Importance of the combined effects of dissolved oxygen and pH on optimization of nitrogen removal in anammox-enriched granular sludge

Accepted Manuscript Title: Importance of the combined effects of dissolved oxygen and pH on optimization of nitrogen removal in anammox-enriched granular sludge Author: Zhixuan Yin Carla Elo´ısa Diniz dos Santos Jaime Gonzalez Vilaplana Dominika Sobotka Krzysztof Czerwionka M´arcia Helena Rissato Zamariolli Damianovic Li Xie Francisco Jes´us Fern´andez Morales Jacek Makinia PII: DOI: Reference:

S1359-5113(16)30148-9 http://dx.doi.org/doi:10.1016/j.procbio.2016.05.025 PRBI 10699

To appear in:

Process Biochemistry

Received date: Revised date: Accepted date:

28-1-2016 15-5-2016 23-5-2016

Please cite this article as: Yin Zhixuan, dos Santos Carla Elo´ısa Diniz, Vilaplana Jaime Gonzalez, Sobotka Dominika, Czerwionka Krzysztof, Damianovic M´arcia Helena Rissato Zamariolli, Xie Li, Morales Francisco Jes´us Fern´andez, Makinia Jacek.Importance of the combined effects of dissolved oxygen and pH on optimization of nitrogen removal in anammox-enriched granular sludge.Process Biochemistry http://dx.doi.org/10.1016/j.procbio.2016.05.025 This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

Importance of the combined effects of dissolved oxygen and pH on optimization of nitrogen removal in anammox-enriched granular sludge

Zhixuan Yin1, Carla Eloísa Diniz dos Santos2, Jaime Gonzalez Vilaplana3, Dominika Sobotka4, Krzysztof Czerwionka4, Márcia Helena Rissato Zamariolli Damianovic2, Li Xie1, Francisco Jesús Fernández Morales3, Jacek Makinia 4

¹ State Key Laboratory of Pollution Control and Resources Reuse, Key Laboratory of Yangtze River Water Environment, College of Environmental Science and Engineering, Tongji University, 1239 Siping Road, Shanghai 200092, PR China 2

Laboratory of Biological Processes, Center for Research, Development and Innovations in Environmental Engineering, School of Engineering of São Carlos, University of São Paulo, Av. João Dagnone 1100, 13563-120, São Carlos, Brazil

3

Department of Chemical Engineering, University of Castilla-La Mancha, Enrique Costa Building, Campus Universitario s/n, 13071, Ciudad Real, Spain

4

Faculty of Civil and Environmental Engineering, Gdańsk University of Technology, ul. Narutowicza 11/12, 80-233, Gdańsk, Poland

1

Graphical Abstract

2

Highlights:  The highest SNRR was achieved in non-aerated conditions due to anammox metabolism.  A local SNRR peak was observed at DO ranging from 0.5 to 0.6 mg O2 L-1.  AOB, NOB and denitrifiers were also active in anammox-enriched granular sludge.  A decline in the SNRR occurred once pH deviated from the optimum range (6.5-8.5).  DO elbow was ultimate indication of nitrogen removal by oxygen-consuming bacteria.

3

Abstract: The combined effects of dissolved oxygen (DO) and pH on nitrogen removal were investigated in a laboratory-scale sequencing batch reactor (SBR) with anammox-enriched granular sludge obtained from a nitritation/anammox system. The highest specific nitrogen removal rate (SNRR) (1.1 gN gVSS-1 d-1) was observed under non-aerated conditions, resulting in the nitrogen removal efficiency of 81.6%. Although nitrogen removal was readily inhibited under aerated conditions, an increased SNRR occurred at the DO concentration of 0.5 mg O2 L-1. This is in contrast with the directional DO suppression on nitrogen removal in the anammox process, indicating that other nitrogen conversion pathways, such as nitrification and endogenous denitrification, were also active in the studied reactor. The highest SNRRs were obtained within a pH range of 6.5–8.5, characterized by low concentrations of free ammonia (FA) and free nitrous acid (FNA). Oxygen-consuming bacteria (nitrifiers) were implicitly inhibited by the low pH so that less oxygen was utilized. In tests carried out using the optimum pH physiological range (6.5-8.5) for microorganisms, an explicit DO elbow was observed when the ammonium and nitrite were almost completely removed. This finding confirms the occurrence of nitrogen removal by the oxygen-consuming bacteria in the system studied.

Keywords: Anammox; granular sludge; dissolved oxygen; pH; nitrogen removal; nitrification; endogenous denitrification.

4

1. INTRODUCTION Biogranulation is an effective strategy for obtaining both stable removal with high loading rates, as well as retaining the biomass in the system due to good settling properties of biogranules. These two features are particularly important for the anaerobic ammonium oxidation (anammox) process which has been widely recognized as a viable alternative to the conventional nitrification-denitrification in sidestream reactors and recently also in mainstream reactors. Due to the fact that the specific growth rate constant of anammox bacteria is very slow, i.e. μ=0.065 d-1 [1, 2], it is important to operate the anammox system at optimum conditions to support the sufficient growth rate of these microorganisms. However, the anammox process has been found to be inhibited by either inhibitory substances, such as substrates (ammonia and nitrite), organic matter (nontoxic organic matter and toxic organic matter), salts, heavy metals, phosphate and sulfide, or under unfavorable operating conditions, including too high dissolved oxygen (DO) concentration and/or non-optimal (too low or too high) pH shock [3].

