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Effects of temperature and dissolved oxygen on hydrolysis of sewer solids
PII: S0043-1354(99)00032-9
Wat. Res. Vol. 33, No. 14, pp. 3119±3126, 1999 # 1999 Elsevier Science Ltd. All rights reserved Printed in Great Britain 0043-1354/99/$ - see front matter
www.elsevier.com/locate/watres
EFFECTS OF TEMPERATURE AND DISSOLVED OXYGEN ON HYDROLYSIS OF SEWER SOLIDS M M JES VOLLERTSEN1** , MARIA DO CEÂU ALMEIDA2* and M THORKILD HVITVED-JACOBSEN1*
1 Environmental Engineering Laboratory, Aalborg University, Sohngaardsholmsvej 57, DK-9000 Aalborg, Denmark; 2Department of Civil Engineering, Imperial College, Imperial College Road, London SW7 2BU, UK
(First received June 1998; accepted in revised form January 1999) AbstractÐEects of temperature and dissolved oxygen (DO) on kinetics of microbial transformation processes of suspended sewer sediment particles, suspended wastewater particles and wastewater were investigated. Microbial activity was estimated by the oxygen uptake rate. The eects were studied under conditions where biomass had not acclimated to changed temperature or DO conditions. Hydrolysis processes limited the availability of organic substrates for the aerobic microbial growth. The eects of temperature were interpreted applying an Arrhenius relationship. The average Arrhenius constants found for sewer sediment particles and wastewater particles diered signi®cantly from the average Arrhenius constant found for wastewater. The eects of dissolved oxygen concentration were interpreted applying a saturation type relationship. No dierences between sewer sediment particles, wastewater particles and wastewater were found for the oxygen saturation coecients. # 1999 Elsevier Science Ltd. All rights reserved Key wordsÐtemperature, dissolved oxygen, hydrolysis, suspended sewer sediments, wastewater particles, wastewater, microbial processes, process kinetics
NOMENCLATURE
a KO OUR SO SO, 1 T#
Arrhenius constant (ÿ) Saturation constant for dissolved oxygen concentration (g O2 mÿ3) Oxygen uptake rate (g O2 mÿ3 hÿ1) Dissolved oxygen concentration (g O2 mÿ3) Nonlimiting dissolved oxygen concentration (g O2 mÿ3) Temperature (8C)
INTRODUCTION
Aerobic microbial transformations of organic matter in sewer systems can be interpreted and simulated using conceptual models addressing process kinetics and stoichiometry of heterotrophic biomass (Bjerre et al., 1995, 1998; Hvitved-Jacobsen et al., 1998a, b,; Vollertsen and Hvitved-Jacobsen, 1998, 1999; Vollertsen et al., 1998). In these models it is assumed that the biomass utilizes organic substrate and dissolved oxygen (DO) for growth and Ð depending on the concept Ð also DO for nongrowth-related energy requirements. The DO concentration and the temperature aect the rates of *Author to whom all correspondence should be addressed. [Tel.: +45-963-58-080; Fax: +45-98-142-555].
aerobic microbial processes. Implementing these conceptual models for simulation of natural systems subject to rapid changing temperature and DO conditions, requires information regarding the in¯uence of short term temperature and DO variations upon the simulated microbial transformations. These concepts may be applied for dierent conditions, i.e. for wastewater, where large fractions of readily biodegradable and hydrolyzable substrates exist together with high concentrations of heterotrophic biomass, wastewater particles and organic matter in sewer sediments where the dominating compounds are slowly hydrolyzable substrate and where heterotrophic biomass is present in only low concentrations. When simulating in-sewer microbial transformation, sewer sediment particles, wastewater particles and wastewater are likely to interchange. Also when assessing combined sewer over¯ow (CSO) impacts, it is a mixture of discharged sewer sediment particles, wastewater particles and diluted wastewater that cause the DO impacts. Relative responses to transient DO and temperature variations are, as an example, of importance. The oxygen uptake rate (OUR) is considered to be the key parameter to be determined for estimation of aerobic microbial transformation of organic matter according to the model concept which
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is used. Therefore, eects of DO and temperature on the OUR are important. The eects of temperature and DO have been studied under conditions where only hydrolyzable substrates are utilized as electron donor for microbial transformations. This is in agreement with concepts presented by e.g. Bjerre et al. (1995), Hvitved-Jacobsen et al. (1998a; Hvitved-Jacobsen and Vollertsen, 1999), Vollertsen and HvitvedJacobsen (1998) and Vollertsen et al. (1998). Furthermore, the hydrolysis processes governing the microbial transformations only allowed for heterotrophic biomass growth rates signi®cantly lower than the maximum growth rates; even negative growth might occur. Therefore, microbial responses to temperature and DO was interpreted as eects of temperature and DO on the hydrolysis processes. Responses of sewer sediment particles, wastewater particles and wastewater due to short-term changes in temperature and DO concentration were studied because sewer solids as well as wastewater may be subject to such changes. Short-term changes of temperature and DO concentration occur for example during transport in sewers and in receiving waters following a CSO discharge. METHODS
Sampling Wastewater and wastewater particles were sampled in Frejlev, Denmark, at the Frejlev Research and Monitoring Station (Schaarup-Jensen et al., 1998). Frejlev has a combined sewer system, 2000 inhabitants with no signi®cant industry and a reduced (impermeable) area of approximately 30 ha. As the catchment is situated on a hillside, with the highest part approximately 40 m above the Research and Monitoring Station, the slope of the sewer is generally high and no signi®cant deposition of sediment occurs in the sewers. In the Frejlev Research and Monitoring Station, the sewer is divided into modules
which are accessible from all sides. In two situations wastewater particles and wastewater were sampled in a pumping station located at an intercepting pipe from a small residential town, Vestbjerg, to the wastewater treatment plant Aalborg East, Denmark. During sampling the sewer pipe in the monitoring station was opened so that large quantities of completely mixed wastewater ¯owed out. Wastewater was collected with buckets and transferred into a 120-litre cylindrical tank. When the tank was full, the wastewater was allowed to settle for 30 min and was subsequently decanted. The settled wastewater was transported to the laboratory for measurement immediately after decanting. Sewer sediments were sampled from a 1000-mm concrete combined sewer pipe in an older residential part of the city of Aalborg, Denmark. The sewer serves approximately 20 households upstream of the sampling locations. Sampling was done using a sewer sediment shovel from three adjacent manholes. In one situation sewer sediments were sampled from a silt trap immediately downstream of the above mentioned pumping station in Vestbjerg. Total solids (TS) and volatile solids (VS) of sewer sediment particles, wastewater particles and wastewater were determined according to Standard Methods (1995). Oxygen uptake rate measurement, OUR Oxygen uptake rates of wastewater and suspended particles or sediments were determined in a series of six stainless steel batch reactors with volumes of 2.2 litres each (Fig. 1). Tap water was added to the samples of particles and sediments up to the reactor volume. The reactors were kept at the required temperature by circulating water from one temperature controlled water bath through caps enclosing approximately 30% of the apparatus surface area. When not determining the temperature dependency of process kinetics, temperature was kept at 20 20.58C. Particles were kept in suspension by using magnetic stirrers. DO of wastewater or suspensions was measured using INGOLD oxygen sensors with 12-mm diameter Te¯on membranes. When the DO concentration was below a preset value, an aeration cycle was started by automatically opening the closing piston between the reactor and the expansion chamber. This process was followed by compressed air injection into the wastewater or particle suspension during a preset time interval. When not determining the in¯uence of DO on the OUR kinetics, this set-
Fig. 1. OUR measuring equipment. Six devices were operated in parallel and temperature controlled by the same water circulation, allowing for six parallel batch experiments under the same temperature conditions.
