- Email: [email protected]
aerobic starvation on aerobic granules
water research 43 (2009) 3622–3632
Available at www.sciencedirect.com
journal homepage: www.elsevier.com/locate/watres
Effect of long term anaerobic and intermittent anaerobic/aerobic starvation on aerobic granules Maite Pijuan, Ursula Werner, Zhiguo Yuan* The University of Queensland, Advanced Water Management Centre (AWMC), Brisbane, QLD 4072, Australia
article info
abstract
Article history:
The effect of long term anaerobic and intermittent anaerobic/aerobic starvation on the
Received 7 January 2009
structure and activity of aerobic granules was studied. Aerobic granular sludge treating
Received in revised form
abattoir wastewater and achieving high levels of nutrient removal was subjected to 4–5 week
7 May 2009
starvation under anaerobic and intermittent anaerobic/aerobic conditions. Microscopic
Accepted 8 May 2009
pictures of granules at the beginning of the starvation period presented a round and compact
Published online 18 May 2009
surface morphology with a much defined external perimeter. Under both starvation conditions, the morphology changed at the end of starvation with the external border of the
Keywords:
granules surrounded by floppy materials. The loss of granular compactness was faster and
Aerobic granules
more pronounced under anaerobic/aerobic starvation conditions. The release of Ca2þ at the
Starvation
onset of anaerobic/aerobic starvation suggests a degradation of extracellular polymeric
Morphology
substances. The activity of ammonia oxidizing bacteria was reduced by 20 and 36% during
Calcium
anaerobic and intermittent anaerobic/aerobic starvation, respectively. When fresh waste-
Nitrification
water was reintroduced, the granules recovered their initial morphology within 1 week of
Enhanced biological
normal operation and the nutrient removal activity recovered fully in 3 weeks. The results
phosphorus removal
show that both anaerobic and intermittent anaerobic/aerobic conditions are suitable for
Recovery
maintaining granule structure and activity during starvation. ª 2009 Elsevier Ltd. All rights reserved.
1.
Introduction
In recent years, aerobic granular sludge has become a promising technology for wastewater treatment. It presents several advantages compared to conventional flocular sludge systems including lower operational costs and lower space requirement (de Bruin et al., 2004). Aerobic granules are defined as aggregates of microbial origin, which do not coagulate under reduced hydrodynamic shear and which subsequently settle significantly faster than activated sludge flocs. Aerobic granules have been mostly cultivated in sequencing batch reactors (SBR) using synthetic wastewater (Tay et al., 2002; Liu et al., 2003; Yang et al., 2003) and most recently using real wastewater (de Villiers and Pretorius, 2001; Arrojo et al., 2004; de Bruin et al., 2004; Yilmaz et al., 2008). Using aerobic granules,
simultaneous removal of organic carbon (denoted as chemical oxygen demand, or COD), nitrogen (N) and phosphorus (P) has been achieved (Cassidy and Belia, 2005; de Kreuk et al., 2005; Yilmaz et al., 2008). The aerobic biogranulation technology will be moving towards industrial use for biological wastewater treatment. Several industry sectors such as the dairy and food processing industries and abattoirs are producers of large volumes of wastewater requiring treatment. The aerobic granulation technology has recently been demonstrated to be suitable for these high-strength wastewaters (Yilmaz et al., 2008). One of the challenges for the treatment of such wastewaters is to cope with the large fluctuations of the wastewater flow linked to the industrial activity. Due to seasonal closure of the industry, the wastewater flow could
* Corresponding author. E-mail address: [email protected] (Z. Yuan). 0043-1354/$ – see front matter ª 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2009.05.007
water research 43 (2009) 3622–3632
even be completely stopped for an extended period of time (weeks). Therefore, the ability of the granules and the biomass to survive under such conditions needs to be investigated in order to achieve reliable biological wastewater treatment. A number of studies have been reported in literature on the effect of starvation on the structure of aerobic granules. Zhu and Wilderer (2003) found that the appearance of granules that developed from a synthetic wastewater (glucose and peptone) did not visually change after a 7 week anaerobic idle phase. As a microbial activity indicator, the oxygen consumption rates were measured, which returned to their pre-starvation levels just 1 week after restarting the reactor. Wang et al. (2008) stored granules treating synthetic wastewater anaerobically for 7 months under low temperatures (in a fridge) and they showed that these granules were able to preserve their granular morphology and recover their nitrification and COD removal ability in approximately 2 weeks. Also, Adav et al. (2007) suggested that storage of aerobic granules at subfreezing temperatures (20 C) was the best way of preserving stability and activity of the granules. However, these preservation conditions are not feasible in large scale systems. In comparison, the impact of starvation on the activities of bacterial community in flocular activated sludge has been investigated more thoroughly. In general, it has been found that anaerobic and intermittent anaerobic/aerobic conditions resulted in the lowest bacteria decay rates, and have therefore been recommended for maintaining bacterial activity in wastewater treatment plants during periods without feed (Siegrist et al., 1999; Lee and Oleszkiewicz, 2003; Yilmaz et al., 2007; Lu et al., 2007). In this study, we investigate the effect of anaerobic and intermittent anaerobic/aerobic starvation conditions on the structure, stability and microbial activity of aerobic granules. Aerobic granular sludge treating real, abattoir wastewater was starved for 4–5 weeks under the above conditions. The external morphology of the granules was monitored using light and scanning electron microscope (SEM). The changes in O2, Ca2þ and pH distribution inside the granules were monitored using microsensors. The activities of ammonia oxidizing bacteria (AOB) and polyphosphate accumulating organisms (PAO) were monitored through batch tests conducted weekly. The recovery of morphology and nutrient removal performance of the granules were examined after fresh wastewater was reintroduced.
2.
Materials and methods
2.1.
Source of granules
The granules used were sourced from a SBR, called the parent reactor, which performed biological COD, nitrogen and phosphorus removal from abattoir wastewater containing approximately 250 mg N/L and 40 mg P/L of total nitrogen and total phosphorus, respectively (Yilmaz et al., 2008). The SBR was operated with a cycle time of 8 h, with 3 L of abattoir wastewater fed in each cycle at the beginning of its 78 min anaerobic period reaching a total working volume of 5 L. Nitrogen and phosphorus were removed in the subsequent
3623
aerobic period, primarily through simultaneous nitrification, denitrification and phosphorus uptake. More details of the operation of the SBR can be found in Yilmaz et al. (2008).
2.2.
Anaerobic/aerobic starvation
One litre of granular sludge was withdrawn from the parent reactor. At the point of sludge sampling, the parent SBR had a mixed liquor suspended solids (MLSS) concentration of 14 g/ L and a mixed liquor volatile suspended solids (MLVSS) concentration of 10 g/L, respectively. It was achieving >99% removal of N and P from an abattoir wastewater (0.4 mg N– 3 in the effluent). The sludge was NHþ 4 and 0.5 mg P–PO4 diluted with 1 L of effluent from the same reactor resulting in an initial MLSS concentration of approximately 7 g/L, and placed to another SBR, called the anaerobically/aerobicallystarved SBR, for 28 days without feed, This SBR was aerated for 15 min in every 6 h. Granules were allowed to settle during the remaining time.
2.3.
Anaerobic starvation
A total of 0.5 L of granular sludge was withdrawn from the parent SBR 1 month after the granules used for the anaerobic/ aerobic starvation was withdrawn to allow the parent SBR to recover from the biomass loss. At the time of sampling, the parent SBR had a MLSS concentration of 16 g/L and a MLVSS concentration of 13.8 g/L and it was achieving >99% of N removal and 53% of P removal (0.7 mg N–NHþ 4 and 16.4 mg P– in the effluent). The sludge was diluted with 1.5 L of PO3 4 effluent from the same reactor providing an initial MLSS concentration of 4 g/L. These 2 L were kept in a batch reactor (called anaerobically-starved SBR) in settling conditions for a period of 32 days. During this period no feed was provided.
2.4.
Monitoring during starvation
During both starvation experiments, liquid phase samples were taken every two days. Samples were taken using a syringe after a short period of mixing, and immediately filtered through disposable Millipore filter units (0.22 mm pore size) for analyses of ammonia, nitrite, phosphate, calcium, magnesium and potassium ions. Sludge samples for granule size distribution and light microscopic analysis was taken weekly, whereas sludge samples for SEM were taken at the beginning and at the end of the starvation period. Granules were removed from the starvation reactors weekly for the measurements of the micro-scale oxygen concentration profiles inside granules. Granules were also taken on days 0, 1 and 6 for the measurements of Ca2þ and pH concentration profiles during both starvation conditions and additionally on day 3 of the anaerobic/aerobic starvation. More details of these measurements will be described later. Batch tests for the assessment of nitrification and P removal activity were performed weekly. Fluorescent in situ hybridization (FISH) was performed on biomass samples taken at the beginning and end of each starvation experiment.
3624
2.5.
water research 43 (2009) 3622–3632
Batch experiments
During both starvation conditions, batch experiments were carried out to monitor the nitrification activity of the biomass by measuring the ammonium oxidation rate over a period of 2 h after a pulse addition of ammonia. In each test, 100 mL of mixed liquor was taken from the starved SBR. After washing the sludge with effluent from the parent SBR (which had very low concentrations of N, P and biodegradable COD), a pulse of a concentrated ammonium solution (2 g N–NHþ 4 /L) was added to the batch reactor, resulting in an initial ammonium concentration in the reactor of 20 mg/L N–NHþ 4 . Mixed liquor samples were taken and filtered every 10 min for analysis of NHþ 4 and NO 2 . The pH was manually controlled in the range 7.4–7.6 through the addition of 0.1 M of NaOH. The ammonia oxidation rate was determined as the slope of the ammonium profile. The decay rate of ammonia oxidizing bacteria (AOB) was determined through fitting an exponential function to the measured ammonia oxidation rates as a function of time (Eq. (1)): y ¼ A ekt
(1)
where y is the biomass specific ammonia oxidation rate (mg N–NHþ 4 /g VSS*h); A is initial biomass specific ammonia oxidation rate (mg N–NHþ 4 /g VSS*h), t is the starvation time (d) and k is the decay rate (d1). The PAO P-uptake activity was also monitored through weekly batch tests carried out in a 100 mL reactor during the intermittent anaerobic/aerobic starvation period. At the start of the experiment, 100 mL granular sludge were withdrawn from the starved SBR at the end of the aeration phase and washed þ with effluent from the parent SBR. A solution of P–PO3 4 and NH4 was added to the batch reactor resulting in initial concentraþ tions of 20 mg P–PO3 4 /L and 10 mg N–NH4 /L/L, respectively. Acetate was added to reach a concentration of 50 mg COD/L. Allylthiourea (ATU) was also added at a concentration of 20 mg/ L to avoid nitrification. During the first hour, nitrogen was sparged to maintain strict anaerobic conditions. After an anaerobic period, the 100 mL reactor was sparged with air for 1 h. Dissolved oxygen was not controlled and was higher than 2.5 mg/L during the entire aerobic period. Samples were taken every 10 min for analysis of acetate and P. The pH was manually controlled in the range 7.4–7.6 through addition of 0.1 M HCl and 0.4 M NaOH.
