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denitrification in a biofilm airlift suspension (BAS) reactor with biodegradable carrier material
water research 43 (2009) 4461–4468
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Simultaneous nitrification/denitrification in a biofilm airlift suspension (BAS) reactor with biodegradable carrier material Evelyn Waltersa,b, Andrea Hillea, Mei Hea, Clemens Ochmannc, Harald Horna,* a
Institute of Water Quality Control, Technische Universita¨t Mu¨nchen, Am Coulombwall, 85748 Garching, Germany Center for Environmental Biotechnology, The Biodesign Institute at Arizona State University, Tempe, AZ 85287, USA c VertUm GmbH, Kirschallee 1, 04416 Markkleeberg, Germany b
article info
abstract
Article history:
Simultaneous nitrification and denitrification in one reactor has been realized with
Received 28 February 2009
different methods in the past. The usage of biodegradable biocompounds as biofilm
Received in revised form
carriers is new. The biocompounds were designed out of two polymers having different
3 July 2009
degradability. Together with suspended autotrophic biomass the biocompound particles
Accepted 7 July 2009
were fluidized in an airlift reactor. Process water from sludge dewatering with a mean
Published online 27 July 2009
ammonium nitrogen concentration of 1150 mg L1 was treated in a two stage system which achieved a nitrogen removal of 75%. Batch experiments clearly indicate that nitrification
Keywords:
can be localized in the suspended biomass and denitrification in the pore structure of the
Biofilm
slowly degraded biocompounds. Images taken with CLSM prove the concept of the pore
Airlift reactor
structure within the biocompounds, which provide both a heterotrophic biofilm and
Biocompound carrier
carbon source.
Simultaneous
ª 2009 Elsevier Ltd. All rights reserved.
nitrification/denitrification CLSM
1.
Introduction
The removal of nitrogen from industrial and municipal wastewater is most often carried out biologically via conventional nitrification and denitrification processes. Nitrification is understood to be the autotrophic oxidation of ammonium (NHþ 4) to nitrate (NO 3 ) via the intermediate product nitrite (NO2 ). Denitrification involves the heterotrophic conversion of nitrate to nitrogen gas under anoxic conditions (EPA, 1993). Although instances of autotrophic denitrification, heterotrophic
nitrification, and anaerobic ammonium oxidation have been described (van Loosdrecht and Jetten, 1998), the conventional conversions predominantly exist in practice. Therefore, in order to provide suitable environments for both nitrifiers and denitrifiers, two separate systems (aerobic and anoxic) are typically utilized. In recent years the process of simultaneous nitrification and denitrification (SND) has been investigated. This process implies that both nitrification and denitrification occur concurrently in one reactor under identical overall operating
Abbreviations: CLSM, confocal laser scanning microscopy; BAS, biofilm airlift suspension; EPS, extracellular polymeric substances; MLSS, mixed liquor suspended solids; PCL, polycaprolactone; PHB, polyhydroxybutyrate; SND, simultaneous nitrification and denitrification. * Corresponding author. Tel.: þ49 89 2891 3713; fax: þ49 89 2891 3718. E-mail address: [email protected] (H. Horn). 0043-1354/$ – see front matter ª 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2009.07.005
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conditions (Muench et al., 1996). If successful, this process could reduce the relatively large reactor volumes and energy costs for recirculation required for separate aerobic and anoxic systems. Studies done with microelectrodes show that for large flocs with diameters of approximately 3000 mm, nitrification and denitrification will take place in different zones of the floc (Satoh et al., 2003). Nevertheless, such flocs are a special case in the activated sludge process. Several types of treatment units have been proposed in which SND can be realized (Guo et al., 2005; Zhang et al., 2007). Zhang et al. (2007) introduced a flexible biofilm reactor having adjustable aerobic, buffer and anoxic zones with liquid circulation being dependent on the aeration flow rate. Both studies were successful in proving the possibility of nitrification and denitrification in one reactor. Biofilm reactors alone or in combination with activated sludge systems are advantageous when slow-growing bacteria such as nitrifiers are required. When immobilized in a biofilm, their growth rates in turn become independent of the hydraulic retention time. Due to the differing growth rates of microorganisms, heterotrophs and autotrophs are typically stratified throughout the biofilm depth (van Loosdrecht et al., 1995; Okabe et al., 1996). Nitrification is adversely affected at high COD:N ratios due to the direct competition for molecular oxygen between autotrophs and heterotrophs. If the COD:N ratio is correspondingly decreased, denitrification will be inhibited due to the deficiency of an electron donor source. In an attempt to overcome the difficulties of SND outlined above, a biofilm airlift suspension reactor operated with biofilm-covered biodegradable carriers and activated sludge was proposed for the treatment of reject water having a low COD:N ratio. A laboratory scale study was conducted to determine whether SND was feasible using biodegradable compounds to compensate for the limited organic loading. Biodegradable
material has already been introduced as a suitable biofilm carrier for water treatment 7 and 9 years ago (Boley et al., 2000; Anton et al., 2002). More recently, biocompounds have been investigated at the microscopic scale with confocal laser scanning microscopy (CLSM) and microelectrodes (Hille et al., 2009). In this study it was postulated that anoxic zones would form both in the deeper layers of the biofilm as well as in the porous structures of the biodegradable carriers despite a broth dissolved oxygen concentration ranging between 0.5 and 3 mg L1. The assumption was made that the carrier material would act as a carbon source for denitrifiers found in these zones, thereby permitting denitrification.
2.
Materials and methods
2.1.
The biofilm airlift suspension reactor
A biofilm airlift suspension reactor cascade consisting of two glass bubble columns (working volume 30 L per column) operated in series and a settling tank was constructed with the goal of achieving simultaneous nitrification and denitrification. The schematic diagram of the reactor cascade is depicted in Fig. 1. A separating plate was inserted into both columns, thus subdividing them into a riser and downcomer. By aerating only one half of the base plate through a membrane in each column, the desired looped-flow and successful fluidization of the carriers was achieved. Both bubble columns were operated at the same oxygen concentration. As the reactor was operated with both suspended biomass and biofilm on the biodegradable carrier, the sludge from the settling tank was recycled at a rate of 200% and reintroduced into the top of the first column together with
Fig. 1 – Schematic diagram of biofilm airlift suspension reactor cascade.
water research 43 (2009) 4461–4468
the influent. To prevent carriers from clogging the tubes transferring the liquid overflows, a sieve was inserted at the top of each reactor. The temperature (21 2.6 C), pH, and dissolved oxygen values in the second column of the reactor system were continuously monitored and recorded by a data acquisition system. Introduction of NaOH solution c(NaOH) ¼ 0.5 mol L1 for pH adjustments (pH was set to 7.0) in the second column was controlled by an electrical control panel. The liquid and aeration flow rates were manually adjustable.
2.2.
Biocompound carriers
Biocompound carriers were used in combination with activated sludge to promote simultaneous nitrification and denitrification of reject water from sludge dewatering at the Gaching (Germany) treatment plant having a low COD:N ratio. The sludge (excess sludge from an activated sludge tank and a trickling filter) was treated in a single stage anaerobic digester. The biodegradable particles were produced by an extrusion process at the Martin-Luther-Universita¨t HalleWittenberg (Chair of Rheology) and were comprised of two components having unequal degradability (Hille et al., 2009). Table 1 outlines the main characteristics of the biocompound carriers. Polyhydroxybutyrate (PHB) is the more readily degradable component of the biocompound carriers and thus, a porous network with increasing tortuosity forms within the polycaprolactone (PCL) over time. As a result, the protected surface area available for slow-growing microorganisms, which are susceptible to detachment from the outer surfaces, becomes greater. Therefore, in addition to acting as an extra carbon source for microorganisms, the biocompound carriers provide microniches for biofilm growth.
2.3.