The DO concentration should be strictly controlled to avoid reversible or even irreversible inhibition. For example, Egli et al [4] observed the reversible and irreversible anammox inhibition at the DO concentration of 0.08 mg O2 L-1 and 1.44 mg O2 L-1, respectively. The irreversible inhibition was also observed in intermittently aerated systems at the DO concentrations in the range of 0.1-0.16 mg O2 L-1 [5, 6]. In order to obtain the overall nitrogen removal effect, the anammox process has to be integrated with the aerobic process such as

5

nitritation [6, 7] or simultaneous nitritation/denitrification [8, 9]. The DO impact on these single biochemical processes (nitritation and denitrification) is well understood, but there has been no clear comprehension regarding the mechanisms of the DO effect on nitrogen removal and the transformation pathways considering the coexistence of diverse microorganisms in anammox-enriched nitrogen removal systems.

With regard to pH, extreme changes in this parameter should be avoided to maintain the anammox metabolism. The pH can directly affect the bacterial growth and its enzymatic activities. Indirectly, the pH affects the anammox activity through changes in the concentrations of pH-dependent, unionized and presumably more toxic forms of free ammonia (FA) and free nitrous acid (FNA) substrates [10]. Strous et al [11] reported the physiological range of pH between 6.7 and 8.3 in a dispersed aggregated anammox system. Egli et al [4] found the optimum anammox activity at pH = 7.5-8.0 in a wider examined range (6.5-9.0) in a rotating biological contactor (RBC) treating leachate. On one hand, stable anammox reactions were obtained at the pH range as high as 8.5-9.3 [12], 9.3-9.5 [13] and even at pH=6.5 [14]. In contrast, a complete inhibition on the anammox activity at pH=9.3 in a fixed-bed reactor was found by Fux et al [15]. These inconsistencies may be attributed to the differences in physical structures of sludge (flocculent sludge, biofilm or granular sludge), microbial population, influent characteristics, and operating conditions during the experiments. Thus, in addition to the overall nitrogen removal effect, the actual nitrogen transformation pathway in anammox-enriched granular sludge should be further investigated.

6

In spite of the significance regarding the anammox inhibition due to either too high DO concentration or unfavorable pH, there is little information available on the combined effects of these two factors on nitrogen removal in the anammox-enriched systems. This study therefore evaluated both effects of DO and pH on nitrogen removal by the anammox-enriched granular sludge, aiming at optimizing the operational conditions in a laboratory-scale sequencing batch reactor (SBR). For this purpose, kinetic bioassays were carried out at different DO concentrations and pH conditions. Furthermore, different models for the pH dependence of nitrogen removal rates were compared in terms of the inhibitory effects of pH itself or actual FA/FNA concentrations.

2. MATERIAL AND METHODS 2.1 Origin and characteristics of the anammox-enriched granular sludge Anammox sludge was obtained from a full-scale sidestream nitritation/anammox system in Zurich (Switzerland) and cultivated for more than one year to obtain a granulation effect [16]. The long-term granulation of anammox biomass was performed in a 10 L SBR (hydraulic retention time = 0.4 days) equipped with a thermostatic bath at 30 ºC, fed with a synthetic autotrophic medium, as recommended by Dapena-Mora et al [17]. The pH was controlled in the range of 7.5–7.8 by adding HCl (1 mol L-1). During the granulation studies, the SBR was operated in 2h cycles divided into four phases: mixed filling (30 min), reaction (60 min), settling (20 min) and drawing (10 min). Aggregates settling at a higher rate than the

7

minimum settling velocity (MSV) of 1.0 m h-1 were not washed out from the SBR. MSV was calculated as the vertical distance of the water volume decanted per cycle divided by the settling time. The cultivated anammox granules had a uniquely carmine color with an average diameter of approximately 0.73 mm. More information about the biomass characteristics and composition can be found elsewhere [18]. The phylogenetic analysis revealed that anammox bacteria (Planctomycetes) were exclusively represented by Candidatus Brocadia. Besides the anammox bacteria, ammonia oxidizing bacteria (AOB), such as Nitrosomonas sp., nitrite oxidizing bacteria (NOB) such as Nitrospira sp., and denitrifying heterotrophic bacteria from Alfa- and Beta-proteobacteria families were also identified by the metagenomic analysis [18].

2.2 Experimental setup and operation Biokinetic assays (Tests 1.1–1.7) at different DO concentration set points (0.0–1.0 mg O2 L-1) were carried out in the SBR with an open space and the pH maintained at 7.6-7.8. To elucidate the biological pathways involved in nitrogen removal at 0.3, 0.5, and 2.0 mg O2 L-1, two series of tests were performed in which only ammonium (Tests 2.1–2.2) or only nitrite (Tests 3.1–3.4) was added to the same SBR. Nitrous oxide (N2O) production was monitored using an online nitrous oxide microsensor (N2O25; Unisense, Aarhus, Denmark) and the data were logged every 30 s. Kinetic bioassays (Tests 4.1–4.10) at different initial pH values (6.0-10.5) were carried out in a batch reactor with a working volume of 4L. The required initial pH was adjusted by adding HCl (1 mol L-1) or NaOH (2 mol L-1). The initial DO concentrations were kept between 0.4 and 0.5 mg O2 L-1. During Tests 4.1–4.10, the pH and

8

DO concentrations remained uncontrolled, whereas Tests 5.1–5.4 were carried out in a completely sealed batch reactor (2.2 L of the working volume) and a N2 gas-tight balloon in order to maintain strictly anoxic conditions and an adequate pressure for sampling. Therefore, those tests were carried out at different initial pH values (6.0–10.5) and an initial DO concentration = 0.03 mg O2 L-1. A mechanical stirrer model (Heidolph, Germany) provided continuous mixing with the minimum stirring speed of 53 rpm to keep the granules in suspension.

Both reactors (SBR and batch reactor) were equipped with automated control systems for DO, pH, and temperature. Probes/electrodes (WTW GmbH, Germany) were used for the continuous monitoring of pH (SenTix21) as well as temperature and DO concentration (CellOx 325). The temperature was kept constant at 30±1ºC by means of the thermostatic bath. Prior to each test, anoxic conditions in the reactors were obtained by gassing the mixed liquor for approximately 5 min with N2.

