BIOAUGMENTATION AND ITS APPLICATION IN

WASTEWATER TREATMENT: A REVIEW

M. Herrero1,2 and D. C. Stuckey1(*)

1 Department of Chemical Engineering, Imperial College, London, SW7 2AZ, UK

and, 2Department of Chemical Engineering and Environmental Technology,

University of Oviedo, Spain.

(*) Corresponding author: Stuckey, D.C. (d.stuckey@imperial.ac.uk)

Chemical Engineering Department, Imperial College London, LONDON SW7 2AZ,

UK

Tel: +44 (0)207 594 5591 Fax: +44 (0) 207 594 5629

2

Abstract

Bioaugmentation (the process of adding selected strains/mixed cultures to

wastewater reactors to improve the catabolism of specific compounds, e.g. refractory

organics, or overall COD) is a promising technique to solve practical problems in

wastewater treatment plants, and enhance removal efficiency. The potential of this

option can now be enhanced in order to take advantage of important advances in the

fields of microbial ecology, molecular biology, immobilization techniques and

advanced bioreactor design. Reports on bioaugmentation in WWT show the difficulties

in evaluating the potential parameters involved, leading frequently to inconclusive

outcomes. Many studies have been carried out on the basis of trial-and-error

approaches, and it has been reported that reactors bioaugmented with pure cultures often

fail to perform as well as the pure cultures under laboratory conditions. As an

interesting technical challenge, the feasibility of bioaugmentation should ultimately be

assessed by data from field implementation, and this review highlights several

promising areas to explore in the future.

Keywords: bioaugmentation, biological wastewater treatment, microbial community

dynamics, advanced bioreactors, immobilization.

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1. Introduction

Advanced bioreactor design and operation in wastewater treatment plants

(WWTPs) is essential in order to develop proper environmental conditions so that the

most desirable microorganisms can be selected for and maintained under adequate

physiological conditions. This extends to the contact with the stream to be treated,

controlling mass transfer and reactor performance to achieve the treatment goal at full

scale. Until recently, biological reactors in wastewater treatment have been considered

as “black boxes”, their functionality depending only on empirical knowledge, and hence

they are difficult to predict and control. Even now their performance can only be

evaluated by applying material balances. Now advanced bioreactor design can take

advantage of compartmentalization and the use of membranes to confine the key

microorganisms in the process to a bioreactor (Barber and Stuckey, 1999; Vyrides et al.,

2010). However, it is also known that effective operation of biological treatment

systems relies on highly active microorganisms carrying out the process, and therefore

on the ratio of resistant/sensitive microorganisms to the contaminant being treated

(depending on both its chemical nature and concentration). Microorganisms can degrade

a wide variety of organic contaminants and can adapt to many different inhospitable

environments. However, there are a variety of pollutants, both man made (xenobiotic)

and natural, that are not easy to degrade biologically, even over long periods of time -

these are referred to as “refractory or recalcitrant”.

In principle, poor bioreactor performance may be due to the lack of a sufficient

number of a specific microorganism harbouring a key metabolic route to transform the

target contaminant into less harmful end products. However, despite functional

redundancy (i.e. different strains that can carry out the same or similar functions), and

metabolic versatility being common in environmental bioprocesses, it is more likely that

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the specific target contaminant (or mixture of compounds) can only be degraded by a

very specific mixture of microorganisms (a consortium) harbouring the key metabolic

pathways, and cooperating in a synergistic way. The importance of microbial consortia

is such that, without their combination, many biodegradation processes could not be

explained thermodynamically on the basis of free energy in a single chemical reaction,

explaining ultimately the shift in fermentation patterns (Melandri, 1997; Jorgensen and

Gallardo, 1999).

Despite non-specific bioaugmentation being used for years in agriculture and

wastewater treatment to reinforce the process, it is still considered a procedure with

unpredictable outcomes. Even now bioaugmentation is considered less predictable and

controllable than other removal techniques which result in the direct destruction of

contaminants (Boon et al, 2000). The approach of bioaugmentation focuses on taking

advantage of microbial consortia designed for the specific physico-chemical properties

of the bioprocess, (van der Gast et al., 2004) since this approach was shown to be more

efficient than using undefined inocula. Potentially higher efficiencies could occur in

systems such as membrane bioreactors which stop bacteria being “washed out”, in

contrast to the natural environment where the environmental conditions can be

manipulated to enhance survival and prolong the activity of the exogenous population

(El Fantroussi and Agathos, 2005).

Bioaugmentation of wastewater treatment has not been reviewed specifically for a

number of years, while the current knowledge base has changed rapidly. Metagenomics

(or Ecogenomics – the application of genomics to ecological and environmental

sciences- (for a review see Maphosa et al., 2010), other molecular methods, microscopy

and flow cytometry- based methods are providing an enormous source of information

for monitoring, detection, quantification and characterization of microorganisms. This

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opens up the possibility of exploring unculturable microorganisms, and exploiting

biodiversity as a means to increasing biodegradation, even through genetic engineering

(though the impact of genetically modified microorganisms on the environment is still

unknown). The question is to determine not only the systematic and taxonomical

structure of the microbial consortia taking part in environmental bioprocesses

(community structure), but also to learn how these biological systems respond to

changes in the influent (understanding the biochemistry and toxicity mechanisms

involved), how the microbial population dynamics evolve (community ecology and

assembly), and in which way these changes can be related to bioprocess efficiency

(community function).

2. Community ecology and assembly

It should be noted that community assembly in biological reactors is a very complex

process since different microorganisms comprising a multitude of cells work together to

treat the effluent, interacting by means of different types of cell signalling such as

quorum sensing and horizontal gene transfer (Verstraete, 2007).

At the moment a complex community cannot be engineered rationally (Curtis et al.,

2003), since is still poorly understood (Sloan et al., 2006). Theoretical ecologists such

as MacArthur and Wilson (1967) and May (1974) assumed microbial communities to be

dynamic systems which were nonlinear, and predicted community stability or

instability. Lawrence and McCarty (1970) developed equations based on biochemistry

and microbial kinetics that enabled key process variables to be predicted in two

different biological systems, aerobic and anaerobic. It has been postulated that dynamic

population behaviour is probably innate to microbial systems. Curtis et al. (2003) also

suggest that it is possible “steady-state” conditions never really exist in bioprocesses

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which have been engineered, even in processes with constant inputs or no apparent

variations in growth conditions. Immigration and chance are important in shaping the

behaviour seen in prokaryotic communities (Sloan et al., 2006). Interestingly, it has

been reported that biological communities in similar econiches may have contrasting

dynamic behaviour depending on the specific chemical, physical, or biological

conditions which the culture operates under (Ayala-del-Rio et al., 2004).