Hydrolysis of sewer solids point was 60% of air saturation. Once the aeration was ®nished, the piston was closed after a preset time delay to ensure that no air bubbles were left in the suspension. Temperature was measured in one of the six batch reactors using a PT100 sensor with a resolution of 0.18C. The temperature in the other ®ve reactors was assumed the same. Time, DO concentration and temperature were logged with a computer. The OUR was calculated from these measurements by linear regressions. Each experiment consisted of duplicate OUR determinations of wastewater, sewer sediment particles and wastewater particles. This resulted in six OUR measurements all performed at the same temperature. Each set of experiments started with a 24-h aeration period, in order to guarantee that only aerobic microbial transformations of hydrolyzable substrates contributed to the OUR during the experiments. Temperature and DO eects on hydrolysis of sewer solids were determined subsequently on the same sample. Eects of temperature on process kinetics Temperature eects of process kinetics was determined by running the OUR experiments under a sequence of periods with dierent reactor temperatures. After an initial period of one day at 208C, the temperature was increased to 23±248C, secondly decreased to 15±168C, then to 9± 108C and ®nally it was increased to the initial temperature of about 208C. At each temperature level the temperature was kept constant for typically 2 h, Fig. 2. Depending on
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the OUR value, 1±2 h of constant temperature was needed. Higher temperatures resulted in higher rates and, hence, the time required for stabilization of the OUR at a new temperature level was shortest for higher temperatures. From the OUR values at adjacent temperature levels, temperature eects on kinetics of aerobic microbial hydrolysis of organic substrates were determined. The interpretation of measured eects of temperature on process kinetics is based on equation (1), presented by Streeter and Phelps (1925) to describe the dependency of wastewater ®rst order deoxygenation kinetics upon temperature. Theriault (1927) showed that there was similarity between the Arrhenius equation for chemical reactions and equation (1). Therefore, equation (1) often simply is called the Arrhenius relationship and a referred to as the Arrhenius constant (Zanoni, 1967). The well-known Arrhenius relationship has traditionally been used in many microbial contexts; e.g. activated sludge (Henze et al., 1987; Barker and Dold, 1997), bio®lms (Characklis and Marshall, 1990), wastewater transformations in gravity sewers (Nielsen et al., 1992; Hvitved-Jacobsen et al., 1998b) and hydrogen sul®de formations in pressure sewers (Boon and Lister, 1974; Nielsen et al., 1998). For this reason, the Arrhenius relationship will be assumed in this study for interpretation of measured temperature eects on process kinetics. OUR T1 a T1 ÿT2 : OUR T2
1
Eects of DO on process kinetics Following completion of the temperature variation period, and after the temperature had been stabilized at 208C, the determination of DO limitation eect on kinetics of aerobic microbial transformations of hydrolyzable substrates was determined. This was achieved by changing the set-point de®ning the end of a measuring period from 60% of air saturation down to 0.1% of air saturation. The resulting DO interval, on which eects of DO concentration on process kinetics was determined, was from 7±9 g O2 mÿ3 down to about 0.01 g O2 mÿ3. These measurements were carried out for 15±20 h. DO limitation eects on kinetics of microbial transformations, under conditions where electron donor availability was limited by hydrolysis, were determined from calculated OUR vs DO values. Eects of DO on microbial transformations of organic substrates, for purposes of modeling microbial transformations of organic substrate in gravity sewers and for activated sludge process modeling, have been simulated with saturation type equations (Henze et al., 1987, 1995; Hvitved-Jacobsen et al., 1998a, b; Bjerre et al., 1998). Therefore, a saturation type equation (Monod) will be assumed in this study equation (2). OUR SO OUR SO, 1
SO : KO S O
2
RESULTS AND DISCUSSION
Fig. 2. Duplicate determination of eects of temperature upon oxygen uptake rate (OUR) for wastewater sampled September 19, 1997. Resulting average Arrhenius constants a were 1.11, 1.10, 1.13 and 1.12 for the temperature changes 20.5±23.6, 23.6±16.3, 16.3±10.3 and 10.3±20.58C, respectively.
For each of the three types of sewage components investigated Ð sewer sediment particles, wastewater particles and wastewater Ð 11 samples were analyzed for the eect of temperature and DO on the OUR kinetics. Kinetics were determined in duplicate on each sample. However, two samples were analyzed only once because of equipment malfunction.
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Table 1. Arrhenius constants a for sewer sediment particles, wastewater particles and wastewater Samples Determinations Average Std. of a a's dev. Sewer sediment particles Wastewater particles Wastewater
11 11 11
43 48 46
1.091 1.089 1.099
0.015 0.016 0.018
Eect of temperature variation on process kinetics Temperature eects on process kinetics were determined from the change in OUR response for each change in temperature level. As an example, the measurements presented in Fig. 2 resulted in duplicate determination of the Arrhenius constant for the four temperature changes: 20.5±23.6, 23.6± 16.3, 16.3±10.3 and 10.3±20.58C. Isolating a from equation (1) and calculating a for all OUR responses to every change in temperature level, duplicate determinations of the Arrhenius constants were obtained (Table 1). The variability of the Arrhenius constants determined was large, with extreme values of 1.053 and 1.142 (Fig. 3). No systematic variability of this constant with temperature was observed. The standard deviation of the duplicate determination was, how-
ever, considerably lower than the standard deviation of the whole sample. The Arrhenius constants obtained where normally distributed for the three types of organic matter investigated (Fig. 3). No signi®cant dierence in the average values of a was found between wastewater particles and sewer sediment particles. However, the average Arrhenius constants for wastewater diered signi®cantly from the average constants found for both wastewater particles and sewer sediment particles. Describing oxygen consumption of wastewater and river water, Streeter and Phelps (1925) found an average Arrhenius constant of 1.047 in the temperature interval 10±37.58C. Zanoni (1967) described oxygen consumption of wastewater at dierent temperatures, measuring BOD at dierent incubation periods. He found that a ®rst order description of the oxygen consumption had shortcomings, primarily because of very high oxygen demands in the beginning of the incubation periods for incubation temperatures of 208C and above. Excluding these high, early oxygen demands, he found an average a of 1.126 for the temperature range 2±158C, of 1.047 for 15±328C and of 0.985 for 32±408C. Looking at the high, initial oxygen demand observed by Zanoni (1967) in the light of
Fig. 3. Normal probability plots of Arrhenius constants (a). The error bars indicate the standard deviation of the duplicate determinations for sewer sediments, wastewater particles and wastewater, respectively.
