Process design manual for nitrogen control

EMBED for wordpress. Want more? Advanced embedding details, examples, and help! This manual presents theoretical and process design criteria for the implementation of nitrogen control technology in municipal wastewater treatment facilities. Design concepts are emphasized through examination of data from full-scale and pilot installations.

Design data are included on biological nitrification and denitrification, breakpoint chlorination, ion exchange, and air stripping. One chapter presents the concepts involved in assembling various unit processes into rational treatment trains and presents actual case examples of specific treatment systems that incorporate nitrogen control processes.

The first extensive study was undertaken in by Battelle Northwest in a federally sponsored demonstration project. Regeneration of the clinoptilolite is undertaken when all the exchange sites are utilized and breakthrough occurs.

Filtration prior to ion exchange is usually required to prevent fouling of the zeolite. Ammonium removals of percent can be expected.

Nitrite, nitrate, and organic nitrogen are not affected by this process. Reducing the partial pressure causes ammonia to leave the water phase and enter the air. Ammonia removal from wastewater can be effected by bringing small drops of water in contact with a large amount of ammonia-free air. This physical process is termed desorption, but the common name is "ammonia stripping. This is accomplished by raising the pH of the wastewater to 10 or 11, usually by the addition of lime. Because lime addition is often used for phosphate removal, it can serve a dual role.

Again, nitrite, nitrate, and organic nitrogen are not affected. The principal problems associated with ammonia stripping are its inefficiency in cold weather, required shutdown during freezing conditions, and formation of calcium carbonate scale in the air stripping tower. The effect of cold weather has been well documented at the South Lake Tahoe Public Utility District where ammonia stripping is used for a 3. The stripping tower is designed to remove 90 percent of the incoming ammonium during warm weather.

During freezing conditions, the tower is shut down. One mechanism of scale formation is attributed to the carbon dioxide in the air reacting with the alkaline wastewater and precipitating as calcium carbonate.

Some factors which may affect the nature of the scale are: orientation of air flow, recirculation of sludge, pH of the wastewater, and chemical makeup of the waste water.

Most are in the experimental stage of development or occur coincidentally with another process. Use of anionic exchange resins for removal of nitrate was developed principally for treatment of irrigation return waters. Oxidation ponds can remove nitrogen through microbial denitrification in the anaerobic bottom layer or by ammonia emission to the atmosphere.

The latter effect is essentially ammonia stripping but is relatively inefficent due to a low surface-volume ratio and low pH. In a study of raw wastewater lagoons in California, removals of percent were reported for well-operated lagoons. If the algal cells are removed from the pond effluent stream, nitrogen removal is thereby effected.

If an organic carbon source such as ethanol or glucose is added to the wastewater, the solids production will be increased and a greater nitrogen removal will be effected. Disadvantages are that large quantities of sludge are produced and that difficulties occur in regulating the addition of the carbon source, with high effluent BOD5 values or high nitrogen levels resulting.

Shown is the effect that the process has on each of the three major forms: organic nitrogen, ammonium, and nitrate. In the last column is shown normal removal percentages which can be expected from that process.

Overall removal for a particular treatment plant will depend on the types of unit processes and their relation to each other. For example, while many processes developed for nitrogen removal are ineffective in removing organic nitrogen, incorporation of chemical coagulation or multimedia filtration into the overall flowsheet can result in a low concentration of organic nitrogen in the plant effluent.

Thus, the interrelationship between processes must be carefully analyzed in designing for nitrogen removal. Further discussion of process interrelationships is presented in Chapter 9. Soluble organic nitrogen, in the form of urea and amino acids, is substantially reduced by secondary treatment. GMay be used to remove particulate organic carbon in plants where ammonia or nitrate are removed by other processes.

Sawyer, C. McCarty, Chemistry for Sanitary Engineers. Christensen, M. Harremoes, Biological Denitrification in Wastewater Treatment. Delwiche, C. Scientific American,. Martin, D. Report No. Sepp, E. McCarty, P. JAWWA, 59, pp Sylvester, R. Alfjae and Metropolitan Wastes, Robert A. Taft Sanitary Engineer- ing Center, Tech. W, Reeves, T. Nitrogenous Compounds in the Environment. Kaufman, W. Weibel, S. JWPCF, 43, p WP, January, Burn, R. JWPCF, 40, pp EPA Report No. JAWWA, 58, pp Johnson, R. SA1, pp November, Ehreth, D.