2.6.
Reactor operation in the recovery period
After 28 days of anaerobic/aerobic starvation, feeding of wastewater to the SBR was started. The 6 h cycle consisted of 4 min feeding (1 L), followed by 1 h anaerobic, 294 min aerobic (DO was controlled between 1.5 and 2.0 mg/L), 1 min settling and 1 min decant periods. The recovery of nitrogen removal was monitored for 1 month through weekly batch experiments as described before. The recovery of phosphorus removal was monitored every 2–3 days for a period of 9 days through batch experiments as described above.
2.7.
Analytical procedures
The ammonium (NHþ 4 –N), nitrate (NO3 –N), nitrite (NO2 –N) and 3 orthophosphate (PO4 –P) concentrations were analyzed using
a Lachat QuikChem8000 Flow Injection Analyzer (Lachat Instrument, Milwaukee). MLSS and MLVSS were analysed according to the standard methods (APHA, 1995). Potassium (Kþ) and magnesium (Mg2þ) ions were measured by Inductively Coupled Plasma – Atomic Emission Spectrometry (ICP–AES Varian Vista- PRO, Varian, Inc.). To determine the size distribution of the granules, 30 mL of well mixed granular sludge were pumped through a Malvern laser light scattering instrument, Mastersizer 2000 series (Malvern Instruments, Worcestershire, UK). The technique of laser diffraction is based on the principle that particles passing through a laser beam will scatter light at an angle that is directly related to their size. This method represents a rapid and robust measurement of particles present in a bulk with a range of 0.02–2000 mm. Microscopic observation of granules confirmed that the absolute majority of the granules in the reactor have sizes in this range (data not presented). Oxygen concentration profiles in the granules were measured using Clark type microelectrodes (Revsbech, 1989). The sensors had an actual sensing surface of 5 mm in diameter and a 90% response time of less than 5 s. The measurements were conducted in a small flow cell with filtered wastewater (GF/A, Whatman) horizontally flowing through the cell (volumetric displacement rate: 25 mL/min). The wastewater used was taken from the parent reactor at the end of the aerobic phase, thus, it had very low levels of ammonium and no detectable levels of readily biodegradable COD. To allow for a comparison of oxygen penetration depths in the granules during the time course of the starvation, the oxygen concentration in the flow cell was adjusted to 3–3.5 mg/L, similar to the oxygen concentration in the parent reactor at the end of the aerobic phase. The oxygen penetration depth was visually determined from replicate profiles. The Ca2þ and pH concentration profiles in the granules were measured using liquid ion-exchange (LIX) microsensors as described in de Beer and Kuhl et al. (2000). These sensors had a 90% response time of less than 5 s and a sensing surface diameter of ca. 5 mm. For these measurements a larger flow cell was used, with a volumetric displacement rate of 55 mL min1. During the anaerobic/aerobic starvation conditions, granules and supernatant were taken directly from the anaerobically/ aerobically-starved reactor. The measurements during anaerobic starvation were conducted on sludge kept in a separate anaerobic reactor to keep disturbances in the main experiment at a minimum. During the microsensor measurements, the pH was stabilized at a level measured in the starvation reactors on the respective day. Oxygen concentrations in the flow cell were adjusted to 0.5–0.7 mg/L for both starvation conditions to achieve similar conditions during the measurements of pH and Ca2þ distribution. This oxygen concentration can be found in the parent reactor at the beginning of the aerobic period and led to an oxygen penetration of less than 50 mm. For all microsensor measurements 25–50 mL of sample were taken from the reactors and up to 10 granules were randomly selected for profiling. Four to ten replicate microsensor profiles were measured with step sizes of 25–100 mm in different granules. The granule morphology was qualitatively observed using a stereo microscope (Olympus SZH10) and a scanning electron microscope (SEM, Jeol JSM-6300F).
3625
water research 43 (2009) 3622–3632
2.8.
Microbial analyses
Sludge samples were fixed and FISH probed as previously described by Amann (1995). Oligonucleotide probes used in this study were EUBmix (Daims et al., 1999) for all bacteria, PAOmix (for Candidatus Accumulibacter phosphatis or Accumulibacter, comprising equal amounts of probes PAO462, PAO651 and PAO846, Crocetti et al., 2000) and GAOmix (for Candidatus Competibacter phosphatis or Competibacter, comprising equal amounts of probes GAOQ431 and GAOQ989, Crocetti et al., 2002). FISH images were collected using a Zeiss LSM 510 Meta Confocal microscope with a 63 Plan-Apochromat oil immersion lens. FISH quantification was performed according to Zhou et al. (2007) where the relative abundance of each group was determined as mean percentage of all bacteria.
3.
Results and discussion
3.1.
Impact of starvation on granular structure
Two different starvation conditions were tested in this study, the continuous anaerobic starvation and the intermittent anaerobic/aerobic starvation. Fig. 1 shows the size distribution profiles of the granules under the two starvation conditions. The initial granule sizes in both studies were between 0.6 and 1.6 mm. The 50 and 90 percentiles remained relatively constant during the 30 days of starvation. This indicates that under both starvation regimes the granular structure was not lost. However, the 10 percentile decreased from 0.6 to 0.05 mm during the first week of anaerobic/aerobic starvation (Fig. 1 left) indicating the release of small particles from granules of all sizes in the reactor. The 10 percentile indicates that 10% of the particles present in the reactor have a diameter that is lower or equal to this value. In contrast, under anaerobic starvation (Fig. 1 right) the 10 percentile started to decrease after 2 weeks and low values similar to those observed in the anaerobic/aerobic starvation after week 1 were reached only at the end of the starvation period.
To monitor changes in the morphology and structure, granules were studied weekly using a light microscope (Fig. 2) and at the beginning and at the end of the starvation period using a SEM (Fig. 3). Granules from the beginning of the starvation period possessed a compact structure with a very well defined external perimeter (Figs. 2A and 3A,E). In contrast, granules after 1 month of anaerobically/aerobically (Fig. 2B) and anaerobically-starved conditions (Fig. 2C), had a layer of fluffy materials at the edges of the granules. They still preserved their structural integrity, but became more transparent in the edges compared to the fresh granules. The appearance of fluffy materials surrounding the granules could be responsible for considerably reduced ten percentile size values observed in both cases. Similar morphology of starved granules was also observed in the parent SBR in several occasions when the wastewater supply was cut for a few days, and the SBR was maintained under alternating anaerobic/ aerobic starvation conditions (data not shown). On the other hand, when the waste sludge from this SBR was stored in a bottle (anaerobic conditions) this fluffy material was not observed. The SEM pictures revealed that granules at the end of the anaerobic/aerobic starvation (Figs. 3B,C) differed substantially in morphology from the ones at the beginning of the starvation, with rough surfaces and irregular shapes. It seems that parts of the granules were consumed or broken off. In contrast, granules kept under anaerobic starvation conditions did not change significantly after 20 days of starvation except for the appearance of significant cracks from the edge to the interior of the granules in some of the granules (Fig. 3F). Keeping granules under long term aerobic starvation conditions, Wang et al. (2005) observed that after 20 days the central part of the granules was becoming more transparent and in some cases the core part became hollow, while the outer shell still remained intact. They attributed this phenomenon to the bacterial consumption of part of the extracellular polymeric substances (EPS) present in the granules. Thus, its own producers would consume the EPS when substrate is severely limited. Conducting a closer investigation on the biodegradability of aerobic granular EPS under aerobic starvation conditions, Wang et al. (2007) quantified that about 50% of EPS produced by aerobic granules could be utilized by their producers and the rest would remain intact to
1800
Granule size (um)
1600 1400 1200 1000 800 600
d(0.1) d(0.5) d(0.9)
400 200 0
0
5
10
15
Time (days)
20
25
0
5
10
15
20
25
30
35
Time (days)
Fig. 1 – Distribution of granule size in the reactor during the entire study. d(0.1), d(0.5), d(0.9) are the 10, 50 and 90 percentiles of the distribution, respectively. Left – anaerobic/aerobic starvation; right – anaerobic starvation.
3626
water research 43 (2009) 3622–3632
Fig. 2 – Light microscope images of the aerobic granules before starvation (A) and 1 month after anaerobic/aerobic (B) and anaerobic starvation (C), showing the change in morphology as a result.
Fig. 3 – SEM images of the outer morphology of granules. Top-anaerobic/aerobic starvation: before (A) and 26 days after (B,C) starvation and (D) after 1 week of recovery. Bottom-anaerobic starvation: before (E) and 20 days after (F, G) starvation.
preserve the structural integrity of the granule. This could explain the changes observed mostly in the anaerobic/aerobic granules from our study, where a significant degradation of the external parts of the granules was detected. The degradation of the external EPS layer in our study may be linked to the distribution of specific EPS components, which can differ depending on the granular type (Chen et al., 2007). Adav et al. (2008) found that the granular structure was most affected by the degradation of b-polysaccharides in phenol fed granules. In another study, the degradation of proteins and a-polysaccharides had most effects on the structural morphology of granules withdrawn from the parent reactor used in our experiment (Seviour et al., 2009). However, a detailed investigation of the distribution of EPS components was not conducted in our study. In comparison, anaerobic starvation conditions had a milder effect on the outer morphology of the granules, compared to the intermittent anaerobic/aerobic
conditions, possibly because the degradation of EPS under strictly anaerobic conditions may be more difficult. Adav et al. (2009) anaerobically stored aerobic granules for 60 days at 8 C in a bottle. They hypothesized that the proteins present in the EPS were hydrolysed and then used by the nearby anaerobic strains, gradually digesting the granules from the inside out as a mechanism of survival. However, even after the anaerobic/aerobic starvation period, the recovery of the initial granular morphology was relatively quick and the granules had a similar shape to the initial ones, just 1 week after the feed was reintroduced (Fig. 3D).
3.1.1.
Effect of starvation on oxygen penetration
In diffusion dominated systems, the penetration of oxygen into microbial aggregates such as granules is determined by the oxygen concentration in the bulk water, the oxygen
3627
water research 43 (2009) 3622–3632
Table 1 – Oxygen penetration depths in granules under anaerobic/aerobic and under anaerobic starvation conditions. O2 penetration (mm) Anaerobic/aerobic starvation Anaerobic starvation
Day 0
Day 7
Day 14
Day 21
Day 28
238.6 42.4 185.7 96.7
250.7 1.2 n.a.
