Evaluating the potential for dissimilatory nitrate reduction by anammox bacteria for municipal wastewater treatment

Accepted Manuscript Evaluating the potential for dissimilatory nitrate reduction by anammox bacteria for municipal wastewater treatment Celia M. Castro-Barros, Mingsheng Jia, Mark C.M. van Loosdrecht, Eveline I.P. Volcke, Mari K.H. Winkler PII: DOI: Reference:

S0960-8524(17)30186-4 http://dx.doi.org/10.1016/j.biortech.2017.02.063 BITE 17635

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Bioresource Technology

Received Date: Revised Date: Accepted Date:

10 December 2016 14 February 2017 15 February 2017

Please cite this article as: Castro-Barros, C.M., Jia, M., van Loosdrecht, M.C.M., Volcke, E.I.P., Winkler, M.K.H., Evaluating the potential for dissimilatory nitrate reduction by anammox bacteria for municipal wastewater treatment, Bioresource Technology (2017), doi: http://dx.doi.org/10.1016/j.biortech.2017.02.063

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1

Evaluating the potential for dissimilatory nitrate reduction by anammox

2

bacteria for municipal wastewater treatment

3

Celia M. Castro-Barros1, Mingsheng Jia1, Mark C.M. van Loosdrecht2, Eveline

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I.P. Volcke1, Mari K.H. Winkler1,2,3,*

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Gent, Belgium.

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2629 HZ Delft The Netherlands.

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3

Department of Biosystems Engineering, Ghent University, Coupure links 653, 9000

Department of Biotechnology, Delft University of Technology, Van der Maasweg 9,

Department of Civil and Environmental Engineering, University of Washington,

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Seattle, WA 98195-2700, USA.

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ABSTRACT

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Anammox bacteria can perform dissimilatory nitrate reduction to ammonium (DNRA)

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with nitrite as intermediate coupled to the oxidation of volatile fatty acids (VFA). Batch

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tests with enriched anammox and a co-culture of anammox and heterotrophic bacteria

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showed the capacity of Candidatus ‘Brocadia fulgida’ to perform the DNRA coupled to

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the anammox reaction (DNRA-anammox) at a high rate although the culture was not

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previously adapted to VFA. From thermodynamic calculations it could be stated that

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low COD/N influent ratios favour the DNRA-anammox transformation over

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heterotrophic conversions since more free energy is gained. A process scheme is

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proposed for an innovative nitrogen removal system in which the nitrate produced by

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nitrite oxidizing bacteria and/or anammox bacteria is converted during DNRA-

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anammox pathway, resulting in a sustainable nitrogen removal from municipal

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wastewater while circumventing the troublesome out-selection of nitrite oxidizing

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bacteria encountered in mainstream applications.

*Corresponding author ([email protected])

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KEYWORDS

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Anammox; dissimilatory nitrate reduction; DNRA; nitrogen removal; VFA

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1. INTRODUCTION

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The anaerobic ammonium oxidation (anammox) conversion comprises the oxidation of

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ammonium with nitrite as electron acceptor to nitrogen gas under anoxic conditions.

31

The implementation of the anammox process needs a preceding step in which half of the

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ammonium in the wastewater is oxidized to nitrite by ammonium oxidizing bacteria

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(AOB) (partial nitritation). For the success of this partial nitritation-anammox system

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(Figure 1A), the further oxidation of nitrite to nitrate by nitrite oxidizing bacteria (NOB)

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must be prevented. Compared to conventional nitrification-denitrification, partial

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nitritation-anammox requires up to 63% less aeration energy, produces 70-80% less

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sludge, emits almost no CO2 and safes up to 100% of costly external organic carbon

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addition (Siegrist et al., 2008).

39

Anammox-based processes were successfully implemented at several full-scale

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plants to treat warm ammonium-rich reject water from anaerobic digesters (sidestream)

41

(Abma et al., 2010; van der Star et al., 2007; Wett, 2007). Due to the advantages of this

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technology in the sidestream (1% of the volume flow of the wastewater treatment plant,

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WWTP), current scientific research is focused on the implementation of anammox

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technology in the mainstream (99% of the volume flow) of the WWTP (Kartal et al.,

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2010). However, different challenges need to be overcome for successful application of

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this process in the mainstream. As opposed to the sidestream, the mainstream has lower

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temperature (down to ca. 10 °C instead of 30 °C), stronger effluent requirements, lower

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ammonium concentrations (ca. 50 g N m-3 instead of 500 g N m-3), and contains higher

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amount of biodegradable organic matter (higher COD/N ratio). Anammox is known to

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optimally thrive at 35 °C, but different studies have shown that an implementation of

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the anammox process at low temperature is also feasible (Hendrickx et al., 2014; Lotti

52

et al., 2014; Vazquez-Padin et al., 2009). In these studies, the nitrite was added directly

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in the anammox reactor whereas, in practice, the nitrite needs to be formed within the

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process. In this regard, Lotti et al. (2015) tried to establish partial nitritation-anammox

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in a pilot-scale reactor to treat a mainstream effluent and found that the formation of

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nitrite (and suppression of NOB) was the limiting factor of the process. Therefore, one

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of the main bottlenecks for the application of anammox-based technologies in the

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mainstream seems to be an effective nitrite supply to the anammox pathway. Although

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NOB have a lower affinity for oxygen than AOB, the presence of NOB cannot be out

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ruled in the main line even if operated under oxygen limiting conditions (Pérez et al.,

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2014). The difficulty to deselect against NOB can be explained by growth rates. At high

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temperatures, AOB have a faster growth rate than NOB. Thus, NOB can be

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outcompeted by operating the system at sludge retention times (SRT) below the

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minimal required growth rate of the nitrite oxidizers (Hellinga et al., 1998). However, at

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low temperatures (such as in the mainstream) NOB possess a faster growth rate than

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AOB and therefore NOB cannot be out selected based on SRT control and hardly

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controlled by oxygen limitation either.

68

Beside nitritation as a manner to provide nitrite to the anammox reaction during

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municipal wastewater treatment, other approaches concern the partial denitrification

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until nitrite only (Figure 1B). This conversion can be performed by either autotrophic or

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heterotrophic microorganisms. For the heterotrophic denitrification, the limitation of

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organic carbon and relatively short anoxic periods positively influence the accumulation

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of nitrite (Du et al., 2015). In addition, recent studies have shown that low amount of

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organic compounds (limitation of organic carbon) in partial nitritation-anammox

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systems can improve the overall nitrogen removal through the partial heterotrophic

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denitrification, which reduces nitrate to nitrite that can be taken up by anammox

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organisms (Mozumder et al., 2014).

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Regarding the presence of organic carbon compounds in the wastewater, it was

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recently discovered that anammox bacteria are able to perform dissimilatory nitrate

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reduction to ammonium (DNRA) with nitrite as intermediate using volatile fatty acids

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(VFA) as electron donor (organotrophic activity) (Guven et al., 2005; Kartal et al.,

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2007a; Kartal et al., 2007b; Shu et al., 2015; Winkler et al., 2012a). The VFA are

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oxidized to CO2 via acetyl-CoA (Russ et al., 2012) but are not assimilated into biomass

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(Guven et al., 2005; Kartal et al., 2007b), hence keeping sludge production low. In some

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experiments, DNRA by anammox bacteria was reported to occur coupled to the

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anammox transformation (from where the bacteria get energy to grow) in two ways:

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converting the nitrate to nitrite only (partial DNRA) and combining the nitrite formed

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with external ammonium present in the environment to yield nitrogen gas in the

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anammox reaction (Figure 1C); and also, anammox bacteria were shown to reduce part

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of the nitrate to the intermediate nitrite and one part to ammonium, the end product of

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DNRA (full DNRA), to then utilize the formed nitrite and ammonium for anaerobic

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ammonium oxidation (Figure 1D) (Kartal et al., 2007a).

