In this article we will discuss about treatment of waste water with advanced methods and processes. Learn about:- 1. Nitrogen and Phosphorous Control 2. Nitrification of Water 3. Denitrification 4. Biological Nitrification Methods 5. Alternative Means of Ammonia-Nitrogen Control 6. Phosphorus 7. Solids Removal.
Contents:
- Nitrogen and Phosphorous Control in Water
- Nitrification of Water
- Denitrification of Water
- Biological Nitrification Methods
- Alternative Means of Ammonia-Nitrogen Control in Water
- Phosphorus in Waste-Water
- Solids Removal from Waste-Water
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1. Nitrogen and Phosphorous Control in Water:
Nitrogen compounds often move within the environment as they change form. Most of the problems caused by nitrogen compounds occur when they enter groundwater or surface water bodies. Nitrogen reaches fresh surface water through precipitation, dustfall, surface runoff, subsurface groundwater entry and the discharge of waste-water effluents. There are also blue-green algae and some bacteria which are able to fix nitrogen from nitrogen gas in the atmosphere.
Discharges of conventionally treated domestic and industrial waste-water effluents, high in nutrients, are the main sources of nitrogen pollution in the form of ammonia-nitrogen and sometimes organic- nitrogen. Nitrogen levels in industrial effluents vary.
Some industries posing significant nitrogen pollution problems include meat processing plants, milk processing plants, petroleum refineries, ice plants, fertiliser manufacturers, synthetic fibers facilities and ammonia scouring and cleansing operations.
The use of fertilisers, concentrated animal growth farms (feedlots) and storm water runoff have also become sources of nitrogen pollution in agricultural areas. Subsurface drainage of agricultural lands carries excess nutrients to the receiving body of water (approximately 19 mg/l).
Runoff from feedlots can contain up to 300 mg/l of ammonium due to the hydrolysis of urea and up to 600 mg/l of organic- nitrogen, compared to fertiliser alone, which can contribute anywhere from 1 to 9 mg/l total nitrogen.
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Stormwater runoff in urban areas also contributes nitrogen to surface waters. Combined sewer (storm and sanitary) overflows during rainstorms can add a significant occasional nitrogen load.
Nitrogen enters soil through precipitation, dustfall, application of waste-water or fertilisers, plant residues, composting or through fixation by bacteria directly from nitrogen gas in the atmosphere. Precipitation and dust fall contain nitrogen compounds due to the combustion of fossil fuels. The planting of legume crops (peas and beans) increases the bacterial fixation of nitrogen gas and can account for 25 per cent of the total nitrogen source in such areas.
Ninety per cent of nitrogen found in soil is in the organic form since it is derived from the decomposition residues of plants and animals. Most of the remaining 10 per cent of nitrogen is as an ammonium ion (NH4+) and tightly bound to soil particles.
Septic tanks in heavily populated areas can contribute high nitrate levels to the soil at rates faster than can be assimilated. When nitrogen is supplied to soil at a faster rate than it can be assimilated by the soil’s bacteria, it will be filtered down to groundwater pockets below. Marine environments receive nitrogen compounds mostly from land drainage. It is the nitrate form which is of concern in such environments.
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Algae are the basic link in the conversion of inorganic compounds in water into organic matter. They perform this function through a mechanism known as photosynthesis in which inorganic compounds and carbon dioxide are converted to a carbohydrate energy source plus oxygen in the presence of light.
The inorganic compounds required include those with hydrogen, nitrogen, phosphorous, sulphur, potassium, magnesium calcium and trace elements such as zinc, copper, iron and molybdenum. Since all of these chemicals except for phosphorous and nitrogen are usually present, phosphorous and nitrogen normally are considered to be the limiting nutrients.
Nitrogen reaching surface waters via the various pathways discussed serve to trigger and/or sustain the growth of algae. While some algal growth is desirable since by day photosynthesis removes carbon dioxide and produces oxygen, dense growths can create problems for the environment.
When the natural balance of nutrients is thrown off by an excess supply of nitrogen and other necessary compounds, algae will grow in large numbers creating a dense mat on the water surface known as an algal bloom or scum.
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Depending on the type of algae present, the water will take on a green, yellow, red, black or turbid appearance ruining the water for recreational purposes. The odours produced by the bacterial decomposition of algae further curb recreational uses.
Similarly, growth of other water plants and some diatoms (fresh and saltwater algae types which form silica shells) also is stimulated by an abundance of nutrients.
Algae growth in water supplies requires taste and odour removal through use of filtration followed by carbon adsorption. In addition, the water supply itself must be chemically treated to kill algae growths.
Algae produces oxygen by photosynthesis during the daylight; however, carbon dioxide is released at night during the algae’s respiration process. While more oxygen is produced than respired, the presence of large quantities of algae can produce wide swings in the water’s oxygen content, which can be harmful to other oxygen-dependent organisms living there.
Similar swings (diurnal) can be found in pH levels and alkalinity of the water as a result of CO2 level changes due to algae photosynthesis and respiration.
During the night, respiration increases dissolved carbon dioxide levels, thus lowering the pH. Waters with high calcium content are able to buffer the pH effects of carbon dioxide reduction by the precipitation of calcium carbonate.
Dense algal growths limit photosynthesis to the top layer of the water. Depths of greater than three feet are shaded from the sunlight. The algae below die and are decomposed by bacteria, which in turn deplete the available oxygen.
In extreme cases, algae production leads to eutrophication, which is defined as a process of enrichment in which a water body is so fertilised with nutrients that aquatic productivity is greater than the decay rate. Eutrophication is more of a problem in relatively slow-moving waters such as surface streams, lakes and reservoirs where nutrients have time to build up.
Oxygen depletion in eutrophying water bodies has a snowball effect as other organisms die as a result, creating a further oxygen depletion problem by their own decay. Once eutrophication starts, it cannot be stopped.
Changes in organism distribution patterns occur in eutrophying water bodies. Predator-prey relationships change and ‘trash’ fish tend to flourish while desirable sport fish disappear, thus further altering the value of the water body.
Besides potentially destroying the aesthetic and recreational value of water bodies, there are some species of algae which can cause gastric disturbances in humans if the water is consumed or accidentally swallowed. Additionally, some people are allergic to planktonic algae.
Certain blue-green algae and chlorella green algae species secrete substances toxic to fish, birds and pets. Toxic blue-green algae water blooms have been known to kill hundreds of birds in just a few hours. Nutrient imbalances in the marine environment result in the production of phytoplankton blooms and dense algal growths. Marine pests such as barnacles, mussels, tube worms, parasitic fungi and jellyfish swarms also result in nitrogen- rich environments.
Nitrogen compounds themselves are known to cause problems in both surface and groundwater. A survey conducted in Illinois indicated that the highest nitrate surface water levels occurred near agricultural areas.
The nitrate form also occurs in groundwater where high concentrations can build up due to over- fertilisation or location of septic tanks near shallow wells. When water containing nitrates is used to prepare formulas for infants less than three months old, an occasional fatal blood disorder can occur called methemoglobinemia (or blue baby).
Nitrate is transformed to nitrite in the infant’s stomach, passes to the bloodstream and attaches itself to the haemoglobin molecule taking the place of oxygen with resulting suffocation. As a result of this danger, the US Public Health Service recommends a limit be placed on both nitrite and nitrate of 10 mg/l as nitrogen for drinking water.
