Applied Microbiology and Biotechnology

, Volume 90, Issue 4, pp 1537–1545 | Cite as

OLAND is feasible to treat sewage-like nitrogen concentrations at low hydraulic residence times

  • Haydée De Clippeleir
  • Xungang Yan
  • Willy Verstraete
  • Siegfried Elias Vlaeminck
Environmental Biotechnology

Abstract

Energy-positive sewage treatment can, in principle, be obtained by maximizing energy recovery from concentrated organics and by minimizing energy consumption for concentration and residual nitrogen removal in the main stream. To test the feasibility of the latter, sewage-like nitrogen influent concentrations were treated with oxygen-limited autotrophic nitrification/denitrification (OLAND) in a lab-scale rotating biological contactor at 25°C. At influent ammonium concentrations of 66 and 29 mg N L−1 and a volumetric loading rate of 840 mg N L−1 day−1 yielding hydraulic residence times (HRT) of 2.0 and 1.0 h, respectively, relatively high nitrogen removal rates of 444 and 383 mg N L−1 day−1 were obtained, respectively. At low nitrogen levels, adapted nitritation and anammox communities were established. The decrease in nitrogen removal was due to decreased anammox and increased nitratation, with Nitrospira representing 6% of the biofilm. The latter likely occurred given the absence of dissolved oxygen (DO) control, since decreasing the DO concentration from 1.4 to 1.2 mg O2 L−1 decreased nitratation by 35% and increased anammox by 32%. Provided a sufficient suppression of nitratation, this study showed the feasibility of OLAND to treat low nitrogen levels at low HRT, a prerequisite to energy-positive sewage treatment.

Keywords

Ammonia-oxidizing bacteria Domestic wastewater Nitrite-oxidizing bacteria Nutrient Sustainable 

Introduction

Biological nitrogen removal is economically preferred above physicochemical nitrogen recovery for wastewaters containing less than 5 g N L−1 (Mulder 2003). Furthermore, if the ratio of biodegradable chemical oxygen demand (bCOD) to nitrogen is relatively low (typically ≤2–3), nitrogen removal with partial nitritation and anammox saves about 60% of the aeration, 90% of the sludge handling and transport, and 100% of the organic carbon addition compared to conventional nitrification/denitrification (Mulder 2003). Overall, some 30–40% of the overall nitrogen removal costs can be saved (Fux and Siegrist 2004). Oxygen-limited autotrophic nitrification/denitrification (OLAND) is a one-stage configuration of this process (Kuai and Verstraete 1998), in which aerobic ammonium-oxidizing bacteria (AerAOB) oxidize about half of the ammonium to nitrite in the outer, aerobic zones of the biomass (partial nitritation), while the anoxic ammonium-oxidizing bacteria (AnAOB) subsequently convert nitrite and the residual ammonium to mainly nitrogen gas (89%) and some nitrate (11%) in the inner, anoxic zones (anammox; Pynaert et al. 2003; Vlaeminck et al. 2010). Oxygen plays a key role in balancing the microbial activities (Fig. 1a), with on the one hand an oxygen requirement of 1.8 g O2 g−1 N to achieve sufficient ammonium oxidation while avoiding excess nitrite production by AerAOB. On the other hand, sufficiently low dissolved oxygen (DO) levels (e.g. 0.3 mg O2 L−1) are needed to suppress excess nitrate production by nitrite-oxidizing bacteria (NOB) (Joss et al. 2009).
Fig. 1

a Conversion of nitrogen species, oxygen, and protons in OLAND, showing balanced and imbalanced contributions of three bacterial groups, i.e., aerobic ammonium-oxidizing, nitrite-oxidizing, and anoxic ammonium-oxidizing bacteria (AerAOB, NOB, AnAOB, respectively); b Conventional and redesigned sewage treatment schemes with OLAND in the side and main line, respectively, simplified and redrafted after Verstraete and Vlaeminck (2011). In the redesigned scheme, energy-positive sewage treatment can, in principle, be obtained by maximizing energy recovery through anaerobic digestion of concentrated organics in the side stream and by minimizing energy consumption for the physicochemical and/or biological concentration step and the residual nitrogen removal step, applying OLAND