Reactor operation
As part of the start-up phase, the reactor system was inoculated with activated sludge from aeration tanks of the Garching, Germany wastewater treatment plant. Initially only municipal wastewater after primary settling was utilized. After a certain period of time the municipal wastewater feed was then mixed with reject water resulting from the sludge dewatering process (Garching WWTP, Germany), whereby the ammonium–nitrogen concentrations of this water varied between 700 and 1400 mg L1. The reject water had a COD
Table 1 – Important characteristics of the biocompound carriers. Characteristics of the carriers Material Diameter Height Specific surface area Density
PHB 60%: PCL 40% composite 3.5 mm 3.5 mm 1731 m2 m3 solid material or 1038.6 m2 m3 bulk carrier volume 1.05 kg L1 carrier material or 0.63 kg L1 for bulk carrier volume
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concentration of approximately 550 mg L1. As this water had already been through the treatment processes, the concentration of readily-degradable organic carbon was very low resulting in a BOD5 concentration of roughly 70 mg L1. The ratio of reject water to municipal wastewater was gradually increased to allow the biocenosis enough time to adjust to the elevated ammonium concentrations. From day 73 onwards, the system was fed only reject water. On day 136 a bulk carrier volume of 3 L was added to the reactor cascade (1.5 L per reactor). Throughout the stabilized phase starting from day 190, the volumetric flow rate of the reject water into the reactor varied between 5 and 20 L day1 (see Fig. 2).
2.4.
Batch experiments
Batch experiments were conducted on the biocompound carriers as well as sludge from the reactor cascade in order to determine the maximum nitrification and denitrification rates. These tests, run at defined conditions, allowed for the differentiation between individual capacities of the sludge and carriers. Nitrification experiments were performed in stirred beakers with a liquid volume of 1.5 L. A pre-defined volume of carriers or sludge was gently rinsed with tap water to remove all traces of the reactor broth. Afterwards, the carriers or sludge were added to tap water containing ammonium in excess (400 mg L1 NH4Cl solution diluted in tap water), whereby the initial ammonium–nitrogen concentration was roughly 100 mg L1. The dissolved oxygen concentration was maintained by aeration between 1 and 3 mg L1 and the pH between 7 and 8 through the addition of NaOH solution c(NaOH) ¼ 0.5 mol L1. Samples were periodically taken throughout the experiment to determine nitrogen compound (NHþ 4 –N, NO2 –N, NO3 –N) concentrations. Denitrification experiments were conducted in stirred flasks having a liquid volume of 0.8 L. A pre-defined volume of carriers or sludge was washed gently in tap water and added to a synthetic medium containing nitrate in excess. The initial nitrate–nitrogen concentration was approximately 60 mg L1. The synthetic substrate medium used for the denitrification experiments consisted of: 450 mg L1 KNO3, 5.4 g L1 K2HPO4, 3.8 g L1 NaH2PO4, 8 mg L1 NaCl, 5 mg L1 MgSO4 $ 7H2O, 16.75 mg L1 KCl, and 23.25 mg L1 KH2PO4. A total of 500 mg L1 C6H12O6 was added to the sludge experiments to determine how the maximum rate was affected when carbon was not the limiting factor. The dissolved oxygen concentration ranged from 0 to 0.9 mg L1. Additional tests were run at higher dissolved oxygen concentrations (1–3 mg L1) with the intent of simulating reactor-like conditions. Samples were collected throughout the experiment and tested for nitrate–nitrogen and COD concentrations.
2.5.
Analytical methods
Wet chemical tests (Hach Lange) were performed to analyze the following parameters: NHþ 4 –N, NO2 –N, NO3 –N, and COD. Flow injection analysis was used to validate the nitrate– nitrogen concentrations yielded by the wet tests.
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Volumetric flow rate
6
20 15
4
10 2 5 0
-1
200
250
300 Day of cultivation [d]
350
20 15 10 5 0
-1
0
25
-1
Ammonium-N load [g d ]
25
8
Volumetric flow rate [L d ]
Ammonium-N load Dissolved oxygen concentration
Dissolved oxygen concentration [mg L ]
30
Fig. 2 – Ammonium–nitrogen mass loading and mean dissolved oxygen (DO) concentration in both reactor cascades during the stabilized phase.
The concentration of mixed liquor suspended solids (MLSS) was measured in accordance with the standard method 2540D after drying for 2 h at 105 C (APHA et al., 1998). The MLSS were separated from the carrier material by sieving. The temperature and dissolved oxygen concentration were measured with a DO meter and probe (WTW Oxi 340i, WTW CellOx 325). The pH was determined with a pH meter and electrode (WTW Microprocessor pH Meter 96/196, WTW pH-Electrode SenTix 41). Online tracking of data was realized with a data acquisition system (Fluke Hydra).
2.6.
Confocal laser scanning microscopy (CLSM)
Biofilm bacteria were stained using the SYBR Green nucleic acid stain (Invitrogen, Eugene, OR). The stock solution of SYBR Green as supplied was used at a dilution of 1:1000 in deionized water. Glycoconjugates of the biofilms were stained with the Aleuria aurantia lectin. The lectin was labeled using the Alexa633 staining kit in accordance with the data sheet of the supplier (Invitrogen). A 1:10 dilution of the stock solution and deionized water was used. Each week carrier samples were removed from the cascade. One portion of the carriers was embedded in a liquid cryosectioning medium (Frozen Section Medium Neg-50, from Thermo Scientific) and allowed to incubate for 20 min. Thereafter, the samples were rapidly frozen on a rapid freezing station and sliced along either the circular or vertical plane in 100 mm increments using the Microtome Cryostat Cryo-Star HM 560 MV (Microm International GmbH, Walldorf, Germany) operated at 20 C. Typically the top 100–500 mm was sliced off the particle to allow investigation of biofilm development within the porous structures. Once the desired depth within the particle was attained (visual inspection), the carriers were defrosted and mounted (along with the uncut particles) in a Petri dish with silicone caulk. Both the cut and uncut samples were incubated at room temperature for 20 min in the Alexa-633-labeled lectin. Thereafter, the samples were rinsed three times with tap water to remove any unbound lectin. Finally, the bacteria in the biofilm were counterstained with SYBR Green and allowed
to incubate for 5 min. The biofilm was examined with confocal laser scanning microscopy (CLSM). The CLSM utilized was an upright laser scanning microscope (Zeiss, LSM510 META) controlled by confocal software, v3.2 (Zeiss, Germany). Laser excitation for the lectin was at 633 nm and for SYBR Green at 488 nm, whereby emission signals were collected between 650 and 800 nm and 500 and 550 nm, respectively. The reflection signal from the carrier particle was excited at 488 nm and the emission signal collected from 475 to 490 nm. The samples were submerged in tap water and scanned in the xy-direction (parallel to the surface) with a 40 (0.8 numerical aperture) water immersible lens. A step size of 0.78 mm was used. The captured images had a resolution of 512 512 pixels and a corresponding edge length of 230.3 mm.
3.
Results and discussion
3.1.
Performance of the reactor
In the following, the results of the reactor operation between days 190 and 380 are presented. This phase was characterized by stable reactor performance. Fig. 2 depicts the operation conditions of the cascade during this period with respect to the mass loading of ammonium–nitrogen and average dissolved oxygen concentration. Between days 235 and 270 the mass loading was steadily decreased with the intention of complete nitrification. For the last 2 weeks the ammonium– nitrogen load was doubled starting on day 341. The aeration rates in both columns during the stabilized phase were roughly 110 L h1 (standard ambient temperature and pressure). The MLSS concentration in the cascade during the period was very stable with a mean value of 3.3 0.5 gMLSS L1. The MLSS concentration was kept constant by the loss of biomass through the effluent. Between days 237 and 242, a bulk carrier volume of 0.5 to 0.7 L (0.31–0.44 kg) was added to each reactor in the cascade to account for their volume losses due to degradation. In response to an increase in the effluent nitrate–nitrogen
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3.1.1.