In order to advance functional organization, the usefulness of Pareto-Lorenz

diagrams have been applied to measure microbial diversity graphically to represent the

structure of a bacterial community. The Pareto 80/20 principle applied in

macroeconomics has been established where 20% of the participants normally acquire

80% of all the energy (labour) flux (Dejonghe et al., 2001). By using Pareto-Lorenz

distribution patterns, Mertens et al. (2005) showed that only a small group of ammonia

oxidising bacteria (AOB) played a dominant role in nitrification, while the less

dominant species remaining were postulated to embody a reserve of AOB which could

grow up to replace the species which dominated (Wittebolle et al., 2008). Wittebolle et

al. (2005) also concluded that certain species which were not numerically very abundant

seemed very active according to their RNA signal, so an rRNA evaluation is useful

when studying the performance and activity of a biological WWT.

3. The aim of bioaugmentation

In order to establish bioreactors which can effectively treat chemically mixed

wastes it is essential that they harbour and stably maintain key microbial consortia with

sufficient activity to degrade the contaminant compounds present (van der Gast et al.,

2004). Communities were identified that had the ability to colonize such harsh niches

with desired catabolic traits, and these provided an opportunity to develop specialised

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inocula that could be exploited by bioaugmentation (van der Gast et al., 2002).

Bioaugmentation is using added microorganisms to “reinforce” biological waste

treatment populations so that they can effectively reduce the contaminant load by

transforming it into less dangerous compounds. However, the introduced strain or

strains may fail to grow or to be active in the bioreactor due to predation or competition

with the indigenous microbiota, presence of bacteriophages, or to a lack of acclimation

to the environmental conditions. In these cases the introduced microorganisms will not

colonize or show any population dominance in such a dynamic context which is

subjected to changing conditions such as starvation periods, pH or load variations, low

temperatures, etc. Microbial degradation of contaminants can also be enhanced by

biostimulation, i.e. adding nutrients or electron acceptors to activate the indigenous

microbiota. Obviously, a profound knowledge of the biosystem is required to support

decision making at any time in this process.

The most common options for performing bioaugmentation (El Fantroussi and

Agathos, 2005) are; the addition of a pre-adapted pure bacterial strain, addition of a pre-

adapted consortia, the introduction of genetically engineered bacteria, or the addition of

biodegradation-relevant genes that are packaged into a vector in order to be transferred

by conjugation into microorganisms already present in the biosystem. In this latter case,

the technique has the advantage of not depending on the survival or growth of the donor

strain/s. However, some of these options have only been performed at the laboratory

scale (El Fantroussi and Agathos, 2005).

When reporting on how effective bioaugmentation is, a clear definition of the

treatment goal must be laid down in order to identify whether it aims to reduce a

particular contaminant, or remove the total overall organic load (total organic carbon,

TOC). A particular contaminant may only amount to a small fraction of the total TOC

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load, however, it may be the major cause of toxicity for the microbial community in the

biosystem unless it can be reduced to the required effluent standard. Effective

bioaugmentation strategies should achieve a quick diminution in toxicity to the

microbial community present, which in turn may help to select “co-operators” for

treating complicated wastewater by synergism (Wang et al., 2009) in comparison with

non bioaugmented reactors. Nevertheless, this effect is not always guaranteed since it

has been reported that effective bioaugmentation strategies can succeed in removing a

particular contaminant, for which the specific microorganisms were selected for on the

basis of their biodegradation capacity, but fail in enhancing the degradation of total

organic carbon (TOC) (Yu and Mohn, 2002). These authors explained that some of the

characterized degraders are apparently nutritional specialists who can only use a small

number of organics, lacking the ability to degrade the majority of the organic

compounds in the wastewater; obviously, this situation does not fulfil all the treatment

goals. However, wider nutritional capacities would allow the bioaugmented

microorganism to shift their metabolism to using other organic substrates rather than the

target pollutant under real operating conditions. Thus, specific consortia rather than a

single strain may be more useful in ensuring successful bioaugmentation, even in the

presence of a single contaminant.

Frequently the presence of compounds with different levels of toxicity

complicates the biotreatment of chemically mixed wastewaters. Under these conditions

bioaugmented bacteria may lose their degradation ability by being inhibited by the more

inhibitory pollutants prior to metabolizing the more easily degradable pollutants. A

“selective” bioaugmentation strategy can help to overcome this problem in mixed

wastewaters by separating (by adsorption) compounds showing different levels of

toxicity. Following this line, a strategy combining selective adsorption and

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bioaugmentation has been reported for treating a mixed wastewater of nitrobenzene and

ρ-nitrophenol, passing through an adsorption column (Hu et al., 2008). Without

nitrobenzene inhibition, ρ-nitrophenol could be easily degraded through

bioaugmentation. Nitrobenzene was adsorbed by the resin, and then it could be desorbed

and degraded; desorption partly recovering the binding capacity of the resin. It was

observed that below a minimum value, the nitrobenzene concentration was too low to

support the growth of nitrobenzene-degrading bacteria.

4. Selection criteria

Different bioaugmentation options have been proven useful, ranging from the

use of commercial products, culture collections, and indigenous or exogenous strain/s or

tailor-made consortia. A review of wastewater treatment bioaugmentation (Stephenson

and Stephenson, 1992) examined independent investigations (as opposed to

manufacturers reports) on the use of commercial products (developed as mixed

populations including autotrophs, heterotrophs, facultative anaerobes and aerobes,

intended to serve a wide range of purposes, marketed as ready-to-use frozen, freeze-

dried or liquid preparations) at laboratory and full scale, along with a discussion of

possible reasons for bioaugmentation failures. Among them, the lack of acclimation

(i.e., physiological adaptation mechanisms that allow microorganisms to survive and

remain active under harsh environmental conditions) was highlighted. The authors

remarked that in order to properly evaluate the amount and type of bioaugmentation

needed to enhance plant performance, process conditions, such as flow and treatment

technologies, and the type of microbial toxicity should be taken into account. The main

strategy used was based on isolation of natural environmental samples, selection of the

most efficient strains after screening for their catabolic traits, further selection after the

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use of mutagenic agents, and finally fermentation tests for single strains following

traditional selective enrichment procedures (Stephenson and Stephenson, 1992).

Several papers have postulated a large number of potential candidates for

bioaugmentation, and these can be found in a variety of sources such as Arctic or

Antarctic environments, laboratory-scale reactors fed with samples taken from full scale

plants, from aerobic or anaerobic wastewater systems, culture collections, etc. (Singer et

al., 2005; Wang et al., 2009; Schauer-Gimenez et al., 2010). Nevertheless, only limited

understanding can be obtained on their biodegradation mechanisms in bioreactors

(Wang et al. 2009), their competitiveness, or their influence on indigenous communities

(Yu and Mohn, 2002), which are considerably more complicated processes than can be

mimicked in pure culture assays.