Barth, Control of Nitrogen in Wastewater Effluents. Nitrogen Removal from Wastewaters. Effects of Pollution Discharges on the Thames Estuary. Parkin, G. Caldwell, D. Uhte, Upgrading Lagoons. An understanding of this subject is useful for developing an appreciation of the factors affecting the performance, design, and operation of nitrification and denitrification processes.

Subsequent chapters deal with design aspects of nitrification Chapter 4 and denitrification Chapter 5. Since these latter chapters are laid out to be used without reference to this chapter, review of the theoretical material in this chapter is not mandatory. Biological processes for control of nitrogenous residuals in effluents can be classified in two broad areas. First, a process designed to produce an effluent where influent nitrogen ammonia and organic nitrogen is substantially converted to nitrate nitrogen can be considered.

This process, nitrification, is carried out by bacterial populations that sequentially oxidize ammonia to nitrate with intermediate formation of nitrite. Nitrification will satisfy effluent or receiving water standards where reduction of residual nitrogenous oxygen demand due to ammonia is mandated or where ammonia reduction for other reasons is required for the treatment system.

The second type of process, denitrification, reduces nitrate to nitrogen gas and can be used following nitrification when the total nitrogenous content of the effluent must be reduced. Both of these groups are classed as autotrophic organisms.

These organisms are distinguished from heterotrophic bacteria in that they derive energy for growth from the oxidation of inorganic nitrogen compounds, rather than from the oxidation of organic matter. Another feature of these organisms is that inorganic carbon carbon dioxide is used for synthesis rather than organic carbon. Each group is limited to the oxidation of specific species of nitrogen compounds.

Nitrosomonas can oxidize ammonia to nitrite, but cannot complete the oxidation to nitrate. On the other hand, Nitrobacter is limited to the oxidation of nitrite to nitrate. Since complete nitrification is a sequential reaction, treatment processes must be designed to provide an environment suitable to the growth of both groups of nitrifying bacteria.

Various reaction intermediates and enzymes are involvedJ More important than an understanding of these pathways is knowledge of the response of nitrification organisms to environmental conditions.

If it assumed that the cell synthesis per unit of energy produced is equal, there should be greater mass of Nitrosomonas formed than Nitrobacter per mole of nitrogen oxidized. As will be seen, this is in fact the case.

These reactions usually take place at pH levels less than 8. Under this circumstance, the production of acid results in immediate reaction with bicarbonate ion HCOs with the production of carbonic acid H2CO3. The consumption of carbon dioxide by the organisms results in some depletion of the dissolved form of carbon dioxide, carbonic acid H2CO3.

Table presents the modified forms of Equations to to reflect the changes in the carbonic acid system. As will be later described in Sections 3.

The equations for energy yielding reactions Equations and can be combined with the equations for organism synthesis Equations and to form overall synthesis- oxidation relations by knowledge of the yield coefficients for the nitrifying organisms. Experimental yield values for Nitrosomonas range from 0. This relatively low yield has some far reaching implications, as will be seen in Section 3.

Oxygen consumption ratios in the equations are 3. Alkalinity and pH Relationships Equation A Table shows that alkalinity is destroyed by the oxidation of ammonia and carbon dioxide H2CO3 in the aqueous phase is produced. When synthesis is neglected, it can be calculated that 7.

The effect of synthesis is relatively small; in Equation , the ratio is 7. A ratio of 7. The effect is mediated by stripping of carbon dioxide from the liquid by the process of aeration and the pH is elevated upwards. If the carbon dioxide is not stripped from the liquid, such as in enclosed high purity oxygen systems, the pH can be depressed as low as 6. It has been calculated that to maintain the pH greater than 6. In many wastewaters there is insufficient alkalinity initially present to leave a sufficient residual for buffering the wastewater during the nitrification process.