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The main challenge to establish DNRA by anammox bacteria is their out-

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competition by heterotrophic bacteria, which compete for nitrate and acetate

95

(heterotrophic denitrification) and which have a much faster growth rate. Heterotrophs

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are also known to perform DNRA. In general, for heterotrophs, DNRA pathway is

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promoted over denitrification under nitrate limiting conditions (i.e. high COD/N ratios)

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and when nitrate (and not nitrite) is the main electron acceptor present (Kraft et al.,

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2014; van den Berg et al., 2015). However, little is known about the conditions that

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promote anammox bacteria performing DNRA over the heterotrophic utilization of

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carbon. Some anammox species, enriched in lab-scale reactors, were found to

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effectively outcompete heterotrophs for propionate (Candidatus ‘Anammoxoglobus

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propionicus’, (Kartal et al., 2007b), and acetate (Candidatus ‘Brocadia fulgida’, (Kartal

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et al., 2008)) in the presence of ammonium but the conclusions remained descriptive.

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The main objective of this study was to gain understanding on the DNRA

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pathway of anammox bacteria. The bacterial populations were determined in three case-

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studies and the factors promoting the DNRA by anammox bacteria over heterotrophic

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transformations evaluated. Thermodynamic calculations were determined to assess the

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energy gained per mole of substrate and electron transferred during different

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transformations involving the conversion of nitrate and acetate. The outcome of this

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research suggests the joint implementation of partial DNRA and anammox processes by

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anammox bacteria, which can circumvent the accumulation of nitrate in the effluent at

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the same time providing nitrite to the anammox reaction by partial DNRA, resulting in

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an innovative and sustainable system to remove nitrogen from municipal wastewater.

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2. MATERIALS AND METHODS

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Nitrate reduction and organotrophic activity by anammox bacteria were assessed in 3

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case-studies involving different conditions. Batch tests were conducted using anammox

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biomass from a full-scale granular sludge reactor (Case A, described in section 2.1) and

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sludge from an anoxic/aerobic sequencing batch reactor (SBR) containing anammox

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bacteria and heterotrophic phosphate accumulating organisms (PAOs) among other

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communities (Case B, described in section 2.2). The third set-up consisted of an

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anammox SBR operated under a semi-continuous exposure to organic carbon (acetate)

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(Case C, described in section 2.3) (see Table 1).

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2.1. Batch tests with anammox granular sludge from a full-scale WWTP

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(Case A)

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A series of 5 different batch tests was performed using biomass taken from the

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anammox granular sludge reactor from the WWTP of Rotterdam (The Netherlands).

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The biomass was first washed with tap water and then used in each batch test, which

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were conducted in a 2 L bioreactor. Before each test, nitrogen gas was supplied to the

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bioreactor for 30 min to ensure anoxic conditions. Different initial substrate

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concentrations were added in each batch test with a syringe through a septum (Table 1,

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Case A), together with a mineral medium according to van de Graaf et al. (1996) and

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Dapena-Mora et al. (2004) (Supplementary Information (SI), Table S1. 1 and Table S1.

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2). Inorganic carbon was provided in the mineral medium in excess as KHCO3 (1 g

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KHCO3 L-1). The temperature was controlled at 30 ± 1 °C by using a heating water

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jacket connected to a circulating bath and the pH was not controlled. Methanol (MeOH)

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was added in some batch tests to assess the potential inhibitory effect of the alcohol on

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the bacterial activities. Liquid samples were collected at different times along the batch

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tests to determine the consumption rates.

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2.2. Batch tests with anammox and heterotrophic bacteria (Case B)

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A series of 8 batch tests was performed with granular biomass collected from a bubble

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column SBR operated under intermittent anoxic/aerobic conditions, carrying out

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nitritation-anammox and organic carbon removal via DNRA process as described in

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Winkler et al. (2012a), but with the difference that the reactor operation was changed to

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a two-fold higher carbon/nitrogen ratio in the influent (from 0.5 to 0.9) (SI, PHASE III

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in Figure S2. 1). The biomass was collected from the SBR, washed with tap water and

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aerated to remove excess of ammonium. All batch tests were completed in glass flasks

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containing 250 mL, in which nitrogen gas was supplied for 25 minutes prior to the start

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of the tests to ensure anoxic conditions. Each test was induced with a different set of

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substrates as listed in Table 1 (Case B) and same mineral medium and trace elements as

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used during the operation of the reactor (Winkler et al., 2012a). Temperature and pH

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were not controlled during the batch tests. As in the batch tests of Case A (section 2.1),

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methanol was added in some tests. Samples were taken over time during each batch test

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and the volatile suspended solids were determined. 2.3. Sequencing batch reactor under semi-continuous acetate supply (Case

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C)

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A SBR of 2.3 L was inoculated with anammox granular sludge from the CANON

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reactor from the wastewater treatment plant of Amersfoort (The Netherlands). The

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reactor was operated in 6 hour cycles with two different regimes (SI, Figure S3. 1).

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‘Anammox regime’ lasted 1 month and compromised 150 min of continuous feeding + mixing (Feed C.1 (Table

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‘Regime under semi-continuous exposure to acetate’ lasted 1 month, during which the

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bacteria were intermittently fed with acetate as the only carbon source. The reactor was

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operated in different phases. The phase, which was operated with acetate consisted of

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90 min of continuous feeding + mixing (Feed C.2 (Table 1, Case C), NH4+ + NO3- +

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acetate) and 30 min of mixing. The phase, which was operated without acetate consisted

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of 120 min of continuous feeding + mixing (Feed C.3 (Table 1, Case C), NH4+ + NO2-)

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and 30 min of mixing. Then a settling phase of 60 min and decanting of 30 min was

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applied.

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For both regimes, 1 L of medium was fed each cycle. The feeding pump for the

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‘Anammox regime’ was set at 6.67 mL min-1, while the feeding flow rate of the ‘regime

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under semi-continuous exposure to acetate’ was 4.76 mL min-1. The modifications on

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the composition of the different feeding conditions used are summarized in Table 1,

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Case C. In addition to the compounds of these feeding strategies, a mineral medium was

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supplied (SI, Table S1. 1 and S1. 2). The effect of methanol was not assessed during the

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operation of the SBR in Case C to avoid the irreversible inhibition to the bacterial

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population during reactor operation. Although the reactor was inoculated with granular

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sludge, the granules disintegrated after 2 days of operation due to the mixing applied by

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a magnetic stirrer to keep homogenous conditions. The temperature was controlled at 30

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± 1 °C by using a heating water jacket connected to a circulating bath and the pH was

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kept at 7.4 ± 0.4.

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2.4. Analytical methods and FISH analysis

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The total suspended solids (TSS) and volatile suspended solids (VSS) were determined

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according to standard methods (APHA, 1998). Liquid samples were taken and

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immediately filtered through disposable Millipore filters (0.45 µm pore size).