COD elimination in the reactor cascade
COD removal efficiency was roughly 60% from days 240 to 380 with an average COD concentration in the influent of 550 mg L1 (results not shown). It was determined that the COD in the reject water contained only a small fraction of readilydegradable organic substances (average BOD5 – concentration was 70 mg L1). It can therefore be assumed that around 220 mg L1 of COD remaining was refractory. Assuming a ratio of 5 gCOD/gNO3–N the degraded COD could have been responsible for less than 10% of the denitrified nitrate– nitrogen. This can be seen from the average values for the main parameters, which are shown in Table 2.
3.1.2.
-1
concentration, on day 321 an additional 2 L bulk carrier volume (1.26 kg) was added to both reactors.
Concentration N-parameters [mg L ]
water research 43 (2009) 4461–4468
1800
+
NH4 -N influent
+
NH4 -N effluent -
1500
NO2 -N effluent
1200
NO3 -N effluent
-
900 600 300 0 200
250
300
350
Day of cultivation [d]
Fig. 3 – Influent and effluent concentrations of nitrogen parameters in the reactor cascade during the stabilized phase.
Nitrogen elimination
The effectiveness of nitrogen removal in the reactor cascade is depicted in Fig. 3. Except for a few fluctuations between days 190 and 221, the mean nitrite–nitrogen concentration in the effluent throughout the stabilized phase was 0.1 mg L1. Additionally, the removal of ammonium–nitrogen was complete. The nitrification efficiencies of nearly 100% indicate that the cascade was not operated at maximum loading and thus, the turnover rates were limited by the substrate loading. Even when the ammonium–nitrogen load was doubled on day 341 (Fig. 2), nitrification was complete. The average sludge-specific nitrification rate in the cascade from days 190 to 380 was 1 and the volumetric determined to be 0.06 gNH4–N g1 MLSS day turnover rate for the same time period was 200 118 gNH4–N m3 day1 (see Table 2). Between days 260 and 290 nitrate–nitrogen was reduced to a concentration between approximately 14 and 40 mg L1. The nitrate–nitrogen elimination is attributed to the denitrification capacity of the biocompound carriers. These results were validated through batch experiments. As seen in Fig. 3, the effluent nitrate–nitrogen concentration increases from days 290 to 321. The decrease in removal efficiency is linked with the steadily diminishing volume of the biocompound carriers in the cascade due to their degradation. Fig. 4 shows the significant carrier degradation after 290 days of cultivation. It can be seen that the carrier material is nearly completely degraded and both the PHB and the PCL are metabolized by the bacteria. A mass balance essentially revealed that most of the organic carbon necessary for denitrification was provided by
the carriers themselves, and not by the COD-degradation (see above). Despite an addition of 0.5 and 0.7 L of carriers on day 237 to the first and second reactors in the cascade, respectively, the volume of the biocompound carriers dropped to 0.5 L per reactor by day 320. Therefore, on day 321 a bulk carrier volume of 2 L (new carriers) was added to each reactor. Denitrification was successfully restored within 1 week after the addition of new carriers. The retention time of the reject water was between 1 and 3 days depending on the volumetric flow rate (see Fig. 2). Therefore, it is difficult to calculate the daily surface-specific denitrification rates as both volumetric flow rate and ammonium concentration fluctuated over time. Nevertheless, the mean surface-specific denitrification rate for the stabilized phase between days 190 and 380 was 2.76 1.48 gNO3–N m2 day1 (see Table 2). Typically, denitrification with methanol in biofilters yields surface degradation rates between 1 gNO3–N m2 day1 (Koch and Siegrist, 1997) and 10 gNO3–N m2 day1 (Horn and Telgmann, 2000), which depends mainly on carrier material and filter velocity. After the increase of ammonium load on day 340 the system was still able to completely nitrify the ammonium, but the nitrate concentration in the effluent increased to approximately 100 mg/L (Fig, 3). To have a better understanding on the maximum denitrification capacity the biocompounds were more closely investigated in batch experiments.
3.2.
Batch experiments
Batch experiments were conducted to determine the extent to which nitrification and denitrification could occur in the Table 2 – Average volumetric flow rate, loading and removal rate with standard deviation. Parameter Volumetric flow rate COD Ammonium Denitrification
L g g g g
1
day m3 m3 m3 m2
day1 day1 day1 day1
Load
Removal
11.05 4.79 101 52.8 210 120 – –
– 56.6 8.44 200 118 158 103 2.76 1.49a
a Surface specific denitrification rate related to biocompound surface.
Fig. 4 – Dried biocompound carriers after 290 days of cultivation in the cascade. Scale bar: 1 mm.
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stabilized reactor cascade. The experiments were specifically performed so that the individual contributions of the biocompound carriers and sludge in the turnover rates could be identified.
3.2.1.
have a stable and complete nitrogen removal. Throughout this 2-month period the particle size decreased (Fig. 4) which, in addition to the increased maximum denitrification rates observed in the batch experiments, indicates the development of more denitrifying bacteria within the microniches on the carrier surface. It is assumed that the heterotrophic denitrifying bacteria utilized the biocompound material as an electron donor source. To permit a direct comparison between the denitrification capacities of sludge and biofilm, the results of the batch experiments were extrapolated to the reactor cascade. Turnover rates of 0.42 and 0.80 gNO3–N day1 were calculated for sludge without and with glucose, respectively. For biofilm on the biocompound carriers a turnover rate of 9.77 gNO3–N day1 was calculated. This indicates that biofilm was responsible for roughly 92% of denitrification in the reactor cascade.
Nitrification experiments
The findings of the nitrification batch experiments performed on both activated sludge and biocompound carriers are presented in Fig. 5. The maximum nitrification rates realized by the biofilm attached to the carriers ranged between 0.12 and 0.40 gNH4–N m2 day1 (average: 0.26 0.10 gNH4–N m2 day1). Maximum nitrification rates of sludge ranging between 0.12 1 (average: 0.32 0.12 gNH4–N and 0.48 gNH4–N g1 MLSS day 1 1 g MLSS day ) were measured. To allow for a direct comparison of the individual contributions of biofilm and sludge to nitrification, the measured rates from the batch experiments were extrapolated to the reactor cascade. These results correspond to sludge and biofilm turnover rates in the cascade of 28.26 and 0.75 gNH4–N day, respectively. It is apparent that sludge was responsible for roughly 98% of nitrification in the reactor cascade. Furthermore, it can be seen that the nitrification capacity of the sludge was not fully used in the reactor as the sludge load was day1 throughout the entire below 0.1 gNH4–N g1 MLSS experiment.
3.2.2.
3.3.
To track biofilm development on the biocompound carriers, CLSM was performed on a weekly basis from days 135 to 350. Examination of the carriers indicated considerable biofilm coverage as early as 2 weeks after introduction into the reactor cascade. It was confirmed that growth was concentrated in the porous structures (Hille et al., 2009). Biofilm coverage on the outer surfaces of the carriers was sparse due to the high abrasional losses (e.g. particle-to-particle collisions) and shear forces in the cascade. Biofilm thicknesses were often less than 100 mm, even inside the porous structures. Maximum thicknesses of 300 mm were recorded, but only rarely. Fig. 7 gives an impression of the biofilm structure within the biocompound carrier material. The porous structure as well as both bacteria and extracellular polymeric substances (EPS) can clearly be identified. Due to oxygen in the bulk phase, anoxic zones for denitrification can only be provided by such structures which additionally offer organic carbon as an electron donor. It is not clear whether the EPS are used as an electron donor by the denitrifying heterotrophic bacteria (Horn et al., 2001). Zhang and Bishop, 2003 conducted degradation experiments with EPS from mixed culture biofilms. The authors could show that the original microorganisms were able to degrade the EPS when they were starved.