Procedures for calculating the operating pH in aeration systems are presented in Section 4. Synthesis has an effect on oxygen requirements; the oxygen requirement is calculated to be, from Equation , 4. An oxygen requirement sufficiently accurate to be used in engineering calculations for aeration requirements is 4. Caution: in virtually all practical nitrification systems, oxygen demanding materials other than ammonia are present in the wastewater, raising the total oxygen requirements of nitrification systems even higher see Section 4.

In the succeeding portions of this section, the impact of a variety of environmental factors on the rates of growth and nitrification are considered. A combined kinetic expression is then formulated which accounts for the effects of ammonia concentration, temperature, pH and dissolved oxygen concentration. At several points, reference is made to data developed from various types of nitrification processes.

Comprehensive descriptions of the various nitrification processes are presented in Chapter 4 and will not be reproduced herein. One distinction that needs to be clearly understood in discussions in this chapter is the difference between combined carbon oxidation-nitrification processes and separate stage nitrification processes. The combined carbon oxidation-nitrification processes oxidize a high proportion of influent organics BOD relative to the ammonia nitrogen content.

This causes relatively low populations of nitrifiers to be present in the biomass. Separate stage nitrification systems, on the other hand, have a relatively low BODs load relative to the influent ammonia load. As a result, higher proportions of nitrifiers are obtained. Separate stage nitrification can be provided in municipal treatment applications when a high level of organic carbon removal is provided prior to the nitrification stage.

This level of treatment is generally greater than provided by primary treatment. Other differences between these classes of processes can be drawn, but these are left for detailed discussion in Section 4. Nitrosomonas' growth is limited by the concentration of ammonia nitrogen, while Nitrobacter 's growth is limited by the concentration of nitrite. For this reason, the rate of nitrifier growth can be modeled with Equation using the rate limiting step, ammonia conversion to nitrite.

For cases where nitrite accumulation does occur, other approaches are available. The solids retention time can be calculated from system operating data by dividing the inventory of microbial mass in the treatment system by the quantity of biological mass wasted daily. Equations applicable for this calculation are presented in Section 4. Further, the maximum growth rate for Nitrosomonas in activated sludge was found to be considerably less than for Nitrosomonas in pure culture.

Kinetic constants found by other investigators are summarized in Tables and This suggests that some additional parameter such as dissolved oxygen DO may have been limiting Downing's activated sludge measurements. For illustrative use in this manual the pure culture values of Downing, et al. Huang and Hopson's summary, with some modifications, is shown in Figure for attached growth systems. Precise measurement of biomass is normally not possible in attached growth systems so other parameters are used such as reaction rate per unit surface or volume.

Therefore, attached growth systems can compensate for colder temperature conditions by the effective slime growth growing thicker. Thus, if rates could be expressed on a unit biomass basis for both system types, reaction rate variation with temperature might be more similar. It could be argued that compensation for low temperature in suspended growth systems could be provided by an increase in mixed liquor level, much as an increase in slime growth occurs in attached growth systems. However, suspended growth systems are limited by reactor-sedimentation tank interactions which at cold temperatures might prohibit this due to reduction of thickening rates of the sludge cf.

Section 4. Other differences in reaction rates shown in Figure may arise from the fact that some determinations were on separate stage nitrification systems while others were made on combined carbon oxidation-nitrification processes. The Monod relationship has been used to model the effect of dissolved oxygen, considering oxygen to be a. U2 British investigators found that the KQ2 value was about 1. Sludge samples were withdrawn, dosed with ammonia, and aerated at various DO levels.

Nitrification rates determined from the data collected are shown in Figure The KQO determined from this data is 2. Temperature was not specified, but indicated to be above 20 C. Several investigations have provided indirect evidence of the importance of the effect of DO on nitrification rate. British investigators found that the nitrifi- cation rates at 2. However, this type of evidence does not indicate that nitrification rate was unaffected, merely that nitrification could be completed in the presence of a low DO level.

Low nitrification rates, depressed by low DO levels, can still be sufficient to cause complete nitrification if the aeration tank detention time is large enough.

Further refinement of KQ2 values can be expected. This value falls in the middle of the range of KQ2 observations 0. This order of reduction in rate could account for most of the difference in growth rate observed by Downing, et al.

