradation, abiotic
transformations, mechanical dispersion, dilution, evaporation, volatilization and
adsorption) that act to limit the migration and reduce the mass, toxicity, mobility,
volume, or concentration of contaminants in soil, sediment, or groundwater [15, 19, 44].
Natural attenuation, as an in-situ technology, is very important because it is often
technically infeasible to clean a contaminated site to regulatory cleanup levels for a
variety of conditions including the presence of low-permeability soils, the inability to
remove all the contaminants from the individual soil particles etc.
42
Artin Hatzikioseyian
Natural attenuation is usually considered as the ‘baseline option’, and although it
takes place without human intervention, the technology is not equal to a “do-nothing” or
“no further action” approach. The difference between the use of “natural attenuation” and
“no further action” as a remedial strategy is that natural attenuation requires thorough
documentation and extensive monitoring of the role of microorganisms and other
attenuation processes in eliminating the target contaminants. This implies that additional
site characterization and development of a groundwater monitoring phase for an
acceptable period of time may be necessary. Natural attenuation, if properly
demonstrated, increases the overall protection of the environment by either containment
or destruction of contaminants. No further action, on the other hand, implies that no
additional investigation is required regardless of whether the contaminants of concern are
degrading or migrating. Natural attenuation also serves as (1) an interim measure until
future technologies are developed, (2) a managerial tool for reducing site risks, and (3) a
bridge from active engineering (i.e., pump-and-treat, vapour extraction, etc.) to no further
action. No further action, however, may be preferable to natural attenuation in certain
instances. Very low risk situations may be better served since it eliminates the need of
continued monitoring and further documentation. Sites with low levels of contaminants
or nondiscernible plumes may be better candidates for no further action. Furthermore,
very minor releases of hydrocarbons to the subsurface may not be sufficient to support
bioremediation [72].
A site-specific, cost–benefit analysis is required to determine if an active remediation
system or MNA would be the most effective remediation option. MNA may be an
appropriate cleanup option when the facility can demonstrate that the remedy is capable
of achieving specific ground-water cleanup levels in a reasonable cleanup time frame. If
MNA is chosen then there are several costs associated with the implementation of it.
These costs include modelling contaminant degradation rates to determine if natural
attenuation is a feasible remedial alternative, subsurface sampling and sample analysis
(potentially extensive) for determining the extent of contamination and confirming
contaminant degradation rates and cleanup status. Regular operation and maintenance (O
& M) costs are required for monitoring to verify degradation rates and maintain data on
contaminant migration. In some cases, such long-term monitoring may be more
expensive than active remediation [44].
When natural attenuation is permitted as a remediation strategy, extensive monitoring
is required. It is beyond the scope of this paragraph to report extensive strategies for
MNA protocols. Interested readers can find information in the literature [35, 40, 72, 122]. While natural attenuation will not be a suitable remedy for all contaminated sites, it
does offer the potential advantages of in-situ technologies:
•
•
•
•
•
Generates less secondary wastes, reduced risk of human exposure during treatment,
reduced potential for cross-media transfer of contamination.
Operates in-situ with minimal site disturbance.
Can be used in conjunction with other remediation technologies.
Reduced need for on-site structures associated with cleanup.
Potentially reduces overall remediation costs.
However, the potential limitations of natural attenuation include [15, 44]:
•
It is well established as a remediation approach for only a few types of contaminants,
(e.g. benzene, toluene, ethylbenzene, and xylene referred as BTEX, oxygenated
hydrocarbons, low-molecular-weight alcohols, ketones, esters, and methylene chloride).
Principles of bioremediation processes
•
•
•
•
•
•
43
Generally requires longer time frame for remediation, (for many years or decades).
Requires more involved site characterization and monitoring.
Toxicity and mobility of transformation products may be greater than that of the
parent compound. Some compounds can form hazardous by-products that in some
cases can persist in the environment.
Changes in environmental or site conditions may allow contaminant migration.
There is a potential for remobilization of previously stabilized metals and
radionuclides.
Public may see natural attenuation as a “do-nothing” approach.
Biotransformation of metals, metalloids and radionuclides
The toxicity and mobility of the elements depend primarily on their speciation, which
is significantly influenced by soil pH, redox conditions, and surface chemistry. All of
these are environmental factors that can be optimized by manipulating microbial
activities to reduce the risk posed by heavy metals in aquifer environments.