Nitrite (N02–) also can be highly toxic to fish and other animals causing methemoglobinemia in them, too. Nitrites are very unstable ions which easily convert to nitrates however. The chloride ion has been shown to provide a protective effect against nitrite toxicity.
In general, nitrite is unimportant in waste-water or pollution. Nitrite indicates a past pollution which is or has been stabilised. Ammonia (NH3) is a significant pollutant in raw water. It reacts with chlorine to form chloroamines which reduce chlorine disinfection properties. Ammonia can be toxic to fish at certain pHs.
It exists as ammonia NH3 at pHs above 7 and as an ammonium ion (NH4+) at pHs below 7. It is the ammonia molecule that is toxic to fish. A maximum concentration of 0.02 ppm has been set as a water quality standard for freshwater aquatic wildlife.
Acute toxicity at a given pH will increase with corresponding increases in dissolved oxygen, carbon dioxide, temperature, or bicarbonate alkalinity levels within a range of 0.01 mg/l to 2.0 mg/l ammonia. At levels up to 25 mg/l, ammonia toxicity can affect all aquatic life. Diurnal pH fluctuations due to photosynthesis therefore can play a significant role in creating toxic conditions.
Nitrogen and algal growth can also interfere with industrial and water treatment operations. Nitrogen in the form of ammonia is corrosive to certain metals. Plankton and filamentous algae can clog sand filters in water treatment plants, cause foam when heated, corrode metal and create undesirable tastes, odours or oily substances which interfere with filter use.
Algae growth is a problem in cooling towers as well. As with untreated carbonaceous matter, ammonia-nitrogen exerts an oxygen demand on receiving water bodies as it slowly oxidises to nitrite and then nitrate.
Nitrification of waste-waters has been a primary concern. BOD testing emphasises control of carbonaceous matter. Concern for control of nitrogen discharges in the environment has received attention and efforts have been directed toward determining design considerations for upgrading existing carbonaceous oxidation plants to accomplish nitrification as well and towards developing independent nitrification or nitrification-denitrification systems.
2. Nitrification
of Water:
Nitrification is the first of two stages in biological nitrogen removal in which ammonia-nitrogen is biologically oxidised to nitrate, a less objectable form which does not exert an oxygen demand on the receiving water.
In the first step of nitrification, ammonium ions are oxidised to nitrite ions according to the following reaction in which 58 to 84 kcal/mole of ammonium is released:
NH4 + 1.5O2 ® H+ + H2O + NO2
The bacteria responsible for this oxidation are usually nitrosomonas, although sometimes nitrosococcus can be involved. These bacteria are aerobic autotrophs. Autotrophs, unlike heterotrophs, which obtain their energy from the oxidation of carbonaceous (organic) matter, get their energy for growth from the oxidation of inorganic nitrogenous matter and use inorganic carbon rather than organic carbon, as heterotrophs do, for cell synthesis. Being aerobic, these autotrophs require the presence of oxygen to convert the nitrogen into a usable form.
In the second step of nitrification, the nitrite ion is further oxidised by nitrobacter bacteria to nitrate releasing only 15-21 kcal/mole of nitrite oxidised as follows:
NO2– + 5O2 ® NO3–
Nitrobacter bacteria are aerobic autotrophs also. The energy freed by the nitrification reactions is used by the bacteria for growth. Since the nitrosomonas obtain more energy than the nitrobacter bacteria per mole of nitrogen oxidised, their mass in any nitrification system is greater. The nitrobacter bacteria require three times the substrate needed by nitrosomonas to get the same energy and therefore their population is 1/3 that of the nitrosomonas.
Nitrifiers grow at a much slower rate than heterotrophs. This growth rate difference can be measured by the BOD test. BOD5, which represents biochemical oxygen uptake after five days, indicates the oxidation of carbon by heterotrophs. BOD20 similarly represents the final oxygen uptake after 20 days by bacteria through nitrification.
While the nitrification reactions appear very simple, there are various intermediates and enzymes involved. The enzymes which control the rates of reactions conducted by the bacteria are sensitive to pH, temperature and substrate concentration.
Conditions necessary for proper functioning of these enzymes are reflected in the overall cell preference. The enzymes are substrate- specific. Therefore, there can be many enzyme reactions involved in the bacteria cell synthesis. The enzymes must convert the nitrogen compounds into an amino acid form before they can be used directly by the bacteria.
The growth of nitrosomonas is limited by the ammonia-nitrogen content of the wastewater, which in turn limits the nitrobacter’s growth. When the food (substrate) supply is plentiful and other conditions are favourable, growth will increase unchecked. As the population begins to exceed the available food supply, the growth rate will decline. As the food supply becomes scarce, the bacteria begin to obtain their nutrition from the dead bacteria through lysis.
Growth rates also increase with rises in temperature. The growth of nitrobacter is more greatly influenced by temperature than that of nitrosomonas. Increases are exponential in nature, reaching a maximum at some optimum temperature and then quickly falling to zero once the optimum temperature is passed.
The alkalinity and pH of a system are also important. The nitrification reaction releases carbon dioxide and free acid (H+) during the oxidation of 1 mg ammonia to nitrate which destroys about 7 mg of alkalinity as calcium carbonate (CaCO3). Depending on the alkalinity available, the reduction in CaCO3 can have a depressing effect on the pH.
When the pH drops below 7, a considerable decrease in the nitrification rate will result. This is true for both acclimated and unacclimated systems, although the short-term effect on an acclimated system is less significant. It has been demonstrated that pH drops from 7.2 to 6.4 have no immediate adverse effects. Drops to 5.8, however, create significant reductions in the nitrification rate.
An abrupt pH change from 5.8 back to 7.2 will cause an immediate rise in the nitrification rate. Therefore, pH has an inhibitory rather than toxic effect. Nitrifiers have been known to adapt to a pH range of 5.5 to 6.0. Many waste-waters do not have a sufficient alkalinity buffer and alkalinity maintenance becomes very important.
In order to obtain complete nitrification of a waste, 4.6 mg of oxygen is required for every mg of ammonia present. Generally, the nitrification rate will increase with an increase in dissolved oxygen of the system if other conditions are favourable. Studies have shown that rates are 10 per cent to 50 per cent lower at dissolved oxygen levels of 1 or 2 mg/l than at 4 to 7 mg/l.
Nitrification therefore is an important factor in stabilising the oxygen demand of the waste. Controlled biological treatment is necessary to obtain nitrification since the population of nitrifying organisms is minimal in surface waters.
Another advantage of nitrification is that a highly nitrified effluent is immune to petrification, thus helping to preserve the aesthetic quality of the receiving body of water.
Nitrification can be required when standards or limitations have been set on the receiving waters or effluent or where the reduction of the residual oxygen demand from ammonia is specifically required. The overall transformation of ammonia to nitrate will depend on how much organic nitrogen has been transformed to ammonia prior to the nitrification process.
When the total nitrogenous content of the effluent must be reduced due to regulatory limitations and/or the growth of algae in the receiving water must be prevented or reduced, denitrification is required.
3. Denitrification
of Water:
Denitrification is the second and final stage in the biological removal of nitrogen. With denitrification, nitrates are reduced to nitrogen gas.