Denitrification experiments
0.4
Sludge Bio compounds
-1
-1
or [gNH4+-N gMLSS d ]
-2
-1
Nitrification rate [gNH4+-N m d ]
The results of the denitrification batch experiments are shown in Fig. 6. For tests conducted both under ideal (DO: <0.9 mg L1) and reactor-like conditions (DO: 1–3 mg L1), no significant denitrification was detected for sludge. On average, maximum denitrification rates with and without addition of 1 glucose of 0.01 and 0.003 gNO3–N g1 MLSS day , respectively, were attained. This indicates that denitrification activity in the cascade can be attributed to the biofilm on the biocompound carriers. Autotrophic microorganisms seemed to dominate the sludge within the reactor cascade. From Fig. 6 it is evident that significant denitrification is achieved by the biocompound carriers. The maximum denitrification rate steadily increased from days 230 to 290. Increasing denitrification rates were not observed in the cascade during this time period (Fig. 6) due to the fact that the ammonium–nitrogen load was steadily decreased (Fig. 2) to
0.5
Anoxic zones within the biofilm structure
0.3 0.2 0.1 0.0 160
180
200
220
240
260
280
300
320
340
Day of cultivation [d] Fig. 5 – Results of batch tests. Maximum nitrification rates of biocompound carriers and sludge related either to the carrier L1 . surface area or amount of MLSS. Oxygen concentration: 1–3 mg LL1; pH: 7–8; initial NHD 4 –N concentration: 100 mg L
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-1
Denitrification rate [gNO3--N m d ]
water research 43 (2009) 4461–4468
4
-1
-1
or [gNO3--N gMLSS d ]
-2
5
Biocompounds Sludge Sludge with glucose
3
DO = 1 – 3 mg L-1
DO = 1 – 3 mg L-1
2 1 0 160
180
200
220
240
260
280
300
320
340
Day of cultivation [d] Fig. 6 – Results of batch tests. Maximum denitrification rates of biocompound carriers and sludge related either to the carrier surface area or amount of MLSS. Oxygen concentration: less than 0.9 mg LL1 (except where otherwise denoted); initial NOL 3 – N concentration: 60 mg LL1. Arrows indicate measured rates of zero.
3.4. Degradation of biocompound material and denitrification To elucidate the origin of the electron donor a estimation on the biocompound carrier consumption has been made for the period between days 220 and 318. It has to be stressed that the measurement of the bulk carrier material at day 318 was not easy as the carrier diameter was reduced (cp. Fig. 4). Approximately 2.1 L of bulk carrier material (from 3.2 L on day 220 to 1.1 L on day 318) was lost and 0.95 kg of nitrogen was eliminated within this period. A total of 2.1 L of bulk carrier material represent 1.3 kg of degradable carrier material. Taking the ratio 60:40 for PHB:PCL, 1 g of dry carrier material accounts for 1.84 g of COD. This gives a highly efficient use of the organic carbon as only a very small amount of nitrate removal (less than 10%) can be attributed to degraded COD (see Section 3.1). For denitrification with methanol in half or full scale biofilters, typically 2.5 g methanol (3.75 g COD) are added for 1 g of nitrate–nitrogen (Horn and Telgmann, 2000). Koch and Siegrist (1997) gave a higher value of 3.2 g methanol per 1 g of nitrate– nitrogen. Due to the situation of limited substrate within the pores, it can be assumed that only very small amounts of biomass are produced and that the main part of the organic carbon is transformed to CO2. An additional option is the direct transformation of nitrite– nitrogen, which should be possible at low oxygen concentrations and a low sludge age (Blackburne et al., 2008). The first condition of low dissolved oxygen was essentially fulfilled in the reactor system, however the sludge age was too high (>10 days) to prevent the development of nitrite oxidizing bacteria (NOB). Perhaps a small fraction of nitrite (mean concentration in the effluent 0.1 mg L1) was directly reduced inside the pore structure. On the other hand, as nitrification was done by the flocs, it is more plausible that ammonium was totally oxidized to nitrate, and the nitrate was then reduced in the pore structure of the biocompound carriers.
3.5.
Possible scale-up for use in wastewater treatment
The presented system is designed to treat wastewater with high ammonium concentrations where nearly no organic carbon is available for denitrification. Reject water with
temperatures of up 20–30 C should be treated with Anammox as there is no organic carbon is necessary (Tho¨le et al., 2005; van der Star et al., 2007). The process presented in this study is advantageous for wastewater with temperatures below 20 C.
Fig. 7 – Example of biofilm growth concentrated in the pore structures of biocompound particles rather than on the external surfaces. (a) Day 378 of particle cut parallel to cylindrical top. (b) Day 350 of whole particle. Green represents EPS, red nucleic acids (e.g. bacteria), and white reflection signal. Scale bar [ 50 mm.
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Compared to Anammox, the control strategies seemed less complicated as oxygen and biocompound concentration are the main and relevant parameter. The system presented does not suffer from complex population dynamics as the growth of nitrite oxidizing bacteria (NOB) will heavily disturb the Anammox process. In this study it has been shown that very low ammonium and nitrate concentrations can be achieved in the effluent, which is important if the system is used as a final treatment step. Up to now the biodegradable biocompounds (10 V per kg) are too expensive to compete with methanol driven denitrification. Nevertheless, in some cases like treatment of recycling water from sensitive fish farming it may be advantageous to operate a single stage nitrification/denitrification unit without storage of methanol.
4.
Conclusions
A biofilm airlift suspension reactor dually operated with both biofilm-covered biodegradable carriers and activated sludge was developed for the treatment of wastewater containing a low COD:N ratio: The main nitrifying activity can be found in the suspended biomass. Denitrification was mainly located on the carrier material where the organic carbon is provided by the carrier itself as was evident by the carrier degradation. Both processes, denitrification and nitrification, could be stabilized for oxygen concentrations of 0.5–3 mg L1 in the bulk liquid. The amount of carrier material restricted the total removal of nitrogen. An average volumetric denitrification rate of 158 gNO3–N m3 day1 was achieved with 5% bulk carrier volume of biocompounds.
references
Anton, W., Jank, M., Schnabel, R., Ochmann, C., 2002. Biocompounds as carrier material for complex wastewater treatment [Biocompounds als Traegermaterial zur komplexen Abwasserreinigung]. KA – Wasserwirtschaft, Abwasser. Abfall 49 (9), 1222–1227. APHA, WEF, AWWA, 1998. Standard Methods for the Examination of Water and Wastewater. American Public Health Association, Washington, DC. Blackburne, R., Yuan, Z., Keller, J., 2008. Demonstration of nitrogen removal via nitrite in a sequencing batch reactor treating domestic wastewater. Water Research 42 (8–9), 2166–2176.