366.7 70.1 157.1 27.8
n.a. 307.5 186.7
350 147.2 356.3 85.6
n.a.: not available. Data are given as average standard deviation.
consumption rate inside the aggregates, the flow rate of the medium and the diffusivity of the matrix (Jørgensen and Marais, 1990; Chiu et al., 2007; Li et al., 2008). Previous research showed that aerobic granules with higher extracellular polymer content had a lower oxygen diffusivity than granules with a lower extracellular polymer content (Chiu et al., 2006). Assuming a degradation of EPS and a decrease in oxygen consumption during both starvation conditions, we expected to see an increase of oxygen penetration depths (under comparable oxygen concentrations and comparable flow conditions in the bulk water) over the time course of the experiments. The depths of oxygen penetration during both starvation periods are summarised in Table 1, while some examples of the measured oxygen profiles are shown in Fig. 4. A diffusive boundary layer above the surface of the granules was clearly visible in all oxygen profiles and all profiles had the parabolic shape of diffusion dominated consumption
profiles, indicating the importance of diffusive transport to deeper layers and the uptake of oxygen inside the granules. Occasionally, increased oxygen concentrations in deeper parts of the profiles indicated that oxygen was also transported through channels and holes into the granules, either by diffusive or advective transport (e.g. Fig. 4d). An increased oxygen penetration depth was measured in the anaerobic/aerobic starving granules from the second week on (Figs. 4a,b, Table 1), whereas during the anaerobic starvation, oxygen penetration increased only from the third week on (Figs. 4c,d; Table 1). At the end of the experiments, the oxygen penetration depths were similar for both starvation conditions (Table 1). The increase in oxygen penetration depths is likely due to a combination of decreased biological oxygen consumption (e.g., heterotrophic respiration using the intrinsic carbon pool and nitrification, to be further discussed below) and increased diffusivity of the EPS matrix resulting from its degradation
-600
b
a -400 -200 0
depth below surface (µm)
200 400 600 0
1
2
3
4
0
1
2
3
4
1
2
3
4
-600
d
c -400 -200 0 200 400 600 0
1
2
3
4
0
O2 concentration (mg/L) Fig. 4 – Oxygen profiles in granules on day 0 (a) and after 14 days of anaerobic/aerobic starvation (b) and on day 0 (c) and after 21 days of anaerobic starvation (d). The line through zero represents the granule surface. The dashed line indicates the approximate height of the diffusive boundary layer. The different symbols represent replicate profiles.
3628
water research 43 (2009) 3622–3632
stability was attributed to the role of Ca2þ in the EPS structure. To assess a possible effect of EPS degradation during starvation conditions on Ca2þ concentrations and distribution patterns, Ca2þ and pH profiles were measured within the granules. Ca2þ and pH microsensor profiles in the granules were measured on days 0, 1 and 6 during both starvation conditions and additionally on day 3 during the anaerobic/ aerobic starvation conditions. The Ca2þ profiles obtained in anaerobic/aerobic starving granules are represented in Fig. 5. On the first day of anaerobic/aerobic starvation a substantial increase of Ca2þ was observed below 100 mm depth, indicating the considerable production of Ca2þ in the outer layer of the granule. On day 3, the Ca2þ concentrations in the granule were already much lower, while they still remained elevated with depth, when compared to the measurements of day 0. The increased concentrations in the first 100 mm indicate the on-going production of Ca2þ in this zone. On day 6, the Ca2þ concentrations profiles showed only minor variations with depth with some profiles showing small production and others showing small consumption zones, the profiles were
during starvation. The respective contribution of both factors in the determination of the oxygen penetration depths can not be deduced from our data set. However, the effects of starvation on the EPS matrix may explain the earlier increase of the oxygen penetration depth during anaerobic/aerobic starvation conditions as compared to the anaerobic starvation conditions (Table 1). The earlier increase of the averaged oxygen penetration depth during anaerobic/aerobic conditions corroborates the observations of more severe granular structure loss (hence, an increased impact on the matrix diffusivity) under this starvation condition.
3.1.2.
Effect of starvation on Ca2þ concentrations and pH
2þ
Ca is an important bridging agent in the EPS and has been recognized to play an important role in the self-immobilization of microbial aggregates (Morgan et al., 1990; Yu et al., 2001). Ren et al. (2008) found that Ca2þ rich aerobic granules had a more rigid structure and a higher shear strength as compared to granules that were grown in water containing lower concentrations of Ca2þ. The enhanced structural
-500
b
a
0
depth below surface (µm)
500
1000 day 1
day 0
1500 6
8
10
12
14
16
20 0
18
-500
100
50
150
200
250
350
300
d
c
0
500
1000 day 3
day 6
1500 6
8
10
12
14
16
18
20
6
8
10
12
14
16
18
20
Ca2+ (mg/L) Fig. 5 – Ca2D concentration profiles in the intermittent anaerobic–aerobically starved granules, (a, b, c, d; days 0, 1, 3 and 6 of starvation, respectively). The zero line represents the granule surface. The different symbols represent replicate profiles. Note the different scale of the x axis in plot b.
3629
water research 43 (2009) 3622–3632
thus comparable to those measured on day 0. The profiles from day 6 demonstrated that large amounts of Ca2þ were no longer produced. The increased concentrations of Ca2þ on days 1 and 3 may originate from a degradation of EPS or may come from a dissolution of CaCO3 precipitates. Dissolution of CaCO3 is expected to cause an increase in the pH at the respective depth layer. However, pH profiles, measured concurrently to the Ca2þ profiles, on day 1 of anaerobic/aerobic starvation (corresponding to the day when the highest Ca2þ production was observed inside the granules) were irregular with depth, some profiles showed small pH decreases and others small pH increases. Additionally, in contrast to the Ca2þ profiles, the pH profiles did not change over the time course of the first week of starvation. These observations support the hypothesis that Ca2þ originated from the biodegradation of the EPS rather than dissolution of CaCO3. Our hypothesis that the EPS degradation, which presumably caused the observed major Ca2þ release during the first 3 days of anaerobic/aerobic starvation corresponds to observations on EPS degradation by Wang et al. (2007). They suggested that the degradation of EPS during aerobic starvation was happening at a faster rate during the first 4 days of their study and slowed down afterwards. Interestingly, no Ca2þ concentration gradients were observed in the profiles measured in the experiments during anaerobic starvation conditions. During all measurements, Ca2þ concentration profiles showed only minor variations with depths and no major Ca2þ production zones were observed. Also the pH profile remained comparable during all measurements, showing only minor variations and irregular trends with depths. The pH profiles measured during anaerobic starvation resembled the pH profiles measured during anaerobic/aerobic starvation conditions. These observations support our hypothesis that the EPS of the granules degraded less in the beginning of the anaerobic starvation conditions. However, the remarkable magnitude of the Ca2þ production observed on day 1 of anaerobic/aerobic starvation indicates that further investigations are necessary to clarify the role of Ca2þ during EPS degradation.
3.2.
Impact of starvation on nutrient removal
3.2.1.
Monitoring nitrification activity
Weekly batch tests were performed with sludge withdrawn from both starved SBRs in order to study the effect of starvation conditions on the activity of nitrifiers.
The nitrification tests were focused on the ammonia oxidizing bacteria (AOB) population, because no nitrate was produced in the parent SBR and the population of nitrite oxidizing bacteria (NOB) was negligible (Yilmaz et al., 2008). Fig. 6 shows the profiles of NHþ 4 oxidation rate measured weekly during both starvation periods. These rates were calculated from the NHþ 4 profiles measured during batch tests as explained previously. The AOB decay rate calculated from the NHþ 4 oxidation rate in the anaerobic/aerobic starvation was 0.0182 d1. This number is very similar to the one obtained by Yilmaz et al. (2007) (0.0165 d1) with a floccular sludge treating abattoir wastewater. The AOB decay rate under anaerobic conditions was determined to be 0.0069 d1. The results show that both conditions are suitable for preserving AOB activities. In comparison, anaerobic starvation seems to further lower the AOB decay rate. The activity of AOB was reduced by 20 and 36% during anaerobic and intermittent anaerobic/aerobic starvation, respectively. These results are in agreement with findings by other researchers (Siegrist et al., 1999).
3.2.2.
3.5
3.5 y = 2.991e-0.0182t R2 = 0.9876
2.5
y = 2.0754e-0.0069t
3.0
mg N/ gVSS h
mg N/ gVSS h
3.0
2.0 1.5 1.0
R2 = 0.799
2.5 2.0 1.5 1.0 0.5
0.5 0.0
PAO activity during starvation
Phosphorus concentrations were periodically monitored in the starved SBRs. Fig. 7 shows the P profiles in the SBR during both starvation periods. P-release increased almost linearly initially but slowed down after around 8 and 7 days in the intermittent anaerobic/aerobic and anaerobic starvation respectively. This behaviour is somewhat different from observations in an earlier study by Yilmaz et al. (2007), where a floccular sludge was submitted to the same anaerobic/ aerobic starvation conditions as used in this study. In that study, no accumulation of P was detected during the first 1– 2 weeks of starvation. The concentration profiles of Kþ and Mg2þ displayed similar trends, with P:Kþ:Mg2þ molar ratios of 3.11:1.05:1 for the anaerobic/aerobic starvation and 3.10:0.92:1 for the anaerobic starvation during the first week of the study, suggesting that phosphate release likely originated from the breakdown of polyphosphate with a theoretical structure of Mg1/3K1/3PO3 (Smolders et al., 1994). This is a confirmation that PAO use their internal polyphosphate as a source of energy during these starvation conditions. The amount of phosphorus, potassium and magnesium release was much higher in the anaerobic/aerobic starvation period (Fig. 7). This may be related to the different starvation conditions applied, but probably is also affected by the changed
0
5
10
15
Time (days)
20
25
30
0.0
0
5
10
15
20
25
30
35
Time (days)
Fig. 6 – Variation of NHD 4 oxidation rate over the anaerobic/aerobic (left) and the anaerobic starvation period (right).
3630
water research 43 (2009) 3622–3632
500
40 30
300
mg/L
mg/L
400
200
10
100 0
20
0
5
10
15
20
25
0
30
0
5
10
Time (days)
15
20
25
30
35
Time (days)
Fig. 7 – Phosphorus (C), potassium (B) and magnesium (;) profiles in the SBR starved under anaerobic/aerobic (left) and under anaerobic conditions (right).
bacterial populations in the sludge. There was a significant increase of the GAO population in the parent reactor during the period between the two starvation studies. FISH analysis showed that the Accumulibacter-PAO decreased from 29% of the entire bacterial population at the beginning of the anaerobic/ aerobic starvation to 21% at the end of the starvation while Competibacter-GAO was present at around 9% and maintained stable. The sludge used for the anaerobic starvation study had 28% of Accumulibacter-PAO and 35% Competibacter-GAO at the beginning of the study and these percentages did not significantly change along the anaerobic starvation period.
3.3.
Recovery after starvation
Granule structure rapidly recovered after the reintroduction of the wastewater (Fig. 3) and the small particles present in the reactor were washed out during the first day of reactor operation due to the short settling time applied. The granule size distribution 1 day into the recovery period became almost identical to the distribution measured at the beginning of the starvation tests. In order to study the ability of the granules to recover their activity in terms of nutrient removal, recovery tests were performed on the anaerobically/aerobically-starved granules. Figs. 8A,B show the recovery of the nitrification rate and the P-
3.4.
