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“ANAMMOX” A NOVEL PROCESS FOR NITRGOEN MANAGEMENT IN BIOREACTOR LANDFILLS – A REVIEW
Obuli. P. Karthikeyan1* and Kurian Joseph2 Centre for Environmental Studies, Anna University, Chennai – 25.
ABSTRACT
Landfilling is still a popular way for Municipal Solid Waste (MSW) disposal. Leachate
generated from landfills is becoming a great threat to surrounding as it contains high
concentration of organic and toxic pollutants. In recent years, due to the advances in knowledge
of landfill behavior and decomposition processes of MSW, there has been a strong thrust to
upgrade existing landfill technology from a storage or containment concept to a process based
approach as a bioreactor landfill. Increasing attention is being given to leachate recirculation in
landfill bioreactor as an effective way to enhance microbial decomposition of biodegradable solid
waste. High concentrations of ammonia in the leachate may become a hindrance to the effective
functioning of bioreactor landfills. Thus, the stabilization of landfill leachate with respect to
ammonia is likely to be the factor that determines when the overall landfill can be considered
stable.
The most likely scenario for ammonia removal is the aerobic treatment of leachate
outside of the landfill to convert ammonia to nitrate, followed by use of the landfill as an
anaerobic bioreactor for denitrification. Anaerobic ammonia oxidation (ANAMMOX) is a novel
process in which nitrite is used as the electron acceptor in the conversion of ammonium to
nitrogen gas. The ANAMMOX process offers great opportunities to remove ammonia in fully
autotrophic systems with biomass retention. No organic carbon is needed in such nitrogen
removal systems, since ammonia is used as electron donor for nitrite reduction. This paper
reviews and summarizes the anaerobic solublization of nitrogen in landfill environment, recent
developments in nitrogen removal, microbial aspects (occurrence, classification, physiology,
biochemistry) of ANAMMOX, followed by a qualitative comparison of several components of
ANAMMOX technology with conventional nitrogen removal systems and finally addresses the
application of the ANAMMOX process for nitrogen management in bioreactor landfills.
Key words: Bioreactor landfills, solid waste, Leachate recirculation, nitrogen
solublization, nitrogen management and ANAMMOX
INTRODUCTION
Municipal Solid Waste (MSW) is one of the major sources of air, water and
soil contamination (Yu et al., 2002). There is a need for alternative waste
management techniques to better utilize the waste and minimize its adverse
environmental impacts. There is a great deal of interest for alternative waste
management techniques, which can accelerate the anaerobic decomposition of
the organic fraction of the MSW (Sponza and Osman, 2005). Landfilling,
composting and co-digestion of solid wastes with sewage sludge represent the
most economical methods for the disposal/treatment of MSW. However,
increasing attention is being given to landfilling in recent years within the waste
management hierarchy (Sosnowski et al., 2003). After depletion of the limited
volume of air available in void spaces of waste bed, decomposition in a landfill
takes place under anaerobic conditions. Biological degradation requires moisture
* Corresponding author 1 Research Scholar – opk_ens@Yahoo.co.uk 2 Assistant Professor – Kuttiani@vsnl.com
and lack of water is generally responsible for retarding stabilization of MSW in
conventional landfills (Chugh et al., 1998). The advantages of leachate
recirculating systems have been well documented at the both bench scale and
pilot scale (Reinhart and Townsend, 1998).
Landfill leachate typically contains high concentrations of ammonia-
nitrogen long after the BOD and COD have decreased to concentrations
representative of well-decomposed refuse. MSW has been reported to contain
3.8 - 4.2 % protein (Barlaz et al. 1990). When protein decomposes under
anaerobic conditions, ammonia, which is very stable under anaerobic conditions,
is produced. Through ammonification and solubilization, nitrogen is removed
from refuse and accumulates as ammonia in leachate (Burton and Watson-Craik
1998). Estimation of the total release of nitrogen requires knowledge about the
proportion of the total nitrogen susceptible to hydrolysis and the rate of its
subsequent ammonification, as well as the transfer of the end-product of
ammonia into leachate (Jokela and Rintala, 2003)
In addition, leachate recirculation in bioreactor landfill may also increase
ammonia concentration; in turn it will affect the waste stabilization process in the
reactors. Thus, the treatment of landfill leachate to remove ammonia-nitrogen is
likely to be a significant factor that determines when a landfill can be considered
stable.