Ammonium chloride (NH4Cl) and sodium nitrite (NaNO2) were used as the electron donor and electron acceptor, respectively, with supplemental sodium bicarbonate (NaHCO3) as an inorganic carbon source. The biomass characteristics, initial ammonium and nitrite concentrations, and operational conditions of the batch experiments are summarized in Table 1.

9

2.3 Analytical methods and calculations Samples of the mixed liquor were collected every 30 min over a period of 2–3 h depending on the observed ammonium and nitrite removal rates. The samples were filtered through 0.45 µm pore-size nitrocellulose membrane (Whatman, Kent, UK) and immediately analyzed. Ammonium (NH4+-N), nitrite (NO2--N), and nitrate (NO3--N) were measured in cuvette tests (Hach Lange, Dusseldorf, Germany) using a Xion 500 spectrophotometer (Dr Lange GmbH, Berlin, Germany). Total nitrogen (TN) concentrations were determined using a total organic carbon TOC/TN analyzer (Shimadzu, TOC-Vcan, Japan) equipped with a total nitrogen measurement unit. All conventional parameters such as pH, total suspended solids (TSS), and volatile suspended solids (VSS) were determined according to Standard Methods [19].

The specific nitrogen removal rate (SNRR) (gN g-1VSS d-1) was calculated based on TN removal measurements using the following equation: SNRR 

TN VSS  T

(1)

Where: ΔTN is TN removed (mg/L) during the reaction time T (d), and VSS (mg/L) refers to the specific measurement in each test. The specific N2O production rate and the ratio between N2O nitrogen emitted and the nitrite nitrogen converted (N2O emission factor) were both determined based on the on-line N2O monitoring data in the liquid phase. In addition to the pH dependence, the pH-related inhibition of anammox activity was attributed to the ionized forms of the substrates such as FA and FNA. Therefore, threshold

10

values for both FA and FNA were calculated based on the original relationships proposed by Anthonisen et al [20].

2.4 Models describing pH effects on nitrogen removal Michaelis model The Michaelis model predicts the pH effect on enzymatic reactions [21]. The model can be adopted for the observed SNRR as follows: SNRR 1   pKS 1  pH  SNRRmax 1  10  10 pH  pKS 2 

(2)

where: SNRR is the observed rate at the given pH (gN g-1VSS d-1), SNRRmax is the maximum rate at the optimum pH (gN g-1VSS d-1), pKS1 is the logarithm of the lowest pH at which SNRR is equal to ½ SNRRmax, and pKS2 is the logarithm of the highest pH at which SNRR is equal to ½ SNRRmax.

The pKS1 and pKS2 values were estimated using the Origin software (version 8.5) to minimize the residuals, i.e., the deviations of the SNRR/SNRRmax prediction curves from the experimental data. The Levenberg-Marquardt (L-M) algorithm was used in the iterative procedure.

Haldane model

11

The Haldane model describes the kinetics of pollutant biodegradation involving the inhibition by high concentrations of substrates [22]: SNNR 

SNNRmax K S 1 m  S Ki

(3)

where: Km is the half-saturation constant (mgL-1), Ki is the substrate inhibition constant (mgL-1), and S is the substrate concentration (mg L-1).

In the anammox process, ammonium and nitrite can be considered as both substrates and inhibitors. Therefore, Equation 3 could be expanded as follows: SNNR 

 K mNH4 1   S NH4 

SNNRmax S NH4  K mNO2 S NO2 1    K iNH4  S NO2 K iNO2

   

(4)

As weak acids and bases, FNA and FA exist in an ionization equilibrium in aqueous solutions, which is determined by the pH relationship. Therefore, Equation 4 could further be expanded as follows: SNNR 1   pH  pH SNNRmax  10 K mNH 4 K b  K mNH 4 K w 10 S NH 4 K b 1    pH  pH  10 S NH 4 K b 10 S NH 4 K b  K w K iNH 4 

 10  pH K mNO 2 10  pH  S NO 2 K a 1    10  pH  S NO K a 10  pH K iNO 2 2 

   

(5) where: SNH4 is the total ammonia concentration, including ionized ammonia and free un-ionized ammonia (mg N L-1), SNO2 is the total nitrite concentration, including ionized nitrite and free un-ionized nitrous acid (mg N L-1), Ka is the dissociation constant of nitrite; Kb is the dissociation constant of ammonia, and Kw is the ionization product constant of water.

12

The unknown parameters were estimated in a similar way to the Michaelis model. The values of the model parameters in the examined pH models were summarized in Table 2.

3. RESULTS AND DISCUSSION 3.1 Effect of DO on nitrogen removal Since Tests 1.1-1.7 were started with fluctuant nitrogen concentrations, the NH4+-N and NO2--N concentration profiles for specific initial DO were normalized. (Figure 1a–b). The highest SNRR (=1.1 gN gVSS-1 d-1) was obtained under non-aerated conditions (DO = 0.03 mg O2 L-1), which resulted in the most efficient ammonium and nitrite removal (81.6%) during the 2-h measurement (Figure 1c). Under aerobic conditions, nitrogen removal was explicitly inhibited, but the actual level of inhibition by the specific DO concentrations was variable. When the DO concentration was set between 0.3-0.4 and 0.6-1.0 mg O2 L-1, similar trends in the effluent ammonium and nitrite concentrations were observed resulting in low SNRRs (<0.5gN g-1VSS d-1) and low nitrogen removal efficiencies (<55%). On the other hand, the SNRRs increased to approximately 0.8gN gVSS-1 d-1 at the DO concentration of 0.5 mg O2 L-1, which resulted in the nitrogen removal efficiency of 80% after 2.5 h. This finding is in contrast with the directional DO suppression on nitrogen removal in the anammox process reported by Egli et al [4]. However, a similar behavior in terms of the nitrogen removal efficiency at the DO range of 0.4-1.0 mg O2 L-1 was observed by Vangsgaard et al [23] in nitritation/anammox biofilm (Figure 2). This suggests that nitritation was an

13

additional efficient nitrogen transformation pathway (besides anammox) in the studied reactor.