Figure presents typical pH relationships from a number of investigations. The results of other investigations have been summarized in the litera- ture. The findings for an attached growth reactor Curve E, Figure are very similar to the findings for an activated sludge Curve C. In neither case were the cultures acclimated to each pH value prior to determining nitrification rates. When a three-week acclimation period was provided for the attached growth reactor, it was found that the rate at pH 6.

For instance, in an activated sludge with insufficient wastewater alkalinity, pH values of 5 to 5. This high acid concentration resulted in a cessation of nitrification; at the same time sludge bulking occurred. The point at which the rate of nitrification decreased was pH 6. In a study of the effect of abrupt changes in pH, it was found that an abrupt change in reactor pH from 7.

However, when the pH was abruptly changed from 7. A return to pH 7. For illustrative use in this manual, the equation of Downing, et alA3, showing the effect of pH on nitrification is adopted. Downing, et al. Using specific values for temperature, pH, ammonia and oxygen, adopted in this manual in Sections 3. The third term in brackets is the Monod expression for the effect of ammonia nitrogen concentration. Similarly, the fourth term in brackets accounts for the effect of DO on nitrification rate.

Equation has been adopted for illustrative use in this manual. As more reliable data becomes available, Equation can be modified to suit particular circumstances.

This rate is expressed per unit of nitrifiers, assuming that there are no other types of bacteria in the population. Nitrification rates of a comparable magnitude have been found experimentally by a number of investigators for laboratory enrichment cultures of nitrifiers.

In all practical applications in wastewater treatment, nitrifier growth takes place in waste treatment processes where other types of biological growth occurs. In no case are there opportunities for pure cultures to develop. This fact has significant implications in process design for nitrification.

In both combined carbon oxidation-nitrification systems and in separate stage nitrification systems, there is sufficient organic matter in the wastewater to enable the growth of heterotrophic bacteria. In this situation, the yield of heterotrophic bacteria growth is greater than the yield of the autotrophic nitrifying bacteria. Because of this dominance of the culture, there is the danger that the growth rate of the heterotrophic organisms will be established at a value exceeding the maximum possible growth rate of the nitrifying organisms.

When this occurs, the slower growing nitrifiers will gradually diminish in proportion to the total population and be washed out of the system. Reduced DO or pH can act to depress the peak nitrifier growth rate and cause a washout condition. As before, the last expression in brackets is taken to be unity above a pH of 7. One way the substrate removal rate can be reduced is to place an organic carbon removal step ahead of the nitrificaiian stage, creating a "separate stage" nitrification process.

The result of this procedure is to reduce the food available to the heterotrophic bacteria and to lessen their dominance in controlling the solids retention time. Separate nitrification stages can have very long solids retention times flf.

Another procedure for reducing the substrate removal rate, without separating the carbon oxidation and nitrification processes, is to increase the biological solids in the system. This can be done by increasing the concentration of biological solids under aeration the MLVSS in the activated sludge system or by increasing the volume of the oxidation tank while maintaining the concentration of biological solids at the same concentration.

The level set for biological solids retention time, 6 c , establishes the biological solids retention time or growth rate of the nitrifiers, since selective wasting of the heterotrophic population is not practical.

Therefore, the design solids retention time can be related to the effluent ammonia level through the Monod relationship and the inverse relationship between nitrifier growth rate and solids retention time Equations and Interestingly, Equation can be manipulated to show that the specification of a SF of 2 will establish an ammonia level equal to KN in the effluent of a complete mix activated sludge plant, if there is a high DO level DO not limiting.

This is because the process will be operating at one-half its maximum growth rate. Furthermore, specific examples of the use of the kinetic expressions developed in this section are presented in Sections 4. Rather than use the sludge growth rate or solids retention time approach described in Section 3.

The other biological solids in the system result from the growth of the hetero trophic population. This latter rate is normally determined experimentally in activated sludge systems, as will be described in Section 4.

Specific analytical techniques for determination of the nitrifier fraction have not as yet been developed. This procedure neglects the ammonia assimilated by heterotrophic growth and therefore is approximate. A further approximation is that the net yield of the heterotrophs has been assumed constant, whereas it is known that it varies with solids retention time. Several examples can be drawn for separate stage nitrification systems as compared to combined carbon oxidation-nitrification systems.