Limitations of the study and future research
This study focused on the changes of morphological structure and microbial activity of aerobic granules subjected to two different starvation conditions. The study demonstrated that both anaerobic and intermittent anaerobic/aerobic strategies are effective in maintaining granular structure and microbial activity of aerobic granules. Both strategies are easily implementable in practice. Microsensor measurements linked with microscopic observations also contributed to the understanding of the survival mechanisms of aerobic granules in starvation. However, the study has several limitations. First of all, the study was performed with only one type of granules developed with abattoir wastewater. The experimental findings should be verified with other types of aerobic granules treating different types of wastewaters. Secondly, the micro-scale studies remain mostly observational and descriptive. Fundamental understanding of the physical and biochemical processes occurring in aerobic granules during starvation
B 3.5
16
3.0
14
2.5
mg P/ gVSS h
mg N/ gVSS h
A
uptake, respectively. The nitrification rate increased from the beginning of the recovery period, but it took 3 weeks for the nitrification rate to fully recover. In comparison, the P-uptake rate, as a measure of the PAO activity, recovered 10 days into the recovery period.
2.0 1.5 1.0 STARVATION PERIOD
0.5 0.0
0
5
RECOVERY PERIOD
10 15 20 25 30 35 40 45 50
Time (days)
STARVATION PERIOD
RECOVERY
12 10 8 6 4 2 0
0
5
10
15
20
25
30
35
Time (days)
Fig. 8 – Recovery of the NHD 4 oxidation rate (A) and the P-uptake rate (B) in the SBR starved under intermittent anaerobicaerobic conditions.
water research 43 (2009) 3622–3632
periods is still missing, and requires further research. Thirdly, the sludge samples used for the anaerobic and the intermittent aerobic/anaerobic starvation tests were taken from the parent reactor at different times and had differences in microbial structures. Therefore, caution has to be taken when comparing the results from the two types of tests.
4.
Conclusions
The effect of a long starvation period (approximately 1 month) under anaerobic and intermittent anaerobic/aerobic conditions on the structure and microbial activity of aerobic granules was studied. The following conclusions are drawn: Aerobic granules can be stored under these conditions for at least 4 weeks with no wastewater supply without loosing their structural integrity. Although the results are not directly comparable since the starting biomass was different, better results were obtained under anaerobic starvation conditions in terms of maintaining the surface structure and microbial activity of granules. The intermittent availability of oxygen likely enhanced the consumption of EPS around the edge of granules, causing rougher surface structures. Full recovery of the external morphology of aerobic granules can be achieved within 1 week of resuming the wastewater addition. The decay rates of ammonia oxidizing bacteria found in aerobic granules are very similar to the ones found in floccular sludge.
Acknowledgement This work was funded by the Environmental Biotechnology Cooperative Research Centre (EBCRC) Pty Ltd, Australia. The authors wish to thank Dr. Romain Lemaire and Mr. Paritam Dutta from the Advanced Water Management Centre for their help on the SEM analysis.
references
Adav, S.S., Lee, D.-J., Tay, J.H., 2007. Activity and structure of stored aerobic granules. Environmental Technol 28, 1227–1235. Adav, S.S., Lee, D.J., Tay, J.H., 2008. Extrapolymeric substances and structural stability of aerobic granules. Water Res. 42, 1644–1650. Adav, S.S., Lee, D.-J., Lai, J.-Y., 2009. Proteolytic activity in stored aerobic granular sludge and structural integrity. Bioresource Technol 100, 68–73. Amann, R.I., 1995. Situ Identification of Microorganisms by Whole Cell Hybridization with rRNA-targeted Nucleic Acid Probes. In: Molecular Microbial Ecology Manual. Kluwer Academic Publications, Dordrecht, Holland, pp. 3.3.6, 1-15. APHA, 1995. Standard Methods for the Examination of Water and Wastewater, nineteenth ed. American Public Health Association, Washington, DC.
3631
Arrojo, B., Mosquera-Corral, A., Garrido, J.M., Mendez, R., 2004. Aerobic granulation with industrial wastewater in sequencing batch reactors. Water Res. 38 (14–15), 3389–3399. Cassidy, D.P., Belia, E., 2005. Nitrogen and phosphorus removal from abattoir wastewater in a SBR with aerobic granular sludge. Water Res. 39 (19), 4817–4823. Chen, M.Y., Lee, D.J., Tay, J.H., 2007. Distribution of extracellular polymeric substances in aerobic granules. Appl. Microbiol. Biotechnol 73, 1463–1469. Chiu, Z.C., Chen, M.Y., Lee, D.J., Tay, S.T.L., Tay, J.H., Show, K.Y., 2006. Diffusivity of oxygen in aerobic granules. Biotechnol. Bioeng 94, 505–513. Chiu, Z.C., Chen, M.Y., Lee, D.J., Wang, C.H., Lai, J.Y., 2007. Oxygen diffusion in active layer of aerobic granule with step change in surrounding oxygen levels. Water Res. 41 (4), 884–892. Crocetti, G.R., Hugenholtz, P., Bond, P.L., Schuler, A., Keller, J., Jenkins, D., Blackall, L.L., 2000. Identification of polyphosphate accumulating organisms and design of 16S rRNA-directed probes for their detection and quantitation. Appl. Environ. Microbiol. 66, 1175–1182. Crocetti, G.R., Banfield, J.F., Keller, J., Bond, P.L., Blackall, L.L., 2002. Glycogen accumulating organisms in laboratory-scale and full-scale activated sludge process. Microbiology 148, 3353–3364. Daims, H., Bruhl, A., Amann, R., Schleifer, K.H., Wagner, M., 1999. The domain-specific probe EUB338 is insufficient for the detection of all bacteria: development and evaluation of a more comprehensive probe set. Syst. Appl. Microbiol. 22, 434–444. de Beer, D., Kuhl, M., et al., 2000. A microsensor study of light enhanced Ca2þ uptake and photosynthesis in the reefbuilding hermatypic coral Favia sp. Marine Ecol.-Progr. Ser 194, 75–85. de Bruin, L.M.M., de Kreuk, M.K., van der Roest, H.F.R., Uijterlinde, C., van Loosdrecht, M.C.M., 2004. Aerobic granular sludge technology: an alternative to activated sludge? Water Sci. Technol 49 (11–12), 1–7. de Kreuk, M., Heijnen, J.J., van Loosdrecht, M.C.M., 2005. Simultaneous COD, nitrogen and phosphate removal by aerobic granular sludge. Biotechnol. Bioeng 90 (6), 761–769. de Villiers, G.H., Pretorius, W.A., 2001. Abattoir effluent treatment and protein production. Water Sci. Technol 43 (11), 243–250. Jørgensen, B.B., Marais, D.J.D., 1990. The diffusive boundary-layer of sediments – oxygen microgradients over a microbial mat. Limnol. Oceanogr 35, 1343–1355. Lee, Y., Oleszkiewicz, J.A., 2003. Effects of predation and ORP conditions on the performance of nitrifiers in activated sludge systems. Water Res. 37 (17), 4202–4210. Li, Y., Liu, Y., Shen, L., Chen, F., 2008. DO diffusion profile in aerobic granule and its microbiological implications. Enzyme Microb. Technol 43, 349–354. Liu, Y., Yang, S.F., Tay, J.H., 2003. Elemental compositions and characteristics of aerobic granules cultivated at different substrate N/C ratios. Appl. Microbiol. Biotechnol 61 (5–6), 556–561. Lu, H., Keller, J., Yuan, Z., 2007. Endogenous metabolism of Candidatus Accumulibacter phosphatis under various starvation conditions. Water Res. 41, 4646–4656. Morgan, J.W., Forster, C.F., Evison, L., 1990. Comparative study of the nature of biopolymers extracted from anaerobic and activated sludges. Water Res. 31, 3191–3194. Ren, T.T., Liu, L., Sheng, G.P., Liu, X.W., Yu, H.Q., Zhang, M.C., Zhu, J.R., 2008. Calcium spatial distribution in aerobic granules and its effects on granule structure, strength and bioactivity. Water Res. 42, 3343–3352. Revsbech, N.P., 1989. An oxygen microsensor with a guard cathode. Limnol. Oceanogr 34, 474–478. Seviour, T., Pijuan, M., Nicholson, T., Keller, J., Yuan, Z.G., 2009. Understanding the properties of aerobic sludge granules as hydrogels. Biotechnol. Bioeng 102, 1483–1493.
3632
water research 43 (2009) 3622–3632
Siegrist, H., Brunner, I., Koch, G., Phan, L.C., Le, V.C., 1999. Reduction of biomass decay rate under anoxic and anaerobic conditions. Water Sci. Technol 39 (1), 129–137. Smolders, G.J.F., Van der Meij, J., van Loosdrecht, M.C.M., Heijnen, J.J., 1994. Model of the anaerobic metabolism of the biological phosphorus removal process: stoichiometry and pH influence. Biotechnol. Bioeng 43 (6), 461–470. Tay, J.H., Liu, Q.S., Liu, Y., 2002. Characteristics of aerobic granules grown on glucose and acetate in sequential aerobic sludge blanket reactors. Environ. Technol 23 (8), 931–936. Wang, Z.-W., Liu, Y., Tay, J.-H., 2005. Distribution of EPS and cell surface hydrophobicity in aerobic granules. Appl. Microbiol. Biotechnol 69, 469–473. Wang, Z.-W., Liu, Y., Tay, J.-H., 2007. Biodegradability of extracellular polymeric substances produced by aerobic granules. Appl. Microbiol. Biotechnol 74, 462–466. Wang, X., Zhang, H., Yang, F., Wang, Y., Gao, M., 2008. Long-term storage and subsequent reactivation of aerobic granules. Bioresource Technol. 99 (13), 8304–8309.
Yang, S.F., Tay, J.H., Liu, Y., 2003. A novel granular sludge sequencing batch reactor for removal of organic and nitrogen from wastewater. J. Biotechnol 106 (1), 77–86. Yilmaz, G., Lemaire, R., Keller, J., Yuan, Z., 2007. Effectiveness of an intermittent aerobic, anoxic/anaerobic strategy for maintaining biomass activity of BNR sludge during long-term starvation. Water Res. 41 (12), 2590–2598. Yilmaz, G., Lemaire, R., Keller, J., Yuan, Z., 2008. Simultaneous nitrification, denitrification and phosphorus removal from nutrient-rich industrial wastewater using granular sludge. Bitotechnol. Bioeng 100 (3), 529–541. Yu, H.Q., Tay, J.H., Fang, H.H.P., 2001. The role of calcium in sludge granulation during UASB reactor start-up. Water Res. 35, 1052–1060. Zhou, Y., Pijuan, M., Yuan, Z.G., 2007. Free nitrous acid inhibition on anoxic phosphorus uptake and denitrification by poly-phosphate accumulating organisms. Biotechnol. Bioeng 98 (4), 903–912. Zhu, J., Wilderer, P.A., 2003. Effect of extended idle conditions on structure and activity of granular activated sludge. Water Res. 37, 2013–2018.