When methanol is used as a source of carbonaceous matter, the reaction for denitrification is:
5CH3OH + 6H+ + 6NO3– ® 5CO2 + 3N2 + 13H2O
It is also possible for nitrites to be converted directly to nitrogen gas. The bacteria responsible for this transformation are heterotrophs, which derive their energy from organic chemicals through the reduction of nitrate or nitrite.
These bacteria include pseudomonas, achromobacter, bacillus and micrococcus, which are facultative, meaning that they can survive with or without the presence of oxygen. The bacteria prefer oxidising organic matter with oxygen rather than by reducing nitrite or nitrate. Therefore, anaerobic (no oxygen) conditions must be maintained in a denitritication system.
With both nitrification and denitritication of wastes, nitrogen removals of 70-90 per cent can be obtained. While such removal will serve to reduce or prevent most algal growth and eutrophication, many blue-green algal blooms cannot be affected since these algae can fix nitrogen gas for their synthesis from the atmosphere.
4. Biological Nitrification Methods
:
Biological nitrification can be achieved by several means and through various add-on treatment and upgrading methods for new and existing treatment systems.
Domestic waste effluent has been one of the main contributors of nitrogenous compounds to our environment. Nitrogen in such wastes usually exists as organic-nitrogen or as free ammonia. The nitrate and nitrite concentrations are generally small in raw wastes in relation to the other forms. Typical values of nitrogen concentrations of raw domestic wastes are shown in Table 12.1.
The most significant of the above compounds is ammonia since it can lower the dissolved oxygen of a receiving stream by nitrification. The ammonia content is derived from urea and to a lesser extent proteins. The organic-nitrogens are in the form of purines, pyrimidines, proteins urea and amino acids.
Much of the organic-nitrogen is transformed to ammonia through hydrolysis before it reaches the wastewater treatment plant. Conversion of organic-nitrogen to ammonia continues to occur within a conventional treatment plant due to the actions of heterotrophic bacteria.
The organic-nitrogen compounds usually are in a soluble form, while most of the ammonia is particulate. Primary sedimentation removes some of the particulate nitrogen forms, which are usually less than 20 per cent of the total nitrogen.
Secondary sedimentation (clarification) removes another 10-20 per cent of the nitrogen. A domestic waste with predominantly the nitrate nitrogen form has been stabilised with respect to its oxygen demand and is considered to be an old waste.
After passing through conventional biological treatment, the secondary sanitary effluent will have a typical nitrogen content of 20-50 mg/l, indicating that nitrogen just passes through such systems. However, biological nitrification does not remove nitrogen any better than conventional treatment.
Biological nitrification can be achieved in conventional carbonaceous removal systems which have been modified to combine nitrification and by the addition of a separate tertiary system. Generally if the BODs/Total Kjeldahl Nitrogen (TKN) ratio is less than 3, a separate nitrification system must be added.
If the BODs/TKN ratio is greater than 5, a combined system should work. There is no special recommendation for wastes with ratios between 3 and 5 at this time. There are varied opinions as to the merits of both methods. There are two basic concepts of biological treatment available.
These are suspended growth, where the bacterial masses are suspended in a mixed liquor and separated via clarification and attached growth, whereby the bulk of the bacterial growth occurs on a plastic or stone media and solids separation is not necessary. The types of suspended growth systems available include various activated sludge setups. Attached growth systems include trickling filters, biodiscs and fluidised beds.
Most of the studies done related to biological nitrification have been in the operation of activated sludge treatment systems, which is one of the major recognised effective methods. Important variables which have been studied extensively include the organic loading and sludge age, pH, dissolved oxygen, temperature and the presence of inhibitory substances.
The organic loading to a system is the single most important factor in nitrification. High organic loadings favour the growth of heterotrophic bacteria which then overrun the system. These bacteria have a faster rate of substrate oxidation than autotrophs.
The heterotroph’s faster growth rate is reflected in the oxygen uptake and sludge production as (bacterial) sludge is wasted faster than the nitrifiers can multiply. Therefore, with high organic (BODs) loadings, little nitrification will occur and at lower BODs loadings, approximately 10 mg/l, higher nitrification rates can occur.
Increases in the organic loading of a waste can be compensated for by increasing the retention time of the activated sludge to prevent washouts of nitrifier populations before they can become established. Sufficient oxygen supplies also must be carefully monitored in such cases to supply both the carbonaceous and nitrogenous demands.
These are important concerns in combined treatment systems. However, most of the carbonaceous oxygen demand already is removed when nitrification serves as tertiary treatment. A shock load of organics to a treatment plant would not have a significant impact on nitrification in a tertiary system.
Retention time is not important by itself, but may be used to moderate effects of changes in other parameters, such as organic loading, sludge age, dissolved oxygen and temperature. The time factor has a direct proportional relationship to the amount of nitrifiers which will be present. The average sludge retention time for a conventional activated sludge system is 3½ days. Six to ten days would be needed in a combined nitrification-carbon removal system to prevent washout of nitrifying populations.
Dissolved Oxygen (DO) has a significant effect on nitrification. The stoichiometry of the nitrification reaction shows that four atoms of oxygen are needed to oxidise one molecule of ammonia to nitrate. This translates to a 50 per cent greater oxygen requirement for good nitritication of a typical domestic waste than is required for carbonaceous removal.
Pilot-plant studies have shown that nitrification is possible at DOs of 1 mg/l and may not occur at all at DOs of 7 mg/l if other important factors are not favourable. Generally though, barring unfavourable conditions, higher DOs will increase the rate of nitrification.
Temperature affects bacterial metabolic activities, gas transfer rates (available DO) and settling characteristics of waste effluents. At temperatures above 40°C and below 5°C, nitrification rates are very slow. The optimum temperatures for nitrifying bacteria are 22°C and 30°C.
Because the temperatures in summer and warm climates are within this optimum range, treatment plants can be operated at less favourable pHs and lower substrate levels that would be required during colder conditions to achieve the same degree of nitrification.
In order to make up for the temperature difference, in winter, up to five times the summer detention time (capacity) may be required. Temperature deficiencies may also be made up by increasing the MLSS of the system and/or adjusting the pH.
Nitrification is most rapid when the pH is maintained at or slightly above neutral. Results have shown that the optimum pH for nitrification is 8.4. With all other conditions favourable, 90 per cent nitrification can be obtained at pHs of 7.8 to 8.9, but less than 50 per cent below 7.0 and above 9.8.
Further reductions in nitrification rapidly occur below a pH of 6.0 and nitrification may cease entirely below pH 5.0. Nitrifiers have a low tolerance to the hydrogen ion concentration. Breakdowns of sludge floes also have occurred when the pH drops below 7.0.
Inorganic loadings and to a lesser extent, the ammonia-nitrogen level of the waste, play a role in affecting nitritication rate. Studies have shown that at a given organic loading, increases in the influent ammonia concentration increase nitrate production levels. These increases are not proportional in nature. There is a point where the nitrate production will be limited.
The ammonia levels found in domestic waste-water are not sufficient to inhibit the rate of nitrification of such effluents. Figure 12.1 shows the rate of nitrification based on mixed liquor volatile suspended solids concentration, ammonia-nitrogen concentration, temperature and pH.