Conventional activated sludge (CAS) systems for sewage treatment have low volumetric carbon and nitrogen loading rates (around 1 g COD L−1 day−1 and 0.08 g N L−1 day−1) and are energy-negative. The aeration required for organic carbon and nitrogen removal constitutes about 60–70% of the total energy consumption of a sewage treatment plant (Zessner et al. 2010). However, if enhanced primary settling is applied to increase physicochemical sludge production and if OLAND is used for nitrogen removal from the digestate of primary and secondary sludge (Fig. 1b), the aeration requirements of the CAS step can be decreased with 25% (Siegrist et al. 2008). Over the last 5 years, several OLAND-type treatments were developed to treat sewage sludge digestates (Joss et al. 2009, Jeanningros et al. 2010). Furthermore, if primary settling is replaced by a highly loaded activated sludge step, where organic matter is converted to biomass at maximal yield, energy neutrality is achievable given the even higher conversion of bCOD by anaerobic digestion into biogas and hence electricity (Wett et al. 2007). Given the high energetic content of the sewage bCOD, energy-positive sewage treatment should be possible (Verstraete et al. 2009; Kartal et al. 2010; Verstraete and Vlaeminck 2011). This requires an advanced biological or physicochemical bCOD concentration step to further increase energy recovery from anaerobic digestion of concentrated organics and a low energy demand for the concentration step and the removal of residual nitrogen (and some bCOD) in the main stream (Fig. 1b). The energy requirement for OLAND is influenced by the reactor configuration: active aeration in sequencing batch reactors requires 1.2 kWh kg−1 N (Wett et al. 2010), whereas passive aeration in rotating biological contactors (RBC) requires down to 0.4 kWh kg−1 N (Mathure and Patwardhan 2005). Depending on the dilution, sewage is typically composed of 30–100 mg N L−1 and 450–1,200 mg COD L−1 rendering a COD/N ratio of about 12 to 15 (Tchobanoglous et al. 2003; Henze et al. 2008). An advanced concentration step is expected to separate up to 75–80% of the COD (Verstraete et al. 2009; Verstraete and Vlaeminck 2011) and about 20% of the sewage nitrogen, mainly consisting of colloidal and particulate organic nitrogen, from which the anaerobically hydrolyzed part is returned to the main stream as ammonium (Fig. 1b). Hence, the OLAND stage would receive nitrogen as ammonium at a COD/N ratio below 4, which is theoretically low enough to avoid the risk that heterotrophs overgrow AnAOB (Lackner et al. 2008).

Until now, the OLAND process has been applied for medium and high-strength nitrogen wastewaters (>0.2 g N L−1) such as landfill leachate and digestates from sewage sludge, specific industrial streams, and concentrated black water at relatively high hydraulic residence times (HRT, Table 1). To obtain reasonably high nitrogen removal rates (400 mg N L−1 day−1), the treatment of low nitrogen levels (<80 mg N L−1) has to occur at low HRT, in the order of some hours, rendering biomass retention an important requirement. In this study, a first and exploratory step towards the implementation of OLAND in the new sewage treatment scheme was tested, feeding sewage-like ammonium influent concentrations without COD addition. Given its low energy consumption for aeration, a RBC was chosen as lab-scale reactor and operated at 25°C, simulating the maximum sewage temperatures in summer (Breda, the Netherlands; Mollen, personal communication). This is one of the first tests on the OLAND treatment of low nitrogen concentrations at such low HRT, a prerequisite to energy-positive sewage treatment.
Table 1

Overview of typical average nitrogen concentrations, volumetric loading/removal rates, and HRT for existing one-step partial nitritation/anammox processes and for the low nitrogen concentration and HRT application in this study

Wastewater

Influent concentration (mg NH4+–N L−1)

N loading rate (g N L−1 day−1)

N removal rate (g N L−1 day−1)

HRT (day)

Reactor type

Reference

Digested black water

1023

0.94

0.71

1.33

RBC

Vlaeminck et al. (2009)

Sewage sludge digestate

800

0.74

0.67

0.93

SBR

Jeanningros et al. (2010)

Sewage sludge digestate

650

0.54

0.51

1.20

SBR

Joss et al. (2009)

Industrial digestate

300

2.0

1.17

0.18

Gas-lift

Abma et al. (2010)

Landfill leachate

209

0.38

0.38

0.55

RBC

Hippen et al. (1997)

Landfill leachate

250

0.67

0.41

0.51

RBC

Siegrist et al. (1998)