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Ammonium, nitrate, nitrite and COD (acetate) were determined with standard test kits

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Hach-Lange and Macherey-Nagel for Cases A and C. The nitrogen compounds in Case

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B were analysed with QuikChem 8500 Flow Injection Analysis (FIA) from Lachat

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Instruments and the acetate was analysed with a High Performance Liquid

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Chromatography (HPLC) system.

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Fluorescence in situ hybridization (FISH) and slicing was performed on the

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granular biomass from the bubble column SBR for Case B according to the method

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proposed by Winkler et al. (2011) in order to observe the population spatial distribution

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inside the granules. Likewise, FISH was accomplished on the floccular biomass from

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the anammox SBR operated with a transient exposure to acetate (Case C).

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3. RESULTS AND DISCUSSION

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Nitrate reduction by heterotrophic and anammox bacteria was evaluated in three

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separate case studies: (i) batch tests with enriched anammox biomass from a full-scale

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reactor (Case A), (ii) batch tests with anammox and heterotrophic biomass from an

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anoxic/aerobic SBR (Case B), and (iii) sequencing batch operation of a bioreactor with

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anammox biomass (Case C) (see Table 1 for more information on the conditions of each

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Case study). The results from bacterial population determination are presented in

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section 3.1, and the anammox and nitrate reduction activities assessed in the absence

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and presence of organic carbon (acetate) are described in section 3.2. The factors

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influencing the acetate and nitrogen transformations of both heterotrophs and anammox

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bacteria are discussed in section 3.3 and 3.4 (i.e. competition heterotrophs vs. anammox

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bacteria). Finally, a scheme to apply the combined partial DNRA-anammox by

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anammox bacteria is presented in section 3.5.

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3.1. Bacterial populations

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FISH analysis was performed to determine the dominant bacterial populations (see SI,

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Figure S4. 1). The anammox species observed in each of the three cases of this study

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was Candidatus ‘Brocadia fulgida’, which is known to have the ability to carry out

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DNRA and outcompete heterotrophs for volatile fatty acids (VFA) (Kartal et al., 2008;

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Winkler et al., 2012a; Winkler et al., 2012b). Concerning their relative abundance,

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anammox bacteria were the dominant organisms present in the experiments of Case A

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and Case C, with negligible amount of other microorganisms. For the Case A, the

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biomass was taken from a full-scale anammox reactor in which the COD/N ratio was

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relatively low, hindering the development of heterotrophic bacteria. The SBR of Case C

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was not exposed continuously to acetate (see Figure S3. 1 in the SI), which disfavoured

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the presence of heterotrophs.

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The granular sludge for the batch tests of Case B was harvested from a lab

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reactor exposed to acetate (Winkler et al., 2012a). The FISH determination of this

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biomass revealed anammox bacteria and polyphosphate-accumulating organisms

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(PAOs) in the inner core of the granules, in addition to AOB in the outer zone. In the

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reactor the produced nitrate from the aerated phase was recycled together with the

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influent containing acetate. However, over time partial nitritation-anammox worked

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well and only little nitrate accumulated. Thus, anaerobic conditions (no electron

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acceptor present) were established and the growth of PAOs promoted, which

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accumulated acetate internally as poly-β-hydroxybutyrate (PHB) under anaerobic

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conditions and under aerobic or anoxic conditions the internally stored PHB was

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oxidized and used for cell growth.

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3.2. Anammox conversion and organotrophic nitrate reduction

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The anammox activity was assessed in the absence (section 3.2.1) and presence of

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acetate (section 3.2.2). Also, the nitrate and acetate consumption rates were measured in

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the experiments with acetate addition.

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3.2.1 Anammox activity in the absence of acetate The anammox activity without acetate present was assessed during Case A (batch tests

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A.1 and A.2 in Table 1), Case B (batch tests B.1 and B.2 in Table 1), and the

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‘Anammox regime’ of Case C (Feed C.1 in Table 1) (Figure 2). The

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determination of the substrate consumption rates is described in section S5 of the

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SI. As expected the ammonium (rNH4) and nitrite (rNO2) specific consumption

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rates (i.e. expressed per g VSS, total biomass) were higher for the reactors

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containing highly enriched anammox biomass (Cases A and C) and accordingly

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lower for the reactor containing also PAOs (Case B) (Figure 2).

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The addition of methanol inhibited the anammox activity (rNH4) by 94.0% and

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89.3% in Cases A and B, respectively (Figure 2A and B). The inhibition observed was

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not complete as previously reported (Guven et al., 2005), likely due to diffusion

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limitations or a protective environment inside the granules.

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3.2.2 Anammox and organotrophic nitrate reduction activities in the presence of acetate

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The anammox activity (rNH4) decreased by 42.1% in Case A and 41.9% in Case

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B (Figure 3.1) in the presence of acetate, nitrate and ammonium, when compared to the

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case with ammonium and nitrite, but not acetate (Figure 2). When nitrite was not added,

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it had to be formed from the reduction of nitrate. Nitrite production would then become

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the limiting reaction, because the nitrate reduction rates (rNO3) for both Cases A and B in

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Figure 3.1 and 3.2 were slower than the nitrite reduction rates (rNO2) performed during

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regular anammox reaction in Figure 2A and B.

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In order to clarify which communities (anammox or heterotrophic bacteria) reduced the nitrate and oxidiz

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the nitrate reduction and acetate oxidation by 99.0% and 98.4%, respectively, compared

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to the test without methanol addition (see batch tests A.3 and A.4 in Figure 3.1). Since

262

earlier studies have shown that heterotrophs are not inhibited at concentrations used in

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this study (12 mM for Case A, Table 1) (Jensen et al., 2007) it could be concluded that

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anammox bacteria were the major responsible microorganisms for the consumption of

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nitrate and acetate in Case A, which is in agreement with the lower abundance of other

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bacteria than anammox on the FISH images (Figure S4. 1). It is important to note that,

267

the anammox biomass used for the experiments in Case A was not adapted to organic

268

matter. This biomass was taken from a full-scale anammox reactor preceded by a

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SHARON reactor in which organic carbon would have been oxidized. Despite this fact,

270

in the present study it was shown that the anammox species Candidatus ‘Brocadia

271

fulgida’ performed DNRA along with anaerobic ammonium oxidation at a high rate. In

272

the present study the nitrate reduction rate by anammox bacteria was 1.9 slower than the

273

conventional anammox activity (ammonium oxidation rate in the test with ammonium

274

and nitrite only), which is significantly higher compared to previous studies (nitrate

11

275

reduction was 21 times slower than regular anammox activity in the tests from Kartal et

276

al. (2007b) and 10 times slower in Kartal et al. (2007a)).

277

In Case B, where experiments were conducted with biomass from a laboratory

278

reactor with high presence of heterotrophic populations and PAOs, it could not be

279

concluded to which extent heterotrophic and/or anammox bacteria contributed to the

280

reduction of nitrate and oxidation of acetate when supplying ammonium, nitrate and

281

acetate (B.3 in Figure 3.1). A higher acetate consumption rate (rCOD) was found during

282

the batch tests with biomass from the lab-scale reactor containing PAOs (Case B, 21.1

283

mg COD (g VSS)-1 h-1) than in the batch tests with anammox biomass from the full-

284

scale system (Case A, 6.99 mg COD (g VSS)-1 h-1). This higher activity can be

285

attributed to PAOs, which are known to have a high acetate uptake rate due to their

286

capacity storing acetate as PHB. Note that during the tests with nitrate and acetate (B3,

287

B4 and B5 in Figure 3) uptake of phosphorus was observed (SI, Table S5. 2),

288

confirming PAO activity.

