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Denitrification of drinking water sources by advanced biological treatment using a membrane bioreactor
DESALINATION Desalination 178 (2005) 211-218
ELSEVIER
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Denitrification of drinking water sources by advanced biological treatment using a membrane bioreactor Gianluigi Buttiglieri a, Francesca MalpeP*, Emilio Daverio b, Mauro Melchiori a, Hans Nieman b, Jos Ligthart b "Politecnico di Milano, DIIAR, Piazza L. da Vinci 32, 20133 Milano, Italy Tel. +39 (02) 2399-6434; Fax +39 (02) 2399-6499; email: [email protected] bjRC, IES, Inland and Marine Waters Unit, European Commission, TP 300, 21020 Ispra (VA), Italy Received 15 October 2004; accepted 22 November 2004
Abstract Nitrate often contaminates groundwater resources due to excessive use of fertilizers and uncontrolled on-land discharges of raw and treated wastewater and can therefore limit the direct use of groundwater for drinking water purposes. In order to investigate the possible application of a membrane bioreactor (MBR) for denitrification of groundwater, the performance of a pilot-scale MBR was tested as a function of hydraulic and biological parameters. For this purpose synthetic groundwater was prepared by taking lake water (Lago Maggiore, Italy) and adding known amounts of ethanol and sodium nitrate to study the nitrate removal capacity of the sludge, to search for an optimum C/N ratio and to measure filtering ability for micro-organisms through the membrane. The optimum C/N ratio was found at 2.2 gC/gN, resulting in an effluent nitrate concentration within the limits stated in EU Directive 98/83 and the US EPA for drinking water use. The effluent nitrite concentration was one order under the EU limit. The membrane module, Zenon ZW-10, was monitored and performed well except for a short stress episode due to low airflow, afterwards rapidly corrected and thus putting the membrane back to its previous stable behavior. Total bacterial count for the treated effluent was lower than influent water, and 100% removal was observed for both total coliforms and E. coli. Calorimetric thermograms related to heat dissipation due to biological denitrification (nitrate and nitrite consumption) and to substrate adaptation are discussed. A maximum nitrate removal rate close to 20 mgN-NO~ gVSS-1 h- ~was measured in the calorimetric tests.
Keywords: Denitrification; MBR; Drinking water; Calorimetry; Ethanol; Nitrate; Nitrite
*Corresponding author.
Presented at the conference on Membranes in Drinking and Industrial Water Production, L 'Aquila, Italy, 15-17 November 2004. Organized by the European Desalination Society. 0011-9164/05/$- See front matter © 2005 Elsevier B.V. All rights reserved doi:10.1016/j.desal.2004.11.038
212
G. Buttiglieri et al. / Desalination 178 (2005) 211-218
1. Introduction
Drinking water sources (surface and groundwater) may be heavily contaminated by nitrates as a result of excessive fertilization or other pollution-mainly anthropogenic--sources. Elevated nitrate concentration in drinking water is a potential risk to public health since it can directly cause methaemoglobinemia (blue baby syndrome) in infants and its correlation with other diseases, including cancer, is not completely clear. Because of these potential adverse effects, nitrate concentration limits in potable water have been set worldwide. For instance, within the European Union, Council Directive 98/83/EC on the quality of water intended for human consumption establishes a maximum allowable nitrate concentration of 50 mg N O 3L -t. This value is presently not met in a considerable amount of drinking water sources in Europe, so that it is mandatory to remove the excess nitrate from the contaminated water in order to ensure potable use. A membrane bioreactor (MBR) process consists of a biological suspended growth reactor coupled to a membrane acting as a separation device solving some of the problems generally associated with traditional activated sludge processes. The full-scale application of this technique is rapidly increasing in the wastewater treatment field [ 1], whereas very few applications are reported for potable water treatment [2,3]. Coupling of biological treatment and biomass separation can be very efficient for groundwater denitrification. Since microfiltration or ultrafiltration membranes can retain almost the totality of the bacterial cells, MBRs may overcome the inconvenience of microbiological contamination of treated water, almost completely reducing the number of post-treatment steps. At the same time, due to their independence from limiting factors such as the sludge age, or the ability of bacteria to form biological aggregates, MBRs favor the development and establishment of a selected biomass in the reactor, which can also act by
removing slowly biodegradable refractory organic contaminants [4]. In this work a Zenon Zee Weed ® 10 (ZW10) MBR pilot plant was used for biological nitrate removal from drinking water sources. The pilot plant was designed in two parts: an anoxic tank where nitrate is biologically reduced to nitrogen, followed by an aerobic tank for excessive carbon removal, containing the ZW 10 filtering unit. The influent was sand-filtered lake water (Lago Maggiore, Italy), spiked with nitrate in order to reach a concentration of 30 mg N L -~. Ethanol was used as carbon source at different C/N ratios.