Hydrolysis of sewer solids
recent studies of wastewater microbial transformations (Bjerre et al., 1995; Hvitved-Jacobsen et al., 1998a), these high oxygen demands probably were due to an initial amount of readily biodegradable organic substrate. When Zanoni (1967) excluded these measurements, the limiting process for the temperature kinetics could be interpreted as hydrolysis, i.e. comparable with the experimental conditions in this study. The authors of the activated sludge model No. 2 (Henze et al., 1995) suggest an Arrhenius constant of 1.07 for all heterotrophic processes. Ekama et al. (1984) have found that the oxygen demand of activated sludge decreases by 5% with a 68C change in temperature, i.e. an Arrhenius constant of 1.009. The Arrhenius constants that have been found in this study are within the range of the ®ndings of Zanoni (1967) for wastewater. However, the Zanoni (1967) ®ndings of a decreasing Arrhenius constant for increasing temperature could not be con®rmed. The value for activated sludge modeling, recommended by the authors of the activated sludge model No. 2 (Henze et al., 1995), is somewhat lower than the results obtained in this study. However, that recommendation is made on the basis that biomass is allowed to acclimatize to the new temperature. This probably results in a change in biomass composition and properties from one temperature level to another, resulting in a dierent a-value. In the present experiments, the biomass was acclimated for 24 h to 208C and its composition during the following hours of experiment expected unchanged. The type of temperature change addressed in this study is expected to occur in sewers. Temperature reductions up to 3.58C were observed during transport in a 5-km long intercepting gravity sewer (Hvitved-Jacobsen and Vollertsen, 1999) and according to the present results this decrease in temperature would correspond to a 28% reduction in wastewater hydrolysis rates during transport. Eects of dissolved oxygen concentration on process kinetics Eects of dissolved oxygen concentration on process kinetics were determined from DO measurements vs calculated OUR for the DO concentration range from 7±9 g O2 mÿ3 down to 0.01 g O2 mÿ3. The number of aeration cycles used by the OURmeasurement equipment depended on the OUR, as the total duration of these experiments were predetermined. Up to four cycles were selected from each batch for simulation with equation (2). OUR at nonlimiting DO levels changed somewhat with time. To minimize errors in KO values due to these variations, cycles where OUR varied considerably at nonlimiting DO levels were omitted. Some typical curves of DO vs OUR are presented in Fig. 4 together with simulations based on equation (2). At low DO values, simulated oxygen
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uptake rates tend to be lower than the measured rates. Furthermore, in some simulations the observed increase in OUR for high DO values was not simulated well, Fig. 4(b). This observation raises the question of whether a saturation type expression is appropriate for simulation of OUR response to decreasing DO concentrations. The argument was, that some of the biomass utilizing DO as electron acceptor might be located within aggregates or in larger particles where DO diusion might be limiting the DO uptake process. If such a process was comparable to bio®lm processes, this should result in some of the DO uptake processes to be better described by a 1/2 order rate expression (Jansen and HarremoeÈs, 1984). Simulation of OUR using a 1/2 order rate expression in DO was tried. However, the results were not better than with equation (2). The hypothesis that only a part of the oxygen uptake followed a 1/2 order rate expression in DO was tested by means of simulation with the sum of equation (2) and a 1/2 order rate expression. This resulted in better agreement between measurements and simulations. However, for a DO level of 1 g O2 mÿ3, equation (2) was on average accounting for 96, 98 and 91% of the OUR measured for sewer sediment particles, wastewater particles and wastewater, respectively. Therefore, the 1/2 order rate expression in DO could only explain a minor part of the measured OUR. Because two dierent expressions would increase the complexity of the simulation, only the saturation type expression (2) was used in the present study. Saturation constants that were determined for each sample showed a rather high standard deviation compared to the standard deviation between dierent samples (Table 2) Ð in average the standard deviation of each sample was 0.04. As the samples have been treated identically using identical equipment, the reason for the poor quality of duplicate determination cannot be explained. The saturation constants for duplicate determinations were found to be normally distributed (Fig. 5). Comparison between average values of KO for the three dierent types of organic matter showed Ð at a 5% signi®cance level and using a t-test Ð no signi®cant dierences between wastewater particles, wastewater or sewer sediment particles. The average KO value for all samples was 0.080 g O2 mÿ3 with a standard deviation of 0.036 g O2 mÿ3. Henze et al. (1995) state that typical values of KO for modeling of activated sludge heterotrophic processes is 0.20 g O2 mÿ3, i.e. somewhat higher than what was found here. Using the same methodology for activated sludge Kappeler and Gujer (1992) found that a saturation type expression is a rather good approximation of the DO dependency for OUR. They found an average KO of 0.25 g O2 mÿ3. For wastewater Bjerre et al. (1998) and GudjoÂnson
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Fig. 5. Normal probability plots of saturation constants KO. The error bars indicate the standard deviation of the duplicate determinations for sewer sediment particles, wastewater particles and wastewater.
et al. (in preparation) found an average KO of 0.05 g O2 mÿ3. Dierences in particle size between activated sludge and the studied particles might explain the somewhat lower KO values found in this study. CONCLUSION
Eects of temperature and dissolved oxygen concentration (DO) on oxygen uptake rate (OUR) of suspended sewer sediment particles, suspended wastewater particles and wastewater have been investigated and compared. Information regarding such eects is required when using models for simulation of microbial transformations of organic substrates, for simulation of transformation processes in sewer systems or for impacts on receiving waters. Eects of temperature and DO concentration on Table 2. Saturation constants KO determined using equation (2) for sewer sediments, wastewater particles and wastewater
Sewer sediment particles Wastewater particles Wastewater
Samples
Average KO
Std. dev.