289

3.3. What triggers DNRA by anammox bacteria over heterotrophic DNRA

290

and heterotrophic denitrification?

291

The present study aims at assessing when denitrification and DNRA take place and if

292

they are carried out by either heterotrophs (including PAOs) or anammox bacteria.

293

Factors influencing the DNRA pathway of anammox bacteria were studied

294

experimentally (Figure 3 and Figure 4) and by thermodynamic calculations (Table 2) to

295

distinguish the carbon and nitrogen transformations among heterotrophs (including

296

PAOs) and anammox bacteria.

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3.3.1. Effect of ammonium in the medium on the organotrophic nitrate reduction

298

by anammox bacteria

299

The effect of ammonium in the medium on the nitrate (rNO3) and COD (rCOD) conversion

300

rates was assessed through batch tests with the enriched anammox biomass in Case A

301

(test A3 vs. A5, Figure 3), in which it was concluded earlier (section 3.2.2) that

302

anammox bacteria were the dominating culture and responsible for the nitrate and

303

acetate consumption. With ammonium in the medium, the reduction of nitrate was

304

faster (tests A.3 vs. A5, Figure 3) and no accumulation of nitrite was observed.

305

Therefore, nitrite for anaerobic ammonium oxidation (rNH4=5.173 mg N (g VSS)-1 h-1,

306

A.3 in Figure 3) must have been supplied through the anammox mediated DNRA

307

pathway in which nitrite is formed as intermediate (Figure 1C). The absence of

308

ammonium (and nitrite) in the medium (A.5 in Figure 3) led to the accumulation of

309

nitrite in the beginning of the test (SI, Figure S7. 1A), which can be explained by the

310

fact that the reduction of nitrate to nitrite by anammox bacteria is faster than the

311

conversion of nitrite to ammonium (Kartal et al., 2007a)). Within this test no

312

ammonium was accumulated, suggesting that anammox bacteria immediately

313

metabolized the ammonium obtained through the full DNRA pathway for anaerobic

314

ammonium oxidation (Figure 1D).

315

The capacity of anammox bacteria of using DNRA expands their growth niche

316

from a pure autotrophic environment to ecological niches in which organic carbon

317

occurs along with nitrate. The DNRA pathway has not been shown to deliver net energy

318

to anammox bacteria but supplies them with their catabolic substrates. Depending on

319

the conditions, only nitrite is produced (via partial DNRA) which is then reduced with

320

ammonium present in the environment (Figure 1C) or DNRA produces both nitrite an

321

ammonium to fuel the anammox reaction (Figure 1D).

13

322

The presence of ammonium enhances the organotrophic reduction rate of nitrate

323

by anammox bacteria since the slow formation of ammonium through DNRA is avoided

324

and the produced nitrite can be immediately utilized for the energy yielding anaerobic

325

ammonium oxidation with nitrite. This is in line with other studies which have shown

326

that the presence of ammonium gives anammox bacteria advantage to perform DNRA

327

over heterotrophic transformations (Guven et al., 2005; Kartal et al., 2007a; Kartal et

328

al., 2007b; Winkler et al., 2012a; Winkler et al., 2012b).

329

3.3.2. Effect of substrate limitation - energy perspective

330

The energy obtained per substrate consumed, N2 formed, and the electrons transferred

331

were calculated for different microbial pathways (see SI, section S6 for description of

332

the calculation) since it is one of the main driving forces explaining microbial

333

competition (Gonzalez-Cabaleiro et al., 2015) (Table 2).

334

According to the energy per electron transferred for the heterotrophic

335

transformations (Table 2), a higher theoretical yield is obtained during the

336

denitrification of nitrite (reaction 6, -120 kJ e--eq-1) than is the case for the

337

denitrification of nitrate (reaction 1, -100 kJ e--eq-1) whereas the DNRA is the least

338

favoured process of the heterotrophic transformations (reaction 2, -62.5 kJ e--eq -1).

339

Therefore, under no substrate limiting conditions, heterotrophic denitrification over

340

nitrite would be more favoured than other heterotrophic transformations of nitrogen if

341

nitrite is available in the medium, as shown by Kraft et al. (2014). With respect to

342

substrate competition, van den Berg et al. (2015) selected heterotrophic communities

343

performing DNRA over denitrification when limiting the nitrate in the system (high

344

COD/N ratio). However, from a thermodynamic point of view, the limitation of nitrate

345

cannot be used to explain the dominance of the heterotrophic DNRA pathway because a

346

similar energy is harvested per mole of nitrate during denitrification (reaction 1) and

14

347

DNRA (reaction 2, Table 2). Instead, the limitation of acetate (low COD/N ratio) would

348

promote the heterotrophic denitrification (-797 kJ mol-acetate-1) over the heterotrophic

349

DNRA (-500 kJ mol-acetate-1).

350

Regarding the conversions involving anammox bacteria, the theoretical Gibbs

351

free energy obtained per N2-mol produced is higher for the partial DNRA-anammox

352

(reaction 7, Figure 1C) and full DNRA-anammox reaction (reaction 8, Figure 1D) than

353

the energy obtained if only performing conventional anaerobic ammonium oxidation

354

(reaction 4). Nevertheless, whether or not anammox bacteria can use this extra energy

355

when performing the DNRA pathway is not yet understood. Current literature suggests

356

that anammox bacteria do not use the organic carbon for cell assimilation and therefore

357

energy usage from this pathway remains unsolved (Kartal et al., 2013). Following this

358

thermodynamic approach, the limitation of acetate could be the key to promote the

359

partial DNRA-anammox (reaction 7, -1984 kJ mol-acetate-1) over both heterotrophic

360

DNRA (reaction 2) and heterotrophic denitrification (reaction 1), as well as the full

361

DNRA-anammox pathway (reaction 8). This can be explained because more energy is

362

generated when supplying externally the ammonium (partial DNRA-anammox, Figure

363

1C) than when forming the ammonium through DNRA and then converting it during the

364

regular anammox reaction (full DNRA-anammox, Figure 1D). Also, the partial DNRA-

365

anammox is the only pathway in which anammox bacteria could outcompete

366

heterotrophic denitrifiers from a thermodynamic perspective when limiting the carbon

367

source (more energy is produced, 1984 kJ acetate-mol-1 during partial DNRA-anammox

368

vs. 797 during heterotrophic denitrification). Although the energy usage of the

369

anammox driven DNRA pathway is not unravelled yet, experimental results have

370

shown that heterotrophic transformations can be outcompeted by anammox bacteria

371

performing DNRA at low COD/N ratios (Winkler et al., 2012b). Overall it must be

15

372

noted that anammox can generate ammonium and/or nitrite from the DNRA pathway

373

and hence support their autotrophic pathway under conditions in which they are lacking

374

their electron acceptor or donor (or both). This therefore yields in an energy gain

375

irrespective from the DNRA pathway.

376

3.3.3. Effect of inorganic carbon

377

The DNRA reaction yields inorganic carbon (reaction 2 in Table 2), while the anammox

378

reaction uses inorganic carbon for biomass growth. The effect of inorganic carbon on

379

the combined partial DNRA-anammox transformation by anammox bacteria was

380

investigated with an anammox SBR, which was operated with alternating feeding

381

regimes of heterotrophic conditions (ammonium, nitrate and acetate), and autotrophic

382

conditions (ammonium and nitrite). The reactor was run without the addition of

383

inorganic carbon and therefore CO2 must have been provided biotically via the

384

oxidation of acetate through the DNRA pathway (Case C, Figure 4).