2. Materials and methods 2.1. Pilot plant
A diagram of the pilot plant used for this research is shown in Fig. 1. The first anoxic tank (200 L) includes a submersible pump to mix the sludge and a gravity drain to the second aerobic tank. Dissolved oxygen and pH were monitored with probes linked to a bio-controller and acquired every 5 min with a PC. The second aerobic tank (190 L) includes a drain valve, a sludge recycling pump device, transmembrane pressure and temperature sensors and overflow valves. The ZW 10 membrane unit timely controls the relaxation or back-wash process by a control panel. Process parameters at start-up were 5 min of suction and 1 min of relaxation (no permeate extraction), increased up to 10 min of operation and 1 min of relaxation. The nominal porosity of the membrane was 0.2 #m and the nominal membrane surface area was 0.93 m 2. The module dimensions were I10 mm wide and almost 700 mm high. Air flow at the bottom of the membrane was maintained in the range of 2-3.5 m 3 h-1 to control fouling, thus resulting in DO concentration in the tank and in the recycle to the anoxic reactor in the range of 6-8 mg L-~. Temperature in the aerobic tank was in the range of 22-24°C during the entire experimental period.
213
G. Buttiglieri et al. / Desalination 178 (2005)211-218
Lake water + NO 3 Dissolved
Ethanol
\
I
TREATEDEFFLUENT
oxygen 14
1 Blocont roller E
Air pump Fig. 1. Pilot plant diagram.
Hydraulic retention times (HRT) varied from 19 to 37 h. The recycle flow was set at a value equal to the influent flow (250-500 L d-l).
2.2. Prepared solutions Ethanol and sodium nitrate were dosed in order to achieve a 30 mg N-NO~ L -I concentration in the influent (with 7.3 g N-NO 3 kg TSS -l d -l and 52 g COD kg TSS -I d -I) and a start-up C/N ratio of 1.8 (7.2 COD/N ratio). The prepared solutions were stored in separated 10 L tanks and fed with two peristaltic pumps.
2.3. Biomass source Inoculum biomass was taken from a municipal wastewater treatment plant. The average anoxic and aerobic sludge concentrations were 1.6 g TSS L-l and 3.2 g TSS L-l, respectively, which were increased to 2 g TSS L -l for the anoxic sludge and 4.1 g TSS L-~ for the aerobic sludge at the final stage of the experiments.
2.4. Water inflow The water inflow was sand-filtered Lago Maggiore water (average characteristics: TSS <0.1 mg L -l, N-NO3 <1 mg L -l, N-NH 4 <0.3 mg L -I, pH = 7.04, DO = 7.8 mg L-l). 2.5. Calorimeter Sludge properties such as substrate adaptation and nitrate and nitrite removal speed were calorimetrically investigated. The Bio-RC1 is a standard 2 L jacketed glass calorimeter with a sensitivity of5-10 mW L-l [5]. In the isothermal mode, a low-viscosity silicone oil is pumped at a high rate (2 L s -l) through the reactor jacket in order to maintain a constant temperature of the reactant medium (Tr) by proper control of the jacket temperature (Tj). The jacket temperature is carefully controlled by blending oil from a "hot" and a "cold" oil circuit via an electronically controlled metering valve. Therefore, when a process dissipates or takes up heat, Tj increases or decreases, respectively. The resulting gradient
214
G. Buttiglieri et al. / Desalination 178 (2005) 211-218
across the reactor wall is directly proportional to the thermal power liberated or absorbed by the process. When a sludge sample is put into the calorimeter, a specific parameter is determined using an internal calibration heater of known power input. The Bio-RC1 does not require any energy balance to be computed and, therefore, it is possible to carefully monitor and control important process variables such as temperature and medium agitation and to operate the vessel under constant and reproducible process conditions. In the anoxic tests a 1.5 L sludge volume was used, bubbled with a CO2 and N2 gas mixture, at a constant 25°C temperature. Once a stable heat flux baseline in endogenous conditions was obtained, a known amount of NO 3 or of ethanol was added to the sludge. The specific heat power (Q) measured is related to the rate of substrate degradation by the equation Q = - YQ/s " d S / d t
[W L-']
where S is the substrate concentration (mgsubstrat e L-l) and Ye/s is the energy generated for the unit substrate degrade (J mgsubs~ate-1),with measured values in the range of 20-35 J (rag N-NOx) -1 for nitrate and nitrite removal.
2. 6. Analytical methods
Nitrate was determined spectrophotometrically at 220 nm and 275 nm. Ammonia and COD analyses were carried out by means of a spectrophotometric test (Spectroquant Merck). In the beginning nitrite was measured by means of a spectrophotometric test (Spectroquant Merck) and later determined spectrophotometrically at 543 nm. Total and volatile solids determinations were performed according to Standards Methods [6]. All chemicals were reagent grade.