Sewage-like nitrogen concentrations

66

0.86

0.44

0.08

RBC

This study

Sewage-like nitrogen concentrations

31

0.84

0.38

0.04

RBC

This study

RBC rotating biological contactor, SBR sequencing batch reactor

Materials and methods

OLAND rotating biological contactor

The lab-scale RBC was based on an airwasher LW14 (Venta, Weingarten, Germany) with a rotor consisting of 40 disks interspaced at 3 mm, resulting in a disk contact surface of 1.32 m2. The reactor had a liquid volume of 3.6 L, immersing the disks for 64%. The reactor temperature was set at 25°C, and the pH was adjusted to be higher than 7.3 by the addition of NaHCO3. The DO concentration was not directly controlled. For continuous rotation, the rotation speed was fixed at 3 rpm, and in the intermittent rotation mode, rotation at the same rotation speed occurred only one third of the time, equally spread over time (1 min on, 2 min off).

Reactor operation

The influent of an OLAND lab-scale RBC, as used by Vlaeminck et al. (2009) to treat digested black water (Table 1), was switched to synthetic wastewater consisting of (NH4)2SO4, NaHCO3, KH2PO4 (10 mg P/L), and 2 mL L−1 of a trace element solution (Kuai and Verstraete 1998). After a long-term stable operation of the reactor treating 537 mg N L−1, the influent ammonium concentration was stepwise decreased to 278, 146, 66, and 31 mg N L−1 over 41, 48, 52, and 60 days, respectively, maintaining a constant loading rate (about 840 mg N L−1 day−1) by a stepwise decrease in HRT (Table 2). Each nitrogen influent concentration was applied for 1.5 to 2 months to obtain enough data points and stabilization for statistical comparison between the phases. Reactor pH, DO, and temperature were daily monitored, and influent and effluent samples were taken at least thrice a week for ammonium, nitrite, and nitrate analyses.
Table 2

OLAND rotating biological contactor conditions and performance (average ± standard deviation) over the periods with stepwise decreases of the ammonium influent concentration and HRT. In periods I–Va, rotation was continuous, whereas this was intermittent in period Vb. For the eight bottom rows, statistical analyses were performed, and the phases that were not significantly different (p > 0.05) are indicated with the number of the similar phase

Period

I

II

III

IV

Va

Vb

Duration (day)

21

41

48

52

31

29

Number of samples (−)

14

18

29

36

23

12

Influent NH4+ level (mg N L−1)

537 ± 13

278 ± 11

146 ± 21

66 ± 5

29 ± 8

31 ± 1

Influent flow rate (L day−1)

5.4 ± 0.2

10.5 ± 0.3

20.5 ± 1.5

42.9 ± 2.3

82.6 ± 2.0

83.6 ± 0.7

HRT (h)

16.0 ± 0.5

8.3 ± 0.3

4.2 ± 0.4

2.0 ± 0.1

1.0 ± 0.0

1.0 ± 0.0

N loading rate (mg N L−1 day−1)

819 ± 30

840 ± 49

832 ± 68

855 ± 56

851 ± 66

840 ± 20

DO level (mg O2 L−1)

1.4 ± 0.2III,Va

1.2 ± 0.2IV,Vb

1.4 ± 0.2I,Va

1.2 ± 0.1II,Va

1.4 ± 0.4I,III

1.2 ± 0.1II,IV

pH (−)

7.6 ± 0.1

7.5 ± 0.1

7.3 ± 0.2IV,Va,Vb

7.4 ± 0.1III

7.3 ± 0.2III,Vb

7.3 ± 0.0III,Va

Free ammonia (mg N L−1)i

0.91 ± 1.58II,III

0.40 ± 0.15III,I

0.40 ± 0.17I,II

0.10 ± 0.03

0.04 ± 0.02Vb

0.04 ± 0.01Va

N removal rate (mg N L−1 day−1)

642 ± 72

565 ± 42

471 ± 88IV

444 ± 84III,Vb

303 ± 75

383 ± 52IV

N removal efficiency (%)

79 ± 9

67 ± 3

58 ± 9

51 ± 8Vb

35 ± 7

46 ± 6IV

NH4+ removal efficiency (%)