385

The rNH4, rNO3 and rCOD were studied for two different COD/NO3--N ratios in the

386

influent (1.1 and 0.6 g COD (g N)-1, Figure 4A and B, respectively). For the feeding

387

ratio COD/NO3--N = 1.1, the consumption rate ratios obtained (rCOD/rNO3 = 1.61 and

388

1.19 in Figure 4A) were very similar to the stoichiometric ratios COD/NO3--N

389

corresponding with partial DNRA-anammox (1.14 in Table 2 reaction 7). This, together

390

with the fact that no significant biomass growth was observed (average biomass

391

concentration during the operation of the SBR = 1.14 ±0.02 g VSS L-1), suggests that

392

partial DNRA-anammox by anammox bacteria dominated over heterotrophic

393

conversions during the feeding ratio COD/NO3--N of 1.1. The inorganic carbon

394

produced during the partial DNRA by anammox bacteria presumably supported the

395

regular autotrophic anammox reaction.

16

When a COD/NO3--N ratio of 0.6 was fed, the rNH4 became negative during the

396 397

mixing step, indicating formation of ammonium through DNRA, and the rNO3 decreased

398

by 62.1%. With a lower feeding ratio, the inorganic carbon formation through partial

399

DNRA was lower since less COD was added and lower oxidation of acetate by DNRA

400

occurred, thus limiting the anammox activity. Overall, the inorganic carbon supplied by

401

the DNRA pathway needs to be sufficient to maintain the energy yielding anaerobic

402

ammonium oxidation. 3.4. Anammox conversion in the presence of heterotrophic bacteria

403 404

Batch tests were performed with nitrite, nitrate, and acetate for Case B (high presence of

405

heterotrophic bacteria, PAOs) (Figure 3) to assess the competitiveness of the anammox

406

pathway in presence of heterotrophic bacteria and organic carbon compounds. Since the

407

addition of methanol did not inhibit the nitrate/nitrite and acetate consumption rates in

408

the absence of ammonium (B.4 vs. B.5 and B.7 vs. B.8 in Figure 3), it is clear that

409

heterotrophic denitrification was the dominant pathway (Jensen et al., 2007) consuming

410

nitrate/nitrite and acetate for these batch tests. When ammonium was added (together

411

with nitrate/nitrite, and acetate) (B.7 and B.8 in Figure 3.3), it was found that anammox

412

bacteria oxidized ammonium in the presence of nitrate (rNH4 = 0.982 mg N (gVSS)-1 h-

413

1

414

0.077 mg N (g VSS)-1 h-1). Thus, anammox bacteria were more competitive for the

415

nitrite formed from the reduction of nitrate (most likely by heterotrophs) than for the

416

nitrite supplied directly in the medium, when having acetate present. This could be

417

explained by a limitation on the heterotrophic denitrification since the acetate was

418

completely consumed after 20 minutes (see SI, Figure S7. 1C, D and E). When nitrite

419

was supplied, it was taken up by the heterotrophs, but when nitrate was added, only

420

partial denitrification (NO3-
NO2-) took place, which was indicated by nitrite

), but only a negligible consumption was found in the experiments with nitrite (rNH4 =

17

421

accumulation in the system (Figure S7. 1C, D and E). Therefore, the organic carbon

422

limitation (low COD/N ratio) together with the presence of nitrate can benefit the

423

anammox reaction, since the heterotrophic denitrification yields nitrite that can be used

424

along with ammonium by anammox bacteria (partial denitrification-anammox, Figure

425

1B). This finding also confirms the results obtained in the modelling study of

426

Mozumder et al. (2014), in which the authors pointed out that low concentrations of

427

COD could benefit the anammox reaction by the heterotrophic reduction of nitrate to

428

nitrite.

429

3.5. Partial DNRA-anammox by anammox bacteria for more sustainable

430

municipal wastewater treatment

431

The main bottleneck to implement anammox technology in the mainstream of the

432

WWTP is an effective supply of nitrite to the anammox reaction (Ma et al., 2016),

433

which is troublesome due to the suppression of NOB at low temperatures (Lotti et al.,

434

2015). If the DNRA pathway of anammox bacteria could be promoted in the

435

mainstream, the suppression of NOB would not be so crucial anymore since nitrate

436

could be reduced by anammox bacteria with some organic carbon, while keeping sludge

437

production low given that anammox bacteria do not grow on organic substrates (Kartal

438

et al., 2007b; Winkler et al., 2012a). Also, stable operational conditions are a critical

439

point during mainstream wastewater treatment, which are more difficult to obtain when

440

aiming at nitritation (NH4 +
NO2-) instead of nitrification (NH4+
NO3-). The

441

feasibility of the partial DNRA-anammox conversion (reaction 7 in Table 2 and Figure

442

1C) by anammox bacteria was demonstrated in this study (Figure 3 and Figure 4) and in

443

previous research (Guven et al., 2005; Kartal et al., 2007a; Winkler et al., 2012a;

444

Winkler et al., 2012b).

18

445

In order to decrease the nitrate accumulation produced by NOB (or by anammox

446

itself) in the mainstream, the partial DNRA-anammox process could be used and

447

implemented for enhanced nitrogen removal according to the scheme shown in Figure

448

5. In the suggested approach the organic matter from the wastewater is removed in a

449

high rate activated sludge (HRAS) system and captured in biomass, which is used to

450

maximize the energy recovery in the digester (Jetten et al., 1997). Since the HRAS

451

system is operated at low hydraulic retention time, nitrification is prevented and the

452

effluent of the HRAS is high in ammonium. This effluent could be treated in two steps:

453

conversion of half of the ammonium into nitrate by means of partial nitrification in an

454

aerobic reactor; and partial DNRA-anammox by anammox bacteria in an anoxic unit

455

(see Figure 5). The advantage would be to not only remove the nitrate generated by

456

NOB but also the nitrate produced during the anammox conversion itself, thus

457

increasing the effluent quality. In order to promote the partial DNRA pathway of

458

anammox bacteria, volatile fatty acids (VFA) need to be supplemented to the system

459

since anammox bacteria have not been reported to be able to utilize e.g. sugars or

460

particulate COD. Therefore, it is suggested to generate VFAs onsite in a small

461

fermentation unit receiving sludge from the HRAS reactor. This external supply of VFA

462

would be associated with some costs but if compared to conventional nitrification-

463

denitrification systems, up to 50% energy savings in aeration would be obtained by

464

applying partial DNRA-anammox, since only half of the ammonium would be oxidized

465

to nitrate. A fraction of 21% of the COD contained in the sludge from the HRAS system

466

would need to be converted into VFA in the fermenter to promote the partial DNRA-

467

anammox (see section S8 in the SI for details on the calculation of the COD needed to

468

carry out the partial DNRA-anammox).

19

469

Regarding the practical implementation of the nitrogen removal system in

470

Figure 5, a two reactor configuration is suggested, since for mainstream applications the

471

use of two reactors would not increase significantly the overall investment costs, while

472

the control in two-unit configuration would be easier than in one reactor. The first

473

aerated reactor could still aim for partial nitritation and the accumulation of nitrite,

474

which has however been shown to be the major bottleneck in the application of

475

anammox in the main line (Lotti et al., 2015). It is therefore a fair assumption that

476

nitrate and not nitrite will be dominant.