Boley, A., Mueller, W.R., Haider, G., 2000. Biodegradable polymers as solid substrate and biofilm carrier for denitrification in recirculated aquaculture systems. Aquacultural Engineering 22 (1–2), 75–85. EPA, 1993. Manual: Nitrogen Control. W. United States Environmental Protection Agency, D.C. EPA/625/R-93/010. Guo, H., Zhou, J., Su, J., Zhang, Z., 2005. Integration of nitrification and denitrification in airlift bioreactor. Biochemical Engineering Journal 23, 57–62. Hille, A., He, M., Ochmann, C., Neu, T.R., Horn, H., 2009. Application of two component biodegradable carriers in a particle fixed biofilm airlift reactor – development and structure of biofilm. Bioprocess and Biosystems Engineering 32, 31–39. Horn, H., Telgmann, U., 2000. Simulation of tertiary denitrification with methanol in an upflow biofilter. Water Science Technology 41 (4-5), 185–190. Horn, H., Neu, T.R., Wulkow, M., 2001. Modelling the structure and function of extracellular polymeric substances in biofilms with new numerical techniques. Water Science Technology 43 (6), 121–127. Koch, G., Siegrist, H., 1997. Denitrification with methanol in tertiary filtration. Water Research 31, 3029–3038. Muench, E., Lant, P., Keller, J., 1996. Simultaneous nitrification and denitrification in bench-scale sequencing batch reactors. Water Research 30 (2), 277–284. Okabe, S., Hiratia, K., Ozawa, Y., Watanabe, Y., 1996. Spatial microbial distributions of nitrifiers and heterotrophs in mixed-population biofilms. Biotechnology and Bioengineering 50, 24–35. Satoh, H., Nakamura, Y., Ono, H., Okabe, S., 2003. Effect of oxygen concentration on nitrification and denitrification in single activated sludge flocs. Biotechnology and Bioengineering 83, 604–607. Tho¨le, D., Cornelius, A., Rosenwinkel, K.-H., 2005. Großtechnische Erfahrungen zur Deammonifikation von Schlammwasser auf der Kla¨ranlage Hattingen. (Full scale experience with deammonification of reject water on WWTP Hattingen.) gwf Wasser/Abwasser 146, 104–109. 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. van Loosdrecht, M.C.M., Jetten, M.S.M., 1998. Microbiological conversions in nitrogen removal. Water Science Technology 38 (1), 1–7. van Loosdrecht, M.C.M., Tijhuis, L., Wijdieks, A.M.S., Heijnen, J.J., 1995. Population distribution in aerobic biofilms on small suspended particles. Water Science Technology 31 (1), 163–171. Zhang, X., Bishop, P., 2003. Biodegradability of biofilm extracellular polymeric substances. Chemosphere 50 (1), 63–69. Zhang, X., Zhou, J., Guo, H., Qu, Y., Liu, G., Zhao, L., 2007. Nitrogen removal performance in a novel combined biofilm reactor. Process Biochemistry 42, 620–626.
Available at www.sciencedirect.com
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Simultaneous nitrification/denitrification in a biofilm airlift suspension (BAS) reactor with biodegradable carrier material Evelyn Waltersa,b, Andrea Hillea, Mei Hea, Clemens Ochmannc, Harald Horna,* a
Institute of Water Quality Control, Technische Universita¨t Mu¨nchen, Am Coulombwall, 85748 Garching, Germany Center for Environmental Biotechnology, The Biodesign Institute at Arizona State University, Tempe, AZ 85287, USA c VertUm GmbH, Kirschallee 1, 04416 Markkleeberg, Germany b
article info
abstract
Article history:
Simultaneous nitrification and denitrification in one reactor has been realized with
Received 28 February 2009
different methods in the past. The usage of biodegradable biocompounds as biofilm
Received in revised form
carriers is new. The biocompounds were designed out of two polymers having different
3 July 2009
degradability. Together with suspended autotrophic biomass the biocompound particles
Accepted 7 July 2009
were fluidized in an airlift reactor. Process water from sludge dewatering with a mean
Published online 27 July 2009
ammonium nitrogen concentration of 1150 mg L1 was treated in a two stage system which achieved a nitrogen removal of 75%. Batch experiments clearly indicate that nitrification
Keywords:
can be localized in the suspended biomass and denitrification in the pore structure of the
Biofilm
slowly degraded biocompounds. Images taken with CLSM prove the concept of the pore
Airlift reactor
structure within the biocompounds, which provide both a heterotrophic biofilm and
Biocompound carrier
carbon source.
Simultaneous
ª 2009 Elsevier Ltd. All rights reserved.
nitrification/denitrification CLSM
1.
Introduction
The removal of nitrogen from industrial and municipal wastewater is most often carried out biologically via conventional nitrification and denitrification processes. Nitrification is understood to be the autotrophic oxidation of ammonium (NHþ 4) to nitrate (NO 3 ) via the intermediate product nitrite (NO2 ). Denitrification involves the heterotrophic conversion of nitrate to nitrogen gas under anoxic conditions (EPA, 1993). Although instances of autotrophic denitrification, heterotrophic
nitrification, and anaerobic ammonium oxidation have been described (van Loosdrecht and Jetten, 1998), the conventional conversions predominantly exist in practice. Therefore, in order to provide suitable environments for both nitrifiers and denitrifiers, two separate systems (aerobic and anoxic) are typically utilized. In recent years the process of simultaneous nitrification and denitrification (SND) has been investigated. This process implies that both nitrification and denitrification occur concurrently in one reactor under identical overall operating
Abbreviations: CLSM, confocal laser scanning microscopy; BAS, biofilm airlift suspension; EPS, extracellular polymeric substances; MLSS, mixed liquor suspended solids; PCL, polycaprolactone; PHB, polyhydroxybutyrate; SND, simultaneous nitrification and denitrification. * Corresponding author. Tel.: þ49 89 2891 3713; fax: þ49 89 2891 3718. E-mail address: [email protected] (H. Horn). 0043-1354/$ – see front matter ª 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2009.07.005
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conditions (Muench et al., 1996). If successful, this process could reduce the relatively large reactor volumes and energy costs for recirculation required for separate aerobic and anoxic systems. Studies done with microelectrodes show that for large flocs with diameters of approximately 3000 mm, nitrification and denitrification will take place in different zones of the floc (Satoh et al., 2003). Nevertheless, such flocs are a special case in the activated sludge process. Several types of treatment units have been proposed in which SND can be realized (Guo et al., 2005; Zhang et al., 2007). Zhang et al. (2007) introduced a flexible biofilm reactor having adjustable aerobic, buffer and anoxic zones with liquid circulation being dependent on the aeration flow rate. Both studies were successful in proving the possibility of nitrification and denitrification in one reactor. Biofilm reactors alone or in combination with activated sludge systems are advantageous when slow-growing bacteria such as nitrifiers are required. When immobilized in a biofilm, their growth rates in turn become independent of the hydraulic retention time. Due to the differing growth rates of microorganisms, heterotrophs and autotrophs are typically stratified throughout the biofilm depth (van Loosdrecht et al., 1995; Okabe et al., 1996). Nitrification is adversely affected at high COD:N ratios due to the direct competition for molecular oxygen between autotrophs and heterotrophs. If the COD:N ratio is correspondingly decreased, denitrification will be inhibited due to the deficiency of an electron donor source. In an attempt to overcome the difficulties of SND outlined above, a biofilm airlift suspension reactor operated with biofilm-covered biodegradable carriers and activated sludge was proposed for the treatment of reject water having a low COD:N ratio. A laboratory scale study was conducted to determine whether SND was feasible using biodegradable compounds to compensate for the limited organic loading. Biodegradable
material has already been introduced as a suitable biofilm carrier for water treatment 7 and 9 years ago (Boley et al., 2000; Anton et al., 2002). More recently, biocompounds have been investigated at the microscopic scale with confocal laser scanning microscopy (CLSM) and microelectrodes (Hille et al., 2009). In this study it was postulated that anoxic zones would form both in the deeper layers of the biofilm as well as in the porous structures of the biodegradable carriers despite a broth dissolved oxygen concentration ranging between 0.5 and 3 mg L1. The assumption was made that the carrier material would act as a carbon source for denitrifiers found in these zones, thereby permitting denitrification.
2.
Materials and methods
2.1.
The biofilm airlift suspension reactor
A biofilm airlift suspension reactor cascade consisting of two glass bubble columns (working volume 30 L per column) operated in series and a settling tank was constructed with the goal of achieving simultaneous nitrification and denitrification. The schematic diagram of the reactor cascade is depicted in Fig. 1. A separating plate was inserted into both columns, thus subdividing them into a riser and downcomer. By aerating only one half of the base plate through a membrane in each column, the desired looped-flow and successful fluidization of the carriers was achieved. Both bubble columns were operated at the same oxygen concentration. As the reactor was operated with both suspended biomass and biofilm on the biodegradable carrier, the sludge from the settling tank was recycled at a rate of 200% and reintroduced into the top of the first column together with
Fig. 1 – Schematic diagram of biofilm airlift suspension reactor cascade.