BIOREACTOR LANDFILL CONCEPT
Many solid waste landfills in developing countries are now being designed
or modified to permit more rapid stabilization of refuse. These “bioreactor
landfills” (Figure 1) are constructed similar to most sanitary landfills, but have
increased moisture content, which allow for an increased rate of biological
activity. This concept was an outgrowth of the practice of leachate recirculation,
which was originally designed as a method to limit the discharge of leachate and
to improve its quality. The leachate recirculation studies showed that leachate
recirculation improved leachate quality, but in addition, the time needed for
landfill stabilization was substantially reduced due to the increased moisture
content (Kelly et al., 2006).
The practice of leachate recirculation was studied during the 1970s and
early 1980s as a method of leachate treatment (Pohland, 1975; Leckie et al.,
1979; Ham and Bookter, 1982). More recently, full-scale bioreactor studies have
been undertaken to determine the relevant design and operational parameters
for these systems (Reinhart and Townsend, 1998; Youcai et al., 2002; He et al.,
2005; Sponza and Osman, 2005; and Wang et al., 2005).
Figure 1. Schematic of Bioreactor landfill design
Regardless of the importance of nitrogenous emissions, landfill bioreactor
studies have largely focused on the factors that affect waste methanization
(Barlaz et al., 1990) and on the characteristics of soluble organic compounds in
leachate (Senior et al., 1990). Previously, Burton and Watson-Craik (1998) has
reviewed ammonia and nitrogen fluxes in landfill, focusing on e.g. nitrogen
Leachate collection/storage tank G as collection pipe
Leachate recirculation pipe
Gas flow monitor
Leachate collection pipeLandfill liner
Landfill coverLeachate collection/storage tank G as collection pipe
Leachate recirculation pipe
Gas flow monitor
Leachate collection pipeLandfill liner
Landfill cover
transformations through the nitrification and the denitrification stages of leachate
recirculation. The characteristics of nitrogen in landfill leachate and its removal
have been reviewed by Lema et al. (1988) and Kettunen (1997). Additionally,
nitrous oxide emissions and anthropogenic nitrogen in wastewater and solid
waste have been reviewed by Barton and Atwater (2002). The Total Kjeldahl
Nitrogen (TKN) content of various MSW from landfill and digestion studies in the
literature is presented in Table 1. It varies from 1.2 to 4% of TS in different
anaerobic digestion and landfill studies were reported by several authors. From
the table it was also evidenced that the segregation of putrecibles does not
reduces the nitrogen content in MSW.
Table 1. Nitrogen contents of MSW from landfill and digestion studies
Waste TKN
(% of TS) System Reference
Landfilled MSW samples 1.2 – 3.8 Landfill Ham et al., 1993
Unsorted MSW 3.3 Landfill lysimeter Pohland, 1980
Unsorted MSW 4.0 Laboratory landfill
lysimeter Leuschner, 1989
Putrescible fraction of MSW 3.2 Anaerobic digestion Cecchi et al., 1992
Putrescible fraction of MSW 1.2 – 2.9 Anaerobic digestion Gallert and winter,
1997
Grey waste fraction MSW 1.2 Laboratory landfill
lysimeter Jokela et al., 2001
LEACHATE QUALITY
Leachate composition is influenced by several factors including waste
composition, operational methods, and climatic conditions. Among these, waste
composition is the most important factor. The participation of organic and
inorganic components in biological, chemical and physical processes define the
general leachate characteristics. The higher the content of degradable material in
the waste, the more important is the biological processes.
For inorganic wastes the solubility of various components plays a major
role in determining leachate composition. Waste components and reaction
products are removed from the waste as it is leached or flushed by leachate and
are subsequently transported out of the landfill with the leachate as solutes or as
landfill gas. The waste and the leachate therefore change composition with time,
both as a result of depletion of various components and of changes in the
chemical environment (e.g. redox-potential, pH, sulphides, and ionic strength).
Very little information is available on the time needed to reach final storage
quality for the various types of landfills. For a number of inorganic pollutants
(e.g. ammonia, chlorides, sulphates, trace elements), the changes of
concentration in the leachate are related to the liquid/solid (L/S) ratio (i.e. the
accumulated amount of leachate produced per unit weight of waste deposited)
than to the chronological age.