11 11 11
0.065 0.087 0.088
0.011 0.040 0.040
OUR have been studied under conditions where availability of the electron donor is limited by hydrolysis and thus also limiting the OUR. Temperature eects on process kinetics were described using an Arrhenius expression, resulting in an average Arrhenius constant of 1.091 for suspended sewer sediments, of 1.089 for suspended wastewater particles and of 1.099 for wastewater. The average Arrhenius constant found for wastewater diered signi®cantly from the two particle suspensions. Investigating eects of DO on OUR, measurements of DO vs calculated OUR were simulated using a saturation type expression. For the calculated saturation coecients no signi®cant dierence between suspended sewer sediment particles, suspended wastewater particles and wastewater was found. The resulting average saturation constant for suspended particles and wastewater was 0.081 g O2 mÿ3. The eects of temperature and DO on microbial transformations were investigated in the short-term, where the biomass did not have time to acclimate. Short-term temperature or DO changes are normal events in sewer systems as well as in the interactions between sewer system and receiving waters. If bio-
Hydrolysis of sewer solids
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Fig. 4. Typical curves of OUR vs DO concentration and simulations based on equation (2). (a) Sewer sediment particles, 15/9/97, (b) wastewater, 8/10/97, (c) wastewater particles, 3/10/97.
mass is allowed to acclimate, rates are likely to be dierent. This may explain the dierences from saturation coecients found in other studies. Extreme changes in temperature, followed by biomass acclimation, might favor the development of other microorganisms, resulting in a change in biomass composition. Other kinetic parameters would be needed under such circumstances. The ®ndings of this study should, therefore, only be used when modeling short-term changes in temperature or DO. AcknowledgementsÐFinancial support for this research project was provided by the Danish Technical Research Council, the framework program on `Solids in Sewage Systems'. REFERENCES
Barker P. S. and Dold P. L. (1997) General model for biological nutrient removal activated-sludge systems: model presentation. Water Env. Res. 69(5), 969±984. Boon A. G. and Lister A. R. (1974) Formation of Sulphide in Rising Main Sewers and Its Prevention. Water Pollution Research Laboratory, Stevenage, U.K. Bjerre H. L., Hvitved-Jacobsen T., TeichgraÈber B. and te Heesen D. (1995) Experimental procedures characterizing transformations of wastewater organic matter in the Emscher river, Germany. Water Sci. Tech. 31(7), 201± 212. Bjerre H. L., Hvitved-Jacobsen T., TeichgraÈber B. and Schlegel S. (1998). Modelling of aerobic wastewater transformations under sewer conditions in the Emscher river, Germany. Water Environ. Res., in press.
Characklis W. G. and Marshall K. C. (1990) Bio®lms. In Wiley Series in Ecological and Applied Microbiology, eds. W. G. Characklis and K. C. Marshall U.S.A. Ekama G. A., Marais G. v. R., Siebritz I. P., Pitman A. R., Keay G. F. P., Buchan L., Gerber A. and Smollen M. (1984) Theory, Design and Operation of Nutrient Removal Activated Sludge Processes. Water Research Commission, Pretoria, South Africa. GudjoÂnson G., Hvitved-Jacobsen T., Vollertsen J. and Nielsen C. R. (in preparation). Microbial transformations in gravity sewers: variability and modeling of dissolved oxygen. Henze M., Grady C. P. L., Jr., Gujer W., Marais G. v. R. and Matsuo T. (1987) Activated Sludge Model No. 1. Scienti®c and Technical Report No. 1. International Association on Water Pollution Research and Control. Henze M., Gujer W., Mino T., Matsuo T., Wentzel M. C. and Marais G. v. R. (1995) Activated Sludge Model No. 2. Scienti®c and Technical Report No. 3. International Association on Water Quality. Hvitved-Jacobsen T., Vollertsen J. and Nielsen P. H. (1998a) A process and model concept for microbial wastewater transformations in gravity sewers. Water Sci. Tech. 37(1), 233±241. Hvitved-Jacobsen T., Vollertsen J. and Tanaka N. (1998b) Wastewater quality changes during transport in sewers: an integrated aerobic and anaerobic model concept for carbon and sulfur microbial transformations. Water Sci. Tech. 39(2), 242±249 Hvitved-Jacobsen T. and Vollertsen J. (1999). Modeling of microbial wastewater transformations in gravity sewers: parameter estimation, calibration and validation. Water Res., in preparation. Kappeler J. and Gujer W. (1992) Estimation of kinetic parameters of heterotrophic biomass under aerobic con-
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ditions and characterisation of wastewater for activated sludge modelling. Water Sci. Tech. 25(6), 125±139. Jansen L. C. J. and HarremoeÈs P. (1984) Removal of soluble substrates in ®xed ®lms. Water Sci. Tech. 17(2-8 part 1), 1±14. Nielsen P. H., Raunkjñr K., Norsker N. H., Jensen N. Aa. and Hvitved-Jacobsen T. (1992) Transformation of wastewater in sewer systems: a review. Water Sci. Tech. 25(6), 17±31. Nielsen P. H., Raunkjñr K. and Hvitved-Jacobsen T. (1998) Sul®de production and wastewater quality in pressure mains. Water Sci. Tech. 37(1), 97±104. Schaarup-Jensen K., Hvitved-Jacobsen T., JuÈtte B., Nielsen B. and Pedersen T. (1998) A Danish sewer research and monitoring station. Water Sci. Tech. 37(1), 197±204. Standard Methods (1995) Standard Methods for the Examination of Water and Wastewater, 19th edn. APHA, Washington.
Streeter H. W. and Phelps E. B. (1925) A study of pollution and natural puri®cation of the Ohio river. Public Health Bulletin No. 146, Washington DC. Theriault, E. J. (1927) The oxygen demand of polluted waters. Public Health Bulletin No. 173, Washington DC. Vollertsen J. and Hvitved-Jacobsen T. (1998) Aerobic microbial transformations of resuspended sediments in combined sewers: a conceptual model. Water Sci. Tech. 37(1), 69±76. Vollertsen J. and Hvitved-Jacobsen T. (1999) Stoichiometric and kinetic model parameters for microbial transformations of suspended solids in combined sewer systems. Water Res. in press Vollertsen J., Hvitved-Jacobsen T., McGregor I. and Ashley R. M. (1998). Aerobic microbial transformations of pipe and silt trap sediments from combined sewers. Water Sci. Tech. 39(2), 234±241 Zanoni A. E. (1967) Waste water deoxygenation at dierent temperatures. Water Res. 1(8/9), 543±566.