Available at www.sciencedirect.com
journal homepage: www.elsevier.com/locate/watres
Effect of long term anaerobic and intermittent anaerobic/aerobic starvation on aerobic granules Maite Pijuan, Ursula Werner, Zhiguo Yuan* The University of Queensland, Advanced Water Management Centre (AWMC), Brisbane, QLD 4072, Australia
article info
abstract
Article history:
The effect of long term anaerobic and intermittent anaerobic/aerobic starvation on the
Received 7 January 2009
structure and activity of aerobic granules was studied. Aerobic granular sludge treating
Received in revised form
abattoir wastewater and achieving high levels of nutrient removal was subjected to 4–5 week
7 May 2009
starvation under anaerobic and intermittent anaerobic/aerobic conditions. Microscopic
Accepted 8 May 2009
pictures of granules at the beginning of the starvation period presented a round and compact
Published online 18 May 2009
surface morphology with a much defined external perimeter. Under both starvation conditions, the morphology changed at the end of starvation with the external border of the
Keywords:
granules surrounded by floppy materials. The loss of granular compactness was faster and
Aerobic granules
more pronounced under anaerobic/aerobic starvation conditions. The release of Ca2þ at the
Starvation
onset of anaerobic/aerobic starvation suggests a degradation of extracellular polymeric
Morphology
substances. The activity of ammonia oxidizing bacteria was reduced by 20 and 36% during
Calcium
anaerobic and intermittent anaerobic/aerobic starvation, respectively. When fresh waste-
Nitrification
water was reintroduced, the granules recovered their initial morphology within 1 week of
Enhanced biological
normal operation and the nutrient removal activity recovered fully in 3 weeks. The results
phosphorus removal
show that both anaerobic and intermittent anaerobic/aerobic conditions are suitable for
Recovery
maintaining granule structure and activity during starvation. ª 2009 Elsevier Ltd. All rights reserved.
1.
Introduction
In recent years, aerobic granular sludge has become a promising technology for wastewater treatment. It presents several advantages compared to conventional flocular sludge systems including lower operational costs and lower space requirement (de Bruin et al., 2004). Aerobic granules are defined as aggregates of microbial origin, which do not coagulate under reduced hydrodynamic shear and which subsequently settle significantly faster than activated sludge flocs. Aerobic granules have been mostly cultivated in sequencing batch reactors (SBR) using synthetic wastewater (Tay et al., 2002; Liu et al., 2003; Yang et al., 2003) and most recently using real wastewater (de Villiers and Pretorius, 2001; Arrojo et al., 2004; de Bruin et al., 2004; Yilmaz et al., 2008). Using aerobic granules,
simultaneous removal of organic carbon (denoted as chemical oxygen demand, or COD), nitrogen (N) and phosphorus (P) has been achieved (Cassidy and Belia, 2005; de Kreuk et al., 2005; Yilmaz et al., 2008). The aerobic biogranulation technology will be moving towards industrial use for biological wastewater treatment. Several industry sectors such as the dairy and food processing industries and abattoirs are producers of large volumes of wastewater requiring treatment. The aerobic granulation technology has recently been demonstrated to be suitable for these high-strength wastewaters (Yilmaz et al., 2008). One of the challenges for the treatment of such wastewaters is to cope with the large fluctuations of the wastewater flow linked to the industrial activity. Due to seasonal closure of the industry, the wastewater flow could
* Corresponding author. E-mail address: [email protected] (Z. Yuan). 0043-1354/$ – see front matter ª 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2009.05.007
water research 43 (2009) 3622–3632
even be completely stopped for an extended period of time (weeks). Therefore, the ability of the granules and the biomass to survive under such conditions needs to be investigated in order to achieve reliable biological wastewater treatment. A number of studies have been reported in literature on the effect of starvation on the structure of aerobic granules. Zhu and Wilderer (2003) found that the appearance of granules that developed from a synthetic wastewater (glucose and peptone) did not visually change after a 7 week anaerobic idle phase. As a microbial activity indicator, the oxygen consumption rates were measured, which returned to their pre-starvation levels just 1 week after restarting the reactor. Wang et al. (2008) stored granules treating synthetic wastewater anaerobically for 7 months under low temperatures (in a fridge) and they showed that these granules were able to preserve their granular morphology and recover their nitrification and COD removal ability in approximately 2 weeks. Also, Adav et al. (2007) suggested that storage of aerobic granules at subfreezing temperatures (20 C) was the best way of preserving stability and activity of the granules. However, these preservation conditions are not feasible in large scale systems. In comparison, the impact of starvation on the activities of bacterial community in flocular activated sludge has been investigated more thoroughly. In general, it has been found that anaerobic and intermittent anaerobic/aerobic conditions resulted in the lowest bacteria decay rates, and have therefore been recommended for maintaining bacterial activity in wastewater treatment plants during periods without feed (Siegrist et al., 1999; Lee and Oleszkiewicz, 2003; Yilmaz et al., 2007; Lu et al., 2007). In this study, we investigate the effect of anaerobic and intermittent anaerobic/aerobic starvation conditions on the structure, stability and microbial activity of aerobic granules. Aerobic granular sludge treating real, abattoir wastewater was starved for 4–5 weeks under the above conditions. The external morphology of the granules was monitored using light and scanning electron microscope (SEM). The changes in O2, Ca2þ and pH distribution inside the granules were monitored using microsensors. The activities of ammonia oxidizing bacteria (AOB) and polyphosphate accumulating organisms (PAO) were monitored through batch tests conducted weekly. The recovery of morphology and nutrient removal performance of the granules were examined after fresh wastewater was reintroduced.
2.
Materials and methods
2.1.
Source of granules
The granules used were sourced from a SBR, called the parent reactor, which performed biological COD, nitrogen and phosphorus removal from abattoir wastewater containing approximately 250 mg N/L and 40 mg P/L of total nitrogen and total phosphorus, respectively (Yilmaz et al., 2008). The SBR was operated with a cycle time of 8 h, with 3 L of abattoir wastewater fed in each cycle at the beginning of its 78 min anaerobic period reaching a total working volume of 5 L. Nitrogen and phosphorus were removed in the subsequent
3623
aerobic period, primarily through simultaneous nitrification, denitrification and phosphorus uptake. More details of the operation of the SBR can be found in Yilmaz et al. (2008).
2.2.
Anaerobic/aerobic starvation
One litre of granular sludge was withdrawn from the parent reactor. At the point of sludge sampling, the parent SBR had a mixed liquor suspended solids (MLSS) concentration of 14 g/ L and a mixed liquor volatile suspended solids (MLVSS) concentration of 10 g/L, respectively. It was achieving >99% removal of N and P from an abattoir wastewater (0.4 mg N– 3 in the effluent). The sludge was NHþ 4 and 0.5 mg P–PO4 diluted with 1 L of effluent from the same reactor resulting in an initial MLSS concentration of approximately 7 g/L, and placed to another SBR, called the anaerobically/aerobicallystarved SBR, for 28 days without feed, This SBR was aerated for 15 min in every 6 h. Granules were allowed to settle during the remaining time.
2.3.
Anaerobic starvation
A total of 0.5 L of granular sludge was withdrawn from the parent SBR 1 month after the granules used for the anaerobic/ aerobic starvation was withdrawn to allow the parent SBR to recover from the biomass loss. At the time of sampling, the parent SBR had a MLSS concentration of 16 g/L and a MLVSS concentration of 13.8 g/L and it was achieving >99% of N removal and 53% of P removal (0.7 mg N–NHþ 4 and 16.4 mg P– in the effluent). The sludge was diluted with 1.5 L of PO3 4 effluent from the same reactor providing an initial MLSS concentration of 4 g/L. These 2 L were kept in a batch reactor (called anaerobically-starved SBR) in settling conditions for a period of 32 days. During this period no feed was provided.
2.4.
Monitoring during starvation
During both starvation experiments, liquid phase samples were taken every two days. Samples were taken using a syringe after a short period of mixing, and immediately filtered through disposable Millipore filter units (0.22 mm pore size) for analyses of ammonia, nitrite, phosphate, calcium, magnesium and potassium ions. Sludge samples for granule size distribution and light microscopic analysis was taken weekly, whereas sludge samples for SEM were taken at the beginning and at the end of the starvation period. Granules were removed from the starvation reactors weekly for the measurements of the micro-scale oxygen concentration profiles inside granules. Granules were also taken on days 0, 1 and 6 for the measurements of Ca2þ and pH concentration profiles during both starvation conditions and additionally on day 3 of the anaerobic/aerobic starvation. More details of these measurements will be described later. Batch tests for the assessment of nitrification and P removal activity were performed weekly. Fluorescent in situ hybridization (FISH) was performed on biomass samples taken at the beginning and end of each starvation experiment.
3624
2.5.
water research 43 (2009) 3622–3632
Batch experiments
During both starvation conditions, batch experiments were carried out to monitor the nitrification activity of the biomass by measuring the ammonium oxidation rate over a period of 2 h after a pulse addition of ammonia. In each test, 100 mL of mixed liquor was taken from the starved SBR. After washing the sludge with effluent from the parent SBR (which had very low concentrations of N, P and biodegradable COD), a pulse of a concentrated ammonium solution (2 g N–NHþ 4 /L) was added to the batch reactor, resulting in an initial ammonium concentration in the reactor of 20 mg/L N–NHþ 4 . Mixed liquor samples were taken and filtered every 10 min for analysis of NHþ 4 and NO 2 . The pH was manually controlled in the range 7.4–7.6 through the addition of 0.1 M of NaOH. The ammonia oxidation rate was determined as the slope of the ammonium profile. The decay rate of ammonia oxidizing bacteria (AOB) was determined through fitting an exponential function to the measured ammonia oxidation rates as a function of time (Eq. (1)): y ¼ A ekt
(1)
where y is the biomass specific ammonia oxidation rate (mg N–NHþ 4 /g VSS*h); A is initial biomass specific ammonia oxidation rate (mg N–NHþ 4 /g VSS*h), t is the starvation time (d) and k is the decay rate (d1). The PAO P-uptake activity was also monitored through weekly batch tests carried out in a 100 mL reactor during the intermittent anaerobic/aerobic starvation period. At the start of the experiment, 100 mL granular sludge were withdrawn from the starved SBR at the end of the aeration phase and washed þ with effluent from the parent SBR. A solution of P–PO3 4 and NH4 was added to the batch reactor resulting in initial concentraþ tions of 20 mg P–PO3 4 /L and 10 mg N–NH4 /L/L, respectively. Acetate was added to reach a concentration of 50 mg COD/L. Allylthiourea (ATU) was also added at a concentration of 20 mg/ L to avoid nitrification. During the first hour, nitrogen was sparged to maintain strict anaerobic conditions. After an anaerobic period, the 100 mL reactor was sparged with air for 1 h. Dissolved oxygen was not controlled and was higher than 2.5 mg/L during the entire aerobic period. Samples were taken every 10 min for analysis of acetate and P. The pH was manually controlled in the range 7.4–7.6 through addition of 0.1 M HCl and 0.4 M NaOH.