Vlaeminck et al [24] and Lotti et al [25] previously reported the stratification of granular or biofilm anammox systems. In the studied cases, AOB and NOB formed a protective barrier to create favorable conditions for anammox development in the inner layers of the granules or biofilms, where oxygen played a role as a co-substrate with nitrogenous compounds that greatly affected the metabolisms of both AOB and NOB. Under DO-limited conditions, AOB and NOB compete for oxygen to oxidize ammonium and nitrite, respectively. Moreover, these microorganisms share a symbiotic relationship since AOB produce the nitrite required as a substrate for NOB. This balance prevents the nitrite accumulation, which can be toxic to anammox bacteria [26]. Nevertheless, low DO concentrations favor the AOB growth over NOB, since NOB are characterized by a lower oxygen affinity [27].

The NO2--N/NH4+-N removal ratio remained stable (approximately 1.3–1.5) for the DO concentration below 0.6 mg O2 L-1 (Figure 1d), which was similar to the reported anammox reaction ratio of 1.32 [1]. At higher DO concentrations (0.8 and 1.0 mg O2 L-1), the NO2--N/NH4+-N removal ratio decreased to 0.8, suggesting that AOB were much more active compared to anammox bacteria. This resulted in the net nitrite production. Furthermore, at the DO concentration below 0.8 mg O2 L-1, the net nitrate consumption was observed which is consistent with the occurrence of endogenous denitrification as reported by Dapena-Mora

14

et al [17]. The net nitrate production was observed only when the reactor was switched from the non-aerated to the aerated phase (0.5 mg O2 L-1). At the DO concentration = 0.5 mg O2 L-1, the increase in the NO3--N production/NH4+-N removal ratio could be attributed to the DO inhibition of endogenous denitrification.

In order to further examine the nitrogen transformation pathways in the anammox-enriched granular sludge, Tests 2.1-2.2 and Tests 3.1-3.4 with only ammonium or nitrite addition, respectively, were carried out. Different behaviors were observed at the DO concentration of 0.3, 0.5 and 2.0 mg O2 L-1 (Figure 3), indicating a variance in the combined nitrogen removal processes in the anammox-enriched granular sludge. In the tests in which NH4+-N (initial concentration of about 80 mg N L-1) was dosed, when the DO concentration was set to 0.3 mg O2 L-1, 20.3% of the initial ammonium load was removed after 3 hours, while the N2O concentration remained at an undetected level (Figure 3a). At the DO concentration = 0.5 mg O2 L-1, a small amount of N2O was produced and then quickly consumed (Figure 3c), and the N2O production level was much lower compared to the nitrifying sludge [28]. This implicitly indicated the occurrence of nitritation and subsequent removal of the produced nitrite and ammonium by anammox bacteria in the granule interior. In the test with nitrite dosage (about 80 mg N L-1), explicit N2O peaks were detected under both examined DO conditions (Figure 3b and 3d), which may be associated with autotrophic denitrification, i.e., the reduction of NO2--N to N2O by AOB [28]. Unlike the N2O accumulation observed within 90 min of the DO concentration = 0.3 mg O2 L-1was very low (Table 3) despite the N2O production rate

15

was slightly lower compared to the DO concentration = 0.5 mg O2 L-1 and a prolonged N2O accumulation was observed (up to 180 min) (Figure 3d). This indicates a much more active reduction in the nitrite concentration and a quick release of N2O from the liquid phase to the gas phase due to the high aeration intensity. Besides the DO concentration, nitrite also played an important role in N2O production. In the case of a lower initial nitrite dosage (about 20 mg N L-1), less N2O accumulation (Figure 3e) and a lower specific N2O production rate was observed (Table 3), indicating the stimulating effect of nitrite on the AOB activity. On the contrary, N2O production in the reactor was hardly detectable at the DO concentration = 2.0 mg O2 L-1 (Figure 3f), indicating that autotrophic denitrification by AOB could easily be suppressed by intensive aeration, thus resulting in halting N2O production (Table 3). The observed N2O production, influenced by both DO and NO2--N concentrations, was in a similar pattern with enriched nitrifying sludge where the autotrophic denitrification pathway was predominant [28, 29]. This confirmed the importance of autotrophic denitrification in anammox-enriched granular sludge for nitrogen transformation under aerated conditions. Furthermore, in comparison with the nitrogen removal performance at the DO concentration = 0.3 mg O2 L-1, higher removal efficiencies of NH4+-N (28.5%) and NO2--N (23.6%) were obtained at the DO concentration = 0.5 mg O2 L-1 (Figure 3b and 3d), which are consistent with the findings of Test 1. It also indicated that nitrogen removal was obtained by various biochemical processes rather than exclusively anammox.

16

The findings obtained in the present study suggest that a complex combination of different nitrogen removal processes was occurring in the studied granular sludge. The actual contribution of each biochemical process was explicitly influenced by the DO concentration. Under non-aerated, anoxic with NO2--N conditions, anammox was the most active process in the anammox-enriched granular sludge, but endogenous denitrification could also be a possible pathway for nitrogen removal under these conditions [17]. At the DO concentration varying from 0.3 to 0.4 mg O2 L-1, the anammox activity was inhibited (Figure 1c), resulting in favorable conditions to activate the AOB and NOB metabolisms. For higher DO concentrations (0.5 to 0.6 mg O2 L-1), a good balance of nitrogen removal by endogenous denitrification and nitritation was observed when AOB outcompeted NOB for oxygen. This situation implies that nitrogen was removed by the combined activity of anammox bacteria, AOB, and denitrifiers, shown by the occurrence of a local SNRR peak (Figure 1c) and nitrogen conversion ratios similar to the reported anammox reaction ratios (Figure 1d). For the DO concentrations higher than 0.8 mg O2 L-1, endogenous denitrification began to be inhibited, whereas AOB are able to more actively use the available oxygen for ammonium oxidation. As a consequence, nitrite accumulation was observed which resulted in severe inhibition of the anammox metabolism.