477

The use of granular sludge technology for the partial DNRA-anammox reactor

478

circumvents the necessity of settlers for the biomass separation, leading to compact and

479

high rate process units. Besides, literature suggests that in granular sludge reactors

480

heterotrophs could be more easily washed out from the system since these

481

microorganisms grow mainly in flocs (Lotti et al. (2015)).

482

Overall, all aspects discussed in section 3.3 and 3.4 would facilitate the success

483

of the partial DNRA-anammox process. The presence of ammonium in the medium

484

would favour the out-competition of heterotrophs, as shown in this study (Figure 3) and

485

by others (Kartal et al., 2007b; Kartal et al., 2008). Also, when ammonium is available

486

in the medium, the conversion of nitrite to ammonium by anammox bacteria is

487

prevented (full-DNRA) since this conversion is slower than the nitrate reduction to

488

nitrite (Figure 3 in this study and Kartal et al. (2007a)), thus promoting partial DNRA-

489

anammox (conversion in Figure 1C instead of D). The presence of nitrate instead of

490

nitrite could make anammox bacteria competitive for the nitrite formed during partial

491

DNRA or partial denitrification (Figure 3 and Mozumder et al. (2014)). However, as

492

shown in Figure 4, the insufficient availability of inorganic carbon for anammox

493

bacteria to grow could impede the combined partial DNRA-anammox process.

20

494

Furthermore, Candidatus ‘Brocadia fulgida’, a common species in full-scale WWTPs

495

and found in this study (Figure S4. 1), has proven to be a suitable organism for

496

performing partial DNRA-anammox, without previous adaptation to VFA (Figure 3.1).

497

This study is the first one suggesting the exploitation of the organotrophic nitrate

498

reduction capacity of anammox bacteria to facilitate the implementation of the regular

499

anammox conversion in the main line of the wastewater treatment plant. However,

500

further research is needed to proof the stability of the process long-term.

501 502

4. CONCLUSIONS

503

The competition among heterotrophs and anammox bacteria for organic carbon and

504

nitrate was studied to gain understanding on the DNRA capacity of anammox bacteria.

505

The species Candidatus ‘brocadia fulgida’ showed high DNRA performance even when

506

the culture was not previously adapted to organic carbon. Gibbs free energy calculations

507

showed that limiting organic carbon (low COD/N influent ratio) favoured the partial

508

DNRA (NO3-
NO2-)-anammox conversion over the heterotrophic DNRA and

509

heterotrophic denitrification. Also, the presence of ammonium and sufficient inorganic

510

carbon enhanced the process. Overall, the combined partial DNRA-anammox could be

511

applied for sustainable nitrogen removal from municipal wastewater.

512 513

ACKNOWLEDGMENTS

514

The research leading to these results has received funding from the People Program

515

(Marie Curie Actions) of the European Union’s Seventh Framework Programme

516

FP7/2007-2013, through the REA agreement 289193 - Project SANITAS; the special

517

research fund (BOF) (no. 01DI4415) 2015-Finalizing doctoral scholarship, Ghent

518

University; and the COST Action ES1202, Water 2020. Mari Winkler was funded by a

21

519

Marie Curie Intra-European fellowship (PIEF-GA-2012-329328). This publication

520

reflects only the authors’ views and the European Union is not liable for any use that

521

may be made of the information contained therein. The authors further thank Udo van

522

Dongen (technician at TUD) for his help during the FISH determinations and the staff

523

of the Rotterdam and Amersfoort WWTPs for proving the biomass used in this study.

524 525

REFERENCES

526 527 528 529 530 531 532 533 534 535 536 537 538 539 540 541 542 543 544 545 546 547 548 549 550 551 552 553 554 555 556 557 558 559 560

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Kartal, B., Kuenen, J.G., van Loosdrecht, M.C.M., 2010. Sewage treatment with Anammox. Science. 328, 702-703. Kartal, B., Kuypers, M.M.M., Lavik, G., Schalk, J., den Camp, H.J.M.O., Jetten, M.S.M., Strous, M., 2007a. Anammox bacteria disguised as denitrifiers: nitrate reduction to dinitrogen gas via nitrite and ammonium. Environmental Microbiology. 9, 635-642. Kartal, B., Rattray, J., van Niftrik, L.A., van de Vossenberg, J., Schmid, M.C., Webb, R.I., Schouten, S., Fuerst, J.A., Damste, J.S.S., Jetten, M.S.M., Strous, M., 2007b. Candidatus “Anammoxoglobus propionicus” a new propionate oxidizing species of anaerobic ammonium oxidizing bacteria. Systematic and Applied Microbiology. 30, 39-49. Kartal, B., van Niftrik, L., Rattray, J., de Vossenberg, J.L.C.M.V., Schmid, M.C., Damste, J.S.S., Jetten, M.S.M., Strous, M., 2008. Candidatus “Brocadia fulgida”: an autofluorescent anaerobic ammonium oxidizing bacterium. Fems Microbiology Ecology. 63, 46-55. Kraft, B., Tegetmeyer, H.E., Sharma, R., Klotz, M.G., Ferdelman, T.G., Hettich, R.L., Geelhoed, J.S., Strous, M., 2014. The environmental controls that govern the end product of bacterial nitrate respiration. Science. 345, 676-679. Lotti, T., Kleerebezem, R., Hu, Z., Kartal, B., de Kreuk, M.K., Kip, C.V.T., Kruit, J., Hendrickx, T.L.G., van Loosdrecht, M.C.M., 2015. Pilot-scale evaluation of anammox-based mainstream nitrogen removal from municipal wastewater. Environmental Technology. 36, 1167-1177. Lotti, T., Kleerebezem, R., Hu, Z., Kartal, B., Jetten, M.S.M., van Loosdrecht, M.C.M., 2014. Simultaneous partial nitritation and anammox at low temperature with granular sludge. Water Research. 66, 111-121. Ma, B., Wang, S.Y., Cao, S.B., Miao, Y.Y., Jia, F.X., Du, R., Peng, Y.Z., 2016. Biological nitrogen removal from sewage via anammox: Recent advances. Bioresource Technology. 200, 981-990. Mozumder, M.S.I., Picioreanu, C., van Loosdrecht, M.C.M., Volcke, E.I.P., 2014. Effect of heterotrophic growth on autotrophic nitrogen removal in a granular sludge reactor. Environmental Technology. 35, 1027-1037. Pérez, J., Lotti, T., Kleerebezem, R., Picioreanu, C., van Loosdrecht, M.C.M., 2014. Outcompeting nitrite-oxidizing bacteria in single-stage nitrogen removal in sewage treatment plants: A model-based study. Water Research. 66, 208-218. Russ, L., Harhangi, H.R., Schellekens, J., Verdellen, B., Kartal, B., Op den Camp, H.J.M., Jetten, M.S.M., 2012. Genome analysis and heterologous expression of acetate-activating enzymes in the anammox bacterium Kuenenia stuttgartiensis. Archives of Microbiology. 194, 943-948. Shu, D.T., He, Y.L., Yue, H., Zhu, L., Wang, Q.Y., 2015. Metagenomic insights into the effects of volatile fatty acids on microbial community structures and functional genes in organotrophic anammox process. Bioresource Technology. 196, 621-633. Siegrist, H., Salzgeber, D., Eugster, J., Joss, A., 2008. Anammox brings WWTP closer to energy autarky due to increased biogas production and reduced aeration energy for N-removal. Water Science and Technology. 57, 383-388. van de Graaf, A.A.V., deBruijn, P., Robertson, L.A., Jetten, M.S.M., Kuenen, J.G., 1996. Autotrophic growth of anaerobic ammonium-oxidizing micro-organisms in a fluidized bed reactor. Microbiology-Uk. 142, 2187-2196.