Microorganisms cannot convert metals to different elements. However, they can modify
the microenvironment around the microbial cell and can catalyze oxidation, reduction,
methylation and dealkylation reactions that affect the solubility and mobility of many
metals. In addition, the microbial cell offers a large number of possible physico-chemical
mechanisms of interaction (e.g. complexation, coordination, chelation, ion exchange,
adsorption, microprecipitation) with soluble metal, radionuclide and metalloid species
resulting in immobilization of them [24, 52, 123-128]. Figure 6, presents different
possible metal mobilization/immobilization mechanisms.
Figure 6. Biological metal mobilization/immobilization mechanisms.
44
Artin Hatzikioseyian
Bioprecipitation
Metabolic mediated processes modify the environment around the microbial cell.
Under aerobic conditions microorganisms grow by transferring the electrons available
from the electron donor molecule (in the case of bioremediation usually an organic
contaminant) to the oxygen dissolved or transferred from the ground to the aquifer.
Organic carbon is mineralized to carbon dioxide and oxygen is reduced to water. The
produced dissolved carbon dioxide increases the alkalinity and pH of the cells
microenvironment and the excess bicarbonate favours the precipitation of metal ions as
metal hydroxides Me(OH)x or carbonate Me2(CO3)x [103, 124].
In subsurface environment, where anaerobic conditions prevail, nitrates can act
alternatively as terminal electron acceptors. The process known as denitrification is
widely used in municipal wastewater treatment units. Removal of nitrates proceeds
through nitrite intermediate, to nitrogen gas. From the oxidation of the carbon source,
bicarbonates are also produced increasing the pH of the medium. Soluble metal ions are
precipitated as metal hydroxides or carbonates with reactions similar to the case of
aerobic metabolism. A wide range of bacteria such as Pseudomonas and Alcaligenes sp.
immobilize dissolved metal species by the previously described action [103, 124, 129].
Finally under anoxic conditions and in the presence of sulphates, sulphate reducing
bacteria (SRBs) can grow using sulphates as terminal electron acceptor. Genera such as
Desulfovibrio, Desulfobacter, Desulfococcus, Desulfosarcina, Desulfomonas and
Desulfotomaculum, Desulfomicrobium, Desulfobulbus can oxidize simple organic
molecules such as ethanol, lactate or acetate to produce excess of sulphides in their
microenvironment [102, 130]. Under such conditions metal ions can immobilize as metal
sulphides, which are the most insoluble forms among the various metal precipitates (e.g.
hydroxides, carbonates and/or phosphates) [103, 123, 124, 129-132]. Successful large
scale ex-situ applications of SRBs have been documented in the literature [126].
Denitrifying microorganisms that use nitrate as an electron acceptor as well as
sulphate-reducing bacteria which reduce sulphate to sulphide have been stimulated in situ
by injecting acetate as a primary substrate and nitrate/sulphate as the electron acceptor in
many bioremediation schemes [29, 103, 130].
Bioreduction – Biooxidation
A wide variety of microorganisms can catalyze the reduction of heavy metals, such
as Fe(III) to Fe(II), Mn(VI) to Mn(II), Cr(VI) to Cr(III), Se(VI) to Se(IV) or Se0, As(V)
to As(III), Mo(VI) to Mo(IV) and U(VI) to U(IV). In such bioreduction processes
reduced elements can serve as electron acceptors in alternative microbial respiration, or
reduced by enzymes without energy production [125, 127]. Few examples for redox
couples are presented below.
Chromium appears in the environment in two oxidative states: Cr(VI) and Cr(III).
Chromium(VI) is mobile in groundwater and is considered carcinogenic and more
hazardous than chromium(III), which is a cation that tends to bind strongly to aquifer
material [29, 133]. A wide range of microorganisms are capable of enzymatic reduction
of Cr(VI) [134]. Some cells also obtain energy from Cr(VI) reduction, although most
cells mediate this reaction cometabolically and do not harvest energy. For example,
dissimilatory reduction of soluble Cr(VI) by Desulfovibrio desulfuricans, D. vulgaris is
mediated by enzymatic reactions [135]. Soluble enzymes are thought to be responsible
for the reduction of chromate by Bacillus sp., Pseudomonas sp., and E. Coli [136].