Only inhibitory effects may be felt from heavy metals concentrations of 10-20 mg/l, provided that the pH is 7.5-8.0. Precipitated metals in the sludge can redissolve if the pH drops down resulting in a system upset. Industrial discharges which are unusually high in ammonia or nitrite can exert a temporary effect on the system also. To screen for a toxicity problem, batch oxygen uptake tests may be used and batch nitrification jar tests may then be run in order to determine the best pre-treatment.
Pre-treatment can afford some protection. Heavy metals may be removed by lime additions, carbon adsorption can be used for organics and two-stage systems are viable where the organic toxics are biodegradable. Perchloroethylene and trichloroethylene are not biodegradable and toxic to nitrifiers. For toxics that come and go, breakpoint chlorination may be used at the end of the system for added safety in ammonia removal.
In activated sludge systems, waste is biologically oxidised under aerobic conditions. Large and easily settleable solids are removed prior to entry of the waste effluent into a reactor. Air (oxygen) is supplied to the reactor by diffusion of mechanical aeration.
After oxidation has occurred, the mass of bacteria which has grown is separated from the liquid in a settling tank or clarifier. Some of the solids are returned to the reactor while the rest are wasted. There are many variations of this method.
Combined Carbon and Nitrogen Removal Systems:
The first nitrification processes developed were combined systems made by modifying extended aeration systems. Combining operations is advantageous in terms of cost for existing carbonaceous systems which can be upgraded to include nitrification.
Combined carbon and nitrogen removal systems have a high proportion of influent organic loading relative to the ammonia-nitrogen concentration. As a result, the population of nitrifiers is small compared to heterotrophs. In addition, the conditions required for the carbon oxidising heterotrophs and nitrifying autotrophs are different and therefore operating parameters must be carefully controlled in combined systems.
Combined systems should be based on the sludge growth rate or solids retention time. This generally means an additional oxygen supply, longer mean cell residence times (about 10 days) and operating temperatures of 21°C to 22°C.
Contact stabilisation systems are shown in Fig. 12.2, where sludge is re-aerated prior to being recycled with the influent, will not provide complete nitrification. Even though the solids can be retained for a longer time, there is an insufficient mass developed in the reactor.
The influent entering the last pass in step aeration systems may not have enough time for hydrolysation of the organic-nitrogen to ammonia to permit nitrification. This problem can be somewhat alleviated by setting up artificial sludge re-aeration zones in the first pass by not feeding influent to that section. However, back-mixing is not prohibited, neither is short-circuiting and it is possible for ammonia bleed-through to occur.
Extended aeration plants are usually operated at such long retention times that, except during cold temperatures below 5°C to 10°C, nitrification usually is obtained if the plant is operated properly. These systems are similar to completely mixed systems except that the hydraulic retention times are 24 to 48 hours rather than 2 to 8 hours.
Completely mixed systems can provide complete nitrification at typical domestic waste concentrations. In such systems, waste is distributed uniformly to all points within the aeration tank.
Extended aeration plants are usually operated at such long retention times that, except during cold temperatures below 5°C to 10°C, nitrification usually is obtained if the plant is operated properly. These systems are similar to completely mixed systems except that the hydraulic retention times are 24 to 48 hours rather than 2 to 8 hours.
Completely mixed systems can provide complete nitrification at typical domestic waste concentrations. In such systems, waste is distributed uniformly to all points within the aeration tank.
Conventional (plug-flow) activated sludge plants can be designed to prevent back-mixing to the head of the tank by the addition of weirs since the first portion of the tank may be ineffective for nitrification.
Theoretically, plug-flow can be more efficient or require less tank volume for the same strength waste than completely mixed or extended aeration plants. However, unless a diffused air system is installed, the carbonaceous oxygen demand can overpower the nitrifier’s needs. If lime is used for flocculation before the plug-flow reactor, carbon dioxide should be added to avoid pH toxicity.
High-purity oxygen systems have been experimented with for nitrification. Since the cover prevents the escape of carbon dioxide, a buildup in the system occurs. pH levels of 6.0 are not uncommon. This has a depressing effect on the nitrification rate and even longer solids retention times are required. If pH is carefully maintained in the system, UNOX plants are no different in the degree of nitrification achievable than conventional aeration plants. The choice must be based on economic and social (odour) considerations.
Two-Stage Carbon and Nitrogen Removal Systems:
Physical separation of carbon and nitrogen removal functions can improve the control and efficiency of nitrification in certain cases. By reducing the BODs load in the first stage significantly to the influent ammonia concentration, more nitrifiers can be established.
The value of separating the heterotrophic and autotrophic populations is realised in the reduced residence time required—6 days total versus 10 days in combined systems. Plug-flow systems are the favoured activated sludge method for obtaining nitrification. Lower effluent ammonia concentrations can be achieved than in completely mixed units.
Combined and Two-Stage Systems:
Figure 12.6 is a typical two-stage activated sludge system.
While it has been demonstrated that both types of systems can be operated to achieve complete nitrification, there are advantages and disadvantages to each type. Results of modifying activated sludge systems for nitrification have been inconsistent. Problems sometimes occur with rising sludge due to the long retention times required where de-nitrification begins to take place due to lack of oxygen. Nitrogen gas bubbles cause the sludge mass to rise.
Combined systems receiving a primary effluent with a weak waste (BOD5) can be operated for nitrification satisfactorily with high MLSS down to temperatures of 10°C. However, two-stage systems are necessary in northern climates when waste-water temperatures often go below 18°C.
It has been demonstrated that two-stage systems can handle seasonal load variations where combined systems cannot. Another advantage of combined rather than single-stage systems is the lower quantity of sludge (about half the amount), which must be handled and the normally better settling characteristics of the sludge produced.
Two-stage systems tend to have more control problems since two systems are involved and the clarifier is the least stable component of a system. However, with careful monitoring, two-stage systems can be managed and a greater degree of control obtained over microbial processes.
Toxins may not be a problem to combined systems if the primary effluent is treated with a coagulant. Some people feel that toxic substances can be reduced by two-stage systems, but others feel that there is an advantage to sacrificing a carbonaceous system over a combined system.
There are some substances, though, which may be toxic to nitrifiers yet biodegradable by heterotrophs; but, there are also indications that the toxic’s advantages may not matter with domestic sewage. The first reactor in a two-stage system also may serve to reduce the possibility of organic surges to the nitrifying bacteria thus preventing overpowering carbonaceous bacterial growth. Combined systems require less land and capital expenditures than separate systems and have lower sludge disposal costs. However, the power costs for separate systems are less.
Trickling filter systems are the major type of attached growth system used to perform nitrification. Due to larger land requirements relative to activated sludge systems, trickling filters are often used for sanitary treatment in smaller cities of less than 10,000 people and less populated areas. Because of the relative stability of trickling filters compared to other biological systems, they are used for treating high-strength industrial wastes too.
Trickling filters are usually circular in form. Waste effluent is distributed by rotary sprays over the media and is collected underneath by an underdrain system.
Media to which organisms attach can be made of rock, plastic or redwood. The liquid waste percolates down through the media and the substrate and inorganic and organic waste matter is assimilated by the organisms attached to the media.
Aerobic degradation takes place on the outer portions of the biological film which develops on the media. As the mass of organisms becomes thicker, anaerobic conditions occur near the media surface. The surface organisms then die and are washed off periodically (sloughing).