3. Results and discussion Nitrate concentration in the permeate and in the denitrification tank monitored during the experiments is presented in Fig. 2 together with the different C/N ratios applied. A start-up C/N ratio of 1.5 using ethanol as substrate was chosen [7], but it was increased to 1.8 gC/gN in order to cope with the DO coming in with the influent and the recycling flow. In the first 2 weeks a necessary period for biomass adaptation to new conditions was required; in fact, effluent nitrate concentrations were high above the normative limits (CD/98/83/EC). Then, with a C/N ratio of 2.4 before and 3.0 after, complete nitrate removal was achieved. Between 24 March and 5 April, with a 2.7 C/N ratio, nitrate concentration was null in the anoxic tank but showed a slight increase in the permeate. With a C/N ratio of 2.2 at the end, an optimum was found. The aim of the project, in effect, was to reach a concentration under the EU limit (11.3 mg N-NO3 L-~), but not excessively low in order to save ethanol, reducing feed costs when applying the system to real scale. From 26 May on, the nitrate (and COD) load were increased by increasing the inflow from 250 to 500 L d-~ in order to test the membrane under different operating conditions. The removal remained very high even when, from 14 June and from 8 July, the loads were raised again, reaching values of about 20 g N-NO3 kg TSS -~ d -t. The permeate flow rate increased from 250 L d -~ to 500 L d -1. A difference in nitrate concentration was observed with an average of 1.6 mg N-NO3 L -1 between permeate and anoxic effluent during the entire experimental period. Nitrite concentration in the treated effluent (Fig. 3), which represents an undesired by-product due to its direct toxicity, was always one order of magnitude below the EU limit (0.15 mg N-NO2 L-~). The limit was mostly respected in the denitrification tank's effluent, too. Only on a few occasions was the nitrite limit
O. Buttiglieri et aL / Desalination 178 (2005) 211-218
CIN ratio
1.8i2. ~ 3 2 . 7 : 2 . 5
COD/N ratio 7.8 19.6', 12
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Time (day) - , - N-NO3" DEN
...........N-NO3" OUT
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Fig. 2. Nitrate concentrationin the treated effluent associated with different C/N ratios. 0.3 "7.
-J 0.2 o
z
o o
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22-jun
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Fig. 3. Nitrite concentration in the treated effluent and in the denitriflcation tank's effluent.
exceeded, and then only when operating problems occurred, for example, when recycling or the inflow pump failed. Fig. 3 very clearly shows the beneficial effect of the aerobic MBR tank to reduce the nitrite concentration. The presence of autothrophic bacteria (ammonia and nitrite oxidizers) was confirmed with respirometric trials [8], and the removal rates obtained were 0.597±0.001 mg N gVSS -t h -t (with K = 0.0193±0.0002 mg N L -~, K = halfsaturation constant) for ammonia and 1.233+ 0.003 m g N gV SS -] h -1 (with K = 0.035 4-
0.001 mg N L -1) for nitrite. The EU limit [N-NO3]/50 + IN-NO2 ]/3 ~ 1 was respected under all conditions in the permeate. Ammonia effluent concentration was always under the EU limit (0.5 mg N-NH~ L-~). Concerning COD, the pilot plant effluent presented a concentration always lower than 10 mg COD L-~ and in most cases lower than 5 mg COD L-L No significant relation between COD removal efficiency and COD loads was detected, probably due to the low values of COD loads and the presence of the aerobic zone that
216
G. Buttiglieri et al. / Desalination 178 (2005) 211-218 120,
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12- jul
Time (day) --,,- COD MBR 0.45 u,m
---*-
CODEffiuent --4,-- g C O D g T S S ' l c r ~
Fig. 4. COD concentration in the permeate and in the aerobic tank (filtered) and COD loads. 35 =_.. 5 ....=
25 15 5
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T i m e (day) •
process pressure
....
clean water pressure
~,
Flow
•
air flow
Fig. 5. Pressure o f the M B R in relation to permeate flow, airflow and suction/relaxation time (7:1 = 7 min suction, 1 rain relaxation).
consumed all the provided substrates. The difference in COD concentration between the filtered (0.45 #m) aerobic mixed liquor and the permeate (Fig. 4) shows the relevant membrane retention capacity to keep back biomass and other organic components over 0.2 #m. The MBR process pressure was monitored during the experiments (Fig. 5) while increasing the water flow up to 500 L d-z, the process period to 10 min against 1 min o f relaxation, and
reducing air flow down to 2 m ~ h -1. The ZW10 behavior was largely regular in comparison with clean water tests made before the beginning for a great part of the trials. On 21 June, nevertheless, a stress event was noticed corresponding to the airflow decrease to 2 m 3 h- ], which affected the pressure significantly by about 10 kPa under the predicted track. Increasing the airflow, however, gradually restored the process pressure.