2.6.
Reactor operation in the recovery period
After 28 days of anaerobic/aerobic starvation, feeding of wastewater to the SBR was started. The 6 h cycle consisted of 4 min feeding (1 L), followed by 1 h anaerobic, 294 min aerobic (DO was controlled between 1.5 and 2.0 mg/L), 1 min settling and 1 min decant periods. The recovery of nitrogen removal was monitored for 1 month through weekly batch experiments as described before. The recovery of phosphorus removal was monitored every 2–3 days for a period of 9 days through batch experiments as described above.
2.7.
Analytical procedures
The ammonium (NHþ 4 –N), nitrate (NO3 –N), nitrite (NO2 –N) and 3 orthophosphate (PO4 –P) concentrations were analyzed using
a Lachat QuikChem8000 Flow Injection Analyzer (Lachat Instrument, Milwaukee). MLSS and MLVSS were analysed according to the standard methods (APHA, 1995). Potassium (Kþ) and magnesium (Mg2þ) ions were measured by Inductively Coupled Plasma – Atomic Emission Spectrometry (ICP–AES Varian Vista- PRO, Varian, Inc.). To determine the size distribution of the granules, 30 mL of well mixed granular sludge were pumped through a Malvern laser light scattering instrument, Mastersizer 2000 series (Malvern Instruments, Worcestershire, UK). The technique of laser diffraction is based on the principle that particles passing through a laser beam will scatter light at an angle that is directly related to their size. This method represents a rapid and robust measurement of particles present in a bulk with a range of 0.02–2000 mm. Microscopic observation of granules confirmed that the absolute majority of the granules in the reactor have sizes in this range (data not presented). Oxygen concentration profiles in the granules were measured using Clark type microelectrodes (Revsbech, 1989). The sensors had an actual sensing surface of 5 mm in diameter and a 90% response time of less than 5 s. The measurements were conducted in a small flow cell with filtered wastewater (GF/A, Whatman) horizontally flowing through the cell (volumetric displacement rate: 25 mL/min). The wastewater used was taken from the parent reactor at the end of the aerobic phase, thus, it had very low levels of ammonium and no detectable levels of readily biodegradable COD. To allow for a comparison of oxygen penetration depths in the granules during the time course of the starvation, the oxygen concentration in the flow cell was adjusted to 3–3.5 mg/L, similar to the oxygen concentration in the parent reactor at the end of the aerobic phase. The oxygen penetration depth was visually determined from replicate profiles. The Ca2þ and pH concentration profiles in the granules were measured using liquid ion-exchange (LIX) microsensors as described in de Beer and Kuhl et al. (2000). These sensors had a 90% response time of less than 5 s and a sensing surface diameter of ca. 5 mm. For these measurements a larger flow cell was used, with a volumetric displacement rate of 55 mL min1. During the anaerobic/aerobic starvation conditions, granules and supernatant were taken directly from the anaerobically/ aerobically-starved reactor. The measurements during anaerobic starvation were conducted on sludge kept in a separate anaerobic reactor to keep disturbances in the main experiment at a minimum. During the microsensor measurements, the pH was stabilized at a level measured in the starvation reactors on the respective day. Oxygen concentrations in the flow cell were adjusted to 0.5–0.7 mg/L for both starvation conditions to achieve similar conditions during the measurements of pH and Ca2þ distribution. This oxygen concentration can be found in the parent reactor at the beginning of the aerobic period and led to an oxygen penetration of less than 50 mm. For all microsensor measurements 25–50 mL of sample were taken from the reactors and up to 10 granules were randomly selected for profiling. Four to ten replicate microsensor profiles were measured with step sizes of 25–100 mm in different granules. The granule morphology was qualitatively observed using a stereo microscope (Olympus SZH10) and a scanning electron microscope (SEM, Jeol JSM-6300F).
3625
water research 43 (2009) 3622–3632
2.8.
Microbial analyses
Sludge samples were fixed and FISH probed as previously described by Amann (1995). Oligonucleotide probes used in this study were EUBmix (Daims et al., 1999) for all bacteria, PAOmix (for Candidatus Accumulibacter phosphatis or Accumulibacter, comprising equal amounts of probes PAO462, PAO651 and PAO846, Crocetti et al., 2000) and GAOmix (for Candidatus Competibacter phosphatis or Competibacter, comprising equal amounts of probes GAOQ431 and GAOQ989, Crocetti et al., 2002). FISH images were collected using a Zeiss LSM 510 Meta Confocal microscope with a 63 Plan-Apochromat oil immersion lens. FISH quantification was performed according to Zhou et al. (2007) where the relative abundance of each group was determined as mean percentage of all bacteria.
3.
Results and discussion
3.1.
Impact of starvation on granular structure
Two different starvation conditions were tested in this study, the continuous anaerobic starvation and the intermittent anaerobic/aerobic starvation. Fig. 1 shows the size distribution profiles of the granules under the two starvation conditions. The initial granule sizes in both studies were between 0.6 and 1.6 mm. The 50 and 90 percentiles remained relatively constant during the 30 days of starvation. This indicates that under both starvation regimes the granular structure was not lost. However, the 10 percentile decreased from 0.6 to 0.05 mm during the first week of anaerobic/aerobic starvation (Fig. 1 left) indicating the release of small particles from granules of all sizes in the reactor. The 10 percentile indicates that 10% of the particles present in the reactor have a diameter that is lower or equal to this value. In contrast, under anaerobic starvation (Fig. 1 right) the 10 percentile started to decrease after 2 weeks and low values similar to those observed in the anaerobic/aerobic starvation after week 1 were reached only at the end of the starvation period.
To monitor changes in the morphology and structure, granules were studied weekly using a light microscope (Fig. 2) and at the beginning and at the end of the starvation period using a SEM (Fig. 3). Granules from the beginning of the starvation period possessed a compact structure with a very well defined external perimeter (Figs. 2A and 3A,E). In contrast, granules after 1 month of anaerobically/aerobically (Fig. 2B) and anaerobically-starved conditions (Fig. 2C), had a layer of fluffy materials at the edges of the granules. They still preserved their structural integrity, but became more transparent in the edges compared to the fresh granules. The appearance of fluffy materials surrounding the granules could be responsible for considerably reduced ten percentile size values observed in both cases. Similar morphology of starved granules was also observed in the parent SBR in several occasions when the wastewater supply was cut for a few days, and the SBR was maintained under alternating anaerobic/ aerobic starvation conditions (data not shown). On the other hand, when the waste sludge from this SBR was stored in a bottle (anaerobic conditions) this fluffy material was not observed. The SEM pictures revealed that granules at the end of the anaerobic/aerobic starvation (Figs. 3B,C) differed substantially in morphology from the ones at the beginning of the starvation, with rough surfaces and irregular shapes. It seems that parts of the granules were consumed or broken off. In contrast, granules kept under anaerobic starvation conditions did not change significantly after 20 days of starvation except for the appearance of significant cracks from the edge to the interior of the granules in some of the granules (Fig. 3F). Keeping granules under long term aerobic starvation conditions, Wang et al. (2005) observed that after 20 days the central part of the granules was becoming more transparent and in some cases the core part became hollow, while the outer shell still remained intact. They attributed this phenomenon to the bacterial consumption of part of the extracellular polymeric substances (EPS) present in the granules. Thus, its own producers would consume the EPS when substrate is severely limited. Conducting a closer investigation on the biodegradability of aerobic granular EPS under aerobic starvation conditions, Wang et al. (2007) quantified that about 50% of EPS produced by aerobic granules could be utilized by their producers and the rest would remain intact to
1800
Granule size (um)
1600 1400 1200 1000 800 600
d(0.1) d(0.5) d(0.9)
400 200 0
0
5
10
15
Time (days)
20
25
0
5
10
15
20
25
30
35
Time (days)
Fig. 1 – Distribution of granule size in the reactor during the entire study. d(0.1), d(0.5), d(0.9) are the 10, 50 and 90 percentiles of the distribution, respectively. Left – anaerobic/aerobic starvation; right – anaerobic starvation.
3626
water research 43 (2009) 3622–3632
Fig. 2 – Light microscope images of the aerobic granules before starvation (A) and 1 month after anaerobic/aerobic (B) and anaerobic starvation (C), showing the change in morphology as a result.
Fig. 3 – SEM images of the outer morphology of granules. Top-anaerobic/aerobic starvation: before (A) and 26 days after (B,C) starvation and (D) after 1 week of recovery. Bottom-anaerobic starvation: before (E) and 20 days after (F, G) starvation.
preserve the structural integrity of the granule. This could explain the changes observed mostly in the anaerobic/aerobic granules from our study, where a significant degradation of the external parts of the granules was detected. The degradation of the external EPS layer in our study may be linked to the distribution of specific EPS components, which can differ depending on the granular type (Chen et al., 2007). Adav et al. (2008) found that the granular structure was most affected by the degradation of b-polysaccharides in phenol fed granules. In another study, the degradation of proteins and a-polysaccharides had most effects on the structural morphology of granules withdrawn from the parent reactor used in our experiment (Seviour et al., 2009). However, a detailed investigation of the distribution of EPS components was not conducted in our study. In comparison, anaerobic starvation conditions had a milder effect on the outer morphology of the granules, compared to the intermittent anaerobic/aerobic
conditions, possibly because the degradation of EPS under strictly anaerobic conditions may be more difficult. Adav et al. (2009) anaerobically stored aerobic granules for 60 days at 8 C in a bottle. They hypothesized that the proteins present in the EPS were hydrolysed and then used by the nearby anaerobic strains, gradually digesting the granules from the inside out as a mechanism of survival. However, even after the anaerobic/aerobic starvation period, the recovery of the initial granular morphology was relatively quick and the granules had a similar shape to the initial ones, just 1 week after the feed was reintroduced (Fig. 3D).
3.1.1.
Effect of starvation on oxygen penetration
In diffusion dominated systems, the penetration of oxygen into microbial aggregates such as granules is determined by the oxygen concentration in the bulk water, the oxygen
3627
water research 43 (2009) 3622–3632
Table 1 – Oxygen penetration depths in granules under anaerobic/aerobic and under anaerobic starvation conditions. O2 penetration (mm) Anaerobic/aerobic starvation Anaerobic starvation
Day 0
Day 7
Day 14
Day 21
Day 28
238.6 42.4 185.7 96.7
250.7 1.2 n.a.