94 ± 10

91 ± 3IV,Va,Vb

72 ± 26

89 ± 4II,Va,Vb

77 ± 31II,IV,Vb

91 ± 5II,IV,Va

NO3 prod./NH4+ cons. (%)ii

12 ± 2

22 ± 2IV

18 ± 8IV

21 ± 6II,III

45 ± 11

32 ± 6

Effluent NH4+ (mg N L−1)

19 ± 10II

25 ± 10I,III

29 ± 12II

7.4 ± 2.7

3.7 ± 1.5Vb

3.0 ± 1.0Va

Effluent NO2 (mg N L−1)

19 ± 5III

10 ± 12

16 ± 6IV,I

14 ± 3III

5.2 ± 1.3Vb

5.1 ± 0.4Va

Effluent NO3 (mg N L−1)

65 ± 4

59 ± 5

21 ± 9

14 ± 2Va

15 ± 2IV

11 ± 0.9

DO dissolved oxygen level, prod production, cons consumption

iCalculated from the measured ammonium level, temperature and pH (Anthonisen et al. 1976)

iiValues exceeding 11% indicate excess nitrate production by NOB

Chemical analyses

Ammonium (Nessler method) was determined according to standard methods (Greenberg et al. 1992). Nitrite and nitrate were determined on a 761 compact ion chromatograph equipped with a conductivity detector (Metrohm, Zofingen, Switzerland). DO and pH were measured with, respectively, an electrode installed on a C833 meter (Consort, Turnhout, Belgium) and an HQ30d DO meter (Hach Lange, Düsseldorf, Germany).

Fluorescent in-situ hybridization

At the start (537 mg N L−1) and at the end (29 mg N L−1) of experiment biomass samples were taken from the disks and bottom of the reactor for identification of the autotrophic nitrogen removing species present. At both time points fluorescent in-situ hybridization (FISH) quantification of AerAOB, AnAOB, and NOB was performed. A paraformaldehyde (4%) solution was used for biofilm fixation, and FISH was performed according to Amann et al. (1990). Relevant target groups were gathered from a recent nitrogen cycle review (Vlaeminck et al. 2011), and probe sequences and formamide concentrations were applied according to Lücker (2010) for Nitrotoga and probeBase for the other targets (Loy et al. 2003): Amx820 for the AnAOB Kuenenia/Brocadia, a mixture of NSO1225 and NSO190 for the β-proteobacterial AerAOB Nitrosomonas/Nitrosospira, and NIT3 (+ competitor), Ntspa662 (+ competitor), and Ntoga221 for the NOB Nitrobacter, Nitrospira, and Nitrotoga, respectively. The AnAOB, AerAOB, and NOB abundance was evaluated by combining the specific probe with an equimolar mixture of EUB338I, II, and III, targeting all bacteria, and 4′–6-diamidino-2-phenylindole, targeting all DNA-containing cells. Image acquisition was done on a Zeiss Axioskop 2 Plus epifluorescence microscope (Carl Zeiss, Germany). For quantification, 20 randomly taken images were analyzed with ImageJ software, and the percentage of the specific group was calculated as the ratio of the specific area to the total DNA-containing area. The EUB338 signals served as a control.

Denaturing gradient gel electrophoresis

At the start and at the end of the experiment, biomass was harvested to compare the community structure (AerAOB, Planctomycetes, and total bacteria) while treating high (537 mg N L−1) and low (29 mg N L−1) nitrogen concentrations, respectively. DNA extraction, nested PCR, and denaturing gradient gel electrophoresis (DGGE) were performed according to Pynaert et al. (2003), based on the primers CTO189ABf, CTO189Cf, and CTO653r for β-proteobacterial AerAOB; PLA40f and P518r for Planctomycetes, the bacterial phylum harboring AnAOB; and GC338 and 518r for all bacteria. The obtained DGGE patterns were subsequently processed with BioNumerics software (Applied Maths, Sint-Martens-Latem, Belgium), and similarities were calculated as the Pearson correlation coefficient.