Furthermore, acclimation in scaling up is enormously important for achieving

effective bioaugmentation, as happens with the use of starter cultures in the food

industry (in dairy or alcoholic beverage fermentations at industrial scale). The

complexity of engineered wastewater biotreatment (in terms of fluctuation in volume or

composition) even requires specific efforts at this key stage. In a similar way, the

interest in “priming” has been highlighted (Singer et al., 2005). Priming has been

described as “predisposing an isolate or population of microorganisms to future

conditions in which they are designed to perform a function” (Singer et al., 2005). Thus

acclimation of the microbial consortia should be required in the flow chart of the

process to predispose them better towards the physico-chemical conditions for which

they were selected and designed for achieving efficient performance.

As reported by Yu and Mohn (2002), microorganisms used in bioaugmentation

should meet at bare minimum three criteria: firstly, to be catabolically able to degrade

the contaminant, even in the presence of other potentially inhibitory pollutants;

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secondly, they must persist and be competitive after their introduction into the

biosystem; and thirdly, they should be compatible with the indigenous communities

present. Therefore, candidates should be carefully selected, since only a few organisms

are suited to bioaugmentation, and these are not necessarily from the community

needing bioaugmentation (Yu and Mohn, 2002). An additional important point is that

for obvious reasons, candidates should not be closely related to human pathogens (same

genus and species, as exemplified by some Pseudomonas aeruginosa strains) when used

in field implementation (Singer et al., 2005).

Strain selection for bioaugmentation is clearly a key issue. Singer at al. (2005)

proposed that substantial progress could be made through exploiting current knowledge,

recent techniques in molecular biology, and worldwide access to culture collections. At

the same time, increasing information from published research supports the most

effective way of overcoming barriers in bioaugmentation is to locate organisms from the

same ecological niche as the pollutant (El Fantroussi and Agathos, 2005).

5. Bioaugmentation failures

Different workers have postulated a number of reasons for the failure of inocula

when used for bioaugmentation (Stephenson and Stephenson, 1992; Vogel and Walters,

2001; El Fantroussi and Agathos, 2005). The major claim is that the selected strains

often fail to show under natural environmental conditions the abilities shown in the

laboratory as pure cultures, even after showing resistance to starvation or lack of

degradation repression in the presence of additional nutrients (Boon et al., 2000).

Assays in test tubes or in laboratory scale bioreactors using axenic cultures, even those

strains with exceptional abilities, seem to be unable to mimic the complexity innate to

natural biosystems. The interaction of the inoculated microorganisms with their new

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biological and non-biological environments in terms of activity, survival, and migration

can be critical in the success of any bioaugmentation strategy (El Fantroussi and

Agathos, 2005). Bioaugmentation can change the composition of the indigenous

microbial community, for example by competition or inhibition (van Veen et al., 1997),

and these changes may be either positive or negative. It is frequently reported that,

despite a significant increase in bioreactor performance soon after bioaugmentation,

very often the positive effects observed are only maintained for a short time after

inoculation (Boon et al., 2000; Patureau et al., 2001; Yu et al., 2010).

Additionally, when different workers reported on bioaugmentation research, the

difficulty in characterising, quantifying, and evaluating all of the potential parameters

involved could lead to general conclusions that may not be widely applicable

(Stephenson and Stephenson, 1992). Some reported failures may be due to choosing

bioaugmentation in the first place, despite the fact that the indigenous microbial

community may contain the appropriate catabolic genes to degrade the target compound

(Thompson et al., 2005). Hence the efficacy of published bioaugmentation results may

frequently be inconclusive.

Members of the same genera are not all equally fit for certain tasks, and hence some

could be competitive under a broader range of conditions, while others may only be

suited to very specialised conditions (Thompson et al., 2005). Differentiation at species

level, even among those showing the same catabolic traits, has been reported as the

cause of failure, along with differences at strain level related to persistence in the

system after inoculation, observations which could lead to more efficient

bioaugmentation designs. Wenderoth et al. (2003) carried out bioaugmentation with

different Pseudomonas spp, and with different strains of the same species, all of which

were capable of chloro- and dichloro-benzene degradation. With some strains,

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bioaugmentation resulted in little improvement, but interestingly, differentiation at

strain level in relation to persistence in the biosystem was observed. Both strains

Pseudomonas putida GJ31 and P. putida F1ΔCC could grow even when the natural

microbial community was present, stimulating the degradation of chlorobenzene.

However, it was noted that the P. putida GJ31 population decreased rapidly when the

xenobiotics were depleted, while P. putida F1ACC survived even when the

chlorobenzene had been degraded.

Frequently, problems were encountered with: growth-limiting conditions due to low

substrate concentration; the presence of inhibitory substances in the stream to be treated,

and even released by other microorganisms showing antagonistic effects (such as

antibiotics or bacteriocins); the presence of bacteriophages; poor biofilm forming ability

(Fu et al., 2009), or as a result of adverse operating conditions such as low temperatures

(Stephenson and Stephenson, 1992). Suitable strains for bioaugmentation could survive

for a long time in a specific habitat, resulting in slow but continuous degradation

(Thompson et al., 2005), as characterised by a low Vmax and Km. Apart from the

traditionally simplified dichotomy based on alive cells versus dead cells when only

culturability tests were applied, there is now evidence that different microbial

physiological states can be observed in environmental bioprocesses (Diaz et al., 2010)

such as metabolically active cells, reproductive growing cells, damaged or

permeabilised cells (Nebe-von-Caron et al., 2000).

Not only is selection a key factor in bioaugmentation, but also the way of introducing

and maintaining the selected microorganisms and/or their activities in the complex

community. Obviously, the biomass concentration introduced should be high enough to

allow for the prevalence of the metabolically active cells bioaugmented. The use of an

enriching bioreactor for bioaugmentation purposes was reported as an interesting

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strategy (Saravanane et al., 2001). In this approach, the main reactor was bioaugmented

by the addition of acclimated cells periodically from a separate enrichment-reactor.

As a result of a failed strategy when introducing a strain or consortia, it is possible

to provoke undesirable effects or unbalanced ecosystems. With the intentional

introduction of genetically engineered bacteria, the greater risk is probably their escape

from the bioreactor to the environment, with unforeseeable results. Additionally, after

inoculation with bioaugmented microorganisms it is possible to unbalance the system.