G. Buttiglieri et al. / Desalination 178 (2005) 211-218
217
160 140 120 •
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Time (day) Fig. 6. Calorimetric response to ethanol pulse: biomass adaptation. Table 1 F/M, maximum nitrate and nitrite removal rates (25°C) and half-saturation constants Date
F/M V~ KNO3 Vm~x KNO~ kg COD kg VSS-1d-' mg N-NO~ g VSS -t h 1 mg N-NO3 L t mg N-NO2 gVSS -I h -1 mg N-NO3 L -I
18 March 25 March 28 April 13 May 24 June 5 July
0.106 0.104 0.092 0.099 0.116 0.146
6.02 7.62 6.50 9.48 15.68 19.65
Total bacterial counts were performed on R2A agar (BD DifcoTM R2A agar low-nutrient medium for enumeration and cultivation of bacteria from potable water). The permeate total count was lower than influent water, and 100% removal was observed for both total coliforms and E. coll. In addition, evaluation ofbiomass adaptation to carbon source and estimation of kinetic parameters were regularly performed by calorimetry. This technique, based on the determination of the heat released by the bacterial culture during substrate degradation, has proven to be very useful for biokinetic characterization of microbial biomass as well as for identification of biological
0.052 0.058 0.104 0.469 0.331 0.184
12.33 11.73 7.65 9.88 13.19 16.08
0.060 0.194 0.164 0.286 0.203 0.124
events such as shifts from one substrate to another or limiting effects [5]. Thermograms (Fig. 6) acquired after ethanol spike (52.5 mg COD) were compared weekly in terms of peak height and duration, thus showing the rapid and progressive adaptation of heterotrophic biomass to substrate. Analogously, experiments were also performed to estimate the maximum substrate consumption rates and half-saturation constants for both nitrate and nitrite reduction. The results are reported in Table 1, together with the COD load (F/M). The maximum nitrate removal rates are also related to the increase in COD loads,
218
G. Buttiglieri et al. / Desalination 178 (2005) 211-218
showing the increase in the active biomass fraction per unit o f VSS. Until 13 May the removal nitrite rate was higher than the one for nitrate (with an average o f 7.40 mg N-NO~ g TSS-~ h-~), allowing no NO~ accumulation. At higher F/M, the nitrate reduction rate was increased to almost 20 mg N-NO3 g TSS-l h-~, while the nitrite removal rate still increased but less than the former. In any case there was no problem of nitrite accumulation as shown before.
4. Conclusions Applying the M B R process to nitrate removal in polluted drinking water sources resulted in a high-quality effluent in terms of nitrate, nitrite and ammonia concentration, with values well below the EC 98/83 Directive standards. After the start-up, the permeate COD concentration was always lower than 10 mg COD L -I and mostly <5 mg COD L -1. The optimum C/N ratio found is 2.2, with a high biomass capacity to remove nitrate, even when loads were increased. Considerable removal o f the total bacterial count was detected with complete removal o f bacterial indicators such as coliforms and E.coli. The permeate extraction flow from the hollow-fiber membrane was in the range of 11.322.6 L m -2 h -~, and during most o f the experimental period the transmembrane pressure was quite close to the value measured under the same operating conditions with lake water as feed. The increase in permeate flow (22.6 Lm -2 h-l), together with a decrease in air flow and the longest suction time, caused a rapid increase in the
suction pressure. However, the situation was quickly corrected by increasing the air flow. In future research, MBRs will be studied for the removal of other types o f microbiological and chemical contaminants found in water resources.
References [1] F. Malpei, L. Bonomo and A. Rozzi, Feasibility study to upgrade a textile WWTP by a hollow fibre MBR for effluent reuse. Water Sci. Tech., 47(10) (2003) 33-39. [2] X. Li and H.P. Chu, Membrane bioreactor for the drinking water treatment of polluted surface water supplies. Water Res., 37 (2003) 4781-4791. [3] A. Nuhoglu, T. Pekdemir, E. Yildiz, B. Keskinler and G. Akay, Drinking water denitrification by a membrane bio-reactor. Water Res., 36 (2002) 11551166. [4] T. Stephenson, S. Judd, B. Jefferson and K. Brindle, Membrane bioreaetors for wastewater treatment. IWA Publishing, London, 2000. [5] E. Daverio, F. Aulenta, J. Ligthart, C. Bassani and A. Rozzi, Application of calorimetric measurements for biokinetic characterization of nitrifying population in activated sludge. Water Res., 37 (2003) 27232731. [6] Standards Methods, 19thed., APHA, AWWA, WEF, United Book Press, Washington, DC, 1995. [7] B. Delanghe, F. Nakamura, H. Myoga, Y. Magara and E. Guibal, Drinking water denitrification in a membrane bioreactor. Water Sci. Tech., 30 (1994) 157-160. [8] M. Ros, Respirometry of activated sludge, Technomic Publishing, Lancaster, PA, 1993. [9] B. Delanghe, F. Nakamura, H. Myoga and Y. Magara, Biological denitrification with ethanol in a membrane bioreactor. Environ. Technol., 15 (1994) 61-70.