3.2 Effect of pH on nitrogen removal Profiles of the normalized concentrations of NH4+-N and NO2--N for the specific initial pH values in Tests 4.1-4.10 are compared in Figure 4 (a-b). The trends were similar when the

17

initial pH was between 6.5 and 8.5. As shown in Figure 4c, the maximum SNRR (approximately 0.5 gN gVSS-1 d-1) was also reached at pH in the same range (6.5-8.5), resulting in high nitrogen removal efficiencies (61.8–78.1%). Thus, the optimal pH range in the present study was consistent with the optimal range (6.7–8.3) for anammox reported by Strous et al [11]. In addition, a decline in the SNRR occurred once the pH deviated from the optimum range and the nitrogen removal activity was hampered. Only minor nitrogen removal efficiency (<7%) was observed for the extreme initial pH values, i.e. 6.0 and 10.5.

It should be noted that the observed maximum SNRR occurred when FA and FNA concentrations were maintained at low levels (Figure 4c). A significant decline in the SNRR occurred when the initial pH dropped from 6.5 to 6.0, and when the FNA concentration increased from 0.02 to 0.08 mg N L-1. A similar behavior was observed when the initial pH raised from 8.5 to 10.5, and when the FA concentrations increased from 6.8 to 31.3 mg N L-1. In both cases, the nitrogen removal efficiencies deteriorated which confirmed the correlation between the high SNRRs and low FA or FNA concentrations. FA and FNA have been previously described as strong anammox inhibitors regardless of the pH in the medium [9, 30]. Jaroszynski et al [30] reported the inhibition of anammox at the FA concentrations as low as 1.7 mg L-1. In the continuous feeding test carried out by Fernandez et al [31], 0.0015 mgFNA L-1 induced a reduction in the nitrogen removal efficiency and disabled the studied biofilm system. Bacterial inhibition by FA has been attributed to its diffusion through the cell

18

membrane into the cytoplasm and the subsequent change in the intracellular pH [32], which could result in neutralization of the membrane potential leading ultimately to cell death.

On the contrary, recent studies have shown that FA is not an important inhibitory factor when its concentration is below 17 mg L-1 [9, 12]. Puyol et al [33] also suggested that in mildly alkaline solutions, the anammox process is primarily inhibited by the pH itself rather than by the FA concentration. Among the cellular organelles of anammox bacteria is the anammoxosome, a membrane-bound organelle where ammonium and nitrite are converted to N2 via a hydrazine intermediate. Between the anammoxosome and the riboplasm (equivalent to the cytoplasm in anammox bacteria), an intracellular proton gradient is responsible for energy production [2]. Therefore, extreme pHs may affect the proton transfer and alter metabolic processes which are dependent on pH gradients, such as energy generation by adenosine triphosphatases (ATP) and active transport of proteins (including NO2- transporters) [34]. Long-term pH alterations in the medium could disrupt the proton motive force, thus inhibiting the associated generation of energy required in certain enzymatic reactions which are strictly pH-sensitive [35]. In addition, exposure to the extreme pH medium may cause denaturation of cellular proteins, original bond cleavage, and destruction of other biological structures leading ultimately to enzyme inactivation.

To confirm the correlation between SNRR and FA/FNA ratio, a two-dimensional contour plot was constructed with experimental FA/FNA and SNRR data within the optimal pH range

19

(6.5-8.5) and low DO concentrations (<0.1mg O2/L) using the Origin software (version 8.5), as shown in Figure 5. The contour of dark grey observed in the figure was an indication of a region of high SNRRs, where the FA and FNA were both at low concentrations. Having an increase in either FA or FNA, the SNRR declined. Furthermore, the decline was sharper with a higher FNA concentration compared to the FA concentration. This indicates that anammox bacteria are more vulnerable to inhibition by FNA, making FNA an important contributor to the inhibition when feeding the bacteria with a high substrate with low pH. Therefore, a slightly alkaline influent was more appropriate for the anammox bacteria. This was consistent with the observation that efficient nitrogen removal remained stable even at a higher pH of 8.5 while a slightly lower pH of 6.5 would lead to 36.1% inhibition in the nitrogen removal activity (Figure 4c). Similar results were also reported in a previous study [12], where the process maintained a steady performance even when the effluent pH was in the range of 8.5–9.3. Due to the fact that pH is an important control parameter during the operation of anammox bioreactors, a control region was determined in this study. This region is in a pH range from 7.0 to 8.5, and FA and FNA concentrations below 7 mg N L-1 and 0.005 mg N L-1, respectively.

In order to further explain the inhibitory effects of pH and FA/FNA, two pH models (Michaelis and Haldane) were applied to describe the observed SNRRs in the studied reactor under the wide range of the examined pH (Figure 6). Considering the substrate inhibitory effects on the biomass, the Haldane model provided a better fit (R2 = 0.93) of the

20

experimental data in comparison with the Michaelis model (R2 = 0.84), which only considered pH effects. As the presence of FA and FNA is pH-dependent, the direct inhibitory effects of these parameters in different pHs are masked. Thus, further studies are needed to elucidate and quantify the actual contribution of these factors concerning the inhibition of the anammox process.