Bouchez et al. (2000) reported on two nitrifying reactors operating under identical

conditions, except that aerobic denitrifying bacteria had been inoculated twice into one

of them. FISH showed that the bacteria which had been initially bioaugmented were

rapidly consumed by protozoa in the bioreactor. The second large inoculation in one

reactor resulted in the ecosystem becoming unbalanced leading to a rapid growth of

protozoa and upsets in nitrification, whereas in the non bioaugmented reactor both

parameters were stable. To solve this problem, two different inoculation strategies were

tested to improve the effectiveness of incorporating the bacteria added to the indigenous

culture. Firstly, chemicals which promoted coagulation and flocculation were added to

the reactor just after bioaugmentation. Secondly, which gave the best results, the

bioaugmented bacteria were encapsulated in alginate beads before being inoculated.

Over time the beads eventually broke-up, although fragments of the alginate containing

the bioaugmented microorganisms were incorporated into the existing biological flocs,

and colonization of the alginate matrix by indigenous bacteria was also observed. Hence

this demonstrated that immobilization in alginate beads offered temporary protection

against grazing, and even after breakage the fragments favoured the attachment of the

bioaugmented microcolonies to the existing flocs.

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It was also reported that after inoculation the bioaugmented population decreased

corresponding to dilution with the fresh feed, eventually resulting in the population

stabilising, but at a lower concentration (McClure et al., 1991). To overcome wash-out,

bioaugmentation can potentially take advantage of compartmentalisation, or the use of

membranes in advanced bioreactors, as a way of keeping the introduced

microorganisms within the biosystem. For example, in the anaerobic baffled reactor

(ABR) the reactor is divided into eight compartments with little mixing between the

biomass in each compartment, promoting separation of the various trophic groups

involved in the process while the gas phases remain separate (Barber and Stuckey,

1999). This design allows for the migration of sensitive anaerobes, such as the

methanogens, from the front of the reactor where unfavourable growth conditions may

occur, to the more protected later compartments. An aerated stage in the penultimate

compartment may oxidise refractory COD and excess sulphide from the anaerobic

stages, enabling nitrification to occur, especially if immobilized cells are used. In the

case of membrane-aerated biofilm reactors (MABRs) (Ohandja and Stuckey, 2007), the

membrane in the reactor supports an active biofilm, maintaining a biofilm of active

bacteria, and the MABR can combine both anaerobic and aerobic layers in the biofilm.

This enables processes to occur simultaneously in the same reactor when aerobic and

anaerobic conditions are needed, such as in the case of nitrification and denitrification

processes, or the degradation of chlorinated organics such as PCE (Ohandja and

Stuckey, 2007).

Failed bioaugmentation at full scale has been reported in coke WWT when using a

consortium of a cyanide-degrading yeast obtained from a culture collection

(Cryptococcus humicolus) and unidentified cyanide-degrading microorganisms

(obtained from the activated sludge of a full-scale coke WWT facility) (Park et al.,

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2008). After laboratory scale cultivation (up to 1.2 m3) enriched for two months with a

huge supply of glucose, KCN and other nutrients, the consortium was inoculated into a

fluidized-bed type process (1280 m3). However, continuous operation of the full-scale

cyanide-degrading bioprocess showed poorer removal efficiency than expected (slow

biodegradation rate of ferric cyanide) owing to the poor settling performance of

microbial flocs, and lack of organic carbon sources within the wastewater. The

concentration of activated sludge within the aeration tanks decreased because of

substantial wash out from the sludge settling tank. In addition, violent bubbling of air

disturbed the formation of a biofilm on the carrier’s surface. The need for further studies

was highlighted in terms of how to solve operating problems in both pilot and full scale

bioaugmentation approaches.

6. Success of bioaugmentation strategies: key factors

6.1. Ecological basis

Before making any decision regarding implementing bioaugmentation, along with

the particular physico-chemical characteristics of the bioprocess, a basic knowledge of

the microbial ecology common to the target biosystem should be acquired in order to

understand the reasons for poor bioreactor performance. As reported by Dejonghe et al.

(2001), it seems necessary to acquire this knowledge in order to understand the main

metabolic processes and, concomitantly, to find out whether a specific species in a

particular ecosystem belongs to the determinative 20% (controlling 80% of the energy

flux) or to the subsisting 80% of the community. These authors state that it is

worthwhile exploring methods of modifying the composition of the ruling fraction so

that other species (or other strains or catabolic genes) might participate in the major

conversion processes. With this approach, the energy flux could be more evenly

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distributed over the community, and even the suppressed bacteria could become active

and fulfil an important role in the community (Dejonghe et al., 2001). In this way it

seems feasible to gain community flexibility in order to shift the electron and carbon

flow through various alternative guilds which is a key factor in enhancing community

function (Fernandez et al., 2000).

6.2. Monitoring techniques

To support successful bioaugmentation strategies, along with enhancing

biodegradation kinetics of the target pollutant, the use of molecular techniques is

required (monitoring both the survival and/or activity of the added microorganisms). In

the last few years different techniques have been used such as PCR-DGGE (Guo et al.,

2010), PCR-temperature gradient gel electrophoresis (TGGE) (Fouratt et al., 2003), T-

RFLP (Tsutsui et al., 2010), ribosomal intergenic spacer analysis (RISA), (Qu et al.,

2009), competitive PCR and Reverse transcription PCR (RT-PCR) (Muttray et al.,

2001), quantitative PCR and DGGE (Watanabe et al., 2002), quantitative PCR and RT-

PCR (Morris et al., 2014), FISH (Patureau et al., 2001; Bartroli et al., 2011), and the use

of marker genes such as the green fluorescence protein (gpf-gene mark) (Boon et al.,

2000; Yu et al., 2010). More detailed information on these techniques is given in Table

1. Even the potential of applying microarrays to environmental studies has been

reported, highlighting the need to face several challenges associated with specificity,

sensitivity and quantification in these types of samples (Zhou and Thompson, 2002).

However, it should be noted that high-throughput sequencing (HTS) is causing a

revolution recently by opening up new pathways in environmental microbiology

(Logares et al., 2012). HTS will provide a new understanding of ecological processes

and microbial community functioning. Applications of single-cell sequencing in

combination with metagenomic analysis have been reviewed recently (Lasken, 2012).

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New powerful bioinformatic tools are improving genome assembly and gene prediction

in anonymous prokaryotic genomes, overcoming bottlenecks in the data analysis work.