ELSEVIER
www.elsevier.com/locate/desal
Denitrification of drinking water sources by advanced biological treatment using a membrane bioreactor Gianluigi Buttiglieri a, Francesca MalpeP*, Emilio Daverio b, Mauro Melchiori a, Hans Nieman b, Jos Ligthart b "Politecnico di Milano, DIIAR, Piazza L. da Vinci 32, 20133 Milano, Italy Tel. +39 (02) 2399-6434; Fax +39 (02) 2399-6499; email: [email protected] bjRC, IES, Inland and Marine Waters Unit, European Commission, TP 300, 21020 Ispra (VA), Italy Received 15 October 2004; accepted 22 November 2004
Abstract Nitrate often contaminates groundwater resources due to excessive use of fertilizers and uncontrolled on-land discharges of raw and treated wastewater and can therefore limit the direct use of groundwater for drinking water purposes. In order to investigate the possible application of a membrane bioreactor (MBR) for denitrification of groundwater, the performance of a pilot-scale MBR was tested as a function of hydraulic and biological parameters. For this purpose synthetic groundwater was prepared by taking lake water (Lago Maggiore, Italy) and adding known amounts of ethanol and sodium nitrate to study the nitrate removal capacity of the sludge, to search for an optimum C/N ratio and to measure filtering ability for micro-organisms through the membrane. The optimum C/N ratio was found at 2.2 gC/gN, resulting in an effluent nitrate concentration within the limits stated in EU Directive 98/83 and the US EPA for drinking water use. The effluent nitrite concentration was one order under the EU limit. The membrane module, Zenon ZW-10, was monitored and performed well except for a short stress episode due to low airflow, afterwards rapidly corrected and thus putting the membrane back to its previous stable behavior. Total bacterial count for the treated effluent was lower than influent water, and 100% removal was observed for both total coliforms and E. coli. Calorimetric thermograms related to heat dissipation due to biological denitrification (nitrate and nitrite consumption) and to substrate adaptation are discussed. A maximum nitrate removal rate close to 20 mgN-NO~ gVSS-1 h- ~was measured in the calorimetric tests.
Keywords: Denitrification; MBR; Drinking water; Calorimetry; Ethanol; Nitrate; Nitrite
*Corresponding author.
Presented at the conference on Membranes in Drinking and Industrial Water Production, L 'Aquila, Italy, 15-17 November 2004. Organized by the European Desalination Society. 0011-9164/05/$- See front matter © 2005 Elsevier B.V. All rights reserved doi:10.1016/j.desal.2004.11.038
212
G. Buttiglieri et al. / Desalination 178 (2005) 211-218
1. Introduction
Drinking water sources (surface and groundwater) may be heavily contaminated by nitrates as a result of excessive fertilization or other pollution-mainly anthropogenic--sources. Elevated nitrate concentration in drinking water is a potential risk to public health since it can directly cause methaemoglobinemia (blue baby syndrome) in infants and its correlation with other diseases, including cancer, is not completely clear. Because of these potential adverse effects, nitrate concentration limits in potable water have been set worldwide. For instance, within the European Union, Council Directive 98/83/EC on the quality of water intended for human consumption establishes a maximum allowable nitrate concentration of 50 mg N O 3L -t. This value is presently not met in a considerable amount of drinking water sources in Europe, so that it is mandatory to remove the excess nitrate from the contaminated water in order to ensure potable use. A membrane bioreactor (MBR) process consists of a biological suspended growth reactor coupled to a membrane acting as a separation device solving some of the problems generally associated with traditional activated sludge processes. The full-scale application of this technique is rapidly increasing in the wastewater treatment field [ 1], whereas very few applications are reported for potable water treatment [2,3]. Coupling of biological treatment and biomass separation can be very efficient for groundwater denitrification. Since microfiltration or ultrafiltration membranes can retain almost the totality of the bacterial cells, MBRs may overcome the inconvenience of microbiological contamination of treated water, almost completely reducing the number of post-treatment steps. At the same time, due to their independence from limiting factors such as the sludge age, or the ability of bacteria to form biological aggregates, MBRs favor the development and establishment of a selected biomass in the reactor, which can also act by
removing slowly biodegradable refractory organic contaminants [4]. In this work a Zenon Zee Weed ® 10 (ZW10) MBR pilot plant was used for biological nitrate removal from drinking water sources. The pilot plant was designed in two parts: an anoxic tank where nitrate is biologically reduced to nitrogen, followed by an aerobic tank for excessive carbon removal, containing the ZW 10 filtering unit. The influent was sand-filtered lake water (Lago Maggiore, Italy), spiked with nitrate in order to reach a concentration of 30 mg N L -~. Ethanol was used as carbon source at different C/N ratios.