Effluent peaks can be taken care of by designing the system with additional surface area and increasing recirculation rates during low-flow periods to keep the media from drying out. Clarification normally is not needed following trickling filters since the solids are maintained within the units.
Trickling filters are the homes of a varied assortment of organisms including: aerobic, anaerobic or facultative bacteria; fungi, during low pHs; algae; protozoans, worms, insect larvae and snails, which in turn feed on the bacteria. Due to the unstable characteristics of the slime, a kinetics theory for the biological activities has not yet been developed. Conclusions regarding nitrification in such systems are based on empirical results.
Nitrification and carbonaceous oxidation can occur simultaneously in trickling filters. It is better to use a media with a lower specific surface and higher voids, such as a maximum of 35 sq. ft./cu. ft., to prevent clogging in combined systems.
When nitrification is separate from carbonaceous oxidation, plugging is less of a problem for the nitrification system and application of a media with a high specific surface (up to 67 sq. ft./cu./ft.) is okay and reduces space requirements.
In two-stage systems where organic carbon and nitrogen activities occur separately, increases in nitrification have been proportional to increases in surface areas. The surface area requirements for two-stage systems increases greatly at temperatures of 7°C to 11°C than at 13°C to 19°C. Surface area requirements for nitrification also increase with the degree of ammonia-nitrogen reduction desired.
If ammonia removal must be below 2.5 mg/l, breakpoint chlorination may be added following the trickling filter, since the cost of removing such ammonia levels is much higher with trickling filters. The important variables in operating a trickling filter for either combined carbonaceous oxidation and nitrification or only nitrification as tertiary treatment include the organic loading, temperature, pH, dissolved oxygen and toxicants.
Organic loading has a significant effect on the ammonia content of the effluent. If it becomes too large, the media will be dominated by heterotrophs and significant nitrification will not occur. For combined systems, the organic loading must be reduced for cold weather operations, which increases the nitrification costs beyond those of adding a separate biological nitrification system or using a physical-chemical treatment.
It has been demonstrated that nitrification efficiency levels of 75 per cent to 100 per cent can be achieved with BODs loadings of less than 10 lbs. /1000 cu. ft./day. Efficiency diminishes at greater loadings.
Greater nitrification can be achieved with higher temperatures. Temperatures within 15° to 30°C are preferable, with 30°C the optimum. Nitrification can be achieved down to 7°C, cost factors make it impractical below 13°C. However, attached growth systems can compensate for cold temperatures better than suspended growth systems by thickening the slime.
pH:
Nitrifying bacteria are limited by pH in attached growth systems as they are in suspended growth systems.
Oxygen mass-transfer limits in bacterial slimes may limit the nitrification reaction. In order to prevent oxygen from being the limiting factor, the dissolved oxygen supply must be 2.7 times the ammonia- nitrogen concentration. This can be achieved by increasing the recirculation rate to dilute the ammonia or by adding a high-purity oxygen to increase the oxygen transfer rate.
Nitrifying bacteria are subject to the same toxicant effects whether they live in an attached or suspended growth system. However, trickling filters can handle shock loads better than activated sludge.
Hydraulic loading of the system can have a profound effect on the degree of nitrification attainable in a trickling filter. An increase from 10 MGAD to 30 MGAD has reduced nitrification from 72 per cent to 52 per cent.
Greater nitrification can be achieved at media depths of six feet.
Trickling Filters vs. Activated Sludge Systems:
Most of the work on nitrification has been done with activated sludge systems; the theory of trickling filter operation is not as precise. However, biofilm models developed indicate that trickling filters can handle adverse loading and lower temperature conditions better than activated sludge systems. Trickling filters do not quickly show ammonia breakthrough with changes in loading rate.
Biodiscs are attached growth systems which consist of a series of large-diameter plastic discs rotating on a horizontal shaft which runs across the top section of a trough-like reactor.
Only 40 per cent of the discs are in the waste-water at any time. Biological films develop on the discs, which is the media. Such systems can be operated for carbonaceous oxidation or combined for nitrification-carbonaceous oxidation. In combined systems, organic oxidation occurs on the first discs and nitrification on the last. Nitrification does not begin until most of the BODs have been removed.
Problems have arisen in application due to diurnal variations or shock loads. These load changes create an increased organic loading and some or all of the nitrifying discs convert to heterotrophic colonies. Depending on the variations involved, this situation can be prevented by derating the disc loading or installing equalisation ahead of the system. This is the disadvantage which has limited acceptance of biodiscs for nitrification.
Temperature shows no effect on nitrification rates above 13°C. The discs are normally housed to reduce the effects of external temperatures, prevent algal growths and to keep out rain or hail which can shear growths off the discs.
Work has been done regarding fluidised-bed systems. Their feasibility in large-scale applications has not as yet been accepted on a wide basis. However, since they have been shown to foster nitrification.
Fluidised-bed systems are purported to combine the best features of activated sludge and trickling filter systems into one process. As in the trickling filters, the degradation organisms coat the media, which for fluidised beds is sand grains in suspension.
Fluidised beds can handle shock loads and toxics as can trickling filters, but there is minimal sloughing of growth. Secondary clarifiers are not needed. In fluidised beds, water is passed up through a bed of sand at a velocity high enough to impart motion to or fluidise the sand. An enormous surface area is obtained using sand—greater than 1000 sq. ft/cu. ft. of reactor.
Using pilot plants handling 80000 gpd domestic waste, 90 per cent nitrification and BODs have been obtained in 45 minutes with MLVSS concentrations between 8000 and 40,000 mg/l, in 5 per cent of the space required for a comparable conventional system.
Oxygen depletion in areas of the reactor has been a significant problem in the combined nitrification organic removal application. In a study of the effectiveness of separate nitrification, 99 per cent nitrification was obtained with a raw waste containing 19.1 mg/l ammonia-nitrogen.
The nitrification performance of a fluidised bed decreased when the pH dropped below 6.0, even if sufficient alkalinity was available. Costs for operating and installing a fluidised-bed system are much less than costs for comparable conventional systems due to reduced space and retention time requirements.
5. Alternative Means of Ammonia-Nitrogen Control
in Water:
An effluent ammonia-nitrogen problem or concern can be eliminated by several other less popular methods as described in this article.
These are shallow lagoons where intensive algal growths are cultivated under aerobic conditions. Algal ponds contain algae which reduce the nitrogen content of the effluent through photosynthesis as could occur in the receiving body of water if the effluent were discharged without such treatment. The algae is harvested from the pond along with the nitrogen which has been assimilated.
This form of nitrogen removal has its application in small cities with plenty of available land. It is a seasonal treatment method dependent on light and temperature. Ice and cold winter weather significantly reduce metabolic activities and ponds go anaerobic.
During the spring which follows, hydrogen sulphide odours are released as ponds return to their aerobic states. As a result, algal ponds must be located as far as possible from existing and future residential communities. While construction and operating costs are low, the land costs and requirements can make ponds prohibitive.
Ion exchange involves the removal of ionic species (in this case, ammonium ions) from an aqueous phase. Nitrite, nitrate and organic-nitrogen cannot be removed by this method. For ammonium removal, effluent is passed through a column of clinoptilolite, a naturally occurring zeolite (resin) with a high selectivity for ammonium.
