“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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