ANAEROBIC SOLUBLIZATION OF NITROGEN FROM MSW
Generally, unsorted MSW has a high concentration of organic carbon due
to its high cellulose content (10-40% of TS) reported by Ham et al. (1993),
whereas the nitrogen concentration is relatively low (between 1.0 and 4.0% of
TS). Proteins are commonly found in MSW (Barlaz et al., 1990) and thus can be
considered a major source of soluble nitrogen in MSW. Basically, the nitrogen
contained in MSW is not normally removed by the other methods of treatment
and thus waste management as it is carried out at present, especially modern
landfills, harbors a substantial amount of fixed nitrogen. The nitrogen flow from
food production to waste management (Figure 2) and its nitrogenous emissions
has been reviewed by Barton and Atwater (2002). This especially noteworthy as
the release of soluble nitrogen from landfilled waste to environment continues
over a long period. It is generally known from the studies of anaerobic digestion
as well as from the analysis of landfill leachate samples that most of the organic
nitrogen released during anaerobic degradation is converted irreversibly into
ammonium ion or as ammonia.
Figure 2. Nitrogen flow to the waste management stream
Normally, high amount of readily soluble nitrogen is present in putrecibles.
The extent of solublization may be affected by the characteristics of the
substrate and mode of reactor operation. Thus, the amount of readily soluble
nitrogen present in putrecibles can be expected to vary between 2 and 4 g NH4+-
N per kg VS. The TKN value is often reported for MSW, and hence the NH4+-
N/TKN ratio may be a more convenient ratio to evaluate the amount of readily
available dissolved nitrogen. Gallert and Winter (1997) have been reported a
LANDFILL as N storage: Norg Solublization into NH4/NH3
Medical wasteWaste
Industrial Emissions
Consumer
Agriculture
Natural or human derived N2
Aquatic environment
NH3, N2O, NOx
NH3, N2O, N2
N2, N2O
NH4+, NO3
-, Norg in leachate
NH4+, NO3
-N2, N2O
NH3, N2O
NH4+-N/TKN ratio of 0.11 for MSW, whereas Held et al. (2002) have been
reported 0.084. Thus, almost 10% of the total nitrogen of putrecibles is readily
solubilized.
It was reported that mesophilic conditions are more favorable for the
solublization of organic nitrogen than thermophilic condition. On the other hand,
the final degree of deamination was higher in the thermophilic conditions. Also, it
was concluded that the thermophilic culture is more tolerant of high ammonia
conditions (Jokela and Rintala, 2003). The high solublization potential of
nitrogen would thus seem to originate in the easily degradable putrescible
fraction of MSW, even though its TKN content does not differ from that of
unsorted MSW. The degradation of organic matter would be expected to release
the hydrolyzed and ammonified organic nitrogen of the waste materials (Jokela
and Rintala, 2003). Hence, most of the amino acids in the proteins of waste
materials are potentially degradable, whereas some proteinaceous material
containing branched chain or aromatic aminoacids may be deaminated
oxidatively at a significantly lower rate. Generally in the case of anaerobic
digestion studies on MSW, it is impossible to assess the extent of nitrogen
solublization (e.g. without knowing characteristics of the inoculum used or the
amount of water added). One approach could be to estimate nitrogen
solublization from the reduction of solids (TS and VS). But the problem with this
approach is that the high content of soluble COD may inhibit the protein
hydrolysis (Glenn, 1976). There is a need for effective anaerobic solublization of
nitrogen contained in MSW.
Lema et al. (1988) have reviewed leachate nitrogen concentration in
young, medium and old landfills. From the results, it was concluded that the slow
leaching of nitrogen from landfilled MSW seems to continue over many decades.
This makes the assessment of the rate of nitrogen solublization in the leachate
very difficult. The nitrogen solublization rate is however slower where the MSW
particle size is higher and where the methanation of organic matter in the MSW
is slower (Jokela and Rintala, 2003).
AMMONIA TOXICITY UPON ANAEROBIC DEGRADATION
It is well known that ammonia/ammonium (NH3/NH4+) can inhibit the
anaerobic processes. This can especially become problem during treatment of
waste containing high protein content. It is especially the methane producing
bacteria that are sensitive to high NH3/NH4+ concentrations (Koster and Lettinga,
1988). The bacteria are more sensitive to NH3/NH4+ at high temperatures. This is
because inhibiting component is NH3 and the equilibrium H+ + NH3 ↔NH4+ will
be shifted towards the left at increasing temperatures. Also an increase in pH will
result in increased ammonia NH3 concentrations. This will cause increased
inhibition of the methane production and an increase in the concentration of
organic acids, which in turn cause the pH drop again. The effect of ammonical
nitrogen/ammonia in the anaerobic digester are positive and negative (Table 2).