366.7 70.1 157.1 27.8
n.a. 307.5 186.7
350 147.2 356.3 85.6
n.a.: not available. Data are given as average standard deviation.
consumption rate inside the aggregates, the flow rate of the medium and the diffusivity of the matrix (Jørgensen and Marais, 1990; Chiu et al., 2007; Li et al., 2008). Previous research showed that aerobic granules with higher extracellular polymer content had a lower oxygen diffusivity than granules with a lower extracellular polymer content (Chiu et al., 2006). Assuming a degradation of EPS and a decrease in oxygen consumption during both starvation conditions, we expected to see an increase of oxygen penetration depths (under comparable oxygen concentrations and comparable flow conditions in the bulk water) over the time course of the experiments. The depths of oxygen penetration during both starvation periods are summarised in Table 1, while some examples of the measured oxygen profiles are shown in Fig. 4. A diffusive boundary layer above the surface of the granules was clearly visible in all oxygen profiles and all profiles had the parabolic shape of diffusion dominated consumption
profiles, indicating the importance of diffusive transport to deeper layers and the uptake of oxygen inside the granules. Occasionally, increased oxygen concentrations in deeper parts of the profiles indicated that oxygen was also transported through channels and holes into the granules, either by diffusive or advective transport (e.g. Fig. 4d). An increased oxygen penetration depth was measured in the anaerobic/aerobic starving granules from the second week on (Figs. 4a,b, Table 1), whereas during the anaerobic starvation, oxygen penetration increased only from the third week on (Figs. 4c,d; Table 1). At the end of the experiments, the oxygen penetration depths were similar for both starvation conditions (Table 1). The increase in oxygen penetration depths is likely due to a combination of decreased biological oxygen consumption (e.g., heterotrophic respiration using the intrinsic carbon pool and nitrification, to be further discussed below) and increased diffusivity of the EPS matrix resulting from its degradation
-600
b
a -400 -200 0
depth below surface (µm)
200 400 600 0
1
2
3
4
0
1
2
3
4
1
2
3
4
-600
d
c -400 -200 0 200 400 600 0
1
2
3
4
0
O2 concentration (mg/L) Fig. 4 – Oxygen profiles in granules on day 0 (a) and after 14 days of anaerobic/aerobic starvation (b) and on day 0 (c) and after 21 days of anaerobic starvation (d). The line through zero represents the granule surface. The dashed line indicates the approximate height of the diffusive boundary layer. The different symbols represent replicate profiles.
3628
water research 43 (2009) 3622–3632
stability was attributed to the role of Ca2þ in the EPS structure. To assess a possible effect of EPS degradation during starvation conditions on Ca2þ concentrations and distribution patterns, Ca2þ and pH profiles were measured within the granules. Ca2þ and pH microsensor profiles in the granules were measured on days 0, 1 and 6 during both starvation conditions and additionally on day 3 during the anaerobic/ aerobic starvation conditions. The Ca2þ profiles obtained in anaerobic/aerobic starving granules are represented in Fig. 5. On the first day of anaerobic/aerobic starvation a substantial increase of Ca2þ was observed below 100 mm depth, indicating the considerable production of Ca2þ in the outer layer of the granule. On day 3, the Ca2þ concentrations in the granule were already much lower, while they still remained elevated with depth, when compared to the measurements of day 0. The increased concentrations in the first 100 mm indicate the on-going production of Ca2þ in this zone. On day 6, the Ca2þ concentrations profiles showed only minor variations with depth with some profiles showing small production and others showing small consumption zones, the profiles were
during starvation. The respective contribution of both factors in the determination of the oxygen penetration depths can not be deduced from our data set. However, the effects of starvation on the EPS matrix may explain the earlier increase of the oxygen penetration depth during anaerobic/aerobic starvation conditions as compared to the anaerobic starvation conditions (Table 1). The earlier increase of the averaged oxygen penetration depth during anaerobic/aerobic conditions corroborates the observations of more severe granular structure loss (hence, an increased impact on the matrix diffusivity) under this starvation condition.
3.1.2.
Effect of starvation on Ca2þ concentrations and pH
2þ
Ca is an important bridging agent in the EPS and has been recognized to play an important role in the self-immobilization of microbial aggregates (Morgan et al., 1990; Yu et al., 2001). Ren et al. (2008) found that Ca2þ rich aerobic granules had a more rigid structure and a higher shear strength as compared to granules that were grown in water containing lower concentrations of Ca2þ. The enhanced structural
-500
b
a
0
depth below surface (µm)
500
1000 day 1
day 0
1500 6
8
10
12
14
16
20 0
18
-500
100
50
150
200
250
350
300
d
c
0
500
1000 day 3
day 6
1500 6
8
10
12
14
16
18
20
6
8
10
12
14
16
18
20
Ca2+ (mg/L) Fig. 5 – Ca2D concentration profiles in the intermittent anaerobic–aerobically starved granules, (a, b, c, d; days 0, 1, 3 and 6 of starvation, respectively). The zero line represents the granule surface. The different symbols represent replicate profiles. Note the different scale of the x axis in plot b.
3629
water research 43 (2009) 3622–3632
thus comparable to those measured on day 0. The profiles from day 6 demonstrated that large amounts of Ca2þ were no longer produced. The increased concentrations of Ca2þ on days 1 and 3 may originate from a degradation of EPS or may come from a dissolution of CaCO3 precipitates. Dissolution of CaCO3 is expected to cause an increase in the pH at the respective depth layer. However, pH profiles, measured concurrently to the Ca2þ profiles, on day 1 of anaerobic/aerobic starvation (corresponding to the day when the highest Ca2þ production was observed inside the granules) were irregular with depth, some profiles showed small pH decreases and others small pH increases. Additionally, in contrast to the Ca2þ profiles, the pH profiles did not change over the time course of the first week of starvation. These observations support the hypothesis that Ca2þ originated from the biodegradation of the EPS rather than dissolution of CaCO3. Our hypothesis that the EPS degradation, which presumably caused the observed major Ca2þ release during the first 3 days of anaerobic/aerobic starvation corresponds to observations on EPS degradation by Wang et al. (2007). They suggested that the degradation of EPS during aerobic starvation was happening at a faster rate during the first 4 days of their study and slowed down afterwards. Interestingly, no Ca2þ concentration gradients were observed in the profiles measured in the experiments during anaerobic starvation conditions. During all measurements, Ca2þ concentration profiles showed only minor variations with depths and no major Ca2þ production zones were observed. Also the pH profile remained comparable during all measurements, showing only minor variations and irregular trends with depths. The pH profiles measured during anaerobic starvation resembled the pH profiles measured during anaerobic/aerobic starvation conditions. These observations support our hypothesis that the EPS of the granules degraded less in the beginning of the anaerobic starvation conditions. However, the remarkable magnitude of the Ca2þ production observed on day 1 of anaerobic/aerobic starvation indicates that further investigations are necessary to clarify the role of Ca2þ during EPS degradation.
3.2.
Impact of starvation on nutrient removal
3.2.1.
Monitoring nitrification activity
Weekly batch tests were performed with sludge withdrawn from both starved SBRs in order to study the effect of starvation conditions on the activity of nitrifiers.
The nitrification tests were focused on the ammonia oxidizing bacteria (AOB) population, because no nitrate was produced in the parent SBR and the population of nitrite oxidizing bacteria (NOB) was negligible (Yilmaz et al., 2008). Fig. 6 shows the profiles of NHþ 4 oxidation rate measured weekly during both starvation periods. These rates were calculated from the NHþ 4 profiles measured during batch tests as explained previously. The AOB decay rate calculated from the NHþ 4 oxidation rate in the anaerobic/aerobic starvation was 0.0182 d1. This number is very similar to the one obtained by Yilmaz et al. (2007) (0.0165 d1) with a floccular sludge treating abattoir wastewater. The AOB decay rate under anaerobic conditions was determined to be 0.0069 d1. The results show that both conditions are suitable for preserving AOB activities. In comparison, anaerobic starvation seems to further lower the AOB decay rate. The activity of AOB was reduced by 20 and 36% during anaerobic and intermittent anaerobic/aerobic starvation, respectively. These results are in agreement with findings by other researchers (Siegrist et al., 1999).
3.2.2.
3.5
3.5 y = 2.991e-0.0182t R2 = 0.9876
2.5
y = 2.0754e-0.0069t
3.0
mg N/ gVSS h
mg N/ gVSS h
3.0
2.0 1.5 1.0
R2 = 0.799
2.5 2.0 1.5 1.0 0.5
0.5 0.0
PAO activity during starvation
Phosphorus concentrations were periodically monitored in the starved SBRs. Fig. 7 shows the P profiles in the SBR during both starvation periods. P-release increased almost linearly initially but slowed down after around 8 and 7 days in the intermittent anaerobic/aerobic and anaerobic starvation respectively. This behaviour is somewhat different from observations in an earlier study by Yilmaz et al. (2007), where a floccular sludge was submitted to the same anaerobic/ aerobic starvation conditions as used in this study. In that study, no accumulation of P was detected during the first 1– 2 weeks of starvation. The concentration profiles of Kþ and Mg2þ displayed similar trends, with P:Kþ:Mg2þ molar ratios of 3.11:1.05:1 for the anaerobic/aerobic starvation and 3.10:0.92:1 for the anaerobic starvation during the first week of the study, suggesting that phosphate release likely originated from the breakdown of polyphosphate with a theoretical structure of Mg1/3K1/3PO3 (Smolders et al., 1994). This is a confirmation that PAO use their internal polyphosphate as a source of energy during these starvation conditions. The amount of phosphorus, potassium and magnesium release was much higher in the anaerobic/aerobic starvation period (Fig. 7). This may be related to the different starvation conditions applied, but probably is also affected by the changed
0
5
10
15
Time (days)
20
25
30
0.0
0
5
10
15
20
25
30
35
Time (days)
Fig. 6 – Variation of NHD 4 oxidation rate over the anaerobic/aerobic (left) and the anaerobic starvation period (right).
3630
water research 43 (2009) 3622–3632
500
40 30
300
mg/L
mg/L
400
200
10
100 0
20
0
5
10
15
20
25
0
30
0
5
10
Time (days)
15
20
25
30
35
Time (days)
Fig. 7 – Phosphorus (C), potassium (B) and magnesium (;) profiles in the SBR starved under anaerobic/aerobic (left) and under anaerobic conditions (right).
bacterial populations in the sludge. There was a significant increase of the GAO population in the parent reactor during the period between the two starvation studies. FISH analysis showed that the Accumulibacter-PAO decreased from 29% of the entire bacterial population at the beginning of the anaerobic/ aerobic starvation to 21% at the end of the starvation while Competibacter-GAO was present at around 9% and maintained stable. The sludge used for the anaerobic starvation study had 28% of Accumulibacter-PAO and 35% Competibacter-GAO at the beginning of the study and these percentages did not significantly change along the anaerobic starvation period.
3.3.
Recovery after starvation
Granule structure rapidly recovered after the reintroduction of the wastewater (Fig. 3) and the small particles present in the reactor were washed out during the first day of reactor operation due to the short settling time applied. The granule size distribution 1 day into the recovery period became almost identical to the distribution measured at the beginning of the starvation tests. In order to study the ability of the granules to recover their activity in terms of nutrient removal, recovery tests were performed on the anaerobically/aerobically-starved granules. Figs. 8A,B show the recovery of the nitrification rate and the P-
3.4.