Results

Treatment of high nitrogen levels

Following the influent shift from digested black water (Vlaeminck et al. 2009) to synthetic wastewater, the OLAND RBC was operated for 96 days at an influent concentration of 537 mg N L−1. Over the last 21 days of this period, the nitrogen removal rate was 642 mg N L−1 day−1, and the nitrogen removal efficiency was 79% (Table 2). The contributions of the different nitrogen pathways were quantified (Fig. 2), using the measured dissolved nitrogen species and assuming that (i) negligible denitrification occurred given the absence of bCOD in the influent, (ii) nitrogen gas was the product of the removed dissolved species, and (iii) AnAOB produced 0.11 g NO3–N per gram NH4+–N converted to nitrogen gas. Initially, OLAND converted 90% of the influent nitrogen, and nitrite and nitrate accumulation were negligible at high nitrogen levels (Fig. 2). Indeed, no NOB could be detected in the biofilm with FISH. The average DO and free ammonia levels are relatively high at 1.4 mg O2 L−1 and 0.9 mg N L−1, respectively (Table 2). The AnAOB and AerAOB communities made up 5% and 23% of the biofilm, respectively (Table 3), and were composed of several species (Fig. 3).
Fig. 2

Contributions of microbial conversions to reactor nitrogen products: nitrogen gas production (nitrogen in–nitrogen out) by anoxic ammonium-oxidizing bacteria (AnAOB) from influent ammonium (upper part) and from nitrite produced by aerobic ammonium-oxidizing bacteria (AerAOB, lower part), nitrate production by AnAOB (11% of ammonium converted by balanced OLAND), nitrate production by nitrite-oxidizing bacteria (NOB: nitrate out–nitrate in–nitrate production by AnAOB), excess nitrite production by AerAOB (nitrite out–nitrite in), and residual ammonium (ammonium in–ammonium out)

Table 3

Abundances of AerAOB, AnAOB and NOB in OLAND biomass, as determined from quantitative FISH. The NOB genera Nitrobacter and Nitrotoga could not be retrieved

Influent

537 mg N L−1

29 mg N L−1

29 mg N L−1

Biomass sample

Biofilm

Biofilm

Settled

AerAOB Nitrosomonas/Nitrosospira (%)

23 ± 18

22 ± 12

30 ± 16

AnAOB Kuenenia/Brocadia (%)

5 ± 5

7 ± 4

8 ± 6

NOB Nitrospira (%)

ND

6 ± 5

5 ± 5

ND not detected

Fig. 3

DGGE gels for β-proteobacterial aerobic ammonium-oxidizing bacteria (AerAOB) and Planctomycetes (Plancto), the phylum harboring anoxic ammonium-oxidizing bacteria (AnAOB). Biomass samples were taken at the end of the treatment period at 537 mg N L−1 (high N; sample from the disk biofilm) and at 31 mg N L−1 (low N; sample from the disk biofilm and from settled biomass). Similarities were calculated as the Pearson correlation coefficient, and plus symbols highlight the three AerAOB and Planctomycetes bands with the highest intensity, indicating shifts of the most dominant species

Treatment of low nitrogen levels

The nitrogen influent concentration and HRT were gradually decreased over the phases II–Va while maintaining a constant nitrogen loading rate. Over these changes, the pH and DO levels remained in the same range, but the total nitrogen removal rate and hence the efficiency decreased significantly (p < 0.05; Table 2). The proportion of ammonium oxidized remained relatively stable, but nitrite and nitrate accumulated in the effluent (Table 2, Fig. 2), indicating that decreased anammox and increased nitratation were responsible for the decreased efficiency. Excess nitrate production by NOB significantly increased from period I to period II, remained constant for periods III and IV, and was followed by a significant increase in period Va (Table 2, Fig. 2). The NOB could be identified as Nitrospira and composed 6% of the biofilm community at the lowest nitrogen concentration (Table 3). The free ammonia concentration decreased sharply over time, whereas DO levels were stable but relatively high (Table 2).

Suppression of nitratation at low nitrogen levels

In an attempt to decrease the DO level in phase Vb in order to decrease nitratation and increase the total nitrogen removal efficiency, discontinuous rotation was introduced. This resulted in a significant decrease of the oxygen concentration from 1.4 to 1.2 mg O2 L−1. Consequently, nitratation decreased with 35% and anammox increased with 32%, restoring the total nitrogen removal efficiency which was previously obtained in phase IV (Fig. 2, Table 2). The lower DO resulted in a total nitrogen effluent concentration of 20 mg N L−1 and a nitrogen removal rate of 383 mg N L−1 day−1, removing 46% of the influent nitrogen (Table 2). The used RBC set-up did not allow to further decrease of the DO levels since rotating more intermittently resulted in a strong decrease of the nitritation rate (data not shown). The relatively stable nitritation indicated few or no influence by the decreasing HRT (Fig. 2, Table 2). The decreasing AnAOB activity at lower HRT could be partly counteracted by a lower DO level and resulting lower nitratation in period Vb (Fig. 2), indicating the influence of DO level on NOB/AnAOB competition rather than a negative effect of low HRT on anammox.