A reasonable estimate for YN is 0. Section 3. Thus even in separate stage systems, the fraction of nitrifiers is relatively low. For the assumed yield values, the fraction is less than 20 percent and greater than 8 percent. It must be emphasized that the values of nitrifier fraction given in Table are estimates only, and not supported by actual measurements of nitrifier fractions.

As the nitrification rates approach their plateau values, nitrification approaches a zero order rate, uninfluenced by ammonia level. It has been shown that for suspended growth processes, the rate of removal approximates a zero order reaction. Figure , demonstrating this effect, was developed similarly to Figure , excepting that the effluent ammonia nitrogen concentration was assumed to be 2. A principal means of increasing the nitrification rate is to increase the fraction of nitrifiers.

It must be emphasized that the nitrification rates developed in this section are only estimated relationships based on theoretical considerations. Actual measured values are presented in Section 4. The effect of BOD5 in the synthetic waste was to displace nitrifiers with heterotrophic bacteria in the bacterial film, thereby reducing the nitrifier fraction.

The growth and oxidation rate relationships presented in Sections 3. While these relationships are operative in attached growth systems, their application is complicated by the fact that oxygen mass transfer limitations through the bacterial slimes may limit reaction rates in some situations.

As a consequence, the design relationships presented for attached growth nitrification in Chapter 4 are empirically based and therefore, less theoretically precise than those developed for suspended growth sytems. Though the design relationships presented are empirical, where possible the loading relationships are presented on a basis that is at least consistent with the biofilm model.

For instance, ammonia nitrogen oxidation rates are expressed on a surface area basis when describing separate stage nitrification in trickling filters and the rotating biological disc process Sections 4. Some of the conclusions that can be drawn from the biofilm model are of interest in considering surface ammonia removal rates in attached growth systems.

The biofilm model shows that the ammonia oxidation rate in attached growth systems should not be decreased as drastically under adverse environmental conditions as in suspended growth systems. The biofilm model also shows that the dissolved oxygen concentration must be 2.

To date, these effects have not been quantitatively incorporated into the kinetic description of nitrifier growth, although such approaches have been used to describe toxicity in other biological systems. A listing of substances toxic to unacclimated nitrifying organisms is presented in Table , which is drawn primarily from the review by Painter.

Such conditions may occur in the sludge collection zone of the secondary clarifier where continuing organism activity may cause low pH values. Alternatively, low pH values may occur when pH control systems fail. Levels of 2 ppb in the influent to the filter were concentrated to 5 ppm in the biomass on the media. This inhibitory effect was found to severely reduce allowable loading rates and result in only partial nitrification.

It has been shown that nitrifiers can adapt to toxic substances when they are consistently present at concentrations higher than cause toxic effects in slug discharges. Under normal municipal conditions of pH and concentrations complete nitrification will occur. When concentrated industrial wastes are present, slug discharges should be avoided; rather, storage facilities should be provided so that wastes can be metered into the collection system at a rate sufficient to ensure dilution to safe loads.

In sum, the possibility of toxic inhibition must be recognized in the design of nitrification systems. Either implementation of source control programs or inclusion of upstream toxicity removal processes may be required, particularly in those cases where significant industrial dischargers are tributary to the collection system. The gaseous product is primarily nitrogen gas but also may be nitrous oxide or nitric oxide. Gaseous nitrogen is relatively unavailable for biological growth, thus denitrification converts nitrogen which may be in an objectionable form to one which has no significant effect on environmental quality.

As opposed to nitrification, a relatively broad range of bacteria can accomplish denitrification, including Psuedomonas, Micrococcus, Archromobacter and Bacillus. These groups accomplish nitrate reduction by what is known as a process of nitrate dissimilation whereby nitrate or nitrite replaces oxygen in the respiratory processes of the organism under anoxic conditions.

Because of the ability of these organisms to use either nitrate or oxygen as the terminal electron acceptors while oxidizing organic matter, these organisms are termed facultative heterotrophic bacteria. Confusion has arisen in the literature in terminology; the process has been termed anaerobic denitrification.