Organics can foul the resin and must be removed prior to this treatment. When all available sites are taken up by the ammonia, breakthrough will occur and the resin must be regenerated. Ammonium removals between 90 to 97 per cent have been obtained by this method.
There are some major disadvantages associated with ion exchange. These include the possibility of organic fouling and high regeneration costs. Additionally there is no ultimate disposal of the ammonium ion since it is contained in the waste brine from regeneration.
Secondary effluent can be disposed of by spray irrigation providing less soil preparation and more crops for farmers, or as a soil conditioner for marginal or drastically disturbed land. Nitrogen removals between 30 per cent and 95 per cent can be obtained.
Toxics management is important in land disposal due to the variability of the soil capacity to filter, buffer, absorb and chemically or biologically react with nitrogen. Land disposal can be a reliable method of nitrogen disposal if care is taken with application and utilisation.
If the nitrogen loading rate is too high, plants and soil bacteria will not be able to assimilate all of it and an increase in groundwater nitrate will result. Other problems associated with land disposal include large land requirements, high management costs, climate dependence, potential health hazards from bacterial contamination in groundwater and accumulation of trace elements (toxicity).
Air Stripping:
Air stripping has been used for ammonia removal from raw waste-water or digester supernatant. Ammonia-nitrogen is achieved by aeration of waste-water in a stripping tower. Effluent is pumped to the top of the tower and as it falls to the bottom, fans force air counter-current to the falling water.
Ammonia is vapourised and discharged to the atmosphere. The ammonia must be in the molecular form of NH3, not as an ammonium ion, NH4+. Up to 98 per cent removal can be obtained; however, residual levels of less than 5 ppm cannot be removed by this method. Nitrite, nitrate and organic-nitrogen levels are not affected by air stripping.
Air temperature and effluent pH also affect the amount of ammonia which can be stripped. The pH must be raised to 10 or 11 with lime and the tower must be shut down during freezing weather. Problems in efficiency and scaling occur with cold weather operation.
The ammonia removed from the effluent is discharged to the air. Rainfall washouts and nearby stormwater runoff can carry the ammonia to a receiving body of water. The net effect to the receiving water is not as bad though as it would be if the effluent were discharged directly.
In breakpoint chlorination or superchlorination, enough chlorine is added to oxidise ammonia-nitrogen to nitrogen gas. Approximately 10-20 mg/l of chlorine is needed to oxidise 1 mg/l of ammonia-nitrogen. With this method, ammonia-nitrogen levels can be brought down near zero. The effect of chlorine on organic-nitrogen is still uncertain.
Nitrite and nitrate are not removed by this method and therefore breakpoint chlorination becomes a possible follow-up to incomplete biological nitrification where low or negligible levels of ammonia are required. The chlorine is rarely added at the actual breakpoint.
Optimum breakpoint chlorination can be obtained at a pH of 10 and temperature of 30°C. Up to 90 per cent ammonia removal can be obtained in 4 to 60 hours. With chlorine gas, ammonia and to a lesser extent, organic-nitrogen, removal can be obtained. Results using chlorine dioxide are not very good for ammonia and are non-existent for organic-nitrogen.
The acidity produced by chlorination must be compensated for by lime or caustic soda addition, which increases the total dissolved solids of the effluent. Thus, breakpoint chlorination can be an expensive operation.
Biological nitrification can provide dependable ammonia removal in warm climates and warmer seasons of northern climates. If dependable nitrification is required in northern climates 365 days a year, it must be supplemented or replaced by physical-chemical treatment. Physical chemical treatments are not without drawbacks either. In many cases though, ammonia removal will be a seasonal requirement, lending itself to biological nitrification.
Nitrification may carry on to denitrification because nitrogen’s nutrient effects are often harmful as well. The performance of denitrification systems are heavily depended upon the efficiency of nitrification.
Several different means of ammonia removal have been discussed. In selecting a method most suitable, the following factors should be considered:
1. Form and concentration of the influent nitrogen compounds.
2. Required effluent quality.
3. Other existing treatment processes.
4. Costs.
5. Degree of reliability required.
6. Flexibility of the system.
6. Phosphorus
in Waste-Water:
Phosphorus in waste-water may be present in three forms: orthophosphate, polyphosphate and organic phosphorus. Typically, the majority of phosphorus enters waste-waters from kitchen grinders, human wastes and inorganic phosphate compounds used in various household detergents.
Phosphate control is of increasing concern because of its contribution to algae and aquatic growth and its interference with coagulation and lime-soda softening in concentrations as small as 0.2-0.4 mg/l.
Approximately 10 per cent of most phosphorus, which corresponds to the portion that is insoluble, is normally removed by primary settling. None of the phosphorus present in waste-water is gaseous at normal temperatures and pressures, so removal must be accomplished by precipitation.
Chemicals used to form these insoluble precipitates are lime, alum, ferric chloride and sulphate. In some cases, polymers are added to lime and alum to enhance their flocculation characteristics.
Human wastes and kitchen wastes account for 30 per cent to 50 per cent of the phosphorus in domestic waste-water. Detergents containing phosphate builders account for the remaining 50 per cent to 70 per cent. Other phosphorus sources originate in industry where they are used to control corrosion and scaling.
Also, discharges from potato processing plants, fertiliser wastes, animal feed lot wastes, dairy wastes, flour processing wastes and metal finishing wastes may contain high concentrations of phosphorus.
Total phosphorus concentration in domestic raw waste-water (using 100 gallons per capita per day waste-water flow) is found to be about 10 mg/l. This figure can be used for rough design information when no phosphorus data are available.
The increased use of phosphates and the resulting discharge into receiving waters has been cited as being responsible for the stimulation of aquatic plant growth and for speeding up the eutrophication process in our lakes. As a result, more and more states have adopted effluent concentration limits ranging from 0.1 to 2.0 mg/l phosphorous with many limits established at 1.0 mg/l for 80 per cent to 95 per cent reduction.
Therefore, there is a need for advanced waste-water treatment to remove phosphorus by chemical means since phosphorus removal obtainable by biological activity is limited. Ion exchange, reverse osmosis and recently, water hyacinths, contribute to phosphorus removal, but are more useful for the removal of nitrogen and dissolved inorganics.
During biological treatment, significant changes take place. As organic materials are decomposed, their phosphorus content is converted to orthophosphate. As a result, in a well-treated secondary effluent, a large fraction of the phosphorus is present as orthophosphate (PO4), which is fortunate since it is the easiest form to precipitate.
Materials found most practical for phosphorous precipitation are the ionic forms of aluminium, iron and calcium. Additions of polymers have also been used effectively in conjunction with lime [a form of calcium and alum (a form of aluminium)].
Aluminium Compounds—Chemical Reactions:
The principle aluminium compound used in phosphorus precipitation is ‘alum’, a hydrated aluminium sulphate.
Its reaction with PO4 is as follows:
Al2 (SO4)3 . 14 H2O + 2 PO4 ® 2 AlPO4 + 3 SO4 + 14 H2O
As you can see, AlPO4 precipitates out. It is determined that it takes 9.6 pounds of alum to remove one pound of phosphorus.