Limitations of the study and future research
This study focused on the changes of morphological structure and microbial activity of aerobic granules subjected to two different starvation conditions. The study demonstrated that both anaerobic and intermittent anaerobic/aerobic strategies are effective in maintaining granular structure and microbial activity of aerobic granules. Both strategies are easily implementable in practice. Microsensor measurements linked with microscopic observations also contributed to the understanding of the survival mechanisms of aerobic granules in starvation. However, the study has several limitations. First of all, the study was performed with only one type of granules developed with abattoir wastewater. The experimental findings should be verified with other types of aerobic granules treating different types of wastewaters. Secondly, the micro-scale studies remain mostly observational and descriptive. Fundamental understanding of the physical and biochemical processes occurring in aerobic granules during starvation
B 3.5
16
3.0
14
2.5
mg P/ gVSS h
mg N/ gVSS h
A
uptake, respectively. The nitrification rate increased from the beginning of the recovery period, but it took 3 weeks for the nitrification rate to fully recover. In comparison, the P-uptake rate, as a measure of the PAO activity, recovered 10 days into the recovery period.
2.0 1.5 1.0 STARVATION PERIOD
0.5 0.0
0
5
RECOVERY PERIOD
10 15 20 25 30 35 40 45 50
Time (days)
STARVATION PERIOD
RECOVERY
12 10 8 6 4 2 0
0
5
10
15
20
25
30
35
Time (days)
Fig. 8 – Recovery of the NHD 4 oxidation rate (A) and the P-uptake rate (B) in the SBR starved under intermittent anaerobicaerobic conditions.
water research 43 (2009) 3622–3632
periods is still missing, and requires further research. Thirdly, the sludge samples used for the anaerobic and the intermittent aerobic/anaerobic starvation tests were taken from the parent reactor at different times and had differences in microbial structures. Therefore, caution has to be taken when comparing the results from the two types of tests.
4.
Conclusions
The effect of a long starvation period (approximately 1 month) under anaerobic and intermittent anaerobic/aerobic conditions on the structure and microbial activity of aerobic granules was studied. The following conclusions are drawn: Aerobic granules can be stored under these conditions for at least 4 weeks with no wastewater supply without loosing their structural integrity. Although the results are not directly comparable since the starting biomass was different, better results were obtained under anaerobic starvation conditions in terms of maintaining the surface structure and microbial activity of granules. The intermittent availability of oxygen likely enhanced the consumption of EPS around the edge of granules, causing rougher surface structures. Full recovery of the external morphology of aerobic granules can be achieved within 1 week of resuming the wastewater addition. The decay rates of ammonia oxidizing bacteria found in aerobic granules are very similar to the ones found in floccular sludge.
Acknowledgement This work was funded by the Environmental Biotechnology Cooperative Research Centre (EBCRC) Pty Ltd, Australia. The authors wish to thank Dr. Romain Lemaire and Mr. Paritam Dutta from the Advanced Water Management Centre for their help on the SEM analysis.
references
Adav, S.S., Lee, D.-J., Tay, J.H., 2007. Activity and structure of stored aerobic granules. Environmental Technol 28, 1227–1235. Adav, S.S., Lee, D.J., Tay, J.H., 2008. Extrapolymeric substances and structural stability of aerobic granules. Water Res. 42, 1644–1650. Adav, S.S., Lee, D.-J., Lai, J.-Y., 2009. Proteolytic activity in stored aerobic granular sludge and structural integrity. Bioresource Technol 100, 68–73. Amann, R.I., 1995. Situ Identification of Microorganisms by Whole Cell Hybridization with rRNA-targeted Nucleic Acid Probes. In: Molecular Microbial Ecology Manual. Kluwer Academic Publications, Dordrecht, Holland, pp. 3.3.6, 1-15. APHA, 1995. Standard Methods for the Examination of Water and Wastewater, nineteenth ed. American Public Health Association, Washington, DC.
3631
Arrojo, B., Mosquera-Corral, A., Garrido, J.M., Mendez, R., 2004. Aerobic granulation with industrial wastewater in sequencing batch reactors. Water Res. 38 (14–15), 3389–3399. Cassidy, D.P., Belia, E., 2005. Nitrogen and phosphorus removal from abattoir wastewater in a SBR with aerobic granular sludge. Water Res. 39 (19), 4817–4823. Chen, M.Y., Lee, D.J., Tay, J.H., 2007. Distribution of extracellular polymeric substances in aerobic granules. Appl. Microbiol. Biotechnol 73, 1463–1469. Chiu, Z.C., Chen, M.Y., Lee, D.J., Tay, S.T.L., Tay, J.H., Show, K.Y., 2006. Diffusivity of oxygen in aerobic granules. Biotechnol. Bioeng 94, 505–513. Chiu, Z.C., Chen, M.Y., Lee, D.J., Wang, C.H., Lai, J.Y., 2007. Oxygen diffusion in active layer of aerobic granule with step change in surrounding oxygen levels. Water Res. 41 (4), 884–892. Crocetti, G.R., Hugenholtz, P., Bond, P.L., Schuler, A., Keller, J., Jenkins, D., Blackall, L.L., 2000. Identification of polyphosphate accumulating organisms and design of 16S rRNA-directed probes for their detection and quantitation. Appl. Environ. Microbiol. 66, 1175–1182. Crocetti, G.R., Banfield, J.F., Keller, J., Bond, P.L., Blackall, L.L., 2002. Glycogen accumulating organisms in laboratory-scale and full-scale activated sludge process. Microbiology 148, 3353–3364. Daims, H., Bruhl, A., Amann, R., Schleifer, K.H., Wagner, M., 1999. The domain-specific probe EUB338 is insufficient for the detection of all bacteria: development and evaluation of a more comprehensive probe set. Syst. Appl. Microbiol. 22, 434–444. de Beer, D., Kuhl, M., et al., 2000. A microsensor study of light enhanced Ca2þ uptake and photosynthesis in the reefbuilding hermatypic coral Favia sp. Marine Ecol.-Progr. Ser 194, 75–85. de Bruin, L.M.M., de Kreuk, M.K., van der Roest, H.F.R., Uijterlinde, C., van Loosdrecht, M.C.M., 2004. Aerobic granular sludge technology: an alternative to activated sludge? Water Sci. Technol 49 (11–12), 1–7. de Kreuk, M., Heijnen, J.J., van Loosdrecht, M.C.M., 2005. Simultaneous COD, nitrogen and phosphate removal by aerobic granular sludge. Biotechnol. Bioeng 90 (6), 761–769. de Villiers, G.H., Pretorius, W.A., 2001. Abattoir effluent treatment and protein production. Water Sci. Technol 43 (11), 243–250. Jørgensen, B.B., Marais, D.J.D., 1990. The diffusive boundary-layer of sediments – oxygen microgradients over a microbial mat. Limnol. Oceanogr 35, 1343–1355. Lee, Y., Oleszkiewicz, J.A., 2003. Effects of predation and ORP conditions on the performance of nitrifiers in activated sludge systems. Water Res. 37 (17), 4202–4210. Li, Y., Liu, Y., Shen, L., Chen, F., 2008. DO diffusion profile in aerobic granule and its microbiological implications. Enzyme Microb. Technol 43, 349–354. Liu, Y., Yang, S.F., Tay, J.H., 2003. Elemental compositions and characteristics of aerobic granules cultivated at different substrate N/C ratios. Appl. Microbiol. Biotechnol 61 (5–6), 556–561. Lu, H., Keller, J., Yuan, Z., 2007. Endogenous metabolism of Candidatus Accumulibacter phosphatis under various starvation conditions. Water Res. 41, 4646–4656. Morgan, J.W., Forster, C.F., Evison, L., 1990. Comparative study of the nature of biopolymers extracted from anaerobic and activated sludges. Water Res. 31, 3191–3194. Ren, T.T., Liu, L., Sheng, G.P., Liu, X.W., Yu, H.Q., Zhang, M.C., Zhu, J.R., 2008. Calcium spatial distribution in aerobic granules and its effects on granule structure, strength and bioactivity. Water Res. 42, 3343–3352. Revsbech, N.P., 1989. An oxygen microsensor with a guard cathode. Limnol. Oceanogr 34, 474–478. Seviour, T., Pijuan, M., Nicholson, T., Keller, J., Yuan, Z.G., 2009. Understanding the properties of aerobic sludge granules as hydrogels. Biotechnol. Bioeng 102, 1483–1493.
3632
water research 43 (2009) 3622–3632
Siegrist, H., Brunner, I., Koch, G., Phan, L.C., Le, V.C., 1999. Reduction of biomass decay rate under anoxic and anaerobic conditions. Water Sci. Technol 39 (1), 129–137. Smolders, G.J.F., Van der Meij, J., van Loosdrecht, M.C.M., Heijnen, J.J., 1994. Model of the anaerobic metabolism of the biological phosphorus removal process: stoichiometry and pH influence. Biotechnol. Bioeng 43 (6), 461–470. Tay, J.H., Liu, Q.S., Liu, Y., 2002. Characteristics of aerobic granules grown on glucose and acetate in sequential aerobic sludge blanket reactors. Environ. Technol 23 (8), 931–936. Wang, Z.-W., Liu, Y., Tay, J.-H., 2005. Distribution of EPS and cell surface hydrophobicity in aerobic granules. Appl. Microbiol. Biotechnol 69, 469–473. Wang, Z.-W., Liu, Y., Tay, J.-H., 2007. Biodegradability of extracellular polymeric substances produced by aerobic granules. Appl. Microbiol. Biotechnol 74, 462–466. Wang, X., Zhang, H., Yang, F., Wang, Y., Gao, M., 2008. Long-term storage and subsequent reactivation of aerobic granules. Bioresource Technol. 99 (13), 8304–8309.
Yang, S.F., Tay, J.H., Liu, Y., 2003. A novel granular sludge sequencing batch reactor for removal of organic and nitrogen from wastewater. J. Biotechnol 106 (1), 77–86. Yilmaz, G., Lemaire, R., Keller, J., Yuan, Z., 2007. Effectiveness of an intermittent aerobic, anoxic/anaerobic strategy for maintaining biomass activity of BNR sludge during long-term starvation. Water Res. 41 (12), 2590–2598. Yilmaz, G., Lemaire, R., Keller, J., Yuan, Z., 2008. Simultaneous nitrification, denitrification and phosphorus removal from nutrient-rich industrial wastewater using granular sludge. Bitotechnol. Bioeng 100 (3), 529–541. Yu, H.Q., Tay, J.H., Fang, H.H.P., 2001. The role of calcium in sludge granulation during UASB reactor start-up. Water Res. 35, 1052–1060. Zhou, Y., Pijuan, M., Yuan, Z.G., 2007. Free nitrous acid inhibition on anoxic phosphorus uptake and denitrification by poly-phosphate accumulating organisms. Biotechnol. Bioeng 98 (4), 903–912. Zhu, J., Wilderer, P.A., 2003. Effect of extended idle conditions on structure and activity of granular activated sludge. Water Res. 37, 2013–2018.