However, the principal biochemical pathways are not anaerobic, but merely minor modifications of aerobic biochemical pathways. The term anoxic denitrification is preferred, since it describes the environmental condition of the absence of oxygen, without implying the nature of the biochemical pathways. Denitrification is a two-step process in which the first step is a conversion of nitrate to nitrite. The second step carries nitrite through two intermediates to nitrogen gas.

This two-step process is normally termed "dissimilation. Ammonia is then used for the bacterial cell's nitrogen requirements.

As will be shown in Section 3. This involves the nitrifiers "electron transport system" and is involved with the release of energy from the carbon source for use in organism growth. It happens that this electron transport system is identical to that used for respiration by organisms oxidizing organic matter aerobically, except for one enzyme. Because of this close relationship, many facultative bacteria can shift between using oxygen or nitrate or nitrite rapidly and without difficulty.

Table compares the energy yields per mole of glucose when oxygen and nitrate are used as electron acceptors. Therefore, denitrification must be conducted in an anoxic environment to ensure that nitrate, rather than oxygen, serves as the final electron acceptor. Nitrate gains electrons and is reduced to nitrogen gas, hence it is the electron acceptor. The carbon source, methanol, loses electrons and is oxidized to carbon dioxide, hence it is the electron donor.

As mentioned in Section 3. Also shown in Table is the equation of synthesis for those organisms deriving energy through nitrate respiration. Also, shown for completeness is the combined expression for oxygen respiration Equation since, if any oxygen is present, it will be used preferentially. Similar expressions can be developed for other organic sources serving as electron donors if organism yields are known.

Including synthesis Equation , the requirement is increased to 2. Biomass production can be calculated similarly: C. Overall nitrate removal Overall nitrite removal Overall deoxygenatlon 1.

The ratio includes the requirements for nitrite and oxygen, which are usually small relative to the nitrate requirement. The stoichiometric quantity of alkalinity produced is 3. Since both the alkalinity concentration is increased and the carbonic acid concentration is reduced, the tendency of denitrification is to at least partially reverse the effects of nitrification and raise the pH of the biological reaction Equation Denitrification only partially offsets the alkalinity loss caused by nitrification, since the alkalinity gain per mg of nitrogen is only one-half the loss caused by nitrification see Section 3.

Measured alkalinity production has been reported to be somewhat lower than indicated theoretically. Experiments with an attached growth process showed that the alkalinity produced averaged 2. A value for alkalinity production suitable for engineering calculations would be 3. Considering alternate commercial sources, methanol seems to continue to be the most economic choice, because price increases in alternate sources have paralleled those for methanol. The use of wastewater organics for denitrification is discussed extensively in Section 5.

Denitrification rates with wastewater oganics are approximately one-third of those when methanol is employed. Therefore, denitrification reactors must be proportionately larger.

Since using wastewater organics adds ammonia and organic nitrogen to the wastewater, the sequence of nitrification-denitrification steps must be modified to ensure that these compounds do not escape from the system.

Thus, wastewater organics are not completely interchangeable with methanol; their attraction, however, is the possible reduction in operating costs with the elimination of the need for methanol in the treatment plant.

In studies conducted for the development of the City of Tampa, Florida's treatment plant, it was shown that brewery wastes could substitute for methanol when used in both suspended growth and column denitrification systems. Solids production was found to be greater with brewery wastes than methanol, but values were not given. Removal efficiencies were similar in a parallel test of brewery wastes and methanol using columnar denitrification.

These rates compare favorably with those measured for use with methanol see Section 5. Volatile acids can be produced from wastewater organics by anaerobic fermentation or by low temperature wet oxidation. In either case, the product will contain varying amounts of ammonia nitrogen which may have to be removed in the process as described in Section 5. This caused a decrease in the settling rates of the sludge when molasses was employed.

Some of the alternatives cause greater sludge production than others. For instance, about twice as much sludge is produced per mg of nitrogen reduced when saccharose is used than when methanol is employed. On the other hand, acetone, acetate and ethanol produced similar quantities of sludge to that produced when methanol is employed.