2. Materials and methods 2.1. Pilot plant
A diagram of the pilot plant used for this research is shown in Fig. 1. The first anoxic tank (200 L) includes a submersible pump to mix the sludge and a gravity drain to the second aerobic tank. Dissolved oxygen and pH were monitored with probes linked to a bio-controller and acquired every 5 min with a PC. The second aerobic tank (190 L) includes a drain valve, a sludge recycling pump device, transmembrane pressure and temperature sensors and overflow valves. The ZW 10 membrane unit timely controls the relaxation or back-wash process by a control panel. Process parameters at start-up were 5 min of suction and 1 min of relaxation (no permeate extraction), increased up to 10 min of operation and 1 min of relaxation. The nominal porosity of the membrane was 0.2 #m and the nominal membrane surface area was 0.93 m 2. The module dimensions were I10 mm wide and almost 700 mm high. Air flow at the bottom of the membrane was maintained in the range of 2-3.5 m 3 h-1 to control fouling, thus resulting in DO concentration in the tank and in the recycle to the anoxic reactor in the range of 6-8 mg L-~. Temperature in the aerobic tank was in the range of 22-24°C during the entire experimental period.
213
G. Buttiglieri et al. / Desalination 178 (2005)211-218
Lake water + NO 3 Dissolved
Ethanol
\
I
TREATEDEFFLUENT
oxygen 14
1 Blocont roller E
Air pump Fig. 1. Pilot plant diagram.
Hydraulic retention times (HRT) varied from 19 to 37 h. The recycle flow was set at a value equal to the influent flow (250-500 L d-l).
2.2. Prepared solutions Ethanol and sodium nitrate were dosed in order to achieve a 30 mg N-NO~ L -I concentration in the influent (with 7.3 g N-NO 3 kg TSS -l d -l and 52 g COD kg TSS -I d -I) and a start-up C/N ratio of 1.8 (7.2 COD/N ratio). The prepared solutions were stored in separated 10 L tanks and fed with two peristaltic pumps.
2.3. Biomass source Inoculum biomass was taken from a municipal wastewater treatment plant. The average anoxic and aerobic sludge concentrations were 1.6 g TSS L-l and 3.2 g TSS L-l, respectively, which were increased to 2 g TSS L -l for the anoxic sludge and 4.1 g TSS L-~ for the aerobic sludge at the final stage of the experiments.
2.4. Water inflow The water inflow was sand-filtered Lago Maggiore water (average characteristics: TSS <0.1 mg L -l, N-NO3 <1 mg L -l, N-NH 4 <0.3 mg L -I, pH = 7.04, DO = 7.8 mg L-l). 2.5. Calorimeter Sludge properties such as substrate adaptation and nitrate and nitrite removal speed were calorimetrically investigated. The Bio-RC1 is a standard 2 L jacketed glass calorimeter with a sensitivity of5-10 mW L-l [5]. In the isothermal mode, a low-viscosity silicone oil is pumped at a high rate (2 L s -l) through the reactor jacket in order to maintain a constant temperature of the reactant medium (Tr) by proper control of the jacket temperature (Tj). The jacket temperature is carefully controlled by blending oil from a "hot" and a "cold" oil circuit via an electronically controlled metering valve. Therefore, when a process dissipates or takes up heat, Tj increases or decreases, respectively. The resulting gradient
214
G. Buttiglieri et al. / Desalination 178 (2005) 211-218
across the reactor wall is directly proportional to the thermal power liberated or absorbed by the process. When a sludge sample is put into the calorimeter, a specific parameter is determined using an internal calibration heater of known power input. The Bio-RC1 does not require any energy balance to be computed and, therefore, it is possible to carefully monitor and control important process variables such as temperature and medium agitation and to operate the vessel under constant and reproducible process conditions. In the anoxic tests a 1.5 L sludge volume was used, bubbled with a CO2 and N2 gas mixture, at a constant 25°C temperature. Once a stable heat flux baseline in endogenous conditions was obtained, a known amount of NO 3 or of ethanol was added to the sludge. The specific heat power (Q) measured is related to the rate of substrate degradation by the equation Q = - YQ/s " d S / d t
[W L-']
where S is the substrate concentration (mgsubstrat e L-l) and Ye/s is the energy generated for the unit substrate degrade (J mgsubs~ate-1),with measured values in the range of 20-35 J (rag N-NOx) -1 for nitrate and nitrite removal.
2. 6. Analytical methods
Nitrate was determined spectrophotometrically at 220 nm and 275 nm. Ammonia and COD analyses were carried out by means of a spectrophotometric test (Spectroquant Merck). In the beginning nitrite was measured by means of a spectrophotometric test (Spectroquant Merck) and later determined spectrophotometrically at 543 nm. Total and volatile solids determinations were performed according to Standards Methods [6]. All chemicals were reagent grade.