Principles of bioremediation processes
45
Aerobic reduction is thought to be a detoxification process where cells use a soluble
enzyme to reduce Cr(VI) to Cr(III) internal or external to the plasma membrane.
Reduction of Cr(VI) may also proceed through the use of CrO42- as a terminal electron
acceptor during anaerobic respiration [124, 137]. The produced Cr(III) can be considered
less mobile as it can be precipitated in the form of insoluble Cr(OH)3. However,
experimental results have shown that in the presence of many organic molecules Cr(III)
remains mobile forming stable soluble metal organic complexes [137-139]. Once
reduced, Cr(III) cannot be oxidized by microbes back to Cr(VI). When considering
natural attenuation of Cr(VI), care should be taken that there are no oxidized forms of
manganese in the aquifer matrix (e.g., MnO2). Such minerals are known to abiotically
oxidize and remobilize Cr(III) back to Cr(VI) [15, 127].
Microbial reduction of soluble U(VI) could be an attractive alternative for in situ
bioremediation of uranium-contaminated groundwater [140-145]. It has been
demonstrated that soluble U(VI) in the form of uranyl ions UO22+ can be reduced to
insoluble U(IV) by microorganisms belonging to the genera Geobacter, Shewanella as
well as Desulfovibrio desulfuricans, Pseudomonas fluorescens, and Deinococcus
radiodurans [143-145]. U(IV) is finally immobilized as black UO2(s) precipitate [146,
147]. However the process is sensitive to nitrate concentration in the ground water,
because nitrate compete uranyl ions in the bioreduction process and affects the stability
of U(IV) precipitate [133, 148, 149].
Enzymatic reduction of Mo(VI) (as molybdate, MoO42-) to Mo(IV) by D. desulfuricans
with both lactate and hydrogen as electron donors has been also reported [141, 150].
Mo(VI) reduction in the presence of sulphide results in the extracellular precipitation of
the black mineral phase, molybdenite MoS2(s).
The most important transformation process in the environmental fate of mercury is
biotransformation. Any Hg form entering sediments, groundwater, or surface water under
the appropriate conditions can be microbially converted to the methylmercuric ion.
Sulfur-reducing bacteria are responsible for most Hg methylation in the environment,
with anaerobic conditions favouring their activity. The methylation of elemental Hg plays
a key role in environmental cycling of Hg. Methylated Hg, the most common Hg form, is
mobile and readily taken up by organisms including some higher plants. Humic
substances are known to mediate the chemical methylation of inorganic Hg by releasing
labile methyl groups [29]. Mercury reduction may be a mechanism for detoxification of
media containing mercury. Reduction of Hg2+ to elemental mercury occurs quite readily
and is enhanced by bacterial enzymes [92]. Energy is probably not obtained by use of
Hg2+ as an electron acceptor [124, 133].
Oxyanions of selenium (SeO42-, SeO32-) can be used in microbial anaerobic
respiration as terminal electron acceptors providing energy for growth and
metabolism. Their reduction can be coupled to a variety of organic substrates, e.g.,
lactate, acetate and aromatics, with the bacteria found in a range of habitats and not
confined to any specific genus. These organisms, and perhaps even the enzymes
themselves, may have applications for bioremediation of selenium contaminated
environments. Microbial reduction of the soluble oxidized form of selenium, Se(VI),
to insoluble elemental selenium, Se0, is possible by microorganisms that conserve
energy to support growth from Se(VI) reduction [141, 151, 152]. This natural
mechanism can be used for the removal of selenium from contaminated surface and
groundwater. This metabolism may be employed in the environment for selenium
bioremediation. Details on the biogeomicrobiology of selenium can be found in the
literature [29, 127, 133, 153, 154].
46
Artin Hatzikioseyian
Bioreduction of elements does not always reduce their mobility and could be
undesirable. For example, under anaerobic conditions, As(V) in the form of arsenate
(AsO43-) may also serve as an alternate electron acceptor and be reduced to more toxic
and more mobile As(III) in the form of arsenite (AsO2-). This process is reversible
because arsenite can be reoxidized to arsenate and thus re-immobilized under aerobic
conditions. Similarly, Mn(IV) is reduced to the more toxic and more mobile Mn(II) under
anaerobic conditions, with a reaction which is also reversible [15]. Details on the
biogeomicrobiology of arsenic can be found in the literature [29, 127, 133, 155-158].