Table 2. Effect of Ammonical Nitrogen in an anaerobic digester
Ammonical Nitrogen Effect
50 – 200 mg/L Beneficial
200 – 1000 mg/L No adverse effect
1500 – 3000 mg/L Inhibitory at pH >7
Free ammonia is toxic to methane forming bacteria. McCarty and
McKinney reported 150 mg/L un-ionized ammonia to be the inhibitory level.
Weigant and Zee man found that NH3 act as a strong inhibitor of the formation
of methane from H2 but had a relatively lower effect on the formation of
methane from acetate. Variations in concentrations of free ammonia toxicity
result from several operational factors. These factors include digester alkalinity or
buffering capacity, temperature and organic loading rates. Parkin et al. (1981)
found that under optimal experimental conditions, 8500 mg/L of total NH4+-N
could be tolerated without decrease in process performance. Soubes et al.
(1994) studied NH4+-N toxicity and found the IC50 to be 4,000 mg/L at neutral
pH.
Ammonium ions perform several important roles in an anaerobic digester.
Ammonium ions are the preferred bacterial nutrient, and they also provide
buffering capacity in an anaerobic digester. However, although ammonium
bicarbonate acts as a buffer, high ammonium bicarbonate concentrations
resulting from the degradation of amino acids, proteins, and highly concentrated
sludges may cause free ammonia toxicity. A common cause of digester failure is
the presence of an unacclimated population of methane forming bacteria at high
ammonia concentrations. Therefore, methane forming bacteria should be
gradually acclimated to increasing concentrations of ammonia (Gerardi, 2003).
AMMONIA REMOVAL FROM LANDFILL LEACHATE
There are three means to remove nitrogen compounds from leachate of
landfilled MSW, namely: ex-situ (nitrification-denitrification-discharge), in-situ
(forced bottom aeration-recirculation) and partially in-situ (ex-situ nitrification-
recirculation-denitrification) (Balslev et al., 2005; Valencia et al., 2005 and Berge
et al., 2006). Ex-situ methods seem to solve the problem of leachate treatment
regarding nitrogen compounds; however, they are not suitable for the Bioreactor
Landfill approach for which recirculation of liquids is essential to achieve optimal
performance. The in-situ approach seems more feasible once the active phase
(waste stabilization) has finished. However, issues related with carbon sources to
perform denitrification could be a problem, as well as the input of air and its
distribution within the waste mass. Therefore, the partial in-situ approach could
be an option to remove the excess of accumulated ammonia, but it can reduce
the buffering capacity of the leachate causing a decrease of pH. Additionally,
aeration of leachate could bring oxygen into the system, which requires anoxic
conditions to perform denitrification. Oxygen could affect the methanogenesis of
waste and could reach explosive levels if not managed properly. Furthermore,
either in-situ or partially in-situ approaches are likely to produce NOx and N2O,
which are significant for their contribution to atmospheric climate change (Hui et
al, 2003 and Price at el, 2003).
To overcome existing limitations, several novel nitrogen removal
processes have been developed, including SHARON process, the ANAMMOX
process, the combination of SHARON and ANAMMOX process, the CANON
process and the OLAND process (Furukawa et al., 2005). Recent research has
permitted the development of new ways of nitrogen removal, such as the partial
nitrification and the anaerobic oxidation of the ammonium (ANAMMOX), which
represent significant advances in the field of biological removal of the nitrogen
pollution (Dominguez et al., 2005). In Table 3, three process options (SHARON,
CANON and ANAMMOX) of the new system are presented and compared to a
conventional nitrogen removal system based on autotrophic nitrification and
heterotrophic denitrification.
The mentioned processes, as well as combinations of both of them,
reduce the power demand and the need of external carbon sources and generate
smaller amounts of sludge production with respect to a traditional
nitrification/denitrification system. The application of a combined partial
nitrification–Anammox process to the treatment of high ammonia nitrogen
content influents, such as leachate, is particularly promising. It would lead to
potential savings of up to 60% in oxygen requirement and 100% in external
carbon, besides significantly reducing the sludge generation and the net emission
of CO2 (Van Dongen et al., 2001), diminishing the total treatment operating cost
up to 90 % ( Jetten et al., 2001).