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Table 1. Examples of different monitoring techniques used in reported bioaugmentation procedures

Treatment Process Strain/Consortium/plasmid/genes Monitoring techniques Contaminant(s) removed/activity detected Reference

Full scale municipal WWT,

(SBR, oxidation ditch, anoxic-

oxic activated sludge-A/O) at

low temperature

Specialised consortium comprising

Proterobacteria, Bacterioides, Nitrospirales and

Cyanobacteria, and the flocculent bacteria

Bacillus sp. F2 and Bacillus sp. F6

PCR-DGGE

BIOLOG

Organics. Effluents quality: COD (≤660 mg L-1), NH4+-N (≤8

mg L-1)

Guo et al., 2010

Laboratory scale SBRs Commercial product, NBP PCR-TGGE Organics (synthetic wastewater), ammonia removal (until

detection limit)

Fouratt et al., 2003

Laboratory scale activated

sludge reactors

Plasmid pJP4 from Escherichia coli and

Pseudomonas putida (as donors) to indigenous

transconjugants

T-RFLP Organics, (COD removal 84-86%) Tsutsui et al., 2010

Laboratory scale anaerobic

bioreactors

mcrA gene and transcripts, coding the α subunit of

methyl coenzyme M reductase

Quantitative PCR and

RT-PCR

Methanogens dynamics, linking mcrA gene copy number to

methane flux

Morris et al., 2014

Laboratory scale membrane

bioreactor (MBR)

Sphingomonas xenophaga QYY RISA Bromoamine acid (dye) in synthetic wastewater, COD

removal 30-50%, decolorization ratio (supernatant and

membrane effluent) 80%-90%

Qu et al., 2009

Laboratory-scale

batch/chemostat

Pseudomonas abietaniphila BKME-9 Competitive PCR and

RT-PCR

Dihydroabietic acid (DhA), degradation rate two times faster

than control

Muttray et al., 2001

Activated sludge, municipal

treatment plant

phc genes coding for multicomponent phenol

hydroxylase mPH, PhcKLMNP and its

transcripcional regulators, PhcR, and PhcT

(phcTRKLMNP)

Quantitative competitive

PCR

Phenol, average value phenol-oxygenating activity 15.3±0.8

U g-1

Watanabe et al.,

2002

SBR anaerobic-aerobic Microvirgula aerodenitrificans (free cells,

different inoculation levels/continuous supply/

immbolised in alginate beads)

FISH N removal (33-43%); or N-oxides (10.2-13.8 mg L-1), a

concomitant phosphate concentration (0-2 mg L-1 ; no

ammonia.

Patureau et al., 2001

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Pilot scale, continuous airlift

reactors, activated carbon for

biofilm formation

Nitrifying activated sludge (from a pilot plant) FISH TAN oxidation near 100%, max nitrifying capacity at steady-

state conditions 1.3gN L-1d-1 (synthetic wastewater)

Bartroli et al., 2011

Laboratory scale, SCAS reactors Comamonas testosterone I2gfp (chromosomically

gfp-marked)

Autofluorescence, PCR-

DGGE

3-chloroaniline (3-CA) (synthetic wastewater) initially

complete removal, 50% in further operation

Boon et al., 2000

Laboratory scale, SBR Pseudomonas putida ONBA-17gfp Autofluorescence, PCR-

DGGE

o-nitrobenzaldehyde (ONBA) complete degradation, and

COD (96.28%) (synthetic wastewater)

Yu et al., 2010

SBR: Sequencing batch reactor; SCAS: semicontinuous activated sludge; PAO: phosphorus accumulating organisms; TAN: total ammonia nitrogen

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6.3. Success in overcoming operational problems and plant management

Sometimes bioaugmentation is used in WWTPs to overcome a significant reduction

in biotreatment efficiency due to seasonal low temperatures, as typically occurs in

nitrification during winter months. In such cases, the reactor needs a greater biomass

concentration in the winter period than in the summer period. It has been reported that

bioaugmentation allowed for a reduction in the necessary volume needed for a stable

process which could be maintained under conditions that would normally have

prevented nitrification, and moreover, its effects could be predicted by the use of a

simple model (Plaza et al., 2001).

Bioaugmentation was also useful in decreasing the recovery time of anaerobic

digesters transiently overloaded (Tale at al., 2011), or exposed to a transient toxic event

(Schauer-Gimenez et al., 2010). In the latter case, results varied depending on the

source of the inoculum to start-up each set of digesters. These authors claimed that

production and distribution of individual bioaugmentation cultures, enriched to degrade

a specific substrate, would be time consuming. Using a novel approach, it was proposed

that it could be more practical to target a key, ubiquitous intermediate that accumulates

during toxic events. To this aim, hydrogen (H2) was chosen since its degradation is

often a rate-limiting step in methane production from many complex substrates, so it is

possible that the more rapid H2 utilization, the more complete the conversion of

propionate and other substrates to methane. Two sets of laboratory-scale digesters were

transiently exposed to the model toxicant (oxygen): the sets differed only in the source

of the original inoculum employed to start operation. In one set, original biomass was

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taken from a mesophilic anaerobic digester fed synthetic municipal sludge. The other

set was originally started with biomass from a mesophilic anaerobic digester fed

synthetic industrial wastewater. The methanogenic culture used for bioaugmentation

was developed using biomass from a mesophilic municipal anaerobic digester after

enrichment. The archaeal community in the bioaugmented culture was analysed,

showing sequences similar to those of H2-utilizing methanogens. However, one set of

digesters produced lingering high propionate concentrations, and bioaugmentation

resulted in significantly shorter recovery periods, while the second set of bioaugmented

digesters did not display lingering propionate and recovered in the same time as the

non-bioaugmented controls. As stated, the contradiction between these results may be

due to differences in the original microbial communities in the two sets of digesters.

Therefore, it could be observed that the microbial community structure within an

existing biological system is just as important as the community structure of any culture

added when bioaugmentation is implemented.

Successful bioaugmentation strategies can also be essential to achieve better

WWTP management when the existing facilities become insufficient to treat increasing

wastewater volumes. As reported by Ma et al. (2009), as both the amount and type of

petrochemical products increase, the existing anoxic–oxic (A/O) activated sludge

process in the WWTP could not meet the demands of the increasingly complicated

petrochemical wastewater, so it was urgent to develop innovative technologies for

proper treatment. The bioaugmentation option chosen to upgrade the existing facilities

was proven to be efficient.

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6.4. Different successful selection criteria

As mentioned above, the strategy of using tailor-made consortia seems to be a

promising approach. Van der Gast et al. (2004) investigated the effectiveness of a

bacterial consortium composed of four species on the basis of their ubiquity in waste

metal-working fluids (MWFs), their degradation ability, and their tolerance to the

fluctuating conditions. The inoculum was found to represent a significant component of

the community in bioreactors both with and without the presence of indigenous MWF

populations. The reduction in the COD by the consortium was approximately 85% of

the total pollution load, and was 30–40% more effective than any other treatment.