3. Results and discussion Nitrate concentration in the permeate and in the denitrification tank monitored during the experiments is presented in Fig. 2 together with the different C/N ratios applied. A start-up C/N ratio of 1.5 using ethanol as substrate was chosen [7], but it was increased to 1.8 gC/gN in order to cope with the DO coming in with the influent and the recycling flow. In the first 2 weeks a necessary period for biomass adaptation to new conditions was required; in fact, effluent nitrate concentrations were high above the normative limits (CD/98/83/EC). Then, with a C/N ratio of 2.4 before and 3.0 after, complete nitrate removal was achieved. Between 24 March and 5 April, with a 2.7 C/N ratio, nitrate concentration was null in the anoxic tank but showed a slight increase in the permeate. With a C/N ratio of 2.2 at the end, an optimum was found. The aim of the project, in effect, was to reach a concentration under the EU limit (11.3 mg N-NO3 L-~), but not excessively low in order to save ethanol, reducing feed costs when applying the system to real scale. From 26 May on, the nitrate (and COD) load were increased by increasing the inflow from 250 to 500 L d-~ in order to test the membrane under different operating conditions. The removal remained very high even when, from 14 June and from 8 July, the loads were raised again, reaching values of about 20 g N-NO3 kg TSS -~ d -t. The permeate flow rate increased from 250 L d -~ to 500 L d -1. A difference in nitrate concentration was observed with an average of 1.6 mg N-NO3 L -1 between permeate and anoxic effluent during the entire experimental period. Nitrite concentration in the treated effluent (Fig. 3), which represents an undesired by-product due to its direct toxicity, was always one order of magnitude below the EU limit (0.15 mg N-NO2 L-~). The limit was mostly respected in the denitrification tank's effluent, too. Only on a few occasions was the nitrite limit
O. Buttiglieri et aL / Desalination 178 (2005) 211-218
CIN ratio
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22- jun
17 jul
Time (day) - , - N-NO3" DEN
...........N-NO3" OUT
- . - gN-NO3" gTSS 1 d "1
Fig. 2. Nitrate concentrationin the treated effluent associated with different C/N ratios. 0.3 "7.
-J 0.2 o
z
o o
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Fig. 3. Nitrite concentration in the treated effluent and in the denitriflcation tank's effluent.
exceeded, and then only when operating problems occurred, for example, when recycling or the inflow pump failed. Fig. 3 very clearly shows the beneficial effect of the aerobic MBR tank to reduce the nitrite concentration. The presence of autothrophic bacteria (ammonia and nitrite oxidizers) was confirmed with respirometric trials [8], and the removal rates obtained were 0.597±0.001 mg N gVSS -t h -t (with K = 0.0193±0.0002 mg N L -~, K = halfsaturation constant) for ammonia and 1.233+ 0.003 m g N gV SS -] h -1 (with K = 0.035 4-
0.001 mg N L -1) for nitrite. The EU limit [N-NO3]/50 + IN-NO2 ]/3 ~ 1 was respected under all conditions in the permeate. Ammonia effluent concentration was always under the EU limit (0.5 mg N-NH~ L-~). Concerning COD, the pilot plant effluent presented a concentration always lower than 10 mg COD L-~ and in most cases lower than 5 mg COD L-L No significant relation between COD removal efficiency and COD loads was detected, probably due to the low values of COD loads and the presence of the aerobic zone that
216
G. Buttiglieri et al. / Desalination 178 (2005) 211-218 120,
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23- may
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Time (day) --,,- COD MBR 0.45 u,m
---*-
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Fig. 4. COD concentration in the permeate and in the aerobic tank (filtered) and COD loads. 35 =_.. 5 ....=
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clean water pressure
~,
Flow
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Fig. 5. Pressure o f the M B R in relation to permeate flow, airflow and suction/relaxation time (7:1 = 7 min suction, 1 rain relaxation).
consumed all the provided substrates. The difference in COD concentration between the filtered (0.45 #m) aerobic mixed liquor and the permeate (Fig. 4) shows the relevant membrane retention capacity to keep back biomass and other organic components over 0.2 #m. The MBR process pressure was monitored during the experiments (Fig. 5) while increasing the water flow up to 500 L d-z, the process period to 10 min against 1 min o f relaxation, and
reducing air flow down to 2 m ~ h -1. The ZW10 behavior was largely regular in comparison with clean water tests made before the beginning for a great part of the trials. On 21 June, nevertheless, a stress event was noticed corresponding to the airflow decrease to 2 m 3 h- ], which affected the pressure significantly by about 10 kPa under the predicted track. Increasing the airflow, however, gradually restored the process pressure.