Table 3. Qualitative comparison of several components of the ANAMMOX technology with conventional nitrogen removal systems
NoneYesYesNoneBiomass retention
YesNoneNoneNonepH control
HighLowNoneLowOxygen requirements
Oxic; anoxicOxygen limitedAnoxicOxicConditions
N2, NO3- ; NO2
-N2, NO3-N2, NO3
-NH4+, NO2
-Discharge
WastewaterWastewaterAmmonium nitrite mixtureWastewaterFeed
2111Number of reactor
Conventional nitrification,
denitrificationCANANONANAMMOXSHARONSystem
NoneYesYesNoneBiomass retention
YesNoneNoneNonepH control
HighLowNoneLowOxygen requirements
Oxic; anoxicOxygen limitedAnoxicOxicConditions
N2, NO3- ; NO2
-N2, NO3-N2, NO3
-NH4+, NO2
-Discharge
WastewaterWastewaterAmmonium nitrite mixtureWastewaterFeed
2111Number of reactor
Conventional nitrification,
denitrificationCANANONANAMMOXSHARONSystem
AEROBIC AND ANAEROBIC AMMONIUM OXIDATION
Ammonium oxidation has been observed in many bacterial species.
Ammonia is oxidized by two pathways: first, ammonia is oxidized to nitrite by
hydroxylamine, which is then oxidized to nitrate by hydroxylamine
oxidoreduxctase; Second, ammonia and nitrite are anaerobically converted to
nitrogen gas. The aerobic chemolithoautotrophic ammonia oxidizing bacteria
(AOB) are specialists that can grow on ammonia and carbon dioxide (Purkhold et
al., 2000) and use ammonia monooxygenase to convert ammonia into
hydroxylamine. Many heterotrophic bacteria, such as P. Pantotropha and
Alcaligenes faecalis strain TUD (Otte et al., 1999), can carry out the same
reaction. Methanotrophs are capable of converting ammonia to hydroxylamine
Source: Jetten et al., 2002
via the methane monooxygenase, whereas the ammonium monooxygenase can
oxidize methane to carbon dioxide. The recently identified lithotrophic
planctomycete possesses the ANAMMOX pathway, which is coupled to nitrite
reduction (Strous et al., 1999).
The ANaerobic AMMonium OXidation (ANAMMOX) process, which was
discovered 10 years ago (Mulder, 1992) but already predicted to exist 30 years
ago (Broda, 1977), could offer an alternative for the treatment of this return
stream. Later, Van de Graff et al. (1997) and Bock et al. (1995) observed that
nitrite was the preferred electron acceptor for the process. Also, other streams
with high nitrogen and low carbon content such as landfill leachates and
evaporator condensates could be treated. In the ANAMMOX process ammonium
is oxidized under anoxic, i.e. oxygen depleted, conditions with nitrite as electron
acceptor. Ammonium and nitrite are consumed on an almost equimolar basis.
The ANAMMOX process should always be combined with a partial nitritation
process, such as the SHARON process (van Dongen et al., 2001a&b), where half
of the ammonium is oxidized to nitrite. Both autotrophic processes will increase
the sustainability of wastewater treatment as the need for carbon addition (and
concomitant increased sludge production) is omitted and oxygen consumption
and the emission of nitrous oxide during oxidation of ammonia are largely
reduced (Jetten et al., 1997). As such, the combined process (partial nitritation
and ANAMMOX) was termed autotrophic nitrogen removal process (Jetten et al.,
2002).
MICROBIOLOGY OF ANAMMOX
Microbial nitrogen metabolism also plays an important role in the global nitrogen
cycle. Microbial activities, such as denitrification and ANAMMOX, are the major
mechanisms that convert combined nitrogen to dinitrogen gas, thereby
completing the nitrogen cycle. The updated nitrogen cycle with ANAMMOX is
depicted in Figure 3 (after Jetten et al., 1999). Nitrification is the aerobic
oxidation of NH3 to NO3-. It consists of two sequential steps carried out by two
phylogenetically unrelated groups of aerobic chemolithoautotrophic bacteria.
Some heterotrophic bacteria can also oxidize ammonium to nitrate, but this is
only a very small contribution to the overall ammonia oxidation (Pynaert, 2003).
No single known autotrophic bacterium is capable of complete oxidation of NH3
to NO3- in a single step (Abeliovich, 1992). In view of coupling a partial
nitrification unit with an Anammox unit, nitrite oxidising activity should be
suppressed and TAN should only be oxidised for about 50 % to TNO2.