The observed nitrogen conversion ratios in terms of the initial pH are shown in Figure 4d. Under the optimal pH conditions (6.5–8.5), the nitrogen conversion ratios were similar to the reported anammox reaction ratio, i.e., 1:1.32:0.26 (NH4+-N:NO2--N:NO3--N), suggesting that under these conditions anammox is the primary process responsible for nitrogen conversions. At the pH values outside the optimal pH range, the nitrogen conversion ratios deviated from the reported reported anammox reaction ratio, resulting in fluctuations in the nitrogen transformation pathways. At the lowest initial pH (=6.0), the NO3--N production/NH4+-N removal ratio was approximately 1, which was much higher in comparison with the reported anammox reaction ratio of 0.26. This indicated that under low pH conditions, NOB were much more active in comparison with other groups of bacteria, resulting in a relatively high rate of nitrate production. At a high initial pH (=9.5), nitrate utilization by heterotrophic denitrifies was faster than nitrate production by anammox and/or NOB. In addition, the higher NO2--N removal/NH4+-N removal ratios at pH 6.5 and 9.5 indicated that anammox bacteria were more sensitive than AOB to pH shocks which is in accordance with the finding

21

of Sin et al [36]. As a result, when the pH began to deviate from the optimum range, AOB were more active and produced more nitrite.

3.3 Combined effects of DO and pH on nitrogen removal As the DO concentration and pH individually influenced the nitrogen removal efficiencies, two-dimensional contour plots were constructed to better visualize the relationship between SNRR and the experimental ranges of these two variables (Figure 7). In this study, the contour plots were characterized by two peak regions, which indicated a significant interaction of DO and pH with the observed SNRR. The highest SNRRs (> 0.9 gN gVSS-1 d-1) occurred within the pH range of 7–8 and at the lowest DO concentrations (0.0–0.1 mg O2 L-1). Moreover, another local SNRR maximum (0.9 gN gVSS-1 d-1) was observed at the pH and DO ranges of 7.5–8.5 and 0.5-0.6 mg O2 L-1, respectively. This suggested that not only anammox bacteria, but also other microbial groups (e.g., AOB, NOB and denitrifiers) were responsible for nitrogen removal in the anammox-enriched granular sludge. The cross-sectional view further confirmed that the combined nitrogen removal reached a relatively higher level at 0.5-0.6 mg O2 L-1 with the optimal pH between 7.5-8.0 (Figure 7).

According to the anammox reaction equation, H+ is consumed when nitrite is used as an electron acceptor to oxidize ammonium, resulting in an increase in the pH by approximately 0.5 (Figure 8a). Therefore, an extremely high pH should be avoided in anammox systems since the anammox bacteria have their own optimum pH habitat [11]. When the pH of the

22

reactor exceeds the optimum range (approximately 7.0–8.5) for anammox bacteria, the metabolic activity of these bacteria becomes suppressed. In addition, the DO concentration also temporarily changed, accompanied by the anammox process in the reactor. As shown in Figure 8b, the DO concentration increased significantly in the abiotic assay, from 0.5 to 4.5 mg O2 L-1. In that case, the DO profile can be represented as a reoxygenation curve due to oxygen dissolution from the air to the liquid phase by a mixing operation. In the bioassays, the increase in the DO concentrations was much smaller in comparison to the abiotic assay, implying the oxygen consumption by some microbial groups in the anammox-enriched granules. Furthermore, at a higher initial pH (8.5–10.5), the DO concentration was maintained at a low level (approximately 0.4–0.5 mg O2 L-1) during the test period, whereas at the lowest pH (= 6.0) the DO concentrations increased to approximately 2.7 mg O2 L-1. This provided further proof of the inhibition of oxygen-consuming bacteria (AOB or NOB) by low pH. An explicit DO elbow could be observed after approximately 150 min at the physiological pH range, when ammonium and nitrite initially present in the reactor were almost completely consumed. This was a clear evidence of the existence of oxygen-consuming bacteria in the anammox-enriched granular sludge.

4. CONCLUSIONS The highest SNRR (1.1 gN gVSS-1 d-1) in the anammox-enriched granular sludge was obtained under non-aerated conditions. Under aerated-conditions, the nitrification and endogenous denitrification processes were also active, resulting in relatively high SNRRs

23

(approximately 0.8 gN gVSS-1 d-1) at the DO concentration of 0.5 mg O2 L-1. At the pH range of 6.5–8.5, an optimal SNRR was achieved in the performed tests, ensured by the maintenance of low FA and FNA concentrations in the reactor. Thus, given the combined effects of DO and pH on the observed SNRR, appropriate ranges of both variables (DO concentration 0.0–0.1 mg O2 L-1 or 0.5-0.6 mg O2 L-1 = and pH=7.0–8.5) should be maintained to optimize the nitrogen removal efficiency by the combined biochemical processes in anammox-enriched granular sludge.

ACKNOWLEDGEMENTS This study was financially supported by the National Science Centre (Poland) under project no. UMO-2011/01/ B/ST8/07289. During this study, Zhixuan Yin and Carla Eloisa Diniz Dos Santos were visiting researchers at Gdansk University of Technology within the framework of the CARBALA Project (CARbon BALAncing for nutrient control in wastewater treatment), People Maria Curie Actions (FP7-PEOPLE-2011-IRSES).

24

REFERENCES [1] Strous M, Heijnen JJ, Kuenen JG, Jetten MSM. The sequencing batch reactor as a powerful tool for the study of slowly growing anaerobic ammonium-oxidizing microorganisms. Appl Microbiol Biotechnol 1998; 50(5): 589-596. [2] Van der Star WR, Dijkema C, de Waard P, Picioreanu C, Strous M, Van Loosdrecht, MCM. An intracellular pH gradient in the anammox bacterium Kuenenia stuttgartiensis as evaluated by 31P NMR. App Microbiol Biotechnol

2010; 86(1): 311-317.