But due to the solubility of Al PO4, which is pH-dependent (optimum pH 5.5 to 6.5), bench, pilot and full-scale studies have shown that considerably higher than stoichiometric quantities of alum are necessary to meet phosphorus removal objectives as follows:
In contrast to alum, the waste-water’s pH is increased by the addition of sodium aluminate. Another possible source of aluminium is aluminium chloride, which is not as readily available as alum or sodium aluminate, but can be considered for phosphorus precipitation.
Commercial dry alum, also known as ‘filter alum’, has the chemical formula Al2(SO4)3. Alum is white to cream in colour and acidic in nature with a pH that varies between 3.0 and 3.5 in aqueous solutions having concentrations of 1 per cent to 10 per cent.
Commercially available grades of dry alum include lump, ground, rice and powdered. Dry alum is not corrosive unless it absorbs moisture from the air as could be encountered in a humid atmosphere. Most municipal treatment plants use ground or rice alum because of their superior flow characteristics.
Dry alum is generally stored in mild steel or concrete bins (with a 30-day supply), fed by conveyor into a dissolver and mixed to the proper dilution. Since alum in solution is corrosive, solution chambers should be constructed with a non-reactive material such as fiberglass or stainless steel. Solution flow from the dissolver to the point of application can be by gravity or pump.
Liquid alum is shipped in insulated tank cars or trucks. Liquid alum is heated in winter to prevent crystallisation, which can occur at 18°F with a Al2O3 strength of 8.3 per cent. Liquid alum is generally more economical than dry alum if the point of use is within 100 miles of the manufacturing plant.
Storing liquid alum is more difficult than dry alum since storage tanks, if outdoors, should be closed, heated and vented. Storage containers may be open if indoors. In any case, since liquid alum is corrosive, the storage tank must be constructed of an inert material. Feeding equipment can be similar to that of dry alum once it is mixed in the dissolver.
Dry sodium aluminate, Na2Al2O4, is non-corrosive with the pH of a 1 per cent solution being about 11.9. Requirements for Na2Al2O4 are similar to those for dry alum. Precautionary measures to be taken are similar to those of strong alkalies.
The main problem with sodium aluminate storage is it deteriorates upon exposure to air. Therefore, care must be taken to avoid the tearing of bags. Dry sodium aluminate is not available in bulk quantities. Dissolvers and feed equipment can be similar to those used for dry alum.
Liquid sodium aluminate is generally available in 30 gallon drums, tank trucks or tank cars. Liquid sodium aluminate is a strong alkali and should be handled with caution. Storage should be in shipping drums or in mild steel tanks. The storage containers should be heated in winter. Feeding equipment is similar to that used for alum.
Aluminium chloride, in most areas of the country, is not as readily available. The majority of facilities using aluminium chloride are located near petrochemical refineries since these industries use three-fourths of the total production.
Therefore, if a treatment facility is located near a refinery, a reliable economical source of aluminium chloride may be available. Solid aluminium chloride is off-white in colour and is derived from direct chlorination of scrap aluminium.
This form is semi-pure anhydrous crystals. The most common form of liquid aluminium chloride contains 28 per cent aluminium chloride by weight. Shipping, storing, handling and feeding requirements of aluminium chloride are similar to those of alum.
Chemical Reaction of Iron Compounds:
Both ferrous (Fe+2) and ferric (Fe+3) ions can be used in the precipitation of phosphate ions on a one-to- one mole ratio as follows:
Fe+3 PO4-3 ® 4 FePO4
The weight ratio of Fe+3 to P is 1.8:1 but in practice, just as in the case of aluminium, a larger amount of iron is required. The reaction with the ferrous ion is much more complicated and not fully understood. Ferrous sulphate; ferric sulphate, ferric chloride and ferrous chloride (pickle liquor) are the primary iron compounds used in phosphorus precipitation.
Liquid ferric chloride is a staining, corrosive, dark brown oily liquid. The pH of a 1 per cent solution is 2. Ferric chloride solutions are normally shipped in 3000 to 4000-gallon truckload lots and in 4000 to 10,000-gallon carload lots. Shipping concentrations of ferric chloride vary since crystallisation temperatures increase as the concentration is increased.
Storage tanks for ferrous chloride should have a free vent or vacuum relief valve and be made out of a non-corrodible material. Normally a 10-day to two-week supply should be kept on hand. Feeding equipment for ferric chloride can be similar to that used for alum except for the metering devices. Glass tube metering devices should not be used because of the tendency of ferric chloride to stain or deposit.
Ferrous Chloride (Waste Pickle Liquor):
FeCl2 as a liquid is generally available in the form of waste pickle liquor from steel processing. Acidic in nature, FeCl2 can vary from 1 per cent to 10 per cent in solution and usually averages about 1.5 per cent to 1.0 per cent in solution. Since ferrous chloride is not normally available on a continuous basis, storage and feeding equipment should be suitable for handling ferric chloride.
Fe2(SO4).XH2O is a dry partially hydrated product. Free acid of ferric sulphate is 2.5 per cent. Since ferric sulphate is actively corrosive in solution, it should be stored dry in shipping bags or in bulk in concrete or steel bins. Bin storage should be as tight as possible to avoid moisture absorption.
Feeding materials for transport of liquid ferrous chloride should be of a non-corrodible material, such as stainless steel, rubber, plastic, ceramic or lead. Dry feeding equipment is similar to that used for dry alum except the feeder should be of closed construction, thereby minimising water absorption.
Copperas or ferrous sulphate FeSO4.H2O is a by-product of pickling steel and is most common in dry form. When dissolved, ferrous sulphate is acidic. Ferrous sulphate is normally transported dry in bulk, bag or drum and is also available in bulk in a wet slate.
Storage feeding and handling systems should be similar to those used for handling ferric sulphate. However, a general precaution should be taken against mixing ferrous sulphate with quicklime since mixing may produce high temperatures and the possibility of fire.
Calcium ions react with the phosphate ion to form hydroxyapatite as follows:
3HPO4 + 5Ca + 4OH ® Ca5 (OH)(PO4)3 + 3H2O
Lime dosage for phosphorus removal is generally not calculated since the lime dose is determined by other reactions that take place when the pH of the waste-water is raised.
CaO has a density of 55 to 75 lb/ft3 and is caustic (when in a slurry). A saturated lime solution has a pH of 12.4. Lime can be purchased bagged and in bulk. The CaO content of commercially available quicklime should not be used if it is less than 75 per cent because of excessive grit and difficulties in slaking.
Bulk lime should be stored in air-tight concrete bins having a 60° slope on the bin outlet. Lime feeding equipment is usually with a belt-type feeder emptying into a lime slaker. The slaker should be of the continuous type and should include one or more slaking compartments, a dilution compartment, a grit separation compartment and a continuous grit remover.
A paste-type slaker should have a water- to-lime ratio of 2:1, an elevated temperature and a five-minute slaking time, whereas a detention-type slaker can operate with a water-to-lime ratio between 2.5:1 and 6:1 at moderate temperatures and a 10-minute slaking time.
The slaked lime is mixed in a holding tank and fed into the plant’s system by a rotor-type feeder.
Hydrated Lime:
Ca(OH)2 is slaked lime and needs only enough water added to form milk or lime. The dust and slurry of hydrated lime is caustic in nature. Storage of hydrated lime is similar to that of quicklime except that bin agitation must be provided. Feeding equipment is usually gravimetric and dilution is not important, therefore, control of the amount of water used in the feeding operation is not considered necessary.