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van den Berg, E.M., van Dongen, U., Abbas, B., van Loosdrecht, M.C.M., 2015. Enrichment of DNRA bacteria in a continuous culture. Isme Journal. 9, 27492749. van der Star, W.R.L., Abma, W.R., Blommers, D., Mulder, J.W., Tokutomi, T., Strous, M., Picioreanu, C., Van Loosdrecht, M.C.M., 2007. Startup of reactors for anoxic ammonium oxidation: Experiences from the first full-scale anammox reactor in Rotterdam. Water Research. 41, 4149-4163. Vazquez-Padin, J., Fernadez, I., Figueroa, M., Mosquera-Corral, A., Campos, J.L., Mendez, R., 2009. Applications of anammox based processes to treat anaerobic digester supernatant at room temperature. Bioresource Technology. 100, 29882994. Wett, B., 2007. Development and implementation of a robust deammonification process. Water Science and Technology. 56, 81-88. Winkler, M.K.H., Kleerebezem, R., Kuenen, J.G., Yang, J.J., van Loosdrecht, M.C.M., 2011. Segregation of biomass in cyclic anaerobic/aerobic granular sludge allows the enrichment of anaerobic ammonium oxidizing bacteria at low temperatures. Environmental Science & Technology. 45, 7330-7337. Winkler, M.K.H., Kleerebezem, R., van Loosdrecht, M.C.M., 2012a. Integration of anammox into the aerobic granular sludge process for main stream wastewater treatment at ambient temperatures. Water Research. 46, 136-144. Winkler, M.K.H., Yang, J.J., Kleerebezem, R., Plaza, E., Trela, J., Hultman, B., van Loosdrecht, M.C.M., 2012b. Nitrate reduction by organotrophic Anammox bacteria in a nitritation/anammox granular sludge and a moving bed biofilm reactor. Bioresource Technology. 114, 217-223.

24

634

FIGURE CAPTIONS

635

Figure 1 Comparison of different nitrogen removal configurations involving the

636

anammox process: A) partial nitritation-anammox; B) partial denitrification-anammox;

637

C) partial DNRA (reduction of nitrate to nitrite only by anammox bacteria) and

638

anammox reaction with the nitrite formed and the ammonium available in the

639

environment; D) partial DNRA and anammox reaction with ammonium formed during

640

full DNRA by anammox bacteria.

641 642

Figure 2 Anammox activities during the A) batch tests of Case A, with biomass from a

643

full-scale anammox granular sludge reactor (A.1, A.2 and A.3); B) batch tests of Case

644

B, with biomass from a SBR with high presence of heterotrophic bacteria (B.1 and B.2),

645

and C) operation of the SBR of Case C under the ‘Anammox regime’ (Feed C.1); when

646

having a media composition with (AN+MeOH) and without (AN) methanol. The

647

corresponding concentrations of each test are listed in Table 1.

648 649

Figure 3 Influence of acetate on the ammonium, nitrate, nitrite and acetate consumption

650

rates (rNH4, rNO3, rNO2 and rCOD). Results during the batch tests with enriched anammox

651

biomass from a full-scale granular sludge reactor of Case A (A.3 to A.5, Table 1) and

652

from a SBR containing anammox and heterotrophic bacteria of Case B (B.3 to B.8,

653

Table 1) under different feeding conditions: nitrate as electron acceptor with (1) and

654

without (2) ammonium in the medium; and nitrite as electron acceptor (3).

655 656

Figure 4 Effect of inorganic carbon absence on the ammonium, nitrate and acetate

657

consumption rates (rNH4, rNO3 and rCOD) for the ‘Regime under semi-continuous

658

exposure to acetate’ in Case C under two COD/NO3--N feeding ratios.

25

659

Figure 5 Suggested scheme for more sustainable municipal wastewater treatment with

660

nitrification of half of the ammonium + partial DNRA-anammox by anammox bacteria

661

as nitrogen removal system.

662

26

663

FIGURES AND TABLES A

B

Partial nitritation-anammox

Partial nitrification Partial denitrification-anammox Partial nitratation

Partial nitritation

1.1NH4+ AOB 1.1NO2 - NOB 1.1NO3-

1.1NH4+ AOB 1.1NO2- NOB 1.1NO3 - AEROBIC

2NH4 + 0.9NH4+

ANX

2NH4+ Anammox reaction

0.55N2

Den

0.9NH4 +

AOB: ammonium oxidizing bacteria NOB: nitrite oxidizing bacteria ANX: anammox bacteria Den: denitrifiers

1.1NO2-

ANX

Den

ANOXIC

Partial denitrification

Anammox reaction

0.9N2 + 0.2 NO3 -

0.9N2 + 0.2 NO3 C

D

Partial nitrification Partial DNRA-anammox

Nitrification Partial DNRA + full DNRA-anammox

Partial nitratation

NOB 1.1NH4+ AOB 1.1NO21.1NO3-

2NH4+

1.1NH4 + 0.9NH4+

ANX

1.1NO2-

ANX Partial DNRA

2NH4 +

AOB

0.9NO2 0.9NH4 + 1.1NO2-

ANX Anammox reaction

0.9N2 + 0.2 NO3

2NO2-

Nitratation NOB

2NO3-

ANX Full DNRA

AEROBIC ANOXIC

ANX Partial DNRA

Anammox reaction -

0.9N2 + 0.2 NO3 -

664 665

Figure 1 Comparison of different nitrogen removal configurations involving the

666

anammox process: A) partial nitritation-anammox; B) partial denitrification-anammox;

667

C) partial DNRA (reduction of nitrate to nitrite only by anammox bacteria) and

668

anammox reaction with the nitrite formed and the ammonium available in the

669

environment; D) partial DNRA and anammox reaction with ammonium formed during

670

full DNRA by anammox bacteria.

671

27

Consumption rate (r) (mg N (g VSS)-1 h-1 )

18 15

A

Case A

B

Case B

C

Case C 16.1

13.8

NH4+ NH4 + 12

11.1

NO2NO2 -

8.94

9 6 3.68

3

0.54 0.95

1.69 0.18 0.36

Not determined

0 AN (A.1)

672

AN+MeOH (A.2)

AN (B.1)

AN+MeOH AN AN+MeOH (B.2) (Feed C.1)

Test

673

Figure 2 Anammox activities during the A) batch tests of Case A, with biomass from a

674

full-scale anammox granular sludge reactor (A.1 and A.2); B) batch tests of Case B,

675

with biomass from a SBR with high presence of heterotrophic bacteria (B.1 and B.2),

676

and C) operation of the SBR of Case C under the ‘Anammox regime’ (Feed C.1); when

677

having a media composition with (AN+MeOH) and without (AN) methanol. The

678

corresponding concentrations of each test are listed in Table 1.