Wat. Res. Vol. 33, No. 14, pp. 3119±3126, 1999 # 1999 Elsevier Science Ltd. All rights reserved Printed in Great Britain 0043-1354/99/$ - see front matter
www.elsevier.com/locate/watres
EFFECTS OF TEMPERATURE AND DISSOLVED OXYGEN ON HYDROLYSIS OF SEWER SOLIDS M M JES VOLLERTSEN1** , MARIA DO CEÂU ALMEIDA2* and M THORKILD HVITVED-JACOBSEN1*
1 Environmental Engineering Laboratory, Aalborg University, Sohngaardsholmsvej 57, DK-9000 Aalborg, Denmark; 2Department of Civil Engineering, Imperial College, Imperial College Road, London SW7 2BU, UK
(First received June 1998; accepted in revised form January 1999) AbstractÐEects of temperature and dissolved oxygen (DO) on kinetics of microbial transformation processes of suspended sewer sediment particles, suspended wastewater particles and wastewater were investigated. Microbial activity was estimated by the oxygen uptake rate. The eects were studied under conditions where biomass had not acclimated to changed temperature or DO conditions. Hydrolysis processes limited the availability of organic substrates for the aerobic microbial growth. The eects of temperature were interpreted applying an Arrhenius relationship. The average Arrhenius constants found for sewer sediment particles and wastewater particles diered signi®cantly from the average Arrhenius constant found for wastewater. The eects of dissolved oxygen concentration were interpreted applying a saturation type relationship. No dierences between sewer sediment particles, wastewater particles and wastewater were found for the oxygen saturation coecients. # 1999 Elsevier Science Ltd. All rights reserved Key wordsÐtemperature, dissolved oxygen, hydrolysis, suspended sewer sediments, wastewater particles, wastewater, microbial processes, process kinetics
NOMENCLATURE
a KO OUR SO SO, 1 T#
Arrhenius constant (ÿ) Saturation constant for dissolved oxygen concentration (g O2 mÿ3) Oxygen uptake rate (g O2 mÿ3 hÿ1) Dissolved oxygen concentration (g O2 mÿ3) Nonlimiting dissolved oxygen concentration (g O2 mÿ3) Temperature (8C)
INTRODUCTION
Aerobic microbial transformations of organic matter in sewer systems can be interpreted and simulated using conceptual models addressing process kinetics and stoichiometry of heterotrophic biomass (Bjerre et al., 1995, 1998; Hvitved-Jacobsen et al., 1998a, b,; Vollertsen and Hvitved-Jacobsen, 1998, 1999; Vollertsen et al., 1998). In these models it is assumed that the biomass utilizes organic substrate and dissolved oxygen (DO) for growth and Ð depending on the concept Ð also DO for nongrowth-related energy requirements. The DO concentration and the temperature aect the rates of *Author to whom all correspondence should be addressed. [Tel.: +45-963-58-080; Fax: +45-98-142-555].
aerobic microbial processes. Implementing these conceptual models for simulation of natural systems subject to rapid changing temperature and DO conditions, requires information regarding the in¯uence of short term temperature and DO variations upon the simulated microbial transformations. These concepts may be applied for dierent conditions, i.e. for wastewater, where large fractions of readily biodegradable and hydrolyzable substrates exist together with high concentrations of heterotrophic biomass, wastewater particles and organic matter in sewer sediments where the dominating compounds are slowly hydrolyzable substrate and where heterotrophic biomass is present in only low concentrations. When simulating in-sewer microbial transformation, sewer sediment particles, wastewater particles and wastewater are likely to interchange. Also when assessing combined sewer over¯ow (CSO) impacts, it is a mixture of discharged sewer sediment particles, wastewater particles and diluted wastewater that cause the DO impacts. Relative responses to transient DO and temperature variations are, as an example, of importance. The oxygen uptake rate (OUR) is considered to be the key parameter to be determined for estimation of aerobic microbial transformation of organic matter according to the model concept which
3119
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is used. Therefore, eects of DO and temperature on the OUR are important. The eects of temperature and DO have been studied under conditions where only hydrolyzable substrates are utilized as electron donor for microbial transformations. This is in agreement with concepts presented by e.g. Bjerre et al. (1995), Hvitved-Jacobsen et al. (1998a; Hvitved-Jacobsen and Vollertsen, 1999), Vollertsen and HvitvedJacobsen (1998) and Vollertsen et al. (1998). Furthermore, the hydrolysis processes governing the microbial transformations only allowed for heterotrophic biomass growth rates signi®cantly lower than the maximum growth rates; even negative growth might occur. Therefore, microbial responses to temperature and DO was interpreted as eects of temperature and DO on the hydrolysis processes. Responses of sewer sediment particles, wastewater particles and wastewater due to short-term changes in temperature and DO concentration were studied because sewer solids as well as wastewater may be subject to such changes. Short-term changes of temperature and DO concentration occur for example during transport in sewers and in receiving waters following a CSO discharge. METHODS
Sampling Wastewater and wastewater particles were sampled in Frejlev, Denmark, at the Frejlev Research and Monitoring Station (Schaarup-Jensen et al., 1998). Frejlev has a combined sewer system, 2000 inhabitants with no signi®cant industry and a reduced (impermeable) area of approximately 30 ha. As the catchment is situated on a hillside, with the highest part approximately 40 m above the Research and Monitoring Station, the slope of the sewer is generally high and no signi®cant deposition of sediment occurs in the sewers. In the Frejlev Research and Monitoring Station, the sewer is divided into modules
which are accessible from all sides. In two situations wastewater particles and wastewater were sampled in a pumping station located at an intercepting pipe from a small residential town, Vestbjerg, to the wastewater treatment plant Aalborg East, Denmark. During sampling the sewer pipe in the monitoring station was opened so that large quantities of completely mixed wastewater ¯owed out. Wastewater was collected with buckets and transferred into a 120-litre cylindrical tank. When the tank was full, the wastewater was allowed to settle for 30 min and was subsequently decanted. The settled wastewater was transported to the laboratory for measurement immediately after decanting. Sewer sediments were sampled from a 1000-mm concrete combined sewer pipe in an older residential part of the city of Aalborg, Denmark. The sewer serves approximately 20 households upstream of the sampling locations. Sampling was done using a sewer sediment shovel from three adjacent manholes. In one situation sewer sediments were sampled from a silt trap immediately downstream of the above mentioned pumping station in Vestbjerg. Total solids (TS) and volatile solids (VS) of sewer sediment particles, wastewater particles and wastewater were determined according to Standard Methods (1995). Oxygen uptake rate measurement, OUR Oxygen uptake rates of wastewater and suspended particles or sediments were determined in a series of six stainless steel batch reactors with volumes of 2.2 litres each (Fig. 1). Tap water was added to the samples of particles and sediments up to the reactor volume. The reactors were kept at the required temperature by circulating water from one temperature controlled water bath through caps enclosing approximately 30% of the apparatus surface area. When not determining the temperature dependency of process kinetics, temperature was kept at 20 20.58C. Particles were kept in suspension by using magnetic stirrers. DO of wastewater or suspensions was measured using INGOLD oxygen sensors with 12-mm diameter Te¯on membranes. When the DO concentration was below a preset value, an aeration cycle was started by automatically opening the closing piston between the reactor and the expansion chamber. This process was followed by compressed air injection into the wastewater or particle suspension during a preset time interval. When not determining the in¯uence of DO on the OUR kinetics, this set-
Fig. 1. OUR measuring equipment. Six devices were operated in parallel and temperature controlled by the same water circulation, allowing for six parallel batch experiments under the same temperature conditions.