However, tailoring inocula for every case and application would just not be

practicable (Thompson et al., 2005). At least some of the broad commercial

bioaugmentation consortia have been proven to be efficient. For example, a generic

commercial consortium was successfully added by Duran et al. (2006) to enhance the

conversion of biosolids to methane and remove odorous compounds, containing

selected strains of bacteria from the genera Bacillus, Pseudomonas, and Actinomycetes,

along with ancillary organic compounds containing various micronutrients. It is known

that certain commercially available products containing enzymes such as lipases,

proteinases, cellulases, and hydrolases enhance enzymatic break down of

macromolecules. Despite the non-specific nature of this consortium, its addition

enhanced methane production (29%), and reduced propionate levels (~50%) compared

to the control. A commercial product suited for nitrification (nitrifying bioaugmentation

product, NBP, presented as a mixed consortium) was useful in enhancing the activity

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and process efficiency in nitrification. Nevertheless, after efficient bioaugmentation

with NBP at 1%, it was not possible to correlate the observed increase in activity with a

detectable shift in the population by means of TGGE (with higher detection level,

requiring at least 5%), as reported by Fouratt et al. (2003).

In any case, a crucial issue for any commercial bioaugmentation consortium is the

preservation and storage of cultures under the best conditions to achieve higher

survival/activity at inoculation time. To this aim it has been reported that there is

considerable interest in adding cryoprotectants to freeze-drying in air methanogenic

cultures to achieve higher activities (Bhattad et al., 2010).

Bioaugmentation with exogenous genes (phc, which code for phenol hydroxylase

and its transcriptional regulators) transferred to the indigenous species dominating the

biosystem turned out to be a successful strategy (Watanabe et al., 2002). With this

approach it seems possible to enhance the activity of the already established population,

showing good abilities to tolerate the prevailing conditions in the target niche. The

introduction of the target catabolic genes into indigenous microorganisms that are

already adapted to survive and proliferate in the environment better guarantees

persistence versus the technologically challenging survival of the exogenous strains.

Genes from Comamonas testosteroni R5, were introduced into an indigenous strain,

Comamonas sp. rN7, which constituted the dominant catabolic population of the

activated sludge community. The high phenol-oxygenating activity showed by the

introduced transformant, rN7(R503) could be established within the sludge, improving

resistance to phenol-shock loading compared to sludge inoculated with no cells, or to

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the host rN7 or to the foreign phenol-degrading strain R5. Nevertheless, it was observed

that the expression of Phc mPH in rN7(R503) cells was less efficient in the sludge than

in pure culture, highlighting the importance of studying factors that govern gene

expression in natural ecosystems. So it is feasible that dominant populations in

wastewater reactors can be good candidates as hosts for desired catabolic activities that

are to be exogenously introduced to accelerate the natural gene exchange, and

recombination events seeking to spread degradation genes.

6.5. Repeated or continuous application versus single inoculation

Interestingly, a key factor ensuring a successful strategy is that bioaugmentation

must be performed on a regular basis due primarily to the temporary stability of the

newly introduced strains (Boon et al, 2000). In Boon’s work it was considered that the

natural level of the indigenous strain was too low to degrade the target contaminant.

After the strain was introduced, community structure was analyzed using DGGE of the

16S rRNA genes indicating that, even with a strain originating from the ecosystem and

able to grow effectively on a selective substrate, bioaugmentation was not permanent

and would probably require regular resupplementation.

As mentioned earlier, an enriching bioreactor in the flow chart of the process may

be useful to guarantee enough acclimated cells to be continuously or intermittently

introduced into the main reactor. The enricher-reactor can operate not only separately,

but also under different conditions from the main reactor. The biomass produced within

this enricher-reactor where optimum growth conditions were maintained and growth-

supporting substrates were added, turned out to be viable in the main reactor for several

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generations, increasing the biomass concentration. Results demonstrated that this

continuous or intermittently transferred inoculant overcame the difficulty of growing

adequate biomass in the main reactor on inhibitory or toxic substances (Saravanane et

al, 2001). Interestingly, with this strategy it is possible to separate the propagation and

acclimation stages from the biodegradation stage, ensuring high cell densities for

inoculation.

Another interesting strategy is using the slow release of immobilized bioaugmented

cells (growing in moulded agar) into the activated sludge by means of encapsulation in

4 mm diameter open-ended silicone tubes (3 cm long) (Boon et al., 2002). The tubes

containing the immobilized bacteria represented about 1% of the volume of the mixed

liquor. The bioaugmentation activity of a reactor containing the immobilized cells was

compared to a reactor inoculated with suspended cells, revealing that from

approximately 30 days after inoculation the reactor with suspended cells failed to

completely degrade the contaminant because of a decrease in metabolic activity.

However, slow release of the growing embedded cells from the agar into the activated

sludge medium resulted in a higher number of active degrading cells, and was

responsible for nearly 90% degradation.

6.6. Success of immobilization techniques

Immobilization techniques of exogenously added bacterial cells might be a solution

to generate protective barriers around microorganisms, and also to increase metabolic

activity. Entrapment can be an efficient way to protect microorganisms from grazing by

protozoa, as well as reducing biomass loss caused by washout. Although attachment by

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adsorption has been the preferred method of immobilization for wastewater treatment,

much higher cell concentrations can be obtained by encapsulation, despite the

attachment methods requiring no chemical addition. It has also been reported that

nitrification rates were generally higher for polyvinyl alcohol (PVA) encapsulation than

for attachment systems (Rostron et al., 2001). Laboratory-scale continuously stirred

tank reactors containing freely suspended and immobilized biomass were operated with

a high-strength synthetic ammonia wastewater. Nitrifiers are slow growing and have a

low yield, and hence without long retention times they will be washed out of a

continuous reactor unless they are kept in by immobilization. A high cell concentration

is possible with immobilization, so the volumetric efficiency is greatly increased. This

can lead to relatively small reactors, and may afford protection from toxic shocks and

adverse temperatures. The freely suspended nitrifiers were washed out of the reactors at

a 1 d hydraulic retention time (HRT), whereas the reactors containing adsorption

particles and PVA-encapsulated nitrifiers continued partially nitrifying down to 12 h. At

that time all reactors suffered a loss of nitrification, with the PVA reactor maintaining

the highest nitrification rate and 30% full nitrification to nitrate.

Other matrices have also been used successfully. Whole-cell immobilization of

selenate-respiring Sulfurospirillum barnesii in polyacrylamide gels was used to treat

selenate contaminated synthetic wastewater with a high molar excess of nitrate (1,500

times) and sulphate (200 times). To validate the bioaugmentation success under

microbial competition, gel cubes with immobilized S. barnesii cells were added to an

upflow anaerobic sludge bed (UASB) reactor, resulting in earlier selenate and sulphate

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removal and higher nitrate/nitrite removal efficiencies compared to a nonbioaugmented

control reactor. The selenate reducing activity was maintained during long-term

operation, and molecular analysis showed that S. barnesii was present in both the sludge

bed and effluent. It could be demonstrated that gel immobilization of specialised

bacterial strains can survive wash-out and out-compete newly introduced strains in

continuous bioaugmented systems (Lenz et al., 2009).