It is free of contaminants, such as nitrogen, and therefore can be used directly in the process without taking the special precautions that must be made for use of with a waste carbon source. Second, the product is of consistent quality while wastewater sources may vary in strength and composition either daily or seasonally, complicating process control and optimization. Use of wastewater sources will require regular assaying of the source to check its purity, strength and biological availability.

Methanol also has the advantage of being nationally distributed while suitable waste carbon sources may not be geographically close to the point of use. Nonetheless, the significant disadvantage of methanol is its cost and this alone mandates the necessity of economic comparisons of alternate carbon sources. Factors considered in subsequent sections are temperature, pH, carbon concentration and nitrate concentration.

A combined kinetic expression incorporating all these factors is presented. Consideration of solids production and solids retention time is an important design consideration. A mass balance of the biomass in a completely mixed reactor yields the relationship. Investigators at the University of California at Davis, found Kj for suspended growth systems to be 0.

In most cases only net yields are reported or can be calculated from the data reported. The data of Stensel, et al. The value of Kj of 0. It is notable that when an aerobic stabilization step was incorporated into the process after anoxic denitrification, net yields reduced by almost an order of magnitude.

Only one investigator has reported growth rates, 74 all others have reported removal rates. To show the effect of temperature on growth and denitrification rates, the available data have been summarized in Figure on a basis that is normalized with respect to the rate at 20 C.

Variable Variable Variable 0. Above 20 C, four out of seven sets of data indicate that the denitrification rates find plateau values at some temperature and do not keep climbing. Differences between attached growth systems and suspended growth systems may reflect differences in the method of measurement rather than, differences in organism reaction rate.

Attached growth system removal rates were expressed on a unit surface basis while suspended growth systems were expressed per unit of biomass MLVSS in Murphy's study. If the biomass level could be measured the rate per unit of biomass may very well be similar. For instance, in one study parallel tests of suspended and attached growth systems were at 30 C. Biomass measurements were made in both systems and peak denitrification rates were found to be comparable, 0.

The earliest investigators used a nonspecific test for methanol, COD. A later more definitive investigation evaluating KM used a specific test for methanol. The value of KM was found to be very low, 0. In other words, great excesses of methanol above stoichiometric requirements need not be in the effluent from a suspended growth denitrification process to achieve nearly the maximum denitrification rates.

While there are some anomalies, it is apparent that denitrification rates are depressed below pH 6. There is some disagreement about the pH of the optima, but the data show the highest rates of denitrification are at least within the range of pH 7.

Relationships for temperature, pH, nitrate and methanol established in Sections 3. Ordinarily, the term for methanol can be neglected Section 3. Removal rates can be related to growth rates through Equation.

The safety factor concept presented in Section 3. The above equations cannot be directly applied to attached growth denitrification because the reactions take place in a more complex environment than is present in suspended growth systems. Rates of nitrate removal in the bacterial films developed in denitrification systems may be affected by the mass transfer of nitrate or methanol through the bacterial film. However, the model indicates that removal rates are most usefully expressed on a unit surface area basis and this is the procedure adopted in Section 5.

The biofilm model usefully predicts certain properties of attached growth denitrification that are significant in design. The model shows that the nitrate removal rate in attached growth sytems should not be drastically affected by adverse environmental conditions compared to effects in suspended growth systems. This has been explained on the basis that the rate of dissimilatory nitrate reduction is considerably slower than the rate of aerobic respiration.

Painter, H. Water Research, 4, No. Haug, R. McCarty, Nitrification with the Submerged Filter. Notes on Water Pollution No. Gujer, W. Mulbarger, M. Horstkotte, G. Newton, D. Wilson, Oxygen Nitrification Process at Tampa. Speece and J. Malena, Jr. Gasser, J. State University, Pa. Osborn, D. This document, "Process Design Manual for Nitrogen Control," EPA, was published in This manual covered a broad range of processes that were being evaluated and applied at the time.

The intent of the manual was to present design information for technologies that appeared to have a viable, practical application to nitrogen control. This EPA document is an interim product in the development of revised design guidance for nitrogen and phosphorus control at municipal WWTPs. This document presents findings from an extensive review of nitrogen and phosphorus control technologies and.