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the influent. To prevent carriers from clogging the tubes transferring the liquid overflows, a sieve was inserted at the top of each reactor. The temperature (21 2.6 C), pH, and dissolved oxygen values in the second column of the reactor system were continuously monitored and recorded by a data acquisition system. Introduction of NaOH solution c(NaOH) ¼ 0.5 mol L1 for pH adjustments (pH was set to 7.0) in the second column was controlled by an electrical control panel. The liquid and aeration flow rates were manually adjustable.
2.2.
Biocompound carriers
Biocompound carriers were used in combination with activated sludge to promote simultaneous nitrification and denitrification of reject water from sludge dewatering at the Gaching (Germany) treatment plant having a low COD:N ratio. The sludge (excess sludge from an activated sludge tank and a trickling filter) was treated in a single stage anaerobic digester. The biodegradable particles were produced by an extrusion process at the Martin-Luther-Universita¨t HalleWittenberg (Chair of Rheology) and were comprised of two components having unequal degradability (Hille et al., 2009). Table 1 outlines the main characteristics of the biocompound carriers. Polyhydroxybutyrate (PHB) is the more readily degradable component of the biocompound carriers and thus, a porous network with increasing tortuosity forms within the polycaprolactone (PCL) over time. As a result, the protected surface area available for slow-growing microorganisms, which are susceptible to detachment from the outer surfaces, becomes greater. Therefore, in addition to acting as an extra carbon source for microorganisms, the biocompound carriers provide microniches for biofilm growth.
2.3.
Reactor operation
As part of the start-up phase, the reactor system was inoculated with activated sludge from aeration tanks of the Garching, Germany wastewater treatment plant. Initially only municipal wastewater after primary settling was utilized. After a certain period of time the municipal wastewater feed was then mixed with reject water resulting from the sludge dewatering process (Garching WWTP, Germany), whereby the ammonium–nitrogen concentrations of this water varied between 700 and 1400 mg L1. The reject water had a COD
Table 1 – Important characteristics of the biocompound carriers. Characteristics of the carriers Material Diameter Height Specific surface area Density
PHB 60%: PCL 40% composite 3.5 mm 3.5 mm 1731 m2 m3 solid material or 1038.6 m2 m3 bulk carrier volume 1.05 kg L1 carrier material or 0.63 kg L1 for bulk carrier volume
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concentration of approximately 550 mg L1. As this water had already been through the treatment processes, the concentration of readily-degradable organic carbon was very low resulting in a BOD5 concentration of roughly 70 mg L1. The ratio of reject water to municipal wastewater was gradually increased to allow the biocenosis enough time to adjust to the elevated ammonium concentrations. From day 73 onwards, the system was fed only reject water. On day 136 a bulk carrier volume of 3 L was added to the reactor cascade (1.5 L per reactor). Throughout the stabilized phase starting from day 190, the volumetric flow rate of the reject water into the reactor varied between 5 and 20 L day1 (see Fig. 2).
2.4.
Batch experiments
Batch experiments were conducted on the biocompound carriers as well as sludge from the reactor cascade in order to determine the maximum nitrification and denitrification rates. These tests, run at defined conditions, allowed for the differentiation between individual capacities of the sludge and carriers. Nitrification experiments were performed in stirred beakers with a liquid volume of 1.5 L. A pre-defined volume of carriers or sludge was gently rinsed with tap water to remove all traces of the reactor broth. Afterwards, the carriers or sludge were added to tap water containing ammonium in excess (400 mg L1 NH4Cl solution diluted in tap water), whereby the initial ammonium–nitrogen concentration was roughly 100 mg L1. The dissolved oxygen concentration was maintained by aeration between 1 and 3 mg L1 and the pH between 7 and 8 through the addition of NaOH solution c(NaOH) ¼ 0.5 mol L1. Samples were periodically taken throughout the experiment to determine nitrogen compound (NHþ 4 –N, NO2 –N, NO3 –N) concentrations. Denitrification experiments were conducted in stirred flasks having a liquid volume of 0.8 L. A pre-defined volume of carriers or sludge was washed gently in tap water and added to a synthetic medium containing nitrate in excess. The initial nitrate–nitrogen concentration was approximately 60 mg L1. The synthetic substrate medium used for the denitrification experiments consisted of: 450 mg L1 KNO3, 5.4 g L1 K2HPO4, 3.8 g L1 NaH2PO4, 8 mg L1 NaCl, 5 mg L1 MgSO4 $ 7H2O, 16.75 mg L1 KCl, and 23.25 mg L1 KH2PO4. A total of 500 mg L1 C6H12O6 was added to the sludge experiments to determine how the maximum rate was affected when carbon was not the limiting factor. The dissolved oxygen concentration ranged from 0 to 0.9 mg L1. Additional tests were run at higher dissolved oxygen concentrations (1–3 mg L1) with the intent of simulating reactor-like conditions. Samples were collected throughout the experiment and tested for nitrate–nitrogen and COD concentrations.
2.5.
Analytical methods
Wet chemical tests (Hach Lange) were performed to analyze the following parameters: NHþ 4 –N, NO2 –N, NO3 –N, and COD. Flow injection analysis was used to validate the nitrate– nitrogen concentrations yielded by the wet tests.
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Volumetric flow rate
6
20 15
4
10 2 5 0
-1
200
250
300 Day of cultivation [d]
350
20 15 10 5 0
-1
0
25
-1
Ammonium-N load [g d ]
25
8
Volumetric flow rate [L d ]
Ammonium-N load Dissolved oxygen concentration
Dissolved oxygen concentration [mg L ]
30
Fig. 2 – Ammonium–nitrogen mass loading and mean dissolved oxygen (DO) concentration in both reactor cascades during the stabilized phase.
The concentration of mixed liquor suspended solids (MLSS) was measured in accordance with the standard method 2540D after drying for 2 h at 105 C (APHA et al., 1998). The MLSS were separated from the carrier material by sieving. The temperature and dissolved oxygen concentration were measured with a DO meter and probe (WTW Oxi 340i, WTW CellOx 325). The pH was determined with a pH meter and electrode (WTW Microprocessor pH Meter 96/196, WTW pH-Electrode SenTix 41). Online tracking of data was realized with a data acquisition system (Fluke Hydra).
2.6.
Confocal laser scanning microscopy (CLSM)
Biofilm bacteria were stained using the SYBR Green nucleic acid stain (Invitrogen, Eugene, OR). The stock solution of SYBR Green as supplied was used at a dilution of 1:1000 in deionized water. Glycoconjugates of the biofilms were stained with the Aleuria aurantia lectin. The lectin was labeled using the Alexa633 staining kit in accordance with the data sheet of the supplier (Invitrogen). A 1:10 dilution of the stock solution and deionized water was used. Each week carrier samples were removed from the cascade. One portion of the carriers was embedded in a liquid cryosectioning medium (Frozen Section Medium Neg-50, from Thermo Scientific) and allowed to incubate for 20 min. Thereafter, the samples were rapidly frozen on a rapid freezing station and sliced along either the circular or vertical plane in 100 mm increments using the Microtome Cryostat Cryo-Star HM 560 MV (Microm International GmbH, Walldorf, Germany) operated at 20 C. Typically the top 100–500 mm was sliced off the particle to allow investigation of biofilm development within the porous structures. Once the desired depth within the particle was attained (visual inspection), the carriers were defrosted and mounted (along with the uncut particles) in a Petri dish with silicone caulk. Both the cut and uncut samples were incubated at room temperature for 20 min in the Alexa-633-labeled lectin. Thereafter, the samples were rinsed three times with tap water to remove any unbound lectin. Finally, the bacteria in the biofilm were counterstained with SYBR Green and allowed
to incubate for 5 min. The biofilm was examined with confocal laser scanning microscopy (CLSM). The CLSM utilized was an upright laser scanning microscope (Zeiss, LSM510 META) controlled by confocal software, v3.2 (Zeiss, Germany). Laser excitation for the lectin was at 633 nm and for SYBR Green at 488 nm, whereby emission signals were collected between 650 and 800 nm and 500 and 550 nm, respectively. The reflection signal from the carrier particle was excited at 488 nm and the emission signal collected from 475 to 490 nm. The samples were submerged in tap water and scanned in the xy-direction (parallel to the surface) with a 40 (0.8 numerical aperture) water immersible lens. A step size of 0.78 mm was used. The captured images had a resolution of 512 512 pixels and a corresponding edge length of 230.3 mm.