G. Buttiglieri et al. / Desalination 178 (2005) 211-218
217
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Time (day) Fig. 6. Calorimetric response to ethanol pulse: biomass adaptation. Table 1 F/M, maximum nitrate and nitrite removal rates (25°C) and half-saturation constants Date
F/M V~ KNO3 Vm~x KNO~ kg COD kg VSS-1d-' mg N-NO~ g VSS -t h 1 mg N-NO3 L t mg N-NO2 gVSS -I h -1 mg N-NO3 L -I
18 March 25 March 28 April 13 May 24 June 5 July
0.106 0.104 0.092 0.099 0.116 0.146
6.02 7.62 6.50 9.48 15.68 19.65
Total bacterial counts were performed on R2A agar (BD DifcoTM R2A agar low-nutrient medium for enumeration and cultivation of bacteria from potable water). The permeate total count was lower than influent water, and 100% removal was observed for both total coliforms and E. coll. In addition, evaluation ofbiomass adaptation to carbon source and estimation of kinetic parameters were regularly performed by calorimetry. This technique, based on the determination of the heat released by the bacterial culture during substrate degradation, has proven to be very useful for biokinetic characterization of microbial biomass as well as for identification of biological
0.052 0.058 0.104 0.469 0.331 0.184
12.33 11.73 7.65 9.88 13.19 16.08
0.060 0.194 0.164 0.286 0.203 0.124
events such as shifts from one substrate to another or limiting effects [5]. Thermograms (Fig. 6) acquired after ethanol spike (52.5 mg COD) were compared weekly in terms of peak height and duration, thus showing the rapid and progressive adaptation of heterotrophic biomass to substrate. Analogously, experiments were also performed to estimate the maximum substrate consumption rates and half-saturation constants for both nitrate and nitrite reduction. The results are reported in Table 1, together with the COD load (F/M). The maximum nitrate removal rates are also related to the increase in COD loads,
218
G. Buttiglieri et al. / Desalination 178 (2005) 211-218
showing the increase in the active biomass fraction per unit o f VSS. Until 13 May the removal nitrite rate was higher than the one for nitrate (with an average o f 7.40 mg N-NO~ g TSS-~ h-~), allowing no NO~ accumulation. At higher F/M, the nitrate reduction rate was increased to almost 20 mg N-NO3 g TSS-l h-~, while the nitrite removal rate still increased but less than the former. In any case there was no problem of nitrite accumulation as shown before.
4. Conclusions Applying the M B R process to nitrate removal in polluted drinking water sources resulted in a high-quality effluent in terms of nitrate, nitrite and ammonia concentration, with values well below the EC 98/83 Directive standards. After the start-up, the permeate COD concentration was always lower than 10 mg COD L -I and mostly <5 mg COD L -1. The optimum C/N ratio found is 2.2, with a high biomass capacity to remove nitrate, even when loads were increased. Considerable removal o f the total bacterial count was detected with complete removal o f bacterial indicators such as coliforms and E.coli. The permeate extraction flow from the hollow-fiber membrane was in the range of 11.322.6 L m -2 h -~, and during most o f the experimental period the transmembrane pressure was quite close to the value measured under the same operating conditions with lake water as feed. The increase in permeate flow (22.6 Lm -2 h-l), together with a decrease in air flow and the longest suction time, caused a rapid increase in the
suction pressure. However, the situation was quickly corrected by increasing the air flow. In future research, MBRs will be studied for the removal of other types o f microbiological and chemical contaminants found in water resources.
References [1] F. Malpei, L. Bonomo and A. Rozzi, Feasibility study to upgrade a textile WWTP by a hollow fibre MBR for effluent reuse. Water Sci. Tech., 47(10) (2003) 33-39. [2] X. Li and H.P. Chu, Membrane bioreactor for the drinking water treatment of polluted surface water supplies. Water Res., 37 (2003) 4781-4791. [3] A. Nuhoglu, T. Pekdemir, E. Yildiz, B. Keskinler and G. Akay, Drinking water denitrification by a membrane bio-reactor. Water Res., 36 (2002) 11551166. [4] T. Stephenson, S. Judd, B. Jefferson and K. Brindle, Membrane bioreaetors for wastewater treatment. IWA Publishing, London, 2000. [5] E. Daverio, F. Aulenta, J. Ligthart, C. Bassani and A. Rozzi, Application of calorimetric measurements for biokinetic characterization of nitrifying population in activated sludge. Water Res., 37 (2003) 27232731. [6] Standards Methods, 19thed., APHA, AWWA, WEF, United Book Press, Washington, DC, 1995. [7] B. Delanghe, F. Nakamura, H. Myoga, Y. Magara and E. Guibal, Drinking water denitrification in a membrane bioreactor. Water Sci. Tech., 30 (1994) 157-160. [8] M. Ros, Respirometry of activated sludge, Technomic Publishing, Lancaster, PA, 1993. [9] B. Delanghe, F. Nakamura, H. Myoga and Y. Magara, Biological denitrification with ethanol in a membrane bioreactor. Environ. Technol., 15 (1994) 61-70.