Hydrolysis of sewer solids point was 60% of air saturation. Once the aeration was ®nished, the piston was closed after a preset time delay to ensure that no air bubbles were left in the suspension. Temperature was measured in one of the six batch reactors using a PT100 sensor with a resolution of 0.18C. The temperature in the other ®ve reactors was assumed the same. Time, DO concentration and temperature were logged with a computer. The OUR was calculated from these measurements by linear regressions. Each experiment consisted of duplicate OUR determinations of wastewater, sewer sediment particles and wastewater particles. This resulted in six OUR measurements all performed at the same temperature. Each set of experiments started with a 24-h aeration period, in order to guarantee that only aerobic microbial transformations of hydrolyzable substrates contributed to the OUR during the experiments. Temperature and DO eects on hydrolysis of sewer solids were determined subsequently on the same sample. Eects of temperature on process kinetics Temperature eects of process kinetics was determined by running the OUR experiments under a sequence of periods with dierent reactor temperatures. After an initial period of one day at 208C, the temperature was increased to 23±248C, secondly decreased to 15±168C, then to 9± 108C and ®nally it was increased to the initial temperature of about 208C. At each temperature level the temperature was kept constant for typically 2 h, Fig. 2. Depending on
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the OUR value, 1±2 h of constant temperature was needed. Higher temperatures resulted in higher rates and, hence, the time required for stabilization of the OUR at a new temperature level was shortest for higher temperatures. From the OUR values at adjacent temperature levels, temperature eects on kinetics of aerobic microbial hydrolysis of organic substrates were determined. The interpretation of measured eects of temperature on process kinetics is based on equation (1), presented by Streeter and Phelps (1925) to describe the dependency of wastewater ®rst order deoxygenation kinetics upon temperature. Theriault (1927) showed that there was similarity between the Arrhenius equation for chemical reactions and equation (1). Therefore, equation (1) often simply is called the Arrhenius relationship and a referred to as the Arrhenius constant (Zanoni, 1967). The well-known Arrhenius relationship has traditionally been used in many microbial contexts; e.g. activated sludge (Henze et al., 1987; Barker and Dold, 1997), bio®lms (Characklis and Marshall, 1990), wastewater transformations in gravity sewers (Nielsen et al., 1992; Hvitved-Jacobsen et al., 1998b) and hydrogen sul®de formations in pressure sewers (Boon and Lister, 1974; Nielsen et al., 1998). For this reason, the Arrhenius relationship will be assumed in this study for interpretation of measured temperature eects on process kinetics. OUR T1 a T1 ÿT2 : OUR T2
1
Eects of DO on process kinetics Following completion of the temperature variation period, and after the temperature had been stabilized at 208C, the determination of DO limitation eect on kinetics of aerobic microbial transformations of hydrolyzable substrates was determined. This was achieved by changing the set-point de®ning the end of a measuring period from 60% of air saturation down to 0.1% of air saturation. The resulting DO interval, on which eects of DO concentration on process kinetics was determined, was from 7±9 g O2 mÿ3 down to about 0.01 g O2 mÿ3. These measurements were carried out for 15±20 h. DO limitation eects on kinetics of microbial transformations, under conditions where electron donor availability was limited by hydrolysis, were determined from calculated OUR vs DO values. Eects of DO on microbial transformations of organic substrates, for purposes of modeling microbial transformations of organic substrate in gravity sewers and for activated sludge process modeling, have been simulated with saturation type equations (Henze et al., 1987, 1995; Hvitved-Jacobsen et al., 1998a, b; Bjerre et al., 1998). Therefore, a saturation type equation (Monod) will be assumed in this study equation (2). OUR SO OUR SO, 1
SO : KO S O
2
RESULTS AND DISCUSSION
Fig. 2. Duplicate determination of eects of temperature upon oxygen uptake rate (OUR) for wastewater sampled September 19, 1997. Resulting average Arrhenius constants a were 1.11, 1.10, 1.13 and 1.12 for the temperature changes 20.5±23.6, 23.6±16.3, 16.3±10.3 and 10.3±20.58C, respectively.
For each of the three types of sewage components investigated Ð sewer sediment particles, wastewater particles and wastewater Ð 11 samples were analyzed for the eect of temperature and DO on the OUR kinetics. Kinetics were determined in duplicate on each sample. However, two samples were analyzed only once because of equipment malfunction.
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Table 1. Arrhenius constants a for sewer sediment particles, wastewater particles and wastewater Samples Determinations Average Std. of a a's dev. Sewer sediment particles Wastewater particles Wastewater
11 11 11
43 48 46
1.091 1.089 1.099
0.015 0.016 0.018
Eect of temperature variation on process kinetics Temperature eects on process kinetics were determined from the change in OUR response for each change in temperature level. As an example, the measurements presented in Fig. 2 resulted in duplicate determination of the Arrhenius constant for the four temperature changes: 20.5±23.6, 23.6± 16.3, 16.3±10.3 and 10.3±20.58C. Isolating a from equation (1) and calculating a for all OUR responses to every change in temperature level, duplicate determinations of the Arrhenius constants were obtained (Table 1). The variability of the Arrhenius constants determined was large, with extreme values of 1.053 and 1.142 (Fig. 3). No systematic variability of this constant with temperature was observed. The standard deviation of the duplicate determination was, how-
ever, considerably lower than the standard deviation of the whole sample. The Arrhenius constants obtained where normally distributed for the three types of organic matter investigated (Fig. 3). No signi®cant dierence in the average values of a was found between wastewater particles and sewer sediment particles. However, the average Arrhenius constants for wastewater diered signi®cantly from the average constants found for both wastewater particles and sewer sediment particles. Describing oxygen consumption of wastewater and river water, Streeter and Phelps (1925) found an average Arrhenius constant of 1.047 in the temperature interval 10±37.58C. Zanoni (1967) described oxygen consumption of wastewater at dierent temperatures, measuring BOD at dierent incubation periods. He found that a ®rst order description of the oxygen consumption had shortcomings, primarily because of very high oxygen demands in the beginning of the incubation periods for incubation temperatures of 208C and above. Excluding these high, early oxygen demands, he found an average a of 1.126 for the temperature range 2±158C, of 1.047 for 15±328C and of 0.985 for 32±408C. Looking at the high, initial oxygen demand observed by Zanoni (1967) in the light of
Fig. 3. Normal probability plots of Arrhenius constants (a). The error bars indicate the standard deviation of the duplicate determinations for sewer sediments, wastewater particles and wastewater, respectively.