679

28

6

2 Case B

N consumption rate (r N) (mg N (g VSS)-1 h-1)

21.1

20.2

4.74

4

3

25

4.83 20

+ 4 NH4+

+ NH4+ 4

NO33

NO22

Acetate

Acetate 2.63

15

2.65 10

2

1.71

6.98

7.32

0.982

1

0.930

0.685

1.07 3.48

1.05

4.12

5

1.62

0.045 0.112

0

680

3 Cas e B

Case B 22.9

5.17

5

Case A

Acetate consumption rate (rCOD) (mg COD (g VSS)-1 h-1)

1 Case A

0.077

0

A.3

A.4

B.3

A.5

B.4

B.5

B.6

B.7

B.8

NH 4+ NO3Acetate

NH4+ NO3Acetate MeOH

NH4+ NO3Acetate

NO3Acetate

NO3Acetate

NO3Acetate MeOH

NH 4+ NO2Acetate

NO2Acetate

NO2Acetate MeOH

Batch test

681

Figure 3 Influence of acetate on the ammonium, nitrate, nitrite and acetate consumption

682

rates (rNH4, rNO3, rNO2 and rCOD). Results during the batch tests with enriched anammox

683

biomass from a full-scale granular sludge reactor of Case A (A.3 to A.5, Table 1) and

684

from a SBR containing anammox and heterotrophic bacteria of Case B (B.3 to B.8,

685

Table 1) under different feeding conditions: nitrate as electron acceptor with (1) and

686

without (2) ammonium in the medium; and nitrite as electron acceptor (3).

687

29

N and acetate consumption rate (r) mg N (or COD) (VSS)-1 h-1

A Feed COD/NO3- -N = 1.1

15 11.2

12

11.5

+ NH4+ 4 NO33 Acetate

13.7

9.90

9 6.98

6 4.02

3

4.02 3.00

2.37

1.55 -1.94

0 -3

688

13.7

B Feed COD/NO3--N = 0.6

1.61 Feeding

1.19 Mixing

2.46 Feeding

4.57 Mixing

Feeding + mixing

Mixing

Feeding + mixing

Mixing

rCOD/rNO3

689

Figure 4 Effect of inorganic carbon absence on the ammonium, nitrate and acetate

690

consumption rates (rNH4, rNO3 and rCOD) for the ‘Regime under semi-continuous

691

exposure to acetate’ in Case C under two COD/NO3--N feeding ratios.

692

30

Nitrogen removal system

High rate activated sludge system

Influent

Settler NH4 +

HRAS (COD removal)

1/2 NH4 +

Nitrification NH4+
NO3-

1/2 NO3 -

Partial DNRAanammox

Effluent

1/2 NH4 +

Sludge

VFA Partial DNRA-anammox

Thickener

Fermenter

Anaerobic digester Dewat.

F

AD

NH4 +

Sludge

693 694

Figure 5 Suggested scheme for more sustainable municipal wastewater treatment with

695

nitrification of half of the ammonium + partial DNRA-anammox by anammox bacteria

696

as nitrogen removal system.

697

31

698

Table 1 Initial substrate concentrations for the batch tests with biomass from a full-

699

scale anammox granular sludge reactor (Case A), from a bubble column SBR

700

containing anammox bacteria and PAOs among other communities (Case B), and feeds

701

used during the operation of the SBR of Case C. NH4+ NO3NO2Acetate (g N m-3) (g N m-3 ) (g N m-3) (g COD m-3) Case A Batch test 200 A.1 NH4+ + NO2200 A.2 NH4+ + NO2- + MeOH 200 A.3 NH4+ + NO 3- + acetate 100 A.4 NH4+ + NO 3- + acetate + MeOH A.5 NO3- + acetate Case B Batch test 147 B.1 NH4+ + NO 2147 B.2 NH4+ + NO 2- + MeOH 147 B.3 NH4+ + NO 3- + acetate B.4 NO3- + acetate B.5 NO3- + acetate + MeOH 147 B.6 NH4+ + NO 2- + acetate B.7 NO2- + acetate B.8 NO2- + acetate + MeOH Case C ‘Anammox regime’ 125-200 Feed C.1 NH4+ + NO 2‘Regime under semi-continuous exposure to acetate’ Feed C.2 NH4+ + NO 3- + acetate (no HCO3-) 50-100 Feed C.3 NH4+ + NO 2- (no HCO3- ) 50-200

200 200 100 50 50

12 460 230 230

50 50 50 50 50 50 50 50

MeOH (mM)

12

5 76 76 76 76 76 76

5

5

100-200 50-100

50-100 50-200

702 703

32

704

Table 2 Catabolic reaction stoichiometry, Gibbs free energy per substrate, N2 produced

705

and electron transferred, and stoichiometric ratio COD/NO3- for different pathways. Biological process

Reaction

kJ (e -eq)-1

kJ (mol NO3 or NO2 -)-1

kJ (mol acetate)-1

kJ (mol N2)-1

Stoichiometric COD/NO3(g COD (g N)-1)

1

Denitrification

8NO3 - + 5CH3 COO- + 8H+ → 5CO2 + 4N2 + 5HCO3 - + 9H2O

-99.60

-498.0

-796.8

-996.0

2.86

2

DNRA

NO3 - + CH3 COO- + 2H+ → CO2 + NH4+ + HCO3-

-62.51

-500.1

-500.1

3

4

5

6 7 8

4.57

Partial DNRA/Denitri 4NO3 - + CH3 COO- → CO2 + 4NO2 - + HCO3- + H2 O -69.05 -138.1 -552.4 1.14 fication* Nitrite reduced via NO2 - + NH4 + → N2 + 2H2 O -119.3 -357.8 -357.8 Anammox Nitrite 4NO2 - + 3CH3COO- + 8H+ + H2 O → 3CO2 + 3HCO3- + reduced via -60.33 -362.0 -482.6 3.43 4NH4 + DNRA Nitrite 8NO2 - + 3CH3COO- + 8H+ → 3CO2 + 3HCO3- + 4N2 + reduced via 7H2O -119.96 -359.9 -959.7 -719.8 1.71 Denitrification Partial DNRA 4NO3 - + CH3 COO- + 4NH4 + → CO2 + HCO3 - + 4N2 + 9H2 O -99.18 -495.9 -1984 -495.9 1.14 + Anammox** + Full DNRA + 8NO3 + 5CH3 COO + 8H → 5CO2 + 5HCO3 + 4N2 + -99.60 -498.0 -796.8 -996.0 2.86 Anammox*** 9H2O * Partial DNRA = partial denitrification, i.e. reduction of nitrate to nitrite only ** All the NH4+ comes from the environment, i.e. only the first step of the DNRA process occurs (NO3 -
NO2-) (Figure 1C) *** All NH4+ comes from the DNRA process, i.e. all NO3 - is converted to NO2- and half of the NO2 - is further converted to NH4+ (Figure 1D)

706 707 708 709

33

DNRAanammox

Nitratation - NOB

NH4+

AOB

NH4+

ANX

2NH4+

NH4+

Conditions

Process

NO3-

NO2

NO2-

ANX

AOB: ammonium oxidizing bacteria NOB: nitrite oxidizing bacteria ANX: anammox bacteria

ANX Partial DNRA

Anammox reaction

N2

Presence of NH4+ Sufficient inorganic C

DNRAanammox

AEROBIC ANOXIC

NO3- e- acceptor

COD/N ↓↓

Heterotrophic DNRA

Heterotrophic Denitrification

COD/N ↑↑

NO3-, NO2- eacceptor

710

Highlights

711

-

Dissimilatory nitrate reduction to ammonium (DNRA) by anammox bacteria

712

-

Limitation of COD and presence of ammonium favoured anammox bacteria to

713 714

pursue DNRA -

DNRA by anammox bacteria could improve nitrogen removal in the main line

715 716 717 718

34