The physiology of anaerobic ammonium oxidizing aggregates cultivated in
a sequencing batch reactor was investigated by Strous et al. (1999). The
maximum specific substrate conversion rate of the ANAMMOX biomass was
measured as a function of temperature and pH in batch experiments. From the
temperature dependency of ANAMMOX activity, the activation energy was
calculated to be 70 kJ/mol. Strous et al. (1998) have also reported that the
affinity constants for the substrates, ammonium and nitrite, are less than 0.1 mg
N/L inhibited ANAMMOX process completely. In another study Strous et al.
N2
TA
TNO
+ O2
+ O2
NO3-
+ COD
+ COD
TNO
N2
TA
TNO
ANAMMOX + O2
Nitrogen Fixation
Classical nitrogen removal Autotrophic nitrogen removal
Figure 3. ANAMMOX process in nitrogen cycle
(1999) have shown that the ANAMMOX process was reversibly inhibited by the
presence of oxygen.
Bacteria capable of anaerobically oxidizing ammonium had not been
known earlier and were referred as the “lithotrophs missing from nature”
(Shivaraman and Geetha, 2003). These missing lithotrophs were discovered and
identified as the new autotrophic members of the order of planctomycete, one of
the major distinct division of bacteria (Strous et al., 1999a). The anaerobic
ammonium oxidation reaction is carried out by two ANAMMOX bacteria that
have been tentatively named as “Brocardia anammoxidans” (Strous et al.,
1999a) and “Kuenenia stuttgartiensis” (Schmid et al., 2000). The high
ANAMMOX activity observed for both bacteria in a pH range between 6.4 and
8.3 and temperature between 20oC and 43oC (Strous et al., 1999b; and Egli et
al., 2001). The ANAMMOX bacterial activity is 25-fold higher than aerobic
nitirifying bacterial oxidation of ammonium under anoxic conditions when using
nitrite as the electron acceptor (Jetten et al., 1999). Acetylene, phosphate and
oxygen are known to be strongly inhibiting ANAMMOX activity (Van De Graaf et
al., 1996).
BIOCHEMISTRY OF ANAMMOX
The possible metabolic pathways for anaerobic ammonium oxidation are
depicted in Figure 4. (Van de Graff et al., 1997). The ANAMMOX process is
based on energy conservation from anaerobic ammonium oxidation with nitrite
as electron accpetor without addition of external carbon source (Jetten et al.,
1999). Hydrazine and hydroxylamine are known to be some intermediates of the
process (Van de Graff et al., 1997; Schalk et al., 1998; and Jetten et al., 1999).
Carbon dioxide is the main source for the growth of ANAMMOX bacteria (Van de
Graff et al., 1997).
Figure 4. Metabolic pathway in ANAMMOX bacteria
Two possible pathways were hypothesized by van de Graaf et al. (1997) for the
ANAMMOX process:
H
NI
4e
HA
H2N=NH2
NH3NO2- N=N
4H+
5H+
NH2 OH
Cytoplasm
Anammoxosome
• Oxidation of ammonium ion to hydroxylamine, that reacts with nitrite which is
further reduced to nitrogen. Hydroxylamine-formation from ammonium ion via
the ammonium monooxygenase, however, seems unlikely because of the strong
oxygen inhibition (van de Graaf et al., 1996; Jetten et al., 1999).
• Partial reduction of nitrite with the formation of hydroxylamine (NH2OH), that
reacts further with ammonium to form hydrazine (N2H4). Hydrazine is further
converted into nitrogen. This oxidation would give the necessary reducing
equivalents for the initial reduction of nitrite.
15N-labeling experiments showed that this second possibility is the correct
one (van de Graaf et al.,1997). The addition of labelled hydroxylamine led to the
formation of labelled nitrogen gas, in contrast to the addition of 15N2O.
Sustained growth on hydroxylamine or hydrazine is however not possible (Schalk
et al., 1998). Strous et al. (1999b) did notice that the addition of at least 50 µM
of these intermediates resulted in complete recovery of the ANAMMOX activity
after inactivation with TNO2. Schalk et al. (2000) succeeded in purifying and
characterizing the hydroxylamine Oxidoreductase/hydrazine reductase
(HAO/HZO) of an ANAMMOX culture. The HAO/HZO was able to oxidize both
hydroxylamine and hydrazine under anoxic conditions to respectively NO, N2O
and N2. The HAO/HZO made up 9 % of the total soluble protein fraction of the
ANAMMOX species Candidatus Brocadia anammoxidans. Schalk et al. (2000)
also found that hydrazine strongly inhibits the oxidation of hydroxylamine.