Biosorption
Biosorption can be defined as the selective sequestering of metal/metalloid or
radionuclide soluble species by microbial cells that results in immobilization of them.
Metal sequestering by different parts of the metabolically active or inactive cells can
occur via various mechanisms: complexation, chelation, coordination, ion exchange,
precipitation, reduction [159, 160]. Biosorption is a process with some unique
characteristics. It can effectively sequester dissolved metals from very dilute complex
solutions with high efficiency. This makes biosorption an ideal candidate for the
treatment of high volume low concentration complex waste-waters. However, today
biosorption is not considered as a competitive stand alone technology because the
industrial applicability of the process is rather limited and pilot applications have shown
the limitations associated with the use of inactive microbial biomass mainly due to the
cost of formulating it into an appropriate biosorbent material. Furthermore, high
concentrations of co-ions present in the treatment water have negative effect on the
uptake of the targeted metals by the immobilized microbial biomass, and the reduced
resilience of the biological material, made recycling and reuse of the biosorbent even
more difficult. However in the cases of metabolically active microbial cells, biosorption
contributes in the overall sequestering and immobilization of metal ions as a parallel
mechanism together with other metabolically mediated mechanisms such as
bioprecipitation and bioreduction [159-163].
Phytoremediation
Phytoremediation is defined as the engineered use of vegetation to contain, sequester,
extract, accumulate, remove, degrade and/or detoxify inorganic and organic contaminants
from soils, sediments, surface waters, and groundwater [29, 53]. Plant-based remediation
systems are generally considered passive, low-cost, low-technology processes and
employ common plants including trees, vegetable crops, grasses, and even annual weeds
to treat heavy metals, inorganic ions, radioactive elements, and organic compounds.
Phytoremediation is considered as an alternative bioremediation option appropriate for
soils having properties that impede the success of conventional technologies (e.g., low
permeability, saturation, dense structure, mixtures of contaminants). When the
appropriate plants are cultivated in contaminated soils the root system functions as a
dispersed uptake system. Contaminants are taken up with soil water and degraded,
metabolized, and/or sequestered in the plant, while evapotranspiration from aerial parts
maximizes the movement of soil water through the plant. Certain plants have been
identified that can take up and concentrate metals and other inorganic molecules from
soil into leaves, stalks, seeds, and roots. Commonly used hyperaccumulators include
sunflower (Helianthus agnus), Indian mustard (Brassica juncea), crucifers (Thlaspi
Principles of bioremediation processes
47
caerulescens, T. elegans), violets (Viola calaminaria), serpentines (Alyssum bertolonii),
corn, nettles, and dandelion, [15]. The cultivated plants subsequently can be harvested
and treated by incineration, composting, or anaerobic digestion to concentrate and/or
recover the pollutants [29].
Phytoremediation technology can be divided into of two main broad categories
depending upon whether the removal of contaminants or stabilisation of geochemical
conditions in the soil is accomplished. Many different mechanisms for pollutants uptake
removal or stabilization have been identified [8, 15, 92]: In phytodegredation, organic
pollutants are converted by internal or secreted enzymes into compounds with reduced
toxicity. Like phytodegradation, rhizosphere degradation or rhizodegradation involves
the enzymatic breakdown of organic pollutants, but through microbial enzymatic activity.