[3] Jin RC, Yang GF, Yu JJ, Zheng P. The inhibition of the Anammox process: a review. Chem Eng J 2012; 197: 67-79. [4] Egli K, Fanger U, Alvarez PJ, Siegrist H, Van der Meer JR, Zehnder AJ. Enrichment and characterization of an anammox bacterium from a rotating biological contactor treating ammonium-rich leachate. Arch Microbiol 2001; 175(3): 198-207. [5] Strous M, Van Gerven E, Kuenen JG, Jetten M. Effects of aerobic and microaerobic conditions on anaerobic ammonium-oxidizing (anammox) sludge. Appl Envir Microbiol 1997; 63(6): 2446-2448. [6] Third KA, Paxman J, Schmid M, Strous M, Jetten MSM, Cord-Ruwisch R. Enrichment of anammox from activated sludge and its application in the CANON process. Microb Ecol 2005; 49(2): 236-244. [7] Cema G, Paza E, Trela J, Surmacz-Górska J. Dissolved oxygen as a factor influencing nitrogen removal rates in a one-stage system with partial nitritation and Anammox process. Water Sci Technol 2011; 64(5): 1009-1015.

25

[8] Chen H, Liu S, Yang F, Xue Y, Wang T. The development of simultaneous partial nitrification, ANAMMOX and denitrification (SNAD) process in a single reactor for nitrogen removal. Bioresour Technol 2009; 100(4): 1548-1554. [9] Wang G, Xu X, Gong Z, Gao F, Yang F, Zhang H. Study of simultaneous partial nitrification, ANAMMOX and denitrification (SNAD) process in an intermittent aeration membrane bioreactor. Proc Biochem 2016; 51: 632-641.

[10] Fernández I, Dosta J, Fajardo C, Campos JL, Mosquera-Corral A, Méndez R. Short-and long-term effects of ammonium and nitrite on the Anammox process. J Environ Manag 2012; 95: S170-S174.

[11] Strous M, Kuenen JG, Jetten MS. Key physiology of anaerobic ammonium oxidation. Appl Envir Microbiol 1999; 65(7): 3248-3250. [12] Tang CJ, Zheng P, Hu BL, Chen JW, Wang CH. Influence of substrates on nitrogen removal performance and microbiology of anaerobic ammonium oxidation by operating two UASB reactors fed with different substrate levels. J Hazard Mat 2010; 181(1): 19-26. [13] Ahn Y, Hwang I, Min K. ANAMMOX and partial denitritation in anaerobic nitrogen removal from piggery waste. Water Sci Technol 2004; 49(5-6): 145-153. [14] Jaroszynski LW, Cicek N, Sparling R, Oleszkiewicz JA. Importance of the operating pH in maintaining the stability of anoxic ammonium oxidation (anammox) activity in moving bed biofilm reactors. Bioresour Technol 2011; 102(14): 7051-7056.

26

[15] Fux C, Marchesi V, Brunner I, Siegrist H. Anaerobic ammonium oxidation of ammonium-rich waste streams in fixed-bed reactors. Water Sci Technol 2004; 49(11-12): 77-82. [16] Sobotka D, Czerwionka K, Makinia J. The effects of different aeration modes on ammonia removal from sludge digester liquors in the nitritation–anammox process. Water Sci Technol 2015; 71(7): 986-995. [17] Dapena-Mora A, Van Hulle S W H, Campos J L, Méndez R, Vanrolleghem P A, Jetten M. Enrichment of anammox biomass from municipal activated sludge: experimental and modeling results. J Chem Technol Biotechnol 2004; 79: 1421-1428. [18] Sobotka D, Tuszynska A, Kowal P, Ciesielski S, Czerwionka K, Makinia J. Long-term performance and microbial characteristics of the anammox-enriched granular sludge cultivated in a bench-scale sequencing batch reactor (in preparation). [19] Standard Methods for the Examination of Water and Wastewater, 21st ed., 2005 APHA, AWWA, WEF, Washington, DC, USA. [20] Anthonisen AC, Loehr RC, Prakasam TBS, Srinath EG. Inhibition of nitrification by ammonia and nitrous acid. J Water Pollut Control Fed 1976; 48(5): 835-852. [21] Segel IH. Enzyme kinetics: behavior and analysis of rapid equilibrium and steady-state enzyme systems. New York: John Wiley and Sons (360); 1975. [22] Chen T, Zheng P, Shen L, Ding S, Mahmood Q. Kinetic characteristics and microbial community of Anammox-EGSB reactor. J Hazard Mat 2011; 190(1): 28-35.

27

[23] Vangsgaard AK, Mauricio-Iglesias M, Gernaey KV, Smets BF, Sin G. Sensitivity analysis of autotrophic N removal by a granule based bioreactor: Influence of mass transfer versus microbial kinetics. Bioresour Technol 2012; 123: 230-241. [24] Vlaeminck SE, Terada A, Smets BF, De Clippeleir H, Schaubroeck T, Bolca S, Verstraete W. Aggregate size and architecture determine microbial activity balance for one-stage partial nitritation and anammox. Appl Envir Microbiol 2010; 76(3): 900-909. [25] Lotti T, Kleerebezem R, Hu Z, Kartal B, de Kreuk MK, Van ErpTaalman KC, Van Loosdrecht MCM. Pilot-scale evaluation of anammox-based mainstream nitrogen removal from municipal wastewater. Environ Technol 2015: 36(9): 1167-1177. [26] Mutlu A. G. Management of microbial community composition, architecture and performance in autotrophic nitrogen removing bioreactors through aeration regimes. Denmark: Department of Environmental Engineering, Technical University of Denmark; PhD thesis; 2015.p. 81. [27] Blackburne R, Yuan Z, Keller J. Partial nitrification to nitrite using low dissolved oxygen concentration as the main selection factor. Biodegrad 2008; 19(2): 303-312. [28] Kampschreur MJ, Van der Star WR, Wielders HA, Mulder JW, Jetten MS, Van Loosdrecht MCM. Dynamics of nitric oxide and nitrous oxide emission during full-scale reject water treatment. Water Res 2008; 42(3): 812-826. [29] Castro-Barros CM, Daelman MRJ, Mampaey KE, Van Loosdrecht MCM, Volcke EIP. Effect of aeration regime on N2O emission from partial nitritation-anammox in a full-scale granular sludge reactor. Water Res 2015; 68: 793-803.