3.
Results and discussion
3.1.
Performance of the reactor
In the following, the results of the reactor operation between days 190 and 380 are presented. This phase was characterized by stable reactor performance. Fig. 2 depicts the operation conditions of the cascade during this period with respect to the mass loading of ammonium–nitrogen and average dissolved oxygen concentration. Between days 235 and 270 the mass loading was steadily decreased with the intention of complete nitrification. For the last 2 weeks the ammonium– nitrogen load was doubled starting on day 341. The aeration rates in both columns during the stabilized phase were roughly 110 L h1 (standard ambient temperature and pressure). The MLSS concentration in the cascade during the period was very stable with a mean value of 3.3 0.5 gMLSS L1. The MLSS concentration was kept constant by the loss of biomass through the effluent. Between days 237 and 242, a bulk carrier volume of 0.5 to 0.7 L (0.31–0.44 kg) was added to each reactor in the cascade to account for their volume losses due to degradation. In response to an increase in the effluent nitrate–nitrogen
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3.1.1.
COD elimination in the reactor cascade
COD removal efficiency was roughly 60% from days 240 to 380 with an average COD concentration in the influent of 550 mg L1 (results not shown). It was determined that the COD in the reject water contained only a small fraction of readilydegradable organic substances (average BOD5 – concentration was 70 mg L1). It can therefore be assumed that around 220 mg L1 of COD remaining was refractory. Assuming a ratio of 5 gCOD/gNO3–N the degraded COD could have been responsible for less than 10% of the denitrified nitrate– nitrogen. This can be seen from the average values for the main parameters, which are shown in Table 2.
3.1.2.
-1
concentration, on day 321 an additional 2 L bulk carrier volume (1.26 kg) was added to both reactors.
Concentration N-parameters [mg L ]
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1800
+
NH4 -N influent
+
NH4 -N effluent -
1500
NO2 -N effluent
1200
NO3 -N effluent
-
900 600 300 0 200
250
300
350
Day of cultivation [d]
Fig. 3 – Influent and effluent concentrations of nitrogen parameters in the reactor cascade during the stabilized phase.
Nitrogen elimination
The effectiveness of nitrogen removal in the reactor cascade is depicted in Fig. 3. Except for a few fluctuations between days 190 and 221, the mean nitrite–nitrogen concentration in the effluent throughout the stabilized phase was 0.1 mg L1. Additionally, the removal of ammonium–nitrogen was complete. The nitrification efficiencies of nearly 100% indicate that the cascade was not operated at maximum loading and thus, the turnover rates were limited by the substrate loading. Even when the ammonium–nitrogen load was doubled on day 341 (Fig. 2), nitrification was complete. The average sludge-specific nitrification rate in the cascade from days 190 to 380 was 1 and the volumetric determined to be 0.06 gNH4–N g1 MLSS day turnover rate for the same time period was 200 118 gNH4–N m3 day1 (see Table 2). Between days 260 and 290 nitrate–nitrogen was reduced to a concentration between approximately 14 and 40 mg L1. The nitrate–nitrogen elimination is attributed to the denitrification capacity of the biocompound carriers. These results were validated through batch experiments. As seen in Fig. 3, the effluent nitrate–nitrogen concentration increases from days 290 to 321. The decrease in removal efficiency is linked with the steadily diminishing volume of the biocompound carriers in the cascade due to their degradation. Fig. 4 shows the significant carrier degradation after 290 days of cultivation. It can be seen that the carrier material is nearly completely degraded and both the PHB and the PCL are metabolized by the bacteria. A mass balance essentially revealed that most of the organic carbon necessary for denitrification was provided by
the carriers themselves, and not by the COD-degradation (see above). Despite an addition of 0.5 and 0.7 L of carriers on day 237 to the first and second reactors in the cascade, respectively, the volume of the biocompound carriers dropped to 0.5 L per reactor by day 320. Therefore, on day 321 a bulk carrier volume of 2 L (new carriers) was added to each reactor. Denitrification was successfully restored within 1 week after the addition of new carriers. The retention time of the reject water was between 1 and 3 days depending on the volumetric flow rate (see Fig. 2). Therefore, it is difficult to calculate the daily surface-specific denitrification rates as both volumetric flow rate and ammonium concentration fluctuated over time. Nevertheless, the mean surface-specific denitrification rate for the stabilized phase between days 190 and 380 was 2.76 1.48 gNO3–N m2 day1 (see Table 2). Typically, denitrification with methanol in biofilters yields surface degradation rates between 1 gNO3–N m2 day1 (Koch and Siegrist, 1997) and 10 gNO3–N m2 day1 (Horn and Telgmann, 2000), which depends mainly on carrier material and filter velocity. After the increase of ammonium load on day 340 the system was still able to completely nitrify the ammonium, but the nitrate concentration in the effluent increased to approximately 100 mg/L (Fig, 3). To have a better understanding on the maximum denitrification capacity the biocompounds were more closely investigated in batch experiments.
3.2.
Batch experiments
Batch experiments were conducted to determine the extent to which nitrification and denitrification could occur in the Table 2 – Average volumetric flow rate, loading and removal rate with standard deviation. Parameter Volumetric flow rate COD Ammonium Denitrification
L g g g g
1
day m3 m3 m3 m2
day1 day1 day1 day1
Load
Removal
11.05 4.79 101 52.8 210 120 – –
– 56.6 8.44 200 118 158 103 2.76 1.49a
a Surface specific denitrification rate related to biocompound surface.
Fig. 4 – Dried biocompound carriers after 290 days of cultivation in the cascade. Scale bar: 1 mm.
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stabilized reactor cascade. The experiments were specifically performed so that the individual contributions of the biocompound carriers and sludge in the turnover rates could be identified.
3.2.1.
have a stable and complete nitrogen removal. Throughout this 2-month period the particle size decreased (Fig. 4) which, in addition to the increased maximum denitrification rates observed in the batch experiments, indicates the development of more denitrifying bacteria within the microniches on the carrier surface. It is assumed that the heterotrophic denitrifying bacteria utilized the biocompound material as an electron donor source. To permit a direct comparison between the denitrification capacities of sludge and biofilm, the results of the batch experiments were extrapolated to the reactor cascade. Turnover rates of 0.42 and 0.80 gNO3–N day1 were calculated for sludge without and with glucose, respectively. For biofilm on the biocompound carriers a turnover rate of 9.77 gNO3–N day1 was calculated. This indicates that biofilm was responsible for roughly 92% of denitrification in the reactor cascade.