Kuenen and Jetten (2001) have suggested the most plausible hypothesis for the
ANAMMOX mechanism. Nitrite reduction by a nitrite reducing enzyme leads to
the formation of hydroxylamine. An unknown hydrazine hydrolase converts
ammonia and hydroxylamine to hydrazine that is converted into nitrogen by
HAO/HZO. This oxidation would give the necessary reducing equivalents for the
initial reduction of nitrite. In the biochemical model, the ANAMMOX reaction
establishes a proton gradient by the effective consumption of protons in the
riboplasm and production of protons inside the anammoxosome, a mechanism
known as separation of charges. This result in an electrochemical proton gradient
directed from the anammoxosome to the riboplasm. Based on isotopic carbon
analysis Schouten et al. (2004) concluded that different ANAMMOX bacteria,
such as Candidatus Scalindua sorokinii and Candidatus Brocadia anammoxidans
use identical carbon fixation pathways, which may be either the Calvin cycle or
the acetyl coenzyme A pathway.
APPLICATION OF ANAMMOX IN LEACHATE TREATMENT
Recent research has permitted the development of new ways of nitrogen
removal, such as the partial nitrification and the anaerobic oxidation of the
ammonium (Anammox), which represent significant advances in the field of
biological removal of the nitrogen pollution. The application of a combined partial
nitrification–ANAMMOX process to the treatment of high ammonia nitrogen
content influents, such as leachate, is particularly promising. It would lead to
potential savings of up to 60% in oxygen generation and 100% in external
carbon, besides significantly reducing the sludge generation and the net emission
of CO2 (Van Dongen et al.,2001), diminishing the total treatment operating cost
up to 90 % (Jetten et al., 2001). The nitrogen removal efficiencies observed in
the aerobic/anoxic biological reactor of the Meruelo landfill leachate pre-
treatment plant are greater than those expected in a conventional nitrification-
denitrification process. Nitrogen losses that cannot be explained by the classical
nitrogen removal phenomena have also been observed in other biological
treatments of leachate (Helmer et al., 1999; and Siegrist et al.,1998). As the
degradation phenomena in Meruelo have not been experimentally characterized
yet, the observed high efficiencies, together with the favourable environment
conditions, lead to the hypothesis of a possible occurrence of Anammox
processes in the reactor.
LIMITATIONS
ANAMMOX coupled to nitrite reduction offers opportunities in the area of
process development of nitrogen removal systems. One of the biggest challenges
is how to accelerate the slow rate of nitrogen removal from these systems (the
rate is less than half that of aerobic nitrification) (Strous et al., 1999; and Jetten
et al., 1998). However, from a commercial application perspective, the more
challenging issue is the extremely slow growth rate (10-14 days) of the bacteria
known to carry out these reactions. Similar to aerobic nitrification, ANAMMOX is
subjected to inhibition. This process requires anaerobic conditions for ammonia
oxidation, but inhibition by oxygen is reversible
FUTURE STUDY
ANAMMOX technology has been evaluated using synthetic
wastewater/sludge digester effluent from domestic WWTP. Research is necessary
to know the feasibility of applying ANAMMOX process technology with other
actual wastewater and leachates using appropriate reactor types and
configuration. The performance of ANAMMOX process in treating actual
wastewater/leachate would not only depend on ANAMMOX bacteria but also on
the co-existence of other important oxygen scavenging and ammonia
generating/ammonia to nitrite oxidizing bacteria. Research is needs to be carried
out to work out optimal conditions for such an ecosystem to sustain in a reactor
and develop, methodologies to monitor the responsible microbial community in
the system.
Applied genomic research can be used to identify genes and patterns of
expression that are critical to the performance of nitrogen metabolism in
responses can be coupled with reporter systems for the development of online
measurement systems. Coupling the advances related to bacterial nitrogen
metabolism with improved monitors of macroscopic performance should lead to
more robust operating strategies for wastewater bioreactors. Genomic
information, in combination with traditional biochemical, genetic and ecological
studies is needed to understand the inorganic nitrogen metabolism, and thus
benefit their industrial applications and landfill leachate treatment.
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