28

[30] Jaroszynski LW, Cicek N, Sparling R, Oleszkiewicz JA. Impact of free ammonia on anammox rates (anoxic ammonium oxidation) in a moving bed biofilm reactor. Chemosphere 2012: 88(2): 188-195. [31] Fernández I, Mosquera-Corral A, Campos JL, Mendez R. Operation of an Anammox SBR in the presence of two broad-spectrum antibiotics. Proc Biochem, 2009; 44(4): 494-498. [32] Wittmann C, Zeng AP, Deckwer WD. Growth inhibition by ammonia and use of a pH-controlled

feeding

strategy

for

the

effective

cultivation

of

Mycobacterium

chlorophenolicum. App Microbiol Biotechnol 1995; 44(3-4): 519-525. [33] Puyol D, Carvajal-Arroyo JM, Sierra-Alvarez R, Field JÁ. Nitrite (not free nitrous acid) is the main inhibitor of the anammox process at common pH conditions. Biotechnol Lett 2014; 36(3): 547-551. [34] Lu HF, Ji QX, Ding S, Zheng P. The morphological and settling properties of ANAMMOX granular sludge in high-rate reactors. Bioresour Technol 2013; 143: 592-597. [35] Carvajal-Arroyo JM, Puyol D, Li G, Sierra‐Álvarez R, Field JÁ. The role of pH on the resistance of resting - and active anammox bacteria to NO2−- inhibition. Biotechnol Bioeng 2014; 111(10): 1949-1956. [36] Sin G, Kaelin D, Kampschreur MJ, Takács I, Wett B, Gernaey KV, Rieger L, Siegrist N, van Loosdrecht MCM. Modelling nitrite in wastewater treatment systems: a discussion of different modelling concepts. Water Sci Technol 2008; 58(6): 1155-1171.

29

Figures caption: Figure 1. Profiles of normalized concentrations of NH4+-N (a) and NO2--N (b); SNRRs (c) and nitrogen conversion ratios in anammox-enriched granular sludge at different DO concentrations (d). Figure 2. Comparison of nitrogen removal in anammox-enriched granular sludge (this study) and biofilm (Vangsgaard et al., 2012) at different DO concentrations. Figure 3. Behavior of nitrogen compounds in the studied reactor at the DO concentrations of 0.3 (a-b), 0.5 (c-e) and 2.0 (f) mgO2 L-1. Figure 4. Profiles of normalized concentrations of NH4+-N (a) and NO2--N (b); specific nitrogen removal rates (SNRRs) (c); nitrogen conversion ratios in anammox-enriched granular sludge under different initial pH conditions (d). Figure 5. The co-inhibition area of FA and FNA on the specific nitrogen removal rates (SNRRs). Figure 6. Effects of pH on the SNRR/SNRRmax ratio. Fitting of the experimental data by the Michaelis model (R2 = 0.84) and Handale model (R2 = 0.93). Figure 7. The interactive effects of pH and DO on the specific nitrogen removal rates (SNRRs). Figure 8. Profiles of pH (a) and DO concentrations (b) in the studied reactors under different initial pH conditions.

30

Table 1. Initial conditions of the performed experiments in terms of biomass, substrates, DO and pH.

Tests No.

Aim of investigation

Biomass

NH4+

NO2-

DO

(gVSS L-1)

(mg N L-1)

(mg N L-1)

(mgO2 L-1)

pH

Volume(L)

0.03-1.0

7.7 ±0.1

10.0

-

0.3 and 0.5

7.7 ±0.1

10.0

80.5±0.9

0.3 and 0.5

7.7 ±0.1

10.0

22.3±1.2

0.5 and 2.0

7.7 ±0.1

10.0

Effects of DO 1.1-1.7

concentration on nitrogen

3.0±0.5

103.9±10.3 127.0±17.4

removal NH4+ utilization pathways 2.1-2.2

under different DO

3.0±0.5

87.3±2.0

concentrations NO2- utilization pathways 3.1-3.4

under different DO

3.0±0.5

-

concentrations Effects of initial pH on 4.1-4.10

nitrogen removal at

1.3±0.3

33.1±2.4

43.0±1.5

0.5±0.1a

6.0-10.5b

4.0

1.3±0.3

39.1±2.7

50.7±3.9

0.03 a

6.0-10.5b

2.2

-1

DO=0.5 mgO2 L

Effects of initial pH on 5.1-5.4

nitrogen removal at -1

DO=0.03 mgO2 L

a. initial DO concentration;

b. initial pH.

31

Table 2. Values of the model parameters in the examined pH models. Michaelis model

Haldane model

Parameters

Unit

Values

Parameters

Unit

Values

pKS1

-

6.47

Ka

-

4.6×10-4

pKS2

-

9.41

Kb

-

1.57×10-5

Kw

-

10-14

KmNH4

mg/L

48.4

KiNH4

mg/L

1123.2

KmNO2

mg/L

6.6

KiNO2

mg/L

720.6

32

Table 3. Biomass specific N2O production rates and N2O emission factors under various DO and NO2- concentrations. DO

NO2-

(mgO2

(mg N

Specific N2O production rate

N2O emission factor

L-1)

L-1)

(mg N g-1VSS h-1)

(%)

Tests 3.1

0.3

81.1

0.73

13.7

Tests 3.2

0.5

79.8

0.61

4.9

Tests 3.3

0.5

23.2

0.19

8.0

Tests 3.4

2.0

21.5

0.00

1.4

Tests No.

33