These breakdown products are either volatilized or incorporated into the microorganisms
and soil matrix of the rhizosphere. The types of plants growing in the contaminated area
influence the amount, diversity, and activity of microbial populations thus there is a
direct independence between the planted vegetation and the dominant microbial species
in the soil matrix of the rhizosphere. Phytoextraction involves the removal of toxins,
especially heavy metals (e.g. Cd, Ni, Hg) metalloids (e.g Se) and radionuclides, by the
roots of the plants with subsequent transport to aerial plant organs. Plants can also
remove toxic substances, such as organics, from the soil through phytovolatization. In
this process, the soluble contaminants are taken up with water by the roots, transported to
the leaves, and volatized into the atmosphere through the stomata. Rhizofiltration
removes contaminants from water and aqueous waste streams, such as agricultural runoff,
industrial discharges, and nuclear material processing wastes. Absorption and adsorption
by plant roots play a key role in this mechanism, and consequently large root surface
areas are usually required. Finally, phytostabilization can be used to reduce the erosion,
leaching and mobilization of soil contaminants which result in aerial or waterborne
pollution of additional sites. In phytostabilization, accumulation by plant roots or
precipitation in the soil by root exudates immobilizes and reduces the availability of soil
contaminants. Plants growing on polluted sites also stabilize the soil and can serve as a
groundcover thereby reducing wind and water erosion and direct contact of the
contaminants with animals. An example of a simple phytoremediation system in use for
many years is the constructed wetland, in which aquatic plants such as water hyacinths
are cultivated to remove contaminants (metals, nitrate, etc.) from municipal or industrial
wastewater [29].
Phytoremediation has been proven useful for soils contaminated with relatively
immobile contaminants to shallow depths and for the case of organic pollutants is
generally applicable for moderately hydrophobic material. Examples of these are toluene,
benzene, PAHs, xylenes, ethylbenzene and many chlorinated solvents. Organic
compounds can be degraded or immobilized in the root zone or incorporated into shoot
tissues and metabolized.
The advantages of phytoremediation technology are summarized below [44]:
•
•
•
•
Operates in-situ and is solar driven.
Costs are approximately 20 to 30% of costs associated with mechanical treatments.
It has high public acceptance.
It is applicable to many remediation scenarios including large contaminated surface areas.
However, some potential disadvantages associated with phytoremediation/plantassisted remediation techniques have been identified [44, 55, 61]:
48
•
•
•
•
•
•
•
•
•
•
•
•
•
•
Artin Hatzikioseyian
Treatment is generally limited to shallow soils within 90 cm from the surface and
groundwater within 3 m from the surface.
High concentrations of hazardous materials can be toxic to plants.
It involves the same mass transfer limitations as other biotreatment technologies.
Treatment time is relatively long (usually more than one growing season).
It may be seasonal, depending on location.
Climatic or hydrologic conditions may restrict the growth rate of certain plants.
Contaminants may enter the food chain via animals (herbivores) or insects that
consume plant material containing contaminants.
Degradation products may be mobilized into ground water or bioaccumulated in
animals.
It can transfer contamination across media, e.g., from soil to air.
It is not effective for strongly sorbed (e.g., PCBs) and weakly sorbed contaminants.
Disposal of secondary waste arising from the harvest of plants is problematic.
The toxicity and bioavailability of biodegradation products is not always known.
It is still in the demonstration stage.
It is unfamiliar to regulators.
Concerning cost aspects, phytoremediation utilizes solar energy, thus it requires less
energy inputs. This factor reduces operating costs. In addition as phytoremediation is
slower treatment process and last longer, expenses are also spread out over a greater time
period than other technologies. The result is lower annual costs. However,
phytoremediation does have several cost components that are unique to the technology.
These components often include the following [44]:
•
•
•
•
•
•
•
•
•
•
Plant or tree stock and/or seeds
Fertilizers, pesticides, and additional soil amendments
Agricultural equipment for amendments application, tilling, and/or harvesting
Irrigation equipment and a water source
Tubes for stimulating deep root growth (collars)
Pest control devices
Supplies, equipment, and/or analyses for testing plant tissues and environmental
conditions
Flow control devices
Plant litter collection, maintenance, pruning, mowing, and/or harvesting
Disposal of plant wastes
The literature cited in this chapter gives more details concerning phytoremediation
technology [11, 29, 35], phytoremediation of metals and radionuclides [9, 29],
phytoremediation cost data as well as comparative economic values between
phytodegradation and a conventional pump-and-treat method [44].
Conclusions
Bioremediation is a multidisciplinary technology and successful application requires
deep understanding of all the relevant scientific fields and attenuation processes. It seems
that nowadays we have entered in the most interesting and intense phase of process
development. Potentials and limitations of the technology are well documented in many
Principles of bioremediation processes
49
resources from the web, books and research papers. Generic and technical information
are given in details. The experience accumulated over the years is promising to design
cost effective successful remediation projects.
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