Hydrolysis of sewer solids
recent studies of wastewater microbial transformations (Bjerre et al., 1995; Hvitved-Jacobsen et al., 1998a), these high oxygen demands probably were due to an initial amount of readily biodegradable organic substrate. When Zanoni (1967) excluded these measurements, the limiting process for the temperature kinetics could be interpreted as hydrolysis, i.e. comparable with the experimental conditions in this study. The authors of the activated sludge model No. 2 (Henze et al., 1995) suggest an Arrhenius constant of 1.07 for all heterotrophic processes. Ekama et al. (1984) have found that the oxygen demand of activated sludge decreases by 5% with a 68C change in temperature, i.e. an Arrhenius constant of 1.009. The Arrhenius constants that have been found in this study are within the range of the ®ndings of Zanoni (1967) for wastewater. However, the Zanoni (1967) ®ndings of a decreasing Arrhenius constant for increasing temperature could not be con®rmed. The value for activated sludge modeling, recommended by the authors of the activated sludge model No. 2 (Henze et al., 1995), is somewhat lower than the results obtained in this study. However, that recommendation is made on the basis that biomass is allowed to acclimatize to the new temperature. This probably results in a change in biomass composition and properties from one temperature level to another, resulting in a dierent a-value. In the present experiments, the biomass was acclimated for 24 h to 208C and its composition during the following hours of experiment expected unchanged. The type of temperature change addressed in this study is expected to occur in sewers. Temperature reductions up to 3.58C were observed during transport in a 5-km long intercepting gravity sewer (Hvitved-Jacobsen and Vollertsen, 1999) and according to the present results this decrease in temperature would correspond to a 28% reduction in wastewater hydrolysis rates during transport. Eects of dissolved oxygen concentration on process kinetics Eects of dissolved oxygen concentration on process kinetics were determined from DO measurements vs calculated OUR for the DO concentration range from 7±9 g O2 mÿ3 down to 0.01 g O2 mÿ3. The number of aeration cycles used by the OURmeasurement equipment depended on the OUR, as the total duration of these experiments were predetermined. Up to four cycles were selected from each batch for simulation with equation (2). OUR at nonlimiting DO levels changed somewhat with time. To minimize errors in KO values due to these variations, cycles where OUR varied considerably at nonlimiting DO levels were omitted. Some typical curves of DO vs OUR are presented in Fig. 4 together with simulations based on equation (2). At low DO values, simulated oxygen
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uptake rates tend to be lower than the measured rates. Furthermore, in some simulations the observed increase in OUR for high DO values was not simulated well, Fig. 4(b). This observation raises the question of whether a saturation type expression is appropriate for simulation of OUR response to decreasing DO concentrations. The argument was, that some of the biomass utilizing DO as electron acceptor might be located within aggregates or in larger particles where DO diusion might be limiting the DO uptake process. If such a process was comparable to bio®lm processes, this should result in some of the DO uptake processes to be better described by a 1/2 order rate expression (Jansen and HarremoeÈs, 1984). Simulation of OUR using a 1/2 order rate expression in DO was tried. However, the results were not better than with equation (2). The hypothesis that only a part of the oxygen uptake followed a 1/2 order rate expression in DO was tested by means of simulation with the sum of equation (2) and a 1/2 order rate expression. This resulted in better agreement between measurements and simulations. However, for a DO level of 1 g O2 mÿ3, equation (2) was on average accounting for 96, 98 and 91% of the OUR measured for sewer sediment particles, wastewater particles and wastewater, respectively. Therefore, the 1/2 order rate expression in DO could only explain a minor part of the measured OUR. Because two dierent expressions would increase the complexity of the simulation, only the saturation type expression (2) was used in the present study. Saturation constants that were determined for each sample showed a rather high standard deviation compared to the standard deviation between dierent samples (Table 2) Ð in average the standard deviation of each sample was 0.04. As the samples have been treated identically using identical equipment, the reason for the poor quality of duplicate determination cannot be explained. The saturation constants for duplicate determinations were found to be normally distributed (Fig. 5). Comparison between average values of KO for the three dierent types of organic matter showed Ð at a 5% signi®cance level and using a t-test Ð no signi®cant dierences between wastewater particles, wastewater or sewer sediment particles. The average KO value for all samples was 0.080 g O2 mÿ3 with a standard deviation of 0.036 g O2 mÿ3. Henze et al. (1995) state that typical values of KO for modeling of activated sludge heterotrophic processes is 0.20 g O2 mÿ3, i.e. somewhat higher than what was found here. Using the same methodology for activated sludge Kappeler and Gujer (1992) found that a saturation type expression is a rather good approximation of the DO dependency for OUR. They found an average KO of 0.25 g O2 mÿ3. For wastewater Bjerre et al. (1998) and GudjoÂnson
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Fig. 5. Normal probability plots of saturation constants KO. The error bars indicate the standard deviation of the duplicate determinations for sewer sediment particles, wastewater particles and wastewater.
et al. (in preparation) found an average KO of 0.05 g O2 mÿ3. Dierences in particle size between activated sludge and the studied particles might explain the somewhat lower KO values found in this study. CONCLUSION
Eects of temperature and dissolved oxygen concentration (DO) on oxygen uptake rate (OUR) of suspended sewer sediment particles, suspended wastewater particles and wastewater have been investigated and compared. Information regarding such eects is required when using models for simulation of microbial transformations of organic substrates, for simulation of transformation processes in sewer systems or for impacts on receiving waters. Eects of temperature and DO concentration on Table 2. Saturation constants KO determined using equation (2) for sewer sediments, wastewater particles and wastewater
Sewer sediment particles Wastewater particles Wastewater
Samples
Average KO
Std. dev.