Nitrification experiments
The findings of the nitrification batch experiments performed on both activated sludge and biocompound carriers are presented in Fig. 5. The maximum nitrification rates realized by the biofilm attached to the carriers ranged between 0.12 and 0.40 gNH4–N m2 day1 (average: 0.26 0.10 gNH4–N m2 day1). Maximum nitrification rates of sludge ranging between 0.12 1 (average: 0.32 0.12 gNH4–N and 0.48 gNH4–N g1 MLSS day 1 1 g MLSS day ) were measured. To allow for a direct comparison of the individual contributions of biofilm and sludge to nitrification, the measured rates from the batch experiments were extrapolated to the reactor cascade. These results correspond to sludge and biofilm turnover rates in the cascade of 28.26 and 0.75 gNH4–N day, respectively. It is apparent that sludge was responsible for roughly 98% of nitrification in the reactor cascade. Furthermore, it can be seen that the nitrification capacity of the sludge was not fully used in the reactor as the sludge load was day1 throughout the entire below 0.1 gNH4–N g1 MLSS experiment.
3.2.2.
3.3.
To track biofilm development on the biocompound carriers, CLSM was performed on a weekly basis from days 135 to 350. Examination of the carriers indicated considerable biofilm coverage as early as 2 weeks after introduction into the reactor cascade. It was confirmed that growth was concentrated in the porous structures (Hille et al., 2009). Biofilm coverage on the outer surfaces of the carriers was sparse due to the high abrasional losses (e.g. particle-to-particle collisions) and shear forces in the cascade. Biofilm thicknesses were often less than 100 mm, even inside the porous structures. Maximum thicknesses of 300 mm were recorded, but only rarely. Fig. 7 gives an impression of the biofilm structure within the biocompound carrier material. The porous structure as well as both bacteria and extracellular polymeric substances (EPS) can clearly be identified. Due to oxygen in the bulk phase, anoxic zones for denitrification can only be provided by such structures which additionally offer organic carbon as an electron donor. It is not clear whether the EPS are used as an electron donor by the denitrifying heterotrophic bacteria (Horn et al., 2001). Zhang and Bishop, 2003 conducted degradation experiments with EPS from mixed culture biofilms. The authors could show that the original microorganisms were able to degrade the EPS when they were starved.
Denitrification experiments
0.4
Sludge Bio compounds
-1
-1
or [gNH4+-N gMLSS d ]
-2
-1
Nitrification rate [gNH4+-N m d ]
The results of the denitrification batch experiments are shown in Fig. 6. For tests conducted both under ideal (DO: <0.9 mg L1) and reactor-like conditions (DO: 1–3 mg L1), no significant denitrification was detected for sludge. On average, maximum denitrification rates with and without addition of 1 glucose of 0.01 and 0.003 gNO3–N g1 MLSS day , respectively, were attained. This indicates that denitrification activity in the cascade can be attributed to the biofilm on the biocompound carriers. Autotrophic microorganisms seemed to dominate the sludge within the reactor cascade. From Fig. 6 it is evident that significant denitrification is achieved by the biocompound carriers. The maximum denitrification rate steadily increased from days 230 to 290. Increasing denitrification rates were not observed in the cascade during this time period (Fig. 6) due to the fact that the ammonium–nitrogen load was steadily decreased (Fig. 2) to
0.5
Anoxic zones within the biofilm structure
0.3 0.2 0.1 0.0 160
180
200
220
240
260
280
300
320
340
Day of cultivation [d] Fig. 5 – Results of batch tests. Maximum nitrification rates of biocompound carriers and sludge related either to the carrier L1 . surface area or amount of MLSS. Oxygen concentration: 1–3 mg LL1; pH: 7–8; initial NHD 4 –N concentration: 100 mg L
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-1
Denitrification rate [gNO3--N m d ]
water research 43 (2009) 4461–4468
4
-1
-1
or [gNO3--N gMLSS d ]
-2
5
Biocompounds Sludge Sludge with glucose
3
DO = 1 – 3 mg L-1
DO = 1 – 3 mg L-1
2 1 0 160
180
200
220
240
260
280
300
320
340
Day of cultivation [d] Fig. 6 – Results of batch tests. Maximum denitrification rates of biocompound carriers and sludge related either to the carrier surface area or amount of MLSS. Oxygen concentration: less than 0.9 mg LL1 (except where otherwise denoted); initial NOL 3 – N concentration: 60 mg LL1. Arrows indicate measured rates of zero.
3.4. Degradation of biocompound material and denitrification To elucidate the origin of the electron donor a estimation on the biocompound carrier consumption has been made for the period between days 220 and 318. It has to be stressed that the measurement of the bulk carrier material at day 318 was not easy as the carrier diameter was reduced (cp. Fig. 4). Approximately 2.1 L of bulk carrier material (from 3.2 L on day 220 to 1.1 L on day 318) was lost and 0.95 kg of nitrogen was eliminated within this period. A total of 2.1 L of bulk carrier material represent 1.3 kg of degradable carrier material. Taking the ratio 60:40 for PHB:PCL, 1 g of dry carrier material accounts for 1.84 g of COD. This gives a highly efficient use of the organic carbon as only a very small amount of nitrate removal (less than 10%) can be attributed to degraded COD (see Section 3.1). For denitrification with methanol in half or full scale biofilters, typically 2.5 g methanol (3.75 g COD) are added for 1 g of nitrate–nitrogen (Horn and Telgmann, 2000). Koch and Siegrist (1997) gave a higher value of 3.2 g methanol per 1 g of nitrate– nitrogen. Due to the situation of limited substrate within the pores, it can be assumed that only very small amounts of biomass are produced and that the main part of the organic carbon is transformed to CO2. An additional option is the direct transformation of nitrite– nitrogen, which should be possible at low oxygen concentrations and a low sludge age (Blackburne et al., 2008). The first condition of low dissolved oxygen was essentially fulfilled in the reactor system, however the sludge age was too high (>10 days) to prevent the development of nitrite oxidizing bacteria (NOB). Perhaps a small fraction of nitrite (mean concentration in the effluent 0.1 mg L1) was directly reduced inside the pore structure. On the other hand, as nitrification was done by the flocs, it is more plausible that ammonium was totally oxidized to nitrate, and the nitrate was then reduced in the pore structure of the biocompound carriers.
3.5.
Possible scale-up for use in wastewater treatment
The presented system is designed to treat wastewater with high ammonium concentrations where nearly no organic carbon is available for denitrification. Reject water with
temperatures of up 20–30 C should be treated with Anammox as there is no organic carbon is necessary (Tho¨le et al., 2005; van der Star et al., 2007). The process presented in this study is advantageous for wastewater with temperatures below 20 C.
Fig. 7 – Example of biofilm growth concentrated in the pore structures of biocompound particles rather than on the external surfaces. (a) Day 378 of particle cut parallel to cylindrical top. (b) Day 350 of whole particle. Green represents EPS, red nucleic acids (e.g. bacteria), and white reflection signal. Scale bar [ 50 mm.
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water research 43 (2009) 4461–4468
Compared to Anammox, the control strategies seemed less complicated as oxygen and biocompound concentration are the main and relevant parameter. The system presented does not suffer from complex population dynamics as the growth of nitrite oxidizing bacteria (NOB) will heavily disturb the Anammox process. In this study it has been shown that very low ammonium and nitrate concentrations can be achieved in the effluent, which is important if the system is used as a final treatment step. Up to now the biodegradable biocompounds (10 V per kg) are too expensive to compete with methanol driven denitrification. Nevertheless, in some cases like treatment of recycling water from sensitive fish farming it may be advantageous to operate a single stage nitrification/denitrification unit without storage of methanol.
4.
Conclusions
A biofilm airlift suspension reactor dually operated with both biofilm-covered biodegradable carriers and activated sludge was developed for the treatment of wastewater containing a low COD:N ratio: The main nitrifying activity can be found in the suspended biomass. Denitrification was mainly located on the carrier material where the organic carbon is provided by the carrier itself as was evident by the carrier degradation. Both processes, denitrification and nitrification, could be stabilized for oxygen concentrations of 0.5–3 mg L1 in the bulk liquid. The amount of carrier material restricted the total removal of nitrogen. An average volumetric denitrification rate of 158 gNO3–N m3 day1 was achieved with 5% bulk carrier volume of biocompounds.
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