An anaerobic sequencing batch biofilm reactor (AnSBBR) was inoculated with

enriched sulphate reducing bacteria (SRB) in an alginate-immobilized matrix for the

enhanced treatment of sulphate bearing chemical wastewater. Following a sound

approach, firstly, the technological and ecological base of the failure was

acknowledged. Poor bioreactor performance could be assessed by the accumulation of

volatile fatty acids, and subsequently this effect was attributed to process inhibition due

to the presence of sulphate, and the non-existence of sulphate reducing bacteria (SRB)

and methanogenic bacteria. Consequently, the reactor was augmented with an enriched

SRB consortia entrapped in an alginate matrix; after augmentation the reactor showed

significant enhancement in overall performance (COD removal efficiency and sulphate

reduction). Microbial diversity in a non-augmented reactor showed the dominance of

acetogenic bacteria over the methanogenic and SRB. After augmentation, the microbial

distribution varied significantly: the introduction of an enriched SRB consortia resulted

in competition between anaerobic bacteria in the system being altering leading to an

improvement in process performance. The entrapping matrix protected the consortia

from possible predation from the new environment prior to adaptation, giving sufficient

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time for the inoculated consortia to acclimatise to the new environment. Interestingly, it

was concluded that biofilm configured systems were generally more favourable for

augmentation due to an increased cell density, the high potential for cell to cell contact,

reduced mass transfer limitations, and good retention of the added consortia (Mohan et

al., 2005).

6.7. Advantages of gene transfer methods

Since introduced microorganisms often do not survive following bioaugmentation,

scientists have investigated the use of naturally occurring horizontal gene transfer

processes for the introduction of catabolic genes to treat wastes. It is known that genes

that encode for the degradation of both naturally occurring and xenobiotic organic

compounds are often located on plasmids, transposons or other mobile and/or

integrative elements (Top et al., 2002). Recent advances in genome sequencing are

revealing the substantial role that horizontal gene transfer has played in microbial

development and adaptation in the environment (Ochman, et al., 2000). Horizontal gene

transfer may occur via uptake of naked DNA by competent cells, which has reached this

particular physiological stage in their life cycle (by transformation), mediation by

bacteriophage (general or specialised transduction, in the latter case when transferred

DNA is adjacent to the phage attachment site), or physical contact and exchange of

genetic material such as plasmids or conjugative transposons between microorganisms

(conjugation). The potential advantage of using gene bioaugmentation over traditional

cell bioaugmentation approaches relies on no requirement for long-term survival of the

introduced host strain. The transfer of plasmids, via conjugation, is the technology most

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studied with respect to bioaugmentation, and gene bioaugmentation may also have

applications for metal contaminated sites (for a review see Top et al. 2002; Tsutsui et al.

2010). The authors concluded that successful cases suggest that the strategy could

indeed work under specific conditions, such as when the in situ degradation potential is

absent, and the pollutant degrading transconjugants can grow and become numerically

dominant populations in the bacterial community. Obviously, further studies in this area

are needed to improve current knowledge on the efficiency of gene dissemination as an

effective tool under real operating conditions.

6.8. Success by using membrane bioreactors

Membrane bioreactors (MB) have been used successfully for bioaugmentation

purposes (Qu et al., 2009; see Table 1). The introduction of Sphingomonas xenophaga

QYY (a specialised degrader of the target compound) into the MBR system

significantly increased contaminant removal in comparison to the non-inoculated MBR.

Also, bioaugmentation could accelerate start-up of the MBR and enable it to run well

despite facing sudden toxic pollutant shock loads. These authors highlighted the fact

that the introduced specialized strain was compatible with the indigenous populations.

Also, the long-term performance of a bioaugmented MBR for treatment of textile

wastewater containing different azo dyes has been reported (Hai et al 2011). Stable

decolourization along with significant TOC removal over 7 months under extremely

high dye loadings demonstrated the feasibility of the process. Thus, the use of an MBR

in combination with bioaugmentation techniques can help to overcome the frequent

claim that bioaugmentation is effective but ephemeral in bioreactors.

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6.9. Success at full scale implementation

Application of bioaugmentation to existing industrial WWT facilities has rarely

been reported. Bioaugmentation with mixed cultures of specialized bacteria targeting

various refractory organics was successfully applied to upgrade a full-scale activated

sludge system into a contact oxidation system (Ma et al., 2009). In this case, application

of bioaugmentation was combined with immobilization through the contact oxidation

process taking advantage of compartmentalization. The petrochemical WWTP influent

was a mixed waste stream from an oil refinery and various petrochemical industries, and

the wastewater contained numerous highly toxic and refractory organics (such as

petroleum hydrocarbons, benzene hydrocarbons, aniline, nitrobenzene, phenols as well

as their derivatives). The consortium (mainly consisting of Pseudomonas, Bacillus,

Acinetobacter, Flavobacterium and Micrococcus) was enriched from the activated

sludge of various petrochemical WWTPs and then acclimated. After bioaugmentation,

the start-up time was shorter than in the non-bioaugmented system; besides the rapid

upgrade period, the bioaugmented system also improved the degradation efficiency of

recalcitrant compounds and the resistance to shock loadings. Rapid removal was

facilitated by temporal and spatial multiple stages accomplished by the collaborative

functions of the bacterial communities formed in each compartment. In conclusion, the

authors highlighted that success relied on the survival of the specialized consortia as the

most significant factor, along with adjustment of DO concentration in the biological

tank, which were considered to create the optimum operational conditions for the

growth and reproduction of the inoculated bacteria.

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7. Prospects

With the recent advent of membrane reactors in WWT, the ability to keep

microbial consortia in bioreactors has been enhanced substantially, which should lead to

better performance and better quality effluents. However, increasing interest in

bioaugmentation ought to be assessed by full scale data which to date is still very

scarce.

Emerging areas of research should help in developing new approaches for novel

applications in bioaugmented bioreactors. Protein engineering has a wide potential for

technological applications. It is known that immobilization techniques can give rise to

higher enzymatic rates by achieving higher concentrations and even promoting

conformational changes. It should also be considered that many of the enzymes and

substrates do not function in solution. Instead, cells use a form of ‘solid-state’

biochemistry, and it has been pointed out that chemical reactions are often ‘channelled’

through these structures so that substrates are not freely diffusible and pass from one

modifying enzyme or enzyme complex to neighbouring functional assemblies,

exhibiting much higher efficiencies than is possible to mimic in solution in a test tube

(Ingber, 2010). This idea could be further investigated to enhance bioaugmentation

strategies by cell-to-cell contact (flocs, biofilms). An in vitro compartmentalisation

(IVC) technique has been developed by creating cell-like structures such as water-in-oil

emulsions (droplets) where single genes are encapsulated in an artificial membrane

along with the components required to transcribe and translate them, adding other

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molecules if required (for a review see Taly et al., 2007). An equivalent procedure is to

compartmentalize bacteria into droplets, allowing each droplet to work as an individual

microbioreactor, not limited by the fitness of the organism expressing the enzyme of

interest, thus permitting conditions that would be detrimental to most living organisms.