While treating 29 mg N L−1, the AnAOB and AerAOB abundances in the biomass were 8% and 25%, respectively (Table 3), which was quite comparable to the abundance while treating high nitrogen concentrations. In the final operation period, part of the biomass was found at the bottom of the reactor, but neither FISH nor DGGE could detect important differences in the microbial composition of settled biomass versus biofilm on the disks (Table 3, Fig. 3), indicating that the biomass was probably the result of detachment of the biofilm from the disks. For the AnAOB communities, the DGGE profiles showed only small changes in the abundant species (88% similarity), while the AerAOB patterns changed more (23% similarity) (Fig. 3). The shift in the most abundant AerAOB could also be observed in the DGGE profiles of all bacteria.

Discussion

OLAND removal rate and efficiency treating low nitrogen levels

In this study, operation of the OLAND RBC on sewage-like nitrogen concentrations (66 and 29 mg N L−1) at low HRT (2 and 1 h) resulted in nitrogen removal rates of 383–444 mg N L−1 day−1, which are reasonably high (Table 1). In the energy-positive sewage treatment scheme (Fig. 1b), 20–25% of the sewage COD remains in the main line following an advanced concentration step (Verstraete et al. 2009; Verstraete and Vlaeminck 2011) as well as about 80% of the sewage nitrogen, partly derived from returning the digestate to the main line. Assuming raw sewage compositions of 825 mg COD L−1 and 65 mg N L−1, the OLAND step would hence receive 165 mg COD L−1 and 52 mg N L−1. Given the overall required removal efficiencies of 50–60% of COD and 75% of the nitrogen according to European standards (European Commission 1991), the OLAND step should remove an additional 36 mg N L−1, and hence, the desired OLAND removal efficiency should be around 70%. In this study, nitrogen removal efficiencies during treatment of low nitrogen levels were 46–51% and hence not sufficiently high to comply with the required standards. The obtained nitrogen removal percentages were lower than previously reported for this type of reactors (Pynaert et al. 2003, 2004; Schmid et al. 2003), mainly due to additional nitratation. Also in absolute terms, the effluent nitrogen concentrations of around 20 mg N L−1 were slightly above the discharge requirements (>15 mg N L−1; European Commission 1991). Since AerAOB and AnAOB have high affinities for their nitrogen substrates, with half-saturation constants of 0.05–2.4 mg N L−1 (Lackner et al. 2008), the microbial capacity should allow further optimization.

Role of DO levels in suppressing nitratation

The DO levels in the RBC were 1.2–1.4 mg O2 L−1 (Table 1) and therefore not low enough to suppress NOB growth (Bernet et al. 2001; Joss et al. 2009), resulting in a substantial nitratation (Fig. 2). Indeed, NOB could not be detected treating 537 mg N L−1, but the NOB genus Nitrospira colonized the biomass at lower nitrogen influent levels, leading to a final abundance of 5–6%. In contrast to Nitrobacter and Nitrotoga, Nitrospira is typically found in systems under oxygen-limited conditions, relatively low nitrite levels, and moderate temperature (Lücker et al. 2010). Free ammonia levels between 0.08 and 0.8 mg N L−1 can inhibit nitratation (Anthonisen et al. 1976) and could have played a role primarily in period I (0.9 mg NH3–N L−1). A DO decrease by 0.2 mg O2 L−1 during phase Vb lowered nitratation with 35% (Fig. 2), demonstrating the link between DO level and NOB activity. It is anticipated that controlled operation at a sufficiently low DO setpoint (e.g., 0.3 mg O2 L−1) will effectively suppress NOB at the long term, as demonstrated for treatment of higher strength OLAND applications (Joss et al. 2009). In an OLAND RBC, it is less straightforward to control the DO experienced by the biomass than in systems based on active aeration in which the biomass is either suspended (e.g., Joss et al. 2009) or attached to submerged carrier material (e.g., Szatkowska et al. 2007). Lower RBC DO levels can generally be obtained by decreasing the rotor speed (e.g., Meulman et al. 2010), which was not possible on the RBC in this study, or by increasing the immersion level of the disks. These two actions influence the biofilm exposure time to atmospheric oxygen and the input turbulence of air in the bulk liquid by rotation. An additional control of the oxygen level in the gas phase of the RBC could further optimize the microbial balance. In practice, the higher oxygen demand in the presence of organics (165 mg COD L−1) will also yield lower bulk DO levels at a similar rotor speed as in the presence of ammonium only.