11 11 11
0.065 0.087 0.088
0.011 0.040 0.040
OUR have been studied under conditions where availability of the electron donor is limited by hydrolysis and thus also limiting the OUR. Temperature eects on process kinetics were described using an Arrhenius expression, resulting in an average Arrhenius constant of 1.091 for suspended sewer sediments, of 1.089 for suspended wastewater particles and of 1.099 for wastewater. The average Arrhenius constant found for wastewater diered signi®cantly from the two particle suspensions. Investigating eects of DO on OUR, measurements of DO vs calculated OUR were simulated using a saturation type expression. For the calculated saturation coecients no signi®cant dierence between suspended sewer sediment particles, suspended wastewater particles and wastewater was found. The resulting average saturation constant for suspended particles and wastewater was 0.081 g O2 mÿ3. The eects of temperature and DO on microbial transformations were investigated in the short-term, where the biomass did not have time to acclimate. Short-term temperature or DO changes are normal events in sewer systems as well as in the interactions between sewer system and receiving waters. If bio-
Hydrolysis of sewer solids
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Fig. 4. Typical curves of OUR vs DO concentration and simulations based on equation (2). (a) Sewer sediment particles, 15/9/97, (b) wastewater, 8/10/97, (c) wastewater particles, 3/10/97.
mass is allowed to acclimate, rates are likely to be dierent. This may explain the dierences from saturation coecients found in other studies. Extreme changes in temperature, followed by biomass acclimation, might favor the development of other microorganisms, resulting in a change in biomass composition. Other kinetic parameters would be needed under such circumstances. The ®ndings of this study should, therefore, only be used when modeling short-term changes in temperature or DO. AcknowledgementsÐFinancial support for this research project was provided by the Danish Technical Research Council, the framework program on `Solids in Sewage Systems'. REFERENCES
Barker P. S. and Dold P. L. (1997) General model for biological nutrient removal activated-sludge systems: model presentation. Water Env. Res. 69(5), 969±984. Boon A. G. and Lister A. R. (1974) Formation of Sulphide in Rising Main Sewers and Its Prevention. Water Pollution Research Laboratory, Stevenage, U.K. Bjerre H. L., Hvitved-Jacobsen T., TeichgraÈber B. and te Heesen D. (1995) Experimental procedures characterizing transformations of wastewater organic matter in the Emscher river, Germany. Water Sci. Tech. 31(7), 201± 212. Bjerre H. L., Hvitved-Jacobsen T., TeichgraÈber B. and Schlegel S. (1998). Modelling of aerobic wastewater transformations under sewer conditions in the Emscher river, Germany. Water Environ. Res., in press.
Characklis W. G. and Marshall K. C. (1990) Bio®lms. In Wiley Series in Ecological and Applied Microbiology, eds. W. G. Characklis and K. C. Marshall U.S.A. Ekama G. A., Marais G. v. R., Siebritz I. P., Pitman A. R., Keay G. F. P., Buchan L., Gerber A. and Smollen M. (1984) Theory, Design and Operation of Nutrient Removal Activated Sludge Processes. Water Research Commission, Pretoria, South Africa. GudjoÂnson G., Hvitved-Jacobsen T., Vollertsen J. and Nielsen C. R. (in preparation). Microbial transformations in gravity sewers: variability and modeling of dissolved oxygen. Henze M., Grady C. P. L., Jr., Gujer W., Marais G. v. R. and Matsuo T. (1987) Activated Sludge Model No. 1. Scienti®c and Technical Report No. 1. International Association on Water Pollution Research and Control. Henze M., Gujer W., Mino T., Matsuo T., Wentzel M. C. and Marais G. v. R. (1995) Activated Sludge Model No. 2. Scienti®c and Technical Report No. 3. International Association on Water Quality. Hvitved-Jacobsen T., Vollertsen J. and Nielsen P. H. (1998a) A process and model concept for microbial wastewater transformations in gravity sewers. Water Sci. Tech. 37(1), 233±241. Hvitved-Jacobsen T., Vollertsen J. and Tanaka N. (1998b) Wastewater quality changes during transport in sewers: an integrated aerobic and anaerobic model concept for carbon and sulfur microbial transformations. Water Sci. Tech. 39(2), 242±249 Hvitved-Jacobsen T. and Vollertsen J. (1999). Modeling of microbial wastewater transformations in gravity sewers: parameter estimation, calibration and validation. Water Res., in preparation. Kappeler J. and Gujer W. (1992) Estimation of kinetic parameters of heterotrophic biomass under aerobic con-
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ditions and characterisation of wastewater for activated sludge modelling. Water Sci. Tech. 25(6), 125±139. Jansen L. C. J. and HarremoeÈs P. (1984) Removal of soluble substrates in ®xed ®lms. Water Sci. Tech. 17(2-8 part 1), 1±14. Nielsen P. H., Raunkjñr K., Norsker N. H., Jensen N. Aa. and Hvitved-Jacobsen T. (1992) Transformation of wastewater in sewer systems: a review. Water Sci. Tech. 25(6), 17±31. Nielsen P. H., Raunkjñr K. and Hvitved-Jacobsen T. (1998) Sul®de production and wastewater quality in pressure mains. Water Sci. Tech. 37(1), 97±104. Schaarup-Jensen K., Hvitved-Jacobsen T., JuÈtte B., Nielsen B. and Pedersen T. (1998) A Danish sewer research and monitoring station. Water Sci. Tech. 37(1), 197±204. Standard Methods (1995) Standard Methods for the Examination of Water and Wastewater, 19th edn. APHA, Washington.
Streeter H. W. and Phelps E. B. (1925) A study of pollution and natural puri®cation of the Ohio river. Public Health Bulletin No. 146, Washington DC. Theriault, E. J. (1927) The oxygen demand of polluted waters. Public Health Bulletin No. 173, Washington DC. Vollertsen J. and Hvitved-Jacobsen T. (1998) Aerobic microbial transformations of resuspended sediments in combined sewers: a conceptual model. Water Sci. Tech. 37(1), 69±76. Vollertsen J. and Hvitved-Jacobsen T. (1999) Stoichiometric and kinetic model parameters for microbial transformations of suspended solids in combined sewer systems. Water Res. in press Vollertsen J., Hvitved-Jacobsen T., McGregor I. and Ashley R. M. (1998). Aerobic microbial transformations of pipe and silt trap sediments from combined sewers. Water Sci. Tech. 39(2), 234±241 Zanoni A. E. (1967) Waste water deoxygenation at dierent temperatures. Water Res. 1(8/9), 543±566.