Nowadays, nanotechnology is an emerging area of research with considerable

potential. Nanomaterials have been extensively used for rapid or cost effective clean-up

of wastes (Shan et al., 2009). Their benefits of application in wastewater treatment are

derived from the nanoparticle characteristics: enhanced reactivity, surface area, sub-

surface transport, and/or sequestration characteristics (Brar et al, 2010). As a practical

application, reverse micellar extraction (RME) has the advantages of ease of scaling-up

and continuous operation. In this case, biomolecules can be recovered from the reverse

micellar phase by exploiting the disassembling nature of RMEs in aqueous media

(George and Stuckey, 2010). A priori, new approaches supported by the current

knowledge could be useful not only for engineered nanomaterials detoxification of a

wide variety of contaminants in waters by adsorptive removal (pesticides, residual

antibiotics, pharmaceutical compounds and endocrine disruptors), but maybe potentially

as delivery systems to facilitate the introduction of gene products in bioreactors treating

highly toxic, refractory or recalcitrant compounds in WWT. However, engineered

nanomaterials could also become water contaminants so concerns regarding the use of

persistent nanoparticles that, on ingestion, will accumulate within the body will require

new methodologies for their evaluation and for establishing adequate criteria for risk

assessment (Brar et al., 2010).

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Single cell analysis (SCA) will be crucial for elucidating cellular diversity and

heterogeneity (Wang and Bodovitz, 2010). A deeper knowledge at the single cell level

(considering the biochemistry and toxicity mechanisms of contaminants) also

complemented at the community level is still needed to understand the success and key

factors controlling bioaugmentation under real operation conditions. It seems evident

that a close interaction between engineering and biotechnology approaches is essential

to promote higher removal efficiencies. It is also necessary that bioreactor design

facilitates the transition from laboratory to full scale operation. Finally, the last step is to

develop proper models, usable at all scales, which are capable of predicting the function

of microbial populations in biological treatment under different process conditions.

8. Conclusions

Community assembly, ecology, and microbial dynamics in bioreactors treating

wastewater are complex processes. Basic knowledge of the biosystem ecology is

required to establish a clear definition of the treatment goal to achieve. Not only should

strain (or tailor-made consortium) selection, but also the way of introducing and

maintaining the selected microorganisms and/or their activities in the community, and

acclimation in scaling up and bioreactor design must be considered. Emerging areas of

research should help in developing new bioaugmented bioreactors: protein engineering;

in vitro compartmentalization; nanomaterials for enhanced reactivity, surface area,

and/or sequestration characteristics; and, single cell analysis and cell-cell interactions

under real operating conditions.

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9. Acknowledgements

M.H acknowledges the support of the Research Vice Chancellor’s Office,

University of Oviedo, for her stay in the Department of Chemical Engineering, Imperial

College, London.

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Table 1. Examples of different monitoring techniques used in reported bioaugmentation procedures

Treatment Process Strain/Consortium/plasmid/genes Monitoring techniques Contaminant(s) removed/activity detected Reference

Full scale municipal WWT, (SBR,

oxidation ditch, anoxic-oxic activated

sludge-A/O) at low temperature

Specialised consortium comprising Proterobacteria,

Bacterioides, Nitrospirales and Cyanobacteria, and the

flocculent bacteria Bacillus sp. F2 and Bacillus sp. F6

PCR-DGGE

BIOLOG

Organics. Effluents quality: COD (≤660 mg L-1), NH4+-N (≤8 mg L-1) Guo et al., 2010

Laboratory scale SBRs Commercial product, NBP PCR-TGGE Organics (synthetic wastewater), ammonia removal (until detection limit) Fouratt et al., 2003

Laboratory scale activated sludge

reactors

Plasmid pJP4 from Escherichia coli and Pseudomonas putida

(as donors) to indigenous transconjugants

T-RFLP Organics, (COD removal 84-86%) Tsutsui et al., 2010

Laboratory scale anaerobic bioreactors mcrA gene and transcripts, coding the α subunit of methyl

coenzyme M reductase

Quantitative PCR and RT-PCR Methanogens dynamics, linking mcrA gene copy number to methane flux Morris et al., 2014

Laboratory scale membrane bioreactor

(MBR)

Sphingomonas xenophaga QYY RISA Bromoamine acid (dye) in synthetic wastewater, COD removal 30-50%,

decolorization ratio (supernatant and membrane effluent) 80%-90%

Qu et al., 2009

Laboratory-scale batch/chemostat Pseudomonas abietaniphila BKME-9 Competitive PCR and RT-PCR Dihydroabietic acid (DhA), degradation rate two times faster than control Muttray et al., 2001

Activated sludge, municipal treatment

plant

phc genes coding for multicomponent phenol hydroxylase

mPH, PhcKLMNP and its transcripcional regulators, PhcR,

and PhcT (phcTRKLMNP)

Quantitative competitive PCR Phenol, average value phenol-oxygenating activity 15.3±0.8 U g-1 Watanabe et al., 2002

SBR anaerobic-aerobic Microvirgula aerodenitrificans (free cells, different inoculation

levels/continuous supply/ immbolised in alginate beads)

FISH N removal (33-43%); or N-oxides (10.2-13.8 mg L-1), a concomitant

phosphate concentration (0-2 mg L-1 ; no ammonia.

Patureau et al., 2001

Pilot scale, continuous airlift reactors,

activated carbon for biofilm formation

Nitrifying activated sludge (from a pilot plant) FISH TAN oxidation near 100%, max nitrifying capacity at steady-state

conditions 1.3gN L-1d-1 (synthetic wastewater)

Bartroli et al., 2011

Laboratory scale, SCAS reactors Comamonas testosterone I2gfp (chromosomically gfp-marked) Autofluorescence, PCR-DGGE 3-chloroaniline (3-CA) (synthetic wastewater) initially complete removal,

50% in further operation

Boon et al., 2000

Laboratory scale, SBR Pseudomonas putida ONBA-17gfp Autofluorescence, PCR-DGGE o-nitrobenzaldehyde (ONBA) complete degradation, and COD (96.28%)

(synthetic wastewater)

Yu et al., 2010

912913

41

SBR: Sequencing batch reactor; SCAS: semicontinuous activated sludge; PAO: phosphorus accumulating organisms; TAN: total ammonia nitrogen

914

915