Polymers are used in phosphorus control to enhance flocculation and settling. Characteristics of individual polymers vary widely and the manufacturer should be consulted for properties and availability.
Polymers normally can be obtained in liquid or dry forms. Dry polymers are normally stored in bags and blended with water to obtain the recommended dilution for efficient action.
Liquid polymer systems differ from dry systems only in the equipment used to blend the polymer to the proper dilution. Liquid systems, in contrast to dry systems, require no ageing and simple dilution is the only requirement for feeding. Piping and accessories are normally stainless steel or plastic.
Since the solubility of AlPO4 and other phosphorus precipitating chemicals are pH-dependent, a discussion of two chemicals frequently used to raise the pH of waste-water is warranted.
Soda ash (Na2CO3) is available in two forms: light and dense soda ash. The pH of a 1 per cent solution of soda ash is 11.2. Soda ash is available in bulk, truck, box car and hopper car. Dense soda ash is generally used in municipal applications because of its superior handling characteristics. It has little dust, good flow characteristics and will not arch in the bin. It is relatively hard to dissolve, therefore, ample dissolver capacity must be provided.
Liquid caustic soda is shipped at two concentrations: 50 per cent and 73 per cent NaOH. The pH of a 1 per cent solution of caustic soda is 12.9. Shipment is normally by tank car or truck, which is transferred to storage and diluted as necessary for feeding.
Liquid caustic soda crystallises at 53°F of when stored at 50 per cent strength. Therefore, storage tanks must be located indoors or provided with heating and suitable insulation if outdoors. If the NaOH is diluted to 20 per cent strength, its crystallisation temperature drops to -20°F. Caustic soda will tend to pick up iron when stored in steel vessels for extended periods, therefore, stainless steel, rubber, nickel alloys or plastics are normally used.
Aquatics:
The use of aquatic plants as a nutrient removal agent is not a new concept. The water hyacinth (Eichlornia crassipes) is the most researched plant in this area to date. The water hyacinth is a floating plant that covers vast areas of water surface and the plant interferes with navigation, causes flood control problems and restricts recreational activities like fishing, boating and water skiing.
Recently, 61 per cent removal of PO4-P was accomplished by growing water hyacinths after a five- day detention time. However, most tests to date have shown that after 25 to 30 days of continuous operation, phosphate removal efficiency declined until only 5 per cent to 8 per cent removal efficiency was observed.
Also, hyacinth removal efficiency is much less during the colder months. Research at Florida, found that the nutrient removal capacity of water hyacinths was directly related to pond surface area. In order to remove 44 per cent of the phosphorus, a one-million-gallon pond with 5.1 acres of water hyacinths was needed.
For the small ponds used in the Gainesville test, influent valued ranged from 3.37 to 3.44 mg/l with effluent values of 1.82 to 1.86 mg/l. Effluent had a four-day detention time and the pond had a depth of one foot. The tests were also done with a deeper pond and it was found that the shallower pond had better PO4-P removal efficiency.
Nutrient uptake by the hyacinths was good during the area growth phase and vertical growth phase, but if the hyacinths’ growth was not limited, lesser efficiencies in PO4-P removal were noted when the pond became overgrown with hyacinths.
This overgrowth could lead to anaerobic conditions in the pond. This experiment showed that nitrogen removal efficiency is better from water hyacinths than PO4-P removal efficiency; however, some phosphorus removal is obtained.
7. Solids Removal
from Waste-Water:
Removal of suspended solids and sometimes dissolved solids, may be necessary in advanced waste- water-treatment systems. The solids removal processes employed in advanced waste-water treatment are essentially the same as those used in the treatment of potable water, although application is made more difficult by the overall poorer quality of the waste-water.
As an advanced treatment process, suspended-solids removal implies the removal of particles and floes too small or too lightweight to be removed in gravity settling operations. These solids may be carried over from the secondary clarifier or from tertiary systems in which solids were precipitated.
Several methods are available for removing residual suspended solids from waste-water. Removal by centrifugation, air flotation, mechanical micro-screening and granular-media filtration have all been used successfully. In current practice, granular-media filtration is the most commonly used process.
Basically, the same principles that apply to filtration of particles from potable water apply to the removal of residual solids in waste-water. Differences in operational modes for application of these principles to waste-water filtration vs. potable water filtration may range from slight to drastic, however and the most commonly used waste-water filtration techniques.
Sand filters have been used to polish effluents from septic tanks. Imhoff tanks and other anaerobic treatment units for decades. Because they are alternately dosed and allowed to dry, the term intermittent sand-filters has been applied to this type of unit. The process is essentially the slow sand filter.
More recently, this type of filter has been applied to the effluent from oxidation ponds with considerable success. Effluent concentrations of less than 10 mg/l of BOD and suspended solids have been reported at filtering rates of 0.37 to 0.56 m3/m2-d. Filter runs in excess of 1 month are possible.
Use of intermittent sand filters in tandem with conventional secondary treatment has not been very successful. The nature of the solids from these processes results in rapid plugging at the sand surface, necessitating frequent cleaning and thus high maintenance costs. The use of intermittent filters for tertiary treatment is usually restricted to plants with small flows.
Granular-media filtration is usually the process of choice in larger secondary systems. Dual or multimedia beds prevent surface plugging problems and allow for longer filter runs. Loading rates depend on both the concentration and nature of solids in the waste-water. Filtering rates ranging from 12 to 30 m3/m2 day have been used with filter runs of up to 1 d.
Other recent innovations in filtration practices hold promise for advanced waste-water treatment. Moving bed filters have been developed which are continuously cleaned and the rate of cleaning can be adjusted to match the solids loading rate.
Another modification called the pulsed-bed filter, uses compressed air to periodically break up the surface mat deposited on a thin bed of fine filter media. Only after a thick suspension of solids has accumulated on the bed, requiring frequent pulsing, is the filter backwashed.
Both the moving bed and the pulsed-bed filters have the capability of filtering raw waste-water. A much higher percentage of solids can be removed by filtration than can be removed in primary settling. The filter effluent, containing lower levels of mostly dissolved organics, responds very well to conventional secondary treatment.
The filtered solids can be thickened and treated by anaerobic digestion, with a resultant increase in overall methane production, a possible source of energy for use within the plant.
Both secondary treatment and nutrient removal decrease the dissolved-organic-solids content of wastewater. Neither process, however, completely removes all dissolved organic constituents nor does neither process remove significant amounts of inorganic dissolved solids. Further treatment will be required where substantial reductions in the total dissolved solids of waste-water must be made.
Ion exchange, microporous membrane filtration, adsorption and chemical oxidation can be used to decrease the dissolved solids content of water. These processes, were developed to prepare potable water from a poor-quality raw water. Their use can be adopted to advanced waste-water treatment if a high level of pre-treatment is provided.
The removal of suspended solids is necessary prior to any of the processes. Removal of the dissolved organic material (by activated carbon adsorption) is necessary prior to microporous membrane filtration to prevent the larger organic molecules from plugging the micropores.
Advanced waste-water treatment for dissolved solids removal is complicated and expensive. Treatment of municipal waste-water by these processes can be justified only when reuse of the waste-water is anticipated.