OLAND operation at low HRT

Compared to described OLAND systems, the applied HRT in this study were very low (Table 1), but this did not seem to have an adverse effect on AerAOB or AnAOB activity. It is not clear whether an expected higher biomass washout at lower HRT could have been responsible for shifts in the microbial community. The sequentially decreasing nitrogen concentrations in the reactor (Table 2) possibly had a stronger influence on establishing an adapted OLAND microbiome which was likely more oligotrophic.

Compared to treatment at high HRT (e.g., 24 h), applying lower HRT (e.g., 1 h) at high volumetric loading rates will have an influence on the design parameters, depending on the type of OLAND reactor. In suspended growth systems, biomass retention is based on settling. In case of an external settler, a lower HRT will result in a higher sludge surface load and hence needs a relative increase of the settler volume compared to the reactor volume to maintain the same sludge residence time. In case of a sequencing batch reactor, decreased HRT will require an increased minimum biomass settling velocity from 1 m h−1 (Joss et al. 2009) to 24 m h−1 to maintain an acceptably low ratio of settling to reaction time, and therefore will require granules rather than flocs (De Clippeleir et al. 2009; Vlaeminck et al. 2010). In contrast to suspended growth configurations, HRT in biofilm-based systems is expected to have only a minor influence on the biomass retention, allowing for compact reactors.

Implementation of OLAND in the main stream

OLAND treatment of pretreated sewage should achieve sufficiently high nitrogen removal rates and efficiencies at low hydraulic residence times and nitrogen concentrations at minimal energy requirements, given the overall aim of an energy-positive sewage treatment. Overall, several decision factors will determine the desirable reactor technology. Passive versus active aeration will determine not only energy requirements, but also the ease of controlling the microbial activity balance, and suspended versus attached biomass growth will determine the ease of maintaining a high biomass retention at low HRT.

The next research challenges for the implementation of OLAND in the main stream of sewage treatment relate firstly to a decrease of the process temperatures from the maximum summer temperature (25°C) over the average year temperature (17°C) to the minimum winter temperature (8°C) (Breda, the Netherlands; Mollen, oral communication). This will elucidate whether OLAND requires a distinct oligotrophic and cold-tolerant autotrophic community and physiology. Secondly, the continued OLAND performance will have to be shown in the presence of moderate bCOD levels (90–240 mg L−1), with COD/N ratios between 2.4 and 3. The latter will likely facilitate DO control at low DO levels due to heterotrophic aerobic activity. However, competition for nitrite will also take place between heterotrophic denitrification and anammox. These processes have however been demonstrated already to successfully co-exist at a COD/N of 2.2 (Desloover et al. 2011). It is anticipated that due to future dilution preventions (Henze 1997; Brombach et al. 2005), higher nitrogen sewage levels together with the higher sewage temperature will facilitate OLAND treatment in the main stream. Finally, a high OLAND performance will have to be shown under realistic temporal variations in sewage composition and in the performance of the preceding advanced concentration.

Notes

Acknowledgments

H.D.C. received a PhD grant from the Institute for the Promotion of Innovation by Science and Technology in Flanders (IWT-Vlaanderen, SB-81068), and S.E.V. was supported as a postdoctoral fellow from the Research Foundation Flanders (FWO-Vlaanderen). The authors gratefully thank Hans Mollen (Waterschap Brabantse Delta, the Netherlands) for sharing temperature data, Siska Maertens for molecular analyses, and Nico Boon, Tom Hennebel, Jan Arends, Yu Zhang, Sebastià Puig, and Samik Bagchi for inspiring scientific discussions.

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Copyright information

© Springer-Verlag 2011

Authors and Affiliations

  • Haydée De Clippeleir
    • 1
  • Xungang Yan
    • 1
  • Willy Verstraete
    • 1
  • Siegfried Elias Vlaeminck
    • 1
  1. 1.Laboratory of Microbial Ecology and Technology (LabMET)Ghent UniversityGentBelgium

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